Animal Biodiversity and Conservation issue 24.2 (2001)

Formerly Miscel·lània Zoològica

2001

and

Animal Biodiversity Conservation 24.2

"La tortue raboteuse" Oeuvres du Comte de Lacépède comprenant L'Histoire Naturelle des Quadrupèdes Ovipares, des Serpents, des Poissons et des Cétacés; Nouvelle édition avec planches coloriées dirigée par M. A. G. Desmarest; Bruxelles: Th. Lejeuné, Éditeur des oeuvres de Buffon, 1836. Pl. 7 Editor executiu / Editor ejecutivo / Executive Editor Joan Carles Senar

Secretaria de Redacció / Secretaría de Redacción / Editorial Office

Secretària de Redacció / Secretaria de Redacción / Managing Editor Montserrat Ferrer

Museu de Ciències Naturals (Zoologia) Passeig Picasso s/n 08003 Barcelona, Spain Tel. +34–93–3196912 Fax +34–93–3104999 E–mail mzbpubli@intercom.es

Consell Assessor / Consejo asesor / Advisory Board Oleguer Escolà Eulàlia Garcia Anna Omedes Josep Piqué Francesc Uribe

Editors / Editores / Editors Pere Abelló Inst. de Ciències del Mar CMIMA–CSIC, Barcelona, Spain Javier Alba–Tercedor Univ. de Granada, Granada, Spain Antonio Barbadilla Univ. Autònoma de Barcelona, Bellaterra, Spain Xavier Bellés Centre d' Investigació i Desenvolupament CSIC, Barcelona, Spain Juan Carranza Univ. de Extremadura, Cáceres, Spain Luís Mª Carrascal Museo Nacional de Ciencias Naturales CSIC, Madrid, Spain Michael J. Conroy Univ. of Georgia, Athens, USA Adolfo Cordero Univ. de Vigo, Vigo, Spain Mario Díaz Univ. de Castilla–La Mancha, Toledo, Spain Xavier Domingo–Roura Univ. Pompeu Fabra, Barcelona, Spain Damià Jaume IMEDEA–CSIC, Univ. de les Illes Balears, Spain Jordi Lleonart Inst. de Ciències del Mar CMIMA–CSIC, Barcelona, Spain Pablo J. López–González Univ de Sevilla, Sevilla, Spain Francisco Palomares Estación Biológica de Doñana, Sevilla, Spain Francesc Piferrer Inst. de Ciències del Mar CMIMA–CSIC, Barcelona, Spain Montserrat Ramón Inst. de Ciències del Mar CMIMA­–CSIC, Barcelona, Spain Ignacio Ribera The Natural History Museum, London, United Kingdom Alfredo Salvador Museo Nacional de Ciencias Naturales, Madrid, Spain José Luís Tellería Univ. Complutense de Madrid, Madrid, Spain Francesc Uribe Museu de Ciències Naturals (Zoologia), Barcelona, Spain Consell Editor / Consejo editor / Editorial Board José A. Barrientos Univ. Autònoma de Barcelona, Bellaterra, Spain Jean C. Beaucournu Univ. de Rennes, Rennes, France David M. Bird McGill Univ., Québec, Canada Mats Björklund Uppsala Univ., Uppsala, Sweden Jean Bouillon Univ. Libre de Bruxelles, Brussels, Belgium Miguel Delibes Estación Biológica de Doñana CSIC, Sevilla, Spain Dario J. Díaz Cosín Univ. Complutense de Madrid, Madrid, Spain Alain Dubois Museum national d’Histoire naturelle CNRS, Paris, France John Fa Durrell Wildlife Conservation Trust, Jersey, United Kingdom Marco Festa–Bianchet Univ. de Sherbrooke, Québec, Canada Rosa Flos Univ. Politècnica de Catalunya, Barcelona, Spain Josep Mª Gili Inst. de Ciències del Mar CMIMA–CSIC, Barcelona, Spain Edmund Gittenberger Rijksmuseum van Natuurlijke Historie, Leiden, The Netherlands Fernando Hiraldo Estación Biológica de Doñana CSIC, Sevilla, Spain Patrick Lavelle Inst. Français de recherche scient. pour le develop. en cooperation, Bondy, France Santiago Mas–Coma Univ. de Valencia, Valencia, Spain Joaquín Mateu Estación Experimental de Zonas Áridas CSIC, Almería, Spain Neil Metcalfe Univ. of Glasgow, Glasgow, United Kingdom Jacint Nadal Univ. de Barcelona, Barcelona, Spain Stewart B. Peck Carleton Univ., Ottawa, Canada Eduard Petitpierre Univ. de les Illes Balears, Palma de Mallorca, Spain Taylor H. Ricketts Stanford Univ., Stanford, USA Joandomènec Ros Univ. de Barcelona, Barcelona, Spain Valentín Sans–Coma Univ. de Málaga, Málaga, Spain Tore Slagsvold Univ. of Oslo, Oslo, Norway Animal Biodiversity and Conservation 24.2, 2001 © 2001 Museu de Ciències Naturals (Zoologia), Institut de Cultura, Ajuntament de Barcelona Autoedició: Montserrat Ferrer Fotomecànica i impressió: Sociedad Cooperativa Librería General ISSN: 1578–665X Dipòsit legal: B–16.278–58 The journal is freely available online at: http://www.bcn.cat/ABC

Animal Biodiversity and Conservation 24.2 (2001)

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Life–history and ecological distribution of chameleons (Reptilia, Chamaeleonidae) from the rain forests of Nigeria: conservation implications G. C. Akani1, O. K. Ogbalu1 & L. Luiselli2,3,*

Akani, G. C., Ogbalu, O. K. & Luiselli, L., 2001. Life–history and ecological distribution of chameleons (Reptilia, Chamaeleonidae) from the rain forests of Nigeria: conservation implications. Animal Biodiversity and Conservation, 24.2: 1–15.

Abstract Life–history and ecological distribution of chameleons (Reptilia, Chamaeleonidae) from the rain forests of Nigeria: conservation implications.— Five species of chameleons were observed in the continuous forest zone of southern Nigeria: Chamaeleo gracilis gracilis Hallowell, 1842, Chamaeleo owenii Gray, 1831, Chamaeleo cristatus Stutchbury, 1837, Chamaeleo wiedersheimi Nieden, 1910, and Rhampholeon spectrum (Bucholz 1874). Many original locality records are presented for each species. One species is apparently rare and confined to montane habitats (C. wiedersheimi), another species is relatively common and its habitat is generalist (C. gracilis), and the other three species are vulnerable and limited to specific micro–habitats. Female R. spectrum had clutch sizes of two eggs each and exhibited a prolonged reproductive season with oviposition likely occurring during the late phase of the dry season. Females of both C. cristatus (clutch sizes: 11–14 eggs) and C. owenii (clutch sizes: 15–19 eggs) have a shorter reproductive season with oviposition occurring most probably at the interphase between the end of the wet season and the onset of the dry season, and female C. gracilis (clutch sizes: 14–23 eggs) appeared to exhibit two distinct oviposition periods (one at the interphase between the end of the wet season and the onset of the dry season, and one at the peak phase of the dry season). Diets of four sympatric species of chameleons consisted almost exclusively of arthropods. There were significant inter–group differences at either intra–specific level (with the females of the two best studied species, i.e. R. spectrum and C. gracilis, having a wider food niche breadth than males) or inter–specific level (with a continuum of dietary specialization from the less generalist (C. cristatus) to the more generalist (C. gracilis). However, “thread–trailing” experiments indicated that activity patterns of Nigerian chameleons were relatively similar among species. The overall abundance of chameleons (as estimated from the number of specimens observed in the time unit of field effort) was relatively similar in three contrasted habitat types, but lizards were more abundant in the mature secondary forest. When greatly altered by massive logging activity, the overall abundance of chameleons in the mature secondary forest habitat declined only slightly, whereas the species diversity declined drastically. This was an effect of (i) the simultaneous extinction of three of the four species originally present in the forest plot, and of (ii) the rapid increase in abundance of a single species (C. gracilis) as a response to habitat alteration. The conservation implications of all these data are also discussed.

Key words: Chameleons, Habitat, Feeding habits, Activity, Comparative ecology, Conservation status.

Resumen Estrategia vital y distribución ecológica de camaleones (Reptilia, Chamaeleonidae) de los bosques húmedos de Nígeria: implicaciones en la conservación.— Se observaron cinco especies de camaleones en la zona de bosque ininterrumpido del sur de Nigeria: Chamaeleo gracilis gracilis Hallowell, 1842, Chamaeleo owenii Gray, 1831, Chamaeleo cristatus Stutchbury, 1837, Chamaeleo wiedersheimi Nieden, 1910 y Rhampholeon spectrum (Bucholz 1874). Se presentan muchos registros de localidad originales para todas las especies. Una especie es aparentemente rara y está confinada a los hábitats montañosos (C. wiedersheimi), otra especie es relativamente común y generalista en cuanto al hábitat (C. gracilis), y las otras tres especies son vulnerables y están limitadas a microhábitats específicos. El tamaño de cada puesta de la hembra de R. spectrum fue de dos huevos, mostrando una prolongada estación reproductora con oviposición durante la última fase de la estación ISSN: 1578–665X

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húmeda. Las hembras de C. cristatus (tamaño de puesta: 11–14 huevos) y C. owenii (tamaño de puesta: 15– 19 huevos) tienen una estación reproductora más corta y la oviposición se da con mayor probabilidad en la interfase entre el final de la estación húmeda y el inicio de la estación seca, y la hembra de C. gracilis (tamaño de puesta: 14–23 huevos) presenta dos periodos distintos de oviposición (uno en la interfase entre el final de la estación húmeda y el inicio de la estación seca y el otro durante el período más seco de la estación seca). Las dietas de cuatro especies simpátricas de camaleones consistían prácticamente de forma exclusiva en artrópodos.

Había diferencias significativas dentro del grupo tanto a nivel intraespecífico (con las hembras de las dos especies mejor estudiadas, es decir R. spectrum y C. gracilis, con un extenso nicho alimentario más amplio que los machos) o interespecíficas (con una continua especialización alimentaria desde los menos generalistas (C. cristatus) a los más generalistas (C. gracilis). Sin embargo los experimentos “thread–trailing” indican que los patrones de actividad de los camaleones de Nigeria eran relativamente similares entre especies. La abundancia de camaleones (estimada a partir de el número de especímenes observado en la unidad de tiempo de esfuerzo de campo) era relativamente similar en tres tipos de hábitats contrastados, pero el bosque secundario maduro fue el hábitat donde los lagartos fueron más abundantes. En cuanto al hábitat del bosque secundario, cuando estaba fuertemente alterado por una fuerte explotación forestal, la abundancia de camaleones disminuía sólo ligeramente, mientras que la diversidad de especies disminuía de forma drástica. Esto era debido a: (i) la extinción simultánea de tres de las cuatro especies originalmente presentadas en el bosque, y (ii) el rápido incremento en abundancia de una única especie (C. gracilis) como respuesta a una alteración del hábitat. Se discuten las implicaciones de estos datos sobre la conservación.

Palabras clave: Camaleones, Hábitat, Hábitos alimentarios, Actividad, Ecología comparada, Conservación.

(Received: 5 VI 01; Conditional acceptance: 17 IX 01; Final acceptance: 10 X 01) 1

Dept. of Biological Sciences, Rivers State Univ. of Science and Technology, P. M. B. 5080, Port Harcourt (Rivers State), Nigeria.

2 F. I. Z. V., via Olona 7, I 00198 Rome, Italy.

3 Inst. of Environmental Studies DEMETRA, Via dei Cochi 48/b, I 00133 Rome, Italy. * Corresponding author: Luca Maria Luiselli, F.I.Z.V., via Olona 7, I 00198 Rome, Italy.

E–mail: lucamlu@tin.it

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Introduction Chameleons are certainly among the most conspicuous lizards of Afro–tropical rainforest habitats (DE WITTE, 1965; BÖHME, 1985) and they have undergone a remarkable adaptive radiation in montane habitats of the central–western region of Africa, particularly Cameroon (BÖHME & KLAVER, 1981; BÖHME, 1985; BÖHME & SCHNEIDER, 1987; KLAVER & BÖHME, 1992; LAWSON, 1993; WILD, 1993, 1994; NECAS, 1994). The rainforest zone of southern Nigeria is ecologically connected with the western Cameroon forests, and the whole region is an important hot–spot for conservation because many species of flora and fauna are endemic in the area (KINGDON, 1990).

Although information on chameleons in Nigeria–Cameroon rainforests is scarce (but see WILD, 1993, 1994), many species are known to suffer from the multiple conservation problems in the area (WILD, 1993, 1994), to the extent that the conservation status of a lot of these species herein is unknown (see NARESCON, 1992 for Nigeria). At Niger Delta, the various species of chameleons are nowadays very rare (POLITANO, 1998; AKANI et al., 1999; AKANI & LUISELLI, 2001) and many populations have recently disappeared from many sites (OJONUGWA, 1973; AKANI et al., 1999). In a recent study (AKANI & LUISELLI, 2001) it was found that over 80% of adult local people interviewed about these species reported never to have seen a chameleon. Information about chameleons in other parts of Nigeria is also very scarce (but see PASQUAL, 1937; DUNGER, 1967a; BUTLER, 1986; REID, 1986).

In order to attain conservation programs for chameleons, this study provides basic information on the ecology and habitat distribution of free– ranging populations of several species inhabiting the rainforest region in Nigeria.

The aims of the study are twofold: to analyse the food niche of sympatric species, their diurnal activity, and changes in the community composition at a forest site before, and after, timbering; and to present records for new sites of these species.

Data on their conservation problems are presented and some solutions suggested. No attempt is made to review the distribution range of species (see for instance, TALBOT, 1912; PASQUAL, 1937; ROMER, 1953; DUNGER, 1967a, 1967b; BUTLER, 1986; REID, 1986; AKANI et al., 1999), but only new records data set are presented.

Materials and methods Study areas Chameleons were studied at different forest habitats of southern Nigeria as follows: moist lowland forest, deltaic freshwater swamp–forest, and coastal mangrove (see LUISELLI et al., 2000 for a detailed description of the area and LUISELLI

& ANGELICI, 2000 for the territories used during field surveys). Field work was performed from September 1994 to April 2001. The whole area has a tropical climate, with the wet season from May to September, and dry season from October to April. The rainfall peak is in June–July, and the driest period between late December and February. Annual precipitation averages between 2,000 and over 3,000 mm per year. The air temperature is generally high (average around 27–28°C), and varies little throughout the year.

The annual range of the monthly average temperature varies only between 3°C and 3.5°C. The human population density is high, and the landscape is characterised by fragmented patches of rainforest interspersed within a sea of urban centres, industry, farmlands and plantations (POLITANO, 1998).

Field methods Observations of free–ranging chameleons were made opportunistically during more general surveys for other forest vertebrates (mainly snakes; for the general survey methods, see LUISELLI & ANGELICI, 2000; LUISELLI et al., 1998, 1999, 2000).

For this paper, the following variables were recorded each time a chameleon was observed: Site A GPS “Garmin 12” was used. Habitat type, time of day, and species of each chameleon sighting were also recorded.

Biometry Snout–vent length (SVL) with a calliper to the nearest ± 1 mm. Chameleons were individually marked with a number painted in white on the back, a useful method for short–term reptile studies (e.g. see LUISELLI et al., 1996), including chameleons (KAUFFMANN et al., 1997; CUADRADO, 1998). The marked specimens were also sexed.

Body mass was not systematically recorded.

Diet Faeces from free–living specimens of chameleons were examined for this study. To obtain faeces from free–ranging specimens while minimizing handling (which may produce stress and damages to the handled animals), chameleons captured from the wild were kept separately into small terraria until defecation occurred. Faeces were analysed in the laboratory.

Prey were sorted, and identified to the lowest taxonomic level possible, and measured (to ± 0.1 mm) under a binocular microscope 10×40 equipped with a micrometer. As in a previous chameleon study (PLEGUEZUELOS et al., 1999), characteristic body parts of arthropods (mandible width for Orthoptera, head width in Coleoptera, Diptera, Hemiptera, Hymenoptera, Mantodea, and Odonata; chelicera length for arachnids) were measured in order to estimate

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the size of the prey, but attempts failed in some cases when prey fragments were broken and/or difficult to size. Diet analysis was performed on a species and sexual basis, and summarised by absolute frequency (i.e. numbers of prey belonging to a given prey type) and by frequency of appearance of a given prey type in the pellets (i.e. numbers of faecal pellets where a given prey type was observed). Diet data reported here were recorded from a single study site (Eket, Akwa–Ibom State).

Reproductive data Clutch sizes of females were determined by palpation in some cases (two cases in C. gracilis, two cases in C. cristatus, two cases in C. owenii, six cases in R. spectrum), and by dissection in other cases, when the females were found dead in the field (four cases in C. gracilis, and two cases in C. owenii).

Daily activity and foraging habits A continuous monitoring procedure of a few specimens encountered in the field was applied, thus devising a monitoring protocol quite equivalent to the “thread–trailing” strategy developed to study activity of tortoises (BREDER, 1927; HAILEY & COULSON, 1999). Three specimens, from three different species (C. owenii, C. gracilis, C. cristatus), were “thread–trailed”, for 14 days each, five hours every day (i.e. for a total of 70 hours of trailing for each specimen), by remaining at approximately 10 m distance, so as not to interfere in the chameleon’s normal activities. Binoculars (8x40) were also used to facilitate observations.

Population abundance and structure Population abundance and structure of chameleons were studied at three forest patches situated in the surroundings of Calabar (Cross River State). All these areas were surveyed during the dry seasons of 2000 (area B) and 2001 (area C), and both 1998 and 2001 (area A). Area A (50x30 m) was a mature secondary forest in the first year of study (1998), but was partially affected by industrial timbering activities during the second year of study (2001, when about 40% of the former wooded surface was cut). This allowed us to test for the effects of habitat reduction on animal abundance and species composition. Area B (approximately 120x10 m) was a riparian woodland growing along the banks of the Rhoko River. Area C (40x40 m) was a secondary bush– grassland mosaic, with plants of 50–150 cm height, surrounded by plantations (of cassava and pineapples) and farmlands. All these areas are virtually flat, at elevations of 300 m a.s.l. (area A), 385 m a.s.l. (area B) and 320 m a.s.l. (area C). In order to compare the abundance patterns of chameleons in the three areas, a “time– constrained–searching” protocol was applied. To do this, each area was carefully explored for a

total of 45 hours by two researchers, each moving independently, both by day (hrs 07.00–16.30 h) and at night (20.30–00.30 h). A balance was maintained between diurnal and nocturnal samplings at each area (50% of survey time for both diurnal and nocturnal searches). Roosting sites of each individual were noted on a scaled map of the study site.

Niche width and niche overlap These variables were estimated by using SIMPSON’s (1949) diversity index and PIANKA’s (1986) symmetric equation index which ranges from 0 (no overlap) to 1 (total overlap). In both cases, the different Orders of–prey types were used as operative taxonomical units to calculate niche widths and overlaps.

Statistical analysis An SPSS (version for Windows) computer package was used for all statistical analyses. All tests were two tailed, and alpha was set at 5%. Mean values +/- one standard deviation are provided.

Results Species distribution and types of habitats Five species of chameleons ( Rhampholeon spectrum, Chamaeleo cristatus, C. owenii, C. gracilis gracilis, C. wiedersheimi) were found at the following sites:

Rhampholeon spectrum spectrum (Bucholz 1874) Sites. Akwa–Ibom State: Eket (riverine forest along the River Kwa–Ibo (= Quo–Ibo); 04°50' N, 07°58' E), Stubbs Creek Forest Reserve (04°49' N, 08°00' E); Cross River State: Iko– Esai Forest (along the Rhoko River banks, 70 km N of Calabar; 05°28’ N, 08°23’ E), Osomba (05°21’ N, 08°24’ E), Oban (05°18’ N, 08°34’ E), Itu (05°14’ N, 07°59’ E), Ikpan Forest (30 km N of Calabar; 05°00'–05°15' N, 08°35'– 08°45' E), Akpabouyoh (04°50' N, 08°22' E); Benue State:: Ogoja (06°40’ N, 08°47’ E).

Habitat. The species is very common in wet forests with closed canopy (primary as well as mature secondary forests) either at sea level (e.g., Stubbs Creek Forest Reserve) or on hills (e.g., Oban). During daylight it was always observed on the ground, whereas it was observed on low bushes at night–time.

Rhampholeon sp.

Sites. Rivers State:: Bonny Island (04°25' N, 07°15' E).

Habitat. Three undetermined Rhampholeon were found in the stomachs of two snakes (the colubrids Rhamnophis aethiopissa and Hapsidophrys lineatus, cf. LUISELLI et al., 2000, 2001) which were captured at the coastal barrier island forest of

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Bonny island. These specimens were most probably spectrum, but as digestion was too advanced to positively identify species level, they have been placed separately from ascertained records of R. spectrum in this section.

Chamaeleo (Trioceros) cristatus Stutchbury, 1837 Sites. Edo State: Oredo Forest, 18 km SW of Benin City (06°03' N, 05°12' E); Bayelsa State: Sagbama (05°10' N, 06°05' E); Rivers State: Upper Orashi Forest Reserve (04°44' N, 07° 10' E); Akwa–Ibom State: Eket (riverine forest along the River Kwa-Ibo (= Quo–Ibo); 04°50' N, 07°58' E); Cross River State: Oban (05°18’ N, 08°34’ E); Osomba (05°21’ N, 08°24’ E); Ikpan Forest (30 km N of Calabar; 05°00'–05°15' N, 08°35'–08°45' E); Obudu Cattle Ranch (06°37' N, 08°46' E); Iko–Esai Forest (70 km N of Calabar; 05°28’ N, 08°23’ E); Okarara (04°50' N, 08°23' E); Ekang (05°23' N, 08°39' E).

Habitat. Uncommon. This species was seldom observed in mature forests and riparian forests, either at the sea level (e.g. in Niger Delta) or in hilly areas (e.g., Oban and Obudu). It was most frequently observed in low, thick, flowering bushes and, much more rarely, ground–dwelling in the leaf litter of the forest floor.

Chamaeleo (Trioceros) owenii Gray, 1831 Sites. Bayelsa State: Sagbama (05°10' N, 06°05' E), Yenagoa (05°12' N, 06°05' E), Taylor Creek Forest Reserve (05°16' N, 06°11' E); Rivers State: Upper Orashi Forest Reserve (04°44' N, 07°10' E), Otari–Abua (04°53' N, 06°41' E), Ahoada (05°04' N, 06°38' E), Buguma Creek (04°43' N, 06°50' E), Elem–Sangama (04°40' N, 06°39' E), Igbeta–Ewoama (04°34' N, 06°21' E), Degema (04°48' N, 06°48' E); Akwa–Ibom State: Eket (riverine forest along the River Kwa–Ibo (= Quo–Ibo); 04°50' N, 07°58' E); Cross River State: Calabar (04°47' N, 08°21' E), Ikpan Forest (30 km N of Calabar; 05°00'–05°15' N, 08°35'– 08°45' E), Itu (05°14' N, 07°59' E).

Habitat. Our original records confirm its presence at several sites east of the River Niger, but we failed to find this species in any locality west of the course of the River Niger. Thus, based on our data, it is possible that C. owenii would be much rarer, if occurring at all, in the western forests of Nigeria, which is also in good agreement with the range of this species at the continental level (e.g., SCHMIDT, 1919). It also seems that C. owenii is found in lowland forests, along river banks, as well as in forest-plantation mosaics and in mature secondary forests. A specimen from the Upper Orashi Forest Reserve was eaten by the snake Rhamnophis aethiopissa (LUISELLI et al., 2000).

Chamaeleo wiedersheimI Nieden, 1910 Sites. Cross River State: Obudu Cattle Ranch (06°37' N, 08°46' E).

Habitat. Our single locality record was relative

to two males observed at an open bush sub– montane area, at the border of a forested site.

Chamaeleo (Chamaeleo) gracilis gracilis Hallowell, 1842 Sites. Lagos State: Lekki (06°30' N, 04°08' E); Ondo State: Ashewele (06°48' N, 04°55' E), Ifetedo (07°27' N, 04°35' E); Delta State: Sapele (05°53' N, 05°42' E), Eku (riverine bushland along the River Benin; 05°49' N, 06°00' E); Edo State: Ologbo Game Reserve (05°55' N, 05°27' E), Oluku (05°59' N, 05°41' E); Bayelsa State: Sagbama (05°10' N, 06°05' E), Yenagoa (05°12' N, 06°05' E), Taylor Creek Forest Reserve (05°16' N, 06°11' E), Zarama–Epie (05°15' N, 06°08' E); Rivers State: Upper Orashi Forest Reserve (04°44' N, 07°10' E), Otari–Abua (04°53' N, 06°41' E), Ahoada (05°04' N, 06°38' E), Odiokwu (05°06' N, 06°37' E), Degema (04°48' N, 06°48' E), Bonny Island (04°25' N, 07°15' E), Peterside (04°29' N, 07°10' E); Anambra State: Onitscha (06°08' N, 06°46' E), Oguta (AGIP Oilfield forest; 05°58' N, 06°44' E); Abia State: Blue River banks (20 km N of Aba; 05°14' N, 07°13' E); Akwa–Ibom State: Eket (riverine forest along the River Kwa–Ibo (= Quo– Ibo); 04°50' N, 07°58' E), Ikot–Ekpene (05°12' N, 07°45' E), Uyo (05°09' N, 07°51’ E); Cross River State: Calabar (04°47' N, 08°21' E), Itu (05°14' N, 07°59' E), Akpabouyoh (04°50' N, 08°22' E), Akamkpa (05°20' N, 08°21' E).

Habitats. Our records came not only from mature lowland forests (e.g., Upper Orashi Forest Reserve), but also from altered forests (Abonnema), bushy spots surrounding farmlands and plantations (Uyo), and even derived savannas (Onitscha).

In general, chameleons of genus Chamaeleo were observed mainly along paths crossing humid secondary forests, where, usually, three vegetation strata were discernible, and with abundance of lianes and patches of undergrowth as general features. Counts of buttresses of commercially felled trees within five hectares from every spot of capture of chameleons were found to average 3.2 ± 2.7 × ha-1 (n = 46, with a range of 0 to 7 × ha-1; e.g. 2 × ha-1 at Oredo, 4 × ha-1 at Zarama, and 5 × ha -1 at Otari). Most forests where chameleons were observed had also been dissected by series of roads created for trucks to cart away the timber. The forest floor of all these sites had damp soils and considerable leaf litters in which eggs or juveniles can be hidden during the reproductive season (e.g., see BRANCH, 1988).

Habitat alteration may be locally high: for instance, Oredo forest had a history of fire episodes from neighbouring farmlands. Chameleons were also found at seasonally flooded areas: for instance, Otari forests were subject to seasonal flood from River Nun and Sombreiro River.

Reproductive biology of females All the species were oviparous. Females with shelled eggs were found in February (n = 1), May

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6

(n = 4), and June (n = 1) in R. spectrum; in August (n =1) and September (n = 3) in C. cristatus; in July (n = 3), August (n = 1), and September (n = 1) in C. owenii; in August (n = 2), September (n = 3), December (n = 1), and January (n = 2) in C. gracilis. Thus, based on these preliminary data, it can be suggested that (i) female R. spectrum show a prolonged reproductive season with oviposition likely occurring during the late phase of the dry season; (ii) females of both C. cristatus and C. owenii have a shorter reproductive season with oviposition occurring most probably at the interphase between the end of the wet season and the onset of the dry season; (iii) female C. gracilis have at least two distinct oviposition periods, one at the interphase between the end of the wet season and the onset of the dry season, and one at the peak phase of the dry season. For this latter species, it could not be excluded that reproduction may in fact take place all the year round. Clutch size was invariably two eggs in six R. spectrum, and was respectively 11 and 14 eggs in two C. cristatus, 17, 15, 17, and 19 eggs in four C. owenii, and 19, 15, 21, 23, 16, and 14 in six C. gracilis.

Diet composition, food niche width and overlap of sympatric chameleons A total of 116 chameleons were examined for food items in their faeces: 47 were R. spectrum (23{, 24}), 20 were C. owenii (14{, 6}), 15 were C. cristatus (9{, 6}), and 34 C. gracilis (16{, 18}). Results are shown in table 1. Diet composition of all the four species of chameleons consisted exclusively of invertebrates (arthropods), although a case of frog–eating was recorded in C. cristatus. In both R. spectrum and C. gracilis, females exhibited a wider food niche width than males (R. spectrum, { B = 5.181, } B = 7.782; differences significant at P < 0.01 at Mantel linear test; C. gracilis, { B = 5.397, } B = 10.020; P < 0.00001 at Mantel linear test) (fig. 1).

Concerning C. owenii, the food niche width of males (B = 6.211) and females (B = 6.897) was very similar (P > 0.7, Mantel linear test), and the same was true for C. cristatus ({ B = 5.618, } B = 4.310; inter–sexual differences: P = 0.372 at Mantel linear test).

At a inter–specific level, the four species proved to be arranged along a continuum in terms of taxonomical food niche width, from the less generalist (C. cristatus B = 4.964, after pooling data for the two sexes) to the more generalist (C. gracilis B = 7.708), with the other two species at intermediate places of this continuum (R. spectrum B = 6.481, and C. owenii B = 6.554). Trophic niche overlap estimates (in terms of taxonomical dietary composition) among species indicated that the highest similarity occurred between C. owenii and C. gracilis, whereas a general low similarity occurred between R. spectrum and each of the three Chamaeleo species (table 2).

In terms of prey size, there were significant differences among the four species (fig. 2; Kruskal–Wallis ANOVA F3,74 = 14.456, P < 0.001), and a Tukey’s HSD post–hoc test indicated that R. spectrum preyed on significantly smaller preys than the other three species, and C. cristatus preyed on significantly larger preys than the other three species.

Daily and foraging patterns As observed in other agamids from Africa (ANIBALDI et al., 1998), the chameleon’s foraging strategy consisted of frequent predatory attempts, also at very short time intervals between two consecutive trials. C. cristatus ingested 276 insects of which 53.6% were taken between 09.00 and 11.00 h local time (fig. 3); C. gracilis ingested 247 insects with 53.4% at the activity peak of 09.00–11.00 a.m. (fig. 3), and C. owenii ingested 232 insects with 50.4% gulped at the same peak period (fig. 3). Daily patterns of foraging activity were not significantly different among individuals (at least P > 0.572 in all comparisons at Mann–Whitney U–test). The feeding strategy pattern exhibited by the three chameleon species indicates that, although these lizards begin hunting very early in the morning, fewer insects are caught at this time. As the sun rises, insect hunting is intensified and the highest number is caught between 09.00 and 11.00 h.

During this period the lizards assumed their brightest colour as they also basked in the morning sun. By midday, the hunting propensity, and consequently the number of prey ingested, has dropped, probably as a result of the intense heat from the sun. By 15.00–17.00 h however, hunting ressumes, but not at the same rates as in the morning hours. They appear to be least successful towards dusk, and hunt more or less on the lower branches and undergrowth of the forest, as they return from the higher branches.

No feeding attempt was recorded during the night hours, which suggests that Nigerian chameleons are strictly diurnal, at least as far as predatory activity is concerned.

Diet of thread–trailed chameleons (fig. 4) was composed of orthopterans, coleopterans and other pterygotes (winged insects), namely Odonata, Hemiptera and Lepidoptera. Green– type insects (grasshoppers, praying mantis and stick insects) are preferred in relation to brown or multicoloured forms (e.g., Z onocerus variegatus ). Prey items identified included: Gryllotalpa africana, Locusta migratoria migratorioides, Zonocerus variegatus, Podagrica sjostedti, Empoasca sp., Rhyncophorus phoenicis, Lagria villosa, Heteroligus sp ., Diacrinia (Spilosoma) maculosa, Acraea eponina, and A. terpsicore. Furthermore, all the species exhibit similar basking habits. At night and early morning hours, chameleons tended to rest climbed on the lower branches, approximately at 30–70 cm from

7

Animal Biodiversity and Conservation 24.2 (2001)

Table 1. Dietary data recorded from faecal pellets of four species of chameleons recorded at a study plot in southern Nigeria (Eket, Akwa–Ibom State). Dietary composition is assessed by: N. Numbers of items; (n). Numbers of pellets containing that prey type. In addition, one frog was eaten by a male C. cristatus. (See text for more details.)

Tabla 1. Datos sobre la dieta registrados a partir de bolas fecales de cuatro especies de camaleones registrados en el área de estudio en el sur de Nigeria (Eket, estado de Akwa–Obom). La composición de la dieta se calcula por: N. Número de unidades; (n). Número de bolas fecales conteniendo este tipo de presa. Además, una rana fue devorada por un macho de C. cristatus. (Ver el texto para más detalles.)

R. spectrum Prey type Miriapoda Chilopoda

C. owenii

{

}

1(1)

3(3)

C. crstatus

C. gracilis

{

}

{

}

{

1(1)

0

1(1)

0

1(1)

} 1(1)

0

0

0

1(1)

0

0

1(1)

0

1(1)

4(3)

0

1(1)

0

0

1(1)

0

0

1 (1)

0

0

0

1(1)

0

0

Araneae

6(6)

15(10)

2(2)

3(2)

1(1)

5(2)

4(4)

5(4)

Opilionidae

1(1)

2(2)

0

1(1)

0

0

1(1)

1(1)

Dermaptera

1(1)

5(4)

1(1)

1(1)

1(1)

0

0

1(1)

Isoptera

23(5)

12(4)

0

0

1(1)

0

0

5(1)

Orthoptera

10(6)

15(8)

21(7)

6(4)

10(5)

8(3)

18(9)

13(8)

Diptera

1(1)

2(2)

8(6)

1(1)

0

0

4(1)

4(2)

Hemiptera

3(2)

3(2)

4(3)

0

8(3)

7(2)

1(1)

8(5)

adults

2(2)

2(2)

4(4)

0

0

0

2 (2)

7(4)

larvae

Isopoda Scorpiones

Lepidoptera 2(2)

5(5)

1(1)

0

0

0

2 (2)

6(4)

Odonata

0

0

3(2)

1(1)

12(6)

0

6 (4)

2(2)

Coleoptera (indet.)

0

3(2)

2(2)

1(1)

2(2)

0

1 (1)

3(3)

1(1)

0

0

0

0

0

0

0

0

1(1)

0

0

0

0

0

0

7(3)

21(6)

11(1)

0

0

0

0

6(2)

Carabidae Tenebrionide Formicoidea Vespoidea

0

0

1(1)

1(1)

1(1)

0

1 (1)

1(1)

Apoidea

0

0

2(2)

0

1(1)

1 (1)

1 (1)

0

Blattoidea

0

0

3(3)

1(1)

1(1)

1 (1)

1 (1)

2(2)

Mantodea

0

0

2 (2)

4 (3)

3 (3)

2 (1)

3 (3)

3(3)

the ground. As the day advances, especially by midday, they tended to migrate towards the higher branches of the trees. Skin colour begins to change until it becomes fully brightened.

After basking for about 15–25 minutes, they often withdraw and hide behind broad leaves, branches of epiphytes, or twigs, to avoid excessive heating.

Hiding behind broad leaves is also the typical antipredatory behaviour exhibited by these animals during daylight hours.

Structure of the population and abundance Densities were very similar at all study sites (table 3), but highest in the area of mature secondary forest (area A) and lowest in the area of bush–grassland mosaic (area C) (fig. 5). The species diversity was different in the three study areas, with four species (i.e. C. owenii, C. cristatus, C. gracilis and R. spectrum) recorded in the mature secondary forest (area A), two species

Akani et al.

Cumulated number of prey categories

8

20 18 16 14 12 10 8 6

R. spectrum C. owenii

4

C. cristatus

2 0

C. gracilis

1

4

7

10

13

16

19 22 25 28 31 Number of specimens

34 37 40

43

46

Fig. 1. Plot showing numbers of chameleons from which faeces were examined against cumulative number of prey categories identified from their faeces. Note that an obvious plateau phase was obtained for Rhampholeon spectrum and Chamaeleo gracilis, whereas the same curve stability was not reached in the other two species.

Fig. 1. Diagrama del número de camaleones cuyas heces se examinaron y según el número acumulado de categorías de presas identificadas a partir de las heces. Nótese que se obtuvo una meseta para Rhampholeon spectrum y Chamaeleo gracilis mientras que la misma estabilidad de la curva no se encontró en las otras dos especies.

Table 2. Food niche overlap estimates (calculated by P IANKA’s (1986) symmetric equation) of taxonomical dietary composition among sympatric species of chameleons from the study area: Rs. Rhampholeon spectrum; Co. Chamaeleo owenii; Cc. Chamaeleo cristatus; Cg. Chamaeleo gracilis.

Tabla 2. Estimación del solapamiento del nicho alimentario (calculado mediante la ecuación simétrica de PIANKA (1986) de la composición alimentaria taxonómica entre especies simpátricas de camaleones del área de estudio: Rs. Rhampholeon spectrum; Co. Chamaeleo owenii; Cc. Chamaeleo cristatus; Cg. Chamaeleo gracilis.

Rs

Co

Cc

Cg

Rs

–

Co

–

0.513

0.583

0.596

–

0.645

0.778

Cc

–

–

–

0.639

Cg

–

–

–

–

(i.e. C. owenii and R. spectrum) in the riparian woodland (area B), and one species (i.e. C. gracilis) in the bush–grassland mosaic (area C).

Concerning area A where logging was undertaken, this manipulation did not substantially reduce the abundance of chameleons (fig. 5), but had dramatic effects on the specific diversity. In fact, three of the four species became extinct after the changes on the initial habitat (i.e. C. owenii, C. cristatus, and R. spectrum), while one substantially increased its abundance (i.e. C. gracilis) (fig. 5).

Discussion Distribution and habitat All the species of chameleons found in our study had been already reported for Nigeria (e.g., TALBOT, 1912; PASQUAL, 1937; ROMER, 1953; DUNGER, 1967a, 1967b; BUTLER, 1986; REID, 1986; AKANI et al., 1999), but information was in most cases anecdotal.

Other species which are known to occur in western Cameroon and south–eastern Nigeria (cf. BÖHME, 1975; JOGER, 1982; KLAVER & BÖHME, 1997; LEBRETON, 1999; WILD, 1993, 1994) were not observed on any occasion during the present study.

9

Animal Biodiversity and Conservation 24.2 (2001)

90 R. spectrum

80

C. owenii

% prey items

70

C. cristatus C. gracilis

60 50 40 30 20 10 0 1

2

3

4 5 Prey size categories

6

7

8

Fig. 2. Prey size distributions for the four species of sympatric chameleons studied in this paper, inferred from faecal pellets. Symbols for prey size categories: 1. 0–3 mm; 2. 3–6 mm; 3. 6–9 mm; 4. 9–12 mm; 5. 12–15 mm; 6. 15–18 mm; 7. 18–21mm; 8. > 21 mm.

Fig. 2. Distribución del tamaño de las presas de cuatro especies simpátricas de camaleones estudiadas en este trabajo, obtenidas a partir de las bolas fecales. Símbolos para cada categoría de presa: 1. 0–3 mm; 2. 3–6 mm; 3. 6–9 mm; 4. 9–12 mm; 5. 12–15 mm; 6. 15–18 mm; 7. 18–21mm; 8. > 21 mm.

180 C. cristatus

Number of insectes ingested

160

C. owenii

140

C. gracilis

120 100 80 60 40 20 0 6–9

9–12

12–15 Hours (Lagos time)

15–18

18–21

Fig. 3. Diet feeding patterns of three “thread-trailed” chameleons at the Niger Delta (n = 1 for all species).

Fig. 3. Patrones alimentarios de la dieta de tres camaleones “thread–trailed” en el delta del Níger (n = 1 para todas las especies).

Akani et al.

10

%

60

C. cristatus C. gracilis

50

C. owenii

40 30 20 10 0 Orthoptera Coleoptera

Odonata Hemiptera Prey categories

Lepidoptera

Others

Fig. 4. Dietary spectrum of three specimens of chameleons at the Niger Delta, based on “thread– trailing” continuous monitoring.

Fig. 4. Espectro alimentario de tres especímenes de camaleones en el delta del Níger, basada en controles continuos en "thread–trailing".

Rhampholeon spectrum The distribution is well known in the extreme south–eastern region of Nigeria (e.g., TALBOT, 1912; REID, 1986), R. spectrum is apparently more common in forests east of the Cross River, although the records from Eket and Stubbs Creek Forest Reserve demonstrate it is not a species for which the Cross River course may represent a geographical barrier (as is the case for numerous other small vertebrates, e.g. see KINGDON, 1990). It is described as euryzonal, and abundant in premontane, submontane, and montane forests (500–1,700 m a.s.l.) of western Cameroon, but much more rare in lowland forest (L AWSON , 1993; W ILD , 1994). It is unknown whether the same distributional pattern may occur also in Nigeria, but according to several experienced hunters interviewed, this species is very common in the montane forests of the northern Cross River State. Its presence was reportedly influenced by other ecological factors, i.e. co– occurrence of food competitors: W ILD (1994) claimed that it may suffer from competition with forest toads (Bufo camerunensis), which may have a very similar dietary spectrum.

Chamaeleo cristatus Well known to occur in south–eastern Nigeria, this species was captured probably around Oban (Cross River State) by TALBOT (1912), and much

more recently both east (at Osomba, see REID, 1986), and west (at Eket, see AKANI et al., 1999) of Cross River. Original records of this study also indicate that it is also found in the western portion of the Nigerian forest zone (i.e. at Oredo, western axis of the Niger Delta). Altitude is likely not an important factor in the distribution of C. cristatus, which was in fact observed both in lowland moist forests and in hilly–montane sites. However, micro–habitat characteristics seem to be important, as both our observations and those of WILD (1994) indicate a strong preference for specific micro–habitats (low, thick, flowering bushes in our case, and “the shrub layer in primary forest” in Wild’s case), and thus a restricted habitat selection.

Chamaeleo owenii Its presence in the forests of south–eastern Nigeria is well documented (see ROMER, 1953, for Port Harcourt (Rivers State, eastern axis of the Niger Delta, and AKANI et al., 1999, for additional localities of the eastern Niger Delta). Records given here suggest that it is found in lowland moist forests and in forest–plantation mosaics.

S CHMIDT (1919) reported similar habitats for conspecifics from the former Belgian Congo.

Chamaeleo wiedersheimi The less common of the five species of chameleons observed in the present study in

11

Animal Biodiversity and Conservation 24.2 (2001)

Table 3. Numbers of chameleons observed in three study plots in southern Nigeria.

Tabla 3. Número de camaleones observados en tres áreas estudiadas en el sur de Nigeria.

Area A

Chamaeleo owenii

Before timbering

After timbering

Area B

4(3{ 1})

0

3{

Area C 0

Chamaeleo cristatus

1{

0

0

0

Chamaeleo gracilis

1}

7(3{ 4})

0

4{

2(1{ 1})

0

4(2{ 2})

0

Rhampholeon spectrum

southern Nigeria. Their distribution is linked specifically to hilly and montane sites, and montane savannas / grasslands in Cameroon (W ILD , 1993; J OGER , 1981; D UNGER , 1967b).

According to WILD (1994), C. wiedersheimi is restricted to the shrub layer in primary forest, the same as C. cristatus and Chamaeleo camerunensis Müller, 1909.

Abundance (Nº specimens / h field ef fort) effort)

0.18 C. owenii

0.16

C. cristatus C. gracilis

0.14

R. spectrum

0.12 0.1 0.08 0.06 0.04 0.02 0

A

A'

B

C

Study areas Fig. 5. Abundance of chameleons (number of specimens observed in relation to the number of hours spent in the field) in three study areas of south–easthern Nigeria. Study areas: A. Mature secondary forest before treatment; A'. Mature secondary forest after treatment; B. Riparian woodland; C. Bush–grasland mosaic.

Fig. 5. Abundancia de camaleones (número de especímenes observados con relación al número de horas empleadas en el campo) en tres áreas de estudio del sudeste de Nigeria. Áreas de estudio: A. Bosque secundario maduro antes de ser sometido a explotación; A'. Bosque secundario maduro después de ser sometido a explotación; B. Bosque maduro; C. Mosaico de bosque y pradera.

Akani et al.

12

Table 4. Data on chameleon trade in local markets of the Niger Delta Basin during the year, 2000: N. Number of chameleons on display; P. Unit selling price (in "Nairas"); $. US dollar equivalent (all of them used for traditional medicine).

Tabla 4. Datos sobre el comercio en mercados locales de la cuenca del delta del Níger a lo largo del año 2000: N. Número de camaleones observados; P. Precio por unidad (en "Nairas"); $.

Equivalente en dólares americanos (todos ellos utilizados en la medicina tradicional).

Date

Location of market

N

Suppliers

17 V 2000

Mile 1, market P. H.

1

Hunter from Bayelsa

2,500

23.0

19 V 2000

Mile 3, market P. H.

3

Hunters from Omoku and Biseni

2,700

25.0

4 VII 2000

Otari

4

Farmers / hunters from Ogbema Abua

3,000

27.0

Yenagoa

2

Farmers / hunters from Zarana Epie

2,800

26.5

Benin City

6

Hunters from Mosoga, Oredo and Aghara

3,200

29.0

23 VIII 2000 27 X 2000 Total

P

($)

16

Chamaeleo gracilis This a typical and common forest species of south–eastern Nigeria (e.g., D UNGER , 1967a; BUTLER, 1986; AKANI et al., 1999), but is certainly found also in the savannas (e.g., see DUNGER, 1967b). As in other countries (e.g., in Cameroon and Liberia, see SCHMIDT, 1919; WITTE, 1965), it is a habitat generalist.

females from Nigeria were considerably smaller than those reported in the layman’s literature for either C. cristatus (16–37 eggs, see KAIWI, 2000) or C. gracilis (20–40 eggs, see KAIWI, 2000), and also a free–ranging female C. gracilis from Belgian Congo had 60 eggs (SCHMIDT, 1919).

Reproductive biology

The four studied chameleon species of tropical Nigeria fed almost exclusively upon arthropods. It is consistent with data on R. spectrum from south– eastern Nigeria (REID, 1986) and Cameroon (WILD, 1994), with data on C. owenii from the former Belgian Congo (SCHMIDT, 1919), and in general with dietary data on chameleons elsewhere (e.g., P LEGUEZUELOS et al., 1999). It is noteworthy however, that a single case of vertebrate–eating (a froglet) by a C. cristatus was recorded. Although this predation event appears exceptionally unusual in the wild (at least considering data presented in table 1), Nigerian C. cristatus are known to readily eat frogs and newly metamorphosed toads in captivity (REID, 1986), which suggests that they have a natural “attitude” for preying upon small amphibians. In any case, it is certainly premature to stress that amphibian–eating is a trophic niche difference between C. cristatus and the other three sympatric species.

In the two better studied species (i.e. R. spectrum and C. gracilis), there were significant inter–sexual differences in dietary habits, with the females exhibiting a wider food niche width than males.

The reasons for this inter–sexual difference are unknown, and are likely not correlated with any sexual size dimorphism (SSD) as SSD is certainly not big enough to justify such an assumption in both

Data on reproduction timing of female R. spectrum are fully consistent with data from conspecifics at Mt. Kupe (western Cameroon, cf. WILD, 1994), and our data on C. cristatus are also in agreement with the single record available on the reproduction of a Nigerian conspecific (REID, 1986).

DUNGER (1967b) recorded seven newly–hatched C. gracilis on 20th May on a low bush in Jos (09°55' N, 08°53' E), and thus suggested that hatching occurs during the early rains. It is quite consistent with a period of oviposition at the peak of the dry season, as indicated by the original data presented in this study. Concerning R. spectrum, an invariable clutch size of two eggs was also detected by WILD (1994) in western Cameroon, which suggests that it is a general pattern for this species (but a clutch size of 2–5 eggs is reported for this species by KAIWI (2000), 1–3 eggs in the closely related Rhampholeon boulengeri from the former Belgian Congo [SCHMIDT, 1919], and 3 eggs in a Tanzanian Rhampholeon uluguruensis [TILBURY & EMMRICH, 1996]). Clutch sizes of 15 and 17 eggs were reported in two C. owenii from the former Belgian Congo (SCHMIDT, 1919), which is very consistent with the original data reported in this study. However, clutch sizes of free–ranging

Feeding ecology

13

Animal Biodiversity and Conservation 24.2 (2001)

species (WILD, 1994). At the inter–specific level, the apparent continuum of dietary specialisation from the less generalist species (C. cristatus) to the more generalist species (C. gracilis) may suggest that a true phenomenon of food resource partitioning occurs between sympatric forest chameleons. In this regard, an obvious pattern of resource partitioning among species was also seen with regard to prey size, with R. spectrum and C. cristatus at the two extremes of the continuum, and C. gracilis being the most generalist species. In general terms, C. gracilis appears thus not only the most generalist species in terms of habitat, but also in terms of prey type and prey size. The lesser interspecific overlap was observed between R. spectrum and C. owenii. It is likely that it depended on the combined effect of size and habits (R. spectrum is mostly terrestrial–dwelling, whereas C. owenii is mainly arboreal).

Despite these interspecific differences found in diet composition, our data should be considered as a “snapshot” of the chameleon diet. A considerably higher variation is likely.

Taking into account that the diet composition of thread–trailed specimens and the composition of faeces was very similar, it is concluded that “thread–trailing” is a very good experimental procedure to study chameleon ecology in sites where these lizards are rare or endangered, and when it is particularly difficult to establish an experimental protocol involving the capture of many specimens.

Structure of the population and abundance Density of chameleons in the forest habitats of southern Nigeria was low and certainly much less than that observed in other African regions (e.g. Madagascar, cf. KAUFFMANN et al., 1997).

Moreover, the species diversity at single sites clearly appears much less than that observed in the forests of the adjacent Cameroon (WILD, 1993, 1994). For instance, R. spectrum sympatric was observed in this study with up to three Chamaeleo species in southern Nigeria, whereas WILD (1994) found up to ten different species in Cameroon forests. However, the diversity of sympatric chameleon species of the Biafran forests is certainly influenced by the the relative elevation: WILD (1994) found seven sympatric species of Chamaeleo in montane areas around Mt. Manenguba, but only four species in lowland Cameroon forests.

Despite the differences in the structure of the vegetation among sites, the overall abundance of chameleons was similar, although densities were higher at mature secondary forests. After logging, the overall abundance of chameleons declined only slightly, whereas the species diversity declined drastically. These findings support the hypothesis that the population of free–ranging chameleons declines in rainforest habitats where rapid changes in the environment are occurring.

Implications for conservation Results of this study provide new insights with important implications for the conservation of chameleons in Nigeria (and probably, West Central Africa). All chameleons in the Niger Delta and in south–eastern Nigeria inhabit very fragile ecosystems such as mature secondary forests and riparian forests. Human alteration of the remnant forest is likely to further destroy this habitat type in the next decades (OLAJIDE & ENIANG, 2000), especially due to the continuing exploration of unaltered biota for oil industry development, which is the main economic income for the country. In addition, hundreds of C. gracilis are captured and later desecated every year from the forests around Lagos and Ibadan, and sent to public markets in Calabar, Uyo (and probably other towns) where they are traded for local medicine or juju practices at a low price (Naira 150–400, i.e. approximately 2–4 US dollars, on March 2001) (Akani et al., unpublished). Both factors are likely to have a tremendous influence on species conservation in the near future.

Species diversity changed dramatically in one of our study sites after human impact. Moreover, the only surviving species (C. gracilis, i.e. the ecologically most versatile species of the forest region of Nigeria), which was rare before timbering, became much more common (i.e. largely dominant in the chameleon community) after logging. The present results fully agree those of GRAY (1989) who suggested that rare and more specialised species tend to disappear with habitat loss, whereas moderately common species (more habitat generalist) would increase in abundance. Accordingly, C. owenii, C. cristatus, and R. spectrum may be dramatically affected by habitat loss and forest fragmentation, whereas C. gracilis might even benefit from this situation, possibly also because of less inter–specific competition with the other chameleon species.

The broad similarity in activity patterns and feeding habits among two “fragile” species (C. owenii and C. cristatus) versus one “versatile” species (C. gracilis) indicate that these life–history attributes cannot influence differently the species–specific persistence in altered habitats, and so it is likely that prey resource availability does not play an important role in determining the scarcity of some particular species in altered habitats. The rarity of chameleons in the study area is attributed to: 1. Habitat destruction / modification through lumbering and cultivation; 2. Local fires in bush areas, especially at the end of the dry season (March–April), for agricultural purposes; 3. Illegal trade because of the great demand for chameleons following the increasing market values (N 2,500–N3,000; approx. 23–29 US dollar (see table 4); 4. Capture and desiccation for traditional medicinal purposes. The authors wish to encourage State and Local Governments to establish some forest reserves specifically

14

targeted at chameleon preservation, and enact and adequately enforce laws to persecute the handling or trading of chameleons.

Acknowledgements We are grateful to “Prime Energy Resources PLC” (Port Harcourt) for logistic support and financial assistance throughout this project. Data were also collected during environmental projects supported by E. N. I. S. p. A. and by “T. S. K. J. Nigeria Ltd.”. F. M. Angelici, D. Capizzi, E. A. Eniang, and E. Politano contributed with many field–data, Ch. Amadi (Port Harcourt) typed the original draft version, and our referees, M. Cuadrado and an anonymous person, immensely improved an earlier version of the manuscript.

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"La tortue greque" Oeuvres du Comte de Lacépède comprenant L'Histoire Naturelle des Quadrupèdes Ovipares, des Serpents, des Poissons et des Cétacés; Nouvelle édition avec planches coloriées dirigée par M. A. G. Desmarest; Bruxelles: Th. Lejeuné, Éditeur des oeuvres de Buffon, 1836. Pl. 7

Editor executiu / Editor ejecutivo / Executive Editor Joan Carles Senar

Secretaria de Redacció / Secretaría de Redacción / Editorial Office

Secretària de Redacció / Secretaria de Redacción / Managing Editor Montserrat Ferrer

Museu de Zoologia Passeig Picasso s/n 08003 Barcelona, Spain Tel. +34–93–3196912 Fax +34–93–3104999 E–mail mzbpubli@intercom.es

Consell Assessor / Consejo asesor / Advisory Board Oleguer Escolà Eulàlia Garcia Anna Omedes Josep Piqué Francesc Uribe

Editors / Editores / Editors Antonio Barbadilla Univ. Autònoma de Barcelona, Bellaterra, Spain Xavier Bellés Centre d' Investigació i Desenvolupament CSIC, Barcelona, Spain Juan Carranza Univ. de Extremadura, Cáceres, Spain Luís Mª Carrascal Museo Nacional de Ciencias Naturales CSIC, Madrid, Spain Adolfo Cordero Univ. de Vigo, Vigo, Spain Mario Díaz Univ. de Castilla–La Mancha, Toledo, Spain Xavier Domingo Univ. Pompeu Fabra, Barcelona, Spain Francisco Palomares Estación Biológica de Doñana, Sevilla, Spain Francesc Piferrer Inst. de Ciències del Mar CSIC, Barcelona, Spain Ignacio Ribera The Natural History Museum, London, United Kingdom Alfredo Salvador Museo Nacional de Ciencias Naturales, Madrid, Spain José Luís Tellería Univ. Complutense de Madrid, Madrid, Spain Francesc Uribe Museu de Zoologia de Barcelona, Barcelona, Spain

Consell Editor / Consejo editor / Editorial Board José A. Barrientos Univ. Autònoma de Barcelona, Bellaterra, Spain Jean C. Beaucournu Univ. de Rennes, Rennes, France David M. Bird McGill Univ., Québec, Canada Mats Björklund Uppsala Univ., Uppsala, Sweden Jean Bouillon Univ. Libre de Bruxelles, Brussels, Belgium Miguel Delibes Estación Biológica de Doñana CSIC, Sevilla, Spain Dario J. Díaz Cosín Univ. Complutense de Madrid, Madrid, Spain Alain Dubois Museum national d’Histoire naturelle CNRS, Paris, France John Fa Durrell Wildlife Conservation Trust, Trinity, United Kingdom Marco Festa–Bianchet Univ. de Sherbrooke, Québec, Canada Rosa Flos Univ. Politècnica de Catalunya, Barcelona, Spain Josep Mª Gili Inst. de Ciències del Mar CMIMA–CSIC, Barcelona, Spain Edmund Gittenberger Rijksmuseum van Natuurlijke Historie, Leiden, The Netherlands Fernando Hiraldo Estación Biológica de Doñana CSIC, Sevilla, Spain Patrick Lavelle Inst. Français de recherche scient. pour le develop. en cooperation, Bondy, France Santiago Mas–Coma Univ. de Valencia, Valencia, Spain Joaquín Mateu Estación Experimental de Zonas Áridas CSIC, Almería, Spain Neil Metcalfe Univ. of Glasgow, Glasgow, United Kingdom Jacint Nadal Univ. de Barcelona, Barcelona, Spain Stewart B. Peck Carleton Univ., Ottawa, Canada Eduard Petitpierre Univ. de les Illes Balears, Palma de Mallorca, Spain Taylor H. Ricketts Stanford Univ., Stanford, USA Joandomènec Ros Univ. de Barcelona, Barcelona, Spain Valentín Sans–Coma Univ. de Málaga, Málaga, Spain Tore Slagsvold Univ. of Oslo, Oslo, Norway

Animal Biodiversity and Conservation 24.1, 2001 © 2001 Museu de Zoologia, Institut de Cultura, Ajuntament de Barcelona Autoedició: Montserrat Ferrer Fotomecànica i impressió: Sociedad Cooperativa Librería General ISSN: 1578–665X Dipòsit legal: B–16.278–58

Animal Biodiversity and Conservation 24.2 (2001)

17

Practical implementation of optimal management strategies in conservation programmes: a mate selection method J. Fernández1, M. A. Toro2 & A. Caballero1,*

Fernández, J., Toro, M. A. & Caballero, A., 2001. Practical implementation of optimal management strategies in conservation programmes: a mate selection method. Animal Biodiversity and Conservation, 24.2: 17–24.

Abstract Practical implementation of optimal management strategies in conservation programmes: a mate selection method.— The maintenance of genetic diversity is, from a genetic point of view, a key objective of conservation programmes. The selection of individuals contributing offspring and the decision of the mating scheme are the steps on which managers can control genetic diversity, specially on "ex situ" programmes. Previous studies have shown that the optimal management strategy is to look for the parents’ contributions that yield minimum group coancestry (overall probability of identity by descent in the population) and, then, to arrange mating couples following minimum pairwise coancestry. However, physiological constraints make it necessary to account for mating restrictions when deciding the contributions and, therefore, these should be implemented in a single step along with the mating plan. In the present paper, a single–step method is proposed to optimise the management of a conservation programme when restrictions on the mating scheme exist. The performance of the method is tested by computer simulation. The strategy turns out to be as efficient as the two–step method, regarding both the genetic diversity preserved and the fitness of the population.

Key words: Inbreeding, Genetic diversity, Genetic drift, Reproductive traits.

Resumen Aplicación práctica de estrategias de manejo óptimo en programas de conservación: un método de selección de apareamientos.— El mantenimiento de la diversidad genética es, desde un punto de vista genético, un objetivo fundamental en programas de conservación. La selección de los individuos que dejarán descendientes y la decisión del esquema de apareamiento son los pasos en los que el conservador puede controlar la evolución de la diversidad, especialmente en programas "ex situ". Se ha demostrado que la estrategia óptima consiste en buscar las contribuciones de los reproductores que den el mínimo parentesco global (probabilidad de identidad por descendencia de la población) y, posteriormente, determinar las parejas utilizando el método de apareamientos de mínimo parentesco. Sin embargo limitaciones fisiológicas y reproductivas pueden impedir que los apareamientos propuestos se lleven a cabo. Por esta razón, sugerimos la aplicación de un procedimiento que decida las contribuciones y el diseño de apareamientos en un solo paso. Mediante simulación con ordenador comparamos la eficiencia de dicho método frente al diseño óptimo en dos etapas. El procedimiento resultó ser tan eficiente como el método en dos pasos, tanto en el mantenimiento de variabilidad genética como en los niveles de eficacia biológica de la población.

Palabras clave: Consanguinidad, Diversidad genética, Deriva genética, Caracteres reproductivos.

(Received: 4 X 01; Conditional acceptance: 23 I 02; Definitive acceptance: 12 II 02) 1Dept.

de Bioquímica, Genética e Inmunología, Fac. de Ciencias, Univ. de Vigo. 36200 Vigo, España (Spain).

Dept. de Mejora Genética Animal, Inst. Nacional de Investigaciones Agrarias, 28040 Madrid, España (Spain). * Corresponding author: Armando Caballero, Área de Genética, Fac. de Ciencias, Univ. de Vigo, Campus Lagoas– Marcosende 36200 Vigo, España (Spain).

2

E–mail: armando@uvigo.es

ISSN: 1578–665X

© 2001 Museu de Zoologia

18

Introduction From a genetic point of view, conservation programmes have two basic objectives: first, to reduce the increase in inbreeding and its collateral effects on fitness and other traits that can threaten the survival of the population; and second, to maintain the highest level of genetic variability in order for the population to be able to face future environmental changes, avoid adaptation to captive conditions (if referring to "ex situ" programmes) and assure a possible long–term response to selection for traits of interest (BALLOU & LACY, 1995; OLDENBROEK, 1999; BARKER, 2001).

Measures of genetic variability To know to which extent a population is threatened from a genetic point of view, and to monitor the performance of a conservation programme, it must be possible to measure the amount of genetic variability present in a group of individuals. From an evolutionary perspective, a straightforward measure of variability is allelic diversity, i.e. the number of different alleles in a locus (or the average over loci) carried by the population. The increasing availability of highly polymorphic neutral molecular markers provides a powerful tool to trace allelic diversity.

Another proposed measure of genetic variability is the expected heterozygosity, usually called gene diversity (NEI, 1973). This represents the proportion of heterozygotes expected if the population were in Hardy–Weinberg equilibrium. Again, if we refer to several loci, gene diversity is the average over loci. Contrary to allelic diversity, gene diversity is not only influenced by the number of alleles present in the population, but also by their frequencies. High levels of heterozygosity also mean high levels of additive genetic variance and, thus, greater potential responses to selection (FALCONER & MACKAY, 1996). As in the case of allelic diversity, the only information available, in most cases, is that from the allelic frequencies in neutral molecular markers.

Another estimate of the amount of diversity preserved in a population can be found via the concept of number of founder genome equivalents. By definition (LACY, 1995), this is the number of founder individuals (individuals on top of the pedigree) required to explain the genetic variability observed in the present population, accounting for the genetic drift occurring during pedigree development. The number of founder genome equivalents is directly related to gene diversity and effective population size (LACY, 1995; CABALLERO & TORO, 2000).

Optimal management The loss of alleles in a small population, like those under conservation, is mainly driven by genetic drift, i.e. the random fluctuation of allelic

Fernández et al.

frequencies due to finite population size (F ALCONER & M ACKAY , 1996). Therefore, any strategy directed to the minimisation of genetic drift will keep the largest number of alleles.

Classical population genetics theory provides such methods, like the minimisation of variance in parents’ contributions to the next generation (GOWE et al., 1959; WANG, 1997).

An indirect measure of genetic variability of the population is provided by the degree of relationship between individuals. It seems logical that a good strategy to maintain genetic diversity is to reduce kinship relationships in the population as much as possible, as less related individuals are more likely to carry different alleles. The common way of controlling relationships is through the coefficient of coancestry (kinship). As first defined

by MALÉCOT (1948), the coefficient of coancestry (ƒij) between individuals i and j is the probability of identity by descent of two alleles taken at random, one from each individual, at any locus.

Two alleles are identical by descent when they are copies of a unique allele of a common ancestor.

If the pedigree of the population is known, coancestries between any pair of individuals can be calculated following very simple rules.

Analogously, using information from markers, we can define the molecular coancestry as the probability that two alleles taken at random at the marker locus, one from each individual, are identical in state (i.e., equal). Therefore, molecular coancestry measures the heterozygosity in a number of known loci. If a model where all alleles in the founder population are different at an infinite number of loci is assumed, molecular and pedigree coancestries are the same.

Otherwise, two alleles of a marker can be identical in state but not by descent. The relationship between both types of coancestries is, in principle, simple and defined by

E(ƒMi ) = pi2 + pi (1 – pi)ƒ where ƒMi is the molecular coancestry due to allele i, pi is the frequency of this allele in the base population, and ƒ is the genealogical coancestry. Different methods to estimate identity by descent from molecular information have been developed from this relationship (see LYNCH & RITLAND, 1999, for a review). Finally, if information from both molecular markers and pedigree are available, these can be used jointly to calculate the coancestry conditional on markers (TORO et al., 1999; WANG, 2000).

Several authors (B ALLOU & L A CY , 1995; M EUWISSEN , 1997; F ERNÁNDEZ & T ORO , 1999; CABALLERO & TORO, 2000) have demonstrated, theoretically and by computer simulation, that the most effective method to maintain genetic diversity is to find the contributions of parents so that global coancestry is minimised. Global coancestry is defined as the average pairwise coancestry among all possible combinations of

19

Animal Biodiversity and Conservation 24.2 (2001)

individuals, including self–coancestries. Every coancestry must be weighted by the product of contributions of the two particular individuals.

In this way, not only are individuals selected, but also the optimal number of offspring they should contribute. The optimality of this method comes from several facts (see CABALLERO & TORO, 2000).

As gene diversity in the population is equal to 1 – ƒ, minimising global coancestry will maximise both expected heterozygosity and the number of founder genome equivalents. Furthermore, this strategy implies equalisation of contributions from all previous generations to the present one, thereby maximising the effective population size. An extension of this result to subdivided populations has been shown by CABALLERO & TORO (in press).

Mating scheme Once the parents of the next generation and their contributions have been determined, the second decision a manager should take, is the way in which those parents should be mated to generate the offspring. Different strategies have been proposed to help in the achievement of conservation aims (WRIGHT, 1921; KIMURA & CROW, 1963; TORO et al., 1988; CABALLERO et al., 1996; SONESSON & MEUWISSEN, 2000). Using computer simulations, FERNÁNDEZ & CABALLERO (2001) showed that, provided that contributions have been arranged to yield the minimum global coancestry, the mating scheme is less determinant, although the mating of pairs with minimum coancestry (TORO et al., 1988) has a slight superiority. This latter procedure consists of finding the combination of couples with the minimum average coancestry between the male and female involved in each mating.

simultaneously, in a single step, and take the restrictions into account.

BALLOU & LACY (1995) proposed a single–step method based on the minimisation of mean coancestry. This is an iterative procedure to find not only which parent will contribute an offspring but also the specific matings among them. As originally proposed, the method is quite efficient in the preservation of genetic diversity, but as it shows a tendency to mate close relatives, populations under this management procedure suffer a great decline in fitness, specially in the first generations, in relation to the high increase of inbreeding (FERNÁNDEZ & CABALLERO, 2001).

Therefore, the implementation of this strategy would increase the probability of population extinction and should thus be discouraged.

Further improvements (avoidance of close relatives’ matings) suggested by BALLOU & LACY (1995) do not completely solve the problem, as some side–effects arise from the influence of mating design on the selection step (FERNÁNDEZ & CABALLERO, 2001).

In the animal breeding field, another method has been proposed to decide the parents and the mating scheme in a single step, the so called "mate selection" (ALLAIRE, 1980; TORO & PÉREZ– ENCISO, 1990; KLIEVE et al., 1994). In the present paper the use of mate selection in conservation programmes is proposed. Using this method, all reproductive and physiological restrictions are taken into account while the Ballou and Lacy method disadvantages are absent. Computer simulations were carried out to compare this strategy with the two–step design. Examples of restricted and unrestricted solutions are also presented to illustrate the performance of the method.

Practical considerations

Methods

Theoretically, the solutions arising from the application of optimal strategies cover a large range of possibilities, from all offspring generated by a single couple to all individuals contributing equally. The same occurs with the mating scheme, where all combinations are possible. However, the practical implementation to particular conservation programmes may be restricted. The first restriction is the number of offspring an individual can contribute. If dealing with plants or animals such as fishes, this may not be a constraint, but programmes on mammals or birds should take into account that a female can provide only one or a few offspring each reproductive season. Physiology also represents a restriction in the mating scheme as a female is generally fertilised by one male only. If performing a two–step conservation programme, it is likely that the optimal contributions will not be compatible with the physiological restrictions on the mating scheme. All these problems may be avoided if selection and mating are set up

Mathematical models Two–step procedure The selection stage of the two–step procedure consists of minimising the global coancestry (from pedigree, from molecular markers or from both, depending on the availability). This process is reduced to find the parental contributions that yield the minimum value of the following function N

N

i=1

j=1

5 5xi xj ƒij

(1)

where xi is the number of offspring to be generated by individual i, ƒij is the coancestry coefficient between individuals i and j, and N is the number of individuals. Some constraints must be included to find reasonable solutions: (i) only positive and integer values of the variables are allowed

xi m 0 xi integer

i = 1,...,N

Fernández et al.

20

as no fractional or negative numbers of offspring are possible; (ii) the sum of contributions from parents must be twice the number of offspring to generate (N, if population size is constant), as each offspring needs two gametes from different parents N

5xi = 2N i=1

(iii) half of the gametes must come from males and half from females Nm

5 xi = N

i=1

assuming males are in the first Nm positions and females in the following Nƒ (N = Nm + Nƒ).

It is obvious that restrictions in the maximum number of offspring per individual are straight– forwardly applied by giving an upper bound to variable x. Moreover, different limits can be given to males and females, if the species’ characteristics point in that direction 0 [ xi [ l m 0 [ xi [ l ƒ

Mate selection procedure To account for restrictions related to mating characteristics, selection and mating design must be arranged simultaneously. The present paper proposes a procedure based on the minimisation of the following combined function i=1

N

N

Nm

Nm

N

l=Nm+1

k=1

i=1

j=Nm+1

Nm

N

i=1

j=1+N m

i = 1,...,Nm j = Nm+1,...,N

5 5 xij = N

N

where lm and lƒ are the maximum possible number of offspring generated by a male and a female, respectively.

Once the optimum contributions per individual are found, those from males and females have to be adjusted in order to determine the exact mating scheme. A linear programming optimisation allows to find the assignation design of male and female contributions yielding the minimum coancestry matings (for details see, e.g., FERNÁNDEZ & CABALLERO, 2001). This mating arrangement, however, can be incompatible with the particular reproductive restrictions of the species or population (see examples below).

N

xij m 0 xij integer

Additional constraints can control the maximum number of offspring per male or female,

i = 1,...,Nm i = Nm+1,...,N

{5 5 [(5 xil)(5 xkj)ƒij]}+ ? {5 5 xijƒij} j=1

those with random contributions from parents, with minimum coancestry matings afterwards.

Because the objective is to apply minimum global coancestry contributions and, only when two solutions have the same ƒ, apply coancestries between couples as a criterion, the optimum should be to use a very small value of ?. In this way, the minimum coancestry mating will be obtained but the individual contributions that yield the minimum global coancestry will be maintained.

As in the two–step procedure, some constraints must be added to find integer positive solutions and to fulfil the restrictions on the total number of offspring

(2)

where x ij is the number of offspring to be generated by the couple between male i and female j, and ? is a weighting factor. The number of variables x is equal to the number of all possible couples between males and females, i.e., Nm x Nƒ. The first term of the function represents the global coancestry (ƒ) as in formula (1), while the second term is the average coancestry between the members of the actual mating pairs. If a value of ? = 0 is given, the solutions obtained are those with minimum global coancestry contributions, as in the selection stage of the two–step method, and random mating of parents afterwards. On the contrary, if ? is very large, the solutions are

5 xij [ lm j=Nm+1 Nm

5 xij [ lf

i=1

i = 1,...,Nm j = Nm+1,...,N

the avoidance of full–sibs among the progeny,

xij [ 1

i = 1,...,Nm j = Nm+1,...,N

or the restriction of a single male mated to a particular female Nm

5 yij = 1 i=1

j = Nm+1,...,N

where yij is a dicotomic dummy variable with a value of one if the couple ij produces any offspring, and zero otherwise.

The problem is then reduced to the minimisation of a quadratic function with the corresponding restrictions. There are mathematical tools available that yield the exact solution, like the integer quadratic programming (MCCORMICK, 1983), but they are difficult to implement in computer simulations.

Some other approximated algorithms, like the genetic algorithms or the simulated annealing (PRESS et al., 1989), allow an easy and quite efficient implementation of optimisation processes into the simulations. In the present work optimisations were performed through the simulated annealing algorithm (further details on the implementation can be found in FERNÁNDEZ & TORO, 1999).

Computer simulations Simulations were performed for a dioecious species, where fitness is controlled by a large number of loci (5800) acting multiplicatively through viability differences among individuals. Mildly or moderately deleterious as well as lethal mutations arose every

21

Animal Biodiversity and Conservation 24.2 (2001)

generation at rates and effects according to estimates in the literature (CROW & SIMMONS, 1983; CABALLERO & KEIGTHLEY, 1994; LYNCH et al., 1999).

Neutral multiallelic loci were simulated in order to monitor changes in neutral genetic variation.

Individuals in the initial sample were assumed to be unrelated, so they carried different alleles at all these neutral loci in order to calculate probabilities of identity by descent and measures of genetic diversity (gene diversity and allelic diversity). A more detailed description of the model and the parameters used can be found in FERNÁNDEZ & CABALLERO (2001).

Management procedures From a large population with frequencies at mutation–selection–drift equilibrium, samples of 8, 24 or 48 individuals were randomly taken. Prior to the implementation of any conservation strategy the population underwent five unmanaged generations in order to generate a complex pedigree and differential coancestries between individuals.

From that point (generation 0), two different schemes were performed for 15 generations and the mean population fitness and diversity measures were calculated each generation, and averaged over 100 replicates:

Table 1. Gene diversity, allelic diversity (averaged over 200 neutral loci), mean population fitness (scaled to that in generation zero) and average inbreeding coefficient at generation 15. All values presented in percentage: N. Population size; ?. Weight given to mating criterium.

Tabla 1. Diversidad génica, diversidad alélica (promediadas para 200 loci neutros), eficacia biológica media de la población (relativa a la de la generación cero) y coeficiente de consanguinidad promedio en la generación 15. Todos los valores aparecen en porcentaje: N. Censo de la población; ?. Ponderación asociada al criterio de apareamineto.

One–step Two–step ?=1 ? =.01 ?=.001

N=8 Gene diversity

13.3 13.7

13.8

72.0

69.5 72.6

71.2

Inbreeding

57.3

58.8 57.8

57.3

One step Contributions and mating design were chosen minimising the joint function (2). Values for ? were ranged from 0.0001 to 1. Two runs were performed restricting to one the number of matings in which a female could be involved, and not allowing more than one offspring per couple.

N = 48

Table 1 shows the level of genetic diversity, measured as gene and allelic diversity, for the

37.0

Fitness

N = 24

Results and Discussion

35.5 36.8

Allelic diversity 13.8

Two step As described in FERNÁNDEZ & CABALLERO (2001), in this method the contribution of every available parent was decided minimising the global coancestry of the population (function [1]).

Minimum coancestry matings were then arranged.

Global and pairwise coancestry were calculated from pedigree records. In both methods, descendants of each couple were evaluated for fitness, calculated as the product of the individual effects of the 5800 loci in each genotype. A random number from 0 to 1 was drawn for each offspring and compared to its viability. If this was lower than the random number, the descendant died and another offspring from the same couple was generated. Population size was constant over generations with equal numbers of males and females. Sex of offspring was assigned at random once all descendants had been obtained.

36.8

Gene diversity

72.6

71.7 72.5

72.7

Allelic diversity 12.3

11.8 12.2

12.3

Fitness

83.7

82.6 83.3

82.5

Inbreeding

23.0

25.2 24.0

23.9

85.2

84.7 85.1

85.3

Allelic diversity 11.8

11.5 11.8

11.8

Fitness

86.5

85.8 86.6

86.3

Inbreeding

12.0

12.6 12.9

13.2

Gene diversity

optimum two–step method and the single–step method using different values of ?.

The amount of genetic diversity preserved was quite similar for both methods, irrespective of the measure of diversity we used, although it was slightly lower for larger values of ?. This behaviour occurs because of the influence of mating criterium on the selection of parents if the weight given to pairwise coancestry is too high. As pointed out by FERNÁNDEZ & CABALLERO (2001), when performing a single–step method, being more strict in the level of coancestry between couples can lead to the use of fewer individuals or those with higher mean coancestry. In this case, the genetic diversity preserved in the population would be smaller and its fitness would suffer from a fall due to inbreeding

Fernández et al.

22

depression. Thus, in the same table, can be seen that mean population fitness is similar for both methods with ? [ 0.01, but slightly lower for large ?, where the inbreeding level is somewhat higher.

The above results suggest that the proposed method is as efficient as the optimal two–step strategy to manage populations under conservation, regarding both the amount of genetic diversity preserved and the fitness of the population.

When restrictions are included, the space of feasible solutions is reduced. But even in this constrained situation the one–step method looks for the solution with the lowest group coancestry and, afterwards, for the mating scheme yielding the lowest pairwise coancestry. Table 2 shows the optimal contributions (table 2A) and the optimal mating design (table 2B), for a particular group of individuals (N = 8), for the unrestricted situation and two restricted cases. The unrestricted situation has the same solution for the one–step and the two–step methods. In the first restricted case (one male, SM), females are allowed to mate to a single male. In the second (NFS), there is a maximum of one offspring per couple, although individuals can be involved in different couples.

This latter restriction implies that no full–sibs are to be found among the progeny. Some authors have suggested this strategy as a way to slow the increase of inbreeding in a population (WANG, 1997; SONESSON & MEUWISSEN, 2000).

In this particular population structure, if we implement minimisation of group coancestry alone to determine the optimal contributions (two–step method), the result is not compatible with any of the restrictions (table 2). Female number 7 should generate six offspring, but there is no male with such a high contribution (impossible to fulfil restriction one male, SM), and there are not six males either (some full–sibs will be created). With the one–step method, restrictions are taken into account when looking for the contributions and, therefore, there are compatible mating schemes (table 2). The group coancestry and coancestry between couples of the unrestricted solution are 0.368 and 0.246, respectively. Group coancestry barely changes to 0.372 and 0.374 for one male (SM) and NFS, respectively. The increase in pairwise coancestry is somehow larger (0.288 and 0.305), as expected, but differences are small.

The use of high values of ? (> 1) leads to worse results (higher inbreeding coefficients), as could be expected for the influence of mating coancestry in the selection of parents’ contributions explained above. Eventually, the average fitness of the population in the very first generations can be slightly higher than that with low values of ?, because the avoidance of a high coancestry between the couples leads to the decrease of inbreeding depression in the offspring. But, over a longer period, both genetic diversity and fitness fall below the levels of the two–step method.

Interestingly, using very small values of the

Table 2. A. Example of optimal contributions of each available individual when no restriction is imposed (UR), when each female can mate with a single male (SM), and when full–sibs are avoided in the offspring (NFS); B. Optimum mating design (male–female), according to contributions obtained in A. Tabla 2. A. Ejemplo de contribuciones óptimas de cada uno de los individuos disponibles cuando no se imponen restricciones (UR), cuando cada hembra puede aparearse con un sólo macho (SM) y cuando se evita la aparición de hermanos en la descendencia (NFS); B. Esquema de apareamientos óptimo (macho– hembra), conforme a las contribuciones obtenidas en A.

A. Parents UR

SM

NFS

Males 1

3

5

3

2

1

1

1

3

1

0

1

4

3

2

3

5

1

1

1

6

1

2

2

7

6

5

4

8

0

0

1

UR

SM

NFS

1

1–7

1–7

1–6

2

1–7

1–7

1–7

3

1–7

1–7

1–8

4

2–7

1–7

2–7

5

3–7

1–7

3–7

6

4–5

2–5

4–5

7

4–6

4–6

4–6

8

4–7

4–6

4–7

Females

B. Offspring

weighting factor for the mating coancestry also produces poorer results (data not shown), contrary to what theory predicts. The reason is the use of an algorithm of random search to perform the optimisations. As the value of ? diminishes, differences between solutions do so, and it is

Animal Biodiversity and Conservation 24.2 (2001)

more difficult for the process to find the global optimum, specially for large populations. As previously pointed out, exact methods exist that can be implemented if only a round of optimisation is necessary, as in the management of a real population for conservation. In this case, therefore, the relative value of ? should be adjusted to the smallest number distinguished by the precision of the computer.

For the population sizes considered, there are no great differences in computing time between procedures, although it is larger for the mate selection method. In principle, this latter implies the optimisation of a function with Nm x Nf variables, while the two–step method needs two optimisation processes with N and N x N variables, respectively. However, the one–step method has a greater feasible space (more time required to find the optimum), and the mating step is just an assignation problem in the two–step method (MCCORMICK, 1983). These factors make the two– step method less demanding.

Traditionally, the way to cope with the issue of physiological restrictions has been to use populations structured in families, with fixed numbers of selected males and females.

Contributions of selected individuals would be equalised, so that it would be straightforward to find a mating design which fitted the restrictions.

However, as several authors have pointed out (MEUWISSEN, 1997; GRUNDY et al., 1998; FERNÁNDEZ & T ORO , 1999), allowing for differential contributions gives a better control of the increase of inbreeding and the loss of genetic information. The method presented in this paper, following this second strategy, is more flexible, has a larger feasible space of solutions and achieves better levels of genetic diversity.

Acknowledgements We thank X. Domingo–Roura and A. Ruíz for helpful comments on the manuscript. This work was supported by grant BOS2000–0896 (Ministerio de Ciencia y Tecnología from Spain), PGIDT01PXI30104PN (Xunta de Galicia) and 64102C124 (Universidade de Vigo).

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Animal Biodiversity and Conservation 24.2 (2001)

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Declining amphibian populations: a global phenomenon in conservation biology T. Gardner

Gardner, T., 2001. Declining amphibian populations: a global phenomenon in conservation biology. Animal Biodiversity and Conservation, 24.2: 25–44.

Abstract Declining amphibian populations: a global phenomenon in conservation biology.— The majority of the recent reductions in the Earth’s biodiversity can be attributed to direct human impacts on the environment. An increasing number of studies over the last decade have reported declines in amphibian populations in areas of pristine habitat. Such reports suggest the role of indirect factors and a global effect of human activities on natural systems. Declines in amphibian populations bear significant implications for the functioning of many terrestrial ecosystems, and may signify important implications for human welfare. A wide range of candidates have been proposed to explain amphibian population declines. However, it seems likely that the relevance of each factor is dependent upon the habitat type and species in question, and that complex synergistic effects between a number of environmental factors is of critical importance. Monitoring of amphibian populations to assess the extent and cause of declines is confounded by a number of ecological and methodological limitations.

Key words: Declining amphibians, Environmental degradation, Indirect human impact, Population monitoring.

Resumen Disminución de las poblaciones de anfibios: un fenómeno global en biología de la conservación.— La mayoría de reducciones recientes en la biodiversidad de la Tierra puede atribuirse al impacto humano sobre el ambiente.

Durante la última década, es cada vez mayor el número de estudios que informan de disminuciones en las poblaciones de anfibios en hábitats inalterados. Dichos estudios sugieren el papel de factores indirectos y un efecto global de las actividades humanas sobre los sistemas naturales. Las disminuciones de las poblaciones de anfibios llevan consigo implicaciones significativas para el funcionamiento de algunos ecosistemas terrestres y pueden tener importantes repercusiones en el bienestar humano. Para explicar la disminución de las poblaciones de anfibios se ha propuesto una amplia gama de posibles factores causales. Sin embargo, parece ser que la relevancia de cada factor depende del tipo de hábitat y de la especie afectada, y que los complejos efectos sinérgicos entre algunos factores ambientales es de importancia crítica. El control de las poblaciones de anfibios con objeto de valorar la dimensión y causa de la disminución está condicionado por una serie de limitaciones ecológicas y metodológicas.

Palabras clave: Disminución de anfibios, Degradación ambiental, Impacto humano indirecto, Control de poblaciones.

(Received: 28 XII 01; Conditional acceptance: 25 III 02; Final acceptance: 10 IV 02) Toby Gardner, Centre for Ecology, Evolution and Conservation, Dept. of Biological Sciences, Univ. of East Anglia, Norwich, NR4 7TJ, United Kingdom.

E–mail: t.gardner@uea.ac.uk

ISSN: 1578–665X

© 2001 Museu de Zoologia

26

Introduction Human alteration of the global environment namely through habitat modification, agricultural practices, anthropogenically induced climate change, and atmospheric pollutants has triggered what is widely regarded as the sixth major extinction event in the history of life (CHAPIN et al., 2000). The extent of loss of biological diversity, and alterations in the distribution of organisms shows considerable variance both with respect to geographic location but also with respect to the ecological and taxonomic characteristics of the species involved. Usually proximal factors such as habitat destruction or modification are easily identified as the responsible cause of local losses of biodiversity, and as such local reductions in biodiversity are most frequently observed across the taxonomic spectrum. Like most terrestrial species amphibians are threatened foremost by habitat destruction (SALA et al., 2000). However, in the past few decades amphibian populations have been threatened by other incompletely understood factors in areas that are perceived to be intact from human disturbance (WALDMAN & TOCHER, 1998; ALFORD & RICHARDS, 1999; CAREY et al., 2001). The suspicion that amphibians are suffering from an unprecedented and abnormally high rate of decline even in protected areas was first voiced at the First World Herpetology Conference in September of 1989, in Canterbury, England (BARINAGA, 1990), although it is clear that widespread concern existed long before this (BURY, 1999). The seriousness with which the scientific community recognised this problem was reflected in the rapid organisation of a NRC sponsored workshop in February of 1990 in Irvine USA, and in light of the perturbing reports presented (B LAUSTEIN & W AKE , 1990; W AKE , 1991), the establishment of a special task force on declining amphibian populations (DAPTF), allied with the Species Survival Commission of the IUCN. During the subsequent decade neither the scale of the problem nor the widespread concern expressed in both the scientific (WAKE, 1998), and public (BLAUSTEIN & WAKE, 1995; MORELL, 2001) community has seen any abatement. In order to tackle any problem in ecology it is essential that one is aware of the present level of understanding of its scale, diagnostic characteristics, and methodologies appropriate to its resolution. This review complements previous similar efforts (e.g.

WALDMAN & TOCHER, 1998; ALFORD & RICHARDS, 1999) by exploring many important advances in the last two years. Edward Wilson recently described conservation biology as the "intensive care ward of ecology" (W ILSON , 2000), and as such a conservation biologist who lacks an up–to–date appreciation of their field is failing the prescription of this definition in inadvertently advocating inefficient, repetitive, or even counterproductive research. This review seeks to provide such a revision, dealing in turn with: 1. The ecological

Gardner

and human importance of amphibians in natural ecosystems; 2. Evidence for population declines and caveats in their interpretation; 3. The range of candidates which have been proposed to explain such declines; 4. Some challenges presently facing conservation biologists in resolving and preventing amphibian declines.

The importance of amphibians in ecological and human environments A world–wide decline of amphibian populations could have a significant and detrimental impact on both natural ecosystems and human welfare.

Amphibians are integral components of many ecosystems, often constituting the highest fraction of vertebrate biomass (BURTON & LIKENS, 1975; BEEBEE, 1996). Their conspicuous role is noted to be of particular importance in tropical forests, where in acting as both predator and prey species, they play a key role in trophic dynamics (TOFT, 1980; BLAUSTEIN et al., 1994c). Their high collective biomass, alongside their high digestion and production efficiencies (WOOLBRIGHT, 1991), go someway to explaining their potential importance in such "functions" as the maintenance ecosystem energetics and carbon flow (PEARMAN, 1997) —namely through the maintenance of arthropod abundance (GUYER, 1990), and the provision of a critical prey base for higher order predators, such as arachnids, snakes, and birds (GUYER, 1990; WOOLBRIGHT, 1991; DUELLMAN & TRUEB, 1994). In identifying the functional significance of amphibians its is clearly of relevance to understand whether species diversity per se plays a unique role over and above species identity —i.e. are a few specific and perhaps more abundant frog species sufficient to maintain the natural integrity and productivity of the ecosystem? In view of the limitations on the world’s resources for the conservation of biodiversity, it would seem sensible to identify the functionally important amphibian species or "guilds" in order to prioritise concern and subsequent potential conservation action following a reported decline in number. However, such an approach could be very dangerous, as it is often extremely difficult if not impossible to identify the functional role or contribution of many species (CHAPIN et al., 2000). Some (limited) empirical evidence exists to offer an explicit justification for the functional importance of species richness per se (LOREAU et al., 2001).

However, a more convincing argument is that differences in the environmental tolerances of many species that may be functionally analogous to dominant species can provide critical insurance or resilience for the system in the face of climate change or altered disturbance patterns (WALKER, 1995; WALKER et al., 1999; NAEEM, 1998). In light of our ignorance of the ecology of most amphibians, and the growing domination of natural systems by human activities (VITOUSEK et al., 1997), it seems that an attitude of concern

Animal Biodiversity and Conservation 24.2 (2001)

towards all populations that could be potentially at risk is prudent, if not essential. It is important to stress that concerns about the functional importance of species in no way detracts from the importance of other values humans can attach to biodiversity —including cultural, existence and intrinsic values (DOLMAN, 2000).

The second most recognised importance of amphibians is their potential role as indicators of global environmental health and resilience (BLAUSTEIN & WAKE, 1990; BARINAGA, 1990; DIAMOND, 1996). They inhabit both aquatic and terrestrial habitats, and are thus exposed to aquatic and terrestrial pollutants —to which they are particularly sensitive due to their highly permeable skin (DUELLMAN & TRUEB, 1994). Furthermore many amphibians interact with a large range of other species in the local environment during their lifetime. For example most anurans (Amphibia, Anura) play dual roles as both herbivores during larval stages and carnivores as adults, making them potentially good indicators of changes in both floristic and faunal community composition —possibly induced through environmental stress.

As BARINAGA (1990) states, the fact that amphibians as a group have remained largely unchanged since the era of the dinosaurs, highlights the potentially disastrous consequences for humans and other species if their suspected demise continues unabated. Finally, from a purely anthropocentric perspective amphibians represent a storehouse of pharmaceutical products waiting to be exploited fully (BLAUSTEIN & WAKE, 1995). Some compounds already extracted are presently being used as painkillers and in the treatment of traumas such as burns and heart attacks, whilst many more undoubtedly await discovery.

Evidence for global amphibian declines Although serious recognition of the potential problem of declining amphibians was not afforded until the last decade, individual anecdotal reports of population declines have been known since the late 18th century (BURY, 1999). However, it was during the 1980’s and early 1990’s that the observations of more dramatic and scientifically credible declines were made. Such declines include notable individual examples such as the Golden toad ( Bufo periglenes) and Harlequin frog (Atelopus varius) (CRUMP et al., 1992; POUNDS & CRUMP, 1994), the Cascades frog (Rana cascadae) (FELLERS & DROST, 1993), and the Yellow and Red–legged frogs (Rana muscosa and Rana aurora) (BLAUSTEIN & WAKE, 1990). All of these declines have occurred in areas considered largely intact from human interference, which explains their common citation in justifying concern for the viability of other seemingly well protected amphibian populations. Further evidence for the apparent vulnerability of the class Amphibia as a whole

27

comes from reports of population declines across whole communities of amphibians at the regional level, also in relatively pristine areas; the Central Valley of California (DROST & FELLERS, 1996; FISHER & SCHAFFER, 1996), the montane forests of Eastern Australia (L AURANCE et al., 1996), and the Monteverde cloud forest of Costa Rica (LIPS, 1998, 1999). It is hard to draw global conclusions from such varied examples, although a number of commonly occurring factors or attributes can be identified (LIPS, 1998; ALFORD & RICHARDS, 1999; CAREY, 2000; CAREY et al., 2001; MIDDLETON et al., 2001). These factors include: 1. Wide geographic spread in presence of declines, accompanied by significant spatial variability in their extent.

2. Significant inter–specific variability in levels of vulnerability to agents of population decline, with many species that are sympatric to others which are threatened or endangered exhibiting no change in population size or dynamics.

3. Many species extinctions or extirpations have occurred at high altitude sites (> 500 m a.s.l.).

4. Many declines have been rapid with population reductions of between 50 and 100% occurring in 1–3 years.

5. Infectious diseases, commonly fungal pathogens have been most frequently identified as the direct cause of decline, whilst a number of indirect environmental factors are thought to play key contributing roles.

Due to the heavily skewed distribution of amphibian biologists towards North America, Europe, and Australia it is possible that a number of these common attributes are at least partly artefacts of research bias. To remove some of this bias and view the declining amphibian problem from a more global perspective, it is necessary to collate information from across many sites and many species.

In attempting to draw global conclusions or patterns about a particular ecological phenomenon (such as population declines) from across different studies, one is commonly faced with two main problems; the inaccessibility of many research reports, and the extreme variability in monitoring techniques used —from the purely anecdotal to the scientifically rigorous. A number of recent reviews have attempted such a difficult collaboration in order to view the plight of amphibians from the widest possible perspective —both at spatial and temporal scales, the two most noteworthy of which are those of ALFORD & RICHARDS (1999) and HOULAHAN et al. (2000). ALFORD & RICHARDS (1999) considered 85 time series of amphibian populations spanning the period between 1951–1997, and following regression analysis concluded that more populations correlated negatively against time than would be expected under their null hypotheses of "normal" population fluctuations, with 67% of relationships being negative. However, they found no evidence that the proportion of populations decreasing

28

changed over time —in other words there was no observation of an increase in the number of susceptible and affected populations which is perhaps what one would expect if the proposed agents of decline where becoming more prevalent or intense. Although continued exposure to stimulants of population declines may produce a residual number of populations and species which show heightened resilience, or adaptive shifts in geographic range to habitat refugia, it seems unlikely that such evolutionary or behavioural changes could occur at comparable speeds to many of the proposed agents of decline which are detailed above. As the authors themselves admit, it is impossible to draw firm conclusions about the global status of amphibian populations due to variance in the size of data sets, their methodological origin, and the inter–specific variance in population dynamics which renders their null model far from optimal for all amphibian species. HOULAHAN et al. (2000) made by far the most exhaustive attempt to date in collating data from 936 populations of 157 species from 6 continents, for studies of between 2 and 31 years duration. Although their results identify marked temporal variation in the speed of the decline, and spatial variation as to its extent, a definite negative relationship is clearly evident, adding perhaps the first real quantitative "weight" to the declining amphibian phenomenon. Criticism has recently been raised as to the validity of the statistical averaging methods used by Houlahan and colleagues (ALFORD et al., 2001; but see HOULAHAN et al., 2001), although re–analysis under the alterative methodology (ALFORD et al., 2001) still concluded that an overall population decline existed and disagreed only in the shape of the relationship —with the more recent interpretation identifying a increase in the rate of declines in the last decade. A number of serious inade–quacies exist in the study by HOULAHAN et al. (2000) —for example it includes only four studies from Latin America, despite the fact that this continent hosts about half the worlds amphibian species (DUELLMAN, 1999). However, a recent synthesis of published and unpublished (> 95% of the total) work from both Central and South America adds strength to the evidence for a global decline (YOUNG et al., 2001). In summarising 118 monitoring projects, population declines were found to be widespread, occurring in 13 countries, with 40 cases of recent extinction or regional extirpation affecting 30 genera and nine families of amphibians.

Candidates for amphibian decline Physical habitat modification The destruction or direct modification of ecological systems is widely held as the primary cause for the observed loss of much of the earth’s biological diversity (SALA et al., 2000), and the loss of

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amphibian species together with reductions in their population size bear no exception (ALFORD & RICHARDS, 1999). In an area directly under human influence, habitat modification can usually be isolated as the cause of an observed population decline, although the actual mechanism is highly variable and can often be obscure. In addition to complete habitat destruction, a number of more subtle environmental modifications can bear particular consequences for amphibians: 1. Fragmentation of habitat. This can have two main deleterious effects. Firstly in the effect on population demographics through the distribution of regional and metapopulation processes (SJOGREN, 1991; SJÖGREN–GULVE, 1994; MARSH & TRENHAM, 2000). Both empirical (SJOGREN, 1991) and theoretical (HALLEY et al., 1996) evidence suggests that the probability of local population extinction increases with increased distance between populations —largely due to the fact that many amphibian species are thought to be highly philopatric (SJOGREN, 1991; WALDMAN & TOCHER, 1998; SCRIBNER et al., 2001). Secondly, the disruption of dispersal mechanisms can produce deleterious effects at the level of genes (e.g.

HITCHINGS & BEEBEE, 1998; SEPPA & LAURILA, 1999).

The genetic consequences of small and declining populations has been adequately reviewed elsewhere (e.g. F RANKHAM , 1995; H EDRICK & KALINOWSKI, 2000), although with relevance to amphibians a recent study has identified a possible relationship between reduced genetic diversity in Southern Leopard frogs (Rana sphenocephala) following restricted migration, and tolerance to insecticide, with possible implications for recent population declines in the western United States (BRIDGES & SEMLITSCH, 2001). In developed countries the deleterious effects of habitat fragmentation on amphibian populations is increasingly apparent with the increase in the number of roads (HITCHINGS & BEEBEE, 1998), a type of habitat modification which has also recently been acknowledged to contribute significantly to population declines through direct mortality (HELS & BUCHWALD, 2001).

2. Forest management operations which can result in a change of microclimate, soil moisture and habitat complexity. Of particular importance is land drainage for reservoirs and other developments, frequently resulting in a removal of breeding sites and fragmentation of populations.

3. The alteration of the biotic environment through the introduction of exotic predators and pathogens (see below).

More obscure and perhaps counter–intuitive examples of the deleterious impacts of habitat modification exist. For example, in the case of the Natterjack toad (Bufo calamita) in Britain, where the removal of modification (grazing) on shrub heathland led to the encroachment of tall vegetation, thus allowing the entrance of the more successful competitor Bufo bufo —the Common toad (BEEBEE, 1977). Such examples serve to emphasise the fragility of many ecological

Animal Biodiversity and Conservation 24.2 (2001)

systems to what we may perceive to be minimal human intervention. However, habitat destruction and modification although of prime concern, are usually easily to isolate, and therefore if possible to rectify. It is the proposal that amphibian declines in largely pristine areas of the world are the result of more indirect and complex reasons that is cause for exceptional concern (WAKE, 1998; WALDMAN & TOCHER, 1998; CAREY, 2000; CAREY et al., 2001).

Ultraviolet radiation Depletion of the stratospheric ozone layer and the observed resultant increases in ultraviolet B (UV–B) radiation at the Earth’s surface (KERR & MCELROY, 1993), has prompted interest as to the possible relationship between the influence of UV–B on amphibian survival and population declines. A number of experimental manipulations of enhanced UV–B have implicated its potential contribution to amphibian declines —e.g. through evidence of; decreased hatching success and enhanced embryonic mortality (BLAUSTEIN et al., 1994a; OVASKA et al., 1997; ANZALONE et al., 1998), decreased larval survival (OVASKA et al., 1997), and negative effects on embryo and larval development (CRUMP et al., 1999). However, all of these studies report significant variation between species as to both the level and type (i.e. embryo, larvae, etc.) of susceptibility. Some resolution of this discrepancy has been proposed through inter– specific variation in the levels of the DNA repair enzyme, photolyase (B LAUSTEIN et al., 1994a; BLAUSTEIN et al., 1996). Indeed a correlation can be made between a number of species whose populations are showing a decline in number (e.g. Bufo boreas and Rana cascade) and which also show significantly low levels of photolyase activity. This can be compared against species such as the Pacific treefrog Hyla regilla which has characteristically high levels of the enzyme and exhibits relative stability in number (BLAUSTEIN et al., 1994a). However, this relationship is clearly not of global relevance as the Red–legged frog Rana aurora , has a relatively high level of photolyase but yet has suffered severe depletions in number (BLAUSTEIN & WAKE, 1990; BLAUSTEIN et al., 1996). A further quite equivocal result is seen in the declining Australian species, the Green and Golden bellfrog Litoria aurea, which although has a lower photolyase activity than two sympatric and non–declining species, the Bleating treefrog L. dentata, and Peron’s treefrog L. peroni, shows no significant difference against them with respect to hatching success under enhanced UV–B exposure (VAN DER MORTEL et al., 1998) —thus pointing to the importance of other, independent agents of decline. Recent work by PAHKALA et al.

(2001) suggests that there may be time–lags in the response of amphibians to UV–B radiation, and that whilst evidence of direct effects of enhanced radiation on early embryonic stages is

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rarely convincing, carry–over effects on later larval development and metamorphosis can be very important.

A number of other factors serve to shed doubt on the responsibility of UV–B in global declines. Most of the field experimental studies cited above have been conducted in shallow, clear high altitude ponds, largely in high latitude locations such as North America and South–eastern Australia. UV–B radiation is largely absorbed in the first few centimetres of the water column (NAGLE & H OFER, 1997; A D A M S et al., 2001), and the depth of penetration is negatively correlated against the dissolved organic carbon content (CRUMP et al., 1999) —factors which suggest that UV–B radiation is unlikely to be a problem in bottom laying species, or in forest (and especially tropical) species. An initial lack of evidence for significant increases in UV–B at tropical or sub–tropical latitudes since the mid–1970’s (MADRONICH & G RUJII, 1993) further diminished its perceived importance in the decline of tropical amphibian populations (e.g. CRUMP et al., 1992; LIPS, 1998). However, recent remote sensing analysis (M IDDLETON et al., 2001) has identified increases in both annual and daily levels of UV–B exposure (average and maximum) between 1978–1998 at Central and South American sites where amphibian population declines have been recorded (e.g. P OUNDS & CRUMP , 1994; LIPS , 1998, 1999; P OUNDS et al., 1999). Further recent work by A DAMS et al.

(2001) provides more correlative evidence for the importance of UV–B radiation in determining amphibian distribution, in identifying the importance of levels of UV–B exposure in determining the spatial pattern of R. cascadae breeding sites in Olympic National Park, USA. Although such studies report only correlative rather then causative evidence, they identify the value and urgency for further field studies on the effects of UV–B radiation on amphibian populations.

Although some of the above evidence is convincing in showing an effect of high UV–B radiation on embryo mortality and larval survival, the ecological significance of such a phenomenon at the population level is far from clear, and equally difficult to assess (ALFORD & RICHARDS, 1999). For example there may be density dependent compensation effects, through the enhanced fitness of competing individuals that survive high levels of UV–B exposure. The potential indirect effects of enhanced UV–B on amphibian dynamics, such as changes in water chemistry and food supplies, are even less well known (ALFORD & RICHARDS, 1999). Finally it is likely that unimodal experiments manipulating only levels of UV–B are inadequate, and that the crucial agent of decline could be in the interaction of UV–B with other key environmental stresses (see below).

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Acidification and other chemical pollutants The existence of amphibian extinctions and population declines in what are otherwise seen to be some of the most pristine environments on Earth has led to the frequent suggestion that atmospheric pollutants may act as indirect agents of decline (LIPS, 1998; CAREY et al., 2001). Such pollutants could originate from neighbouring and foreign agricultural depositions, as well as from factory emissions of industrialised nations, and are able to travel vast distances and persist for considerable periods of time. Recent analysis of remote sensing data in Puerto Rico has shown spatial correlations between urban and agricultural pollutants and amphibian population declines (STALLARD, 2001).

One of the most acknowledged remote impacts of human activity is increased acidity of rainfall, a phenomenon of great potential importance in light of the importance of the annual water regime to amphibians. Increased acidity of ground and pond water is suspected to have both lethal and sub–lethal effects on amphibian populations through a number of factors; enhanced embryo and larval mortality, reduced egg and larval growth, reduced reproductive output, delayed hatching times, reduced adult body size, alterations in geographic distribution, and alterations in predator–prey ratios through indirect effects on plant growth and pH sensitive competitors and predators (FREDA & DUNSON, 1986; WALDMAN & TOCHER, 1998; ALFORD & RICHARDS, 1999). Some field and laboratory work has provided evidence for such detrimental effects, for example; reduced ion exchange and larval growth in the Wood frog Rana sylvatica (FREDA & DUNSON, 1986), and a significant reduction in range size of Natterjack toad Bufo calamita following long term acidification of many British ponds (BEEBEE et al., 1990).

Observational and experimental evidence also exists for the potential role of a wide range of industrial and agricultural pollutants in precipitating amphibian population declines. Contamination from a number of major agricultural pollutants (pesticides, herbicides and fertilisers) has been correlated with observed spatial patterns of decline in a number of amphibian species (RUSSELL et al., 1995; SPARLING et al., 2001; STALLARD, 2001), with early embryonic stages being particularly vulnerable (CAREY & BRYANT, 1995). This correlative evidence for the negative impact of agricultural practices on many amphibian populations is strongly supported by a number of experimental studies.

Negative effects of nitrate fertiliser including ammonium nitrate, one of the most commonly applied chemicals, have been observed on the larval mortality and development, feeding behaviour, growth rates and physical abnormalities of a number of amphibian species —including many pond and stream breeders (HECNAR, 1995; OLDHAM et al., 1997; MARCO et al., 1999; MARCO & BLAUSTEIN, 1999), and treefrogs and forest–dwelling

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species (SCHUYTEMA & NEBEKER, 1999; MARCO et al., 2001). Responses have been observed to be both dose–dependent and cumulative over time (MARCO et al., 1999; MARCO et al., 2001), and although there is clearly significant interspecifc variation in patterns of susceptibility (HECNAR, 1995; MARCO et al., 1999; MARCO et al., 2001), levels of fertiliser application observed to be sufficient to cause significant negative effects on individual survival and fitness are frequently no higher than officially recommend levels for field application (HECNAR, 1995; OLDHAM et al., 1997) or even for drinking water (MARCO et al., 1999).

In addition to the effects of fertiliser a number of other chemical pollutants have been identified as being of potential importance in explaining observed patterns of amphibian population decline. In the Sierra Mountains of California, a region exhibiting a high level of amphibian population declines across several species during the last 10–15 years, correlative evidence suggests the importance of pesticide contamination from the heavily agricultural downwind San Joaquin Valley (SPARLING et al., 2001). Furthermore a recent experimental study identified negative effects of ambient concentrations of atrazine —a common pesticide— on the length and weight of H. versicolor larvae at metamorphosis (DIANA et al., 2000), although another more recent experimental study also concerned with testing the effects of atrazine reported more equivocal results for other species, with no observed effect on either hatching success or post hatching larval morality (ALLRAN & KARASOV, 2001). The fact that such studies rarely consider longer–term or secondary effects, or even in this case report analogous measures of fitness and reproductive success, makes it difficult to make generic conclusions of the overall significance of such contaminants at the population level. Finally a number of other non–agricultural chemical pollutants have been isolated as being of potential importance in explaining population declines, including, namley: 1. Negative effects of endocrine disrupting chemicals on reproductive success, and larval development (FOX, 2001), and 2. An increasing number of experimental studies reporting negative effects of polychlorinated biphenyls (PCBs) on the larval development and feeding rates for a number of species (GUTLEB et al., 2000; GLENNEMEIER & DENVER, 2001), although the magnitude of the effect depends critically on the length of the observation period (GUTLEB et al., 1999). Alongside cases of direct mortality (CAREY & BRYANT, 1995), empirical evidence has also identified important indirect influences of agricultural pollution on populations, e.g. through altered recruitment and predator response behaviours (COOKE, 1971; BRIDGES & S EMLITSCH , 2000). The marked interspecific variation of amphibians in their susceptibility to pesticides alongside geographic variation in their deposition is highlighted as a potential

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Animal Biodiversity and Conservation 24.2 (2001)

explanation for the heterogeneous nature of observed declines at both taxonomic and geographic scales (BRIDGES & SEMLITSCH, 2000).

Although it is clear from the above examples that an increase in environmental acidity or other pollutants can have a negative effect on some amphibian populations, the actual physiological mechanisms remain unclear (CAREY et al., 1999).

One recent experimental study provides evidence that a commonly used insecticide, endosulfan, causes impairment of the pheromonal system in Red-spotted newts (Notophthalmus viridescens) at very low exposure–concentrations (PARK et al., 2001). This provides one potential mechanism which could help explain reduced mating success —through the disruption of mate choice.

Furthermore, and perhaps most crucially, it is once again not evident what the overall consequences of such environmental changes at the population level would be, and there is very little data implicating contaminants on the recent catastrophic population declines (A LFORD & RICHARDS, 1999). However, as in the case for enhanced UV–B levels, it is possible that the critical role of environmental contamination in population declines comes from its interaction with other causative agents (CAREY et al., 2001; STALLARD, 2001; see below).

Predation by exotics or introduced species Biotic interactions amongst and between species can play a critical role in determining their relative local abundance, distribution and population dynamics (RICKELFS & SCHLUTER, 1993; HUSTON, 1994).

Although perhaps more appropriately considered as human induced habitat modification, the introduction of exotic predators to amphibian environments has been implicated as the factor responsible for many population declines, including the collapse of whole communities (FELLERS & DROST, 1993; FISHER & SCHAFFER, 1996; HECNAR & M’CLOSKEY, 1996a). Two recent studies have analysed the spatial distribution and abundance of amphibian species against that of introduced fish stocks in mountain and alpine lakes at the landscape scale. KNAPP et al. (2001) found that the Yellow–legged frog (Rana mucosa) exhibited dramatic reductions in both distribution and abundance in lakes which had received artificial stocks of predatory fish when compared to those that remained naturally fishless. Also PILLIOD & PETERSON (2001) found lower abundance of both the Long–toed salamander (Ambystoma macrodactylum) and the Columbia Spotted frog (Rana luteiventris) in alpine lakes that had received artificial fish stocks, and predicted that the range restriction of amphibians to remnant shallow lakes unsuitable for fishing, in addition to severely inhibited migration patterns, could lead to the extirpation of amphibians from entire landscapes —including from sites that remained in a natural– fishless condition. Finally, although the majority

of studies reporting such clear negative effects of exotic predators on amphibian populations are from temperate regions, the phenomenon is also prevalent in the tropics —for example in South America where some 30% of the amphibians are classified by the IUCN as threatened by alien invaders (RODRIGUEZ, 2001).

Aside from such convincing but co–incidental evidence, experimental manipulations of predator and amphibian distributions provide firm support as to their devastating effect on amphibian populations.

Powerful examples include: 1. The significant reduction in survival of the endangered Red–legged frog (Rana aurora) in California, following the introduction of the two larval predators —Mosquitofish ( Gambusia affinis), and Bullfrogs (Rana catesbeiana) (LAWLER et al., 1999); 2. The severe impact of both Mosquitofish, and a crayfish (Procambarus clarki) on the eggs and larvae of the Californian newt, Taricha torosa (GAMRADT & KATS, 1996); and 3. Significantly enhanced predation pressure on Spotted treefrog larvae ( Litoria spenceri) from south–east Australia when exposed to two alien trout species —the Brown trout Salmo trutta, and the Rainbow trout Onchorhynchus mykiss— as opposed to when in the presence of the native mountain fish (GILLESPIE, 2001). Although the introduction of exotic predators such as the above is considered to be a prime cause of population decline across North America (FISHER & SCHAFFER, 1996), their role is comparatively easy to identify, and as such seems unlikely to be a global factor, especially in largely pristine tropical areas.

Disease The remote nature of many amphibian population declines, in addition to the frequent observations of larval and adult growth abnormalities, has led to the perhaps unsurprising and widespread implication of disease (CAREY, 1993, 2000). In particular the wave–like pattern of population decline across the range of many threatened species seem to implicate the role of a biotically induced agent —as observed in both the Atlantic forest of Brazil (HEYER et al., 1988), the Eastern montane forests of Australia (LAURANCE et al., 1996), and the forests of Panama and Costa Rica (LIPS, 1998, 1999). Perhaps the most confident proposition as to the culpability of disease in precipitating the collapse of an entire amphibian community is in Australia, where 14 endemic species have decreased by more than 90% in the last 15 years (LAURANCE et al., 1996). The authors note the extreme virulence of the disease as being evidence of its potentially exotic nature, and report histological changes in infected tissue of diseased individuals as being consistent with viral infection. LIPS (1998, 1999) identified a fungal infection found on dead individuals as being the most likely cause of population decline in the forests of Panama, between 1993 and 1997.

Furthermore she notes that the similarity in timing

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of declines, regional climatic factors, frog phylogenies, and clinical symptoms, as being suggestive of the same causal factor being responsible for catastrophic amphibian declines noted in nearby Costa Rica (POUNDS & CRUMP, 1994; LIPS, 1998). Support for this theory comes from BERGER et al. (1998) who identified the same chytridomycete (Chytridiomycota, Chytridiales) fungus on dead anurans from forests of both Central America, and also of Queensland, Australia —adding some further strength to the argument of LAURANCE et al. (1996), although the disease origin differs. The same fungus has been more recently identified to the species level (LONGORE et al., 1999), and an increasing number of reports have confirmed its presence in all of the six continents that are inhabited by amphibians (DASZAK et al., 1999; CAREY, 2000; FELLERS et al., 2001). A recent report of the expanding geographic distribution of this species, documents its arrival in Europe where it is implicated as being responsible for the disappearance of the Common Midwife toad (Alytes obstretricans) from more than 85% of its breeding sites in an a protected area in central Spain (BOSCH et al., 2001). The relatively sudden observation of catastrophic declines in such disparate areas of the world is suggestive of either a recent increase in virulence, or decrease in amphibian immuno–activity, perhaps due to a key interaction with a changing global climate —although the potential mechanisms behind any such interaction are poorly understood (CAREY, 2000; see below). In addition to a decrease in amphibian immuno–activity or an increase in pathogenic virulence, an increase in the level of pathogenic activity could be affecting amphibian populations through changes in the food supply or competitive ability of species (CAREY et al., 2001). Further convincing evidence as to the role of disease in population declines comes from the Pacific north– west of America, where a different species of fungus, Saprolegina ferax (a globally distributed fish pathogen), has been implicated as responsible for declines in the Boreal toad Bufo boreas, through increased egg mortality (BLAUSTEIN et al., 1994b; KIESECKER & BLAUSTEIN, 1997).

There is clearly enough convincing evidence to support the two facts that disease agents can be highly detrimental to amphibian fitness and survival, and that furthermore, evidence of them can be found in many areas where catastrophic declines have occurred. However, as for most if not all agents of decline, it is very difficult to attribute what contribution they make to the overall population dynamics of declining amphibians.

Strong circumstantial evidence exists as to the role of disease in mass declines such as that observed in Australia (LAURANCE et al., 1996). However, when experimental proof is difficult to obtain, it is easy to argue for competing hypotheses which may produce equally parsimonious statistical comparisons of a potential agent of decline against the spatial distribution of population declines

(ALFORD & RICHARDS, 1997; HERO & GILLESPIE, 1997).

Despite the ubiquitous presence of a large range of competing hypotheses to explain any one population decline, it is crucial to the progress of science that plausible hypotheses are voiced, if only for their heuristic value in targeting future research and formulating further, refined hypotheses (LAURANCE et al., 1997). As noted below, in the case of disease it is even more likely than in other agents of decline, that interactions of disease vectors with other environmental factors plays a crucial role in determining their impact on amphibian populations (CAREY, 2000).

Climate and weather As discussed already, amphibians are particularly sensitive to changes in their external environment, both due to their biphasic lifestyle in existing as both aquatic larvae and terrestrial adults, and due to their highly permeable skins. Perhaps the most important component of the abiotic environment to both amphibian fitness and population dynamics is the maintenance of a stable and predictable water–temperature regime (P OUNDS & C RUMP , 1994; L IPS , 1998). Many amphibians are subject to both water and temperature sensitive physiological limitations on locomotive and reproductive activities. As a consequence of this the balancing of evaporative water loss against direct absorption through the skin is a critical functional attribute, as has been observed in the Marine toad Bufo marinus (POUNDS & CRUMP, 1994). Aside from detrimental effects of disrupting this balance (i.e. through desiccation), at the individual level, the water regime in particular can play a vital role in many other aspects of amphibian ecology, including: 1.

Determination of phenological patterns of reproductive activity (WELLS, 1977; AICHINGER, 1987; GASCON, 1991); 2. Determination of the spatial distribution of community assemblages (INGLER & VORIS, 1993); and 3. In the provision of suitable breeding sites and conditions (e.g. PYBURN, 1970).

The suspected role of alterations in the annual water regime of amphibians in global population declines, follows increasing recognition of gradual changes in the global climate due to human activities. There has been a discernible human influence on world temperatures during the last century, with average temperatures projected to increase by between 1.4 and 5.8°C by 2,100, with considerably greater regional variation (IPCC, 2001). One consequence of this that is relevant here is a projected increase in activity of the tropical hydrological cycle, with the prediction of erratic and frequently severe weather patterns (GRAHAM, 1995; IPCC, 2001). The effect of climatic change on ecological systems has been observed at all levels, from population and life history alterations, to shifts in geographic range, and subsequent changes in community composition resulting in disruption of ecosystem structure and

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function (HUGHES, 2000; MCCARTY, 2001). It is therefore perhaps unsurprising that many changes in the population dynamics of amphibians, organisms which are so closely coupled with their environment, have been attributed to changing climatic and weather patterns.

A number of notable reports documenting multiple amphibian declines have implicated the potential role of synchronously observed climatic extremes, and in particular, periods of reduced or abnormally distributed rainfall (CORN & FOGELMAN, 1984; HEYER et al., 1988; BERVEN, 1990; CRUMP et al., 1992; FELLERS & DROST, 1993; STEWART, 1995). Owing to the sensitive response of amphibian breeding cycles it is easily conceivable that a simple shift in the commencement of the wet season in seasonal environments could either trigger premature spawning and subsequent desiccation of eggs, or if early rains are abnormally intense, the flooding of breeding ponds, and an equally disastrous loss of an entire breeding attempt (WELLS, 1977; CRUMP et al., 1992). POUNDS & CRUMP (1994) executed a detailed analysis of the infamous declines in number of Golden toad and Harlequin frog populations in the Monteverde cloud forests of Costa Rica, and concluded that coincidentally low periods of rainfall during phases of population decline were at least in part responsible. It was clear that depletions in number of the Harlequin frog populations (for which demographic data was available) matched climatic records of reduced rainfall during both the 1982–1983 and 1986–1987 El Niño induced drought periods. The potential role of long–term warming and increased intensity of precipitation patterns, when coupled with intense warm periods of El Niño —Southern oscillation cycles, has been noted to be of severe consequence for many biological communities (MCCARTY, 2001), and in light of the above, particularly so for amphibians (POUNDS, 2001).

With reference to the example of CRUMP et al.

(1992), it has been recently calculated that the effect of El Niño events in Central America is expected to be through severe drought periods rather than increased rainfall (HOLMGREN et al., 2001).

Although the juxtaposition of the timing and extent of population declines in Harlequin frogs with the timing and intensity of periods of drought suggests that they are causally linked, it is much more difficult to identify either the environmental variable that is of crucial ecological significance, and further, the exact mechanism by which that change acts to reduce amphibian populations. POUNDS et al. (1999) in a further analysis of the situation in the Monteverde cloud forests, isolate a perhaps rather unintuitive climatic variable as being closely correlated with not only amphibian declines but also with demographic changes in many other taxa, including birds and reptiles. The climatic variable is that of decreasing "dry season mist frequency", which suggests that the important water–related mechanism affecting amphibian populations is likely to be an increase in desiccating conditions affecting egg hatching in non–aquatic species, alongside

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individual survival, rather than changes in reproductive phenology and breeding behaviour.

Human induced climate change can be implicated here, as dry season mist frequency is negatively correlated with sea surface temperatures of the equatorial Pacific, which have increased dramatically since the mid–1970’s (STILL et al., 1999). Such examples highlight the importance of studying the environment of a species under threat so as to identify the ecologically important variables, and allow an assessment of future population stability through the parameterisation of predictive ecological models —both verbal and mathematical (MCCARTY, 2001).

Aside from the above, a number of other mechanisms have been suggested to explain the potential role of climatic change in precipitating amphibian population declines. A shift in rainfall patterns could result in a change in availability of breeding sites, a reduction in which could increase levels of competition and predation, and even vulnerability to disease, resulting in a reduced overall reproductive output for that year (DONNELLY & CRUMP 1998). An increased frequency of drought periods, coupled with increased temperatures, have also been identified as having potentially severe effects on leaf litter species which don’t congregate to breed, through alteration of their arthropod prey base and an increase in soil desiccation (DONNELLY & CRUMP, 1998). Finally, there is evidence of changes in spring spawning times of amphibian species in England, showing that amphibian reproductive cycles are highly sensitive to climate warming, with possible long–term consequences for population dynamics through alterations of biotic interactions (BEEBEE, 1995). However, a recent study of a number of other temperate–zone anuran populations suggests that this sensitivity of breeding patterns to changes in temperature exhibits marked inter–specific differences, although sufficient detailed monitoring information necessary to confidently describe such patterns of susceptibility is notably lacking (BLAUSTEIN et al., 2001).

However, as with all the potential agents of amphibian decline listed above, changes in climatic patterns cannot always be found to explain observed declines (LAURANCE, 1996; ALEXANDER & EISCHEID, 2001). Furthermore, due to the close coupling of amphibian population dynamics to their ecological environments, it is likely that any climatic change would affect amphibians through interactions with other biotic and abiotic factors, to which both the external climate and amphibians themselves are closely linked.

Interaction effects amongst environmental factors Frequently, separation of the almost myriad of current hypotheses for amphibian declines in any one situation can be almost impossible, although some recent advances have been made

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using GIS technology to compare spatial patterns of decline with spatial patterns of potentially causal agents (DAVIDSON et al., 2001). However, as noted above, most of the studies to date that have considered a single causal mechanism behind amphibian declines have invoked a critical interaction between multiple factors (ALFORD & RICHARDS, 1999; CAREY et al., 2001; MIDDLETON et al., 2001). Such acceptance means that despite the urgency of explaining observed declines, it is important to realise that interacting suites of environmental change could produce complex effects that are often difficult or even inappropriate to isolate (ADAMS, 1999).

A number of recent experimental and observational studies offer support to the importance of interaction and synergistic effects between different hypothetical agents of decline.

Increased UV–B exposure has been shown to increase the susceptibility of some amphibian species to disease (KIESECKER & BLAUSTEIN, 1995, 1997). Furthermore, an increase in UV–B can act synergistically with reduced pH levels to reduce embryo survival, when each factor alone is shown to have no significant effect (LONG et al., 1995).

Normally harmless diseases may increase their effective virulence under increased environmental pollution by contaminants such as pesticides (CAREY & BRYANT, 1995; BRIDGES & SEMLITSCH, 2000), and even different diseases themselves can be seen to act in concert in order to produce a detrimental effect (CUNNINGHAM et al., 1996). Temperature and water pH have been shown to interact to increase the detrimental effect of pathogenic fungi on reproductive success and survival in amphibians (BEATTIE et al., 1991; BANKS & BEEBEE, 1988). An interaction between a changing environment and either the virulence and distribution of a pathogen or the immuno–activity of amphibians may not be sufficient to increase mortality directly. However, through differential responses of both different amphibian species and predators it may significantly alter the competitive and predatory dynamics resulting in a shift in the species composition or abundance rank (KIESECKER & BLAUSTEIN, 1999). The presence of carbaryl pesticide has been shown to dramatically increase the level of predation stress felt by the Gray treefrog Hyla versicolor with mortality being found to be 2–4 times greater when individuals were subject to predatory cues in addition to the pesticide (RELYEA & MILLS, 2001).

Another recent study on the interaction effects of carbaryl pesticide has identified complex interactions between chemical exposure, larval competition, predation and pond drying, with results differing between species —although interestingly higher tadpole survival was observed in high density (competition) treatments which were exposed to carbaryl than in low density or control environments (BOONE & SEMLITSCH, 2002).

Although the mechanisms are not well understood (CAREY, 2000; CAREY et al., 2001), it is likely that global climate change can interact

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importantly with virtually all local environmental factors with respect to their effect on amphibian populations (P OUNDS, 2001). For example an increase in drought events and the subsequent loss of many ponds and breeding sites, could greatly exacerbate the effects of local predators (DROST & FELLERS, 1996). Also in relation to drought stress a recent experimental study on H. versicolor has identified a negative effect on larval survivorship and mass at metamorphosis from the interaction between pond drying and susceptibility to infection from the digenetic trematode parasite Telorchis sp. (KIESECKER & SKELLY, 2001). Furthermore, an increase in temperature can increase the volatility of potentially harmful chemical deposits, the aerial concentration of which may then be increased due to a reduced frequency of rainfall events (POUNDS & CRUMP, 1994). Due to the sensitivity of many ecological systems to climatic change (MCCARTY, 2001), it is likely that alterations of key environmental variables such as rainfall patterns and temperature, have the effect of reducing or even removing an important constraint on the potential of many agents of decline, both with respect to their geographic distribution but also in their physiological or biotic effect (e.g. UV–B, pH, disease) (POUNDS, 2001).

Just how complex an effect such changes in climatic parameters can precipitate has been illustrated in a very recent study by KIESECKER et al.

(2001). Their findings illustrate that climatic induced reductions in water depth at amphibian oviposition sites have caused a high level of mortality in embryos, by increasing their exposure to UV–B radiation, and consequently their vulnerability to infection by disease. The implication of this is that elevated sea surface temperatures in the tropical Pacific, which drive large scale climatic patterns, could be the precursor for many pathogen–mediated amphibian declines world–wide (KIESECKER et al., 2001). One common theme with respect to the implication of synergistic effects in amphibian population declines is that the direct or proximate mechanism which increases mortality is thought to frequently be disease following immosuppression (CAREY, 1993, 2000).

It should be clear from the above discussion that observed amphibian population declines seem unlikely to be the result of a small number of independent global agents, but rather the complex interaction of local effects in the context of varying regional influences and global climatic change. In order to study the existence of such effects in natural populations, and thus elucidate the relative stability and integrity of such populations, wellplanned programs of observation and experimentation are needed (ALFORD & RICHARDS, 1999; CAREY, 2000). Furthermore, in light of the importance of both abiotic but also biotic interactions, it is important to gain an understanding of the interactions of the populations under study with both other species (amphibians and other potential

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competitors and predators), and their physical environment. Finally in order to identify the existence of a real population decline with confidence, development of specific "null" hypotheses or models is needed to describe how amphibian populations behave in the absence of external pressures (ALFORD & RICHARDS, 1999; MARSH, 2001).

Monitoring of amphibian populations: directions and challenges Biological considerations: observations of temporal and spatial variability in amphibian population dynamics In studying the proposed phenomenon of global declining amphibian populations, there are perhaps three main questions in which uncertainty remains: 1. How to determine real declines from natural population fluctuations? 2. Whether human induced agents can be isolated as the potential cause of the decline? 3. Whether global agents are responsible for the majority of observed declines? In light a growing recognition of the important implications of the mounting extinction crisis (see above), we cannot afford to be either complacent or conservative in our approach towards answering such central questions. In view of this there is a desperate need for comprehensive monitoring studies on amphibian populations world–wide (BLAUSTEIN et al., 1994c; WAKE, 1998; YOUNG et al., 2001). As can be seen from recent compilations by ALFORD & RICHARDS (1999) and HOULAHAN et al. (2000), existing studies exhibit a notable disparity with respect to length, scope, and detail. In order to draw firm conclusions at both the local and global level, it is imperative that future studies build upon previous work, and where possible incorporate recent advances in our understanding of amphibian species and their population dynamics. It is therefore instructive here to draw attention to a number of considerations, both biological and methodological, which are central to planning amphibian– monitoring programs.

The detection of real population declines which are deserving of concern, from purely natural population fluctuations can pose a serious problem in monitoring programs. It is essential that we understand the natural levels of variability inherent in amphibian populations, so as not to invoke unnecessary conservation and management action —a result that could severely compromise support for conservation in other situations (PECHMANN et al., 1991). An understanding of the levels of variability inherent in population dynamics is central to calculating both the statistical power of a monitoring program (MARSH, 2001), and the level of extinction risk from stochastic events (LEIGH, 1981; ENGEN & SAETHER, 1998; MARSH, 2001). Furthermore, an appreciation of population variability is fundamental to

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understanding the processes that drive population fluctuations (SEMLITSCH et al., 1996).

There exist both temporal and spatial aspects of amphibian population dynamics that can serve to confound the attempts of many monitoring programs to elucidate real declines in number.

A number of ecological variables have been proposed as predictors of variability in amphibian populations, including; habitat type (W ILLIAMS & HERO , 1998, 2001), reproductive mode and density dependent processes (SEMLITSCH et al., 1996; ALFORD & RICHARDS, 1999; MARSH, 2001), rainfall, taxanomic family, and latitude (MARSH, 2001). Firstly, there is strong evidence of intraspecific density dependence in many amphibian populations (B ERVEN , 1990; PECHMANN et al., 1991; MEYER et al., 1998; ALFORD & RICHARDS, 1999). Crucial life history factors that appear to be regulated by density dependence include larval survival, larval size and time to metamorphosis. In a highly heterogeneous environment such as a forest, variance in such factors could produce seemingly chaotic fluctuations in population size (TURNER, 1962; BERVEN, 1990; PECHMANN et al., 1991). In the wood frog (Rana sylvatica) BERVEN (1990) recorded variation in R0 (the net population replacement rate) between 0.009–7.49 over only 7 years, and as monitoring programs are rarely longer than this (BLAUSTEIN et al., 1994c; ALFORD & RICHARDS, 1999) it is easy to see how a short term population decline may be interpreted with unwarranted concern. Secondly, it is possible that at the level of the population, density dependent effects following fluctuations in resource levels may override the effect of any density independent environmental stress factors that may act to reduce juvenile or adult survival. However, as the judgement of "natural" levels of stability in biological populations is exceedingly difficult to make (CONNELL & SOUSA, 1983), it is consequently difficult to identify the ecological significance of any such contribution to mortality or reduction in reproductive success to overall population dynamics, even though they may be non–trivial.

Aside from density dependence, a second important consideration of the temporal dynamics of amphibians is in the fact that fluctuations in breeding aggregations may be much greater than fluctuations in total population size, due to intra– population variance in breeding behaviour (PECHMANN et al., 1991). This point is of particular relevance, as due to severe logistical constraints most censuses of amphibians and especially frogs and toads (Amphibia, Anura), are conducted on aggregations at breeding sites (ALFORD & RICHARDS, 1999; see below).

Aside from temporal considerations one must also take into account the spatial aspects of amphibian population dynamics in making any conclusions about population stability or integrity.

It is becoming increasingly recognised that many amphibian populations often exist in a metapop-

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ulation structure, where regional processes affecting extinction and colonisation of habitat patches, play a dominant role in determining local species assemblage composition and population size (HANSKI & GILPIN, 1991; HECNAR & M’CLOSKEY, 1996b; ALFORD & RICHARDS, 1999; MARSH & TRENHAM, 2000). It is important to recognise that from the perspective of a monitoring program effective habitat "patches" which are subject to such regional influences can represent the breeding sites or transects under human surveillance, and with respect to the actual amphibian population are often not ecologically distinct. The importance of regional processes to the persistence of local populations means that local extinction can occur due to essentially stochastic factors that are unrelated to the local environmental (abiotic or biotic) quality (SJOGREN, 1991; MARSH, 2001). Local amphibian populations are predisposed to stochastic extinctions due to the susceptibility of a peak–breeding attempt to climatic conditions (i.e. droughts or floods), their relatively short life spans (M ARSH & T RENHAM , 2000), and their philopatric behaviour (WALDMAN & TOCHER, 1998).

However, the essential point is that although devastating reductions in population size may be observed at a particular monitoring site, at the regional spatial scale the species may be perfectly healthy, adding doubt to the extrapolation of many population censuses to conclusions about the viability of an entire species. In order to confidently assess the stability of an amphibian population, and attribute a reason to any observed decline, it is important that such spatial factors are considered (see below).

Methodological considerations: challenges and pre–requisites for effective amphibian population monitoring The natural variability and complexity that is inherent in both temporal and spatial amphibian population dynamics has already been highlighted above. It is crucial to recognise that such factors introduce serious practical considerations and caveats in the construction, execution, and analysis of amphibian monitoring programs. A direct consequence of such natural levels of variability is that the failure to find a significant decline in number of a particular population may frequently not be due to a lack of real decline, but rather to a lack of statistical power (GIBBS, 1995; REED & BLAUSTEIN, 1995; HAYES & STEIDL, 1997; ALFORD & RICHARDS, 1999; MARSH, 2001). The statistical power of a test for a population decline can be defined as the probability of rejecting the null hypotheses of no decline given that the null hypothesis is false and the alternative hypothesis of a declining population is true. Calculation of power requires knowledge of a number of factors, namely the sample size, the desired alpha level for avoiding Type I errors, the natural variance in sample size, and the effect size (PETERMAN, 1990). The value of

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conducting a prospective power analysis (HAYES & STEIDL, 1997) is severely limited by the high level of uncertainty inherent in the last two factors: 1.

The natural coefficient of variability in amphibian population size —in particular in the context of the specific monitoring approach being used; 2.

The level of population decline (effect) which bears ecological significance for the future stability of an amphibian population. Uncertainty in these values produces an equal level of uncertainty in level of power calculated (GIBBS, 1995). Attempts should be made to calculate the confidence in intervals associated with estimates of power, and furthermore there may be considerable merit in using Bayesian approaches to estimate levels of uncertainty (HILBORN & MANGEL, 1997; WADE, 2000).

It is suggested here that in light of the serious logistical and financial limitations imposed on many, if not most amphibian monitoring projects, the "guestimating" of such variables is a dangerous game as it may render void many otherwise valuable projects which are lacking in apparent statistical rigour (and crucially lacking in ability to expand the project’s sample size to achieve a satisfactory level of power). Except in situations where the species under surveillance is well studied, it may be of greater ecological significance to take a comprehensive approach to monitoring which incorporates a number of key ecological, as well as methodological considerations. This will hopefully achieve an increased understanding of the environmental requirements and population dynamics of the specific focal species, and afford greater confidence in any data interpretation.

This does not nullify the clear value of prospective power analysis, but rather suggests that there is a great deal of merit in carefully considered monitoring projects which do not yet hold the minimum level of information needed to make such a preliminary analysis worthwhile. Identified below are some of the considerations deemed central to amphibian population monitoring.

For reasons emphasised earlier the two main problems facing monitoring projects are the logistical constraints on their temporal and spatial focus. Clearly in order to elucidate real declines from stochastic fluctuations, a long time series is highly favourable, although as seen from recent literature reviews few studies are longer than five years, and even less are more than 10 (ALFORD & RICHARDS, 1999; HOULAHAN et al., 2000; YOUNG et al., 2001). It is important to note however that an increase in the length of the study period will undoubtedly increase the perceived level of variability in population size and distribution due to the incorporation of a greater range of environmental conditions (PECHMANN & WILBUR, 1994; M ARSH, 2001). Secondly, in light of a commonly metapopulation structure and the critical role of processes such as emigration and colonisation in amphibian populations, a regional monitoring perspective is important in order to distinguish overall regional declines from local

Animal Biodiversity and Conservation 24.2 (2001)

(and essentially stochastic) extinctions (MARSH & TRENHAM , 2000; MARSH, 2001). For most amphibians this requires an appreciation as to the importance of the spatial arrangement, and degree of isolation between different breeding sites. It is appropriate to note here that according to a recent review of techniques used to quantify amphibian populations, most attempts focus on direct or indirect (e.g. vocal calls, egg masses) counts at breeding sites (ALFORD & RICHARDS, 1999) —utilising the fact that most species congregate en masse to breed (BEEBEE, 1996). It thought that the population dynamics of a species are determined primarily by recruitment processes occurring at breeding ponds, and that such a focus can accurately determine the cause of any local or regional decline (MARSH & TRENHAM, 2000).

However, such an exclusive focus carries a number of caveats in data interpretation: 1. Variation in population size at breeding ponds can as well be due to variation in breeding behaviour as to actual variation in population number (PECHMANN et al., 1991); 2. It is often impossible to clearly distinguish variation in population size from simply variation in the size of breeding aggregations —i.e. the degree of "openness" of the population (MCARDLE & GASTON, 1993). Both such measures represent useful information but it is important to note that they are not synonymous; 3. In terms of adult survival and distribution an exclusive focus on breeding sites ignores the potential importance of the intervening terrestrial habitat which may be of ecological significance (MARSH & TRENHAM, 2000).

A final note with respect to natural variability in amphibian populations is interspecific or taxonomic variance. As emphasised earlier different species exhibit different levels of susceptibility to different agents of decline (e.g.

DROST & FELLERS, 1996). Accordingly, any attempt to assess the stability or vulnerability of an amphibian fauna at any one regional site should consider not only the breadth of species present, but also a number of different populations of each. With relevance to all levels of variability that can serve to confound attempts to identify declining populations, a high number of intra– annual repeat visits to each monitoring site (especially during the peak breeding season) can add important, if not essential strength to the results (ALFORD & RICHARDS, 1999).

Alongside data on the population dynamics of the focal species, it is important to gain an appreciation of the differential importance of key ecological variables in both the biotic and abiotic environment (e.g. climate, water quality, floral composition, and predator abundance).

Such a multidimensional approach to monitoring helps to identify any potential agents of decline, but also to helps predict any secondary effects or feedbacks following a potential change in the structure of the amphibian community. This information, when integrated into demographic

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data of the amphibian populations can be invaluable in building a null model to predict the range and patterns of population behaviour in the absence of external pressures or agents of decline (ALFORD & RICHARDS, 1999). Such null models can be used to reduce the subjectivity and ambiguity that often surrounds the evidence for a population decline (POUNDS et al., 1997).

The way ahead: past lessons and future potential In conclusion it is fair to say that conservation biology is still far from providing confident answers to the three questions posed above —how to determine real declines from simply natural population fluctuations, how to isolate the causal agents of a decline, and whether any particular factor is of global relevance. None of these are trivial questions, although each poses a significantly different challenge. With respect to the latter two questions it is becoming increasingly clear that yes human-induced agents can frequently be isolated as being causal factors behind population declines, but also that there exists a multitude of such factors operating at different scales, many of which exhibit complex interactions with both other factors and the local environment. The importance of a particular agent of decline in any one area or for any one species is likely to be context dependent, with synergistic effects that are difficult if not practically impossible to tease apart. However, as has been shown above for both specific and more general cases significant progress has been made.

It is likely that further progress is only really possible through the interaction of both the many disciplines of ecology and environmental science, but also of ecosystem management, public policy and economics (e.g. LUDWIG et al., 2001) —all of which contribute towards the precipitation, identification and mitigation of amphibian population declines.

In allocating limited conservation resources to the problem of declining amphibian populations the first of the three questions outlined above takes paramount importance —when are we observing real population declines and when are we just measuring natural population fluctuations? As was discussed in the final section of this review, this question has arisen through observing high levels of natural variability in amphibian population dynamics across both temporal and spatial scales —variability which serves to confront the fieldworker with a number of severe methodological challenges. The central distillation of this problem reveals a trade–off between needing enough statistical power to effectively reject the null hypothesis of no decline in cases where a decline truly exists, and the simple truth that conservation biology has insufficient funds (or historically accurate and detailed population data sets) to conduct and analyse exhaustively long monitoring programs in

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every case of suspected population decline.

However, although this problem requires careful and objective assessment, recent progress and hard– won experience provides a number of possible alternatives for its confrontation: 1. Firstly as discussed in the previous section existing and proposed monitoring programs could be greatly strengthened in their ability to identify regional amphibian declines if they undertook a more multi–dimensional approach. Recent results from such a monitoring program in Belize provides evidence that natural variability exists in all of; species presence, relative abundance and calling activity (often taken as a surrogate of audible abundance) —across a range of temporal (within and between nights, across season, and between years), spatial (between ponds, even of similar habitat), and environmental scales (between different habitat types and climatic conditions) (GARDNER & FITZHERBERT, 2001; Gardner et al., unpublished data). Although not a substitute for long time series our ability to identify regional declines in amphibian populations would be greatly enhanced by the simultaneous monitoring of a range of both breeding sites and species, and across as many temporal scales as possible. Furthermore, such information provides a much better understanding of the underlying mechanisms which produce the observed variation.

2. Despite the value of the above recommended comprehensive approach to monitoring it still demands levels of resource allocation which may frequently be unavailable to practising conservationists. A recent application of a genetic test for bottlenecks (CORNUET & LUIKART, 1996) to distinguish between natural oscillations and true population declines in British Natterjack toads (Bufo calamita) (BEEBEE & ROWE, 2001) presents one potentially very useful alternative to resource intensive monitoring programs. BEEBEE & ROWE (2001) analysed a range of Natterjack populations, including ones which have experienced a recent decline, and ones which have remained comparatively stable. Microsatellite allele frequency data from these populations were tested for heterozygote excess and shifts in allele frequency distributions, and inferences from these computations about bottlenecks (i.e., persistently smaller population sizes than the recent means) were compared with demographic information. The genetic test accurately differentiated between declining and relatively stable populations (BEEBEE & ROWE, 2001). Recent theoretical (LUIKART et al., 1998, 1999) and empirical (SPENCER et al., 2000) work on the requirements for such tests suggests that to achieve sufficiently high power they only require samples of 5 to 20 polymorphic loci and approximately 30 individuals. The same work has also identified allelic diversity and temporal variation in and temporal variance in allele frequencies were most sensitive to genetic changes that resulted from the bottlenecks —but not the proportion of polymorphic loci (SPENCER et al., 2000).

3. The identification of declines in extant

populations often requires a simple historical record of prior distributions of species occurrence. Recent work on a wide range of plants and animals provides encouragement that museum collections can be successfully used analyse declines, at least at a coarse spatial scale (SHAFFER et al., 1998).

4. As noted above a recent meta–analysis has been conducted on fluctuations in amphibian populations (MARSH, 2001). This work identified a number of predictive correlates of natural variability in amphibian population dynamics, notably life history type, family and latitude —correlates which could provide a rough but useful guide to the regions and species groups in which we may expect either greater or less than average natural variability in population fluctuations, and therefore help separate cases of particular concern.

5. Recent work has employed the use of skeletochronology to describe the differences in demographic composition between different populations of amphibians (e.g. DRISCOLL, 1999; REASER, 2000; KHONSUE et al., 2001). When coupled with mark–recapture data skeletochronology can provide invaluable information on age structure of a population, the stability of such age cohorts, and therefore the potential of the population to undergo large fluctuations in population size (e.g. DRISCOLL, 1999). Aside from helping to identify the potential for population variability, this technique can help isolate populations which have a skewed–senile age distribution —thus indicating a lack of recent recruitment and an accompanying higher risk of local extinction.

The above list of methodological and analytical techniques provides some undeniably valuable tools for the conservation biologist who is faced with identifying declining amphibian populations which are cause for concern, while at the same time is equipped with a limited budget. However, the list is not exhaustive, and another message that needs to be emphasised is that it is of utmost importance to maintain an open and vigilant mind with respect to new and evolving ideas and techniques. Only by adopting a flexible and holistic approach to conservation, can we profitably employ and integrate the diversity of knowledge and experience that exists in the many disciplines of ecology, environmental science and management —and thereby provide an increasingly effective response to dealing with the declining amphibian phenomenon.

A note of caution A final note of caution in studying the declining amphibian phenomenon needs to be emphasised.

Although both scientific (WAKE, 1998), and public (MORELL, 2001) opinion recognises the severity of declining amphibian populations, it is important to maintain a broad appreciation of other conservation problems and priorities (HALLIDAY, 2001). Two points should be considered at this

Animal Biodiversity and Conservation 24.2 (2001)

junction. Firstly, amphibian declines have occurred in a number of pristine habitats and protected areas —removed from areas of direct human impact. This evidence bears serious implications for the effectiveness of the protected areas approach to conservation, and as such the study of amphibian populations should be integrated wherever possible into the wider context of conservation science and action. Secondly, it is important in world where resource allocation to conservation biology is seriously inadequate, that an focus or even over–emphasis on amphibians does not eclipse the equally worrying status of many other taxanomic groups (e.g. GIBBONS et al., 2000; GROOMBRIDGE & JENKINS, 2000) from both the scientific and public eye.

Acknowledgements I would like to express my thanks to all the members and advisors of Project Anuran during the last 3 years, who have provided the forum for discussion of many of the issues raised in this paper. Particular thanks goes to Emily Fitzherbert who co–founded our DAPTF monitoring project in Belize. I am also grateful to M. Tejedo, A. Salvador and one anonymous referee for providing helpful and constructive comments which greatly improved an original draft of this review.

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Tinautius exilis sp. n.

(Coleoptera, Carabidae, Pterostochini) de la Alpujarra almeriense (SE España) J. Mateu

Mateu, J., 2001. Tinautius exilis sp. n. (Coleoptera, Carabidae, Pterostochini) de la Alpujarra almeriense (SE España). Animal Biodiversity and Conservation, 24.2: 45–49.

Abstract Tinautius exilis n. sp. (Coleoptera, Carabidae, Pterostochini) from the Alpujarra of Almeria (SE Spain).— A new troglobitic species belonging to the genus Tinautius Mateu 1997, from the Sierra de Cazorla in the province of Jaen is described. This new species (T. exilis n. sp.) was found in the Alpujarra mountains in Almería. It is anophthalmous and differs from the anterior species in several morphological features. The female genitalia and the particular shape of the metatrochant enable simple differentiation with the two species of the genus Tinautius known to date.

Key words: Coleoptera, Carabidae, Pterostichini, Tinautius exilis n. sp., Almería, Andalucia, Spain.

Resumen Tinautius exilis sp. n. (Coleoptera, Carabidae, Pterostochini) de la Alpujarra almeriense (SE España).— Se describe una nueva especie troglobia perteneciente al género Tinautius Mateu 1997, procedente de la Sierra de Cazorla, provincia de Jaén. La nueva especie (T. exilis sp. n.) proviene de la Alpujarra almeriense. Es anoftalma y difiere de T. troglobius por diferentes caracteres morfológicos y anatómicos. La genitalia femenina y la especial conformación del metatrocánter permiten separar fácilmente las dos especies de Tinautius conocidas hasta ahora.

Palabras clave: Coleoptera, Carabidae, Pterostichini, Tinautius exilis sp. n., Almería, Andalucía, España.

(Received: 16 V 01; Conditional acceptance: 12 IX 01; Final acceptance: 15 XI 01) Joaquín Mateu, Museu de Zoologia, Psg. Picasso s/n., Parc. de la Ciutadella, 08003 Barcelona, España (Spain).

ISSN: 1578–665X

© 2001 Museu de Zoologia

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Introducción En la cueva de La Corraliza, término municipal de Fondón, provincia de Almería, han sido capturados 5 ejemplares (2{ y 3}) de una nueva especie del género Tinautius Mateu, 1997, cuya especie principal (Tinautius troglophilus Mateu) fue hallada en la cueva Peal del Becerro, provincia de Jaén, hace unos pocos años.

La nueva especie (Tinautius exilis sp. n.) proviene de la Alpujarra almeriense y se trata de una especie absolutamente troglobia (anoftalmia, despigmentación del tegumento, delgadez y alargamiento de los apéndices, etc.), mientras que en T. troglophilus los caracteres troglobios están todavía poco evolucionados, o sea, primarios; sin embargo, ese grado de especialización parece ser ya suficiente para mantener la especie confinada y dependiente, del medio subterráneo. En oposición, T. exilis sp. n., presenta un alto grado de evolución hipogea: patas y antenas largas y finas, tegumentos rojizos y despigmentados, cabeza voluminosa suborbicular carente de ojos, sedas principales de la seda umbilicada flageliformes y más o menos onduladas, cuerpo alargado, estrecho y paralelo, etc. A primera vista tal conformación general podría inducirnos a considerar T. exilis sp. n. como representante de un género inédito pero, a pesar de la diferencias observadas, éstas en realidad son más cuantitativas que cualitativas. Finalmente, el estudio de la genitalia hembra de las dos especies que constituyen el género Tinautius, nos ha confirmado, por sus peculiares características comunes en ambas especies, la bondad de reunirlas en una misma entidad genérica.

Resultados

Tinautius exilis sp. n.

Material estudiado Holotipo: 1{ de la cueva de la Corraliza, Fondón, provincia de Almería (España), 21 V 2000, T. G. Pardo & M. Piquer leg., colección J. Mateu.

Paratipos: de la misma localidad que el holotipo: 1}, 21 V 2000, D. Ortega & J. G. Mayoral leg.; 2}, 29 X 2000, J. G. Mayoral & D. Ortega leg.; 1{, 23 XII 2000, D. Ortega & J. G. Mayoral leg.; col. J. Mateu.

Descripción Insecto estrecho, alargado (fig. 1), élitros paralelos, color rojo testáceo más o menos oscurecido. Cabeza gruesa, más larga que ancha, poco más ancha que el pronoto, anoftalmo.

Pronoto alargado poco dilatado por delante, con los ángulos posteriores grandes y algo aguzados. Élitros paralelos con estrías completas.

Metatrocánteres largos y atenuados en su parte apical, algo más largos que la mitad de la longitud

Mateu

del fémur. Aparato genital de la } con los gonópodos presentando una corona con fuertes y largas espinas quitinizadas (fig. 6).

Cabeza grande, alargada y gruesa, ligeramente suborbicular y glabra. Desprovisto de ojos, éstos simplemente indicados por una mancha blancuzca lisa y sin onmatidias visibles.

Sienes largas, glabras y algo abultadas; cuello grueso. Mandíbulas bastante largas, robustas y poco curvadas en su extremidad. Antenas largas, pubescentes a partir del cuarto artejo, sobrepasando hacia atrás la base del pronoto de poco más de dos artejos. Dos sedas suborbitales de cada lado y una seda epistomal. Labro truncado provisto de 6 sedas sobre el borde anterior. Superficie finamente punteada, más densamente entre las impresiones frontales poco arqueadas y que a su vez limitan una depresión interocular. Labio articulado, diente labial bastante grande y bífido y dos grandes sedas basales. Paraglosas largas y estrechas sobrepasando largamente la lengüeta, que tiene el borde truncado y bisetulado.

Pronoto alargado, 1,20 veces más largo que ancho, nada cordiforme y poco más ancho que la cabeza, convexo, pero aplanado por encima a lo largo del surco mediano. Ángulos anteriores pequeños, obtusos y algo salientes; lados poco redondeados por delante y débilmente sinuosos cerca de los ángulos posteriores algo salientes hacia afuera. Reborde lateral fino y estrecho.

Base rectilínea apenas escotada en el medio.

Fositas basales superficiales y punteadas. Dos sedas laterales, la primera muy adelantada, la posterior en el ángulo.

Prosterno fino y espaciadamente punteado por debajo; apófisis prosternal no rebordeada, levemente incisa en el centro, plana y apenas pubescente. Élitros largos, estrechos y paralelos con una estría escutelar bien incisa sobre el segundo intervalo. Base rebordeada. Aun que convexos, los élitros están aplanados por encima y la convexidad es muy fuerte y brusca a partir de la séptima estría. Húmeros redondeados, espaldas altas. Todas las estrías elitrales son completas y bien marcadas y finamente punteadas. Intervalos débilmente convexos y lisos, el tercero con tres poros, el primero contra la tercera estría, los otros dos contra la segunda; en otros ejemplares la segunda y tercera estrías se anastomosan.

Serie umbilicada de 13 poros: 6–1–6.

Sobre la cabeza y el pronoto la microescultura es prácticamente nula, en cambio los élitros presentan una fina y regular microescultura compuesta por mallas estrechas y transversales.

Patas largas y finas con el profémur hinchado y el metafémur delgado, casi estrangulado en su parte basal próxima al trocánter; metatrocánter muy largo, atenuado hacia el ápice, sobrepasando la mitad del metafémur y con algunos gránulos quitinosos en su extremidad.

Los tres primeros artejos de los protarsos dila-

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1mm

1mm

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Fig. 1. Habito de Tinautius exilis sp. n. de la cueva de la Corraliza de Fondón, provincia de Almería.

Fig. 2. Habito de Tinautius troglophilus Mateu, de la cueva PB4, de Peal del Becerro, Cazorla.

Fig. 1. Habitus of Tinautius exilis n. sp. from Corraliza de Fondón cave, province of Almeria.

Fig. 2. Habitus of Tinautius troglophilus Mateu from PB4 cave, Peal del Becerro, Cazorla.

tados y con faneras adhesivas por debajo en el macho. Todos los tarsos son glabros y subcarenados por encima. Oniquio también glabro por debajo. Uñas lisas.

Edeago (fig. 3) de tipo Pterostichini, acodado débilmente en ángulo agudo casi recto. Ápice visto de perfil largo y bastante afilado, arremangado y provisto de unos pocos gránulos dorsales.

El orificio del saco interno, visto por encima, se muestra algo caído hacia su lado izquierdo. El parámero izquierdo en forma de paleta redondeada, el parámero derecho es corto, robusto, subparalelo y romo en la extremidad.

Aparato sexual de la hembra (fig. 6) parecido al del Tinautius troglophilus Mateu, pero con el gonocoxito con punta mucho más grande y menos arqueada; su base presenta cuatro o cinco espinas envolventes en lugar de las seis del T. troglophilus . Espermateca también digitiforme, notablemente más larga en la nueva especie y con un muy largo conducto espermatecal que le une al saco vaginal (más de dos veces y media más largo que la espermateca

que en T. troglophilus, es solamente vez y media más largo que el receptaculum seminis) (fig. 6). Glándula suplementaria un tercio más larga que éste último y terminada en una gran bola piriforme.

Observaciones Las desigualdades que presentan las dos especies del género Tinautius pueden tal vez interpretarse como el resultado de un diferente grado de evolución entre ambas especies (genepistasia).

Así, la diversidad, por lo que atañe a la forma general de las genitalias del macho y de la hembra, permitiría suponer que ésta haya sido inducida en su origen por la presión del medio ambiente subterráneo y por la distinta velocidad evolutiva o heteropistasia.

El T. troglophilus Mateu es, como su nombre indica, una especie troglófila en tránsito para alcanzar plenamente la categoría de troglobia a parte entera. Los caracteres discriminativos que caracterizan esta última categoría ya se perciben

Mateu

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3

4

5

Figs. 3–5. Edeagos de: 3. Tinautius exilis sp. n. de la cueva de la Corraliza de Fondón, Almería; 4. Tinautius troglophilus Mateu, de la cueva PB4 de Peal del Becerro, Cazorla, Jaén; 5.

Metafémur y trocánter del Tinautius exilis sp. n.

Figs. 3–5. Aedeagus of: 3. Tinautius exilis n. sp. from Corraliza de Fondón cave, Almeria; 4.

Tinautius troglophilus Mateu, from PB4 cave of Peal del Becerro, Cazorla, Jaen; 5. Metafemur and trochanter of Tinautius exilis n. sp.

Fig. 6. Genitalia femenina de Tinautius exilis sp. n. de la cueva de la Corraliza de Fondón, Almería.

Fig. 7. Genitalia femenina de Tinautius troglophilus Mateu de la cueva de Peal del Becerro, Cazorla, Jaén.

Fig. 6. Female genitalia of Tinautius exilis n. sp. from Corraliza de Fondón cave, Almeria.

Fig. 7. Female genitalia of Tinautius troglophilus Mateu from Peal del Becerro cave, Cazorla, Jaen.

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“pro parte” en T. troglophilus: reducción ocular con atenuación y aun desaparición del número de ommatidias y también de las carínulas suborbitales y genales, despigmentación de los tegumentos, etc. En oposición, T. exilis sp. n., ofrece dichos caracteres en su fase extrema: anoftalmia total, especialización del sistema sensorial (sedas flageliformes), cuerpo largo estrecho y paralelo, alargamiento de los apéndices y patas, despigmentación, etc. Estos caracteres troglobios son sin duda autapomórficos y se encuentran al unísono con aquellos otros plesiomorfos (o juveniles) de la forma ancestral epigea y, por derivación, con esta otra más evolucionada, es decir con T. troglophilus (figs. 2, 4 y 7).

Una notable autapomorfía se ha desarrollado en T. exilis, es el considerable alargamiento del metatrocánter. Este carácter se conoce solamente de algunos géneros de Carabidae, y siempre como carácter sexual secundario. No es el caso en T. exilis, puesto que tal anomalía se verifica en ambos sexos y en todos los ejemplares (fig. 5).

Es factible que la presión selectiva del medio hipogeo y de los factores genéticos del mismo, además de la segregación geográfica inicial entre ambas poblaciones (de Sierra Nevada y Sierra de Cazorla), sean en gran parte responsables de la diversificación morfogenética del género que nos ocupa. Hasta ahora solamente dos especies son conocidas, pero no es imposible que otras puedan aparecer un día, especies o poblaciones intermedias, en cuevas no prospectadas todavía.

Derivatio nominis exilis, en latín delgado, magro.

Ubicación geográfica Aproximadamente unos 120 km, separan en

línea recta la localidad almeriense de Fondón en la provincia de Almería, patria del T. exilis sp. n., de aquella otra jienense de Peal del Becerro (provincia de Jaén) de la Sierra de Cazorla, patria del T. troglophilus Mateu.

El nuevo taxon fue capturado en una gruta situada en calizas triásicas béticas de la Alpujarra almeriense; el segundo procede de una caverna ubicada en las unidades pre–béticas calcáreas de Cazorla. Ambos fueron recolectados en cavidades situadas a mediana altitud, dentro del complejo serrano de las cordilleras béticas de Sierra Nevada y de las del Segura a 1.050 m y 1.710 m, respectivamente.

Datos sobre el área de estudio pueden ser consultados en la obra colectiva titulada “Parque Natural de Sierra Nevada”.

Agradecimientos Agradecemos al profesor D. Pablo Barraco Vega y a sus colaboradores los señores Daniel Ortega Sánchez y Jaime García Mayoral por el material recolectado en sus propecciones en la Corraliza.

Esta investigación ha podido realizarse gracias a un proyecto subvencionado por la Federación Andaluza de Espeleología y al Espeleoclub Almería.

Referencias MATEU, J., 1997. Tinautius n. gen troglophilus n. sp., nuevo Pterostichini hipogeo del sur de España (Coleoptera, Carabidae). Bol. Mus. Reg. Sci, nat. Torino, 15: 137–146.

MOLERO MESA, J., PÉREZ RAYA, F. & VALLE TENDERO, F. (Eds.), 1992. Parque Natural de Sierra Nevada.

Editorial Rueda, Madrid.

"La tortue greque" Oeuvres du Comte de Lacépède comprenant L'Histoire Naturelle des Quadrupèdes Ovipares, des Serpents, des Poissons et des Cétacés; Nouvelle édition avec planches coloriées dirigée par M. A. G. Desmarest; Bruxelles: Th. Lejeuné, Éditeur des oeuvres de Buffon, 1836. Pl. 7

Editor executiu / Editor ejecutivo / Executive Editor Joan Carles Senar

Secretaria de Redacció / Secretaría de Redacción / Editorial Office

Secretària de Redacció / Secretaria de Redacción / Managing Editor Montserrat Ferrer

Museu de Zoologia Passeig Picasso s/n 08003 Barcelona, Spain Tel. +34–93–3196912 Fax +34–93–3104999 E–mail mzbpubli@intercom.es

Consell Assessor / Consejo asesor / Advisory Board Oleguer Escolà Eulàlia Garcia Anna Omedes Josep Piqué Francesc Uribe

Editors / Editores / Editors Antonio Barbadilla Univ. Autònoma de Barcelona, Bellaterra, Spain Xavier Bellés Centre d' Investigació i Desenvolupament CSIC, Barcelona, Spain Juan Carranza Univ. de Extremadura, Cáceres, Spain Luís Mª Carrascal Museo Nacional de Ciencias Naturales CSIC, Madrid, Spain Adolfo Cordero Univ. de Vigo, Vigo, Spain Mario Díaz Univ. de Castilla–La Mancha, Toledo, Spain Xavier Domingo Univ. Pompeu Fabra, Barcelona, Spain Francisco Palomares Estación Biológica de Doñana, Sevilla, Spain Francesc Piferrer Inst. de Ciències del Mar CSIC, Barcelona, Spain Ignacio Ribera The Natural History Museum, London, United Kingdom Alfredo Salvador Museo Nacional de Ciencias Naturales, Madrid, Spain José Luís Tellería Univ. Complutense de Madrid, Madrid, Spain Francesc Uribe Museu de Zoologia de Barcelona, Barcelona, Spain

Consell Editor / Consejo editor / Editorial Board José A. Barrientos Univ. Autònoma de Barcelona, Bellaterra, Spain Jean C. Beaucournu Univ. de Rennes, Rennes, France David M. Bird McGill Univ., Québec, Canada Mats Björklund Uppsala Univ., Uppsala, Sweden Jean Bouillon Univ. Libre de Bruxelles, Brussels, Belgium Miguel Delibes Estación Biológica de Doñana CSIC, Sevilla, Spain Dario J. Díaz Cosín Univ. Complutense de Madrid, Madrid, Spain Alain Dubois Museum national d’Histoire naturelle CNRS, Paris, France John Fa Durrell Wildlife Conservation Trust, Trinity, United Kingdom Marco Festa–Bianchet Univ. de Sherbrooke, Québec, Canada Rosa Flos Univ. Politècnica de Catalunya, Barcelona, Spain Josep Mª Gili Inst. de Ciències del Mar CMIMA–CSIC, Barcelona, Spain Edmund Gittenberger Rijksmuseum van Natuurlijke Historie, Leiden, The Netherlands Fernando Hiraldo Estación Biológica de Doñana CSIC, Sevilla, Spain Patrick Lavelle Inst. Français de recherche scient. pour le develop. en cooperation, Bondy, France Santiago Mas–Coma Univ. de Valencia, Valencia, Spain Joaquín Mateu Estación Experimental de Zonas Áridas CSIC, Almería, Spain Neil Metcalfe Univ. of Glasgow, Glasgow, United Kingdom Jacint Nadal Univ. de Barcelona, Barcelona, Spain Stewart B. Peck Carleton Univ., Ottawa, Canada Eduard Petitpierre Univ. de les Illes Balears, Palma de Mallorca, Spain Taylor H. Ricketts Stanford Univ., Stanford, USA Joandomènec Ros Univ. de Barcelona, Barcelona, Spain Valentín Sans–Coma Univ. de Málaga, Málaga, Spain Tore Slagsvold Univ. of Oslo, Oslo, Norway

Animal Biodiversity and Conservation 24.1, 2001 © 2001 Museu de Zoologia, Institut de Cultura, Ajuntament de Barcelona Autoedició: Montserrat Ferrer Fotomecànica i impressió: Sociedad Cooperativa Librería General ISSN: 1578–665X Dipòsit legal: B–16.278–58

Animal Biodiversity and Conservation 24.2 (2001)

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Changes in abundance of birds in a Neotropical forest fragment over 25 years: a review W. D. Robinson

Robinson, W. D., 2001. Changes in abundance of birds in a Neotropical forest fragment over 25 years: a review.

Animal Biodiversity and Conservation, 24.2: 51–65.

Abstract Changes in abundance of birds in a Neotropical forest fragment over 25 years: a review.— Few data are available to evaluate the long term effects of habitat isolation on species richness or abundances in the tropics.

Barro Colorado Island (BCI), Panama, has been studied for more than 80 years since its isolation from surrounding lowland forest when the Panama Canal was constructed. Thirty-five percent of the originally present 200 resident species have disappeared. Although the loss of species is well–studied, changes in abundance that might help predict future losses have not been evaluated. One study in 1970 and the present study conducted 25 years later estimated abundances of most bird species on BCI. Comparisons indicate at least 37 species have declined by at least 50%. Twenty–six species of edge habitats are expected to decline as forest maturation proceeds, yet 11 forest species that are now rare may be lost soon. All 26 species that were present in 1970 but not detected in the mid–1990s were rare in 1970. Thus, rarity appears to be a good predictor of extinction risk in this tropical habitat fragment.

Key words: Barro Colorado Island, Extinction, Faunal relaxation, Habitat fragmentation, Neotropics, Panama.

Resumen Cambios en la abundancia de aves en un fragmento de bosque neotropical durante un período de 25 años: una revisión.— Hay pocos datos disponibles para evaluar los efectos a largo plazo que supone el aislamiento del hábitat con respecto a la riqueza o la abundancia de especies en el trópico. La Isla de Barro Colorado (BCI), Panamá, se ha estado estudiando durante más de 80 años, desde que la construcción del Canal de Panamá la dejara aislada de los bosques de las tierras bajas circundantes. El treinta y cinco por ciento de las 200 especies residentes inicialmente presentes ha desaparecido. Aunque la pérdida de especies se ha estudiado a fondo, no se han evaluado los cambios en abundancia que podrían ayudarnos a predecir pérdidas futuras. Un estudio de 1970 y el presente estudio, realizado 25 años después, han estimado la abundancia de la mayoría de especies de aves presentes en la BCI. Las comparaciones indican que al menos 37 especies han disminuido en un 50%, como mínimo. Se prevé que 26 especies pertenecientes a hábitats de las orillas vayan disminuyendo con la maduración del bosque, si bien 11 especies del bosque que ahora son poco frecuentes podrían extinguirse muy pronto. La totalidad de las 26 especies existentes en 1970, pero que no se detectaron a mediados de la década de 1990, ya eran raras entonces. Así pues, el hecho de que una especie sea rara parece constituir un buen indicador del riesgo de extinción en este fragmento de hábitat tropical.

Palabras clave: Isla de Barro Colorado, Extinción, Relajamiento de la fauna, Fragmentación del hábitat, Neotrópico, Panamá.

(Received: 26 II 02; Conditional acceptance: 12 III 02; Definitive acceptance: 9 IV 02) W. Douglas Robinson, Dept. of Fisheries and Wildlife, 104 Nash Hall, Oregon State Univ., Corvallis, OR 97331, U.S.A. E–mail: wdrobins@acesag.auburn.edu ISSN: 1578–665X

© 2001 Museu de Zoologia

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Introduction The dynamics of species communities in habitat patches have been a focus of interest among ecologists and conservation biologists (MACARTHUR & WILSON, 1967; SIMBERLOFF, 1995; LAURANCE & BIERREGAARD, 1997; TERBORGH et al., 2002). In particular, the process of species loss from human– created habitat patches, such as forest fragments, has been of interest because understanding that process may allow predictions of the size of reserves necessary to maintain viable populations (ZIMMERMAN & BIERREGAARD, 1986). A common pattern observed among recently isolated habitat patches is where an initially high species richness relaxes through time to some lower level of species richness (DIAMOND, 1972; SOULÉ et al., 1979). This process of species loss, also called faunal relaxation, apparently occurs most quickly immediately after isolation, but may continue to occur for an indeterminate amount of time (LOVEJOY et al., 1984), possibly until an equilibrium between extinction and immigration is reached (MACARTHUR & WILSON, 1967). Because of a lack of long–term data sets, however, efforts to understand the process of faunal relaxation have been impeded. Consequently, conservation biologists trying to preserve habitat remnants in regions threatened by further habitat destruction lack sufficient information from which to understand how the size of nature preserves may influence persistence of species over long time periods. Such problems are particularly acute in the tropics where species abundance patterns are very different from temperate sites; that is, a greater proportion of species in the tropics are rare, and thus may be more likely to be lost from habitat remnants by stochastic effects of population fluctuations (KARR , 1982a; L ANDE, 1987) or by environmental variation (LEIGH 1981).

Of particular interest to conservation biologists is predicting which species are at greatest risk of being lost from preserves so that steps might be taken to reduce risks of extinction. A key predictor of extinction risk is population size (TERBORGH & WINTER, 1980; DAVIES et al., 2000); species with low population sizes tend to have greater risks of extinction (PIMM et al., 1988; BELOVSKY et al., 1999). One useful method, therefore, for predicting the extinction risk of a population would be to assess temporal changes in population size within a conservation reserve.

Ideally, long–term monitoring of population sizes would allow conservation biologists to examine long temporal series of annual population size estimates and then to evaluate statistically the probability of extinction resulting from stochastic or deterministic factors. However, in most cases, such long–term data remain scarce or non– existent. In the Neotropics, for example, loss of bird species from isolated conservation reserves or forest fragments is common, with local extinctions often exceeding 35% of the species

originally present (KATTAN et al., 1994; CHRISTIANSEN & PITTER, 1997; STOUFFER & BIERREGAARD, 1995; ROBINSON, 1999), yet no temporal sequence of community–wide bird census data spanning more than five years is available. Several sites have been surveyed through use of mistnets for 15 to 30 years (KARR et al., 1990a; BRAWN et al., 1995; 1999; STOUFFER & BIERREGAARD, 1995), but such studies only effectively sample the understory of forest bird communities (K ARR , 1971, 1981).

Notwithstanding the limitations of presently available data, pressing conservation needs require use of extant data to help predict species that are sensitive to anthropogenic disturbances of Neotropical habitats, particularly forest fragmentation (BRAWN et al., 1998).

The objective of this study was to assess changes in population sizes of forest birds on Barro Colorado Island (BCI), Panama, over the last 25 years. Studies of the BCI avifauna provide the longest–running data set on the process of species loss from a tropical habitat fragment (ROBINSON, 1999). BCI is a 1,562 ha land–bridge island in Gatun Lake, which forms part of the Panama Canal. Originally isolated from the mainland in 1914 by creation of the Canal, BCI has been inventoried repeatedly by ornithologists since 1923. The island hosted as many as 200 resident species of the forest and forest edge (WILLIS & EISENMANN, 1979; KARR, 1982b; ROBINSON, 1999). Although the exact number of species lost is debated, positively known extinctions represent at least 35% of the species present at the first inventory in the 1920s.

Furthermore, the extinctions were not concentrated in the time period immediately after isolation; rather, they have occurred throughout the 80 years since the first inventories (CHAPMAN, 1929; WILLIS & EISENMANN, 1979; ROBINSON, 1999).

Although species inventories were conducted several times, only once were population sizes for most species on the island estimated. WILLIS (1980) spent the entire year of 1970 working on BCI and estimated population sizes for most species he encountered that year. An extensive series of point counts on BCI were conducted from 1994 to 1996 to generate population estimates for comparison with Willis’s, so that species experiencing large changes in population size could be identified, the tendency for rare species to disappear could be evaluated, and predictions could be made regarding which currently extant species might be most likely to disappear from BCI in the future.

Methods WILLIS (1980) generated island population estimates by studying intensively several species of understory birds on BCI from 1960 to 1971. By 1970, he knew all calls and songs of all species on the island and generated island–wide estimates

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BCI

Panama Lab

1 km

Fig. 1. Distribution of point count locations across Barro Colorado Island, Panama. Random (closed circles) and supplemental (open circles) points are all separated by at least 200 m: BCI. Barro Colorado Island.

Fig. 1. Distribución de los emplazamientos de recuento de la Isla de Barro Colorado, Panamá. Los puntos al azar (círculos negros) y los suplementarios (círculos blancos) distan entre sí un mínimo de 200 m: BCI. Isla de Barro Colorado.

of abundance for that year. Those estimates were generated from spending hundreds of hours in the field and gauging abundances of all species relative to the antfollowing species for which he knew abundances rather precisely because of following color–marked birds and mapping their territories (WILLIS, 1980). Thus, if certain species were encountered twice as often as a focal antfollower species which had an abundance of 200 pairs, Willis estimated the BCI population to be 400 pairs. Although this method is not repeatable and error can not be estimated, given the enormous time spent in the field to generate the numbers reasonable confidence can be had that the numbers are probably in the general vicinity of actual population sizes for most species.

An exhaustive inventory of bird species on BCI from 1994 to 1996 was conducted using three methods: point counts, line transects, and ad lib observations (R OBINSON , 1999). To generate estimates of population sizes, the point count data was used for several reasons. First, the location of points was randomized, improving my ability to extrapolate results from the subset of island space actually surveyed during point counts.

Sixty–four points were randomly distributed across the island so that no point was closer than 200 m to the next nearest point. In addition, the random points were supplemented with 65 other points distributed between the random points, sometimes along established foot paths and sometimes

between random points that were distributed off the trail system. To determine if the random and supplemental points differed in the number of individuals of each species detected, the mean number of individuals detected for each resident bird species at random and supplemental points were compared with ANOVA. In 7 of 199 resident species, number of detections differed (p < 0.05) between random and supplemental points. This is no different than expected by chance (p = 0.035) when so many tests are involved. Therefore, the results from random and supplemental points were combined and used all 129 to generate estimates of abundance (fig. 1).

Each point was visited for 8 minutes between January and July, the period of the year during which peak singing activity occurs (ROBINSON et al., 2000a). Eight–minute visits were used extensively during surveys on the nearby mainland (ROBINSON et al., 2000a) and are of sufficient duration that few new individuals are detected during the last two minutes of point counts. Points were conducted within four hours of dawn. Although some supplemental points were visited more than once, no random point was, so only data from the first visit to each point were used. During point counts, each bird seen or heard was identified and its distance and direction from the observer estimated. Distances were calibrated based on experience conducting more than 1,000 similar point counts throughout

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central Panama in the previous 2 years (ROBINSON et al., 1999, 2000a; CONDIT et al., 2001). Briefly, a series of point counts were conducted over a six– month period in 1993 and early 1994 where distances to vocalizing birds were paced off to ensure accuracy of distance estimation. Distance estimates were also frequently checked during line transect surveys conducted on BCI. Although accuracy of distance estimates diminishes with distance from the observer, distances to most species can be accurately judged within 200 m.

No birds detected more than 200 m from the observer were included in analyses.

Not all species can be heard as far as 200 m.

For example, several small insectivores have weak voices that carry no more than 30 to 50 m in tropical forest (e.g., Southern bentbill [see table 1 for scientific names] and Golden–crowned spadebill; ROBINSON et al., 2000a). Therefore, when calculating the area that was surveyed at each point count location, the maximum detection distance for each species was used as the radius of the circular area being inventoried. The maximum distance at which an individual of each species was noted was considered the maximum radius of the circular area for which surveys of that particular species could be considered effective.

To estimate island population size for each species, the area surveyed for each species was calculated separately because maximum detection distance varied among species. Thus, the area surveyed during point counts was calculated as the number of points surveyed times the area per point (o x [maximum detection distance]2).

Because points were distributed randomly, a proportional relationship between the number of birds detected at points and number of birds on the entire island was assumed. Therefore, to estimate abundances, the mean number of birds of a species detected per point was multiplied by the island area (1,562 ha) and then that product divided by the area surveyed at the 129 points. That quotient equals the mean number of birds on the island. An important assumption of this simple approach is that all species are equally detectable; in other words, the assumption is made that probability of detection if a bird is within detectable range of a point equals one. Given the wide range of behaviors among tropical birds, this is clearly not the case. Nearly 98% of detections during point counts in Panama are from auditory cues (ROBINSON et al., 2000a). Some species vocalize several times per minute during the morning, whereas others may vocalize much less regularly.

Thus, probability of detecting a species when it is within detectable range is not always high. To adjust for such detectability differences among species would involve a huge effort to assess how vocalization patterns of each species vary with respect to time of day, season, and breeding phenology. In a tropical bird community with

nearly 200 resident species, such data are not yet available. Therefore, I have taken the conservative approach in this study and assumed probability of detection is equal across species.

For most species, the approach will lead to an underestimate of total island population size.

Because extreme changes in abundance over the 25 years since Willis’ survey were assessed and not minor changes, evaluations of population declines should still be of interest. Any dramatic increases in population size should be of even greater interest because of the likelihood that the method employed here underestimates population sizes. A better method for analyzing temporal changes in abundance would be to institute a standardized point count scheme where particular points are surveyed each year in the same season by the same or comparable observers (V ERNER, 1985).

That design would be statistically more robust, would allow assessment of temporal trends within a species so that detectability issues would be minor, and would provide the most precisely repeatable protocol currently available for monitoring bird populations.

Residency status, preferred habitat, and ecological guild membership of bird species were categorized according to criteria established by ROBINSON et al. (2000a). Here, migratory species were excluded and only population estimates of year–round residents were compared.

Results and Discussion Population increases. Population estimates of fifteen species increased by 100% or more (table 2). Four species (Plain xenops, Checker– throated antwren, Dot–winged antwren and White–shouldered tanager) associate with one another in mixed–species foraging flocks. Two other species that forage with them also increased: Slaty antshrike increased by 40% and White– flanked antwren by 50%. Several canopy species (Paltry tyrannulet, Lesser greenlet and Blue dacnis) increased by 100–150%. Since 1970, the number of colonial icterids foraging on the island increased from 40 to 175 birds, presumably because of the locations of several nesting colonies in dead trees standing in lake coves around the island. Another lake margin species, Common tody–flycatcher, which builds its pendant nest from a branch tip hanging over water, increased from 6 to 100 individuals. Whether this increase is real or if lake margin habitats were under–surveyed by Willis is unclear. Crested guan populations tripled since 1970 as protection of BCI from hunters has increased in effectiveness over the last few decades. Guans have been hunted nearly to extirpation in mainland forests of the Canal watershed (ROBINSON et al., 2000a). Lastly, the increase in numbers of three hummingbird species is enigmatic.

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Animal Biodiversity and Conservation 24.2 (2001)

Table 1. Resident bird species detected on Barro Colorado Island, Panama, by WILLIS (1980) and ROBINSON (1999). Habitat affinities, guilds, and body masses are reported in addition to island– wide population size estimates. Species (names and sequence follow RIDGELY & GWYNNE (1989): nd. Not detected; ne. Abundance not estimated). Habitat (H): f. Forest; e. Edge or open habitats such as the island–lake interface, canopy edge, or aerial space above the island. Mass (M, in g): data from mistnet captures in Soberania National Park (ROBINSON et al. [1999b, 2000a]; STILES & SKUTCH [1989]; KARR et al. [1978, 1990]; WILLIS [1980]). Guilds (assignments are based on personal observations and data in KARR et al. (1990b): Aquat. Aquatic species found primarily along forest streams or, in the case of some vagrants, near larger bodies of water; Carr. Carrion consumers; FA. Arboreal frugivores; FAS. Sallying arboreal frugivores; FT. Terrestrial frugivores; GA. Arboreal granivores; GT. Terrestrial granivores; IADL. Arboreal insectivores that search primarily dead leaf clusters; IAer. Aerial insectivores (species that capture and consume insect while in flight); IAF. Army ant followers; IAG. Gleaning arboreal insectivores; IAS. Sallying arboreal insectivores; IBI. Insectivores that extract food from the interior of bark substrates (e.g., woodpeckers); IBS. Insectivores that glean food from the surface of bark (e.g., woodcreepers); ITG. Gleaning terrestrial insectivores (e.g., leaftossers); ITS. Sallying terrestrial insectivores (e.g., common pauraque); N. Nectarivores; most species also consume some small arthropods; OA. Arboreal omnivores; OAG. Gleaning arboreal monivores; OAS. Sallying arboreal omnivores; OT. Terrestrial omnivores; OTG. Gleaning terrestrial omnivores; RD. Raptors diurnal; RN. Raptors nocturnal.

Tabla 1. Especies de aves residentes detectadas en la Isla de Barro Colorado, Panamá, por WILLIS (1980) y ROBINSON (1999). Se informa de afinidades de hábitats, agrupaciones y masas corporales, además de estimaciones en cuanto al tamaño de la población presente en toda la isla. Las especies (nombres y orden según RIDGELY & GWYNNE (1989): nd. No detectadas; ne. Abundancia no estimada). Hábitat (H): f. Bosque; e. Orillas o hábitats abiertos, como la interfase entre el lago y la isla, el extremo de la bóveda o el espacio aéreo que cubre la isla. Masa (M, en g): datos obtenidos en capturas mediante redes en el Parque Nacional Soberanía (ROBINSON et al. [1999b, 2000a]; STILES & SKUTCH [1989]; KARR et al. [1978, 1990]; WILLIS [1980]). Agrupaciones (las asignaciones se basan en observaciones personales y datos de KARR et al. (1990b): Aquat. Especies acuáticas encontradas principalmente a lo largo de arroyos del bosque o, en el caso de algunas especies vagabundas, cerca de masas de agua de mayor tamaño; Carr.

Carroñeros; FA. Frugívoros arbóreos; FAS. Frugívoros arbóreos que cazan insectos al vuelo; FT. Frugívoros terrestres; GA. Granívoros arbóreos; GT. Granívoros terrestres; IADL. Insectívoros arbóreos que buscan principalmente acumulaciones de hojas muertas; IAer. Insectívoros aéreos (especies que capturan y consumen insectos durante el vuelo); IAF. Rastreadores de hormigas–ejército; IAG. Insectívoros arbóreos que rebuscan; IAS. Insectívoros arbóreos que cazan insectos al vuelo; IBI. Insectívoros que extraen alimento del interior del sustrato de las cortezas (por ej.: los pájaros carpinteros); IBS. Insectívoros que recogen alimentos de la superficie de las cortezas (por ej.: los pájaros trepadores); ITG. Insectívoros terrestres que rebuscan (por ej.: la hojarasca); ITS. Insectívoros terrestres que cazan insectos al vuelo (por ej.: los tapacaminos picuyos); N. Nectarívoros; la mayor parte de las especies también consumen pequeños artrópodos; OA. Omnívoros arbóreos; OAG. Omnívoros arbóreos que rebuscan; OAS. Omnívoros arbóreos que cazan insectos al vuelo; OT. Omnívoros terrestres; OTG. Omnívoros terrestres que rebuscan; RD. Aves de rapiña diurnas; RN. Aves de rapiña nocturnas.

Species

H

M

Robinson

Willis

Guild

Tinamidae Great tinamou – Tinamus major

f

1,160

100

200

FT / GT

Little tinamou – Crypturellus soui

e

250

2

nd

FT / GT

e

840

10

10

Aquat

Black vulture – Coragyps atratus

e

1,800

20

70

Carr

Turkey vulture – Cathartes aura

e

1,300

10

50

Carr

King vulture – Sarcoramphus papa

e

3,200

2

10

Carr

Ardeidae Rufescent tiger–heron – Tigrisoma lineatum Cathartidae

Robinson

56

Table 1. (Cont.)

Species

H

M

Robinson

Willis

Guild

Gray–headed kite – Leptodon cayenensis

f

500

4

5

RD

Hook–billed kite – Chondrohierax uncinatus

e

270

<1

10

RD

Double–toothed kite – Harpagus bidentatus

f

185

20

30

RD

Tiny hawk – Accipiter superciliosus

f

100

nd

1

RD

Crane hawk – Geranospiza caerulescens

f

377

<1

2

RD

Semiplumbeous hawk – Leucopternis semiplumbea

f

278

4

20

RD

White hawk – Leucopternis albicollis

f

736

2

20

RD

Short–tailed hawk – Buteo brachyurus

e

495

nd

1

RD

Zone–tailed hawk – Buteo albonotatus

e

565

nd

1

RD

Crested eagle – Morphnus guianensis

f

1,750

nd

2

RD

Black hawk–eagle – Spizaetus tyrannus

f

1,005

2

10

RD

Ornate hawk–eagle –Spizaetus ornatus

f

1,305

nd

5

RD

Collared forest–falcon – Micrastur semitorquatus

f

1,325

2

1

RD

Bat falcon – Falco rufigularis

e

150

1

nd

RD

Accipitridae

Falconidae

Cracidae Gray–headed chachalaca – Ortalis cinereiceps

e

536

nd

20

OAG

Crested guan – Penelope purpurascens

f

1,000

150

50

FA

Great currasow – Crax rubra

f

3,800

1

nd

FT / GT

f

405

20

20

IT

f

210

nd

5

Aquat

Rallidae Gray–necked wood–rail – Aramides cajanea Eurypygidae Sun bittern – Eurypyga helias Columbidae Pale–vented pigeon – Columba cayennensis

e

210

12

10

FA

Scaled pigeon – Columba speciosa

e

259

20

5

FA

Short–billed pigeon – Columba nigrirostris

e

160

100

300

FA

Blue ground–dove – Claravis pretiosa

e

69

2

nd

FA / GA

White–tipped dove – Leptotila verrauxi

e

130

2

6

FT

Gray–chested dove – Leptotila cassinnii

f

155

400

400

FT

Violaceous quail–dove – Geotrygon violacea

f

102

6

100

FT

Ruddy quail–dove – Geotrygon montana

f

128

100

200

FT

Orange–chinned parakeet – Brotogeris jugularis

e

63

60

400

FA / GA

Brown–hooded parrot – Pionopsitta haematosis

f

145

5

nd

GA

Blue–headed parrot – Pionus menstruus

e

235

50

150

GA

Red–lored amazon – Amazona autumnalis

f

416

150

200

GA

Mealy amazon – Amazona farinosa

f

687

150

250

GA

Squirrel cuckoo – Piaya cayana

f

105

150

300

IAG

Pheasant cuckoo – Dromococcyx phasianellus

f

86

nd

5

ITG

Greater ani – Crotophaga major

e

170

50

50

IADL

Smooth–billed ani – Crotophaga ani

e

110

nd

15

IAG

Psittacidae

Cuculidae

57

Animal Biodiversity and Conservation 24.2 (2001)

Table 1. (Cont.)

Species

H

M

Robinson

Willis

Guild

Vermiculated screech–owl – Otus guatemalae

f

100

ne

100

RN

Crested owl – Lophostrix cristata

f

510

nd

20

RN

Spectacled owl – Pulsatrix perspicillata

f

850

16

40

RN

Mottled owl – Ciccaba virgata

e

300

nd

30

RN

Black–and–white owl – Ciccaba nigrolineata

e

458

ne

10

RN

Striped owl – Rhinoptynx clamator

e

420

nd

1

RN

Short–tailed nighthawk – Lurocalis semitorquatus

e

75

5

5

IAer

Common pauraque – Nyctidromus albicollis

e

53

2

2

ITS

f

585

nd

10

IAS

White–collared swift – Streptoprocne zonaris

e

80

nd

1

IAer

Short–tailed swift – Chaetura brachyura

e

19

7

5

IAer

Band–rumped swift – Chaetura spinicauda

e

15

20

20

IAer

Lesser Swallow–tailed swift – Panyptila cayennensis

e

20

5

10

IAer

Long–tailed hermit – Phaethornis superciliosus

f

6

200

60

N

Little hermit – Phaethornis longuemareus

f

3

150

10

N

White–necked jacobin – Florisuga mellivora

e

6.3

100

60

N

Black–throated mango – Anthracothorax nigricollis

e

7

nd

5

N

Rufous–crested coquette – Lophornis delattrei

f

3

nd

5

N

Garden emerald – Chlorostilbon canivetii

e

4

nd

1

N

Crowned woodnymph – Thalurania colombica

f

5

200

100

N

Violet–bellied hummingbird – Damophila julie

f

4

100

100

N

Sapphire–throated hummingbird – Lepidopyga coeruleogularis e

4

1

nd

N

Strigidae

Caprimulgidae

Nyctibiidae Great potoo – Nyctibius grandis Apodidae

Trochilidae

Blue–chested hummingbird – Amazilia amabilis

f

3.3

100

150

N

Rufous–tailed hummingbird – Amazilia tzacatl

e

5

2

5

N

White–vented plumeleteer – Chalybura buffoni

f

6.0

5

nd

N

Purple–crowned fairy – Heliothryx barroti

e

6

5

50

N

Long–billed starthroat – Heliomaster longirostris

e

6

nd

1

N

White–tailed trogon – Trogon viridis

f

80

40

100

OA

Violaceous trogon – Trogon violaceus

f

57

60

100

OA

Black–throated trogon – Trogon rufus

f

52

250

300

OA

Black–tailed trogon – Trogon melanurus

f

115

2

6

OA

Slaty–tailed trogon – Trogon massena

f

140

175

200

OA

Blue–crowned motmot – Momotus momota

e

105

ne

nd

OA

Rufous motmot – Baryphthengus martii

f

162

300

200

OA

Broad–billed motmot – Electron platyrhynchum

f

62

100

100

OA

e

290

2

10

Aquat

Trogonidae

Momotidae

Alcedinidae Ringed kingfisher – Ceryle torquata

Robinson

58

Table 1. (Cont.)

Species

H

M

Robinson

Willis

Guild

Green kingfisher – Chloroceryle americana

e

25

Amazon kingfisher – Chloroceryle amazona

e

130

nd

30

Aquat

nd

4

Aquat

Green–and–rufous kingfisher – Chloroceryle inda

f

59

American pygmy kingfisher –Chloroceryle aenea

f

16

nd

2

Aquat

5

20

Aquat

Black–breasted puffbird – Notharchus pectoralis

f

68

150

600

IAS

Pied puffbird – Notharchus tectus

f

33

25

50

IAS

White–whiskered puffbird – Malacoptila panamensis

f

44

300

200

IAS

Collared aracari – Pteroglossus torquatus

f

65

125

300

FA

Keel–billed toucan – Ramphastos sulfuratus

f

375

200

300

FA

Chestnut–mandibled toucan – Ramphastos swainsoni

f

750

150

300

FA

Black–cheeked woodpecker – Melanerpes pucherani

f

54

150

150

IBI

Cinnamon woodpecker – Celeus loricatus

f

74

2

nd

IBI

Lineated woodpecker – Dryocopus lineatus

e

180

24

20

IBI

Crimson–crested woodpecker – Campephilus melanoleucos f

225

30

50

IBI

Bucconidae

Ramphastidae

Picidae

Furnariidae Plain xenops – Xenops minutus

f

11

800

400

IBS

Scaly–throated leaftosser – Sclerurus guatemalensis

f

34

100

150

ITG

Plain–brown woodcreeper – Dendrocincla fuliginosa

f

41

90

100

IAF / IBS

Wedge–billed woodcreeper – Glyphorynchus spirurus

f

15

500

800

IBS

Buff–throated woodcreeper – Xiphorhynchus guttatus

f

47

175

250

IBS

Black–striped woodcreeper – Xiphorhynchus lachrymosus

f

51

200

500

IBS

Fasciated antshrike – Cymbilaimus lineatus

f

37

nd

5

IAG

Slaty antshrike – Thamnophilus punctatus

f

22

3,500

2,500

IAG

Spot–crowned antvireo – Dysithamnus puncticeps

f

15

175

200

IAG

Checker–throated antwren – Myrmotherula fulviventris

f

10

4,000

1,500

IADL

White–flanked antwren – Myrmotherula axillaris

f

8

3,000

2,000

IAG

Dot–winged antwren – Microrhopias quixensis

f

8

4,000

1,000

IAG

Dusky antbird – Cercomacra tyrannina

e

17

20

20

IADL / IAG

White–bellied antbird – Myrmeciza longipes

e

28

nd

5

ITG

Chestnut–backed antbird – Myrmeciza exsul

f

27

1,500

1,050

IAG / ITG

Spotted antbird – Hylophylax naevioides

f

17

750

700

IAF / ITG

Bicolored antbird – Gymnopithys leucaspis

f

30

60

60

IAF

Ocellated antbird – Phaenostictus mcleannani

f

51

nd

6

IAF

Streak–chested antpitta – Hylopezus perspicillata

f

42

nd

4

ITG

Paltry tyrannulet – Tyranniscus vilissimus

e

9

2,500

1,000

OAS

Brown–capped tyrannulet – Ornithion bruneicapillum

f

7

250

600

IAS

Southern beardless–tyrannulet – Camptostoma obsoletum e

8.5

6

150

IAS

8

700

400

OAS

Dendroccolaptidae

Formicariidae

Tyrannidae

Yellow–crowned tyrannulet – Tyrannulus elatus

e

59

Animal Biodiversity and Conservation 24.2 (2001)

Table 1. (Cont.)

Species

H

M

Robinson

Willis

Guild

Forest elaenia – Myiopagis gaimardii

f

14

90

80

IAS

Yellow–bellied elaenia – Elaenia flavogaster

e

25

nd

1

IAS

Ochre–bellied flycatcher – Mionectes oleaginea

f

13

100

100

OAS

Black–capped pygmy–tyrant – Myiornis atricapillus

f

5

125

300

IAS

Southern bentbill – Oncostoma olivaceum

f

7

800

400

IAS

Common tody–flycatcher – Todirostrum cinereum

e

7

100

6

IAS

Olivaceous flatbill – Rhynchocyclus olivaceus

f

22

200

300

IAS

Yellow–margined flycatcher – Tolmomyias assimilis

f

14

400

500

IAS

Golden–crowned spadebill – Platyrinchus coronatus

f

9

200

200

IAS

Ruddy–tailed flycatcher – Terenotriccus erythrurus

f

7

400

400

IAS

Bright–rumped attila – Attila spadiceus

f

38

25

50

IAS

Speckled mourner – Laniocera rufescens

f

49

4

30

OAS

Rufous mourner – Rhytipterna holerythra

f

38

50

70

OAS

Dusky–capped flycatcher – Myiarchus tuberculifer

f

20

1,000

1,000

IAS

Panama flycatcher – Myiarchus panamensis

e

33

8

nd

IAS

Lesser kiskadee – Philohydor lictor

e

25

100

100

IAS

Great kiskadee – Pitangus sulphuratus

e

50

50

nd

IAS

Boat–billed flycatcher – Megarhyncus pitangua

e

60

100

150

IAS

Rusty–margined flycatcher – Myiozetetes cayanensis

e

28

75

70

IAS

Social flycatcher – Myiozetetes similis

e

24

300

600

IAS

Streaked flycatcher – Myiodynastes maculatus

e

45

20

50

IAS

Piratic flycatcher – Legatus leucophaius

e

26

10

2

OAS

Tropical kingbird – Tyrannus melancholicus

e

40

400

500

IAS

White–winged becard – Pachyramphus polychopterus

e

18

nd

ne

IAS

Masked tityra – Tityra semifasciata

e

80

40

300

OAS

Black–crowned tityra – Tityra inquisitor

e

41

15

50

OAS

Rufous piha – Lipaugus unirufus

f

86

1

80

OAS

Blue cotinga – Cotinga nattererii

f

55

50

80

FA

Purple–throated fruitcrow – Querula purpurata

f

104

180

250

OAS

Thrushlike mourner – Schiffornis turdinus

f

33

2

nd

OAS

Golden–collared manakin – Manacus vitellinus

e

17

50

150

FAS

Red–capped manakin – Pipra mentalis

f

15

800

1,000

FAS

Gray–breasted martin – Progne chalybea

e

39

25

10

IAer

Mangrove swallow – Tachycineta albonotata

e

15

100

ne

IAer

Southern rough–winged swallow – Stelgidopteryx serripennis e

16

ne

10

IAer

f

220

nd

1

OAG

e

18

2

20

IAG

f

10

30

100

IAG

Cotingidae

Pipridae

Hirundinidae

Corvidae Black–chested jay – Cyanocorax affinis Troglodytidae Plain wren – Thryothorus modestus Sylvinae Long–billed gnatwren – Ramphocaenus rufiventris

Robinson

60

Table 1. (Cont.)

Species

H

M

Robinson Willis

Guild

Tropical gnatcatcher – Polioptila plumbea

f

7

300

200

IAS

e

80

1

nd

OTG

Yellow–green vireo – Vireo flavoviridis

e

17

nd

5

OAG

Scrub greenlet – Hylophilus flavipes

e

13

nd

5

IAG

Lesser greenlet – Hylophilus decurtatus

f

10

3,000

1,500

IAG

Green shrike–vireo – Vireolanius pulchellus

f

25

1

10

IAG

e

9

100

300

OAG

Plain–colored tanager – Tangara inornata

e

18

100

200

OAG

Bay–headed tanager – Tangara gyrola

e

22

ne

nd

OAG

Golden–hooded tanager – Tangara larvata

e

19

50

100

OAG

Scarlet–thighed dacnis – Dacnis venusta

e

16

5

40

OAG

Blue dacnis – Dacnis cayana

e

13

1,000

500

OAG

Green honeycreeper – Chlorophanes spiza

e

18

200

400

OAG

Shining honeycreeper – Cyanerpes lucidus

e

12

75

50

FAG

Red–legged honeycreeper – Cyanerpes cyaneus

e

13

200

300

FAG

Yellow–crowned euphonia – Euphonia luteicapilla

e

12

1

nd

FAG

Thick–billed euphonia – Euphonia laniirostris

e

15

2

nd

FAG

Fulvous–vented euphonia – Euphonia fulvicrissa

f

11

250

1,000

OAG

White–vented euphonia – Euphonia minuta

e

11

nd

2

FAG

Blue–gray tanager – Thraupis episcopus

e

30

20

150

OAG

Palm tanager – Thraupis palmarum

e

35

10

100

OAG

Gray–headed tanager – Eucometis penicillata

f

30

40

ne

OAF / OAG

White–shouldered tanager – Tachyphonus luctuosus

e

15

300

100

OAG / OAS

Red–throated ant–tanager – Habia fuscicauda

e

39

50

40

OAG

Crimson–backed tanager – Ramphocelus dimidiatus

e

28

6

nd

OAG

Slate–colored grosbeak – Pitylus grossus

f

43

4

500

OAG

Blue–black grosbeak – Cyanocompsa cyanoides

f

32

80

300

OAG

Orange–billed sparrow – Arremon aurantiirostris

e

31

ne

nd

OTG

Variable seedeater – Sporophila aurita

e

10

2

30

GT

Yellow–bellied seedeater – Sporophila nigricollis

e

10

ne

2

GT

Giant cowbird – Scaphidura oryzivora

e

187

4

nd

OTG

Yellow–backed oriole – Icterus chrysater

e

43

40

40

OAG

Yellow–rumped cacique – Cacicus cela

e

68–113

100

20

OAG

Chestnut–headed oropendola – Zarhynchus wagleri

e

113–214

75

20

OAG

Turdinae Clay–colored thrush – Turdus grayii Vireonidae

Coerebinae Bananaquit – Coereba flaveola Thraupinae

Cardinalinae

Emberizinae

Icterinae

61

Animal Biodiversity and Conservation 24.2 (2001)

Table 2. Species for which island population sizes increased at least 100% between 1970 and the mid–1990s. (See table 1 for scientific names.

Table 3 Species, or species groups, for which island population sizes declined at least 50% between 1970 and the mid–1990s.

(See table 1 for scientific names.)

Tabla 2. Especies cuya población en la isla aumentó al menos en un 100% desde 1970 hasta mediados de la década de 1990. (Para los nombres científicos, véase la tabla 1.)

Tabla 3. Especies, o grupos de especies, cuya población en la isla disminuyó al menos en un 50% desde 1970 hasta mediados de 1990.

(Para los nombres científicos, véase la tabla 1.)

Species

Species

Crested guan Long–tailed hermit Little hermit

1970

1995 %

1970

1995

%

50

150

200

Great tinamou

200

100

50

233

Hawks

108

<37

66

Vultures

130

32

75

Short–billed pigeon

300

100

67

Violaceous quail–dove

100

6

94

Ruddy quail–dove

200

100

50

60 10

200

150 1,400

Crowned woodnymph

100

200

100

Plain xenops

400

800

100

Checker–throated antwren

1,500

4,000

167

Orange–chinned parakeet

400

60

85

Dot–winged antwren

1,000

4,000

300

Blue–headed parrot

150

50

67

Paltry tyrannulet

1,000

2,500

150

Squirrel cuckoo

100

Owls

Southern bentbill Common tody–flycatcher Lesser greenlet

400 6

800

100 1,567

300

150

50

>200

>16

92

Purple–crowned fairy

50

5

90

White–tailed trogon

100

40

60

6

2

67

66

7

89

600

150

75

50

25

50

1,500

3,000

100

Blue dacnis

500

1,000

100

Kingfishers

White–shouldered Tanager

100

300

200

Black–breasted puffbird

20

100

400

Pied Puffbird

275

Black–striped woodcreeper

500

200

60

Brown-capped tyrannulet

600

250

58

Southern beardless tyrannulet

150

6

96

Black–capped pygmy–tyrant

300

125

58

Bright–rumped attila

50

25

50

Speckled mourner

30

4

87

600

300

50

Yellow–rumped cacique Chestnut–headed oropendola

20

75

Population declines. Thirty–seven species and four species groups (hawks, vultures, owls, and kingfishers) declined by at least 50% (table 3).

Those species represent an ecologically diverse subset of the island avifauna. Several major patterns are apparent.

First, consumers of vertebrates declined as a whole. Hawks declined by at least two–thirds; owls by 92%; carrion–eating vultures by 75%; and piscivorous kingfishers by 89%. Despite extensive pre–dawn surveys (ROBINSON, 1999), extremely few owls were detected on the island, with only 3 of the 5 forest species Willis detected being found by me in the mid–1990s. A single Striped owl was observed by Willis on a small satellite island near the BCI laboratory facilities, which was dominated by grass and shrubs at that time. By the mid–1990s the habitat had matured and was unsuitable for that species. Similarly, 5 species of hawk found by Willis were not found by me, although Double–toothed kite continued to be the most numerous raptor. The paucity of hawks and owls is curious. The lack of kingfishers has been explained by the introduction of Peacock

Black–throated trogon

Social flycatcher Streaked flycatcher

50

20

60

300

40

87

Black–crowned tityra

50

15

70

Rufous piha

80

1

99

150

50

67

Masked tityra

Golden–collared manakin Plain wren Long–billed gnatwren Green shrike–vireo

20

2

90

100

30

70

10

1

90

Bananaquit

300

100

67

Plain–colored tanager

200

100

50

Scarlet–thighed dacnis

40

5

90

Green honeycreeper Fulvous–vented euphonia

400

200

50

1,000

250

75

Blue–gray tanager

150

20

87

Palm tanager

100

10

90

Slate–colored grosbeak

500

4

99

Blue–black grosbeak

300

80

73

30

2

93

Variable seedeater

62

bass (Cichla sp.) into Gatun lake, which are predaceous on smaller, native fishes (ZARET & PAINE, 1973). The hypothesis suggests that as bass populations rose, native fishes were driven locally extinct and the food resource base of kingfishers was reduced. However, strong tests of the hypothesis have not been conducted.

Second, 11 species are usually commonest in edge or young forest habitats: Orange–chinned parakeet, Southern beardless–tyrannulet, Golden–collared manakin, Plain wren, Long– billed gnatwren, Plain–colored tanager, Fulvous– vented euphonia, Blue–gray tanager, Palm tanager, Blue–black grosbeak, and Variable seedeater. Nearly all those species are now most frequently observed near the laboratory clearing or along the island–lake interface and go virtually undetected within the interior of the island.

Continued successional maturation of the forest probably explains the declines of those species.

Third, several species commonly nest in standing trees in Gatun lake. Such trees were numerous after lake waters rose to isolate the island in the early 1900s and have steadily disappeared as they have rotted and fallen into the lake. Social flycatchers commonly build their bulky nests in the branches of such trees, whereas the tityras and Streaked flycatcher usually occupy cavities in dead trees.

Fourth, several frugivorous or granivorous species are highly mobile and numbers on the island could fluctuate from year to year depending on fruit availability so much that comparisons of two point estimates may not indicate long–term declines in numbers. Short– billed pigeon, Ruddy quail–dove, Blue–headed parrot, and Green honeycreeper all move easily across long distances and may track resource abundance (WRIGHT, 1985). Violaceous quail–dove has declined the most precipitously, by 94%, but this decline may reflect a real regional decline.

Nowhere in central Panama can one now find population densities like the 100 individuals Willis estimated on BCI in 1970. Instead, the species is rare and sparsely distributed (Robinson, unpublished data).

Fifth, including Violaceous quail–dove, six species of forest interior birds have declined so dramatically that they are on the brink of extinction from BCI. Speckled mourner, which is a persistent singer and is easy to detect has dropped to four individuals on BCI. Rufous piha, a loud and formerly common species with an estimated 80 birds in 1970, was down to one individual in 1994; none have been seen or heard since then.

Green shrike–vireo, also a very loud and persistent singer that vocalizes all day long, was never common, but is now represented by one singing male; it is unknown if a female accompanied the male. Scarlet–thighed dacnis was rarely found in the mid–1990s and its total island population size was estimated at five individuals. That species remains high in the canopy with other tanager

Robinson

and honeycreeper species, however, and could have been underestimated. Slate–colored grosbeak, however, has a very distinctive song uttered at regular intervals as it forages in the mid–story of tall forest. Its population collapsed from 500 to four (two pairs).

Finally, in contrast to increases in numbers of guans, Great tinamous, another species often hunted in Panama, declined on BCI by about 50%. Given the level of protection from hunting, an increase would be expected; however, it is possible that this ground–nesting species has experienced elevated levels of nest predation.

Tinamou nests have low survival rates on the mainland (ROBINSON et al., 2000b) and their eggs are consumed by snakes and monkeys (ROBINSON et al., 2001). BCI has somewhat higher densities of monkeys than nearby mainland forests (WRIGHT et al., 1994), suggesting that predation on tinamou nests might be greater. A handful of canopy species, such as Black–breasted and Pied puffbirds, Black–striped woodcreeper, Brown–capped tyrannulet and Black–capped pygmy–tyrant have declined for unknown reasons.

Extinctions and colonizations. Nineteen species were detected on BCI during the mid–1990s but not found by Willis in 1970 (table 4). Great curassow, Brown–hooded parrot, Cinnamon woodpecker and Thrush–like mourner are all forest–dwelling species, whereas the remaining species all prefer edge habitats. ROBINSON (1999) provides details on these colonizations.

Twenty–seven species were found by Willis but not the author of this study. ROBINSON (1999) discusses possible reasons for the loss of those species. Rarity, however, appears to be strongly related to extinction probability. The maximum estimated abundance of any species lost since 1970 was 30, which was Green kingfisher, a species that utilizes only the margins of the island. Island–wide abundance for 21 of the 27 species not found in the mid–1990s was 5 or fewer individuals. Although KARR (1982b) argued that rarity is not a good predictor of extinction probability among birds on BCI, his conclusions were based on comparisons of mainland and BCI species lists built from general impressions of abundance on the mainland only; no abundance estimates derived from censuses, on either the mainland or island, were used. Evidence supporting the hypothesis that rarity increases extinction risk has come from the British island avifauna, where population size was the most important predictor of risk of extinction (PIMM et al., 1988).

If rarity is defined as any species whose population size is less than 10, then 26 of the rare species present in the mid–1990s were species of edge habitats (table 1). Losses of those species could occur naturally as forest maturation continues, but many may not disappear permanently from the island. As ROBINSON (1999) indicated, edge species disperse well and

63

Animal Biodiversity and Conservation 24.2 (2001)

Table 4. Resident species not detected in both 1970 and the mid–1990s. Those undetected in 1970 probably colonized by the mid-1990s and those detected in 1970 but not found in the mid–1990s had probably disappeared from Barro Colorado Island: A. Present in 1970, absent in 1995; B. Absent in 1970, present in 1995. (See table 1 for scientific names, habitat affiliations, and abundance estimates.)

Tabla 4. Especies residentes no detectadas, ni en 1970 ni a mediados de la década de 1990.

Las especies que no se detectaron en 1970 probablemente se establecieron en colonias a mediados de la década de 1990, y aquellas que se detectaron en 1970, pero que no pudieron encontrarse a mediados de la década de 1990, probablemente habían desaparecido de la Isla de Barro Colorado: A. Presentes en 1970, ausentes en 1995; B. Ausentes en 1970, presentes en 1995. (Para los nombres científicos, afiliaciones de hábitat y abundancias estimadas, véase la tabla 1.)

A

B

Tiny hawk

Little tinamou

Short–tailed hawk

Bat falcon

Zone–tailed hawk

Great curassow

Crested eagle

Brown–hooded parrot

Ornate hawk–eagle

Sapphire–throated hummingbird

Gray–headed chachalaca Long–billed starthroat Sunbittern

White–vented plumeleteer

Pheasant cuckoo

Blue–crowned motmot

Smooth–billed ani

Cinnamon woodpecker

Great potoo

Panama flycatcher

White–collared swift

Great kiskadee

Black–throated mango

Thrush–like mourner

Rufous–crested coquette Clay–colored robin Garden emerald

Bay–headed tanager

Green kingfisher

Yellow–crowned euphonia

Amazon eingfisher

Thick–billed euphonia

Green–and–rufous kingfisher Crimson–backed tanager Fasciated antshrike

Orange–billed sparrow

White–bellied antbird

Giant cowbird

Ocellated antbird Streak–chested antpitta Yellow–bellied elaenia White–winged becard

recolonize when appropriate habitat becomes available. Thus, their disappearances from BCI, despite local rarity, may typically be temporary.

In contrast, rare species of the forest may be absent for long time intervals after local extinction from BCI. Many forest species appear to disperse poorly and have difficulties recolonizing isolated forest remnants (WILLIS, 1974; ROBINSON, 1999). ROBINSON (1999) showed that most forest species that had disappeared from BCI were never encountered on the island again. Thus, the 11 forest species currently present in very low numbers may disappear from BCI permanently once they become locally extinct.

That total excludes six forest raptor species whose abundances are less than 10, but whose home ranges are so large that they must forage off the island as well.

If the remaining forest species are common enough that they are relatively unlikely to disappear over the next 25 years, the rate at which forest species are being lost from BCI may be slowing. BCI may now be acting as a reserve for common species and failing to preserve many rare species that comprise a significant proportion of the species in the bird community of central Panama. Thus, with continued loss of bird species from the island, BCI will be filled with common and widely distributed species and, like other small tropical reserves (DIAMOND et al., 1987), will not act as an effective preserve of regional avian diversity.

Future directions No method of surveying tropical birds is perfect or complete. The great diversity and variety of life histories causes some species to be much more easily detected than others so that complete community surveys must involve multiple methods (TERBORGH et al., 1990; ROBINSON et al., 2000a). However, for many species, a simple point count scheme, where points are surveyed annually by the same or comparable observers at the same season, will provide the best information for detecting long–term population trends with a minimum of bias (VERNER, 1985).

Such schemes are now possible in some Neotropical locations such as Panama where knowledge of bird vocalizations is reasonably complete. With the accumulation of such information, conservation biologists will be better able to predict accurately the long–term effects of habitat fragmentation on bird communities, particularly the likelihood of extinction as a function of population fluctuations.

Black–chested jay Yellow–green vireo

Acknowledgements

Scrub greenlet White–vented euphonia

The Smithsonian Tropical Research Institute supported the study with a pre–doctoral

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fellowship to W. D. Robinson. Suggestions for improvements to the study were provided by S. J. Wright, N. G. Smith, and E. G. Leigh. The manuscript was improved by comments from E. G. Leigh, G. Rompré, R. Moore, and T. Robinson.

ANAM kindly allowed me to study birds in Panama.

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"La tortue greque" Oeuvres du Comte de Lacépède comprenant L'Histoire Naturelle des Quadrupèdes Ovipares, des Serpents, des Poissons et des Cétacés; Nouvelle édition avec planches coloriées dirigée par M. A. G. Desmarest; Bruxelles: Th. Lejeuné, Éditeur des oeuvres de Buffon, 1836. Pl. 7

Editor executiu / Editor ejecutivo / Executive Editor Joan Carles Senar

Secretaria de Redacció / Secretaría de Redacción / Editorial Office

Secretària de Redacció / Secretaria de Redacción / Managing Editor Montserrat Ferrer

Museu de Zoologia Passeig Picasso s/n 08003 Barcelona, Spain Tel. +34–93–3196912 Fax +34–93–3104999 E–mail mzbpubli@intercom.es

Consell Assessor / Consejo asesor / Advisory Board Oleguer Escolà Eulàlia Garcia Anna Omedes Josep Piqué Francesc Uribe

Editors / Editores / Editors Antonio Barbadilla Univ. Autònoma de Barcelona, Bellaterra, Spain Xavier Bellés Centre d' Investigació i Desenvolupament CSIC, Barcelona, Spain Juan Carranza Univ. de Extremadura, Cáceres, Spain Luís Mª Carrascal Museo Nacional de Ciencias Naturales CSIC, Madrid, Spain Adolfo Cordero Univ. de Vigo, Vigo, Spain Mario Díaz Univ. de Castilla–La Mancha, Toledo, Spain Xavier Domingo Univ. Pompeu Fabra, Barcelona, Spain Francisco Palomares Estación Biológica de Doñana, Sevilla, Spain Francesc Piferrer Inst. de Ciències del Mar CSIC, Barcelona, Spain Ignacio Ribera The Natural History Museum, London, United Kingdom Alfredo Salvador Museo Nacional de Ciencias Naturales, Madrid, Spain José Luís Tellería Univ. Complutense de Madrid, Madrid, Spain Francesc Uribe Museu de Zoologia de Barcelona, Barcelona, Spain

Consell Editor / Consejo editor / Editorial Board José A. Barrientos Univ. Autònoma de Barcelona, Bellaterra, Spain Jean C. Beaucournu Univ. de Rennes, Rennes, France David M. Bird McGill Univ., Québec, Canada Mats Björklund Uppsala Univ., Uppsala, Sweden Jean Bouillon Univ. Libre de Bruxelles, Brussels, Belgium Miguel Delibes Estación Biológica de Doñana CSIC, Sevilla, Spain Dario J. Díaz Cosín Univ. Complutense de Madrid, Madrid, Spain Alain Dubois Museum national d’Histoire naturelle CNRS, Paris, France John Fa Durrell Wildlife Conservation Trust, Trinity, United Kingdom Marco Festa–Bianchet Univ. de Sherbrooke, Québec, Canada Rosa Flos Univ. Politècnica de Catalunya, Barcelona, Spain Josep Mª Gili Inst. de Ciències del Mar CMIMA–CSIC, Barcelona, Spain Edmund Gittenberger Rijksmuseum van Natuurlijke Historie, Leiden, The Netherlands Fernando Hiraldo Estación Biológica de Doñana CSIC, Sevilla, Spain Patrick Lavelle Inst. Français de recherche scient. pour le develop. en cooperation, Bondy, France Santiago Mas–Coma Univ. de Valencia, Valencia, Spain Joaquín Mateu Estación Experimental de Zonas Áridas CSIC, Almería, Spain Neil Metcalfe Univ. of Glasgow, Glasgow, United Kingdom Jacint Nadal Univ. de Barcelona, Barcelona, Spain Stewart B. Peck Carleton Univ., Ottawa, Canada Eduard Petitpierre Univ. de les Illes Balears, Palma de Mallorca, Spain Taylor H. Ricketts Stanford Univ., Stanford, USA Joandomènec Ros Univ. de Barcelona, Barcelona, Spain Valentín Sans–Coma Univ. de Málaga, Málaga, Spain Tore Slagsvold Univ. of Oslo, Oslo, Norway

Animal Biodiversity and Conservation 24.1, 2001 © 2001 Museu de Zoologia, Institut de Cultura, Ajuntament de Barcelona Autoedició: Montserrat Ferrer Fotomecànica i impressió: Sociedad Cooperativa Librería General ISSN: 1578–665X Dipòsit legal: B–16.278–58

Animal Biodiversity and Conservation 24.2 (2001)

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Passerine bird communities of Iberian dehesas: a review J. L. Tellería

Tellería, J. L., 2001. Passerine bird communities of Iberian dehesas: a review. Animal Biodiversity and Conservation, 24.2: 67–78.

Abstract Passerine bird communities of Iberian dehesas: a review.— The Iberian dehesas are a man–made habitat composed of scattered oaks (Quercus spp.) and extensive grass cover occupying three million ha in south– western Iberia. This paper compares the structure of the passerine bird communities in this region with other bird assemblages of Iberian woodlands. Although forest bird numbers in the southern half of the Iberian peninsula are decreasing, the dehesas show the highest richness in breeding birds, seemingly as the result of the increased presence of border and open–habitat birds. A low intra–habitat turnover of species was observed in the dehesas, with birds recorded at a sampling point accounting for a high percentage of the total richness of the community. This can be related to the low spatial patchiness of this habitat. In winter, the dehesas continued to maintain many bird species, but showed bird densities similar to other woodlands. This pattern, as well as the scarcity of some common forest passerines during the breeding period, could result from the removal of the shrub layer typical of Mediterranean woodlands.

Key words: Dehesas, Forest bird communities, Iberian peninsula, Passerines, Seasonal changes, Species richness.

Resumen Las comunidades de pájaros de las dehesas ibéricas: una revisión.— En este trabajo se revisa la composición y la estructura de las comunidades de aves de las dehesas ibéricas. Estos pastizales arbolados, cubiertos de encinas dispersas, ocupan unos tres millones de hectáreas del cuadrante suroccidental ibérico. Pese a la disminución de las aves forestales hacia el sur ibérico, las dehesas presentan un número elevado de especies si se las compara con otros bosques del área. Este rasgo parece ser consecuencia de la abundancia de aves típicas de medios abiertos o ecotónicos. La homogeneidad espacial de la estructura de las dehesas da lugar, además, a una baja tasa de cambio en la riqueza de especies dentro de este hábitat, de forma que las aves vistas en una estación de censo son un porcentaje elevado del total de la comunidad. En invierno, las dehesas siguen siendo un hábitat diverso, aunque presentan densidades similares a otros bosques del suroeste ibérico. Esta relativa escasez de aves pudiera deberse a la falta de arbustos fruticosos típicos del Mediterráneo sobre los que se alimentan muchas aves migradoras.

Palabras clave: Dehesas, Comunidades de aves forestales, Península ibérica, Paserinos, Cambios estacionales, Riqueza específica.

(Received: 15 XI 01; Definitive acceptance: 8 I 02) José Luis Tellería, Dept. of Animal Biology (Zoology), Fac. of Biology, Univ. Complutense, 28040 Madrid, España (Spain).

E–mail: telleria@bio.ucm.es

ISSN: 1578–665X

© 2001 Museu de Zoologia

Tellería

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Introduction The Spanish dehesa, and its Portuguese counterpart the “montado”, is an extensive agroecosystem situated in the south–western quarter of the Iberian peninsula and occupying around three millions ha (DÍAZ et al., 1997). It is composed of scattered oaks (e.g. holm oaks Quercus ilex and cork oaks Quercus suber) resulting from the clearing of former forests and the subsequent grazing and ploughing to prevent bush development and to maintain extensive grass cover.

Grasses and acorns produced by these wooded pastures thereby sustain herds of cattle, sheep and pigs on large private estates (mean size approx. 500 ha; CAMPOS,1993). The landscape resulting from such forest management resembles an African savannah dotted with small pools that provide water for livestock during summer.

The dehesas are important for wildlife conservation because, despite their own biodiversity resources (DÍAZ et al., 1997; TUCKER & EVANS, 1997), they form a wooded matrix interspersed with hills and undisturbed patches of Mediterranean forests and shrublands. These habitat patches maintain several endangered species (Iberian Lynx Lynx pardina, Imperial Eagle Aquila adalberti, Black Stork Ciconia nigra, etc.) which, like other animals (raptors, carnivores, wild boars, deer, etc.), move between forests and dehesas (RODRIGUEZ & DELIBES, 1990; GONZÁLEZ, 1991). In addition, dehesas are a main wintering habitat for several migratory birds, some of concern for conservation (e.g. wood pigeons Columba palumbus, cranes Grus grus; PURROY, 1988; ALONSO & ALONSO, 1990). As the dehesas prove compatible with birds and other wildlife, a significant percentage of this habitat has been included in areas of national and international conservation strategies (TUCKER & EVANS, 1997; VIADA, 1999).

The bird fauna of Iberian dehesas has been studied from different perspectives. The study of HERRERA (1978a, 1980) on the structure and seasonal evolution of bird communities in two dehesas was the first and most complete approach to this issue to date. Other approaches have been focussed on describing the composition of bird communities (PERIS, 1991; SÁNCHEZ 1991), the effects of management strategies on birds (PASCUAL et al., 1991; CABELLO DE ALBA, 1992; DÍAZ & PULIDO, 1995, PULIDO & DÍAZ, 1992, 1997; DÍAZ et al., 2002), or different issues on the biology of individual bird species (e.g. LÓPEZ–GORDO et al., 1976, HERRERA, 1977, 1978b; ALONSO et al., 1991; DÍAZ & PULIDO, 1993; DÍAZ et al., 1996; DÍAZ & MARTÍN, 1998).

However, bird communities in the dehesas have not been compared with those of other Iberian woodlands even though such an approach is adequate to disclose the specific features of bird assemblages (WIENS, 1989). This paper reviews the available information on the structure of dehesa bird communities to interpret their traits in the light of current hypotheses on the factors affecting

the distribution of Iberian forest birds. To do so, specific traits of the dehesa bird assemblages are compared with the bird communities of other Iberian woodlands. More specifically, the study analyses: (i) the main features of the Iberian pool of forest birds from which the dehesa bird communities have been assembled; (ii) the effects of habitat structure of these wooded pastures on their species richness; and (iii) the way seasonal changes affect the structure of their bird assemblages.

Material and methods Two data sets resulting from former studies on the structure of Iberian bird communities have been used to evaluate some relevant features of the dehesa bird assemblages (fig. 1). They refer to forest passerines (crows excluded) provided that these species are abundant and potentially ubiquitous along the Iberian gradient and homogeneous from a methodological point of view (censuses). The first data set records the results of point counts used to assess the effects of geographical location, climate and tree density on the structure of breeding bird communities (see TELLERÍA et al., 1992, 1999; TELLERÍA & SANTOS, 1993 for further details). Point counts were used to record the number of species observed over 10– minute long periods in the early morning at 20 sampling stations randomly distributed across each study woodland. These data are useful to obtain richness scores per sampling point (point diversity or internal alpha diversity) and the accumulative total richness (alpha or within–habitat diversity) of the studied woodlands (WHITTAKER, 1977 for further details). In addition, the mean point richness / total richness ratio can be used as an index of the intra–habitat turnover of species (internal beta or pattern diversity). A second data set refers to line transect counts where birds were recorded within a 50 m wide band (25 m at either side from the observer). They are useful to evaluate inter–habitat distribution and seasonal changes of bird abundance. These data were obtained from a review on the abundance distribution of Iberian passerines (see TELLERÍA et al., 1999 for details).

The statistical analyses were designed for specifically testing the relationships between the parameters of bird communities and some environmental effects, as well as the differences of dehesa bird communities with bird assemblages of conifer and broad–leaved woodlands, the two main groups of trees according to habitat preferences of Iberian forest birds (TELLERÍA & SANTOS, 1994).

Regression analyses on log–transformed data were performed and ANOVA / ANCOVA comparisons in which dehesas were compared to the other woodlands were planned. All analyses were performed using the Multiple Regression and ANOVA / ANCOVA modules implemented in STATISTICA 5.5 (STATSOFT, 1999).

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Animal Biodiversity and Conservation 24.2 (2001)

Pyrenees

Iberian peninsula

12º 16º

Atlantic Ocean 16º

Mediterranean Sea

Fig. 1. Distribution of woodlands sampled to evaluate the structure of Iberian forest bird communities. The isotherm of 12ºC approximately delimits the distribution of Northern Iberian Plateau and mountains. White dots represent those woodlands where bird richness was evaluated by point counts during the spring. Black dots are sites where the structure of bird communities was evaluated by line transects during spring and winter. White and black stars show the situation of the dehesas sampled in both cases (see text for further details).

Fig. 1. Distribución de los bosques muestreados para estudiar la estructura de las comunidades de aves forestales ibéricas. La isoterma de 12ºC delimita, aproximadamente, la distribución de la Meseta y montañas del norte ibérico. Los puntos blancos sitúan los bosques censados en primavera mediante estaciones de escucha. Los puntos negros son bosques censados en primavera e invierno mediante transectos. Las estrellas blancas y negras representan la ubicación de las dehesas en ambos casos (ver el texto para más detalles).

Results

Discussion

Tree density of Iberian woodlands decreased southwards along the Iberian peninsula, a pattern associated to a concomitant increase of bird richness (figs. 2, 3A). This was related to the changing composition of bird assemblages provided that arboreal and open–habitat birds decreased and increased respectively in the more open woodlands (fig. 3B). The high species richness of dehesa bird assemblages in relation to other Iberian woodlands could thus be related to the presence of large numbers of border birds (figs. 4A, 4B; table 1). Dehesas showed higher point and internal beta diversities during the breeding period (fig. 4C, 4D; see however the marginal statistical significance for internal beta diversity in table 1) but did not differ in spring and winter densities (fig. 4F; table 1). However, they showed higher species richness than other bird assemblages, even after controlling for the effects of altitude (an index of winter hardness along the Iberian gradient; fig. 5; table 2).

Putting dehesas in a biogeographical context Processes operating on larger spatial and temporal scales are important determinants of the structure of bird communities as they determine, for instance, the characteristics of the species pool from which local communities can be assembled (CALEY & SCHLUTER, 1997). This claim is particularly important when analysing the structure of bird communities in the Iberian peninsula given that many forest birds are adapted to environmental conditions of central Europe and become increasingly scarce in the Mediterranean (BLONDEL, 1990; MÖNKKÖNEN, 1994; TELLERÍA & SANTOS, 1993; HAGEMEIJER & BLAIR, 1997).

From a historical perspective, this pattern of decreasing bird richness can be interpreted as the outcome of paleo–environmental fluctuations experienced by the Western Palearctic from the Quaternary (BLONDEL & MOUVER–CHAUVIRÉ, 1998).

Throughout the late Pleistocene period, forests

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70

A dehesas broadleaved conifer

Tree density (Nº / ha)

400

300 r = -0.47

P < 0.001

200

100

0 0

200

400 Distance (km)

600

800

B 40 36

Nº species

32 28 24 20 16 r = 0.37 P = 0.005 12

0

200

400 Distance (km)

600

800

Fig. 2. Geographical distribution of tree density and bird richness in some Iberian woodlands.

These patterns have been obtained by recording the mean density of trees (mean number of trunks > 20 cm dbh per ha) and the total richness of forest passerines recorded in the study woodlands (white dots in fig. 1). Distance refers to the number of km from each study area to the western side of the Pyrenees at the Spanish–French frontier. It describes a rough geographical and environmental gradient in which the arrow shows the transition between the northern plateaus and the southern half of Iberia (southern border of the 12ºC isotherm in fig. 1). Dehesas are represented by black dots.

Fig. 2. Distribución de la densidad del arbolado y de la riqueza de aves forestales en algunos bosques ibéricos. Estos patrones han sido obtenidos registrando la densidad media del arbolado (número medio de troncos de más de 20 cm de diámetro por ha) y el número total de especies de pájaros en los bosques estudiados (puntos blancos en fig. 1). La distancia se refiere al número de km desde cada bosque a la frontera franco–española. Describe un gradiente geográfico y ambiental en el que la flecha marca el límite meridional de la meseta norte (borde sur de la isoterma de los 12ºC en la fig. 1). Las dehesas están representadas por puntos negros.

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Animal Biodiversity and Conservation 24.2 (2001)

A

40 dehesas broadleaved conifer

Nº species

34

r = -0.41 P < 0.001

28

22

16

10 -50

B

50

150 250 T ree density (Nº / ha) Tree

350

450

16 14

Nº species

12 10 8 6 r = -0.42 P < 0.001

4 2 0 -2

-50

50

150 250 T ree density (Nº / ha) Tree

350

450

Fig. 3. Effects of tree density on bird richness of Iberian woodlands. A. Relationship between tree density and the total richness of passerines. B. Relationship between tree density and the richness of arboreal and border passerines (black and white dots respectively). Arboreal species are those that feed and nest in trees (e.g. tits Parus, treecreepers Certhia, nuthatches Sitta, etc.) while border species are those which usually breed in open habitats or woodlands with very scarce tree cover (larks Galerida, waigtails Mortacilla, wheathears Oenanthe, stonechats Saxicola, some Sylvia warblers, shrikes Lanius, buntings Miliaria, etc.).

Fig. 3. Efectos de la densidad del arbolado sobre la riqueza de pájaros en los bosques ibéricos.

A. Relación entre la densidad de árboles y la riqueza total de pájaros. B. Relación entre la densidad de árboles y la riqueza de pájaros arborícolas y ecotónicos (puntos negros y blancos, respectivamente). Las especies arborícolas son aquellas que crían y se alimentan en el arbolado (carboneros Parus spp., agateadores Certhia, trepadores Sitta, etc.) mientras que las ecotónicas son aquellas que crían en medios abiertos o muy aclarados (cogujadas Galerida, lavanderas Motacilla, collalbas Oenanthe, tarabillas Saxicola, alcaudones Lanius, escribanos Miliaria, etc.).

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72

A

B

32

32

28 n = 24

26

n = 28

Nº species

Nº species

30

24 16 8

24 22 dehesas

broadleaved

0

conifer

C Point / total richness

11 Nº species

dehesas

boradleaved

conifer

D

12

10 9 8 7

dehesas

broadleaved

conifer

0.40 0.30 0.36 0.34 0.32 0.30

E

dehesas

boradleaved

conifer

F 26

18

n = 21

n = 17

16 dehesas

broadleaved

conifer

Nº birds / 10 ha

winte winterr

22

14

80

spring

24 n = 8 Nº species

other arboreal border

n = 6

spring 70

winter

60 50 40 30

dehesas

boradleaved

conifer

Fig. 4. Structure of forest bird communities in the Iberian peninsula. Patterns of bird richness during the breeding period as reflected by sampling point censuses: A. The mean total richness (± s.e.); B. Mean richness of arboreal and border passerines; C. Mean point richness; D. The relationship between the point and the total richness (an index of the internal turnover of species are represented). Other species in B refer to birds that are neither arboreal nor border passerines (see fig. 2). Seasonal changes of Iberian forest bird communities: E. The mean scores (± s.e.) of the number of species; F. Densities recorded by line transects are shown. Border species were originally excluded from these censuses. Broad–leaved woodlands are mainly composed of oaks (Quercus) and conifer woodlands by pines (Pinus).

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Animal Biodiversity and Conservation 24.2 (2001)

concentrated in the mild Mediterranean peninsulas but, as global warming progressed, they shifted northwards producing the drawback of forest optimum to central Europe (MOREAU, 1954; HUNTLEY, 1993; T ABERLET et al., 1998). In the Mediterranean, these changes were coupled with severe human pressure (forest fragmentation, clearings, etc.) since ancient times, reinforcing the effects of summer drought. The structure of mature forests was thus replaced by a predominantly shrubby vegetation or clearings to favour pasture growth. This structural modification changed the original microclimatic conditions favouring the propagation of heliophytic, schlerophyllous or pyrophytic trees, such as the holm oaks (Quercus ilex) and cork oaks (Quercus suber) of the Iberian dehesas (COSTA et al., 1990).

These environmental changes produced a concomitant expansion of Mediterranean birds adapted to xeric conditions and open habitats (e.g. several Mediterranean Sylvia warblers) and the retreat to moist habitats or sectors (river banks, rainy mountains, etc.) of birds adapted to more mesic conditions (e.g. Erithacus rubecula, Sylvia atricapilla, Phylloscopus collybita, Prunella modularis, Sylvia borin, Troglodytes troglodytes, Turdus philomelos, etc; see TELLERÍA & SANTOS, 1994; PURROY, 1997 for the Iberian peninsula). This depletion of forest conditions was particularly strong in the southern half of Iberia, warmer and drier than northern mountains and plateaux (FONT, 1983) where woodlands are today composed of central European trees (Fagus sylvatica, Quercus robur, Pinus sylvestris) or by trees adapted to Mediterranean mountains (e.g. Quercus pyrenaica, Juniperus thurifera; BLANCO et al., 1997). In addition, as many of these northern forests have been managed for wood production, they presently show higher tree development and density (fig. 2). Consequently, forest birds are more abundant and widespread in the north as compared to southern Iberia where several species common to European woodlands are lacking (e.g.

Anthus trivialis, Prunella modularis, Regulus regulus, Sylvia borin, Parus palustris, Turdus philomelos, etc.; PURROY, 1997; TELLERÍA et al.,

Table 1. Results of ANOVA planned comparisons testing for differences between "dehesas" and other woodlands (broadleaved and conifer) in the community traits described in figure 4 (contrast vector: dehesas -2, broadleaved 1, conifer 1).

Tabla 1. Resultados de las comparaciones con ANOVA para analizar las diferencias entre dehesas y otros bosques (de hoja ancha y coníferas) en los rasgos de la comunidad descritos en la figura 4 (vectores de contraste: dehesas –2, hoja ancha 1, coníferas 1).

d.f.

F

P

Total richness

1,55

3.77

0.057

Arboreal species

1,55

2.07

0.156

1,55

6.85

0.011

Border species

1,55

0.14

0.708

Point richness

Other species

1,55

11.33

0.001

Point / total richness

1,55

2.93

0.093

1,43

0.45

0.504

Spring density Winter density Spring richness Winter richness

1,43

3.07

0.087

1,43

0.09

0.761

1,43

10.16

0.003

1999). However, and despite this contrasting suitability of northern and southern Iberia for forest avifauna, the study showed that total bird richness in woodlands increased southwards along the Iberian gradient with dehesas showing the highest scores (figs. 2, 4). These noteworthy results can be explained, however, as an outcome of the particular structure of dehesas.

Fig. 4. Estructura de las comunidades de aves de los bosques ibéricos. Patrones de riqueza de pájaros durante el periodo reproductor reflejados por las estaciones de censo: A. Riqueza total media (± error estándar); B. Riqueza media de especies arborícolas y ecotónicas; C. Riqueza puntual media; D. Media de la relación entre la riqueza puntual y la riqueza total de cada comunidad (un índice del recambio interno de especies). Otras especies en B se refiere a las que no son ni arborícolas ni ecotónicas según los criterios explicados en la fig. 2. Cambios estacionales de las comunidades de aves forestales: E. Valores medios (± error estándar) del número de especies; F. Densidades registradas mediante taxiados. Las especies ecotónicas han sido eliminadas por no haber sido incluídas en muchos de los estudios revisados. Los bosques caducifolios están principalmente compuestos de robles y encinas (Quercus) y los de coníferas, de pinos (Pinus).

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A Winter–spring richness (Nº species)

12

dehesas broadleaved

8

conifer 4

r = -0.49 P < 0.001

0 -4 -8 -12 -16

0

400

800 1200 Altitude (m a.s.l.)

1600

2000

Winter–spring density (Nº birds / 10 ha)

B 150 r = -0.57 P < 0.001 100

50

0

–50

–100 0

400

800 1200 Altitude (m a.s.l.)

1600

2000

Fig. 5. Relationship between altitude and the seasonal change of Iberian forest bird communities as determined by subtracting spring scores from winter scores: A. Richness; B. Densities. Declined in winter as altitude increased (data from line transects in fig. 1).

Fig. 5. Relación entre la altitud y la evolución estacional de las comunidades de pájaros forestales caracterizada por la resta de los valores invernales a los primaverales: A. Riqueza; B. Densidad.

Disminuyen en invierno al aumentar la altitud (datos de los transectos en la fig. 1).

Breeding bird communities of Iberian dehesas: the role of habitat structure In addition to the effects of the regional pool of species, the structure of bird communities is determined by processes operating at local scales,

such as the floristic and physiognomic structure of habitats (WIENS, 1989). As a rule, “the greater the habitat variety, the greater the species diversity” (ROSENZWEIG, 1995). Consequently, bird communities finally result from a balance between the available regional pool of species and the

75

Animal Biodiversity and Conservation 24.2 (2001)

local structure of habitats. The observed high total species richness of the dehesas is an unexpected result (see above) that can be explained because, despite their decreasing tree density and diversity and the concomitant reduction of some arboreal birds (fig. 4), their open physiognomy favours the presence of many birds from open and border habitats (fig. 3). The dehesas can therefore be viewed as an ecotonic habitat where a mixture of forest and nonforest passerines occurs (figs. 3, 4; see also DÍAZ et al., 2002). This may explain why despite the strong role of climate (e.g. rainfall), floristic composition (conifer vs. broadleaved woodlands) and geographical location (distance to northern core areas for many forest birds) on the distribution of individual forest passerines (TELLERÍA & SANTOS, 1994), forest physiognomy (tree density) has been described as the main correlate of woodland bird diversity in the Iberian Peninsula (TELLERÍA et al., 1992).

An additional, distinctive feature of dehesas is the high point richness (internal alpha diversity) and the low intra–habitat turnover of species (internal beta diversity) as reflected by the elevated percentage of the total bird richness observed in each sampling point (fig. 4; table 1).

This feature can be related to the diverse structure of each sampling point, where grasslands and scattered oaks occur, and the strong spatial constancy of this physiognomy across a habitat where internal patchiness has been eliminated to produce a homogeneous cover of trees and grasses (BLONDEL & ARONSON, 1999).

As management of dehesas decreases and bushes occupy the understorey, an internal turnover of species occurs in which birds from open habitats retreat (larks Galerida, Lullula, starlings Sturnus, sparrows Passer, Petronia, and some finches Carduelis and buntings Emberiza ) while other birds from shrubby habitats expand (e.g.

Mediterranean Sylvia warblers; PULIDO & DÍAZ, 1992). Unfortunately, unmanaged dehesas do not recover the floristic diversity of former Mediterranean forests and shrublands given that are usually colonised by oak ramets and gum cistus (Cistus ladanifer). It has been suggested that this lack of a diverse cohort of trees and shrubs typical of more undisturbed Mediterranean woodlands (other Quercus spp, Olea, Pistacia, Phyllirea, Rhamnus,... BLANCO et al., 1997), together the low tree density that could impair the ability of some arboreal birds to feed (PULIDO & DÍAZ, 1997), strongly affect the ability of this habitat to maintain some forest passerines common to European forests. Some of these birds (e.g. robins Erithacus rubecula, blackcaps Sylvia atricapilla ) have sedentary populations in some undisturbed forests of Southern Iberia which show biological and morphological differences with their migratory counterparts (TELLERÍA & CARBONELL, 1999; PÉREZ– TRIS et al., 2000; TELLERÍA et al., 2001). These

Table 2. Results of ANCOVA planned comparisons testing for differences between dehesas and other woodlands (broadleaved and conifer) in the seasonal changes of density and richness observed in fig. 6 (effect: habitat; contrast vector: dehesas -2, broadleaved 1, conifer 1; covariate: altitude).

Tabla 2. Resultados de las comparaciones con ANCOVA para evaluar las diferencias entre dehesas y otros bosques (de hoja ancha y coníferas) en los cambios estacionales de densidad y riqueza observada en la figura 6 (efecto: hábitat; vector de contraste: dehesas –2, hoja ancha 1; coníferas 1; covariante: altitud).

d.f.

F

P

Habitat

1,42

0.49

0.489

Altitude

1,42

20.21

<0.001

Habitat

1,42

9.13

0.004

Altitude

1,42

10.71

0.002

Density

Richness

endemic populations, scarcer than the migratory ones, may become extremely scarce in areas covered by these extensive wooded pastures (PURROY, 1997).

Seasonal changes of bird communities Migratory behaviour is a dynamic response of birds to environmental opportunities so that many populations and species track seasonal productive outputs moving across environmental gradients. Latitudinal trends in productivity are considered responsible for latitudinal migrations, while changing productivity with elevation also shapes similar selective pressures on birds inhabiting at middle and low latitudes (ALERSTAM, 1990). The contrasting climatic patterns between northern Iberian highlands and southern lowlands are strongly related to seasonal changes in abundance and distribution of some forest passerines (TELLERÍA et al., 1999, 2001). While mountains and plateaux of northern Iberia are better for breeding because of their productive spring output, the southern half is a more suitable region for wintering. In addition to mild temperatures, the arrival of autumn rains in these southern lowlands produces a concomitant sprouting of primary productivity and the ripening of fruit on many shrubs and trees which

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local and migratory birds rely on (HERRERA, 1984; FUENTES, 1992). Because of these environmental conditions, south–western Iberia is one of the best wintering grounds for European migratory birds (TELLERÍA, 1988).

The dehesas seem a suitable place for wintering given that bird communities in this habitat maintain density and richness at the spring levels (fig. 4).

This pattern can be related, however, to the location of dehesas at the south–western quarter of the Iberian peninsula and not to any particular ability to lodge large numbers of wintering passerines.

After controlling for the effects of altitude —a surrogate of winter hardness— on the seasonal changes of Iberian bird communities, dehesas continued to have higher richness than other habitats but did not show any particular ability to maintain densities of wintering birds if compared to other woodlands located at the same altitudinal levels (fig. 5; table 2). This lack of seasonal differences can be related to the relatively low degree of seasonality in food supply to paserines (HERRERA, 1978a). It can be suggested that the lack of fruit–bearing shrubs in dehesas (e.g. Olea, Pistacia, Phyllirea, Rhamnus) as the result of bush removal may prevent the occupation of this habitat by the huge number of frugivorous passerines that winter in adequate forests and shrublands of the Mediterranean region (e.g. thrushes Turdus , blackcaps Sylvia atricapilla, robins Erithacus rubecula; HERRERA, 1998).

Conclusions A first conclusion resulting from this review concerns the high number of bird species recorded in the Iberian dehesas despite their location at a considerable distance from the core area of European forest birds. This feature results from the particular physiognomy of dehesas, where border and open–habitat bird species compensate for the loss of forest birds. As this increased richness has been described by comparing dehesas with a large sample of Iberian woodlands, and woodlands usually show the highest richness scores of land bird communities (WIENS, 1989), it can be presumed that dehesas preserve one of the most diverse bird assemblages of Iberian habitats. It is interesting to point out that this increased richness, as well as the observed high point richness (another distinctive feature of this habitat), has been recorded in other organisms. Dehesa pastures, for instance, may possess 120–180 plant species, and some grass plots hold one of the highest plant diversities ever measured (DÍAZ– PINEDA et al., 1981; FERNÁNDEZ–ALÉS et al., 1993).

These features, which enlarge the conservation value of dehesas (see Introduction), are the result of the traditional management of this habitat.

From here it follows that misuse of these wooded grasslands (e.g. overgrazing, scrub invasion, degradation of tree cover, etc.) may lead to the

depletion of these biodiversity resources (DÍAZ & PULIDO, 1995; BEAUFOY, 1998). A second conclusion refers to some evident shortcomings related to the open physiognomy of dehesas. A lack of developed tree and shrub covers have relegated some forest birds to remnants of Mediterranean forests so that the preservation of these patches inside the wooded matrix of dehesas benefits these birds in this extensive agro–ecosystem.

Finally, it should be emphasised that the seasonal changes of bird abundance on dehesas are quite similar to those observed in other woodlands of the Iberian peninsula, a region where climatic harshness (and its productive correlates) strongly affects the seasonal distribution of birds. This feature contrasts with the spectacular numbers of woodpigeons (millions) and cranes (many thousands) which winter in dehesas (PURROY, 1988; ALONSO & ALONSO, 1990) relying on acorns and grasses (the two main resources produced by this farming system).

Acknowledgements I thank Dr. Mario Díaz for constructive criticisms and for providing of some unpublished papers.

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DÍAZ, M., CAMPOS, P. & PULIDO, F. J., 1997. The Spanish dehesas: a diversity in land–use and wildlife. In: Farming and bird in Europe: The Common Agricultural Policy and its implications for wildlife conservation: 178–209 (D. J. Pain & M. W. Pienkowski, Eds.). Academic Press, San Diego.

DÍAZ, M., GONZÁLEZ, E., MUÑOZ-PULIDO, R. & NAVESO, M. A., 1996. Habitat selection patterns of common cranes Grus grus wintering in holm oak Quercus ilex dehesas of Central Spain: effect of human management. Biological Conservation, 75: 119–123.

DÍAZ, M. & MARTÍN, P., 1998. Habitat selectivity by wintering woodpigeons (Columba palumbus) in Holm–Oak (Quercus ilex) dehesa of Central Spain. Gibier et Faune Sauvage, 15: 167–181.

DÍAZ, M. & PULIDO, F., 1993. Relaciones entre la abundancia de artrópodos y la densidad del Herrerillo Común (Parus caeruleus) en dehesas durante el período reproductor. Ardeola, 40: 33–38.

– 1995. Wildlife–habitat relationships in the Spanish dehesa. In: Farming on the edge: The nature of traditional farmland in Europe: 103–111 (D. I. McCracken, E. Bignal & S. E. Wenlock, Eds.). Joint Nature Conservation Committee, Peterborough.

DÍAZ, M., PULIDO, F. J. & MARAÑÓN, T., 2001. Diversidad biológica en los bosques mediterráneos ibéricos: relaciones con el uso humano e importancia para la sostenibilidad de los sistemas adehesados. In: Beneficios comerciales y ambientales de la repoblación y la regeneración del arbolado del monte mediterráneo (P. Campos & G. Montero, Eds.). CIFOR–INIA, Madrid.

DÍAZ–PINEDA, F., NICOLAS, J. P., RUIZ, M., PECO, B. & G ONZÁLEZ –B ERNÁLDEZ , F., 1981. Sucession, diversité et amplitude de niche dans les paturages du centre de la péninsule Iberique.

Vegetatio, 47: 267–277.

FERNÁNDEZ–ALÉS, R., LAFFARGA, J. M. & ORTEGA, F., 1993. Strategies in Mediterranean grassland annuals in relation to stress and disturbance.

Journal of Vegetation Science, 4: 313–322.

FONT, I., 1983. Atlas climático de España. Instituto Nacional de Meteorología, Madrid.

FUENTES, M., 1992. Latitudinal and elevational variation in fruit phenology among Western European bird–dispersed plants. Ecography, 15: 177–183.

GONZÁLEZ, L. M., 1991. Historia Natural del Aguila Imperial Ibérica (Aquila adalberti Brehm 1861).

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Monografía del ICONA, Madrid.

HAGEMEIJER, W. & BLAIR, M. J., 1997. The EBCC Atlas of European Breeding Birds. T. & A. D. Poyser, London.

HERRERA, C. M., 1977. Ecología alimenticia del petirrojo ( Erithacus rubecula ) durante su invernada en encinares del Sur de España.

Doñana, Acta Vertebrata, 4: 35–59.

– 1978a. Ecological correlates of residence and non–residence in a Mediterranean passerine bird community. Journal of Animal Ecology, 47: 871–890.

– 1978b. Datos sobre la dieta invernal del Colirrojo Tizón ( Phoenicurus ochruros ) en encinares de Andalucía Occidental. Doñana, Acta vertebrata, 5: 61–71.

– 1980. Composición y estructura de dos comunidades mediterráneas de passeriformes.

Doñana Acta Vertebrata, 7: 1–340.

– 1984. A study of avian frugivores, bird-dispersed plants, and their interaction in Mediterranean scrublands. Ecological Monographs, 54: 1–23.

– 1998. Long–term dynamics of Mediterranean frugivorous birds and fleshy fruits: a 12–year study. Ecological Monographs., 68: 511–538.

H UNTLEY, B., 1993. Species–richness in northtemperate zone forests. Journal of Biogeography, 20: 163–180.

LÓPEZ–GORDO, J. L., LÁZARO, E. & FERNÁNDEZ–JORGE, A., 1976. Comparación de las dietas de Strix aluco, Asio otus y Tyto alba en un mismo biotopo de la provincia de Madrid. Ardeola, 23: 189–211.

M ÖNKKÖNEN , M., 1994. Diversity patterns in Palaearctic and Nearctic forest bird assemblages.

Journal of Biogeography, 21: 183–195.

MOREAU, R. E., 1954. The main vicisitudes of the European avifauna since the pliocene. Ibis, 96: 411–431.

PASCUAL, J. A., PERIS, S. & ROBREDO, F., 1991. Efectos de tratamientos forestales con cipermetrina y malatión sobre el éxito de cría del Herrerillo Común (Parus caeruleus). Ecología, 5: 359–374.

P ÉREZ–TRIS, J., CARBONELL, R. & TELLERÍA, J. L., 2000. Abundance distribution, morphological variation, and juvenile condition of robins Erithacus rubecula (L.) in their Mediterranean range boundary. Journal of Biogeography, 27: 879–888.

PERIS, S., 1991. Las aves. In: El libro de las dehesas: 749–802 (J. M. Gómez, Ed.). Consejería de Medio Ambiente y Ordenación del Territorio, Salamanca.

PULIDO, F. & DÍAZ, M., 1992. Relaciones entre la estructura de la vegetación y comunidades nidificantes de aves en las dehesas: influencia del manejo humano. Ardeola, 39: 63–72.

– 1997. Linking individual foraging behavior and population spatial distribution in patchy environments: a field example with Mediterranean blue tits. Oecologia, 111: 434–442.

– 2000. Foraging behaviour of bluetits in a patchly environment under contrasting levels

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of natural food suply. Journal of Avian Ecology, 31: 81–86.

PURROY, F. J., 1988. Sobre la invernada de la Paloma Torcaz (Columba palumbus) en Iberia In: Invernada de aves en la Península Ibérica: 137–151 (J. L. Tellería, Ed.). Monografías SEO, Madrid.

– (Ed.), 1997. Atlas de las aves de España (1975– 1995). SEO–Lynx Edicions, Barcelona.

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Seasonal changes in abundance and flight– related morphology reveal different migration patterns in Iberian passerines. Ardeola, 48: 27–46.

TELLERÍA, J. L. & SANTOS, T., 1993. Distributional pattern of insectivorous passerines in the Iberian forests: does abundance decrease near the border? Journal of Biogeography, 20: 235–240.

– 1994. Factors involved in the distribution of forest birds on the Iberian Peninsula. Bird Study, 41: 161–169.

– 1999. Distribution of birds in fragments of Mediterranean forest: the role of ecological densities. Ecography, 22: 13–19.

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W IENS , J. A., 1989. The ecology of bird communities. 2 vols. Cambridge University Press, Cambridge.

I

Animal Biodiversity and Conservation 24.2 (2001)

Animal Biodiversity and Conservation Animal Biodiversity and Conservation (abans Miscel·lània Zoològica) és una revista interdisciplinària publicada, des de 1958, pel Museu de Zoologia de Barcelona. Inclou articles d'investigació empírica i teòrica en totes les àrees de la zoologia (sistemàtica, taxonomia, morfologia, biogeografia, ecologia, etologia, fisiologia i genètica) procedents de totes les regions del món amb especial énfasis als estudis que d'una manera o altre tinguin relevància en la biología de la conservació. La revista no publica catàlegs, llistes d'espècies o cites puntuals. Els estudis realitzats amb espècies rares o protegides poden no ser acceptats tret que els autors disposin dels permisos corresponents. Cada volum anual consta de dos fascicles.

Animal Biodiversity and Conservation es troba registrada en la majoria de les bases de dades més importants i està disponible gratuitament a internet a http://www.museuzoologia.bcn.es/ servis/servis3.htm, de manera que permet una difusió mundial dels seus articles.

Tots els manuscrits són revisats per l'editor executiu, un editor i dos revisors independents, triats d'una llista internacional, a fi de garantir–ne la qualitat. El procés de revisió és ràpid i constructiu. La publicació dels treballs acceptats es fa normalment dintre dels 12 mesos posteriors a la recepció.

Una vegada hagin estat acceptats passaran a ser propietat de la revista. Aquesta es reserva els drets d’autor, i cap part dels treballs no podrà ser reproduïda sense citar–ne la procedència.

Normes de publicació Els treballs s'enviaran preferentment de forma electrònica (mzbpubli@intercom.es). El format preferit és un document Rich Text Format (RTF) o DOC que inclogui les figures (TIF). Si s'opta per la versió impresa, s'han d'enviar quatre còpies del treball juntament amb una còpia en disquet a la Secretaria de Redacció. Cal incloure, juntament amb l'article, una carta on es faci constar que el treball està basat en investigacions originals no publicades anteriorment i que està sotmès a Animal Biodiversity and Conservation en exclusiva.

A la carta també ha de constar, per a aquells treballs en que calgui manipular animals, que els autors disposen dels permisos necessaris i que compleixen la normativa de protecció animal vigent. També es poden suggerir possibles assessors.

Quan l'article sigui acceptat, els autors hauran d'enviar a la Redacció una còpia impresa de la versió final acompanyada d'un disquet indicant el programa utilitzat (preferiblement en Word).

Les proves d'impremta enviades a l'autor per a la correcció, seran retornades al Consell Editor en el termini de 10 dies. Aniran a càrrec dels autors les despeses degudes a modificacions ISSN: 1578–665X

substancials introduïdes per ells en el text original acceptat.

El primer autor rebrà 50 separates del treball sense càrrec a més d'una separata electrònica en format PDF. Manuscrits Els treballs seran presentats en format DIN A–4 (30 línies de 70 espais cada una) a doble espai i amb totes les pàgines numerades. Els manuscrits han de ser complets, amb taules i figures. No s'han d'enviar les figures originals fins que l'article no hagi estat acceptat.

El text es podrà redactar en anglès, castellà o català. Se suggereix als autors que enviïn els seus treballs en anglès. La revista els ofereix, sense cap càrrec, un servei de correcció per part d'una persona especialitzada en revistes científiques.

En tots els casos, els textos hauran de ser redactats correctament i amb un llenguatge clar i concís.

La redacció del text serà impersonal, i s'evitarà sempre la primera persona.

Els caràcters cursius s’empraran per als noms científics de gèneres i d’espècies i per als neologismes intraduïbles; les cites textuals, independentment de la llengua, seran consignades en lletra rodona i entre cometes i els noms d’autor que segueixin un tàxon aniran en rodona.

Quan se citi una espècie per primera vegada en el text, es ressenyarà, sempre que sigui possible, el seu nom comú.

Els topònims s’escriuran o bé en la forma original o bé en la llengua en què estigui escrit el treball, seguint sempre el mateix criteri.

Els nombres de l’u al nou, sempre que estiguin en el text, s’escriuran amb lletres, excepte quan precedeixin una unitat de mesura. Els nombres més grans s'escriuran amb xifres excepte quan comencin una frase.

Les dates s’indicaran de la forma següent: 28 VI 99; 28, 30 VI 99 (dies 28 i 30); 28-30 VI 99 (dies 28 a 30).

S’evitaran sempre les notes a peu de pàgina.

Format dels articles Títol. El títol serà concís, però suficientment indicador del contingut. Els títols amb designacions de sèries numèriques (I, II, III,...) seran acceptats previ acord amb l'editor.

Nom de l’autor o els autors.

Abstract en anglès que no ultrapassi les 12 línies mecanografiades (860 espais) i que mostri l’essència del manuscrit (introducció, material, mètodes, resultats i discussió). S'evitaran les especulacions i les cites bibliogràfiques. Estarà encapçalat pel títol del treball en cursiva.

Key words en anglès (sis com a màxim), que orientin sobre el contingut del treball en ordre d’importància.

Resumen en castellà, traducció de l'Abstract. De la traducció se'n farà càrrec la revista per a © 2001 Museu de Zoologia

II

aquells autors que no siguin castellanoparlants.

Palabras clave en castellà.

Adreça postal de l’autor o autors.

(Títol, Nom, Abstract, Key words, Resumen, Palabras clave i Adreça postal, conformaran la primera pàgina.)

Introducción. S'hi donarà una idea dels antecedents del tema tractat, així com dels objectius del treball.

Material y métodos. Inclourà la informació pertinent de les espècies estudiades, aparells emprats, mètodes d’estudi i d’anàlisi de les dades i zona d’estudi.

Resultados. En aquesta secció es presentaran únicament les dades obtingudes que no hagin estat publicades prèviament.

Discusión. Es discutiran els resultats i es compararan amb treballs relacionats. Els suggeriments de recerques futures es podran incloure al final d’aquest apartat.

Agra decimientos (optatiu).

Agradecimientos Refer enc Referenc enciia s. Cada treball haurà d’anar acompanyat de les referències bibliogràfiques citades en el text. Les referències han de presentar–se segons els models següents (mètode Harvard): * Articles de revista: CONROY, M. J. & NOON, B. R., 1996. Mapping of species richness for conservation of biological diversity: conceptual and methodological issues. Ecological Applications, 6:: 763–773. * Llibres o altres publicacions no periòdiques: SEBER, G. A. F., 1982. The estimation of animal abundance. C. Griffin & Company, London. * Treballs de contribució en llibres: MACDONALD, D. W. & JOHNSON, D. P., 2001. Dispersal in theory and practice: consequences for conservation biology. In: Dispersal: 358–372 (T. J. Clober, E. Danchin, A. A. Dhondt & J. D. Nichols, Eds.). Oxford University Press, Oxford. * Tesis doctorals: MERILÄ, J., 1996. Genetic and quantitative trait variation in natural bird populations. Tesis doctoral, Uppsala University. * Els treballs en premsa només han d’ésser citats si han estat acceptats per a la publicació: RIPOLL, M. (in press). The relevance of population studies to conservation biology: a review. Anim.

Biodivers. Conserv.

La relació de referències bibliogràfiques d’un treball serà establerta i s’ordenarà alfabèticament per autors i cronològicament per a un mateix autor, afegint les lletres a, b, c..., als treballs del mateix any. En el text, s’indicaran en la forma usual: “...segons WEMMER (1998) ... ”, “...ha estat definit per ROBINSON & REDFORD (1991)...”, “...les prospeccions realitzades (BEGON et al., 1999)...” Quan en el text s’anomeni un autor de qui no es dóna referència bibliogràfica el nom anirà en rodona: “...un altre autor és Caughley...”

Taules. Les taules es numeraran 1, 2, 3, etc. i han de ser sempre ressenyades en el text. Les taules grans seran més estretes i llargues que amples i curtes ja que s'han d'encaixar en l'amplada de la caixa de la revista.

Figures. Tota classe d’il·lustracions (gràfics, figures o fotografies) entraran amb el nom de figura i es numeraran 1, 2, 3,... i han de ser sempre ressenyades en el text. Es podran incloure fotografies si són imprescindibles. La mida màxima de les figures és de 15,5 cm d'amplada per 24 cm d'alçada. S'evitaran les figures tridimensionals. Tant els mapes com els dibuixos han d'incloure l'escala. Els ombreigs preferibles són blanc, negre o trama. S'evitaran els punteigs ja que no es reprodueixen bé.

Peus de figura i capçaleres de taula. Els peus de figura i les capçaleres de taula seran clars, concisos i bilingües en la llengua de l’article i en anglès.

Els títols dels apartats generals de l’article (Introducción, Material y métodos, Resultados, Discusión, Conclusiones, Agradecimientos y Referencias) no aniran numerats. No es poden utilitzar més de tres nivells de títols.

Els autors procuraran que els seus treballs originals no passin de 20 pàgines (incloent–hi figures i taules).

Si a l'article es descriuen nous tàxons, caldrà que els tipus estiguin dipositats en una institució pública.

Es recomana als autors la consulta de fascicles recents de la revista per tenir en compte les seves normes.

III

Animal Biodiversity and Conservation 24.2 (2001)

Animal Biodiversity and Conservation Animal Biodiversity and Conservation (antes Miscel·lània Zoològica ) es una revista interdisciplinar, publicada desde 1958 por el Museo de Zoología de Barcelona. Incluye artículos de investigación empírica y teórica en todas las áreas de la zoología (sistemática, taxonomía, morfología, biogeografía, ecología, etología, fisiología y genética) procedentes de todas las regiones del mundo, con especial énfasis en los estudios que de una manera u otra tengan relevancia en la biología de la conservación. La revista no publica catálogos, listas de especies sin más o citas puntuales. Los estudios realizados con especies raras o protegidas pueden no ser aceptados a no ser que los autores dispongan de los permisos correspondientes. Cada volumen anual consta de dos fascículos.

Animal Biodiversity and Conservation está registrada en todas las bases de datos importantes y además está disponible de forma gratuita en internet en http://www.museuzoologia.bcn.es/ servis/servis3.htm, lo que permite una difusión mundial de sus artículos.

Todos los manuscritos son revisados por el editor ejecutivo, un editor y dos revisores independientes, elegidos de una lista internacional, a fin de garantizar su calidad. El proceso de revisión es rápido y constructivo, y se realiza vía correo electrónico siempre que es posible. La publicación de los trabajos aceptados se realiza con la mayor rapidez posible, normalmente dentro de los 12 meses siguientes a la recepción del trabajo.

Una vez aceptado, el trabajo pasará a ser propiedad de la revista. Ésta se reserva los derechos de autor, y ninguna parte del trabajo podrá ser reproducida sin citar su procedencia.

Normas de publicación Los trabajos se enviarán preferentemente de forma electrónica (mzbpubli@intercom.es). El formato preferido es un documento Rich Text Format (RTF) o DOC, que incluya las figuras (TIF). Si se opta por la versión impresa, deberán remitirse cuatro copias juntamente con una copia en disquete a la Secretaría de Redacción. Debe incluirse, con el artículo, una carta donde conste que el trabajo versa sobre investigaciones originales no publicadas anteriormente y que se somete en exclusiva a Animal Biodiversity and Conservation.

En dicha carta también debe constar, para trabajos donde sea necesaria la manipulación de animales, que los autores disponen de los permisos necesarios y que han cumplido la normativa de protección animal vigente. Los autores pueden enviar también sugerencias para asesores.

Cuando el trabajo sea aceptado los autores deberán enviar a la Redacción una copia impresa de la versión final junto con un disquete del ISSN: 1578–665X

manuscrito preparado con un procesador de textos e indicando el programa utilizado (preferiblemente Word). Las pruebas de imprenta enviadas a los autores deberán remitirse corregidas al Consejo Editor en el plazo máximo de 10 días.

Los gastos debidos a modificaciones sustanciales en las pruebas de imprenta, introducidas por los autores, irán a cargo de los mismos.

El primer autor recibirá 50 separatas del trabajo sin cargo alguno y una copia electrónica en formato PDF. Manuscritos Los trabajos se presentarán en formato DIN A–4 (30 líneas de 70 espacios cada una) a doble espacio y con las páginas numeradas. Los manuscritos deben estar completos, con tablas y figuras. No enviar las figuras originales hasta que el artículo haya sido aceptado.

El texto podrá redactarse en inglés, castellano o catalán. Se sugiere a los autores que envíen sus trabajos en inglés. La revista ofrece, sin cargo ninguno, un servicio de corrección por parte de una persona especializada en revistas científicas.

En cualquier caso debe presentarse siempre de forma correcta y con un lenguaje claro y conciso.

La redacción del texto deberá ser impersonal, evitándose siempre la primera persona.

Los caracteres en cursiva se utilizarán para los nombres científicos de géneros y especies y para los neologismos que no tengan traducción; las citas textuales, independientemente de la lengua en que estén, irán en letra redonda y entre comillas; el nombre del autor que sigue a un taxón se escribirá también en redonda.

Al citar por primera vez una especie en el trabajo, deberá especificarse siempre que sea posible su nombre común.

Los topónimos se escribirán bien en su forma original o bien en la lengua en que esté redactado el trabajo, siguiendo el mismo criterio a lo largo de todo el artículo.

Los números del uno al nueve se escribirán con letras, a excepción de cuando precedan una unidad de medida. Los números mayores de nueve se escribirán con cifras excepto al empezar una frase.

Las fechas se indicarán de la siguiente forma: 28 VI 99; 28, 30 VI 99 (días 28 y 30); 28–30 VI 99 (días 28 al 30).

Se evitarán siempre las notas a pie de página.

Formato de los artículos Título. El título será conciso pero suficientemente explicativo del contenido del trabajo. Los títulos con designaciones de series numéricas (I, II, III, etc.) serán aceptados excepcionalmente previo consentimiento del editor.

Nombre del autor o autores.

Abstract en inglés de 12 líneas mecanografiadas (860 espacios como máximo) y que exprese la © 2001 Museu de Zoologia

IV

esencia del manuscrito (introducción, material, métodos, resultados y discusión). Se evitarán las especulaciones y las citas bibliográficas. Irá encabeza do por el título del trabajo en cursiva.

Key words en inglés (un máximo de seis) que especifiquen el contenido del trabajo por orden de importancia.

Resumen en castellano, traducción del abstract. Su traducción puede ser solicitada a la revista en el caso de autores que no sean castellano hablantes.

Palabras clave en castellano.

Dirección postal del autor o autores.

(Título, Nombre, Abstract, Key words, Resumen, Palabras clave y Dirección postal conformarán la primera página.)

Introducción. En ella se dará una idea de los antecedentes del tema tratado, así como de los objetivos del trabajo.

Material y métodos. Incluirá la información referente a las especies estudiadas, aparatos utilizados, metodología de estudio y análisis de los datos y zona de estudio.

Resultados. En esta sección se presentarán únicamente los datos obtenidos que no hayan sido publicados previamente.

Discusión. Se discutirán los resultados y se compararán con otros trabajos relacionados. Las sugerencias sobre investigaciones futuras se podrán incluir al final de este apartado.

Agradecimientos (optativo).

Referencias. Cada trabajo irá acompañado de una bibliografía que incluirá únicamente las publicaciones citadas en el texto. Las referencias deben presentarse según los modelos siguientes (método Harvard): * Artículos de revista: CONROY, M. J. & NOON, B. R., 1996. Mapping of species richness for conservation of biological diversity: conceptual and methodological issues. Ecological Applications, 6:: 763–773 * Libros y otras publicaciones no periódicas: SEBER, G. A. F., 1982. The estimation of animal abundance. C. Griffin & Company, London. * Trabajos de contribución en libros: MACDONALD, D. W. & JOHNSON, D. P., 2001. Dispersal in theory and practice: consequences for conservation biology. In: Dispersal: 358–372 (T. J. Clober, E. Danchin, A. A. Dhondt & J. D. Nichols, Eds.). Oxford University Press, Oxford. * Tesis doctorales: MERILÄ, J., 1996. Genetic and quantitative trait variation in natural bird populations. Tesis

doctoral, Uppsala University. * Los trabajos en prensa sólo se citarán si han sido aceptados para su publicación: RIPOLL, M. (in press). The relevance of population studies to conservation biology: a review.

Anim. Biodivers. Conserv.

Las referencias se ordenarán alfabéticamente por autores, cronológicamente para un mismo autor y con las letras a, b, c,... para los trabajos de un mismo autor y año. En el texto las referencias bibliográficas se indicarán en la forma usual: "...según WEMMER (1998)...", "...ha sido definido por ROBINSON & REDFORD (1991)...", "...las prospecciones realizadas (BEGON et al., 1999)..."

Cuando en el texto se mencione un autor no incluido en la bibliografía el nombre irá en redonda: "...otro autor es Caughley..."

Tablas. Las tablas se numerarán 1, 2, 3, etc., y se reseñarán todas en el texto. Las tablas grandes deben ser más estrechas y largas que anchas y cortas ya que deben encajarse en la caja de la revista.

Figuras. Toda clase de ilustraciones (gráficas, figuras o fotografías) se considerarán figuras, se numerarán 1, 2, 3, etc., y se citarán todas en el texto. Pueden incluirse fotografías si son imprescindibles. El tamaño máximo de las figuras es de 15,5 cm de ancho y 24 cm de alto. Deben evitarse las figuras tridimensionales. Tanto los mapas como los dibujos deben incluir la escala. Los sombreados preferibles son blanco, negro o trama. Deben evitarse los punteados ya que no se reproducen bien.

Pies de figura y cabeceras de tabla. Los pies de figura y cabeceras de tabla serán claros, concisos y bilingües en castellano e inglés.

Los títulos de los apartados generales del artículo (Introducción, Material y métodos, Resultados, Discusión, Agradecimientos y Referencias) no se numerarán. No utilizar más de tres niveles de títulos.

Los autores procurarán que sus trabajos originales no excedan las 20 páginas incluidas figuras y tablas.

Si en el artículo se describen nuevos taxones, es imprescindible que los tipos estén depositados en alguna institución pública.

Se recomienda a los autores la consulta de fascículos recientes de la revista para seguir sus directrices.

V

Animal Biodiversity and Conservation 24.2 (2001)

Animal Biodiversity and Conservation Animal Biodiversity and Conservation (formerly Miscel·lània Zoològica) is an interdisciplinary journal which has been published by the Zoological Museum of Barcelona since 1958. It includes empirical and theoretical research in all aspects of Zoology (Systematics, Taxonomy, Morphology, Biogeography, Ecology, Ethology, Physiology and Genetics) from all over the world with special emphasis on studies that stress the relevance of the study of Conservation Biology.

The journal does not publish catalogues, lists of species (with no other relevance) or punctual records. Studies about rare or protected species will not be accepted unless the authors have been granted all the relevant permits. Each annual volume consists of two issues.

A nimal Biodiversity and Conservation is registered in all principal data bases and is freely available online at http://www.museuzoologia.bcn.es /servis/servis3.htm, thus assuring world–wide access to articles published therein.

All manuscripts are screened by the Executive Editor, an Editor and two independent reviewers in order to guarantee the quality of the papers. The process of review is rapid and constructive. Once accepted, papers are published as soon as practicable, usually within 12 months of initial submission.

Upon acceptance, manuscripts become the property of the journal, which reserves copyright, and no published material may be reproduced without quoting its origin.

Information for authors Electronic submission of papers is encouraged (mzbpubli@intercom.es). The preferred format is a document Rich Text Format (RTF) or DOC, including figures (TIF). In the case of sending a printed version, four copies should be sent together with a copy in a computer disc to the Editorial Office. A cover letter stating that the article reports on original research not published elsewhere and that it has been submitted exclusively for consideration in Animal Biodiversity and Conservation is also necessary. When animal manipulation has been necessary, the cover letter should also make explicit that the authors follow current norms on the protection of animal species and that they have obtained all rellevant permissions. Authors may suggest referees for their papers.

Once an article has been accepted, authors should send a printed copy of the final version together with a disc of the manuscript prepared on a word processor. Please identify software (preferably Word). Proofs sent to the authors for correction should be returned to the Editorial Board within 10 days. Expenses due to any substantial alterations of the proofs will be charged to the authors.

ISSN: 1578–665X

The first author will receive 50 reprints free of charge and an electronic version of the article in PDF format.

Manuscripts Manuscripts must be presented on A–4 format page (30 lines of 70 spaces each) with double spacing. Number all pages. Manuscripts should be complete with figures and tables. Do not send original figures until the paper has been accepted.

The text may be written in English, Spanish or Catalan. Authors are encouraged to send their contributions in English. The journal provides a FREE service of correction by a professional translator specialized in scientific publications.

Care should be taken in using correct wording and the text should be written concisely and clearly. Wording should be impersonal, avoiding the use of the first person.

Italics must be used for scientific names of genera and species as well as untranslatable neologisms. Quotations in whatever language used must be typed in ordinary print between quotation marks. The name of the author following a taxon should also be written in small print.

The common name of the species should be written in capital letters. When referring to a species for the first time in the text, both common and scientific names must be given when possible.

Place names may appear either in their original form or in the language of the manuscript, but care should be taken to use the same criteria throughout the text.

Numbers one to nine should be written in full in the text except when preceding a measure.

Higher numbers should be written in numerals except at the beginning of a sentence.

Dates must appear as follows: 28 VI 99, 28,30 VI 99 (days 28th and 30th), 28–30 VI 99 (days 28th to 30th).

Footnotes should not be used.

Formatting of articles T itle. The title must be concise but as informative as possible. Part numbers (I, II, III,...) should be avoided and will be subject to the Editor’s consent.

Name of author or authors.

Abstract in English, no longer than 12 typewritten lines (840 spaces), covering the contents of the article (introduction, material, methods, results and discussion). Speculation and literature citation must be avoided. Abstract should begin with the title in italics.

Key words in English (no more than six) should express the precise contents of the manuscript in order of importance.

Resumen in Spanish, translation of the Abstract.

Summaries of articles by non–Spanish speaking authors will be translated by the journal on request.

Palabras clave in Spanish. © 2001 Museu de Zoologia

VI

Address of the author or authors.

(Title, Name, Abstract, Key words, Resumen, Palabras clave and Address should constitute the first page.)

Introduction. The introduction should include the historical background of the subject as well as the aims of the paper.

Material and methods. This section should provide relevant information on the species studied, materials, methods for collecting and analysing data and the study area.

Results. Report only previously unpublished results from the present study.

Discussion. The results and their comparison with related studies should be discussed. Suggestions for future research may be given at the end of this section.

Acknowledgements (optional).

References. All manuscripts must include a bibliography of the publications cited in the text. References should be presented as in the following examples (Harvard method): * Journal articles: CONROY, M. J. & NOON, B. R., 1996. Mapping of species richness for conservation of biological diversity: conceptual and methodological issues. Ecological Applications, 6:: 763–773. * Books or other non–periodical publications: SEBER, G. A. F., 1982. The estimation of animal abundance. C. Griffin & Company, London. * Contributions or chapters of books: MACDONALD, D. W. & JOHNSON, D. P., 2001. Dispersal in theory and practice: consequences for conservation biology. In: Dispersal: 358–372 (T. J. Clober, E. Danchin, A. A. Dhondt & J. D. Nichols, Eds.). Oxford University Press, Oxford. * Ph. D. Thesis: MERILÄ, J., 1996. Genetic and quantitative trait variation in natural bird populations. Ph. D. Thesis, Uppsala University. * Works in press should only be cited if they have been accepted for publication: RIPOLL, M. (in press). The relevance of population studies to conservation biology: a review.

Anim. Biodivers. Conserv.

References must be set out in alphabetical and chronological order for each author, adding the letters a, b, c,... to papers of the same year. Bibliographic citations in the text must appear in the usual way: "...according to W EMMER (1998)...", "...has been defined by R OBINSON & R EDFORD (1991)...", "...the prospections that have been carried out (B EGON et al., 1999)..." When an author is mentioned in the text but no bibliographical reference is given, the name must appear in ordinary print: "...another of these authors is Caughley..."

Tables. Tables must be numbered in Arabic numerals with reference in the text. Large tables should be narrow (across the page) and long (down the page) rather than wide and short, so that they can be fitted into the column width of the journal.

Figures. All illustrations (graphs, drawings or photographs) must be termed as figures, numbered consecutively in Arabic numerals and with reference in the text. Glossy print photographs, if essential, may be included.

Maximum size of figures is 15.5 cm width and 24 cm height. Figures will not be tridimensional. Both maps and drawings must include scale. The preferred shadings are white, black and bold hatching. Avoid stippling, which does not reproduce well.

Legends of tables and figures. Legends of tables and figures must be clear, concise, and written both in English and Spanish.

Main headings (Introduction, Material and methods, Results, Discussion, Acknowledgements and References) should not be numbered. Do not use more than three levels of headings.

Manuscripts should not exceed 20 pages including figures and tables.

If the article describes new taxa, type material must be deposited in a public institution.

Authors are advised to consult recent issues of the journal and follow its conventions.

Animal Biodiversity and Conservation 24.2 (2001)

VII

Animal Biodiversity and Conservation Subscription Form  Please enter my subscription to Animal Biodiversity and Conservation  66.11 € Spain  68.52 € Europe  69.12 € rest of world  Personal subscription:  21.04 € Spain  23.4 4 € Europe  24.04 € rest of world  Please despatch my issues by air mail (supplement of 6.01 € for outside Europe)  Please send me the Instructions to authors

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"La tortue greque" Oeuvres du Comte de Lacépède comprenant L'Histoire Naturelle des Quadrupèdes Ovipares, des Serpents, des Poissons et des Cétacés; Nouvelle édition avec planches coloriées dirigée par M. A. G. Desmarest; Bruxelles: Th. Lejeuné, Éditeur des oeuvres de Buffon, 1836. Pl. 7

Editor executiu / Editor ejecutivo / Executive Editor Joan Carles Senar

Secretaria de Redacció / Secretaría de Redacción / Editorial Office

Secretària de Redacció / Secretaria de Redacción / Managing Editor Montserrat Ferrer

Museu de Zoologia Passeig Picasso s/n 08003 Barcelona, Spain Tel. +34–93–3196912 Fax +34–93–3104999 E–mail mzbpubli@intercom.es

Consell Assessor / Consejo asesor / Advisory Board Oleguer Escolà Eulàlia Garcia Anna Omedes Josep Piqué Francesc Uribe

Editors / Editores / Editors Antonio Barbadilla Univ. Autònoma de Barcelona, Bellaterra, Spain Xavier Bellés Centre d' Investigació i Desenvolupament CSIC, Barcelona, Spain Juan Carranza Univ. de Extremadura, Cáceres, Spain Luís Mª Carrascal Museo Nacional de Ciencias Naturales CSIC, Madrid, Spain Adolfo Cordero Univ. de Vigo, Vigo, Spain Mario Díaz Univ. de Castilla–La Mancha, Toledo, Spain Xavier Domingo Univ. Pompeu Fabra, Barcelona, Spain Francisco Palomares Estación Biológica de Doñana, Sevilla, Spain Francesc Piferrer Inst. de Ciències del Mar CSIC, Barcelona, Spain Ignacio Ribera The Natural History Museum, London, United Kingdom Alfredo Salvador Museo Nacional de Ciencias Naturales, Madrid, Spain José Luís Tellería Univ. Complutense de Madrid, Madrid, Spain Francesc Uribe Museu de Zoologia de Barcelona, Barcelona, Spain

Consell Editor / Consejo editor / Editorial Board José A. Barrientos Univ. Autònoma de Barcelona, Bellaterra, Spain Jean C. Beaucournu Univ. de Rennes, Rennes, France David M. Bird McGill Univ., Québec, Canada Mats Björklund Uppsala Univ., Uppsala, Sweden Jean Bouillon Univ. Libre de Bruxelles, Brussels, Belgium Miguel Delibes Estación Biológica de Doñana CSIC, Sevilla, Spain Dario J. Díaz Cosín Univ. Complutense de Madrid, Madrid, Spain Alain Dubois Museum national d’Histoire naturelle CNRS, Paris, France John Fa Durrell Wildlife Conservation Trust, Trinity, United Kingdom Marco Festa–Bianchet Univ. de Sherbrooke, Québec, Canada Rosa Flos Univ. Politècnica de Catalunya, Barcelona, Spain Josep Mª Gili Inst. de Ciències del Mar CMIMA–CSIC, Barcelona, Spain Edmund Gittenberger Rijksmuseum van Natuurlijke Historie, Leiden, The Netherlands Fernando Hiraldo Estación Biológica de Doñana CSIC, Sevilla, Spain Patrick Lavelle Inst. Français de recherche scient. pour le develop. en cooperation, Bondy, France Santiago Mas–Coma Univ. de Valencia, Valencia, Spain Joaquín Mateu Estación Experimental de Zonas Áridas CSIC, Almería, Spain Neil Metcalfe Univ. of Glasgow, Glasgow, United Kingdom Jacint Nadal Univ. de Barcelona, Barcelona, Spain Stewart B. Peck Carleton Univ., Ottawa, Canada Eduard Petitpierre Univ. de les Illes Balears, Palma de Mallorca, Spain Taylor H. Ricketts Stanford Univ., Stanford, USA Joandomènec Ros Univ. de Barcelona, Barcelona, Spain Valentín Sans–Coma Univ. de Málaga, Málaga, Spain Tore Slagsvold Univ. of Oslo, Oslo, Norway

Animal Biodiversity and Conservation 24.1, 2001 © 2001 Museu de Zoologia, Institut de Cultura, Ajuntament de Barcelona Autoedició: Montserrat Ferrer Fotomecànica i impressió: Sociedad Cooperativa Librería General ISSN: 1578–665X Dipòsit legal: B–16.278–58

Les cites o els abstracts dels treballs d’aquesta publicació es resenyen a / Las citas o los abstracts de los trabajos de esta publicación se mencionan en / This publication is cited or abstracted in: Abstracts of Entomology, Agrindex, Animal Behaviour Abstracts, Anthropos, Aquatic Sciences and Fisheries Abstracts, Behavioural Biology Abstracts, Biological Abstracts, Biological and Agricultural Abstracts, Current Primate References, Ecological Abstracts, Ecology Abstracts, Entomology Abstracts, Environmental Abstracts, Environmental Periodical Bibliography, Genetic Abstracts, Geographical Abstracts, Índice Español de Ciencia y Tecnología, International Abstracts of Biological Sciences, International Bibliography of Periodical Literature, International Developmental Abstracts, Marine Sciences Contents Tables, Oceanic Abstracts, Recent Ornithological Literature, Referatirnyi Zhurnal, Science Abstracts, Serials Directory, Ulrich’s International Periodical Directory, Zoological Records.

ISSN 1578–665X

Animal Biodiversity and Conservation 24.2 (2001)

Índex / Índice / Contents

1–15 Akani, G. C., Ogbalu, O. K. & Luiselli, L. Life–history and ecological distribution of chameleons (Reptilia, Chamaeleonidae) from the rain forests of Nigeria: conservation implications

45–49 Mateu, J. Nueva especie de carábido troglobio de la Alpujarra almeriense (Coleoptera, Carabidae, Pterostochini)

17–24 Fernández, J., Toro, M. A. & Caballero, A. Practical implementation of optimal management strategies in conservation programmes: a mate selection method

51–65 Robinson, W. D. Review Changes in abundance of birds in a Neotropical forest fragment over 25 years: a review

25–44 Gardner, T. Review Declining amphibian populations: a global phenomenon in conservation biology

67–78 Tellería, J. L. Review Passerine bird communities of Iberian dehesas: a review