Animal Biodiversity and Conservation issue 28.2 (2005) by Museu Ciències Naturals de Barcelona MCNB

Formerly Miscel·lània Zoològica

2005

and

Animal Biodiversity Conservation 28.2

"Le psettus rhomboidal (Psettus rhombeus, Cav. Nat.)" Le Règne Animal par Georges Cuvier; Paris: Fortin, Masson et Cie, Librairies; Pl. 42 Poissons 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 Passeig Picasso s/n 08003 Barcelona, Spain Tel. +34–93–3196912 Fax +34–93–3104999 E–mail abc@mail.bcn.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 José Antonio Donazar Estación Biológica de Doñana–CSIC, Sevilla, Spain Gary D. Grossman Univ. of Georgia, Athens, USA Damià Jaume IMEDEA–CSIC, Univ. de les Illes Balears, Spain Jordi Lleonart Inst. de Ciències del Mar CMIMA–CSIC, Barcelona, Spain Jorge M. Lobo Museo Nacional de Ciencias Naturales–CSIC, Madrid, Spain Pablo J. López–González Univ de Sevilla, Sevilla, Spain Juan José Negro Estación Biológica de Doñana–CSIC, Sevilla, Spain Vicente M. Ortuño Univ. de Alcalá de Henares, Alcalá de Henares, Spain Miquel Palmer IMEDEA–CSIC, Univ. de les Illes Balears, Spain Francisco Palomares Estación Biológica de Doñana–CSIC, 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 Nacional de Ciencias Naturales–CSIC, Madrid, Spain Pedro Rincón Museo Nacional de Ciencias Naturales–CSIC, Madrid, Spain Alfredo Salvador Museo Nacional de Ciencias Naturales–CSIC, Madrid, Spain José Luís Tellería Univ. Complutense de Madrid, Madrid, Spain Francesc Uribe Museu de Ciències Naturals 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, 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 28.2, 2005 © 2005 Museu de Ciències Naturals, 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

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Preferencias de hábitat, densidad y diversidad de las comunidades de aves en Tenerife (Islas Canarias) L. M. Carrascal & D. Palomino

Carrascal, L. M. & Palomino, D., 2005. Preferencias de hábitat, densidad y diversidad de las comunidades de aves en Tenerife (Islas Canarias). Animal Biodiversity and Conservation, 28.2: 101–119. Abstract Species–specific habitat preferences, density and species richness of bird communities in Teneriffe (Canary Islands).— Bird distribution and abundance are described and analyzed in Teneriffe (Canary Islands). Inter– habitat differences in density, diversity and species richness are shown in table 1. Figure 2 shows the main determinants of bird species richness in Teneriffe, and tables 2 and 3 and figure 3 show the species–specific patterns of spatial variation abundance (more detailed for Anthus berthelotii, Fringilla coelebs canariensis, Fringilla teydea, Parus caeruleus teneriffae, Phylloscopus canariensis, Regulus teneriffae, Serinus canarius and Turdus merula cabrerae). Deeply transformed environments due to human impact (urban habitats, agricultural mosaics, banana plantations) have high bird densities and species richness, even higher than those measured in native, unmodified habitats such as laurel forests or mature pinewoods. Urban environments in Teneriffe are very permeable to native bird fauna, as they have been occupied by many widespread endemic species/subspecies. Many of the endemic, well defined species or subspecies of island birds have high population densities within native, untransformed habitats. Density compensation and niche expansion is not a common phenomenon in the avifauna of Teneriffe. Nevertheless, all species/subspecies broadening the inter–habitat or altitudinal distribution are endemic of the Canary Islands. Key words: Altitudinal distribution, Avifauna, Bird density, Habitat preferences, Island vs. mainland comparisons, Teneriffe Island. Resumen Preferencias de hábitat, densidad y diversidad de las comunidades de aves en Tenerife (Islas Canarias).— Mediante el empleo de transectos lineales, se describen las preferencias de hábitat, la distribución altitudinal y la abundancia de la avifauna reproductora de Tenerife (Islas Canarias). Los hábitats profundamente transformados debido a la acción humana (e.g., áreas urbanas, mosaicos agrícolas, plantaciones de plátanos) tienen elevadas densidades y riquezas de especies, que llegan a ser tan altas o mayores que las observadas en medios autóctonos no transformados como laurisilvas y pinares maduros. Muchas especies/subespecies taxonómicamente bien diferenciadas de las poblaciones continentales están distribuidas mayoritaria o exclusivamente en hábitats autóctonos poco degradados. Las hipótesis de la compensación de densidades y la expansión de nicho en poblaciones insulares no parecen cumplirse de modo generalizado en Tenerife. No obstante, todas las especies o subespecies que muestran una mayor amplitud de distribución en Tenerife son endémicas del archipiélago canario. Palabras clave: Distribución altitudinal, Avifauna, Densidad de aves, Preferencias de hábitat, Tenerife. (Received: 1 VI 04; Conditional acceptance: 16 IX 04; Final acceptance: 30 IX 04) L. M. Carrascal & D. Palomino, Dept. Biodiversidad y Biología Evolutiva, Museo Nacional de Ciencias Naturales–CSIC, c/ José Gutiérrez Abascal 2, 28006 Madrid, Spain. Corresponding author: L. M. Carrascal. E–mail: mcnc152@mncn.csic.es

ISSN: 1578–665X

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Introducción En la actualidad los archipiélagos e islas oceánicos constituyen escenarios de investigación ornitológica tan habituales en contextos de teoría ecológica como de biología de la conservación. Por una parte, sus condiciones de aislamiento geográfico a lo largo de enormes periodos de tiempo son idóneas para el examen de hipótesis complejas y/o poco susceptibles de experimentación en el campo de la ecología evolutiva. Algunos ejemplos de teorías y conceptos biogeográficos basados en observaciones en ecosistemas insulares serían el equilibrio dinámico de especies (MacArthur & Wilson, 1967), la compensación de densidades poblacionales (MacArthur et al., 1972; Blondel et al., 1988), la expansión/contracción de nichos ecológicos (Blondel et al., 1988; Martin, 1992; Prodon et al., 2002), o el ciclo del taxón (Wilson, 1961; Ricklefs & Bermingham, 1999). Por otra parte, los trabajos de corte conservacionista y aplicado sobre comunidades insulares de aves son también muy frecuentes, dados el alto nivel de endemismos que acogen estos ambientes (Johnson & Stattersfield, 1990; Stattersfield et al., 1998) y su predisposición a padecer elevados niveles de amenaza y tasas de extinción (Milberg & Tyrberg, 1993; Collar et al., 1994; Pimm et al., 1988; pero ver también Manne et al., 1999). A pesar del extenso conocimiento ornitológico del archipiélago Canario (ver la extensa revisión bibliográfica de Martín & Lorenzo, 2001), el número de trabajos cuantitativos sobre la distribución, abundancia y biogeografía de su avifauna es aún reducido. No obstante, los trabajos ya acumulados abordan una notable diversidad de temas, incluyendo descripciones generales de distribución y abundancia en el archipiélago (p.e., Báez, 1992; Fernández–Palacios & Andersson, 1993; Marshall & Baker, 1999; Carrascal & Palomino, 2002), estructura y composición de comunidades en determinados hábitats (p.e., Suárez, 1984; Carrascal, 1987), procesos de selección de hábitat y uso del espacio en especies concretas (p.e., Carrascal et al., 1992; Valido et al., 1994; Illera, 2001), fenómenos de interacción planta–animal (p.e., Nogales et al., 2001; 2002) o paleontología (p.e., Rando et al., 1999; Rando, 2002). La disponibilidad de información empírica sobre los patrones de distribución y abundancia de las aves es fundamental para análisis biogeográficos detallados y para una adecuada gestión de la biodiversidad. Esta información puede servir para establecer de modo objetivo la rareza de los organismos y de esta manera poder definir categorías de amenaza y Listas Rojas (ver por ejemplo los criterios y listas SPEC para aves europeas y UICN para todo el planeta; Tucker & Heath, 1994; UICN, 2001). Para llevar a cabo esta tarea según los criterios cuantitativos, actuales es necesario enfrentarse a preguntas como ¿cuán abundante es cada especie? ¿cuál es su valencia ecológica? y ¿en qué medida están creciendo o decreciendo sus poblaciones? Considerando estos hechos, este trabajo utiliza el

potencial que tienen los datos obtenidos a partir de censos extensivos de aves para examinar hipótesis biogeográficas y macroecológicas, y cuantificar la rareza actual de la avifauna de la isla de Tenerife, para de este modo proporcionar información que pueda contribuir a una mejor gestión de este recurso natural. Para ello, elaboramos modelos descriptores de la distribución y abundancia de las especies de aves reproductoras en Tenerife, identificando para cada una de ellas los rasgos ambientales concretos que más les favorecen. Además, se examinan algunas hipótesis relativas a avifaunas insulares: expansión de nicho y compensación e incrementos de densidad en poblaciones de especies insulares, y empobrecimiento diferencial de la avifauna en distintos hábitats atendiendo a la estructura de la vegetación (Blondel, 1979; Wiens, 1989; Brown & Lomolino, 1998). Material y métodos Área de estudio Tenerife es una isla de 2.059 km2 localizada en el archipiélago canario y distante 288 km de la costa africana. Su gran extensión y la existencia de un amplio gradiente altitudinal (desde el nivel del mar hasta el Teide a 3.718 m) determina una gran diversidad de condiciones climatológicas y formaciones vegetales diferentes (Anónimo, 1980; González et al., 1986). Los principales medios autóctonos que pueden distinguirse son (1) tabaibales y cardonales–tabaibales dominados por plantas marcadamente xerófilas de porte arbustivo y subarbóreo (Euphorbia spp., Plocama pendula, Kleinia neriifolia) localizados en el piso basal (0–500 m); (2) monteverde constituido por laurisilvas y diferentes etapas seriales de su degradación (brezales y fayales–brezales), localizado fundamentalmente en el norte de la isla entre los 500 y 1.200 m de altitud y dominado por diferentes especies de árboles y arbustos arborescentes (Erica arborea, Myrica faya, Persea indica, Ocotea foetens, Laurus azorica, Picconia excelsa e Ilex canariensis); (3) diversas formaciones de porte arbustivo distribuidas por encima del cardonal–tabaibal hasta los 2.500 m de altitud (jarales, codesares, retamares; dominados principalmente por Spartocytisus spp., Chamaecytisus proliferus, Adenocarpus spp., Cistus spp. y Micromeria spp.); (4) pinares de Pinus canariensis distribuidos desde los 1.200 m en el norte de la isla y los 600 m en el sur hasta los 2.100 m de altitud; (5) formaciones alpinas localizadas por encima de los 2.500 m y caracterizadas por una escasísima cubierta vegetal relegada a unos pocos caméfitos que se desarrollan sobre malpaises y otros suelos volcánicos. A estos grandes tipos de paisaje hay que añadir formaciones agropecuarias derivadas de las actividades humanas que se localizan desde el nivel del mar hasta los 1.000 m principalmente (plataneras, mosaicos de cultivo, pastizales) y áreas urbanas de distinto tamaño y desarrollo urbanístico.

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3170 3160

Latitud (UTM)

3150 3140 3130 3120 3110 3100 3090 310

320

330

340 350 360 370 Longitud (UTM)

380

390

400

Fig. 1. Mapa de la localización en Tenerife de los 592 transectos lineales de 0,5 km. La latitud y longitud se expresa en coordenadas (miles de metros) dentro del bloque 28R. Fig. 1. Location of the 592 line transects of 0.5 km censused in Teneriffe Island. Latitude and longitude in UTM coordinates (in thousands of metre) within the block 28R.

Método de censo Durante Abril de 2002 y 2003 se efectuaron censos a lo largo de toda la isla para cuantificar la distribución y abundancia de las aves de Tenerife (fig. 1). Los censos sólo incluyeron medios terrestres, habiéndose descartado la línea de costa y las zonas húmedas artificiales (balsas, embalses). El método elegido fue el del transecto lineal, contabilizándose todas las aves vistas u oídas a lo largo del trayecto. Se distinguieron los contactos efectuados hasta una distancia de 25 m a cada lado del observador con el objeto de efectuar estimas de densidad. Sólo se censó en días sin viento ni precipitaciones, entre las 7:00–11:00 y las 16:00–17:30 GMT. La velocidad media de progresión andando fue de 1–3 km/h. Para más detalles acerca de esta metodología consúltese Bibby et al. (2000). Debido a las horas de censo, las aves nocturnas quedaron excluidas de los inventarios aunque fueran observadas (casos del Búho Chico–Asio otus, Lechuza Común–Tyto alba y Chocha Perdiz–Scolopax rusticola). Los vencejos tampoco pudieron ser distinguidos con toda seguridad durante los censos, por lo que sólo se anotaron los individuos contados sin identificarlos a nivel de especie (vencejos Común– Apus apus, Unicolor–A. unicolor y Pálido–A. pallidus en el área de estudio; Martín & Lorenzo, 2001) y no se incluyeron en los análisis de datos. Tampoco fueron considerados en los análisis las aves migradoras observadas sin constancia de reproduc-

ción segura y habitual en Tenerife. En el caso de las palomas de laurisilva (Turqué–Columba bolli y Rabiche–C. junoniae) y las tórtolas recientemente introducidas (Turca–Streptopelia decaocto y de Cabeza Rosa–S. roseogrisea) no siempre fue posible identificar al nivel de especie a todos los individuos. Sin embargo, las aves identificadas como "indeterminadas" fueron asignadas específicamente manteniendo las proporciones observadas en aquellos individuos identificados dentro de cada localidad (6 localidades en zonas de monteverde, y 7 localidades en áreas urbanas). Los transectos fueron divididos en unidades de 500 m. El punto medio de cada uno de ellos fue georeferenciado mediante un GPS Garmin 12 (latitud, longitud y altitud) utilizando la función promedio permaneciendo inmóvil durante 2 minutos. Los transectos se definieron en unidades de paisaje y tipos de hábitat lo más homogéneos posible, mediante el estudio de mapas 1:25.000 y visitas previas a las áreas de censo. En cada unidad de transecto se efectuaron tres estimas de la estructura de la vegetación (a 125 m, 250 m y 375 m dentro del transecto de 500 m) que fueron promediadas para caracterizarlo. Se midió la cobertura (en porcentaje) de herbáceas en el suelo, la cobertura del estrato arbustivo, la altura media del matorral, la cobertura del estrato arbóreo, la altura promedio del arbolado y la cobertura de suelo urbano. Se establecieron 9 categorías de porcentajes (0, 1,

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2–5, 5–12, 13–25, 25–50, 50–75, 75–95, 95–100%) para establecer las coberturas. También se caracterizó cada transecto atendiendo a si incluían barrancos o suelo con uso agropecuario (diversos tipos de cultivo o praderas de siega). Los grandes tipos de bosques se codificaron atendiendo a si eran o no (si 1; no 0) pinares o monteverde, o a si los arbustos eran tabaibas (Euphorbia spp), brezos, codesos/escobones (Chamaecytisus proliferus, Adenocarpus spp.), o retamas del Teide (Spartocytisus supranubius). En total se efectuaron 296 km de censo repartidos en 592 unidades de censo de 0,5 km. Para cada especie se proporcionó el porcentaje de unidades en que se observó (independientemente de la distancia a la que fuese contactada; i.e., sin considerar bandas de censo). Estas 592 unidades fueron agrupadas en 26 formaciones vegetales – tipos de paisaje atendiendo a su localización geográfica (principalmente norte o sur de la isla y posición altitudinal), proximidad y estructura de la vegetación (tabla 1). Cada una de ellas fue representada por al menos 20 has de censo (8 unidades de 0,5 km). Debido a las características particulares de algunas muestras (mixtas entre distintas formaciones ambientales; p.e., transiciones pinar–fayal, laurisilvas degradadas, tabaibales–cultivos, brezales–huertos, áreas urbanas muy dispersas) o escasez numérica total o por localidad (menos de 20 ha censadas), 103 unidades de censo no fueron utilizadas en la tabla 1 al definir los 26 hábitats principales de Tenerife. En ellos se estableció la densidad de cada especie, excluyendo a las aves nocturnas, vencejos y migrantes no reproductores. Estos valores de densidad deben considerarse medidas mínimas de abundancia, ya que el método del taxiado no detecta todos los individuos existentes, al oscilar generalmente las detectabilidades (probabilidad de detectar un ave estando presente), según las especies, entre un 33% y 80% (Bibby et al., 2000). Las 592 unidades de censo también fueron agrupadas en seis bandas altitudinales cada 500 m. En cada una de ellas se calculó la abundancia relativa de las especies expresada en aves/km, utilizando para ello todas las aves vistas u oídas sin tener en cuenta la banda principal de recuento de 25 m a cada lado del observador. Análisis de datos Las comunidades de aves en cada uno de estos 26 hábitats fue caracterizada por la densidad total de aves, la riqueza de especies y la diversidad. La riqueza se midió mediante el número de especies cuya densidad era mayor de 0,5 aves/10 ha (S0,5). De este modo se evitó el efecto de la diferente superficie muestreada en distintas comunidades y la inclusión de especies accidentales o muy raras. La diversidad se estimó mediante el índice de Shannon (H’ = – Spi · ln pi, donde pi es la proporción de la densidad de la especie i dentro de la densidad total de aves). Las estimas de diversidad no se han visto influidas por la distinta superficie

muestreada en cada unidad ambiental, ya que H’ y superficie de censo (en logaritmo) no están significativamente relacionados en la muestra de las 26 formaciones de la tabla 1 (r = 0,016, p = 0,939). Otro tanto ocurre al analizar el efecto que la superficie muestreada tiene sobre la estima de riqueza estandarizada a un mínimo de densidad (> 0,5 aves/10 ha; r = 0,008, n = 26, p = 0,970). Con la abundancia relativa en las seis bandas de distribución se calculó la amplitud de distribución altitudinal utilizando la siguiente fórmula: 6

–1

(3 p ) 2

i=1

Amplitud = 6 donde pi es la proporción de la abundancia relativa en cada una de las seis bandas altitudinales. Este índice varía entre 0,17 y 1, de manera que a mayor valor del índice se corresponde una mayor amplitud de distribución de la especie. Los factores influyentes sobre la distribución de las aves se identificaron con árboles de regresión (De’Ath & Fabricius, 2000) aplicados al número total de aves observadas en cada una de las unidades de censo de 0,5 km (i.e., sin considerar las bandas principales de recuento de 25 m). Los árboles de regresión someten a la variable respuesta (índice kilométrico de abundancia en este caso; aves/0,5 km) a sucesivas divisiones dicotómicas para obtener grupos homogéneos de muestras. Tales divisiones se hacen según criterios determinados por las variables predictoras (las que describen las características de cada unidad de censo). Mediante este procedimiento (1) se obvia la necesidad de establecer "a priori" patrones lineales homogéneos a todo el conjunto de datos (caso de la regresión múltiple), (2) se evita el ajuste "forzado" a distribuciones canónicas concretas a los cuales no tienen por qué ajustarse los datos, y (3) se definen modelos de efectos jerarquizados que particionan la variabilidad original en subconjuntos de datos en los cuales pueden estar operando de modo distinto variables predictoras diferentes (i.e., estima de interacciones). Los árboles de regresión permiten enfrentarse con éxito a las complejidades inherentes de los datos en ecología, como son las relaciones no lineares entre las variables respuesta y los predictores, o las interacciones entre predictores, por lo que son muy adecuados para explorar los patrones de distribución y variación de la abundancia en aves. Debido a la escasez de datos para algunas especies (menos de 5 presencias en las 592 unidades de censo), no fue posible obtener árboles de regresión en todos los casos. La complejidad de los árboles de regresión fue limitada atendiendo a las siguientes condiciones: árboles que redujesen la devianza significativamente, con un máximo de once criterios de clasificación (i.e., ramificación) que definían doce puntas con, al menos, 5 unidades de censo (los árboles incluyen principalmente

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5–8 puntas, con unos 7–15 transectos de 0,5 km por punta). Se muestran y comentan los árboles de regresión de 8 especies / subespecies endémicas de las Islas Canarias, ampliamente extendidas y que ilustran los grandes modelos de distribución de las aves en Tenerife. Para el resto de las especies, los resultados de los árboles de regresión se sintetizan mediante la selección de los criterios que maximizan la abundancia de cada especie en Tenerife. Estos criterios definen las configuraciones del paisaje que permiten el asentamiento de las poblaciones más densas dentro de la isla. La variación espacial de la riqueza de especies en los 592 transectos lineales de 0,5 km también se analizó utilizando los árboles de regresión, siguiendo los criterios expuestos en el párrafo anterior. En esta ocasión, la riqueza se mide como el número de especies observado en 0,5 km de transecto sin utilizar distancias límite de detección (i.e., se incluyen todas las especies observadas). Resultados y discusión Comunidades de aves La tabla 1 muestra los valores de densidad, riqueza y diversidad de la avifauna terrestre de la isla de Tenerife en 26 formaciones ambientales. La densidad de aves osciló entre 0 y 95 aves/10 ha. Las mayores densidades se midieron en zonas urbanas de gran extensión (e.g., ciudades, 75– 100 aves/10 ha) y en una gran variedad de medios situados en el área de influencia de los alisios en el norte de la isla (pastizales, zonas agrícolas, brezales, fayales–brezales y laurisilva, 60–70 aves/10 ha). Por el contrario, las menores densidades (< 10 aves/ 10 ha) se obtuvieron en zonas de alta montaña (> 2.200 m), en áreas bajas del sur de la isla con muy poca cobertura vegetal (< 250 m y cobertura de arbustos y herbáceas < 25%) y en pinares abiertos altimontanos o jóvenes (> 1.650 m y < 11 m de altura del arbolado). La riqueza de especies con densidades mayores de 0,5 aves/10 ha osciló entre 0 y 15 especies. El mayor valor de riqueza se midió en las áreas agropecuarias del norte de la isla situadas entre 500 y 1.000 m de altitud. También se obtuvieron elevados valores (10–12 especies) en una gran variedad de formaciones vegetales que incluyen los tabaibales–cardonales del norte de la isla (especialmente en los barrancos), otras zonas agrícolas (cultivos del sur y plataneras) y algunas zonas urbanas (tanto pueblos pequeños como ciudades). Las menores riquezas (< 5 especies) se observaron en áreas desarboladas de alta montaña (> 2.200 m), formaciones xéricas de las zonas bajas del sur de la isla (< 250 m), y pinares jóvenes poco densos. Un patrón similar se obtuvo para la diversidad de aves (correlación entre riqueza y diversidad: r = 0.823, n = 26, p < 0.001). El empobrecimiento de los medios insulares es muy patente en los hábitats de alta montaña de Tenerife, que

105

no están representados por ninguna especie típicamente alpina o cuyos máximos poblacionales se den allí. Este hecho contrasta fuertemente con la presencia de varias especies alpinas en los grandes macizos montañosos de la península ibérica (e.g., Lagópodo Alpino–Lagopus mutus, Bisbita Alpino–Anthus spinoletta, Pechiazul–Luscinia svecica, Treparriscos–Tichodroma muraria, Chova Piquigualda–Pyrrhocorax graculus, Gorrión Alpino– Montifringilla nivalis o Acentor Alpino–Prunella collaris; Martí & Del Moral, 2003), algunos de los cuales tienen menor superficie por encima de 2.000 m que Tenerife (e.g., Sierra Nevada, Cordillera Cantábrica). A continuación se analizan los principales determinantes de la variación espacial de la riqueza de especies en Tenerife. El 67,8% de la variabilidad observada (0–14 especies/0,5 km) es explicada por el árbol de regresión de la figura 2 (χ2 = 3451, df = 12, p << 0.001). Los principales factores que influyen sobre la riqueza de especies son la situación latitudinal y altitudinal dentro de la isla. En la mitad meridional de la isla (LAT<3.135; 3,5 especies/0,5 km) la riqueza es menor que en la mitad septentrional (6,9 especies/0,5 km). La altitud no tiene un efecto lineal sobre la riqueza en el sur de Tenerife, sino que alcanza su máximo entre 328 y 1.100 m, siendo menor por encima de 1.656 m que en el piso basal (ALT<328 m; ver distintas ramificaciones para la variable ALT en las ramas de la izquierda del árbol de regresión; ver también los valores medios de especies/0,5 km en la tabla 2). En el norte de la isla existe una asociación negativa con esta variable, de manera que hay más especies por debajo de 1.270 m que por encima de esta altitud. Los efectos de la altitud y la latitud son matizados por el desarrollo del estrato arbóreo, la presencia de cultivos y la existencia de núcleos urbanos. El desarrollo del arbolado tiene un marcado efecto positivo sobre la riqueza de especies en el sur de la isla (tanto considerando la cobertura como la altura de árboles; criterios CARB<22% y HARB<3,2 m). La existencia de áreas agrícolas incrementa ligeramente el número de especies en el norte de la isla (criterio AGR=0). La cobertura de suelo urbano tiene un efecto distinto según la localización dentro de Tenerife. En el sur de la isla, y a altitudes menores de 328 m (LAT<3.135 y ALT<328), la riqueza de especies es mayor en áreas urbanas densas (CURB>92%; 5 spp/0,5 km) que en el resto de los ambientes disponibles en esta área. Por el contrario, en el norte de Tenerife el suelo urbano, aunque sea disperso (CURB>58%) disminuye el número de especies. Patrones específicos de distribución y abundancia La tabla 2 muestra la variación de la abundancia relativa de las aves a lo largo de un gradiente altitudinal, así como su amplitud de distribución. La tabla 3 ilustra la frecuencia de aparición de las especies en los 592 transectos efectuados, la den-

Carrascal & Palomino

106

Tabla 1. Densidades (aves/10 ha) de las aves en 26 medios diferentes en la isla de Tenerife. En la parte inferior de la tabla se proporcionan las variables descriptoras de cada hábitat. A continuación se proporciona una breve descripción de los 26 medios censados (n, situación en el norte de la isla; s, situación en el sur de la isla): Am. Alta montaña; Rm. Retamar de Spartocitysus supranubius sobre malpais; R. Retamar de Spartocitysus supranubius; Mo. Matorrales occidentales (Cistus spp., Echium spp., Sonchus spp., 1.100–1.400 m); TCn y TCs. Tabaibales–cardonales situados en el norte y sur de la isla; TCxs. Tabaibales–cardonales xéricos; Txs. Tabaibales xéricos; Bn. Barrancos cubiertos de tabaibales–cardonales y restos de arbolado termófilo; Br. Brezales; FB. Fayal–brezal; L. Laurisilva; Pn y Ps. Pinares de Pinus canariensis situados en el norte y sur de la isla; Pjn. Pinares jóvenes de Pinus canariensis; Pm. Pinares maduros de Pinus canariensis; Pa. Pinares altitudinales de Pinus canariensis; Pah. Pastizales húmedos; Es. Campos de cultivo abandonados de porte estepárico en el sur de la isla; Pl. Plataneras; Cn y Cs. Mosaicos de cultivos situados en el norte y sur de la isla; Pbn y Pbs. Pueblos situados en el norte y sur de la isla; Cun y Cus. Cascos urbanos extensos situados en el norte y sur de la isla. S0,5. Número de especies con densidad mayor de 0,5 aves/10 ha. HAS. Hectáreas censadas; CHB. Cobertura de herbáceas; CMAT. Cobertura de arbustos; HMAT. Altura media de los arbustos; CARB. Cobertura del arbolado (árboles mayores de 3 m de altura); HARB. Altura media del arbolado; AGR. Porcentaje de suelo dedicado a uso agrícola; CURB. Cobertura de suelo urbano.

TCn

TCs

TCxs Txs

Accipiter nisus

–

Alectoris barbara

–

0.21

3.00

–

0.27

–

Anthus berthelotii

–

0.78

1.92

0.31

0.84

3.40

5.47 4.46

–

Bucanetes githagineus

–

– 0.31

–

Burhinus oedicnemus

–

0.21

–

Buteo buteo

–

0.11

–

0.08

–

Carduelis cannabina

–

0.63

–

1.07

–

Carduelis carduelis

–

Carduelis chloris

–

Columba bollii

–

Columba junoniae

–

Columba livia

–

1.26

2.00

– 1.23

6.13

–

Corvus corax

–

Coturnix coturnix

–

Dendrocopos major

–

Erithacus rubecula

–

0.11

0.32

–

3.16

–

Falco tinnunculus

–

0.21

0.42

Fringilla coelebs

–

0.42

–

– 3.56 3.00

Fringilla teydea

–

Gallinula chloropus

–

0.53

–

Lanius excubitor

–

0.22

0.48

–

+ 0.77

–

Miliaria calandra

–

Motacilla cinerea

–

0.80

–

Myiopsitta monachus

–

Parus caeruleus

–

0.11

–

3.38

2.74

1.60

0.21

–

Passer hispaniolensis

–

– 0.46

–

Petronia petronia

–

1.07 0.44 2.67 – 1.00

0.27 11.56 11.00 –

–

– 0.67

7.73 6.67 3.67

107

Animal Biodiversity and Conservation 28.2 (2005)

Table 1. Density (birds/10 ha) of bird species in 26 different habitats in Teneriffe Island. In the lower part of the table are shown the main characteristics of these habitats. Small letters with habitat names (n, located in the northern Teneriffe; s, located in the south of the island): Am. Poorly vegetated alpine habitats; Rm. Shrubland of Spartocitysus supranubius on volcanic outcrops and lava fields; R. Shrubland of Spartocitysus supranubius; Mo. Montane shrublands in the western part of the island (Cistus spp., Echium spp., Sonchus spp., 1,100–1,400 m); TCn and TCs. Scrublands of several Euphorbia species in north or south of Teneriffe; TCxs. Dry scrublands of several Euphorbia species in southern Teneriffe; Txs. Dry scrublands of several Euphorbia species in southern Teneriffe lacking Euphorbia candelabrum; Bn. Deep gullies covered by Euphorbia shrubs and some thermophilic trees and shrubs; Br. Heathlands of Erica arborea; FB. "Monteverde" mainly composed of tall heaths and trees of Myrica faya; L. Laurel forests; Pn and Ps. Pinewoods of Pinus canariensis in north or south of Teneriffe; Pjn. Young pinewoods of Pinus canariensis; Pm. Ancient pinewoods of Pinus canariensis; Pa. High altitude pinewoods of Pinus canariensis; Pah. Grasslands; Es. Abandoned agricultural fields, poorly vegetated, located in southern Teneriffe; Pl. Banana plantations; Cn and Cs. Mosaic of agricultural fields devoted to several crops in north or south of Teneriffe; Pbn and Pbs. Small villages in north or south of Teneriffe; Cun y Cus. Large cities in north or south of Teneriffe; Densidad. Density; S0,5. Number of species with densities higher than 0.5 birds/10 ha; Diversidad. Diversity; HAS. Hectares censused; CHB. Herbaceous layer cover; CMAT. Shrub cover; HMAT. Average shrub height; CARB.Tree layer cover (trees higher than 3 m); HARB. Average height of tree layer; AGR. Percentage cover of ground devoted to agricultural use; CURB. Urban cover.

Pjn

–

– 0.53

–

– 0.27

– 0.80

–

Pah

Pbn

Pbs Cun Cus

–

– 0.44

–

0.40

–

0.10

0.85

–

9.07 8.60

4.44

3.12

3.15

–

1.33

–

7.47

–

4.20

3.15

–

0.33

–

2.67

0.10

0.12

0.89

–

0.59

–

0.44

–

1.11

–

5.08

–

1.85

–

– 0.27

–

2.80

–

3.88

0.44 10.67 44.89

6.00

–

0.59

–

0.29 0.27 1.78

–

– 0.63 6.46 4.42

– 2.67 1.78

–

3.80

–

+ 0.27 0.89

0.40

0.44

0.20

0.12

4.31

–

1.37

–

0.29 6.93 4.44 1.33

–

– 1.26

–

– 0.53

–

0.60

–

10.40

–

1.56

–

1.33

0.49

1.58

1.78

1.33

0.89

–

5.54 5.68

3.14 7.73 8.44 0.80

–

0.44

–

0.53

–

2.22

0.88

2.79

2.22

2.33

1.33

0.20

–

4.00

0.78

–

7.33 13.78 39.00

–

15.73

–

4.00

–

Carrascal & Palomino

108

Tabla 1. (Cont.) Am

Phylloscopus canariensis

–

0.44

Psittacula krameri

–

Regulus teneriffae

–

Scolopax rusticola

–

Serinus canarius

–

Streptopelia decaocto

–

Streptopelia roseogrisea

–

Streptopelia turtur

–

Sylvia atricapilla

TCn

TCxs

Txs

3.40

0.84

–

0.42

–

– 12.22 10.33

–

2.77

1.68

1.00

–

1.60

2.44

–

0.20

–

0.62

0.40

0.42

–

1.60

–

2.11

0.20

–

3.73

1.56

–

Sylvia conspicillata

–

0.32

–

0.60

–

Sylvia melanocephala

–

1.85

2.11

2.40

–

2.13

–

Turdus merula

–

0.11

–

2.95

–

3.73 12.67 14.33

Upupa epops

–

0.0

1.8

5.9

19.1

26.8

18.2

8.0

7.8

41.4

65.8

61.7

Diversidad

0.00

1.52

1.23

1.33

2.23

2.11

1.15 1.33

2.18

1.90

1.86

HAS

27.5

90.0

62.5

32.5

47.5

50.0

47.5 65.0

37.5

45.0

30.0

Altitud

2593 2202 2200 1256

338

445

220

189

0.6

3.7

21.0

1.7

23.2 13.3

63.2

40.8

39.9

Densidad S0,5

2.88 10.15 8.00

TCs

– 10.67 14.67 15.00

0.42 0.62

–

915 1006

CHERB

0.0

9.4

26.6

2.6

CMAT

0.6

9.9

46.8

45.4

61.2

40.6

HMAT

0.1

0.8

1.1

1.0

0.8

0.6

0.5

0.9

2.3

2.6

CARB

0.0

1.2

0.3

0.0

2.1

54.3

63.8

HARB

0.0

0.2

2.3

1.2

0.0

3.6

5.0

9.3

AGROPEC

0.0

CURB

0.0

1.5

0.0

sidad ecológica máxima medida en los 26 medios distinguidos, y las configuraciones del paisaje que determinan las máximas abundancias de las especies de aves en Tenerife (obtenidas mediante árboles de regresión). La combinación de estos resultados con los de la tabla 1 perfilan los principales patrones de preferencias de hábitat, abundancia y amplitud de distribución de las especies en Tenerife. A continuación se resaltan los principales resultados relativos a la rareza de las aves en la isla. Las especies observadas en más del 33% de todos los transectos efectuados fueron el Anthus berthelotii (Bisbita Caminero), Turdus merula (Mirlo Común), Phylloscopus canariensis (Mosquitero Canario), Parus caeruleus (Herrerillo Común) y Serinus canarius (Canario). Otras especies ampliamente distribuidas fueron Falco tinnunculus

(Cernícalo Vulgar), Streptopelia turtur, Motacilla cinerea (Lavandera Cascadeña), Erithacus rubecula (Petirrojo), Sylvia atricapilla (Curruca Capirotada), Sylvia melanocephala (Curruca Cabecinegra) y Sylvia conspicillata (Curruca Tomillera). Columba livia (Paloma Bravía) y Passer hispaniolensis (Gorrión Moruno) alcanzaron densidades ecológicas máximas muy elevadas (aprox. 40 aves/10 ha). Streptopelia decaocto, Turdus merula, Erithacus rubecula, Regulus teneriffae (Reyezuelo Canario), Phylloscopus canariensis, Petronia petronia (Gorrión Chillón), Serinus canarius y Miliaria calandra (Triguero) fueron también localmente muy abundantes (10–25 aves/10 ha). Por el contrario, las especies autóctonas con menores frecuencias de aparición (FREC<2,5%) en los censos fueron Coturnix coturnix (Codorniz Común), Accipiter nisus

109

Animal Biodiversity and Conservation 28.2 (2005)

13.69

1.47

–

Pjn

0.86 2.40

Pan

Pbn

Pbs

Cun

Cus

4.44 2.13 0.53

–

4.00

15.51

5.94

4.89

8.33

6.00

3.60

–

2.00

17.08 13.05 0.29 2.67

4.00

–

0.44

–

0.92 0.84 1.14 1.07

4.89

– 22.93

–

7.56

15.61

1.21

1.33

0.33

4.67

–

4.33 11.56 19.00

–

0.44

0.29

–

3.56

–

3.11

4.40

0.31 0.21 0.29

–

0.44

–

0.89

0.88

0.24

0.44

1.33

–

0.15

–

1.33

3.32

1.45

3.11

1.67

0.67

1.00

–

– 1.07 0.20

1.78

0.20

2.18

–

0.33

–

0.15

–

0.20

0.44

2.15

0.61

0.44

0.33

–

– 0.53

–

– 0.53

–

3.11

6.93

0.36

2.22

5.67

5.11

0.80

–

0.44

–

0.33

–

0.40

6.3 25.3

31.6

5.9 68.3 13.2

35.1

62.6

27.6

21.8

50.0

95.3

76.4

1.98 1.63 1.49 1.85

2.00 1.61 1.68 1.09

2.39

2.26

2.33

2.20

2.28

1.78

1.44

65.0 47.5 35.0 37.5

22.5 37.5 37.5 50.0

22.5

102.5

82.5

22.5

30.0

45.0

50.0

744

727

356

547

358

13.69 1.47 –

–

69.2 29.1 9

– –

888 1268 1640 1640 1653 1987 722 1.2

2.7

0.6

0.0 91.3 15.7

15.5

68.5

42.1

2.6

12.9

4.8

0.0

13.4 28.8 12.0 11.4

26.0

9.2

5.3

9.4

16.7

13.0

28.8

11.0

16.9

1.9

6.4

1.3

1.1

0.6

1.3

1.0

0.8

0.9

0.5

0.2

0.5

38.7 21.9

0.2

0.0

46.6

2.1

1.1

2.8

3.3

4.7

1.1

1.5

9.5 1.8

2.1

0.9

1.6

85.5 72.3 35.3 65.1 9.9 17.0 11.0 14.9

18.6

8.2

0.5

0.0

3.0

3.4

2.4

5.7

3.8

6.6

6.3

0.0

0.0 100.0

100.0

93.9

0.0

8.3

0.0

1.8

6.8

81.4

56.8

84.4

92.5

(Gavilán Común), Columba junoniae, Dendrocopos major (Pico Picapinos), Petronia petronia, Bucanetes githagineus (Camachuelo Trompetero) y Carduelis cannabina (Pardillo Común). De la combinación de los valores observados de frecuencia de aparición, la abundancia ecológica máxima y la restricción en la distribución a unos pocos hábitats, se obtiene que las especies más raras de la avifauna terrestre de Tenerife (no introducidas recientemente) fueron Coturnix coturnix, Alectoris barbara (Perdiz Moruna), Burhinus oedicnemus (Alcaraván Común), Accipiter nisus, Columba junoniae, Upupa epops (Abubilla), Dendrocopos major, Corvus corax (Cuervo), Miliaria calandra, Petronia petronia, Bucanetes githagineus, Carduelis carduelis (Jilguero) y Carduelis chloris (Verderón Común).

1.8

La figura 3 ilustra los resultados de los árboles de regresión de ocho especies/subespecies endémicas de Canarias, ampliamente extendidas y representativas de los principales modelos de distribución de las aves terrestres en Tenerife. A continuación se exponen muy sucintamente los principales resultados relativos a la variación de sus abundancias. Phylloscopus canariensis es una especie de amplísima distribución en Tenerife, cuya variación de abundancia se ve influida principalmente por la posición geográfica en la isla, siendo la latitud la variable que más le influye. En la mitad meridional de la isla (LAT<3.136) es bastante más escasa que en la mitad septentrional, donde es especialmente abundante en Anaga (LAT>3.153). En este sector de Tenerife es donde la especie alcanza sus mayores

Carrascal & Palomino

110

LAT<3135

ALT<1656

ALT<328 CURB<92 2.4

CARB<22 1.3

ALT<1270 CURB<58 LONG<316 AGR=0 5.1 5.1 7.1 8.6

5.1

3.3

ALT<1100 5.0

5.5

7.3

4.3

Fig. 2. Árbol de regresión de la variación espacial de la riqueza de especies en Tenerife. Los criterios hacen referencia a las ramas de la izquierda (los valores son contrarios en las ramas derechas). La longitud de las ramas es proporcional a la devianza explicada por cada criterio (i.e., a mayor longitud, mayor variabilidad explicada). Los valores de las puntas expresan el número medio de especies observado en 0,5 km de transectos lineales. Para el significado de las siglas véase la tabla 3. Fig. 2. Regression tree analyzing the spatial variation in species richness in Teneriffe. The split criteria refer to the left side of each dichotomy. Branch lengths are proportional to deviance explained by each split criterium. Figures in the tips of the tree ("leaves") measure the average number of bird species observed per 0.5 km of line transect (without limiting the detection belt). See table 3 for acronyms.

abundancias, especialmente cuando la altura del estrato arbustivo (HMAT; mayoritariamente brezos) es mayor de 1,9 m. En la mitad meridional de Tenerife su abundancia está influida positivamente por el desarrollo del estrato arbustivo (CMAT>26), siendo siempre más densa en el extremo más occidental de la isla (LONG<330 en dos ramificaciones). El desarrollo del estrato arbóreo (en cobertura –CARB– y altura –HARB–) y la existencia de barrancos (BCO) son las variables que más influyen sobre la variación de la abundancia de Parus caeruleus, otra especie de muy amplia distribución en Tenerife. Alcanza sus mayores densidades en medios arbolados muy desarrollados (CARB>28% y HARB>16,5 m; 7,8 aves/km), o en barrancos con muy alta cobertura de arbustos (BCO=1 y CMAT>78%; 9,1 aves/km) en ausencia de bosques. En medios arbustivos no localizados en barrancos, la especie puede llegar a ser medianamente abundante si la cobertura de matorral es muy alta (>78%). Turdus merula es más abundante en aquellas áreas con presencia de brezos (BRZ=1) y ausencia de pinares (PIN=0), en especial en el sector más septentrional de Tenerife (LAT>3157; i.e., brezales, fayales y laurisilvas de Anaga donde alcanza un promedio de 15,1 aves/km). Si no hay brezos en el

estrato arbustivo (BRZ=0), la especie es mucho más abundante en el tercio norte de la isla (LAT>3135), donde alcanza sus mayores densidades en medios con cobertura de matorral (CMAT) mayor del 68%, situados a altitudes (ALT) superiores a 330 m (11,2 aves/km). Si no se cumplen los requisitos anteriores de presencia de brezos en el estrato arbustivo y elevadas coberturas de matorral, la tercera configuración del paisaje donde la especie tiene elevadas densidades es en áreas agrícolas (AGR=1) situadas por encima de 486 m de altitud (7,7 aves/km). En los dos tercios meridionales de Tanerife (LAT<3135) es muy escaso (promedio de 0,4 aves/km). La abundancia de Serinus canarius es mayor en zonas agropecuarias (AGR=1) que en otros ambientes naturales de Tenerife (los índices kilométricos de abundancia –IKA– de las ramas derechas de su árbol de regresión son mayores que las situadas a la izquierda). Dentro de las zonas agrícolas, sus mayores densidades se han observado en el cuadrante noroccidental de la isla (LONG<351 y LAT>3.135) a altitudes superiores a 683 m. En áreas sin uso agrícola su abundancia es mayor en la mitad oriental de la isla (LONG>340) si la cobertura de arbustos (CMAT) es superior al 24%.

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Animal Biodiversity and Conservation 28.2 (2005)

Tabla 2. Abundancia relativa (aves/km) de las aves de Tenerife a lo largo de un gradiente altitudinal de seis bandas a intervalos de 500 m: 0–500 m (A); 500–1.000 m (B); 1.000–1.500 m (C); 1.500– 2.000 m (D); 2.000–2.500 m (E); > 2.500 (F); AA. Amplitud de distribución altitudinal (mínima 0,17; máxima 1); Transectos. Número de transectos de 0,5 km de longitud con los que se han calculado las abundancias relativas; Altitud media. Altitud media de los transectos efectuados en cada banda altitudinal; Especies/0,5 km. Riqueza específica en cada banda altitudinal. Table 2. Relative abundance (birds observed per km of line transect) of birds in Teneriffe Island across an altitudinal gradient: six belts at 500 m interval: 0–500 m (A); 500–1,000 m (B); 1,000–1,500 m (C); 1,500–2,000 m (D); 2,000–2,500 m (E); > 2,500 (F); AA. Altitudinal breadth (minimum 0.17; maximum 1). Transectos. Number of transects of 0.5 km censused in each altitudinal belt; Altitud media. Average altitude of transects within each belt. Especies/0,5 km. Average species richness (spp/0.5 km of line transect) within each altitudinal belt.

Alectoris barbara Anthus berthelotii Bucanetes githagineus Burhinus oedicnemus Buteo buteo Carduelis cannabina Carduelis carduelis Carduelis chloris Columba bollii Columba junoniae Columba livia Coturnix coturnix Dendrocopos major Erithacus rubecula Falco tinnunculus Fringilla coelebs Fringilla teydea Lanius excubitor Miliaria calandra Motacilla cinerea Parus caeruleus Passer hispaniolensis Petronia petronia Phylloscopus collybita Regulus teneriffae Serinus canarius Streptopelia decaocto Streptopelia roseogrisea Streptopelia turtur Sylvia atricapilla Sylvia conspicillata Sylvia melanocephala Turdus merula Upupa epops Altitud media Transectos Especies/0,5 km

A 0.38 3.24 0.14 0.03 0.06 0.26 0.10 0.02 0.09 0.00 4.91 0.00 0.00 0.37 0.39 0.02 0.00 0.16 0.00 0.32 1.70 3.18 0.13 4.30 0.06 0.99 2.05 0.43 0.48 1.25 0.41 0.69 1.28 0.05 192 185 4.8

B 0.08 2.27 0.00 0.00 0.10 2.30 0.05 0.28 0.47 0.13 3.20 0.19 0.00 2.23 0.30 0.92 0.00 0.00 1.17 0.42 2.30 0.79 0.69 8.82 2.24 6.01 0.16 0.00 0.59 1.43 0.45 0.66 5.84 0.02 746 183 6.7

Bandas altitudinales C D 0.28 0.03 0.50 0.03 0.00 0.00 0.00 0.00 0.06 0.00 0.25 0.00 0.00 0.00 0.03 0.00 0.44 0.00 0.25 0.00 0.38 0.03 0.00 0.00 0.09 0.48 4.06 1.01 0.06 0.24 0.47 0.00 0.53 1.81 0.00 0.00 0.00 0.00 0.09 0.00 3.63 3.57 0.00 0.00 0.00 0.00 6.25 2.05 3.50 1.01 2.59 1.73 0.00 0.00 0.00 0.00 0.47 0.43 0.16 0.00 0.03 0.00 0.34 0.00 3.97 0.24 0.00 0.00 1216 1720 64 75 5.0 3.4

E 0.00 1.38 0.00 0.00 0.00 0.00 0.00 0.00 0.00 0.00 0.00 0.00 0.00 0.34 0.10 0.00 0.08 0.18 0.00 0.00 0.05 0.00 0.00 1.19 0.00 0.00 0.00 0.00 0.00 0.00 0.08 0.00 0.05 0.00 2187 77 1.2

F 0.00 0.50 0.00 0.00 0.00 0.00 0.00 0.00 0.00 0.00 0.00 0.00 0.00 0.00 0.25 0.00 0.00 0.00 0.00 0.00 0.00 0.00 0.00 0.00 0.00 0.00 0.00 0.00 0.00 0.00 0.00 0.00 0.00 0.00 2644 8 0.2

AA 0.42 0.58 0.17 0.17 0.48 0.24 0.31 0.23 0.39 0.30 0.35 0.17 0.23 0.47 0.80 0.31 0.27 0.33 0.17 0.40 0.62 0.24 0.23 0.60 0.42 0.46 0.19 0.17 0.66 0.37 0.41 0.46 0.42 0.28

Carrascal & Palomino

112

Tabla 3. Patrones de distribución y abundancia de las especies de aves terrestres en Tenerife: Dmax. Densidad (aves/10 ha) maxima medida en Tenerife en los 26 hábitats de la tabla 1; Frec. Frecuencia de aparición de las especies en los transectos de 0,5 km de longitud (en %); D2%. Devianza explicado por los árboles de regresión (en %); para cada especie también se proporcionan las características del espacio que describen los lugares donde son más abundantes, teniendo en cuenta los criterios de los árboles de regresión que conducen a las puntas con máxima densidad; IKA. Medida de abundancia máxima de cada especie en las configuraciones del hábitat definidas por los árboles de regresión (aves observadas/km de transecto; n. Número de muestras de 0,5 km con las cuales se obtiene el IKA). Tras cada sigla se indica el criterio cuantitativo que define el hábitat (+, presencia de ese atributo; –, ausencia del atributo): LAT, LONG. Latitud y longitud en coordenadas UTM (en kilómetros dentro del bloque 28R); ALT. Altitud (en m); CHB. Cobertura de herbáceas (en %); CMAT. Cobertura de arbustos (en %); HMAT. Altura media del matorral (en m); CARB. Cobertura de arbolado (en %); HARB. Altura media del arbolado (en m); AGR+. Suelo dedicado a la agricultura; LAUR+. Monteverde; PIN+. Pinares de Pinus canariensis; PIN–. Ausencia de pinares; BRZ+. Presencia de brezos en el estrato arbustivo; BRZ–. Ausencia de brezos; BCO+. Presencia de barrancos; CURB. Cobertura de suelo urbano (en %). Table 3. Distribution and abundance of terrestrial birds in Teneriffe Island: Dmax. Maximum density (birds/10 ha) measured in 26 hábitats in Teneriffe (table 1); Frec. Frequency of occurrence of each species in 592 transects of 0.5 km spread out in Teneriffe (in %); D2%. Deviance explained by regression trees analyzing the spatial variation of relative abundance of birds in line transects of 0.5 km (in %); the criteria of tree regression analyses defining the spatial configuration of the areas where the species are more abundant is provided for each species; for some species it was not possible to obtain regression models due to sample limitations; IKA. Average abundance of the species in the leaves defined by regression trees (birds per km of lineal transect; n. Number of transects of 0.5 km with which IKA’s are obtained). Quantitative criteria are provided after acronyms (+, presente of that habitat attribute; – absence of the attribute): LAT, LONG. Latitude and longitude in UTM coordinates (in km with UTM block 28R); ALT. Altitude above the sea level (in m); CHB. Cover of the herbaceous layer (in %); CMAT. Shrub cover (in %); HMAT. Average height of shrubs (in m); CARB. Cover of the tree layer (in %); HARB. Average height of the tree layer (in m); AGR+. Presence of groung devoted to agriculture; LAUR+. "Monteverde" (laurel forest and/or woodlands mainly composed by tall heaths and Myrica faya trees); PIN+. Pinewood of Pinus canariensis; PIN–.Absence of pinewoods; BRZ+. Presence of Erica arborea in the shrub layer; BRZ–. Absence of Erica arborea in the shrub layer; BCO+. Deep gullies; CURB. Urban cover (in %).

D 2%

Dmax

Frec

Accipiter nisus

0.53

0.5

Alectoris barbara

3.00

3.7

15.4

Anthus berthelotii

9.07

39.9

Bucanetes githagineus 0.31

0.8

Burhinus oedicnemus

0.21

0.5

Buteo buteo

0.11

2.2

Carduelis cannabina

7.47

Carduelis carduelis

2.67

Carduelis chloris Columba bollii Columba junoniae

IKA

CHB>48, LAT<3124

6.9

47.6

HARB<4.2, LAT<3116, CHB>32

9.8

22.0

LAT<3103, ALT<32

1.7

22.5

LONG>380, LAT<3158

1.4

14.5

54.3

CHB>42, AGR+, LAT>3140

9.5

0.8

13.7

AGR+, CARB>34

1.7

1.11

2.4

24.6

LAT>3140, CHB>65

2.5

5.08

5.1

57.0

LAUR+, LAT<3158, HARB>11.5

4.8

1.85

1.5

35.9

CARB>92, LAT<3136

Columba livia

44.89

23.5

48.5

CURB>4, LAT>3144, LONG<371

3.2

54.0

Corvus corax

0.05

0.3

Coturnix coturnix

0.59

1.5

27.1

CHB>68, LAT>3140

1.8

Dendrocopos major

1.78

2.4

44.6

PIN+, HARB>19.5

2.8

Erithacus rubecula

11.56

23.3

64.3

BRZ+, HMAT>1.7, CARB<72, LAT>3137 11.6

Falco tinnunculus

0.89

11.3

16.2

ALT<819, CURB<5.5, LAT<3102

1.8

Fringilla coelebs

4.31

9.0

31.8

BRZ+, PIN-, HARB>9.5

3.6

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Animal Biodiversity and Conservation 28.2 (2005)

Tabla 3. (Cont.) Dmax

Frec

D 2%

Fringilla teydea

6.93

8.4

62.7

Gallinula chloropus

0.53

0.2

Lanius excubitor

0.77

2.7

Miliaria calandra

10.40

Motacilla cinerea

1.78

Myiopsitta monachus

0.44

0.2

Parus caeruleus

8.44

Passer hispaniolensis 39.00

IKA

HARB>13.5, ALT>1594, LONG>346

5.2

18.2

LAT<3102, HMAT>0.65

1.2

5.1

58.9

CHB>88, CMAT<4, LAT>3137

12.3

9.0

35.7

CURB 1-52, CHB<51, CMAT<28, LON<352 3.2

42.1

41.9

CARB<28, BCO+, CMAT>78

9.1

7.9

52.6

CURB>82, LAT<3118, CMAT<1

36.3

Petronia petronia

15.73

0.8

16.2

CHB>92.5, LAT<3136

16.3

Phylloscopus canar.

15.51

67.1

48.9

LAT>3153, HMAT>1.9

19.4

Psittacula krameri

2.00

0.5

Regulus teneriffae

17.08

16.6

75.2

BRZ+, LAT>3144, LONG<375

15.2

Serinus canarius

22.93

32.4

52.2

AGR+, LONG<351, LAT>3135, ALT>683 18.1

Streptopelia decaocto

19.00

6.9

60.0

CURB>94, ALT<35

23.2

Streptopelia roseogrisea 4.40

2.4

43.2

CURB>94, ALT<32

8.0

Streptopelia turtur

1.60

14.2

31.7

CHB>8, CARB>0.5, LAT>3151

3.4

Sylvia atricapilla

3.73

23.0

43.3

LAT>3136, CARB<32, HMAT>0.6 HARB<6, ALT<180

5.1

Sylvia conspicillata

2.18

8.6

22.7

CHB>8, LAT<3137, LONG>358, HMAT>0.8 3.7

Sylvia melanocephala

2.40

14.4

31.6

ALT<1045, CMAT<29, AGR+, LONG>319 3.7 CARB<0.5, CHB>88

Turdus merula

14.33

38.3

68.8

BRZ+, PIN-, LAT>3157

Upupa epops

0.44

1.0

Anthus berthelotii es una especie generalista de medios deforestados, que se ve influida principalmente por el desarrollo del arbolado (negativamente), la posición latitudinal en la isla (más abundante en el sur que en el norte) y el desarrollo del estrato herbáceo (positivamente). Es muy escasa (0,3 aves/km) en áreas arboladas (altura de árboles > 4,2 m). En los medios con muy poco desarrollo del arbolado (< 4,2 m) es más abundante en el tercio meridional de la isla (LAT<3116; ver localización en la figura 1), que en el norte. Su abundancia aumenta con el desarrollo del estrato herbáceo, aunque requiere más cobertura de este estrato en el norte (68% a LAT>3116) que en el sur de Tenerife (32% cuando LAT<3116). En las áreas con poco desarrollo del arbolado del norte de la isla, donde la cobertura de herbáceas es mayor del 68%, la especie es considerablemente más abundante en la mitad occidental (LONG<345). Regulus teneriffae es una especie forestal generalista cuya abundancia está supeditada principalmente a la presencia de un estrato arbustivo de brezos bien desarrollado. Si no existe brezo (BRZ=0), la especie sólo estará presente, con

15.1

relativamente bajas abundancias, si el estrato arbóreo está muy desarrollado (HARB>16,5 m; 3,4 aves/km) o la cobertura de arbustos es elevada (CMAT>48%; 1,4 aves/km). En hábitats con presencia de brezos (BRZ=1), Regulus teneriffae alcanza sus mayores abundancias (>7,9 aves/km) en el tercio norte de la isla (LAT>3144), aunque su densidad decrece en la península de Anaga (LONG>375). La densidad de Fringilla teydea (Pinzón Azul) alcanza su máximo en Tenerife (5,2 aves/km) en los bosques con árboles de más de 13,5 m de altura localizados a más de 1594 m de altitud en la mitad oriental de la isla (LONG>346). Estos ambientes coinciden exclusivamente con los pinares de Pinus canariensis del sector oriental de la corona forestal de Tenerife. Cuando el arbolado no está muy desarrollado (HARB<13.5 m) la especies está virtualmente sólo presente en pinares (PIN=1), siendo considerablemente más abundante en los orientales (LONG>350; 2,5 aves/km) que en los occidentales (promedio de 0,5 aves/km). Fringilla coelebs (Pinzón Vulgar) es una especie típica de monteverde, cuya abundancia se ve influida principalmente por la presencia de brezos en el

114

estrato arbustivo. Incluso existiendo brezos, la especie está virtualmente ausente de pinares (BRZ=1 y PIN=1; 0,0 aves/km). En el monteverde, su abundancia se asocia positivamente con el desarrollo del estrato arbóreo, tanto en altura (HARB>9,5 m; 3,6 aves/km), como en cobertura si los árboles no son muy altos (CARB>55% cuando HARB<9,5 m; 2,1 aves/km). En ausencia de brezos como planta dominante del estrato arbustivo, la especie sólo es medianamente abundante en el norte de la península de Anaga (LAT>3.158; 1,7 aves/km) donde puede ser observado en tabaibales–cardonales bien desarrollados (ver CULTn en tabla 1). También está presente en algunas áreas agrícolas cuando éstas tienen arbolado disperso desarrollado (AGR=1 y HARB>6,5 m; 1,5 aves/km). Rareza, endemicidad e impacto humano Varias especies autóctonas de la avifauna de Tenerife son hoy marcadamente más escasas de lo que fueron hace 25–50 años (véase Martín, 1987; Martín & Lorenzo, 2001). Entre ellas se encuentran Upupa epops, Coturnix coturnix, Calandrella rufescens (Terrera Marismeña; no observada durante este estudio), Corvus corax, Petronia petronia, Carduelis cannabina, Carduelis carduelis y Miliaria calandra. A ellas hay que añadir los ya extintos Milvus milvus (Milano Real) y Neophron percnopterus (Alimoche). Por la información recopilada por diferentes ornitólogos en la primera mitad del siglo XX (ver revisión de Martín & Lorenzo, 2001) todas ellas fueron abundantes, llegando a alcanzar elevados efectivos poblacionales en zonas rurales y sus áreas agropecuarias colindantes, siendo escasas en hábitats naturales con poca influencia humana. Como consecuencia de los cambios en los usos tradicionales del suelo de los últimos 25–50 años (reducción o abandono de la agricultura de subsistencia, recuperación de la vegetación autóctona de tabaibales, matorrales de medianía y brezales, reducción de la cabaña ganadera de cabras, implantación de monocultivos industriales de plátano en áreas bajas cercanas a la costa, urbanismo masivo) y el efecto de los plaguicidas usados en las décadas de 1950– 1960, sus máximos efectivos poblacionales se han reducido. Consistentemente con lo descrito en el

Carrascal & Palomino

pasado (primeros 75 años del siglo XX; Martín & Lorenzo, 2001), en este estudio sus máximas densidades no se han medido en hábitats autóctonos poco degradados, sino en medios fuertemente impactados por el hombre: mosaicos de cultivos, plataneras, pastizales y núcleos urbanos pequeños. Por otro lado, las últimas poblaciones de Calandrella rufescens de Tenerife (menos de 100 parejas) se asentaban en hábitats profundamente degradados (pastizales del aeropuerto de Los Rodeos, campo de golf, áreas de cultivo de tomate abandonados; Martín & Lorenzo, 2001). La casi total destrucción por el hombre del hábitat original de estas especies en el piso basal de la isla (zonas estepáricas o áreas termófilas de medianía) quizás las hizo desplazarse a unos hábitats secundarios muy alterados donde llegaron a ser abundantes y se encuentran hoy (áreas agrícolas). Por tanto, lo que observamos hoy día sería una disminución de sus tamaños poblacionales respecto a sus efectivos de hace unos 100–50 años, como consecuencia de los cambios en los usos del suelo (nuevas alteraciones y abandono de prácticas tradicionales; ver también Tucker & Heath, 1994 para la avifauna europea), pero muy posiblemente sus poblaciones ya venían mermadas desde un pasado. Esto no parece haber ocurrido con las especies más claramente endémicas cuyas preferencias de hábitat se establecen en medios autóctonos. Incluso especies estenoicas que en el pasado fueron consideradas muy escasas, hoy día aumentan sus efectivos paralelamente a la recuperación de sus hábitats. Este es el caso de las palomas endémicas Columba bollii y Columba junoniae propias de laurisilva y de los ya desaparecidos bosques termófilos (Martín & Lorenzo, 2001). También es el caso de otras dos especies propias de pinares autóctonos de Pinus canariensis: Dendrocopos major canariensis y Fringilla teydea teydea (Martín & Lorenzo, 2001). Durante los últimos 20 años, la densidad de Dendrocopos major ha aumentado sustancialmente en numerosas zonas de pinar donde antes no existía. Así, en pinares censados en 1986 por Carrascal (1987) donde la especie no estaba presente (pinares por encima del valle de la Orotava), en la actualidad se han medido densidades de 0,2–1 aves/10 ha. Las únicas especies

Fig. 3. Árboles de regresión para 8 especies / subespecies endémicas ampliamente distribuidas en Tenerife. Los criterios hacen referencia a las ramas de la izquierda (los valores son contrarios en las ramas derechas). La longitud de las ramas es proporcional a la devianza explicada por cada criterio (i.e., a mayor longitud, mayor variabilidad explicada). Los valores de las puntas expresan la abundancia de las especies en aves/km. Para el significado de las siglas véase la tabla 3. Fig. 3. Regression trees analyzing the spatial variation of abundance in 8 endemic species/subspecies widely distributed in Teneriffe. The split criteria refer to the left side of each dichotomy. Branch lengths are proportional to deviance explained by each split criterium. Figures in the tips of the trees (‘leaves’) measure the abundance expresed in birds per 1 km of line transect (without limiting the detection belt). See table 3 for acronyms.

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Animal Biodiversity and Conservation 28.2 (2005)

Parus caeruleus

Phylloscopus canariensis CARB<28

LAT<3136

BCO=0

HARB<16.5 LAT<3153

CMAT<26 LONG<385 COD=0 2.7 CMAT<78 ALT<1974 ALT<174 0.05 0.2 1.6

7.1

4.6

7.8 LONG<328

CMAT<78 4.3

4.6

6.0

LONG<330 1.4

LONG<316

10.6

4.5

HMAT<1.9 AGR=0

0.6

10.2 11.4 4.1

11.4

Serinus canarius

Turdus merula BRZ=0

AGR=0

CMAT<24 0.5

19.4

CMAT<55

9.1

LONG<351

LONG<340 3.6

1.1

LAT<3135 LAT<3135 ALT>683 6.1 7.4

3.0

PIN=0

CMAT<68 AGR=0 LONG<317 ALT<486 4.9 1.8 0.7 1.3 7.7

0.4

18.1

Anthus berthelotii

LAT<3157 10.7 15.1 11.2

2.3

Regulus teneriffae HARB<4.2

BRZ=0

LAT<3116 0.3

CHB<32

CHB<68

LONG<332 2.3

CARB<48 0.1

9.8

5.4

LONG<345

HARB<16.5 1.4

3.4

1.4 6.8

LAT<3144 HMAT<1.7 2.4

2.7

BRZ=0

PIN=0 ALT<1594

PIN=0

0.0 0.2

7.9

HARB<13.5

LAT<3158 HARB<6.5 1.7

15.2

Fringilla teydea

Fringilla coelebs

AGR=0

LONG<375 7.2

HARB<9.5 0.0

1.5

0.5

CARB<55 1.1

LONG<350

2.1

3.6

LAT<3143

0.01 2.5

0.6

LONG<346 2.6 2.0

5.2

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endémicas de hábitats autóctonos que en la actualidad son mucho más escasas que en el pasado son Burhinus oedicnemus distinctus y Bucanetes githagineus amantum. Este declive se ha debido principalmente a la alteración de hábitats semidesérticos cercanos a la costa del sur de la isla (urbanismo y monocultivos industriales de plátano y tomate; Martín & Lorenzo, 2001). A pesar de su carácter artificial, producto de la degradación de formaciones vegetales autóctonas, las áreas agrícolas de Tenerife mantienen altos efectivos poblacionales de numerosas especies de aves en Tenerife. Así, Motacilla cinerea y Streptopelia turtur (Tórtola Europea) alcanzan en estos ambientes densidades tan altas como las observadas en formaciones autóctonas con casi nula influencia humana. Los escasos pastizales del norte de la isla sustentan las poblaciones más densas de Carduelis cannabina y de algunas otras especies bastante raras en Tenerife: Miliaria calandra, Petronia petronia y Carduelis chloris. Los mosaicos agrícolas de norte de la isla, mantienen áreas de praderas suficientemente extensas como para albergar las últimas poblaciones densas de Coturnix coturnix. Las plataneras y áreas semi–urbanas dispersas del norte de la isla acogen los mayores efectivos de Carduelis carduelis. Las zonas urbanas de Tenerife se caracterizan por acoger una considerable riqueza y diversidad de especies de aves. Así, los valores urbanos de estos parámetros son similares o incluso más altos que los medidos en formaciones vegetales autóctonas (i.e., matorrales de montaña, tabaibales–cardonales, laurisilva y pinares maduros), y atípicos si los comparamos con lo observado en áreas continentales, tanto del paleártico como del neártico (Melles et al., 2003; Marzluff et al., 2001; Clergeau et al., 1998). Este fenómeno se puede explicar atendiendo a la permeabilidad que tienen las zonas urbanas tinerfeñas para captar numerosas especies de la avifauna de su entorno. Así, las ciudades y pueblos de Tenerife acogen elevadas densidades de especies que en Europa son raras en medios urbanizados (e.g., Streptopelia turtur, Upupa epops, Motacilla cinerea), o menos frecuentes que lo observado en Tenerife (e.g. Turdus merula, Parus caeruleus, Phylloscopus canariensis, Sylvia atricapilla y Serinus canarius; Fernández–Juricic, 2000; Jokimäki, 1999; Palomino y Carrascal, en preparación). Este hecho incrementa la variabilidad, y con ello la diversidad y riqueza, del grupo de especies más típicamente urbanas (i.e., Columba livia, Passer hispaniolensis, y las recientes colonizadoras durante los últimos 30 años Streptopelia decaocto, S. roseogrisea, Carduelis chloris y Sturnus vulgaris–Estornino Pinto). Considerando estos patrones de distribución, proponemos una hipótesis que vincula el carácter de endemicidad de las poblaciones de aves de Tenerife con las preferencias de hábitat de las especies, atendiendo a la ocupación preferente de medios con fuerte impacto humano frente a aquellos autóctonos no degradados. Esta hipótesis predice que es más probable que hoy existan poblaciones tinerfeñas de especies claramente diferenciadas (i.e., subespecies

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o especies distintas) si su distribución y máximos de abundancia coinciden con formaciones vegetales climácicas (retamares altimontanos, matorrales montanos, tabaibales–cardonales, pinares maduros y laurisilva) y no penetran medios antrópicos a no ser que tengan elevadas amplitudes de hábitat. Por el contrario, sería poco probable que especies que son muy escasas o están ausentes de hábitats autóctonos poco modificados y cuyas densidades ecológicas máximas están en medios producto de la degradación ambiental (áreas agropecuarias y urbanas) tuviesen poblaciones en Tenerife que hayan subespeciado respecto a las poblaciones continentales. La base teórica de este fenómeno debe encontrarse en la hipótesis del "ciclo del taxón" (Ricklefs & Cox, 1978; Williamson, 1981; Ricklefs & Bermingham, 1999), que postula que las poblaciones insulares tienden a lo largo del tiempo a la diferenciación especializándose en ambientes isleños concretos (ver no obstante Wiens, 1989 para una controversia respecto a esta hipótesis). Esto se debe a que la diferenciación taxonómica de poblaciones implica unas escalas de tiempo muy superiores a las implicadas en la transformación histórica de hábitats naturales por el hombre (últimos 2.000 años, frente a las decenas de miles de años requeridos en procesos de especiación en aves –Klicka & Zink, 1997). Ejemplos del primer grupo de especies (endémicas distribuidas en medios autóctonos) serían Accipiter nisus granti, Burhinus oedicnemus distinctus, Dendrocopos major canariensis, Erithacus rubecula superbus, Regulus [regulus] teneriffae, Fringilla coelebs canariensis, Phylloscopus [collybita] canariensis, Parus caeruleus teneriffae. La única excepción a este patrón parece ser Motacilla cinerea canariensis (muy probablemente debido a la desaparición de casi todos los cursos de agua naturales en los barrancos debido a las canalizaciones actuales; Martín & Lorenzo, 2001). Evidencias del segundo grupo de especies serían aquellas cuyo estatus taxonómico de subespecies endémicas quedan descartados considerando recientes análisis de taxonomía molecular (Calandrella rufescens rufescens, Coturnix coturnix confisa–Consejería de Política Territorial y Medio ambiente, J. M. Naranjo, com. pers.). La validez y grado de generalización de esta hipótesis podrá ser comprobado una vez que sean desarrollados estudios taxonómicos exhaustivos con sólidas bases moleculares y morfométricas. La combinación del conocimiento del estatus taxonómico de las poblaciones insulares, junto con sus preferencias de hábitat y ocupación de medios autóctonos vs. degradados por la acción humana servirá para definir prioridades de conservación a escala del archipiélago (Dennis, 1997; Thomas et al., 1999; Gordon & Ornelas, 2000; Sangster, 2000). Tenerife vs. península ibérica: compensación de densidades y amplitud de distribución altitudinal Debido a la menor cantidad de especies presentes en las islas que en el continente, varios autores han postulado que en las islas disminuye la presión

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de depredación (i.e., menos depredadores presentes) y se reduce el solapamiento interespecífico de los nichos ecológicos, mitigándose considerablemente el efecto de la competencia interespecífica. En este escenario de reducción de potenciales depredadores y competidores, las poblaciones de especies insulares aumentarían su abundancia respecto al continente, produciéndose el fenómeno denominado "compensación de densidad". No obstante, este incremento de abundancia también podría darse como consecuencia de que la dispersión en las islas está limitada debido al efecto barrera impuesto por el mar y la existencia de hábitats desfavorables dentro de las islas (Diamond, 1970; Yeaton, 1974; Emlen, 1979; Crowell, 1983; Wiens, 1989). Este fenómeno ha sido observado en el suroeste del Paleártico Occidental con la avifauna de Córcega (Blondel et al., 1988). Utilizando los valores de densidades ecológicas máximas obtenidos en este trabajo (tabla 3), y los obtenidos para la península ibérica (Martí & Del Moral, 2003; textos de cada especie y datos del apéndice 1) se puede comprobar si esta hipótesis se cumple de modo generalizado con la avifauna de Tenerife. Las únicas especies que alcanzan densidades máximas sustancialmente más elevadas en Tenerife que en la península ibérica (valores expresados en aves/10 ha) son: Anthus berthelotii (9,1 frente a 2,4 de Anthus campestris–Bisbita Campestre; ver Voelker, 1999 para una justificación del parentesco filogenético entre estas dos especies), Sylvia conspicillata (3,7 vs. 2.08), Petronia petronia (15,7 vs. 1,7) y Serinus canarius (22,9 vs. 13,3 de Serinus serinus–Verdecillo en la península). Las densidades ecológicas máximas en Tenerife de las especies/subespecies endémicas Lanius excubitor [meridionalis] koenigi (Alcaudón Real; 0,77 aves/ 10 ha), Regulus teneriffae (17,1), Phylloscopus canariensis (15,5) y Carduelis cannabina meadewaldoi (7,5) son muy parecidas a las obtenidas por sus equivalentes congenéricos en Iberia (Lanius excubitor [meridionalis] 0,74 aves/10 ha; Regulus regulus– Reyezuelo Sencillo 14,8 aves/10 ha, ver Sturmbauer et al., 1998 para una justificación del parentesco filogenético con Regulus teneriffae; Phylloscopus ibericus–Mosquitero Ibérico 13,4 aves/10 ha; Carduelis cannabina: 7,7 aves/10 ha). Otras especies o subespecies tinerfeñas claramente diferenciadas de las formas continentales muestran densidades considerablemente menores en Tenerife (valores expresados en aves/10 ha en Tenerife vs. península ibérica): Dendrocopos major canariensis (1,78 vs. 2,4 como media de los tres valores máximos medidos en Iberia), Motacilla cinerea canariensis (1,8 vs 3,0), Erithacus rubecula superbus (11,6 vs. 22,2), Turdus merula cabrerae (14,3 vs 19,6), Sylvia melanocephala (2,4 vs. 15,3), Parus caeruleus teneriffae (8,4 vs. 25,5), Fringilla coelebs canariensis y F. teydea (4,3 y 6,9, respectivamente, vs. 23,2 de F. coelebs ibéricos). Esto mismo ocurre con especies de Tenerife no claramente diferenciadas taxonómicamente de sus equivalentes ibéricos: Sylvia atricapilla (3,7 vs. 12,2), Carduelis carduelis (2,7 vs. 7,8) y Carduelis chloris

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(1,1 vs. 10,4). Por tanto, los datos no parecen apoyar de modo generalizado la hipótesis de compensación de densidades en la avifauna de Tenerife. Un paradigma clásico de la biogeografía insular es la expansión del nicho ecológico de las especies en las islas respecto al continente, aunque existen evidencias tanto a favor como en contra de esta hipótesis (véase Blondel, 1979; Wiens, 1989 y referencias allí dadas). Para la isla de Córcega, Blondel et al. (1988) y Prodon et al. (2002) encuentran evidencias tanto apoyando como rechazando esta hipótesis: incremento de la amplitud de hábitat en algunas especies (p.e., Parus spp.), y contracción generalizada en la amplitud de distribución altitudinal, aunque las subespecies endémicas manifiestan un incremento de ésta, expandiéndose por las zonas elevadas de la isla. La ampliación del espectro de hábitats ocupados es un hecho muy característico en unas pocas especies (Parus caeruleus, Phylloscopus canariensis, Regulus teneriffae, Sylvia atricapilla, Serinus canarius; tabla 1), que ocupan distintos tipos de bosques, tabaibales, áreas agrícolas y zonas urbanas. Por otro lado, se produce una expansión de las preferencias de hábitat de las aves forestales hacia medios estructuralmente más simples (caso de Parus caeruleus, Phylloscopus canariensis y Sylvia atricapilla), aspecto que también se ha observado en la avifauna de Córcega (ver Blondel, 1979; Blondel et al., 1988 para una discusión de este tema). No obstante, en el resto de las especies no se observa tal incremento de su amplitud de hábitat en Tenerife (ver en Martí & Del Moral, 2003 los gráficos de distribución entre hábitats de las especies de Passeriformes en la península ibérica). En Tenerife destacan los enormes rangos altitudinales de algunas especies que se distribuyen desde el nivel del mar hasta 2.500 m de altitud (e.g., Falco tinnunculus, Anthus berthelotii, Lanius excubitor [meridionalis], Erithacus rubecula, Turdus merula, Phylloscopus canariensis, Sylvia conspicillata, Parus caeruleus; ver tabla 2 y Martín, 1987). Las amplitudes de distribución altitudinal de estas especies son muy grandes y aparentemente mayores a las observadas en la península ibérica (ver para comparación Sánchez, 1991; Pleguezuelos, 1992; Martí & Del Moral, 2003). Todas estas especies que manifiestan grandes rangos de distribución son subespecies o especies endémicas, algo parecido a lo obtenido por Prodon et al. (2002) en Córcega. Otra especie que presenta en Tenerife una gran distribución altitudinal, superior a la observada en la península es Streptopelia turtur (ver datos en la tabla 2 y en Sánchez, 1991; Pleguezuelos, 1992). Sin embargo, hay otras especies y subespecies endémicas que muestran una reducida dispersión altitudinal ya que establecen sus preferencias de hábitat en formaciones ambientales muy localizadas (e.g, Burhinus oedicnemus y Bucanetes githagineus en tabaibales xéricos; Columba bollii, Columba junoniae y Fringilla coelebs en laurisilvas; Dendrocopos major y Fringilla teydea en pinares; Miliaria calandra y Coturnix

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coturnix en praderas húmedas bajo influencia de los alisios). Otras especies distribuidas altitudinalmente de modo relativamente amplio en la península ibérica, tienen en Tenerife un rango altitudinal bastante restringido (e.g., Upupa epops, y las tres especies del género Carduelis; véase para comparación Pleguezuelos, 1992). Por tanto, los resultados obtenidos en Tenerife no parecen apoyar de modo claro y generalizado la hipótesis de expansión de distribución en las poblaciones insulares. (véase Prodon et al., 2002 para un resultado similar obtenido con la avifauna de Córcega). Agradecimientos Este manuscrito se ha beneficiado de los comentarios de Alfredo Valido, Álvaro Ramírez y Mario Díaz. Referencias Anónimo, 1980. Atlas Básico de Canarias. Interinsular Canaria, Sta. Cruz de Tenerife. Báez, M., 1992. Zoogeography and evolution of the avifauna of the Canary Islands. Natural History Museum of Los Angeles County Science Series, 36: 425–431. Bibby, C. J., Burgess, N. D., Hill, D. A. & Mustoe, S. H., 2000. Bird census techniques (2nd edition). Academic Press, London. Blondel, J., 1979. Biogéographie et ecologie. Masson, Paris. Blondel, J., Chessel, D. & Frochot, B., 1988. Bird species impoverishment, niche expansion, and density inflation in Mediterranean island habitats. Ecology, 69: 1899–1917. Brown, J. H. & Lomolino, M. V., 1998. Biogeography. Sinauer, Sunderland. Carrascal, L. M. & Palomino, D., 2002. Determinantes de la riqueza de especies de aves en las islas Selvagem y Canarias. Ardeola, 49: 211–221. Carrascal, L. M., 1987. Relación entre avifauna y estructura de la vegetación en las repoblaciones de coníferas de Tenerife (Islas Canarias). Ardeola, 34: 193–224. Carrascal, L. M., Tellería, J. L. & Valido, A., 1992. Habitat distribution of Canary chaffinches among islands: competitive exclusion or species–specific habitat preferences? Journal of Biogeography, 19: 383–390. Clergeau, P., Savard, J. P. L., Mennechez, G. & Falardeau, G., 1998. Bird abundance and diversity along an urban–rural gradient: a comparative study between two cities on different continents. Condor, 100: 413–425. Collar, N. J., Crosby, M. J. & Stattersfield, A. J., 1994. Birds to Watch 2. Birdlife International– Smithsonian Institution Press, Washington, D.C. Crowell, K. L., 1983. Islands–insight or artefacts? Population dynamics and habitat utilization in

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Martín, A. & Lorenzo, J. A., 2001. Aves del archipiélago canario. Francisco Lemus Editor, La Laguna. Martín, A., 1987. Atlas de las aves nidificantes de la isla de Tenerife. Instituto de Estudios Canarios, Sta. Cruz de Tenerife. Martin, J–L., 1992. Niche expansion in an insular bird community: an autoecological perspective. Journal of Biogeography, 19: 375–381. Marzluff, J. M., Bowman, R. & Donnelly, R., 2001. Avian conservation and ecology in an urbanizing world. Kluwer Academic Publications, Boston. Melles, S., Glenn, S. & Martin, K., 2003. Urban bird diversity and landscape complexity: species–environment associations along a multiscale habitat gradient. Conservation Ecology, 7: 5. [online] URL: http://www.consecol.org/vol7/iss1/art5 Milberg, P. & Tyrberg, T., 1993. Naive birds and noble savages–a review of man–caused prehistoric extinctions of island birds. Ecography, 16: 229–250. Nogales, M., Medina, F. M., Quilis, V. & Gonzáles– Rodríguez, M., 2001. Ecological and biogeographical implications of Yellow–Legged Gulls (Larus cachinnans Pallas) as seed dispersers of Rubia fruticosa Ait. (Rubiaceae) in the Canary Islands. Journal of Biogeography, 28: 1137–1145. Nogales, M., Quilis, V., Medina, F. M., Mora, J. L. & Trigo, L. S., 2002. Are predatory birds effective secondary seed dispersers? Biological Journal of the Linnean Society, 75: 345–352. Pimm, S. L., Jones, H. L. & Diamond, J. M., 1988. On the risk of extinction. American Naturalist, 132: 757–785. Pleguezuelos, J. M., 1992. Avifauna nidificante de las sierras Béticas orientales y depresión de Guadix, Baza y Granada. Monográfica Tierras del Sur, Univ. de Granada. Prodon, R., Thibault, J–C. & Dejaifve, P–A., 2002. Expansion vs. compression of bird altitudinal ranges on a Mediterranean island. Ecology, 83: 1294–1306. Rando, J. C., 2002. New data of fossil birds from El Hierro (Canary Islands): probable causes of extinction and some biogeographical considerations. Ardeola, 49: 39–49. Rando, J. C., López, M. & Seguí, B., 1999. A new species of extinct flightless Passerine (Emberizidae: Emberiza) from the Canary Islands. Condor, 101: 1–3. Ricklefs, R. E. & Bermingham, E., 1999. Taxon cycles in the Lesser Antillean avifauna. Ostrich, 70: 49–59. Ricklefs, R. E., & Cox, G. W., 1978. Stage of taxon

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cycle, habitat distribution, and population density in the avifauna of the West Indies. American Naturalist, 112: 875–895. Sánchez, A., 1991. Estructura y estacionalidad de las comunidades de aves de la Sierra de Gredos. Ardeola, 38: 207–231. Sangster, G., 2000. Taxonomic stability and avian extinctions. Conservation Biology, 14: 579–581. Stattersfield, A. J., Crosby, M. J., Long, A. J. & Wege, D. G., 1998. Endemic Bird Areas of the World: Priorities for Biodiversity Conservation. Birdlife International–Smithsonian Institution Press, Washington, D.C. Sturmbauer, C., Berger, B., Dallinger, R. & Föger, M., 1998. Mitochondrial phylogeny of the genus Regulus and implications of the evolution of breeding behaviour in sylvoid songbirds. Molecular Phylogenetics and Evolution, 10: 144– 149. Suárez, F., 1984. Estructura y composición de las comunidades de aves invernantes en las zonas semiáridas de Lanzarote y Fuerteventura (Islas Canarias). Ardeola, 30: 83–91. Thomas, C. D., Glen, S. W. T., Lewis, O. T., Hill, J. K. & Blakeley, D. S., 1999. Population differentiation and conservation of endemic races: the butterfly, Plebejus argus. Animal Conservation, 2: 15–21. Tucker, G.M., & Heath, M.F. 1994. Birds in Europe: their conservation status. Birdlife International (Birdlife Conservation Series No. 3), Cambridge, U.K. UICN, 2001. IUCN Red List Categories: Version 3.1 IUCN. Species Survival Commission (http:// www.iucn.org/themes/ssc/redlists/ssc–rl–c.htm). Valido, A., Tellería, J. L. & Carrascal, L. M., 1994. Between and within habitat distribution of the Canary Common Chaffinch (Fringilla coelebs ombriosa): a test of the food abundance hypothesis. Ardeola, 41: 29–35. Voelker, G., 1999. Molecular evolutionary relationships in the avian genus Anthus (Pipits: Motacillidae). Molecular Phylogenetics and Evolution, 11: 84–94. Wiens, J. A., 1989. The ecology of bird communities. Vol. 1. Cambridge Univ. Press, Cambridge. Williamson, M., 1981. Island populations. Oxford Univ. Press, Oxford. Wilson, E. O., 1961. The nature of the taxon cycle in the Melanesian ant fauna. American Naturalist, 95: 169–193. Yeaton, R. I., 1974. An ecological analysis of chaparral and pine forest bird communities on Santa Cruz island and mainland California. Ecology, 55: 959–973.

"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

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Minimizing extinction risk through genetic rescue T. A. Waite, J. Vucetich, T. Saurer, M. Kroninger, E. Vaughn, K. Field & S. Ibargüen

Waite, T. A., Vucetich, J., Saurer, T., Kroninger, M., Vaughn, E., Field, K. & Ibargüen, S., 2005. Minimizing extinction risk through genetic rescue. Animal Biodiversity and Conservation, 28.2: 121–130. Abstract Minimizing extinction risk through genetic rescue.— According to the genetic rescue hypothesis, immigrants can improve population persistence through their genetic contribution alone. We investigate the potential for such rescue using small, inbred laboratory populations of the bean beetle (Callosobruchus maculatus). We ask how many migrants per generation (MPG) are needed to minimize the genetic component of extinction risk. During Phase 1, population size was made to fluctuate between 6 and 60 (for 10 generations). During this phase, we manipulated the number of MPG, replacing 0, 1, 3, or 5 females every generation with immigrant females. During Phase 2, we simply set an upper limit on population size (.10). Compared with the 0–MPG treatment, the other treatments were equivalently effective at improving reproductive success and reducing extinction risk. A single MPG was sufficient for genetic rescue, apparently because effective migration rate was inflated dramatically during generations when population size was small. An analysis of quasi–extinction suggests that replicate populations in the 1–MPG treatment benefited from initial purging of inbreeding depression. Populations in this treatment performed so well apparently because they received the dual benefit of purging followed by genetic infusion. Our results suggest the need for further evaluation of alternative schemes for genetic rescue. Key words: Extinction risk, Founder events, Genetic rescue, Inbreeding. Resumen Minimización del riesgo de extinción mediante el rescate genético.— Según la hipótesis del rescate genético, los inmigrantes pueden mejorar la persistencia de una población mediante su contribución genética. Hemos investigado el potencial de un rescate de este tipo, utilizando pequeñas poblaciones endogámicas de laboratorio del gorgojo del haba Callosobruchus maculatus. Nos preguntamos cuántos migrantes por generación (MPG) son necesarios para minimizar el componente genético del riesgo de extinción. Durante la Fase 1, se hizo fluctuar el tamaño de la población entre 6 y 60 (durante 10 generaciones). En dicha fase manipulamos el número de MPGs, reemplazando 0, 1, 3, o 5 hembras nativas por hembras inmigrantes en cada generación. Durante la Fase 2, nos limitamos a poner un límite superior al tamaño de la población (.10). Comparados con el tratamiento de 0–MPG, los otros tratamientos resultaron ser igualmente efectivos en la mejora del éxito reproductivo y la reducción del riesgo de extinción. Un único MPG era suficiente para el rescate genético, aparentemente debido a que la tasa de migración efectiva quedaba espectacularmente sobredimensionada durante generaciones, cuando el tamaño de la población era pequeño. Un análisis de cuasi–extinción sugiere que las poblaciones replicadas durante el tratamiento 1–MPG se beneficiaron de un saneamiento inicial por la disminución de la endogamia. Aparentemente, las poblaciones de este tratamiento se comportaron tan bien debido a que recibieron el doble beneficio del saneamiento seguido de la inyección genética. Nuestros resultados sugieren la necesidad de posteriores evaluaciones del rescate genético mediante esquemas alternativos. Palabras clave: Riesgo de extinción, Acontecimientos de hundimiento, Rescate genético, Endogamia. (Received: 19 VIII 04; Conditional acceptance: 8 X 04; Final acceptance: 9 XI 04) T. A. Waite, T. Saurer, M. Kroninger, E. Vaughn, K. Field & S. Ibargüen, Dept. of Evolution, Ecology, and Organismal Biology, Ohio State Univ., Columbus, Ohio 43210–1293, U.S.A.– J. Vucetich, School of Forest Resources and Environmental Science, Michigan Technological Univ., Houghton, Michigan 49931, U.S.A. Corresponding author: T. A. Waite. E–mail: waite.1@osu.edu

ISSN: 1578–665X

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Introduction Small isolated populations are subject to loss of genetic diversity through drift and inbreeding. Despite a large body of findings implicating inbreeding as a contributor to extinction risk (reviewed by Hedrick & Kalinowski, 2000), the strength of any causal linkage between inbreeding and extinction remains a point of contention. Until recently, there was no direct evidence that genetic deterioration contributes to extinction of wild populations (Frankham & Ralls, 1998). Lacking such evidence, some workers have argued that stochastic demographic and environmental events may typically drive small populations to the brink of extinction before genetic deterioration poses a serious threat (Lande, 1988; Pimm et al., 1988; Caro & Laurenson, 1994). Even so, there is widespread agreement that loss of genetic diversity can lead to extinction. Support for this perspective comes from theoretical studies (Mills & Smouse, 1994; Lande, 1998; Tanaka, 2000; Finke & Jetschke, 1999; Fowler & Whitlock, 1999), laboratory experiments (Frankham, 1995a; Bryant et al., 1999; Bijlsma et al., 2000; Reed & Bryant, 2000; Nieminen et al., 2001), field experiments (e.g., Newman & Pilson, 1997), a landmark study of a metapopulation in nature (Saccheri et al., 1998), and meta–analyses (Frankham, 1999). A recent review summarizes evidence, based on new pedigree data and new data made possible by molecular and analytical tools for estimating inbreeding, that inbreeding can adversely affect population performance (Keller & Waller 2002; see also Goudet & Keller 2002). Meanwhile, a flurry of recent experimental (e.g., Bryant et al., 1999; Reed & Bryant, 2000, 2001; Newman & Tallmon, 2001) and theoretical studies (e.g., Fu et al., 1998; Bataillon & Kirkpatrick, 2000; Kirkpatrick & Jarne, 2000; Wang, 2000; Whitlock, 2000; Linklater, 2003) have explored ways to minimize the genetic component of extinction risk. What kind of genetic intervention, if any, is needed? Ideally, genetic risks could be minimized without intervention, simply by maintaining populations above minimum viable size (reviewed by Reed & Bryant, 2000; see also Lande, 1995; Lynch et al., 1995; Gilligan et al., 1997). However, when this approach is not feasible or has already failed, genetic diversity can be maintained or restored by facilitating gene flow via translocation of individuals or propagules (e.g., Madsen et al., 1999). Because the mere arrival of immigrants could forestall local extinction, to demonstrate unequivocally that gene flow per se is beneficial, one must perform experiments in which genetic diversity is introduced without a simultaneous increase in population size. Recent studies have sought to provide evidence for such genetic rescue (i.e., increase in fitness due to gene flow) of recently fragmented or newly colonized populations (reviewed by Ingvarsson, 2001; see also Vila et al., 2003). For example, pollen–mediated gene flow improved fitness in small populations of a dioecious weedy

plant (Silene alba) (Richards, 2000). Likewise, gene flow via immigration improved various fitness components in the self–incompatible mustard (Brassica campestris) (Newman & Tallmon, 2001), and improved fitness and reduced extinction risk in the house fly (Musca domestica) (Bryant et al., 1999). Lastly, gene flow facilitated by an alien pollinator (African honeybee, Apis mellifera scutellata) is apparently responsible for improved reproductive output in an Amazonian tree (Dinizia excelsa [Fabaceae]) in pastures and forest remnants, where native pollinators are absent (Dick, 2001). Although some earlier studies provided contradictory findings (references in Newman & Tallmon, 2001), these recent studies indicate that pollen– or immigrant–mediated gene flow can dramatically improve fitness in small inbred populations. Here, we describe an experiment that extends these recent findings. Using inbred laboratory populations of the bean beetle (Callosobruchus maculatus), we manipulated the number of MPG by replacing 0, 1, 3, or 5 females with immigrant females each generation. The experiment allowed us to evaluate: (1) whether even a single MPG could lead to genetic rescue, and (2) how many migrants are needed to minimize the genetic component of extinction risk. Methods Subjects C. maculatus is an important pest species. The beetles used in our experiment were derived from a genetic strain from southern India and reared at Ohio State University. Several features make this species a suitable model organism (e.g., Vucetich et al., 2000): (1) it has a short generation time (4–6 weeks); (2) females oviposit on beans and offspring emerge synchronously, with the adults typically dying before the next generation emerges; and (3) because only one beetle typically emerges from each mung bean (Vigna radiata), carrying capacity can be controlled simply by limiting the number of beans available. Overview and rationale The experiment was designed to quantify the requisite number of MPG to minimize extinction risk in small inbred populations. It comprised two phases. During Phase 1, population size was made to fluctuate between 6 and 60 individuals across 10 generations. In each generation, the &ð:%ð sex ratio was 5:1. During this phase, we manipulated the number of MPG by replacing 0, 1, 3, or 5 females every generation with immigrant females from a large outbred population. At the end of this phase, we measured the reproductive fitness and founding success of each replicate population. During Phase 2, we limited N by simply providing 10 mung beans to each replicate population for 10 generations. During

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this phase, we documented extinctions. The rationale for each of these procedures is described below. Most tests of the genetic rescue hypothesis have treated the infusion of new genetic material dichotomously, testing the effect of a single level of infusion versus no infusion. In an attempt to titrate the level of genetic infusion minimizing extinction risk, we assessed the effects of several levels on reproductive fitness and extinction risk. To avoid confounding the results with effects attributable to demographic rescue, these migrants were replacements, not additions. Due to fluctuations in population size (FPS) (Vucetich et al., 1997; Vucetich & Waite, 1999), skewed sex ratio, and extra–Poisson variation in fecundity, most real populations exhibit Ne/N ratios that are less than unity (Frankham, 1995b). Therefore, during Phase 1, we manipulated the FPS and sex ratio to achieve Ne/N ratios typical of real populations. This approach resulted in an Ne/N ratio of - 0.2, which is close to the median of surveyed populations (Frankham, 1995b). The most straightforward way to perform genetic rescue is to infuse a population with genetic material for a pulse (i.e., one, two, or a few generations). In Phase 2 of the experiment, we assess the residual impact of genetic infusion. That is, we assess the effects of prior genetic management (imposed during Phase 1) on extinction risk. Detailed protocol Preliminary steps To establish replicate inbred populations, we began by orchestrating two successive full–sibling matings. Eighty female–male pairs, representing 80 unique pairs of founders, were used. Offspring of these pairs comprised the parental generation. The inbreeding coefficient, F, in these progeny was 0.375. This procedure served several purposes. First, we were interested in investigating the effectiveness of genetic rescue of already–inbred populations. Second, we intended to purge the genetic load such that further purging would not confound our results. Finally, our prior work (unpubl. results) showed that additional full–sibling matings would push F beyond the extinction quasi–threshold (Frankham, 1995a). To establish large outbred source populations of potential immigrants, we created five populations each comprising - 5,000 individuals. Through Phase 1(described below), we housed these five populations separately. Because the timing of emergence in the five source populations diverged over time, this procedure was used to ensure a continuous supply of immigrants. Phase 1 We placed a female and a male with the same full– sibling parents into each of 80 petri dishes, each containing 40 pristine (eggless) mung beans. Following oviposition, we placed each of egg–laden bean in a separate Eppendorf tube. As adults

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emerged, we placed one male and a specified number of females (5, 4, 2, or 0 in the 0–, 1–, 3–, and 5–MPG treatments, respectively) from the same replicate population along with a complimentary number of immigrant females (0, 1, 3, or 5) into petri dishes containing 250 pristine beans. For example, in each replicate in the 1–MPG treatment, one male was placed together with four females from the same replicate population along with one immigrant female. Thus, each of 80 dishes (20 replicates in each treatment) contained five females and one male in all odd–numbered generations. On the 21st day after the adults had been put together, any surviving adults were removed and each egg– laden bean was placed in a separate tube. Upon emergence, we repeated the protocol, except that 10 males were put together with 50, 49, 47, or 45 females (in the 0–, 1–, 3–, and 5–MPG treatments, respectively) from the same replicate population and with 0, 1, 3, or 5 immigrant females. Each replicate population thus comprised 60 adults (50 females and 10 males) in generation 2 (and all even–numbered generations in Phase 1). We then repeated the above alternation between N = 6 in odd generations and N = 60 in even generations through the 10th generation. Throughout this phase, a pool of immigrant females was kept available by placing egg–laden beans from the source population singly into 200 tubes every generation. By matching female immigrants by date of 4th emergence in candidate recipient populations, we ensured that female immigrants were approximately the same age (i.e., within 7 days) as most members of the recipient population. Any potential female immigrant not assigned within two weeks following her emergence was excluded. Females satisfying the criteria for inclusion were transferred to appropriate populations according to the following rules. Egg– laden beans (one in each of # 250 tubes in each replicate) were monitored daily for onset of emergence. We designated the day of 4th emergence as Day 0. On Day 7, we determined whether at least one male had emerged. If so and if the criterion numbers of females and males had emerged, they were combined in a petri dish with the specified number of female immigrants. Mating was allowed to proceed. On Day 21, we transferred each egg– laden bean to a tube. If the criterion numbers of females and males had not been reached by Day 7, we placed the male(s) together with females (including immigrants) and added newly emerging individuals daily. This process continued until the criterion was met or until 3 consecutive days passed with no emergence. Then, 7 days after the last individual was added or 2 weeks after initially putting beetles together (whichever was longer), we transferred each egg–laden bean to a tube. Finally, if no males had emerged by Day 7, we waited until the first male emerged and then followed the just–described procedure. Some populations failed to reach the criterion numbers of adults, particularly in even–numbered genera-

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Proportion surviving

1.0

0.8 0.6 0.4 0.2 0.0

Proportion surviving

1.0

0.8 0.6 0.4 0.2 0.0 16 Generation

Fig. 1. Survivorship curves for small experimental populations of bean beetles (Calosobruchus maculatus) in four migrant–per–generation (MPG) treatments, with 20 replicate populations per treatment at the start of the experiment: A. Survivorship curves during Phase 1 of the experiment, when each population fluctuated between 6 individuals (1 male and 5 females) in odd generations and 60 individuals (10 males and 50 females) in even generations; B. Survivorship curves during Phase 2 of the experiment, when each population was subjected to an approximate carrying capacity of 10 individuals during every generation. Fig. 1. Curvas de supervivencia para pequeñas poblaciones experimentales del gorgojo del haba Calosobruchus maculatus, en tratamientos de cuatro migrantes por generación (MPG), con 20 poblaciones replicadas en cada tratamiento al inicio del experimento: A. Curvas de supervivencia durante la Fase 1 del experimento, cuando cada población fluctuaba entre 6 individuos (1 macho y 5 hembras) en las generaciones impares y 60 individuos (10 machos y 50 hembras) en las generaciones pares; B. Curvas de supervivencia durante la Fase 2 del experimento, cuando cada una de las poblaciones estaba sujeta a una capacidad de carga aproximada de 10 individuos por generación.

tions, when numerous populations failed to produce 45 females. We refer to these failures as quasi–extinctions. We compare the incidence of quasi–extinction between treatments and across generations during Phase 1, when very few true extinctions took place. Phase 2 After generation 10, replicate populations across all treatments were treated uniformly. Every population was subjected to a constant carrying capacity. No further immigration was orchestrated and no population variability was imposed. This phase lasted 10 generations. For each extant

population at the end of Phase 1, we placed all egg–laden beans (up to 210) in a large petri dish and then added 10 pristine beans. (Several replicates [10 in 0–MPG: 4 in 1–MPG , 2 in 3–MPG, 1 in 5–MPG] had gone extinct during Phase 1; others were lost to human error [2 in 0–MPG, 1 in 3–MPG].) Following oviposition, we discarded the original beans and placed 10 pristine beans in the dish with the 10 egg–laden beans. Following the next emergence and oviposition, we replaced the 10 old beans with 10 pristine beans. We repeated this process until extinction occurred or until the 20 th generation (10th in Phase 2). For each population, time to extinction (in genera-

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tions) was recorded. A population was judged to have gone extinct if: (1) no oviposition took place, (2) no beetles emerged, or (3) beetles of only one sex emerged. Fitness measurement To estimate individual reproductive success at the end of Phase 1, we began by randomly selecting 40 (of a possible 250) egg–laden beans from each population and placing two such beans in each of 20 tubes. Next, we monitored emergence and transferred five (whenever possible) female–male pairs to five petri dishes, each containing 100 pristine beans. We then allowed mating and oviposition to occur. Following emergence, we tallied the offspring produced by each pair. We used this quantity as our primary measure of fitness, but we also took advantage of the fact that some pairs failed to produce at least one adult offspring of each sex. We considered any such case to be a failed founding event. We thus compare both reproductive success and founding success across MPG treatments. Data analysis Survival analysis was performed using S–PLUS 2000 (1999). Kaplan–Meier nonparametric survival models were used to estimate mean time to extinction in each MPG treatment. Cox proportional hazards models were used to evaluate the effect of MPG treatment on risk of extinction. All pairwise comparisons were performed. Nominal P–values are reported, with an indication of whether each test is significant at the experimentwise ±–level of 0.05 (following Bonferroni correction). Other analyses were performed using SPSS (1999). Fisher’s exact tests were used to perform pairwise comparisons of incidence of extinction and quasi–extinction between MPG levels and between generations during Phase 1. One–way ANOVA was used to compare fitness (number of offspring produced per female–male pair) across MPG levels. Pairwise comparisons were made using Tukey’s HSD method. Finally, we compared the incidence (arcsin square root transformed proportion) of successful founding (production of at least one offspring of each sex) across MPG levels. Because the normality test failed (P < 0.001), we used Kruskal–Wallis nonparametric one–way "analysis of variance" on ranks. Pairwise comparisons were made using Dunn’s method, with the critical ±–level set at 0.05. Results Survival analysis Figure 1 shows the survival of replicate populations. No significant differences in incidence of extinction emerged among treatments by the end of Phase 1 (all Ps > 0.15, Fisher’s exact test), when few extinctions occurred (i.e., 8 of 67 populations). In

Table 1. Results of pairwise Cox proportional hazards comparisons (test of effect of manipulating number of migrants per generation). For each comparison, $ (= slope), exp($), and P are shown. Each of the first three comparisons is significant following Bonferroni adjustment for the number of pairwise tests performed (i.e., the nominal P– value is less than 0.05/6). Any negative value of $ corresponds with a value of exp($) < 1, which indicates the decrease in relative risk of extinction associated with a unit increase in the experimental factor MPG. Tabla 1. Resultados de las comparaciones al azar de Cox de los riesgos proporcionales por parejas (test del efecto de la manipulación del número de inmigrantes por generación). Para cada comparación se muestran $ (= pendiente), exp($), y P. Cada una de las tres primeras comparaciones es significativa según el ajuste de Bonferroni para el número de test por parejas llevados a cabo (es decir, el valor nominal de P es menor de 0.05/6). Cualquier valor negativo de $ se corresponde con un valor de exp($) < 1, lo que indica un descenso del riesgo relativo de extinción asociado con un incremento unitario del factor experimental MPG.

Between–treatment comparison $

exp($)

0– vs 1–MPG

–2.48

0.08

1.7×10–5

0– vs 3–MPG

–0.76

0.47

3.6×10–6

0– vs 5–MPG

–0.61

0.54

1.7×10–7

1– vs 3–MPG

0.10

1.10

0.65

1– vs 5–MPG

–0.002

0.99

0.98

3– vs 5–MPG

–0.12

0.88

0.53

Phase 2, mean persistence of populations in the 0– MPG treatment (6.4 generations) was substantially shorter than in every other treatment (1–MPG: 8.9, 3–MPG: 8.7, and 5–MPG: 8.9), based on Kaplan– Meier survival analysis. Cox proportional hazards analysis revealed that extinction risk declined significantly with increasing MPG ( $ = –0.23, exp[$ ] = 0.80 [95% CI: 0.68–0.94], P = 0.007), where exp($) quantifies the proportional effect of a unit increase in the experimental factor (MPG). Pairwise comparisons revealed significant effects of increasing MPG from 0 to any other level (i.e., 1, 3, or 5) (table 1). No other pairwise comparison was significant. Number of MPG was a significant predictor of extinction risk only when the 0–MPG treatment was included. The 1–, 3–, and 5–MPG treatments appeared to reduce extinction risk with equivalent effectiveness.

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Proportion quasi–extinction

0.8 0 1 3 5

0.6

MPG MPG MPG MPG

0.4

0.2

0.0 2

6 Generation

Fig. 2. The incidence of quasi–extinction (proportion of populations failing to produce at least 45 female offspring during even–numbered generations) in four migrant–per–generation treatments (MPG). Any extant population could experience quasi–extinction repeatedly. Fig. 2. Incidencia de la cuasi–extinción (proporción de poblaciones que no consiguen producir al menos 45 descendientes hembra durante las generaciones pares) en tratamientos de cuatro migrantes por generación (MPG). Cualquier población existente podría experimentar la cuasi–extinción repetidas veces.

Quasi–extinction analysis Figure 2 shows the incidence of quasi–extinction (defined as failure to produce at least 45 females in even–numbered generations during Phase 1). The incidence of quasi–extinction showed a pronounced temporal pattern in the 0–MPG treatment, decreasing from generation 2 to 4 (P = 0.003; Fisher’s exact test) and then increasing (generation 6 vs. 8: P = 0.009; generation 4 vs 8: P = 0.003) to the initial level (generation 2 vs 10: P = 1.0). This temporal pattern of quasi–extinction in the 0–MPG treatment suggests an initial purging of inbreeding depression followed by onset of inbreeding depression. An initial decline, from generation 2 to 4, was detectable in the 1–MPG treatment (Ps = 0.044 for comparisons between generation 2 vs 4, 6, 8, and 10), suggesting an initial purging of inbreeding depression with no subsequent onset of inbreeding depression by the end of Phase 1. Other pairwise comparisons were nonsignificant (Ps = 1.0). No significant between–generation differences in incidence of quasi–extinction emerged for either the 3– or 5–MPG treatment (Ps > 0.48), suggesting neither an initial purging of inbreeding depression nor a subsequent onset of inbreeding depression in these treatments. Within generations, several between–treatment differences emerged. In generation 2, quasi–extinction risk was reduced by the one–time immigration of a single female (i.e., incidence of quasi–extinction was lower in 1– than 0–MPG; P = 0.025).

Quasi–extinction risk was further reduced by the immigration of additional females (i.e., 1– vs 3– MPG: P = 0.008; 1– vs 5–MPG: P = 0.02). Incidence of quasi–extinction was minimized equivalently in the 3– and 5–MPG treatments. In generation 4, only one comparison (0– vs 5–MPG) was nominally significant. Within subsequent generations, incidence of quasi–extinction was higher in the 0–MPG treatment than in any other treatment (Ps < 0.001). All other pairwise comparisons were nonsignificant. Fitness analysis Figure 3A summarizes the results of the fitness test conducted at the end of Phase 1. Significant heterogeneity emerged across treatments (F3,43.7 = 11.86, P < 0.001), but not among replicates (F19,37.9 = 1.33, P = 0.22). Tukey’s HSD tests revealed that the number of offspring produced per female–male pair was significantly lower in the 0–MPG treatment than in any other treatment (all Ps < 0.001). Although a visual inspection of Figure 3A suggests a weak tendency for female–male pairs in the 5–MPG treatment to produce more offspring (mean = 55.0) than pairs in the 1–MPG (49.3) and 3–MPG treatments (48.6), neither of these comparisons was significant (1– vs 5–MPG: P = 0.20; 3– vs 5–MPG: P = 0.15) (nor was comparison between 1– and 3–MPG: P = 1.0). Figure 3B shows the founding success (i.e., proportion of pairs that produced at least one adult

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Reproductive success

0 0 1.0

Founding success

0.8 0.6 0.4 0.2 0.0 0

1 3 5 Migrants per generation

Fig. 3. A. Fitness measurements (mean number of offspring produced per female–male pair) in four migrant–per–generation (MPG) treatments at the end of Phase 1 of the experiment; the thin line within each box indicates the median, the thick line within each box indicates the mean, the box represents the interquartile interval (25th to 75th percentile), and the whiskers show the 10th and 90th percentiles. B. The founding success (defined as proportion of female–male pairs that produced at least one adult offspring of each sex) in four migrant–per–generation (MPG) treatments at the end of Phase 1 of the experiment. Symbols indicate means and the error bars represent standard errors. For both analyses, the numbers of replicate populations were as follows: 0–MPG, 8; 1–MPG, 16; 3–MPG, 17; and 5–MPG, 19 replicates. Fig. 3. A. Mediciones de aptitud (número promedio de descendientes producidos por cada par macho–hembra) en los tratamientos de cuatro migrantes por generación (MPG), al final de la Fase 1 del experimento; la fina línea del interior de cada rectángulo indica la mediana, y la línea gruesa la media, mientras que los rectángulos representan los intervalos intercuartiles (los percentiles 25 a 75); los extremos de las barras verticales muestran los percentiles 10 y 90. B. Éxito de fundación (definido como la proporción de parejas macho–hembra que produjeron al menos un descendiente adulto de cada sexo) en tratamientos de cuatro migrantes por generación (MPG) al final de la Fase 1 del experimento. Los símbolos indican las medias y las barras de error los errores estándar. Para ambos análisis, los números de poblaciones replicadas fueron los siguientes: 0–MPG, 8; 1–MPG, 16; 3–MPG, 17; y 5–MPG, 19.

offspring of each sex) of female–male pairs. Significant heterogeneity emerged among treatments (Kruskal–Wallis test: H3 = 24.75, P < 0.001). Dunn’s multiple–comparison procedure revealed that found-

ing success was significantly lower in the 0–MPG treatment than in any other treatment (qs > 3.54, Ps < 0.05). Other pairwise comparisons were nonsignificant.

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Discussion Our survival analysis suggests that the introduction of a small number of migrants per generation (MPG) was sufficient for genetic rescue of small, inbred populations of the bean beetle (fig. 1, table 1). Compared with the control (0–MPG treatment), all of the other treatments (1–, 3–, and 5–MPG) improved population persistence. Because immigrant females were replacements rather than extras, this finding may be attributable to beneficial effects of gene flow per se, perhaps including the masking of fixed deleterious mutations. That is, our manipulation apparently led to genetic rescue, as distinguished from demographic rescue. In agreement with other studies, our findings suggest that even a single MPG can lead to improved fitness (e.g., Newman & Tallmon, 2001; Vila et al., 2003) and reduced extinction risk (e.g., Bryant et al., 1999). Moreover, our results suggest that the extent of genetic rescue was independent of number of MPG, provided there was at least one MPG. Compared with the 0–MPG treatment, the other treatments were equivalently effective both at improving fitness (fig. 3) and reducing extinction risk (table 1). This finding is superficially puzzling in light of recent theoretical arguments that one actual migrant per generation will often be inadequate (Mills & Allendorf, 1996), particularly when the recipient population fluctuates (Vucetich & Waite, 2000). However, this theory does not apply here (Kalinowsky & Waples, 2002) because the populations did not merely fluctuate; they also fell to a very small size (in alternate generations). When this happens, the migration rate associated with a fixed number of migrants can be dramatically inflated (Vucetich & Waite, 2001). Because we did not determine parentage, we cannot use genealogical data to calculate the realized genetically effective migration rate each generation. However, as a first approximation, we estimate that the average migration rate, úm in the 1–, 3– and 5–MPG treatments was 4.5, 14, and 23 in odd generations (and 0.5, 1.7, and 2.8 in even generations), where ú is mean population size and m is actual number of immigrants divided by current size of the recipient population. Thus, it appears that the 1–MPG treatment performed so well because the rate of genetic infusion was adequate after all. Our analysis of quasi–extinction (fig. 2) suggests that populations in the 1–MPG treatment might have benefited also from initial purging of the genetic load (e.g., Fu et al., 1998; Fu, 1999; Wang, 2000; Reed & Bryant, 2001). Incidence of quasi– extinction in the 1–MPG treatment was high initially, but then decreased dramatically and remained low, suggesting initial purging followed by fitness– enhancing gene flow. By contrast, incidence of quasi–extinction in the 0–MPG treatment decreased initially but then increased, suggesting purging followed by onset of inbreeding depression in the absence of gene flow. Further evidence that 1– MPG populations benefited, in part, from gene flow

is provided by the observation that quasi–extinction risk was reduced by the first introduction of a single female (i.e., incidence of quasi–extinction was lower in 1–MPG than 0–MPG treatment in generation 2; fig. 2) (see also Spielman & Frankham, 1992). This result suggests that even a one–time immigration by a single individual can make a sufficient genetic contribution to provide a rescue effect (Ball et al., 2000; Vila et al., 2003). Our results also indicate, though, that quasi–extinction risk was further reduced by additional immigrants (fig. 2). Thus, the 1–MPG treatment did not perform as well as the 3– and 5–MPG treatments at this stage. Taken together, these results suggest a duel benefit for populations in the 1–MPG treatment: initial purging of inbreeding depression combined with subsequent fitness–enhancing gene flow. In summary, our results suggest that even a single MPG may sometimes be useful for genetic management of small, inbred populations. A single actual MPG may sometimes be sufficient, particularly if the recipient population is small (Vucetich & Waite, 2001; see also Kalinowsky & Waples, 2002) and if inbreeding depression is purged initially (e.g., Backus et al., 1995). The adequacy of one MPG could be further enhanced if offspring of immigrants exhibit heterosis (Ingvarsson & Whitlock, 2000) or if immigrants are characterized by outbred vigor (Ball et al., 2000) and/or a mating advantage. Yet, it would be premature to promote the introduction of just one MPG as a general practice. Additional work should build upon experimental and theoretical studies (cited in Introduction) that have attempted to identify strategies for minimizing extinction risk. Acknowledgements We thank D. Blazer, J. Vetter, D. Fletcher, S. Milne, E. Lawyer, and especially S. Reaser for help with the experiment; G. Keeney for supplying the beetles; and D. Fowler and Robin Waples for comments on the ms. JAV was supported, in part, by the U.S. National Science Foundation (DEB– 9317401, DEB–9903671). References Backus, V. L., Bryant, E. H., Hughes, C. R. & Meffert, L. M., 1995. Effect of migration or inbreeding followed by selection on low–founder– number populations: implications for captive breeding programs. Conservation Biology, 9: 1216–1224. Ball, S. J., Adams, M., Possingham, H. P. & Keller, M. A., 2000. The genetic contribution of single male immigrants to small, inbred populations: a laboratory study using Drosophila melanogaster. Heredity, 84: 677–684. Bataillon, T. & Kirkpatrick, M., 2000. Inbreeding depression due to mildly deleterious mutations

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– 2000. Is one migrant per generation sufficient for the genetic management of fluctuating populations? Animal Conservation, 3: 261–266. – 2001. Migration and inbreeding: the importance of recipient population size for genetic management. Conservation Genetics, 2: 167–171. Vucetich, J. A., Waite, T. A. & Nunney, L., 1997. Fluctuating population size and the ratio of effective to census population size. Evolution, 51: 2015–2019. Vucetich, J. A., Waite, T. A., Qvarnemark, L. & Ibargüen, S., 2000. Population variability and extinction risk. Conservation Biology, 14: 1704– 1714. Wang, J. L., 2000. Effects of population structures and selection strategies on the purging of inbreeding depression due to deleterious mutations. Genetical Research, 76: 75–86. Whitlock, M. C., 2000. Fixation of new alleles and the extinction of small populations: drift load, beneficial alleles, and sexual selection. Evolution, 54: 1855–1.

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Long–term change of species richness in a breeding bird community of a small Mediterranean archipelago A. Martínez–Abraín, D. Oro, R. Belenguer, V. Ferrís & R. Velasco

Martínez–Abraín, A., Oro, D., Belenguer, R., Ferrís, V. & Velasco, V., 2005. Long–term changes of species richness in a breeding bird community of a small Mediterranean archipelago. Animal Biodiversity and Conservation, 28.2: 131–136. Abstract Long–term changes of species richness in a breeding bird community of a small Mediterranean archipelago.— We analyzed the pattern of species richness changes in a bird–breeding bird community on a small western Mediterranean archipelago (Columbretes Islands) over a 40–year period (1964–2003). The aim of this study was to qualitatively account for the relative roles of local and regional factors in shaping the community. As expected, we found that regional factors (at the metapopulation spatial scale) increased diversity whereas local factors (i.e. ecological) probably prevented further increases in diversity. We found that the archipelago gained four new species (two seabirds and two falconids) during the study period, whereas no extinctions were recorded. The community seems partially or completely closed to some groups of species (e.g. small–sized birds such as passerines and storm–petrels), probably owing to predatory exclusion by Eleonora falcons (Falco eleonorae). As newly arrived species have breeding calendars that do not fully overlap with those of resident species, competition for space in a rather saturated area is prevented. Preservation of rare species which increase gamma (regional) diversity rather than alpha diversity with common species should be the main local conservation goal. Key words: Colonization, Extinction, Diversity, Columbretes Archipelago, Conservation, Metapopulation. Resumen Cambios a largo plazo en la riqueza de especies en una comunidad de aves nidificantes en un pequeño archipiélago mediterráneo.— Este trabajo analiza los patrones de cambio en la riqueza de especies en una comunidad de aves nidificantes de un pequeño archipiélago mediterráneo (las islas Columbretes, Castellón) durante un periodo de 40 años (1964–2003). El estudio pretende valorar cualitativamente la influencia relativa de los factores locales y regionales. Como se esperaba, se encontró que los factores regionales (a la escala espacial de la metapoblación) aumentaron la diversidad, mientras que los factores ecológicos locales evitaron mayores incrementos. El archipiélago ganó cuatro especies durante el periodo de estudio (dos aves marinas y dos falcónidos), mientras que no se produjo ninguna extinción. La comunidad parece parcial o totalmente cerrada a ciertos grupos de especies, tales como las aves de pequeña talla (p.ej. Paseriformes y paíños) probablemente debido a la depredación excluyente por parte de los halcones de Eleonor (Falco eleanorae). Dado que las especies que son colonizadoras recientes tienen calendarios de cría que no se solapan completamente con los de las especies residentes, se evita la competencia por el espacio de cría en un área bastante saturada. La principal meta conservacionista debe ser la protección de las especies raras, que incrementan la diversidad gamma (regional), más que la diversidad alpha de las especies comunes. Palabras clave: Colonización, Extinción, Diversidad, Columbretes, Conservación, Metapoblación. (Received: 15 VI 04; Conditional acceptance: 17 XII 0; Final acceptance: 12 I 05) A. Martínez–Abraín, D. Oro, Instituto Mediterráneo de Estudios Avanzados, IMEDEA (CSIC–UIB), c/ Miquel Marquès 21, 07190–Esporles, Mallorca, Spain.– R. Belenguer, V. Ferrís & R. Velasco, Reserva Natural de las Islas Columbretes, Conselleria de Territorio y Vivienda, Avda. Hermanos Bou 47, 12003–Castellón, Spain. Corresponding author: A. Martinez–Abrain. E–mail: a.abrain@uib.es

ISSN: 1578–665X

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Introduction Much debate has taken place among ecologists during the last decades about the nature of communities and the regulation of their structure. Typically, a community is now defined as an association of populations of species in a certain area, with no fixed boundaries, and whose structure is shaped by the environment, by the interactions of the populations within the community, and also by historical, regional and global processes (Ricklefs & Latham, 1993). One of the main structural properties of communities is the number and types of species in the community (see Magurran, 1988). This property is time–dependent and long–term records of species richness are thus necessary to properly characterize the diversity of a given system and to assess the role of both local and external factors in the pattern of change. The latter also involve a spatial dimension and hence demographic spatial processes of dispersal, typically emigration and immigration, which are not necessarily correlated with local factors. Examples of this are colonization of empty patches or extinction of occupied patches, in the framework of the metapopulation theory (e.g. Hanski, 1999). We have extensively monitored the population dynamics of most avian species breeding on a small western Mediterranean archipelago but we have not studied community properties as a whole to date. The aim of this paper is to describe and discuss long–term changes in species richness in a small Mediterranean island community, to derive information of conservation interest. This goal is especially relevant from a regional perspective as many of the thousands of islands in the Mediterranean are very small and home to bird faunas which are among the most endangered in the world. (Rodríguez, 1982; Blondel & Aronson, 1999). Material and methods The Columbretes Islands are a small volcanic archipelago located close to the continental slope in the north–western Mediterranean, some 50 km from the continental coast (see fig. 1). The total area of the Columbretes archipelago is 19 ha, divided in four major groups of islands. The largest island (Grossa Island, with 13 ha. or 68% of the emerged land) holds most breeding birds. This island is horseshoe– shaped, has a maximum length of about 1,200 m, a maximum width of 220 m, a minimum width of 17 m, and a maximum height of 67 m, with steep cliffs all around. As the archipelago is very small it is possible to perform reliable surveys on breeding species. Vegetation mainly consists of a number of shrubs adapted to the arid conditions and high salinity, and annual plant species. Vegetation on the main island (Grossa Island) was deeply altered by humans during the construction of the light house in the mid– 19th century when it was burnt down and pigs were introduced to deal with the abundant snakes (vipers?) on the island (Serrano, 1987). The present

breeding bird community includes only eight species (see table 1). Interestingly, there are no breeding passerines on the island although they are very abundant during migration (Giménez, 1987). Local seabird populations have been monitored annually since the early 80’s and annual records of breeding success have been available since 1989. Data from the 60’s and 70’s comes from sporadic visits to the islands by expert ornithologists. The closest seabird colonies are located 80–100 km away, at the Ebro Delta (Tarragona) and the island of Ibiza (Balearic Islands). Information on the number and type of breeding species, as well as competitive interactions and environmental changes comes to a large extent from a literature review, including unpublished reports of the regional government (i.e. Generalitat Valenciana), over a 40–year period. Much of the recent literature on population dynamics however has been generated by our group (Martínez–Abraín et al., 2001, 2002a, 2002b, 2003a, 2003b). Following the automation of the light house in 1975 over a decade elapsed without human habitation until a permanent crew of wardens (three per 15–day shift) was established on the islands. Results As seen in table 1 the "basal" community was made up of two oceanic seabirds (Cory’s Shearwater Calonectris diomedea diomedea and the European Storm–Petrel Hydrobates pelagicus), a raptor commonly found in the Mediterranean islands (Eleonora’s Falcon Falco eleonorae) and the Yellow–legged Gull Larus cachinnans, a gull species known to breed on these islands at least since the 19th century (Salvator, 1895; Bernis & Castroviejo, 1968). The first species to join this community was Audouin’s Gull Larus audouinii, in 1974 (Pechuán, 1974, 1975; Gómez, 1987). A second seabird species (European Shag, Phalacrocorax aristotelis) became established on the islands in 1991, although an isolated breeding attempt occurred in 1985 (see Martínez–Abraín et al., 2001). More recently, two new raptor species have colonized the archipelago. A pair of peregrine falcons, Falco peregrinus, bred successfully in 2002 and again in 2003 after several failed breeding attempts in the past, and a pair of European Kestrels Falco tinnunculus bred successfully during their first attempt in 2003. Hence, during a 40–year period the archipelago has experienced four colonization events, two of them by seabirds and two of them by birds of prey. Seabird species can be considered as established breeders after several decades of continuous reproduction, whereas falconids cannot, owing to their short breeding record to date. We have not included the presence in the colony of several breeding Cory’s Shearwaters from the Atlantic subspecies Calonectris diomedea borealis (see Martínez–Abraín et al., 2002a) as a colonization event because their taxonomic identity (i.e. species or subspecies) is under discussion.

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Discussion Grossa Island

Regional factors increase alpha (local) diversity Increases in diversity at Columbretes are better explained by factors external to the archipelago than by changes in local conditions. For example, colonization by Audouin’s gulls at the beginning of the 70’s took place before the islands were legally protected, when they were suffering a number of human– related disturbances. In addition, fishermen and lighthouse keepers had free access to the islands which were also used as targets for military exercises at the time (Serrano, 1987). Audouin’s gull dynamics is known to be very dependent on dispersal processes within metapopulations and the growth of this colony cannot be explained without attending to immigration from the outside (Oro & Pradel, 2000; Oro & Ruxton, 2001; Oro et al., 2004). The growth of the colony has recently been influenced by external factors and especially by the rescue effect of immigrant individuals from the neighbour colony of the Ebro Delta (Oro et al., 2004). Colonization by shags coincided with a steep decline of a large colony located in Majorca, owing to sand mining close to it that probably affected local feeding grounds (Martínez–Abraín et al., 2001). Alternatively, since shag colonization occurred after island protection at the end of the eighties, it might also be related to a reduction in human disturbance. Both Audouin’s gulls and shags are good examples of the importance of emigration and immigration processes in the colonization rates at the metapopulation level (Oro, 2003). At the same time, the recent decline in Audouin’s gull numbers is also a consequence of dispersal to higher quality sites, and extinction, in the absence of high immigration rates from the outside, is only buffered by the high survival of old adult philopatric breeders (Cam et al., in press). Colonization by falconids may respond to increasing numbers of these two species on the continent and major nearby islands during the last decades (g.o.b, 1997; Marti & Del Moral, 2003), and might be supported by eventually abundant and extended migratory flows of passerines during the spring. The magnitude of these flows is also independent of local features. Ecological factors prevent further increases in alpha diversity Local factors (i.e. competitive or predatory exclusion) are known to reduce diversity, either by removing species or preventing the invasion of new species. In Columbretes, predatory exclusion probably plays a role in preventing the settlement of small– sized birds (e.g. small Passeriformes adapted to arid conditions such as warblers which are present during migration periods) owing to the large density of breeding Eleonora’s Falcons, that are known to prey upon all sorts of small–sized birds (Walter, 1979). In fact, during the study period falcon numbers in-

Ferrera

Foradada

Bergantín 1 km

Fig. 1. Map showing the location of the Columbretes Islands within the western Mediterranean and the major groups of islands within the archipelago. Fig. 1. Mapa de localización de las islas Columbretes en el Mediterráneo occidental y los principales grupos de islas de dicho archipiélago.

creased from ca. 20 pairs in 1964–1985 (Bernis & Castroviejo, 1968; Dolz & Díes, 1987) to the 40–45 present pairs (Generalitat Valenciana, unpub. data). Falcons are present on the islands between April and November, excepting June when most birds fly to the mainland to feed on insects (own unpublished observations) and hence they are present on the islands during the breeding season of passerines and storm–petrels. As early as the 19th century early visitors to the archipelago suggested that Eleonora’s Falcons might have prevented colonization by small birds (see Salvator, 1895). Alternatively it could be argued that small species of long–distance migratory birds exhibit a great deal of biogeographic regionalism and are generally bad colonizers

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Table 1. Historical changes in the number of all breeding species recorded on the Columbretes Islands (NW Mediterranean) between 1964–2003. A positive sign indicates that a species was present as a breeder and a negative sign indicates the opposite. The archipelago became legally protected in 1988. Tabla 1. Cambios históricos en el número de todas las especies nidificantes observadas en las islas Columbretes (Mediterráneo nordoccidental), entre 1964 y 2003. Los signos positivos indican que la especie estaba presente como nidificante, y los negativos, la situación opuesta. El archipiélago empezó a estar bajo protección legal en 1988.

Species

1964–1973

1974–1983

1984–1993

Calonectris diomedea

Hydrobates pelagicus

Larus cachinnans

Falco eleonorae

Larus audouinii

–

Phalacrocorax aristotelis

–

Falco peregrinus

–

Falco tinnunculus

–

(Boehning et al., 1998). It is also possible that the low vegetation cover on some of the smaller islands, together with the small total surface area and the considerable distance to the continental coast, might make Columbretes an unsuitable target for non– migrant passerines. In addition, the small populations of Passeriformes that the island could hold would be very susceptible to local extinction owing to stochastic phenomena (Legendre et al., 1999). Eleonora’s Falcon might also impose some pressure on the small European Storm–Petrel, which is scarce at Columbretes despite the seemingly favourable conditions of the islands. Although storm–petrels were not found in the diet of Columbretes falcons by Dolz & Díes (1987), Columbretes wardens have found remains of storm–petrels predated by falcons in several occasions (G. Urios, pers. comm.). In addition, they have appeared as a component in the Eleonora’s Falcon diet at other colonies such as the Cabrera archipelago (Balearic Islands) where foraging activity at dusk on nocturnal insects, night– dwelling arthropods such as scorpions and even on bats, has been reported (Araujo et al., 1977; Suárez, 2000). Additional support to our hypothesis comes from Bernis & Castroviejo (1968) who reported on a storm–petrel being attacked by Eleonora’s falcon when it was released after ringing at Columbretes. Interestingly, the largest colonies of storm–petrel along the eastern Iberian coast (e.g. Benidorm Island) occur on islands where Eleonora’s Falcons are absent, despite the presence of suitable breeding habitat. Evidence of competitive exclusion also comes from the breeding calendar of newly arrived species. Interestingly, new species colonizing the

1994–2003

Columbretes Islands do not overlap in time with resident species, segregating over the breeding season. Shags coincide with Eleonora’s falcons in their preference for cliffs and crevices located out of the direct influence of solar radiation (Urios, 2003; own observations) and indeed some shag nests are placed on ledges previously used by falcons. The optimal falcon nesting site is virtually occupied at the moment (op. cit.), but there is no conflict with shags regarding nesting sites because the latter start breeding in December–January, whereas falcons do so in July. Syntopic breeding of Audouin’s and Yellow–legged Gulls has been possible to a large extent because the breeding calendar of the former species is delayed about one month in relation to that of the latter. Audouin’s Gulls therefore occupy the space not taken by the Yellow– legged gull latter. Similarly, newly arrived Falco species (table 1) complete their reproduction well in advance of Eleonora’s Falcons starting to build their nests, thus avoiding conflict even though nesting–site preferences seem to coincide. Island diversity and human activities There is no evidence in the presently available literature on bird extinctions during the profound environmental transformations associated with the construction of the light house and later human occupancy. Similarly, no extinction took place during the 40–year period considered by this study. This is at least partially explained by the life history traits of most birds breeding at the study site: high adult survival, large generation times and relatively low rates of population growth (negative or posi-

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tive). However, human disturbance is an important component to understand patterns of species diversity, especially in the Mediterranean basin (Blondel & Aronson, 1999). In this region, population dynamics and distribution patterns of many species have been shaped more by human activities than by evolutionary determinants (op. cit.). Once direct human impacts were removed at Columbretes by conservation laws and especially by effectively protecting the site, only colonizations have occurred. However, human activities are changing rapidly and could drive several species to local extinction. Indeed since 1991, Audouin’s Gulls have locally faced a steep decline owing to reduced food availability during the chick–rearing period when a trawling fishing moratorium was established, depriving gulls from fishing discards, their main food source in the area since colonization of the archipelago (Oro et al., 2004). The declining colony is now suffering high levels of disturbance by Yellow–legged Gulls, probably as a consequence of changed predator/prey ratios (Oro et al., in press). Additionally, Cory’s Shearwaters seem to be following a decreasing trend at Columbretes because of adult mortality in long–line fishing gear (Belda & Sánchez, 2001; Cooper et al., 2003). Island biodiversity and conservation goals Increases in diversity are not good per se. Provided that our hypothesis is true, a decrease in breeding numbers of Eleonora’s Falcons could lead to the colonization of the islands by small passerines and probably to larger numbers of breeding storm–petrels, increasing local diversity. However, the preservation and growth of the colony of this endemic Mediterranean raptor is a more desirable conservation goal than the gain of common species which are abundant elsewhere. Similarly, colonization by a common raptor species, such as the Common Kestrel, could result in increased predation of endemic lizards and beatles reducing the diversity of other animal taxa. Local management efforts should hence focus on promoting the persistence of rare species such as Audouin’s Gulls and Eleonora’s Falcons which increase gamma (regional) diversity. Acknowledgements This study is a contribution to the LIFE02NATURE/ E/8608 for the conservation of Audouin’s Gull in the Comunidad Valenciana, financed by the Generalitat Valenciana and the European Union. We are most grateful to all the wardens of the Columbretes Islands who have monitored bird populations since 1988. We are also grateful to Juan Jiménez, José Vicente Escobar and Josep Carda for promoting the scientific study of the Columbretes birds. R. E. Ricklefs, M. Giménez, J. L. Tella and an anonymous reviewer read early drafts of the manuscript and provided substantial suggestions.

References Araujo, J., Muñoz–Cobo, J. & Purroy, F. J., 1977. Las rapaces y aves marinas del archipiélago de Cabrera. Naturalia Hispanica, 12. Madrid. Belda, E. & Sánchez, A., 2001. Seabird mortality on longline fisheries in the western Mediterranean: factors affecting bycatch and proposed mitigating measures. Biological Conservation, 98: 357–363. Bernis, F. & Castroviejo, J., 1968. Aves de las islas Columbretes en primavera. Ardeola, 12: 143–163. Blondel, J. & Aronson, J., 1999. Biology and wildlife of the Mediterranean region. Oxford Univ. Press, Oxford. Boehning, G. K., Gonzalez–Guzman, L. I. & Brown, J. H., 1998. Constraints on dispersal and the evolution of the avifauna of the Northern Hemisphere. Evolutionary Ecology, 12: 767–783. Cam, E., Oro, D., Pradel, R. & Jimenez, J., in press. Assessment of hypotheses about dispersal in a long–lived seabird using multistate capture–recapture models. Journal of Animal Ecology, 73. Cooper, J., Baccetti, N., Belda, E. J., Borg, J. J., Oro, D., Papaconstantinou, C. & Sánchez, A., 2003. Seabird mortality from longline fishing in the Mediterranean sea and Macaronesian waters: a review and a way forward. Scientia Marina, 67: 57–64. Dolz, J. C. & Dies, N., 1987. El halcón de Eleonor (Falco eleonorae, Gené) en las islas Columbretes. In: Islas Columbretes: contribución al estudio de su medio natural: 241–262 (L. A. Matilla, J. L. Carretero & A. García–Carrascosa, Eds.). Generalitat Valenciana, Valencia. Giménez, M., 1987. Notas sobre migración de aves en las islas Columbretes. In: Islas Columbretes: contribución al estudio de su medio natural: 205– 214 (L. A. Matilla, J. L. Carretero & A. García– Carrascosa, Eds.). Generalitat Valenciana, Valencia. G.O.B., 1997. Atles dels aucells nidificants de Mallorca i Cabrera. G.O.B., Palma de Mallorca. Gómez, J. A., 1987. Láridos nidificantes en las Islas Columbretes: Gaviota Patigualda (Larus cachinnans) y la gaviota de Audouin (Larus audouinii). In: Islas Columbretes: contribución al estudio de su medio natural: 215–222 (L. A. Matilla, J. L. Carretero & A. García–Carrascosa, Eds.). Generalitat Valenciana, Valencia. Hanski, I., 1999. Metapopulation Ecology. Oxford Univ. Press, Oxford. Legendre, S., Clobert, J., Møller A. P & Sorci, G., 1999. Demographic stochasticity and social mating system in the process of extinction of small populations: the case of passerines introduced to New Zealand. The American Naturalist, 153: 449–463. Magurran, A. E., 1988. Ecological diversity and its measurement. Princeton Univ. Press, Princeton, NJ. Marti, R. & Del Moral, J. C., 2003. Atlas de las aves reproductoras de España. Dirección General de Conservación de la Naturaleza–Sociedad

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Española de Ornitología, Madrid. Martínez–Abraín, A., González–Solís, J., Pedrocchi, V., Genovart, M., Abella, J. C., Ruiz, X., Jiménez, J. & Oro., D., 2003a. Kleptoparasitism, disturbance and predation of yellow–legged gulls on Audouin’s gulls in three colonies of the western Mediterranean. Scientia Marina, 67 (Suppl. 2): 89–94. Martínez–Abraín, A., Oro, D., Forero, M. G. & Conesa, D., 2003b. Modelling temporal and spatial colony–site dynamics in a long–lived seabird. Population Ecology, 45: 133–139. Martínez–Abraín, A., Oro, D., Freís, V. & Belenguer, R., 2002b. Is growing tourist activity affecting the distribution or number of breeding pairs in a small colony of the Eleonora’s Falcon? Animal Biodiversity and Conservation, 25.2: 47–51. Martínez–Abraín, A., Oro, D. & Jiménez, J., 2001. The dynamics of a colonization event in the European shag: the roles of immigration and demographic stochasticity. Waterbirds, 24: 97–102. Martínez–Abraín, A., Sánchez, A. & Oro, D., 2002a. Atlantic Cory’s shearwaters breeding in a colony of Mediterranean Cory’s shearwaters. Waterbirds, 25: 221–224. Oro, D., 2003. Managing seabird metapopulations in the Mediterranean: constraints and challenges. Scientia Marina, 67: 13–22. Oro, D., Cam, E., Pradel, R. & Martínez–Abrain, A., 2004. Influence of food availability on demography and local population dynamics in a long– lived seabird. Proceedings of the Royal Society London, Series B, 271: 387–396. Oro, D., Martínez–Abraín, A., Paracuellos, M., Nevado, J. C. & Genovart, M., in press. Influence of density dependence on predator–prey seabird interactions at large spatio–temporal scales. Proceedings of the Royal Society of London, Series B. Oro, D. & Pradel, R., 2000. Determinants of local

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recruitment in a growing colony of Audouin’s gull. Journal of Animal Ecology, 69: 119–132. Oro, D. & Ruxton, G. D., 2001. The formation and growth of seabird colonies: Audouin’s gull as a case study. Journal of Animal Ecology, 70: 527–535. Pechuán, L., 1974. La colonia de Larus audouinii en las islas Columbretes. Ardeola, 20: 358–359. – 1975. Nidificación de Larus audouinii en las islas Columbretes. Ardeola, 21: 407–408. Ricklefs, R. E. & Latham., R. E., 1993. Global patterns of diversity in mangrove floras. In: Species diversity in ecological communities: historical and geographical perspectives: 215–219 (R. E. Ricklefs & D. Schluter, Eds.). Univ. of Chicago Press, Chicago. Rodríguez, J., 1982. Oceanografía del mar Mediterráneo. Editorial Pirámide, Madrid. Salvator, L. von., 1895. Columbretes. Heinr. Mercy und Druck Verlag, Praga. (Traducción española: G. Urios & J. Nachtwey, 1990. Columbretes. Ayto. de Castellón, Castellón). Suárez, M., 2000. Las rapaces nidificantes en el archipiélago de Cabrera. In: Las aves del Parque Nacional marítimo–terrestre del archipiélago de Cabrera (Islas Baleares, España): 233–252 (G. X. Pons, Ed.). G.O.B., Colecciones Técnicas del Ministerio de Medio Ambiente, Madrid. Serrano, R., 1987. Historia de los asentamientos humanos en Columbretes. In: Islas Columbretes: Contribución al estudio de su medio natural: 13–18 (L. A. Matilla, J. L. Carretero & M. García–Carrascosa, Eds.). Generalitat Valenciana, Valencia. Urios, G., 2003. Aplicación de los sistemas de información geográfica y modelos digitales del terreno a la cartografía ambiental en las islas Columbretes. Tesis doctoral, Univ. de València. Walter, H., 1979. Eleonora’s Falcon. Adaptations to prey and habitat in a social raptor. The Univ. of Chicago Press, Chicago.

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Occurrence and abundance of fungus–dwelling beetles (Ciidae) in boreal forests and clearcuts: habitat associations at two spatial scales A. Komonen & J. Kouki

Komonen, A. & Kouki, J., 2005. Occurrence and abundance of fungus–dwelling beetles (Ciidae) in boreal forests and clearcuts: habitat associations at two spatial scales. Animal Biodiversity and Conservation, 28.2: 137–147. Abstract Occurrence and abundance of fungus–dwelling beetles (Ciidae) in boreal forests and clearcuts: habitat associations at two spatial scales.— Insect material (> 30,000 individuals) reared from the fruiting bodies of wood–decaying Trametes fungi was compared between old–growth boreal forests and adjacent clearcuts in Finland. Sulcacis affinis and Cis hispidus occurred more frequently and were, on average, more abundant in the clearcuts. Interestingly, Octotemnus glabriculus and Cis boleti had a slightly higher frequency of occurrence in the forests, despite lower resource availability. The former also showed a higher average abundance. On average, the cluster size of Trametes fruiting bodies occurring on woody debris was higher in the clearcuts than in the forests and had a positive effect on species occurrence and abundance in these clusters. The independent effect of the macrohabitat (forest or clearcut) underscores the importance of the macrohabitat where specific resources occur, and this may override the positive effects of resource availability. Key words: Forest landscape, Boreal forest, Coarse woody debris, Wood–decaying fungi, Trametes, Ciidae. Resumen Presencia y abundancia de los escarabajos fungícolas (Ciidae) en los bosques y claros de tala boreales: asociaciones al hábitat según dos escalas espaciales.— Se compararon las cantidades de insectos (> 30,000 individuos) que se alimentan de los cuerpos fructíferos de los hongos desintegradores de la madera Trametes en los bosques boreales maduros y los claros adyacentes en Finlandia. Sulcacis affinis y Cis hispidus aparecían con mayor frecuencia, y en promedio eran más abundantes en los claros. Llama la atención la frecuencia ligeramente mayor de Octotemnus glabriculus y Cis boleti en los bosques, a pesar de una menor disponibilidad de recursos. El primero también presentaba una abundancia promedio mayor. En promedio, el tamaño de las masas de cuerpos fructíferos de Trametes de los restos de árboles era mayor en los claros que en los bosques, y tenía un efecto positivo en la presencia y abundancia de especies en dichas masas. El efecto independiente del macrohábitat (bosque o claro) subraya la importancia del macrohábitat cuando los recursos específicos aparecen, pudiendo anular los efectos positivos de la disponibilidad de recursos. Palabras clave: Paisaje forestal, Bosque boreal, Restos gruesos de madera, Hongos desintegradores de la madera, Trametes, Ciidae. (Received: 20 IX 04; Conditional acceptance: 17 XI 04; Final acceptance: 12 I 05) Atte Komonen & Jari Kouki, Fac. of Forest Sciences, P. O. Box 111, FI–80101, Univ. of Joensuu, Finland. Corresponding author: A. Komonen. E–mail: atte.komonen@joensuu.fi

ISSN: 1578–665X

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Introduction Ecological research on habitat fragmentation has often metaphorically viewed suitable habitats as "islands" in a hostile "sea" (Haila, 2002). Because there is a low degree of deforestation in the boreal forest, landscapes rarely appear simply as a black–and–white contrast between habitat and non–habitat, rather the many shades of g r e y r e f l e c t d i ff e r e n t i a l h a b i ta t s u i ta b i l i t y (Mönkkönen & Reunanen, 1999). Forestry practices in boreal forests create highly dynamic landscapes which remain forested while undergoing spatial and temporal changes in structure and dynamics (Kouki et al., 2001; Schmiegelow & Mönkkönen, 2002). The structural heterogeneity in the forest landscape is manifested in the amount and quality of living and dead wood (Siitonen, 2001), forest–stand age structure and naturalness (Uotila et al., 2002), and in the distribution of different stand types in the landscape (Löfman & Kouki, 2003). Dead wood is a key resource for thousands of organisms in boreal forests (Esseen et al., 1997; Siitonen, 2001). Dead wood is scarce in managed forests and many dead–wood dependent organisms are consequently absent or occur in low numbers. This has contributed to the misconception that the majority of these species would require natural forests. However, many dead–wood dependent species can persist in clearcuts if critical resources are left in adequate densities and qualities in management

operations (Kaila et al., 1997; Jonsell et al., 2001; Jonsson et al., 2001; Martikainen, 2001; Sverdrup–Thygeson & Ims, 2002). However, the importance of the macrohabitat and landscape context where these resources occur is poorly understood, particularly for species that have high habitat specificity and limited powers of dispersal, such as fungus–dwelling insects (Jonsell et al., 1999; Jonsson et al., 2001; Komonen et al., 2000). It is important to disentangle the large– and small–scale effects of forestry on habitat suitability to fully understand the limiting factors for species occurrences and abundances (Mönkkönen & Reunanen, 1999). Trametes fruiting bodies occur in a variety of forest environments where deciduous dead wood is available. Nevertheless, it is not known how Trametes–dwelling beetles respond to the different forest surroundings. There are generally marked ecological differences among the fungus–dwelling insects in dispersal ability (Jonsell et al., 1999; Jonsson, 2003), habitat requirements (Nilsson, 1997; Guevara et al., 2000a; Thunes et al., 2000; Jonsell et al., 2001) and host–fungus specificity (Lawrence, 1973; Økland, 1995; Fossli & Andersen, 1998; Guevara et al., 2000b; Komonen, 2001). In this paper, the occurrence and abundance of four fungivorous beetle species (Ciidae) co–occurring in Trametes fruiting bodies is investigated at two spatial scales. The small–scale effects of fungal–cluster size and the large–scale effects of macrohabitat (forest–clearcut) are tested.

Table 1. Study site characteristics. Area of forest refers to the area (ha) with > 60 m3 fallen woody debris (diameter at breast height ≥ 7 cm) ha–1: As. Area sampled; Af. Area of the forest; Ac. Area of the clearcut, refers to the area (ha) logged at the same time and consists of interconnected openings rather than a single large opening; CWD. Number of stumps, snags, logs and branches (diameter ≥ 10 cm) of birch and aspen ha–1; Yl. Year of logging; MAl. Mean age at the time of logging, refers to the mean age of trees belonging to the dominant canopy storey; Vl. Volume at logging, refers to the volume of living spruce, pine and birch at the time of logging. Tabla 1. Características del área de estudio. Área forestal se refiere al área (ha) con > 60 m3 de residuos de madera caída (diámetro a la altura del tórax ≥ 7 cm) ha–1: As. Área muestreada; Af. Área de bosque; Ac. Área de claros, se refiere al área (ha) talada al mismo tiempo, y consiste en varios claros interconectados, más que en uno solo de mayor tamaño; CWD. Número de tocones, cepas, troncos y ramas (diámetro ≥ 10 cm) de abedules y álamos temblones ha–1; YI. Edad de talado; MAI. Edad media de talado, se refiere a la edad media de los árboles del piso dominante del dosel forestal; Vl. Volumen cuando la tala, se refiere al volumen de pinos, álamos y píceas en el momento de la tala.

Forest Study site

Clearcut

CWD

MAl

CWD

1. Ruunavaara

4.0

1994

155

239

2. Pieni Hovinvaara

2.0

1990/1993

145/128

226/268

3. Haapahasianvaara

3.5

1993

127

197

111

4. Vankonvaara

4.5

1994

132

223

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Animal Biodiversity and Conservation 28.2 (2005)

1.4

Mean log. fungal weight (SE)

Fungal density ha–1 (SE)

1.2

1.0

0.8

0.6

0 1

2 3 Study sites

Fig. 1. A. The mean ± SE fungal density in the study sites (density is measured as the number of occupied pieces of woody debris ha–1); B. The mean ± SE fungal–cluster weight in the study sites, note that y–axis starts from 0.6: Black circles represent forests and open circles indicate clearcuts. Fig. 1. A. Densidad fúngica media ± EE en las áreas de estudio (la densidad se mide como el número de trozos de restos de madera ocupados ha–1); B. Peso medio ± EE de las masas fúngicas en las áreas de estudio; obsérvese que el eje y parte de 0,6: Círculos negros representan los bosques; círculos vacíos, los claros.

Materials and methods Study system Four species of Trametes (formerly included in Coriolus Quél.) occur in the study region in eastern Finland. Of these, Trametes ochracea (Pers.) Gilb. & Ryvarden is the most common species inhabiting dead deciduous trees, mainly aspen (Populus tremula L.) and birch (Betula spp.) (Niemelä, 2001). Over 95% of our samples were T. ochracea (K. Junninen det.), but as the fruiting bodies of all Trametes species are very similar in physical structure (Ryvarden & Gilbertson, 1993) it was impossible to identify some of the heavily– consumed samples with certainty. As far as it is documented, there are no great differences in the Ciidae fauna associated with the different species of Trametes found in Fennoscandian boreal forests (Fossli & Andersen, 1998; Selonen, 2004). Trametes fruiting bodies are annual and typically occur in a relatively early phase of decay succession (3–7 yrs; Hintikka, 1993). They are common in woody debris and stumps, and typically form clusters of fruiting bodies. Although the fruiting bodies of Trametes are annual, dead fruiting bodies can remain attached to wood for one to two years and become entirely consumed by insects, mainly larvae and adults of Ciidae.

Information on Ciidae life history is limited. The life history is completed within the same piece of fungus and several generations may occur before the adults emigrate to find a fresh piece of fungus (Entwistle, 1955). The adults are long–lived for an insect (up to seven months; Klopfenstein, 1971) and during the summer all life–stages may be found at the same time. Three of the species included in this study [Sulcacis affinis (Gyllenhal), Octotemnus glabriculus (Gyllenhal) and Cis boleti (Scopoli)] lay eggs singly. The adults copulate at intervals during the oviposition period of the female and are not monogamous (Entwistle, 1955). The latter two species, in addition to Cis hispidus (Paykull) encountered in this study, are specialists in Trametes (Fossli & Andersen, 1998; Guevara et al., 2000b; Selonen, 2004). Study sites The study area is located in North Karelia, eastern Finland, in the middle boreal zone (63o 00’ – 30’ N, 30 o – 31 o E). First, 12 spruce–dominated old– growth stands were visited. They were rich in aspen (Kouki et al., 2004) and thus assumed to have a sufficiently high density of Trametes to provide adequate sample size. Only four stands were adjacent to clearcuts and these were selected for the study: 1. Ruunavaara; 2. Pieni

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Table 2. Number of individuals and percent of samples occupied by ciid beetles in this study in forests and clearcuts. Tabla 2. Número de individuos y porcentaje de muestras ocupadas por escarabajos cíidos en este estudio, en los bosques y en los claros de tala.

Forest Species

Indiv.

18,125

4,096

3,529

Cis hispidus

508

3,719

Cis boleti

239

570

Sulcacis fronticornis

254

Cis comptus

Cis glabratus

Sulcacis affinis Octotemnus glabriculus

Indiv.

Clearcut

Ennearthron laricinum

0.4

Orthocis alni

Cis lineatocribratus

Cis jacquemartii

Ennearthron cornutum

0.4

Hovinvaara; 3. Haapahasianvaara; 4. Vankonvaara (table 1). As not all stands had large enough (≥ 2 ha) adjacent clearcuts, two of the sampling quadrats (see below) in site 4, and one in site 2 were located further from the studied forest, yet within 1–km distance and adjacent to other patches of old– growth. All the clearcut areas had once been part of the larger old–growth forest but had been logged 7 or 8 years earlier; part of the clearcut in site two had been logged 11 years earlier (data from Forest and Park Service, Lieksa). This oldest clearcut area was the only one with 2–3 m tall birch trees, the others were fully open. The small difference between the mean fungal density in site two results from this difference (fig. 1). Sampling Fungal samples were collected between 22 IX and 6 X 01, the optimal time given the species phenology. One hectare study quadrats were randomly positioned and marked in the field; in the forests, quadrats were at least 50 m from the forest–clearcut edge. An equal area was sampled in a given old– growth forest and the adjacent clearcut (table 1). The sampled areas varied in size among forest– clearcut pairs, because larger forest stands allowed larger sampling coverage, thus increasing sample size and statistical power. However, the area sampled does not consistently follow the area of forests due to the discrepancy between the forest area and the area and shape of the adjacent clearcuts. Due to the shape and size of the clearcut in site 4, it was

impossible to establish all the sample units as 1–ha quadrats; instead we used 200 m x 50 m strips. In sites 3 and 4, two 50 m x 50 m quadrats were used for sampling the area of 0.5 ha. In all study quadrats we examined all woody debris (dbh ≥ 10 cm) of deciduous trees for Trametes fruiting bodies. As Trametes fruiting bodies often occur in tight clusters, insect larvae could potentially move from one fruiting body to another. Thus, all the fruiting bodies in a cluster were considered one sample and carefully removed. If the fungal clusters were located in separate parts of the same individual tree, for example in a stump and a trunk not attached to each other, these were considered separate samples. Samples were transferred to mesh– net covered plastic boxes and kept in outdoor conditions. On 8 II 02, all the samples were transferred to room temperature and weighed after two weeks, a period which was considered adequate for excess water to evaporate and make the weight of samples collected under different daily weather conditions more comparable. The fungal cluster size was determined by weighing, reflecting both the number and size (g) of fruiting bodies. All the fruiting bodies were carefully dissected and the adult insect individuals were removed and identified. Statistical analyses Generalized Linear Models (GLM; McCullagh & Nelder, 1989) were used to analyze the data. In all models, site was introduced as a random effect and management category (forest or clearcut) as

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Table 3. GLM results on the fungal–cluster occupancy for four ciid beetles. The change in deviance indicates the model improvement when a given term is included in the model. The full model was compared to a model with only a constant term: F. Full model; Mng. Management nested within site; S. Site; W. Weight. Rd. Residual deviance

Table 4. Parameter estimates ± SE give the change in the log of the odds for a ciid species occurring or not occurring in a fungal cluster sampled from forests, holding fungal–cluster weight constant. Estimates are taken from the GLM in table 3: Mng. Management; W. Weight. (* Unstable estimate as this ciid species was absent from the forest in this site.)

Tabla 3. Resultados GLM de la ocupación de las masas fúngicas por parte de cuatro coleópteros cíidos. El cambio de la desvianza indica la mejora del modelo cuando se incluye en éste un término dado. El modelo completo se comparó con un modelo con un único término constante: F. Modelo completo; Mng. Manejo anidado dentro del área; S. Área; W. Peso. Rd. Desvianza residual.

Tabla 4. Las estimas de los parámetros ± EE nos dan el cambio en el logaritmo de los valores predichos/observados ("odds") para cada especie de cíido, que aparece o no, en las masas fúngicas muestreadas en los bosques, siendo constante el peso de dichas masas fúngicas. Las estimas se tomaron de los valores GLM de la tabla 3: Mng. Control; W. Peso. (* Estima inestable cuando esta especie de cíido estaba ausente del bosque en ese lugar.)

Change in deviance Ciid species

d.f.

χ2

P Ciid species

Sulcacis affinis

Estimate ± SE

Sulcacis affinis

277.86

0.000

Mng

185.76

0.000

Mng (site 1)

–1.63 ± 0.35 –3.08 ± 0.79

2.87

0.412

Mng (site 2)

47.99

0.000

Mng (site 3)

–7.28 ± 26.6*

276

156.74

1.000

Mng (site 4)

–2.10 ± 0.35

Octotemnus glabriculus

2.88 ± 0.48

Octotemnus glabriculus

80.18

0.000

Mng

27.11

0.000

Mng (site 1)

1.25 ± 0.30

12.83

0.005

Mng (site 2)

0.31 ± 0.45 0.20 ± 0.30

55.43

0.000

Mng (site 3)

276

262.94

0.704

Mng (site 4)

0.82 ± 0.32

Cis hispidus

2.38 ± 0.36

111.05

0.000

Cis hispidus

Mng

24.12

0.000

Mng (site 1)

–0.68 ± 0.29 –1.34 ± 0.62

1.66

0.647

Mng (site 2)

61.30

0.000

Mng (site 3)

–0.35 ± 0.31

276

225.17

0.988

Mng (site 4)

–0.85 ± 0.26

Cis boleti

2.63 ± 0.38

Cis boleti

99.49

0.000

Mng

12.57

0.014

Mng (site 1)

0.79 ± 0.28

8.11

0.044

Mng (site 2)

0.30 ± 0.41

91.53

0.000

Mng (site 3)

0.54 ± 0.29

0.052

Mng (site 4)

0.17 ± 0.24

2.74 ± 0.34

276

315.30

a fixed effect nested within site. Fungal–cluster weight was included as a continuous covariate. For the distribution of a species (i.e. the presence or absence of a species in a cluster; a binary

response) a binomial error distribution and a logit link–function were assumed. For the abundance of a species (the average number of individuals per cluster) an identity link–function and normally dis-

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142

% of clusters occupied

100

Sulcacis affinis

100

20 0

0 1 100

Octotemnus glabriculus

Cis hispidus

1 100

Cis boleti

0 1

2 3 Study sites

Fig. 2. Percent of fungal clusters occupied by ciid species in the four study sites; black bars represent forests and grey bars show clearcuts. Fig. 2. Porcentaje de masas fúngicas ocupadas por especies de cíidos en las cuatro áreas estdiadas; las barras negras representan los bosques, y las grises los claros.

tributed errors were assumed. The significance of each term was evaluated based on the increase in deviance when the term was dropped from the full model containing all explanatory parameters. Fungal–cluster weight and the number of individuals were log10(x+1)–transformed in all analyses, unless otherwise stated. Results Host availability Altogether, 106 and 245 pieces of woody debris occupied by dead fruiting bodies of Trametes were recorded and sampled from forests and clearcuts, respectively. The density of woody debris occupied by Trametes was consistently lower in the forest sites (mean ha–1 ± SE = 7.34 ± 0.63) than in the clearcut sites (16.00 ± 2.97; fig. 1A).

The average fungal–cluster weight was also consistently lower in the forest than in the clearcut sites (fig. 1B; change in deviance = 1.760, F4, 343 = 8.560, P < 0.000). Patterns of beetle occupancy A total of 32,193 insect individuals were removed from the sampled fruiting bodies. These included 12 species of Ciidae, the four most common species of which (96% of all insect individuals) were used in the analyses (table 2). Fungal–cluster weight had a positive effect of similar magnitude on the probability of a cluster being occupied for all four species (tables 3, 4). Sulcacis affinis and Cis hispidus had a very consistent pattern of occurrence in all sites, in comparison with Octotemnus glabriculus and Cis boleti, in that they were more frequent in clearcuts (fig. 2, table 4). Interestingly, O. glabriculus and C. boleti were

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Predicted probability of occurrence

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1.0

0.8

0.6

0.4

0.4 Sulcacis affinis 1.107 (1.053–1.165) 1.064 (1.026–1.103)

0.2 0.0

100

120

0.2 0.0

1.0

0.8

0.6

0.4

100

120

0.4 Cis hispidus 1.111 (1.055–1.171) 1.121 (1.053–1.193)

0.2 0.0

Octotemnus glabriculus 1.052 (1.028–1.077) 1.133 (1.022–1.255)

40 60 80 100 Fungal weight (g)

120

Cis boleti 1.048 (1.029–1.067) 1.139 (1.065–1.218)

0.2 0.0

40 60 80 100 Fungal weight (g)

120

Fig. 3. The predicted probability of occurrence of the four ciid species in the fungal clusters as a function of cluster weight (untransformed), based on a logistic regression model. Figures give the estimated odds ratios and 95% CIs; black dots represent forests, grey dots clearcuts. Fig. 3. Probabilidad predicha de la presencia de las cuatro especies de cíidos en las masas fúngicas, como función del peso de dichas masas (no transformado), basada en un modelo logístico de regresión. Las cifras se refieren a la estima de la razón entre los valores predichos y observados ("odds ratio") y los CIs 95%; los círculos negros representan los bosques, y los grises los claros.

more likely to be found from forests, as indicated by signs of parameter estimates (table 4), despite lower fungal availability (fig. 1). For the two "forest species", the predicted probability of occurrence increased more slowly as a function of fungal– cluster weight in the clearcut clusters than in the forest clusters (fig. 3). The opposite pattern was observed for the two "clearcut species". Patterns of beetle abundance Sulcacis affinis and C. hispidus were more abundant on average in the clearcut than in the forest clusters, and O. glabriculus was more abundant in the forest clusters, after controlling for fungal weight (table 5; fig. 4). Cis boleti did not show significant difference in abundance between a forest and a clearcut. It was further tested whether there were interspecific differ-

ences in average abundances in forest and clearcut clusters. In the forests, O. glabriculus had a larger mean population size in the clusters than S. affinis, C. hispidus and C. boleti (F3, 200 = 35.273, r2 = 0.346, P < 0.05; Dunnett’s C for pairwise comparisons). In the clearcuts, S. affinis had a significantly larger population size than the other three species and C. boleti population size was significantly smaller (F3, 722 = 112.343, r2 = 0.318, P < 0.05; Dunnett’s C for pairwise comparisons). Discussion Findings from this study demonstrate that the macrohabitat where specific resources occur is important for ciid beetles in Trametes fruiting bodies (c.f. Thunes & Willassen, 1997; Jonsell et al., 2001;

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Table 5. GLM results on the mean abundance of ciid beetles in fungal clusters. The change in deviance indicates the model improvement when a given term is included in the model. The interaction term between management category and weight was non–significant for each species (F < 2.63, P > 0.1), and thus it was excluded from the final models: η2. Proportion of explained variance; Mng. Management nested within site; W. Weight; Rd. Residual deviance; * d.f. is only 3 because one site did not include any occupied fungal clusters). Tabla 5. Resultados GLM de la abundancia media de coleópteros cíidos en las masas fúngicas. El cambio de la desvianza indica la mejora del modelo cuando se incluye en éste un término dado. El término de interacción entre la categoría de gestión y el peso para cada especie (F < 2,63, P > 0,1) no fue significativo, por lo que se excluyó de los modelos finales: η2. Proporción de varianza explicada; Mng. Manejo anidado dentro del área; W. Peso; Rd. Desvianza residual; * d.f. es sólo de 3 debido a que uno de los lugares no incluía ninguna masa fúngica ocupada).

Species Source

d.f.

Change in deviance

η2

Sulcacis affinis Mng

3.775

17.113

0.000

0.193

Site

3.064

0.376

0.779

0.287

27.736

125.723

0.000

0.370

214

0.221

3.405

16.415

0.000

0.207

Octotemnus glabriculus Mng Site

0.263

0.080

0.968

0.056

14.665

70.686

0.000

0.219

252

0.207

Mng

0.552

4.525

0.001

0.065

Site

0.383

0.705

0.597

0.344

16.675

136.666

0.000

0.344

261

0.122

Mng

0.128

1.568

0.185

0.036

Site

0.207

1.634

0.311

0.539

2.443

29.815

0.000

0.151

168

0.082

Residual deviance Cis hispidus

Residual deviance Cis boleti

W Residual deviance

Jonsson et al., 2001). Most interestingly, some ciids seem to prefer forests over clearcuts despite lower fungal density and smaller size of fungal clusters in the former. In the absence of experiments, however, we cannot make explicit causal inferences. We therefore discuss two potential ecological factors that could contribute to the difference observed in beetle species occurrence and abundance between forests and clearcuts. Small–scale effects of fungal cluster For all ciid species of the present study, fungal– cluster weight contributed positively to the prob-

ability of a sample being occupied, both in the forest and clearcut (see also Midtgaard et al., 1998). The higher frequency of occurrence of ciids can be explained by the greater probability of detecting larger fungal clusters (actively or by chance only), as well as by a longer expected persistence time of the local beetle population in larger clusters. There is evidence that both walking and flying ciids are attracted to the volatile compounds of their host fungus rather than finding the fungi accidentally (Jonsell & Nordlander, 1995; Guevara et al., 2000a, 2000b). If larger fungal clusters emit greater amounts of volatile compounds, then they would also attract more

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Mean log. population size (SE)

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2.0

Sulcacis affinis 2.0

1.5 1.5 1.0

1.0

0.5

0.5 0.0

0.0 1

Mean log. population size (SE)

Octotemnus glabriculus

1.4

Cis hispidus

1 0.80

2 3 Study sites

Cis boleti

0.75 1.2

0.70 0.65

1.0

0.60 0.8

0.55 0.50

0.6

0.45 0.4

0.40 1

2 3 Study sites

Fig. 4. The mean ± SE population size per fungal cluster occupied by a given ciid species in the study sites; black circles represent forests and open circles clearcuts. Fig. 4. Tamaño medio ± EE de la población por masa fúngica ocupada por una especie dada de cíido en las áreas de estudio; los círculos negros representan los bosques, y los vacíos los claros.

ciids, partly explaining the positive effect of fungal weight on the probability of occurrence. In this study it is shown that "forest species" have a higher probability of occurrence in a fungal cluster of a given size in forests (optimal macrohabitat) vs clearcuts (suboptimal macrohabitat), and "clearcut species" show an opposite pattern. These results suggest that species may compensate for adverse environmental conditions in the suboptimal macrohabitat by utilizing larger fungal clusters. As the measure of fungal cluster weight used here incorporates both the number and the weight of fruiting bodies, it is difficult to say much about the possible mechanisms that cause the observed difference in species incidence. The observed pattern may result from a higher colonization rate or from a longer ciid population persistence time in larger fungal clusters. Besides, larger clusters probably have greater variation in the quality of fruiting bodies, thus increasing the likelihood that a given beetle species meets its

specific ecological requirements. In larger clusters there is also more potential for resource partitioning, which could facilitate coexistence by reducing interspecific competition between ciid species (Guevara et al., 2000a). Of the species recorded in this study, only S. affinis is known to inhabit —in great numbers— fruiting bodies of fungal species other than Trametes, namely Pycnoporus cinnabarinus (Jacq.: Fr.) P. Karsten (Økland, 1995). The fruiting bodies of this fungus occur exclusively in warm, open areas (Niemelä, 2001) and were also abundant in the clearcuts of this study. This supplementary host fungus could contribute to the very high frequency of occurrence of S. affinis in the clearcuts. Larger fungal clusters inherently support larger ciid populations (Midtgaard et al., 1998), which in turn could affect local occurrence patterns via increased number of dispersing individuals. Evidence on the dispersal ability of ciid beetles is scarce and indirect, making it impossible to assess the dispersal rate

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between fungal clusters. Jonsell et al. (1999) demonstrated that some ciid species were absent from fruiting bodies placed out in forest fragments 350– 2,000 m from the natural forest. Similarly, Rukke (2000) found out that the incidence of ciid species in fruiting bodies was negatively affected by increased isolation at the scale of individual trees, 15 to over 500 m depending on the species. Given the high density of Trametes fruiting bodies it seems unlikely there would have been great difficulties for ciids in moving between clusters in clearcuts. However, restricted movement between clusters in forests and between adjacent forests and clearcuts is possible. Large scale effects of macrohabitat The forest management category had an independent effect on species frequency of occurrence, even less so than fungal–cluster weight. Similarly, the management category contributed positively to the abundance of S. affinis and C. hispidus in the clearcuts and O. glabriculus in the forests. In this study, it is difficult to explicitly distinguish between the effect of fungal cluster density and management category on the frequency of occurrence of the ciid beetles in the clusters, as the density was consistently higher in the clearcuts. The apparent inconsistent response of the two "forest species" to the management category may be due to opposing forces of microclimate and resource availability (Jonsell et al., 2001; Jonsson et al., 2001). However, a larger sample of forest and clearcut sites is needed to clarify whether such trade–off exists, or if the inconsistency results from inadequate sampling. Microclimatic conditions are one of the most important abiotic differences between forests and clearcuts. Therefore, varying interspecific responses of ciid species to forest management may result from different microclimatic optima. Other studies have documented that increased sun–exposure and dryness of fruiting bodies increase the probability of occurrence of some fungus–dwelling insect species (Midtgaard et al., 1998; Rukke & Midtgaard, 1998), whilst other species occur more frequently in moist fruiting bodies or in shady conditions (Økland, 1996; Jonsell et al., 2001; Thunes et al., 2000). Sverdrup– Thygeson & Ims (2002) showed that among the beetle species in dead aspen there are clear preferences concerning the degree of sun–exposure. Their window–trap material also included O. glabriculus and, as in the present study, the species preferred shady conditions being more abundant in traps on shady aspen logs. Fossli & Andersen (1998) collected Trametes fruiting bodies from forests and, again, S. affinis was very rare. In Germany, S. affinis occurs readily in clearcuts, whereas O. glabriculus and C. boleti manage well in more shady conditions (Reibnitz, 1999). Microclimate can affect the abundance of ciids directly by speeding up the individual development and indirectly by affecting the quality of fruiting bodies. However, there are no studies linking population growth to the quality of fruiting bodies.

Conclusions Many dead–wood dependent organisms can successfully occur in clearcuts if critical resources are left in adequate densities and qualities (Jonsell et al., 2001; Jonsson et al., 2001; Martikainen, 2001; Sverdrup–Thygeson & Ims, 2002). Prior to the extensive fire suppression in Finnish forests in the 1900s, many of these species may have favored open areas created by forest fires. Nevertheless, sweeping generalizations about species responses should be avoided even for common species, as demonstrated here. Despite higher Trametes density, clearcuts are more ephemeral environments for Trametes–dwelling insects in comparison with old–growth forests. The rationale is that most woody debris becomes unsuitable for Trametes over the course of years (3–7 years; Hintikka, 1993), after which it takes decades before new woody debris is available. Retaining green and dead deciduous trees in clear–cutting makes woody debris available for Trametes and many other more demanding species after the logging residues have become unsuitable. Acknowledgements We thank the Forest and Park Service (Lieksa) for providing data on the study areas. Tomas Roslin and Mats Jonsell kindly commented on the earlier version of the manuscript. This study was funded by the Academy of Finland, Centre of Excellence Programme (# 64308). References Entwistle, H. M., 1955. The biology and morphology of the fungus beetles of the family Ciidae and their parasites. MSc thesis, Univ. of London. Esseen, P.–A., Ehnström, B., Ericson, L. & Sjöberg, K., 1997. Boreal forests. Ecol. Bull., 46: 16–47. Fossli, T.–E. & Andersen, J., 1998. Host preference of Ciidae (Coleoptera) on tree–inhabiting fungi in northern Norway. Entomol. Fennica, 9: 65–78. Guevara, R., Hutcheson, K. A., Mee, A. C., Rayner, A. D. M. & Reynolds, S. E., 2000a. Resource partitioning of the host fungus Coriolus versicolor by two ciid beetles: the role of odour compounds and host ageing. Oikos, 91: 184–194. Guevara, R., Rayner, A. D. M. & Reynolds, S. E., 2000b. Orientation of specialist and generalist fungivorous ciid beetles to host and non–host odours. Physiol. Entomol., 25: 288–295. Haila, Y., 2002. A conceptual genealogy of fragmentation research: from island biogeography to landscape ecology. Ecol. Appl., 12: 321–334. Hintikka, V., 1993. Occurrence of edible fungi and other macromycetes on tree stumps over a sixteen–year period. Acta Bot. Fenn., 149: 11–17. Jonsell, M. & Nordlander, G., 1995. Field attraction of Coleoptera to odours of the wood–decaying polypores Fomitopsis pinicola and Fomes

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fomentarius. Ann. Zool. Fenn., 32: 391–402. Jonsell, M., Nordlander, G. & Ehnström, B., 2001. Substrate associations of insects breeding in fruiting bodies of wood–decaying fungi. Ecol. Bull., 49: 173–194. Jonsell, M., Nordlander, G. & Jonsson, M., 1999. Colonization patterns of insects breeding in wood– decaying fungi. J. Insect Conserv., 3: 145–161. Jonsson, M., 2003. Colonization ability of the threatened tenebrionid beetle Oplocephala haemorrhoidalis and its common relative Bolitophagus reticulatus. Ecol. Entomol., 28: 159–167. Jonsson, M., Jonsell, M. & Nordlander, G., 2001. Priorities in conservation biology: a comparison between two polypore–inhabiting beetles. Ecol. Bull., 49: 195–204. Kaila, L., Martikainen, P. & Punttila, P., 1997. Dead trees left in clear–cuts benefit saproxylic Coleoptera adapted to natural disturbances in boreal forest. Biodivers. Conserv., 6: 1–18. Klopfenstein, P. C., 1971. The ecology, behavior, and life cycle of the mycetophilous beetle, Hadreule blaisdelli (Casey) (Insecta: Coleoptera: Ciidae). Thesis, Bowling Green State Univ., Ohio. Komonen, A., 2001. Structure of insect communities inhabiting two old–growth forest specialist bracket fungi. Ecol. Entomol., 26: 63–75. Komonen, A., Penttilä, R., Lindgren, M. & Hanski, I., 2000. Forest fragmentation truncates a food chain based on an old–growth forest bracket fungus. Oikos, 90: 119–126. Kouki, J., Löfman, S., Martikainen, P., Rouvinen, S. & Uotila, A., 2001. Forest fragmentation in Fennoscandia: linking habitat requirements of wood–associated threatened species to landscape and habitat changes. Scand. J. Forest Res. Suppl., 3: 27–37. Kouki, J., Arnold, K. & Martikainen, P., 2004. Long– term persistence of aspen –a key host for many threatened species– is endangered in old–growth conservation areas in Finland. Journal for Nature Conservation, 12: 41–52. Lawrence, J. F., 1973. Host preference in Ciid beetles (Coleoptera: Ciidae) inhabiting the fruiting bodies of basidiomycetes in North America. Bull. Mus. Comp. Zool., 145: 163–212. Löfman, S. & Kouki, J., 2003. Scale and dynamics of transforming forest landscape. For. Ecol. Manage., 175: 247–252. Martikainen, P., 2001. Conservation of threatened saproxylic beetles: significance of retained aspen Populus tremula on clearcut areas. Ecol. Bull., 49: 205–218. McCullagh, P. & Nelder, J. A., 1989. Generalized linear models. 2nd ed. Chapman & Hall, New York. Midtgaard, F., Rukke, B. A. & Sverdrup–Thygeson, A., 1998. Habitat use of the fungivorous beetle Bolitophagus reticulatus (Coleoptera: Tenebrionidae): effects of basidiocarp size, humidity and competitors. Eur. J. Entomol., 95: 559–570.

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Mönkkönen, M. & Reunanen, P., 1999. On critical thresholds in landscape connectivity: a management perspective. Oikos, 84: 302–305. Niemelä, T., 2001. Guide to the polypores of Finland. 13th ed. Botanical Bulletins of the Univ. of Helsinki 179. Nilsson, T., 1997. Survival and habitat preferences of adult Bolitophagus reticulatus. Ecol. Entomol., 22: 82–89. Økland, B., 1995. Insect fauna compared between six polypore species in a southern Norwegian spruce forest. Fauna Norv. Ser. B, 42: 21–26. – 1996. Unlogged forests: important sites for preserving the diversity of mycetophilids (Diptera: Sciaroidea). Biol. Conserv., 76: 297–310. Reibnitz, J., 1999. Verbreitung und Lebensräume der Baumschwammfresser Südwestdeutschlands (Coleoptera: Cisidae). Mitt. Ent. Stuttgart, 34: 1–76. Rukke, B. A., 2000. Effects of habitat fragmentation: increased isolation and reduced habitat size reduces the incidence of dead wood fungi beetles in a fragmented forest landscape. Ecography, 23: 492–502. Rukke, B. A. & Midtgaard, F., 1998. The importance of scale and spatial variables for the fungivorous beetle Bolitophagus reticulatus (Coleoptera, Tenebrionidae) in a fragmented forest landscape. Ecography, 21: 561–572. Ryvarden, L. & Gilbertson, R. L., 1993. European Polypores 1. Abortiporus–Lindtneria. Synopsis Fungorum 6. Fungiflora A/S, Oslo. Schmiegelow, F. K. A. & Mönkkönen, M., 2002. Habitat loss and fragmentation in dynamic landscapes: avian perspectives from the boreal forest. Ecol. Appl., 12: 375–389. Selonen, V. A. O., 2004. Polypores and associated beetles in a forest–clearcut ecotone. MSc thesis, Univ. of Jyväskylä. Siitonen, J., 2001. Forest management, coarse woody debris and saproxylic organisms: Fennoscandian boreal forests as an example. Ecol. Bull., 49: 11–41. Sverdrup–Thygeson, A. & Ims, R. A., 2002. The effect of forest clearcutting in Norway on the community of saproxylic beetles on aspen. Biol. Conserv., 106: 347–357. Thunes, K. H. & Willassen, E., 1997. Species composition of beetles (Coleoptera) in the bracket fungi Piptoporus betulinus and Fomes fomentarius (Aphyllophorales: Polyporaceae): an explorative approach with canonical correspondence analysis. J. Nat. Hist., 31: 471–486. Thunes, K. H., Midtgaard, F. & Gjerde, I., 2000. Diversity of coleoptera of the bracket fungus Fomitopsis pinicola in a Norwegian spruce forest. Biodivers. Conserv., 9: 833–852. Uotila, A., Kouki, J., Kontkanen, H. & Pulkkinen, P., 2002. Assessing the naturalness of boreal forests in eastern Fennoscandia. For. Ecol. Manage., 161: 257–277.

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

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Stachorutes cabagnerensis n. sp., Collembola (Neanuridae) from Central Spain, and a preliminary approach to phylogeny of genus J. C. Simón Benito, D. Espantaleón & E. García–Barros

Simón Benito, J. C., Espantaleón, D. & García–Barros, E., 2005. Stachorutes cabagnerensis n. sp., Collembola (Neanuridae) from Central Spain, and a preliminary approach to phylogeny of genus. Animal Biodiversity and Conservation, 28.2: 149–157. Abstract Stachorutes cabagnerensis n. sp., Collembola (Neanuridae) from Central Spain, and a preliminary approach to phylogeny of genus.— A new species of the genus Stachorutes, Stachorutes cabagnerensis n. sp., from central Spain is described. It is characterized by the presence of 6+6 eyes in the head, retinaculum 2+2 teeth, dentes with 5 hairs, and the absence of mucron. A phylogenetic analysis of this genus was attempted. Potential synapomorphies supporting the monophyly of Stachorutes are presented. One member of the genus (the Nearctic S. navajellus) appears as a basal form, phylogenetically distant from the remaining (Old World) species. There is evidence for a monophyletic infrageneric clade with the species S. dematteisi, S. jizuensis and S. sphagnophilus. However, more information is required for further phylogenetic resolution. Key words: Collembola, Stachorutes, Spain, Phylogeny. Resumen Stachorutes cabagnerensis sp. n., Collembola (Neanuridae) de la región central de España, y una aproximación preliminar a la filogenia del género.— Se describe una nueva especie del genero Stachorutes de la region central de España. Stachorutes cabagnerensis nov. sp. se caracteriza por la presencia de 6+6 ojos en la cabeza, retinaculum con 2+2 dientes y 5 sedas en cada rama del dentes; la furca carece de mucrón. Se ha efectuado un análisis filogenético. Las sinapomorfias potenciales establecen la monofilia del género. Una especie del mismo, S. navajellus, aparece como forma basal, filogenéticamente distante del resto de especies (Viejo Mundo). Se podría establecer un clado infragenérico con las especies S. dematteisi, S. jizuensis y S. sphagnophilus. Sin embargo, se precisa de mayor información para poder confirmarlo. Palabras clave: Colembolos, Stachorutes, España, Filogenia. (Received: 13 VIII 04; Conditional acceptance: 18 XI 04; Final acceptance: 13 I 05) José Carlos Simón Benito, David Espantaleón & Enrique Garcia–Barros, Unidad de Zoología, Depto. de Biología, Fac. de Ciencias, Univ. Autónoma de Madrid, Cantoblanco 28049, Madrid, Spain. Corresponding author: J. C. Simón Benito. E–mail: carlos.simon@am.es

ISSN: 1578–665X

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Introduction The genus Stachorutes was established by Dallai (Dallai, 1973) from Italian specimens from the province of Canzo, according to the type species Stachorutes dematteisi. Dallai characterized it for its small Antenna IV with simple sensillae, jaw with two teeth, simple maxillae, 2+2 eyes disposed as in the genus Micranurida, postantennal organ present, claw without empodial appendage, tenent hairs absent, reticulum and furca present, and without anal spines. Deharveng & Lienhard (1983) established two new species taken from the Eastern French Pyrenees and the Swiss Alps, and re–defined the genus. A further re–definition was later proposed by Jordana et al. (1997) in their monograph on Iberian Collembola. Thibaud & Palacios– Vargas (2000) added new characters, which resulted in an even sharper definition of the genus. A summary of the morphological features of potential use for defining Stachorutes is given in table 1. At present, Stachorutes Dallai consists of 16 species distributed all around the world. Eight species occur in Europe: S. cabagnerensis n. sp.; S. dematteisi Dallai, 1973; S. longirostris Deharveng & Lienhard, 1983; S. ruseki Kovac,1999; S. scherae Deharveng & Lienhard, 1983, S. sphagnophilus Slawaska, 1996; S. tatricus Smolis & Skarzynski 2001 (Smolis & Skarzynski, 2001), S. valdeaibarensis Arbea & Jordana, 1991. One species is known from Africa (S. dallai Weiner and Najt, 1998), three from North America: S. escobarae (Palacios–Vargas, 1990), S. maya Thibaud & Palacios–Vargas, 2000 and S. navajellus Fjellberg, 1994 and four in Asia: S. ashrafi (Yosii, 1966), S. jizuensis Tamura and Zhao, 1997; S. tieni Pomorski & Smolis, 1999 and S. triocelatus Pomorski & Smolis, 1999. The primary purpose of the present study was to describe a new taxon, S. cabagnerensis n. sp. However, since no attempt has currently been directed either at testing the monophyly of the genus, or at resolving its internal relationships, a preliminary approach to these questions was attempted. This task was complicated by the generally low degree of phylogenetic resolution of the group. However, given the small number of species, and the lack of information on the majority of them, any insight into the problem may help in facilitating taxa and character selection for further phylogenetic studies on pseudachorutine springtails. Material and methods

vided into six different units. In each of them, samples were collected according to the method designed. Individuals belonging to the genus Stachorutes were obtained in wooded units only. These locations were: Unit 1. Natural wood with vegetation of holly oaks Quercus ilex ilex L., oaks Quercus pyrenaica Wild. and cork oaks Quercus suber L., with undergrowth of cistus Cistus ladanifer L., heather Erica australis L. and Erica arborea L., arbutus Arbutus unedo L., located in the province of Ciudad Real in Navas de Estenas, UTM: 30SVJ5754. Samples 106–H, fallen leaves from holly oak, with heather and moss, nine specimens. Unit 2. Reforestation pinewood, Pinus pinaster Aiton, also in the province of Ciudad Real, in Horcajo de los Montes, UTM: 30SVJ6973. Sample 204–H2, fallen pine leaves after second year, three specimens. Sample 207–H, fallen pine leaves, two specimens. All samples were taken during the months of April and May (2001), a period characterised by a seasonal maximum in the numbers of both individuals and species (Simón, unpubl. data). Cladistic analysis The data matrix was analysed using the program Henning 86 (Farris, 1988; options h*, mh* and bb*), treating all multistate characters as unordered. As an alternative exploratory option, the successive weighting approach (Farris, 1969, 1989) was attempted with the same programme. All other analyses, as well as character state optimisation, were completed through Winclada (Nixon, 2002) and Nona (Goloboff, 1993) (strict, majority consensuses, as well as bootstrap and jacknife tests with 100 replicates). Since the number of species currently included in Stachorutes is low, an effort was made to include all of them in the analyses. This resulted in two problems related to the character state coding (see character list below): First, only partial information from S. ashrafi was available, and as many as 5 of the 13 characters were coded as unknown. Second, character 12 was found to be variable among individuals of S. longirostris from Pyrenean samples. Given the inability of the cladistic packages used in this study to make a different treatment of unknown vs. non comparable character states, this species was entered as four different taxa (a,b,c,d). We believe that these decisions may be acceptable given the prospective nature of our approach.

The new species of Stachorutes

Description

Soil samples were taken following the standard method designed for the research project "Bioasses". An area of 1 km2 was chosen in each selected land unit and a sample was taken every 200 m. A total of 16 units was prospected, each consisting of six spots representing different stages in development of the vegetation. In Spain, the National Park of Cabañeros was chosen and di-

Stachorutes cabagnerensis n. sp. Length of 0.66 mm in adult/s and 0.44 mm in young, dark blue in the adult/s, and light blue in the juveniles. Dorsal setae reduced, subequal, with sensillae longer and thicker than in normal hairs. Integument with thickly grained surface. Reduced mouth parts, jaw with 4 teeth, maxilla

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Table 1. Anatomical features of potential use in the definition of the genus Stachorutes, as stated by different authors (+ feature present; – feature absent): A. Dallai, 1973; B. Deharveng & Lienhard, 1983; C. Jordana et al., 1997; D. Thibaud & Palacios–Vargas, 2000. Note that some of these characters display no variation within Stachorutes, and were not coded for phylogenetic analyses. Tabla 1. Carácteres anatómicos de uso potencial en la definición del género Stachorutes, según distintos autores (+ caracter presente; – caracter ausente): A. Dallai, 1973; B. Deharveng & Lienhard, 1983; C. Jordana et al., 1997; D. Thibaud & Palacios–Vargas, 2000. Nótese que algunos de estos caracteres no presentan variación alguna dentro del género Stachourutes, y no fueron utilizados para los análisis filogenéticos.

Anatomical features

Antenna IV with 5–6 sensilla cylindrical and one microsensilla

Antenna IV with sensilla in flame–shape and one microsensilla

Antenna IV without hair–brush

Antenna IV and III join

Postantennal organ with vesicles simple never moruliform

Less than 8+8 eyes

Maxilla styliform

Empodial appendage

–

Tenent hairs

–

Furca reduced

Tenaculum

Mucron

–

+/–

–

Chaetotaxy dorsal reduced Macrochaetae Anal spines

styli form with two teeth,one ending in two apical teeth and the other in a hook with 5–6 teeth along the apical area. Labium without hair L, its relation with the length of the nail is 3 ( fig. 2). Ocular spot with 6 eyes, 3 anterior (A, B, C) and 3 posterior (E, F, G), the H and D are lacking (fig. 4). Oval post antennal organ with 13 to 15 vesicles similar in size, disposed in one line, twice longer than the nearest corneola and with approximately the same diameter (fig. 4). Antenna quite thick, the relation with the head diameter is 0.54 in adults and 0.58 in juveniles. Antenna I with 7 hairs, II with 12, III and IV are joined dorsally; segment III shows about 17 hairs and a sensorial organ formed by two microsensilla (si) as a war club and long thick lateral sensilla: (sgd) and (sgv) and an extra ventral microsensilla (sa). Antenna IV with normal hairs, straight, some small, without sensorial hair–brush in the ventral area. With 6 olfactory sensilla, 4 in the dorsal area (S1 to S4) forming a rhomb, and 2 in the ventro–apical area (S7, S8). Furthermore, there is a dorsal external microsensilla (m) and a very small distal organ. An apical tri–lobed vesicle is located at the apex of the antenna (fig. 5).

–

Reduced dorsal chaetotaxy (fig. 1), the position of the most internal sensilla S is: 3, 3/4, 4, 4, 4, 2 hairs, from thorax II to abdomen V. The formula of the dorsal inner hairs is 1, 3, 3/3, 3, 3, 2. Head without hair a0, the hairs d0 odd. Pronotum with 3+3 hairs. Mesonotum with a2. Tibiotarsi I, II, III with 19, 19, 18 hairs disposed in two whorls, the apical with 11 hairs and the basal with 8 hairs, except in the third pair of legs which shows 7 hairs, without tenent hairs (fig. 3). Ventral tube with 4+4 hairs, 2+2 basal and 2+2 apical. Claw without teeth and empodial appendage. Tenaculum with 2+2 teeth. Dens without mucron, and five hairs, manubrium with 7–9+7–9 hairs (fig.6). Genital orifice in the male with 18 hairs, in the female with 8 hairs in the anterior margin of the genital orifice, plus two central ones. Discussion The new species shows 6+6 eyes like two other species of this genus, ruseki Kovac, 1999 from Slovakia and ashrafi (Yosii, 1966) from Nepal. This latter species may however not belong to this genus because of the number of olfactory

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6 P4

Figs. 1–6. 1. Dorsal chaetotaxy; 2. Labium; 3.Tibiotarsus; 4. Eyes and postantennal organ; 5. IV and III antennal segment; 6. Furca. Figs. 1–6. 1. Quetotaxia dorsal; 2. Labio; 3. Tibiotarso; 4. Ojos y órgano postantenal; 5. Segmento antenal III y IV; 6. Furca.

hairs. Stachorutes cabagnerensis differs from both in the number of teeth of the tenaculum (2 compared to 3), and in the structure of the dentes (5 hairs against 6), and in the lack of mucron. Phylogenetic analysis A matrix of 13 characters (8 binary and 5 multistate) was prepared (table 2). The state 0 corresponds to the plesiomorphic character state. Rooting was done with reference to Pseudachorutes parvulus Börner, 1901.

Results Henning analysis produced 626 trees with a length of 50 steps (CI = 50, RI = 63). All characters except 2, 6 and 8 showed a certain degree of homoplasy, (table 3). The strict consensus of these cladograms is shown in fig. 7 (tree length, 76; CI, 32; RI, 25). On the assumption that Stachorutes represents a monophyletic assemblage, its monophyly can be supported by the synapomorphies 0:1 and 4:2 (that is, state 1 of character 0, and state 2 of

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Table 2. Matrix of species x character state. Pseudachorutes parvulus was used as outgroup: 0. Jaw (number of teeth: 0 = 4 teeth; 1 = 3 teeth; and 2 = 2 teeth); 1. Olfactory hairs of the Antenna IV (0 = tubular; 1 = in the shape of a sparkling flame); 2. PAO (shape of the postantenna organ: 0 = circular; 1 = elliptic); 3. Number of vesicles in the postantennal organ (0 = less than 10; 1 = 11 or more); 4. Number of eyes (0 = 8 eyes; 1 = 6 eyes; 2 = 5 eyes; 3 = 3 eyes; 4 = 2 eyes; 5 = 1 eye; 6 = 0 eyes); 5. Hair a0 in the head (0 = with this hair; 1 = without it); 6. Hair d0 in the head (0 = odd or without it; 1 = pair); 7. Number of hairs in the pronotum (0 = 3+3 hairs; 1 = 2+2 hairs; 2 = 4+4 hairs); 8. Hair a2 in the mesonotum (0 = with hair; 1 = without hair); 9. Number of hairs in the tibiotarsum I, II, III (0 = 19, 19, 18; 1 = another condition); 10. Number of teeth in the retinaculum (0 = 3+3 teeth; 1 = 2+2 teeth); 11. Number of hairs in the dentes (0 = 6 hairs; 1 = 5 hairs; 2 = 4 hairs; 3 = 3 hairs; 4 = 2 hairs; 5 = 1 hairs); 12. Mucron (0 = mucron separated from the dentes; 1 = mucrodens; 2 = absent). ? Indicates that the character state is unknown. Tabla 2. Matriz de especies x estados del carácter estudiado. Se utilizó a Pseudachorutes parvulus como grupo externo: 0. Mandíbula (número de dientes: 0 = 4 dientes; 1 = 3 dientes; and 2 = 2 dientes); 1. Sedas olfactorias de la antena IV (0 = tubular; 1 = en forma de llama de bujía); 2. PAO (forma del órgano postantenal: 0 = circular; 1 = elíptica); 3. Numero de vesículas del órgano postantenal (0 = menos de 10; 1 = 11 o más); 4. Número de ojos (0 = 8 ojos; 1 = 6 ojos; 2 = 5 ojos; 3 = 3 ojos; 4 = 2 ojos; 5 = 1 ojo; 6 = 0 ojos); 5. Seda a0 de la cabeza (0 = con esta seda; 1 = sin ella); 6. Seda d0 de la cabeza (0 = impar o sin seda; 1 = par); 7. Número de sedas del pronoto (0 = 3+3 sedas; 1 = 2+2 sedas; 2 = 4+4 sedas); 8. Seda a2 del mesonoto (0 = con seda; 1 = sin seda); 9. Número de sedas del tibiotarso I, II, III (0 = 19, 19, 18; 1 = otra condición); 10. Número de dientes del retináculo (0 = 3+3 dientes; 1 = 2+2 dientes); 11. Número de sedas del dentes (0 = 6 sedas; 1 = 5 sedas; 2 = 4 sedas; 3 = 3 sedas; 4 = 2 sedas; 5 = 1 sedas); 12. Mucrón (0 = mucrón separado del dentes; 1 = mucrodens; 2 = ausente). ? indican que se desconoce el estado del carácter citado.

Species

Pseudachorutes parvulus 0

10 11 0

12 0

Cosmopolitan

Distribution

Stachortes ashrafi

Nepal

S. cabagnerensis n. sp.

Spain

S. dallai

Tanzania

S. dematteisi

Italy

S. escobarae

Mexico

S. jizuensis

China

S. longirostris (a)

France (Pyrenees)

S. longirostris (b)

France (Pyrenees)

S. longirostris (c)

France (Pyrenees)

S. longirostris (d)

France (Pyrenees)

S. maya

Mexico

S. navajellus

USA and Canada

S. ruseki

Slovakia

S. scherae

Switzerland

S. sphagnophilus

Poland

S. tatricusa

Poland

S. tatricusb

Poland

S. tieni

Vietnam

S. triocellatus

Vietnam

S. valdeaibarensis

Spain

character 4). It is true, however, that the definition of the consensus tree was low, and that the monophyly of this and other pseudachorutine gen-

era awaits further and more thorough reassessment based on a taxonomically wider sample. All the species except S. navajellus seem to belong to

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Table 3. Performances of the characters 1–13 in the initial parsimony analysis (above; characters unordered, without weight), and in the strict consensus derived from these trees (below). The figures given are the number of steps in the tree, consistency index (CI) and retention index (RI). Tabla 3. Comportamiento de los caracteres 1–13 en el análisis de parsimonia inicial (arriba; caracteres no ordenados, sin peso), y en el consenso estricto derivado de estos árboles (abajo). Las cifras son el número de escalones en el árbol, el índice de consistencia (CI) y el índice de retención (RI).

Characters 1

Best fits (all trees) Steps

100

Worst fits (all trees) Steps

100

Number of steps, CI and RI in the consensus tree Steps

100

one clade supported by the synapomorphies 2:1, 5:1, 8:1 y 12:1, while navajellus shows the plesiomorphic state for most of these characters. Two clades stand out from the main group: the one formed by (S. dematteisi + (S. jizuensis + S. sphagnophilus)), characterized by 11:4, and (S. triocellatus + S. tieni), which can only be defined through homoplasies.

Bootstrap and jacknife analyses rendered the same results (not presented in detail). Due to the low resolving power of the data, speculations may be ventured on the basis of three alternative procedures: setting all multistate characters as ordered (with analysis proceeding as beforehand), successive weighting (multistate characters unordered), and majority consensus based on the set of

Figs. 7–10. 7. Unordered multistate characters, strict consensus (L = 76, Cl = 32, RI = 25, derived from 626 trees with L = 50, Cl = 50, RI = 63). Each circle represents one change in one character (filled = homoplay–free apomorphies, empty = convergences or reversals). The number above each circle is the number of characters, the one below it is the state of that character at that node. 8. Strict consensus tree, multistate characters ordered; consensus (L = 88, Cl = 28, RI = 33, derived from 235 trees with L = 60, Cl = 41, RI = 63). 9. Strict consensus, based on rescaled consistency index (L = 72, Cl = 34, RI = 30, derived from 11 trees with L = 177, Cl = 63, RI = 74). 10. Majority rule consensus, based on more than 600 trees. Figs. 7–10. 7. Caracteres multiestados no ordenados, consenso estricto (L = 76, CI = 32, RI = 25, derivados de 626 árboles con L = 50, CI = 50, RI = 63). Cada círculo representa un cambio en un carácter (lleno = apomorfias sin homoplasia, vacíos = convergencias o inversiones). El número que se halla sobre cada círculo es el número del carácter, y el de debajo el estado de dicho carácter en el nodo. 8. Árbol de consenso estricto, caracteres multiestados ordenados. Consenso (L = 88, CI = 28, RI = 33, derivados de 235 árboles con L = 60, CI = 41, RI = 63). 9. Consenso estricto, basado en un índice de consistencia re–escalado (L = 72, CI = 34, RI = 30, derivados de 11 árboles con L = 177, CI = 63, RI = 74). 10. Consenso de la mayoría, basado en más de 600 árboles.

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pseudac

pseudac 10

navajel

navajel 1

1 0

2 4

4 8 10 11

escobar

1 3 0 3 4 6 7 8 9 11

3 3 4 6 7 8 9 11

0 1 4 1 7 11

maya 0 4

2 2 4 6 8 10 11

longib

1 2 0 2 6 7 9 11

1 2

dallai

1 2 2 0 4 1 0 1 1 7 11

longic

1 2 8

0 0 0 3 4 8 10 12

2 6 1

2 1 0 3 4

0 1 1 0 1 2 0 4 11

1 1 1

2 1 0 0 3 4 11

tatricb 2 7

2 1 5 0 7 11

1 5

1 8

longici

1 0 8 11

scherae

0 0 3 4 11

0 3 4

ashrafi ruseki 2 6

0 4 10 11

2 6 0

3 1 3 0 11

1 1 1

escobar 6

4 11 0 4 2

4 10 11

4 2

1 9

5 0

1 1

dematte

2 11 12

dematte

0 1

dallai 4

2 8 10

0 2 4 11 12

0 1

2 4 0 1

tieni 2

2 0 4

0 4 8 10

triocel

7 9

tatricb

scherae

1 0 0

ruseki

tatrica

longid

1 1 0 4 11 12

ashrafi

1 1 4 12 1

cabagno

6 8 11 12

1 1

tatrica

8 12

1 1 1

cabagno

longic valdeai

valdeai

0 0 1 0 3 4 8 10 11 12 1 0 1 1

tieni

2 0 1 1 1 2 7

1 1 2 8 11

0 1 0 4

maya

0 1 4 1 2 0 1 2 7 11

2 0 1 2

longib 1 0

0 4

triocel

1 3

longia

longia 1 0

6 7 8 10 11

0 3 1 1 1 7 11

2 6 0 1 3 7 11

1 8 9

jizuen

5 5 0 10

1 1 1

sphagno

10 pseudac 10

pseudac navajel

navajel

1 0 4 10 11

valdeai

escobar

2 6 1 3 7 6 11

scherae

longia

1 1 3 0 3 4 6 7 9 11

tatrica

maya

0 1 4 1 2 1 2 7 9 11

100

longib

0 4

1 1 2 0 2 4 6 10

tatricb

ashrafi dallai

1 2

ruseki

2 0 4 1 1 7 8 1 1 2

longic valdeai

0 0 3 4 10 12

2 5 11 12 1 1 1 1

escobar

0 1 1 1 2 7 8 11

100

dallai 84

cabagno

dematte

longid

100

1 1 0 11

scherae

4 10 11 2 5 7 8 9 11 0 1 1 1

1 2

3 1 3 0

sphagn

triocel longia

tieni

longib

dematte 0

2 4 9 10 1112

2 0 4

1 1 4 2

4 11 1 8 1 1

5 5 0 10 0

longic

jizuens

longid maya

sphagn 84

tatrica 4

8 11

1 0

2 12

100

ruseki 0 4

triocel tieni

0 8

jizuens

100

tatricb

2 5

ashrafi

jizuens sphagn

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cladograms obtained in the first trial. Treating the multistate characters as ordered resulted in 235 cladograms (L = 60, CI = 41, RI = 63); the consensus tree L = 88, CI = 28, RI = 33 is shown in fig. 8. In contrast with the former results, S. dallai and S. escobarae were associated to the clade formed by (S. dematteisi + (S. jizuensis + S. sphagnophillus)), although no synapomorphy free of homoplasy was found to define this group. Further possible tree structures were suggested by the results of successive weighting (rescaled CI) and majority consensus. None of these trees represent a maximum parsimony solution. However, they are interesting to the extent that either the addition of new characters, or the recoding of characters where some states are non–applicable or unknown, might not support the present most parsimonious solution. These results suggested a potential link between S. escobarae, S. dallai and S. cabagnensis, and the clade formed by (S. dematteisi + (S. jizuensis + S. sphagnophilus)) (fig. 7), as well as a degree of relatedness between S. longirostris and S. maya with the clade (S. triocellatus + S. tieni) (fig. 9). Finally, although not strongly supported and based on homoplastic features, some relationship between the species (S. tatricus + S. ashrafi + S. ruseki) could be determined (fig. 10). Conclusions Although the overall resolution of the cladograms is low, a few points can be highlighted: (1) from a parsimonious point of view, the anatomical information currently available does not permit a detailed, highly resolved, phylogenetic hypothesis. This could only be solved by adding new characters; (2) some partial conclusions may be of interest for further studies: 1. The monophyly of the group can be provisionally supported on the basis of two synapomorphies (fig. 7, characters 0 and 4: jaw with 3 teeth, and five ocelli present). This interpretation requires further assessment as there is no out–group taxon to polarise character state changes at the basal node (i.e., external to Stachorutes + Pseudachorutes). Characters 2, 5, 8 and 12 support the monophyly of an infrageneric clade including all the other members considered in this study except navajellus. 2. The position of navajellus is peculiar for it apparently belongs to an isolated basal group or,it may have lost some features of the rest of the group due to reversal in characters 2, 5 , 8 and 12. 3. Within the ingroup, the resolution of the consensus is low. Only one clade with three species (dematteisi, jizuensis y sphagnophilus) can be defined with some accuracy, and even so, on the basis of one single nonhomoplasic synapomorphy (character 11, state 4). There is some evidence for the existence of two or three additional clades, but this is supported by homoplastic features only. The combination of the inferred phylogenetic information (e.g., fig. 10) and the known geographic

distributions of the species dealt with here results in strikingly broad geographic ranges at the supra– specific clades, in contrast with rather local specific distributions. Thus, for instance, the clade comprising S. tatricus, S. ashrafi and S. ruseki could be classified as of wide Palaearctic distribution (or Eurasian, e.g. Cox, 2001; Morrone, 2002), ranging from Western Europe to the Himalayas (Yosii, 1966; Smolis & Skarzynski, 2001). The clade comprising S. longirostris and S. tieni includes species of the Western Palaearctic, Nearctic, and Oriental regions. The three species represented in the best supported clade (S. dematteisi + S. jizuensis + S. sphagnophilus) were described from Central Europe, Italy, and the Yunnan region in South–Western China. The closest relatives of these three taxa include one East African and one Nearctic member (Dallai, 1973; Palacios–Vargas, 1990), together with the new species S. cabagnerensis for Spain. A highly conservative interpretation of such patterns is recommended. Moreover, the authors’ feeling is that either (a) very little insight on the phylogeny of the group has actually been gained, (b) the distributions of these springtail species are still very poorly known, or (c) an important number of related collembolan species in each of the main geographic areas mentioned have not yet been described. It is quite likely that all three hypotheses are equally pertinent. Acknowledgements This work was conducted with financial support from the European Union for the Research Project "Biodiversity assessment tools" (Grant Nº 0028, 2000). References Cox, C. B., 2001. The biogeographic regions reconsidered. Journal of Biogeography, 28: 511–523. Dallai, R., 1973. Ricerche sui Collemboli. XVI. Stachorutes dematteisi n. gen., s. sp., Micranurida intermedia n. sp. e considerazioni sul genere Micranurida. Redia, 54: 3–31. Deharveng, L. & Lienhard, C., 1983. Deux nouvelles espèces du genre Stachorutes Dallai, 1973 (Collembola). Revue suisse de Zoologie, 90: 929–934. Farris, J. S., 1969. A successive approximations approach to character weighing. Systematic Zoology, 18: 374–385. – 1988. Henning version 1.5, Reference guide. Published by the autor. Admiral Street, Port Jefferson Station, New York. – 1989. The retention index and the rescaled consistency index. Cladistics, 5: 417–419. Fjellberg, A., 1984. Collembola from the Colorado Front Range U.S.A. Arctic and Alpine Research, 16: 193–209. Jordana, R., Arbea, J. I., Simón, C. & Luciáñez, M.

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J., 1997. Collembola, Poduromorpha. In: Fauna Ibérica, V. 8: Museo Nacional de Ciencias Naturales, Madrid. Goloboff, P. A., 1993. Nona, version 2.0 for Windows. Inst. Miguel Lillo. Miguel Lillo 205, 4000 S. M. Tucumán, Argentina. Kovac, L., 1999. Stachorutes ruseki sp. n. (Collembola, Neanuridae) from Slovakia. Biologia, Bratislava, 54: 35–138. Morrone, J. J., 2002. Biogeographic regions under track and cladistic scrutiny. Journal of Biogeography, 29: 149–152. Nixon, K. C., 2002. WinClada version 1.00.08. Published by the autor. Ithaca, New York. Palacios–Vargas, J. G., 1990. Nuevos Collembola del estado de Chichuahua, México. Folia Entomológica Méxicana, 79: 5–32. Pomorski, R. J. & Smolis, A., 2000. Two new species of Stachorutes Dallai, 1973 from North Vietnam (Collembola, Neanuridae). Annales zoologici, 49: 151–156.

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Smolis, A. & Skarzynski, D., 2001. A new species of the genus Stachorutes Dallai, 1973 from Poland (Collembola: Neanuridae). Genus, 12: 407–410. Slawska, M., 1996. Stachorutes sphagnophilous n. sp. from Northern Poland (Collembola: Neanuridae). Genus, 7: 325–329. Tamura, H. & Zhao, L., 1997. Two new species of the family Pseudachorutidae from Mt. Jizu, western Yunnan, southwest China (Insecta: Collembola). Natural History Bulletin Ibaraki University, 1: 45–50. Thibaud, J. M. & Palacios–Vargas, J. G., 2000. Remarks on Stachorutes (Collembola: Pseuda– chorutidae) with a new Mexican species. Folia Entomologica Mexicana, 109: 107–112. Weiner, W. M. & Najt, J., 1998. Collembola (Entognatha) from East Africa. European Journal Entomology, 95: 217–237. Yosii, R., 1966. Collemboles of Himalaya. Journal of the College of Arts and Sciences, Chiba Univ., 4: 461–531.

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

Animal Biodiversity and Conservation 28.2 (2005)

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The status of marine turtles in Montserrat (Eastern Caribbean) C. S. Martin, J. Jeffers & B. J. Godley

Martin, C. S., Jeffers, J. & Godley, B. J., 2005. The status of marine turtles in Montserrat (Eastern Caribbean). Animal Biodiversity and Conservation, 28.2: 159–168. Abstract The status of marine turtles in Montserrat (Eastern Caribbean).— The status of marine turtles in Montserrat (Eastern Caribbean) is reviewed following five years of monitoring (1999–2003). The mean number of nests recorded during the annual nesting season (June–October) was 53 (± 24.9 SD; range: 13–43). In accordance with earlier reports, the nesting of hawksbill (Eretmochelys imbricata) and green (Chelonia mydas) turtles was confirmed on several beaches around the island. Only non–nesting emergences were documented for loggerhead turtles (Caretta caretta) and there was no evidence of nesting by leatherback turtles (Dermochelys coriacea); however, it is possible that additional survey effort would reveal low density nesting by these species. Officially reported turtle capture data for 1993–2003 suggest that a mean of 0.9 turtle per year (± 1.2 SD; range: 0–4) were landed island–wide, with all harvest having occurred during the annual open season (1 October to 31 May). Informed observers believe that the harvest is significantly under–reported and that fishermen avoid declaring their catch by butchering turtles at sea (both during and outside the open season). Of concern is the fact that breeding adults are potentially included in the harvest, and that the open season partially coincides with the breeding season. The present study has shown that although Montserrat is not a major nesting site for sea turtles, it remains important on a regional basis for the Eastern Caribbean. Key words: Caribbean, Eretmochelys imbricata, Hawksbill sea turtle, Chelonia mydas, Green sea turtle, Conservation. Resumen Estatus de las tortugas marinas en Montserrat (Caribe oriental).— Se ha estudiado la situación de las tortugas marinas en Montserrat (Caribe oriental) mediante un seguimiento de cinco años (1999–2003). El número medio de nidos registrados durante la estación anual de nidificación (junio–octubre) fue de 53 (± 24.9 SD; rango: 13–143). En concordancia con informes anteriores, se confirmó la nidificacón de las tortugas carey (Eretmochelys imbricata) y verde (Chelonia mydas) en varias playas alrededor de la isla. En la tortuga boba (Caretta caretta) sólo se registraron salidas sin nidificación, y no se encontraron pruebas de que la tortuga laúd (Dermochelys coriacea) nidificase; sin embargo, es posible que ulteriores estudios pongan de manifiesto una baja densidad de nidificación de esta especie. Los datos oficiales de capturas de tortugas (1993–2003) sugieren que en toda la isla llegaban a tierra una media de 0.9 tortugas anuales (± 1.2 SD; rango: 0–4), produciéndose todas las capturas cuando se había levantado la veda. Observadores bien documentados creen que las cifras de recolección están significativamente falseadas a la baja, y que los pescadores evitan declarar sus capturas sacrificando las tortugas en el mar (con la veda abierta o cerrada). Es preocupante que en esta caza puedan incluirse tortugas que crían, y que el período de captura permitida coincide en parte con la estación reproductora. Este estudio demuestra que aunque Montserrat no es un lugar principal de nidificación de las tortugas marinas, sigue siendo importante a escala regional en el Caribe oriental. Palabras clave: Caribe, Eretmochelys imbricata, Tortuga carey, Chelonia mydas, Tortuga verde, Conservación. ISSN: 1578–665X

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(Received: 14 X 04; Conditional acceptance: 14 II 05; Final acceptance: 25 IV 05) Corinne S. Martin, Dept. of Geographical and Life Sciences, Canterbury Christ Church Univ., Canterbury CT1 1QU, U.K.– John Jeffers, Dept. of Fisheries, Ministry of Agriculture, Government of Montserrat, Brades, Montserrat, West Indies.– Brendan J. Godley, Marine Turtle Research Group, Centre for Ecology and Conservation, Univ. of Exeter in Cornwall, Tremough Campus, Penryn TR10 9EZ, U.K. Corresponding author B. J. Godley. E–mail: bgodley@seaturtle.org

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Introduction Four species of sea turtles have been reported as nesting in Montserrat (Eastern Caribbean). Early studies suggested that the green (Chelonia mydas) and hawksbill (Eretmochelys imbricata) turtles nested in small numbers, whilst loggerhead (Caretta caretta) and leatherback (Dermochelys coriacea) turtle nests were only occasionally encountered (Meylan, 1983; John, 1984; Groombridge & Luxmore, 1989). A recent review of hawksbill turtle nesting in the Caribbean region (Meylan, 1999) reported that nesting in Montserrat is "incidental" although this result was based on reconnaissance of beaches and interviews cited in Meylan (1983). Meylan (1983) concluded that nesting levels were low, presumably because of constant human activity on the island’s beaches (which were widely used for boat storage and recreational purposes). Both adult and juvenile hawksbill and green turtles are found in Montserrat’s inshore waters (Meylan, 1983; John, 1984). Montserrat’s Turtle Ordinance (1951) states that turtles can be captured, sold and bought during an annual open season (1 October to 31 May). Although there are no quota or species restrictions, harvested turtles must weigh at least 20 lbs (ca 9.1 kg), and there are no restriction on the maximum size of harvested turtles. For several years now, the island’s fisheries authorities have been attempting to raise awareness about biodiversity conservation and turtle stock management issues among the island’s local fishermen. During these conversations, local fishermen are often verbally encouraged by the fisheries authorities to report any sea turtle catch to them (along their fish catches). It is not known, however, what proportion of fishermen actually report their turtle catches to the authorities. We present a five–year marine turtle monitoring dataset gathered with limited resources to elucidate spatial and temporal patterns of marine turtle nesting in Montserrat. The first estimates of sea turtle nest numbers for Montserrat are provided. In addition, available turtle capture data are presented, offering preliminary insights into the local marine turtle fishery. Material and methods Study site The Caribbean Island of Montserrat (62° 12’ W, 16° 45’ N) is part of the Leeward Islands of the Lesser Antilles. It is 104 km² in area and situated approximately 35 km southwest of Antigua and 60 km northwest of Guadeloupe (fig. 1; Blankenship, 1990). Apart from Trant’s and Farm beaches (east coast), all of Montserrat’s sandy beaches are located on the western side of the island (fig. 1). The island is of volcanic origin and all but one of its sandy beaches consist of black volcanic sand; white calcareous

sand dominates at Rendez–vous beach, the northern most beach on the island’s western side (Anonymous, 1993). The volcanic origin of the island was dramatically exposed in 1995 when the Soufrière Hills’ volcano located in the southern part of the island began exhibiting signs of volcanic activity. Since then, there has been an ongoing volcanic crisis, with an evacuation of the southern part of the island (including the Capital, Plymouth), a safety "exclusion zone" that covers almost two thirds of the island (fig. 1; Gell & Watson, 2000) and widespread human emigration and economic disruption. Nesting populations Day–time monitoring of marine turtle nesting The Fisheries Department of Montserrat’s Ministry of Agriculture has been coordinating the monitoring of beaches for turtle activities (including nesting, hatchling emergences, and nest excavations) since 1999. Although ad–hoc day–time beach monitoring has been carried out by dedicated island residents who regularly check local beaches for turtle emergences and nests, the bulk of the monitoring effort has been carried out by the Fisheries Department (J. J.). Monitoring frequency of nesting beaches has been uneven, being especially patchy (i.e. a few times a year) on the beaches located in the exclusion zone (fig. 1). Safe, accessible beaches were walked and checked for turtle tracks and nests on a fairly regular basis (i.e. up to twice a week at the peak of the nesting season). Beach monitoring datasheets were completed (by J. J.) each time a beach was visited, even if no nesting activity had taken place; other island residents did so only when they detected nesting activity. As a result, the number of beach monitoring sheets filled during a given period of time was only loosely indicative of the monitoring effort. Nests (N), i.e. adult emergences thought to have resulted in the deposition of a clutch of eggs, were individually counted. Non–nesting emergences (NNE) were not counted individually but, instead, their presence or absence on any given survey day was recorded. No distinction was made among species based on track morphology, as in many cases the nature of the beach, the type of substratum, the age of the tracks, and the relative inexperience of some the recorders precluded reliable species identification. All island beaches were monitored a minimum of once a week for one month (mid–August / mid– September) in 2003. Although one month of comprehensive survey was insufficient to accurately assess, on an annual basis, the extent of the spatial bias caused by uneven monitoring effort, it was thought sufficient for detecting any major underestimation of nesting activity for beaches relatively less monitored during the five year dataset (1999– 2003). Due to the relatively low nesting activity, monitoring beaches a minimum of once a week was sufficient to detect all activities occurring during the preceding week. More frequent monitoring

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Montserrat 1 2 4 km

Rendez–vous Bay Little Bay Carr's Bay Soldier Ghaut Bay Trant's Bay

Bunkum Bay

Farm Bay

Woodlands Bay Lime Kiln Bay

Fox Bay Bransby Point

Caribbean Sea

Blackburne Airport (abandoned)

Old Road Bay Illes Bay

Daytime Entry Zone

Hot Water Pond

Exclusion Zone (closed)

Plymouth (abandoned)

Montserrat

South America

Soufriere Hills Volcano

Kinsale

Sugar Bay

N O'Garro's Estate

Germans Bay

Fig. 1. The island of Montserrat in the Eastern Caribbean, showing nesting sites and the Exclusion Zone. Fig. 1. Isla de Montserrat en el Caribe oriental, mostrando los lugares de nidificación y la Zona de Exclusión.

(i.e. more than once a week) facilitated species identification based on track morphology (following Pritchard & Mortimer, 1999). Beaches were either walked or checked from a distance with binoculars (e.g. from a helicopter/boat). Special permission was granted from the authorities to access and walk some of the beaches of the exclusion zone (at the time Trant’s, Farm, Fox’s, Bransby Point, Hot Water Pond). In these surveys, individual non– nesting emergences and nests were counted. Night–time monitoring of marine turtle nesting In 2002 and 2003, logistics permitting, beaches were monitored at night for the presence of nesting turtles. When possible, nesting turtles were measured (Curved Carapace Length, CCL) and tagged subcutaneously with Passive Integrated Transponder (PIT) tags. Fishery harvest data Records of turtle harvests were obtained from Montserrat’s Fisheries Department in the form of a list detailing the month and year (1993–2003) of capture, the turtle species (if known), and the weight of the animal (in lb, if measured). The list had been compiled, over the years, by officers working at the island’s main harbours (Plymouth then Carr’s Bay). No other information is available, hence it is not known what percentage of the turtle catch these represent or if certain forms of fishing are over or under represented.

Results Nesting populations Day–time monitoring of marine turtle nesting For the five year dataset (1999–2003), data originating from a total of 453 beach monitoring forms were analysed. The mean annual number of nests was 53 (± 24.9 SD, range: 13–143). Records of non–nesting emergences (NNE) and numbers of nests (N) followed patterns similar to the monitoring effort (as defined by the number of completed beach monitoring datasheets) (fig. 2A). As could be predicted, the seasonality of nesting closely follows the seasonality of the monitoring effort (fig. 2B). The inventory of completed beach monitoring datasheets reveals that relatively little survey effort was expended annually during the five months between January and May, inclusive, and, given the seasonal pattern of nesting of the leatherback and loggerhead in the region, may in part explain the absence of documentation of theses species. Nevertheless, the collected information revealed that nesting activities followed a strong seasonal pattern, with 97% of activities (non–nesting emergences and nests) recorded between the months of June and October, clearly peaking in September (fig. 2B). During the monitoring period (1999–2003), Woodlands beach demonstrated the greatest nesting intensity of all beaches, but was also the most monitored beach on the island (fig. 3A). The three other

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B 200

Beech monitoring sheets

NNE

200

NNE

150 Frequency

Beech monitoring sheets

100

0 1999

2000

2001 Year

2002

2003

Mr Ap My Jn Jl Ag S O N Month of the year

Fig. 2. The total numbers of completed beech monitoring sheets, records of non–nesting emergences (NNE), and records of nesting emergences (N) by year and cumulatively (A) and by month (B) for the period 1999 to 2003: J. January; F. February; Mr. March; Ap. April; My. May; J. June; Jl. July; Ag. August; S. September; O. October; N. November; D. December. Fig. 2. Cifras totales de las hojas de control de las playas, registros de salidas sin nidificación (NNE) y registros de las salidas con nidificación (N) por: A. Año y acumulativamente; B. Mes para el período 1999–2003. (For abbreviations of fig. 2B see above.)

key nesting beaches appeared to be Rendez–vous, Fox’s Beach/Bransby Point and Old Road/Iles Bay beaches (fig. 3A). Preliminary results for the 2004 season indicated that Fox’s Bay was no longer a prime nesting site, while beaches near Plymouth (Hot Water Pond, Sugar Bay, Kinsale) showed increased nesting activities. During the study period (1999–2003), 594 nesting attempts (including 263 successful nests) were documented (table 1). During the more intensive monitoring period between mid–August and mid–September 2003, a total of 79 nesting attempts, including 19 successful nests were recorded (table 2). There were 21 non–nesting emergences and six nests from green turtles, and 17 non–nesting emergences and three nests from hawksbill turtles. Because of their relatively large widths, four asymmetrical tracks observed on Trant’s beach were attributed to loggerhead turtle(s), despite no nest being observed. The spatial distributions of non– nesting emergences and nests for mid–August/ mid–September 2003 (fig. 3B) showed patterns similar to those shown when all data are pooled for 1999–2003 (fig. 3A). The numbers of non–nesting emergences for mid–August/mid–September 2003 were highly correlated with the total numbers of recorded non–nesting emergences for the period 1999 to 2002 (Spearmans rank correlation Rs = 0.84; P < 0.01). This relationship in spatial pattern was also detected between the numbers of nest for mid–August/mid–September 2003 and the total number of nests for the period 1999 to 2002 (Rs = 0.57; P < 0.05).

Night–time monitoring of marine turtle nesting In 2002 and 2003, a total of 28 individual nesting turtles were measured: 16 green turtles (12 in 2002, four in 2003; mean CCL (cm) = 106.9 ± 6.3 SD, range: 103–118) and 11 hawksbill turtles (nine in 2002, two in 2003; mean CCL (cm) = 87.8 ± 6.8 SD; range: 79–103). A total of nine hawksbill (eight in 2002, one in 2003) and 13 green turtles (11 in 2002, two in 2003) were PIT tagged. All were tagged on Woodlands beach, with the exception of three hawksbill turtles tagged on Carr’s Bay (two in 2002, one in 2003). In 2002, two green turtles were re– sighted on Woodlands beach, 11 and 12 days respectively, after having been PIT tagged on that beach. These data were supplemented by one sighting (by a member of the public) of a loggerhead turtle nesting on Woodlands beach in August 2002 and hatchling leatherback turtles being discovered and filmed on the same beach in the mid 1990’s (J. J., unpublished data). Fishery harvest data For the period 1993 to 2003, the harvest of 10 turtles was declared to the Fisheries Department (fig. 4), hence a mean of 0.9 harvest per year (± 1.2 SD; range: 0–4). All captures took place during the open season (October to May). One green turtle (9.1 kg) and seven hawksbill turtles (13.6 kg, 18.1 kg, 29.5 kg, 45.4 kg, 45.4 kg, 63.1 kg, 90.9 kg; mean mass (kg) = 43.7 ± 26.9 SD) were declared to the authorities. There were two

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A 200

Beach monitoring sheets

NNE

18 Exclusion Zone

100 50

Frequency

150

NNE

Exclusion Zone

0 Rv Lt Cr SG Bn Wl LK OrI FB HSK GG TF Beach code

Rv Lt Cr SG Bn Wl LK OrI FB HSK GG TF Beach code

Fig. 3. The total numbers of completed beach monitoring sheets, records of non–nesting emergences (NNE) and numbers of nests (N), per beach (A) during the years 1999–2003 and (B) for the period mid–August to mid–September 2003. Beach codes: Rv. Rendez–vous; Lt. Little Bay; Cr. Carr’s Bay; SG. Soldier Ghaut; Bn. Bunkum Bay; Wl. Woodslands Beach; LK. Lime Kiln Bay; OrI. Old Road/Iles Bay; FB. Fox’s Bay/Bransby Point; HSK. Hot Water Pond/Sugar/Kinsale; GG. German’s/O’Garro’s; TF. Trant’s/Farm Bay); * One hawskbill turtle nest; ** Two hawksbill turtle nests. (For other abbreviations see figure 2.) Fig. 3. Cifras totales de las hojas de control de las playas, registros de las salidas sin nidificación (NNE) y números de nidos (N) por playa (A) durante los años1999–2003 y (B) para el período de mediados de agosto–mediados de septiembre del 2003. Códigos de las playas: Rv. Rendez–vous; Lt. Little Bay; Cr. Carr’s Bay; SG. Soldier Ghaut; Bn. Bunkum Bay; Wl. Woodlands Beach; LK. Lime Kiln Bay; OrI. Old Road/Iles Bay; FB. Fox’s Bay/Bransby Point; HSK: Hot Water Pond/Sugar/Kinsale; GG. German’s/ O’Garro’s; TF. Trant’s/Farm Bay; * Un nido de tortuga carey; ** Dos nidos de tortuga carey. (Para las otras abreviaturas ver la figura 2.)

declared captures for which the species was not recorded. All reported landings were of turtles that met the legal minimum size criteria (> 20 lbs, or 9.1 kg). Using a published regression equation between mass and CCL for hawksbill turtles (Log10 (mass) = 2.8966 * Log10 (CCL) – 3.8534, with mass in kg and CCL in cm, Limpus et al., 1983), the masses of nesting hawksbill turtles that were measured in Montserrat were estimated to range from 43.9 to 94.8 kg. When compared to the masses of harvested turtles, it appeared that four out of the seven harvested hawksbill turtles declared to authorities could have been adults. These four potentially adult turtles were captured during the months of February (N = 1 turtle), October (N = 1 turtle) and November (N = 2 turtles). Discussion Although marine turtle monitoring had been ongoing since preliminary studies in the early 1980’s (Meylan, 1983; John, 1984), almost all relevant data were lost, along with many government records, in the volcanic flows that engulfed Plymouth in

1997. Monitoring efforts documented by this study (1999–2003) were intermittent and uneven, meaning that caution is warranted in making any recommendation regarding population status. There are, however, a few key points that can be extracted from the existing data. Green and hawksbill turtles nest in modest yet regionally important numbers for the Eastern Caribbean, probably every year. Leatherback and loggerhead turtles may also nest, but at lower densities. The lack of documented leatherback nesting may be attributed to a comparatively low level of monitoring during peak nesting months (April–June), however it is unlikely that nesting of this species is more frequent than occasional. The data are in concord with the wider literature which suggests that green, hawksbill and leatherback turtles (and loggerheads to a much lesser extent) are the most common species of nesting sea turtles in the Lesser Antilles (e.g. Carr et al., 1982; Meylan, 1983, 1999; Eckert et al., 1992; Eckert & Honebrink, 1992; Fuller et al., 1992; Sybesma, 1992; D’Auvergne & Eckert, 1993; Scott & Horrocks, 1993; Richardson et al., 1999; Chevalier & Lartiges, 2001). The magnitude of nesting data recorded was closely correlated with survey frequency in time

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Table 1. Breakdown of the records of non–nesting emergences (NNE), and the numbers of nests (N), per beach and year, for the period 1999–2003. Tabla 1. Detalle de los registros de salidas del mar sin nidificación (NNE), y número de nidos (N), por playa y por año, para el período 1999–2003.

1999

2000

2001

2002

2003

NNE

Rendez–vous

Little

Carr’s

Soldier Ghaut

Bunkum

Woodlands

Lime Kiln

Old Road/Iles

Hot Water Pond/Sugar/Kinsale

German’s/O’Garro’s

Fox’s/Bransby Point

Trant’s/Farm Total

165

143

Table 2. Breakdown of the numbers of non–nesting emergences (NNE), and the numbers of nests (N), per beach and by species, for the period mid–August to mid–September (2003). Tabla 2. Detalle de los registros de salidas del mar sin nidificación (NNE), y número de nidos (N), por playa y por especie, para el período mediados de agosto–mediados de septiembre (2003).

Green NNE

Hawksbill N

NNE

Loggerhead NNE

Undetermined NNE

Rendez–vous

Little

Carr’s

Soldier Ghaut

Bunkum

Lime Kiln

Old Road/Iles

Fox’s/Bransby Point

Hot Water Pond/Sugar/Kinsale

German’s/O’Garro’s

Trant’s/Farm

Woodlands

Total

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Hawksbill

Green

Undetermined

Turtle captures

2 Closed season 1

0 J

Mr Ap My Jn Jl Ag Month

Fig. 4. The temporal distribution of reports of turtle captures (1993–2003; N = 10 turtles). The closed season is highlighted. (For abbreviations see figure 2.) Fig. 4. Distribución temporal de los registros de capturas de tortugas (1993–2003; N = 10 tortugas). Se ha destacado la estación de veda. (Para las abreviaturas ver la figura 2.)

and space. It is likely that recorders more frequently carried out surveys at times and locations when the probability of recording turtle nesting activity was more likely. Although this may have resulted in spatial and temporal biases in the dataset, the seasonality of the Montserrat nesting season as described by the data set is plausible, peaking from June to October, if we assume that hawksbill and green turtles are the dominant nesting species. Although because of the nature of the data, it was not possible to discriminate between the seasonality of the different species, the temporal distribution of the data are consistent with seasonality of nesting reported for hawksbill and green turtles (Fuller et al., 1992; Hirth, 1997) and hawksbill turtles (Eckert & Honebrink, 1992; Corliss et al., 1989; Scott & Horrocks, 1993) in the Eastern Caribbean region. Additionally, data collected during the period of intensive monitoring in 2003 generated a spatial distribution of nesting broadly similar with that of the data gathered in previous years. Notwithstanding, it is likely that comprehensive (e.g. once weekly, year– around) island–wide surveys would reveal more complex patterns of habitat use by gravid females. The key nesting beaches for green and hawksbill turtles in Montserrat appeared to be Woodlands (so far unreported in the literature), Rendez–vous, Fox’s/Bransby Point and Old Road/Iles Beaches. Even though green turtles left tracks on many of the island’s beaches, actual nesting by this species was only confirmed for Rendez–vous, Woodlands and Fox’s/Bransby Point beaches. Based on interviews with island residents and beach reconnaissance, Meylan (1983) reported that green turtles might also be nesting at Little and Iles beaches.

Actual nesting by hawksbill turtles was solely confirmed in the present study for Rendez–vous and Old Road/Iles Beaches, although Meylan (1983) also quotes Carr’s, Little and Soldier Ghaut beaches as nesting sites for this species. On Trant’s beach, tracks possibly left by loggerhead turtles were reported, in agreement with the belief that loggerhead turtles occasionally nest on the island (John, 1984). It is thought that the turtle fishery has declined significantly in magnitude since the extensive emigration from the island in recent years. Only ten turtles were declared to the fishing authorities for the period 1993 to 2003. Popular accounts suggest that it is likely that this low total is the result of significant under–reporting. Fishermen are said to avoid declaring their catch to the authorities by butchering turtle carcasses at sea both in and outside the open season. Of great concern, as evidenced by the temporal distribution of declared turtle capture records and the fact that potential breeding adults are being captured, is that the open season for the turtle fishery overlaps partially with the nesting season. Consequently, in planned regulations, it has been suggested that the closed season be defined as 1 March (the beginning of leatherback nesting season in the central Eastern Caribbean) to 1 December. Other suggested changes in the regulation include the prohibition of catching turtles on land and an increase of the minimum weight of harvested turtles from 20 pounds (9.07 kg) to 50 pounds (22.68 kg). However, a recent report to the UK Government (Godley et al., 2004) recommended that legislation be further revised to "ensure a permanent and complete prohibition of harvest of any large,

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reproductively valuable turtles by instigating a maximum size limit". It was suggested that this threshold should be based upon further research into the fishery and turtle stocks and that a curved carapace length threshold was developed. A shift from weight–based to size–based limits enables a fisherman to more easily determine the legality of the catch while still at sea. The harvest information suggests that a wide size range of green and hawksbill turtles could be present year round in Montserrat’s waters. Relatively little is known of the current state of Montserrat’s marine and coastal habitats with regards to suitability as marine turtle foraging areas. The area of the coastal shelf is relatively small (140 km2) and only generalized distributions of primary substrata types are available (Meylan, 1983; Anonymous, 1993). Before 1995, coral communities (foraging habitats for hawksbill turtles) were found in small patches interspersed with sand and sediment on the north, south and west coasts (Gell & Watson, 2000). The harmful consequences of sediments on coral reef communities and associated organisms are well documented (e.g. Rogers, 1990). In Montserrat, volcanic sediments are thought to have had a severe impact on reef growth, particularly those in the east and southwest of the island (Gell & Watson, 2000). Direct deposits of ash and waterborne sediments have led to some coral bleaching and disintegration of large sponges. Some reef areas, however, are thought to be recovering (Wolfe Krebs, pers. comm. 2003). In recent times, three main seagrass beds (foraging habitats for green turtles) were known: the largest, 750 ha, being located at the northern tip of the island and the other two on the east and west coasts (Gell & Watson, 2000). It is thought that seagrass beds suffered considerable damage during Hurricane Hugo in 1989, although the effect on the spatial distribution of foraging habitat for green turtles is not known. Montserrat presents a relatively narrow coastal shelf, dropping off rapidly to nearly 200 m only 650 m from the shoreline along the southern half of the island, whilst in the north, northeast and west, the shelf slopes more gently (the 200 m contour is approximately 5 km offshore, Gell & Watson, 2000). The result is a high energy, erosion prone coastline, with generally intermittent beaches (Anonymous, 1993). For this reason, the quality of Montserrat’s beaches with regards to sea turtle nesting appears to be naturally poor. Although only assessed qualitatively to date, beach erosion destroys incubating eggs and periodically prevents gravid turtles from nesting. Additional factors of concern are linked to the volcanic eruptions and include ash deposits and beach mining. Occasional ash deposits cover nesting beaches, rendering them less suitable or wholly unsuitable for nesting until they are cleared by heavy storms. For Montserrat’s rebuilding after the catastrophic eruptions of 1997, extraction of beach sediments, largely of volcanic origin, are commonplace. Such extraction has ceased at Isle’s Bays (in 2003) but is ongoing at Trant’s Bay. It is important

that the integrity of Trant’s Bay be maintained and that ongoing sea turtle monitoring, preferably on a more frequent basis, include the relocation of clutches from high risk to lower risk beach areas. Nest predation by feral pigs and feral/domestic dogs has also been recorded (J. J. and B. J. G. pers. obs.), but the actual levels are yet to be quantified. The present study has drawn a more accurate picture of the status of marine turtles in Montserrat. Further studies involving species identification with increased survey effort will more fully elucidate the status of nesting populations. Of high priority for marine turtle conservation are a revision of the regulatory framework to feature a more restricted harvest season (and one that does not coincide with the turtle breeding season), maximum rather than minimum size limits, new measures to encourage fishermen to report their turtle catches, the full protection of nesting adults, their eggs and young, the careful management of beach sediment extraction, and the control of feral pigs and feral/domestic dogs. Acknowledgements The authors would like to thank the staff of Montserrat Fisheries Department, Montserrat Governors Office, Montserrat Ministry of Agriculture, Montserrat National Trust, Montserrat Volcano Observatory, Royal Society for the Protection of Birds, Sea Wolf Diving School, and the following individuals: Crystal & Dean Archer, Mrs Hilda Blake, Helen & Gerard Cooper, Bo Dalsgaard, Mr & Mrs Darby, Alfred Edwards, Lexvern Fenton, Anne–Marie & David Graham, Gerard Gray, Linda Green, John Keller, Mr & Mrs Krebs, Melissa O’Garro, Geoff Patton, Joe Philips, Sarah Sweeney , Mr & Mrs Walker. Much of the fieldwork for this study was carried out as part of the Turtles in the Caribbean Overseas Territories (TCOT) project funded by DEFRA and the FCO’s UK Overseas Territories Environment Fund. BG is a NERC Research Fellow. Time to support final manuscript preparation was provided through funding by the Overseas Territories Environment Programme (OTEP) for the Turtles in the UK Overseas Territories (TUKOT). The manuscript considerably benefited from the comments of Catherine Bell, Annette Broderick,, Claude Gerald, Matthew Godfrey and Kartik Shanker and two reviewers. References Anonymous, 1993. Environmental Profile, An Assessment of the Critical Environmental Issues Facing Montserrat with an Action Agenda for the Future. United Nation Development Program (UNDP), Project No. MOT/92/002/A/01/99 Blankenship, J. R., 1990. The wildlife of Montserrat (including an annotated bird list for the island). Montserrat National Trust, Montserrat, West Indies.

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Carr, A., Meylan, A. B., Mortimer, J., Bjorndal, K. A. & Carr, T., 1982. Survey of sea turtle populations and habitats in the Western Atlantic. NOAA Technical Memorandum NMFS–SEFC 91, U.S. Department of Commerce. Chevalier, J. & Lartiges, A., 2001. Les Tortues Marine des Antilles, Etude Bibliographique. Office National de la Chasse et de la Faune Sauvage, CNERA Faune d’Outre Mer. Corliss, L. A., Richardson, J. I., Ryder, C. & Bell, R., 1989. The hawksbills of Jumby Bay, Antigua, West Indies, In: Proceedings of the Ninth Annual Workshop on Sea Turtle Conservation and Biology: 33–35 (S. A. Eckert, K. L. Eckert, T. H. Richardson, Eds.). NOAA Tech. Memo. NMFS– SEFC–232. U. S. Department of Commerce. D’Auvergne, C. & Eckert, K. L. 1993. WIDECAST Sea turtle Recovery Action Plan for St Lucia, CEP Technical Report n°26. In: UNEP Caribbean Environment Programme: 1–70 (K. L. Eckert, Ed.). Kingston, Jamaica. Eckert, K. L. & Honebrink, T. D., 1992. WIDECAST Sea turtle Recovery Action Plan for St Kitts and Nevis, CEP Technical Report n°17. In: UNEP Caribbean Environment 292 Programme: 1–92 (K. L. Eckert, Ed.). Kingston, Jamaica. Eckert, K. L., Overing, J. A. & Lettsome, B. B., 1992. WIDECAST Sea turtle Recovery Action Plan for British Virgin Islands, CEP Technical Report n°15. In: UNEP Caribbean Environment Programme: 1–116 (K. L. Eckert, Ed.). Kingston, Jamaica. Fuller, J. E., Eckert, K. L. & Richardson, J. I., 1992. WIDECAST Sea turtle Recovery Action Plan for Antigua and Barbuda, CEP Technical Report n°16. In: UNEP Caribbean Environment Programme: 1– 88 (K. L. Eckert, Ed.). Kingston, Jamaica. Gell, F. & Watson, M., 2000. UK Overseas Territories in the Northeast Caribbean: Anguilla, British Virgin Islands, Montserrat. In: Sea at the Millennium: an Environmental Evaluation: 615–626 (C. Sheppard, Ed.). Pergamon, Elsevier Science Ltd., United Kingdom. Godley, B. J., Broderick, A. C., Campbell, L. M., Ranger, S., Richardson, P. B., 2004. An Assessment of the Status and Exploitation of Marine Turtles in the UK Overseas Territories in the Wider Caribbean. Final Project Report to the Department of Environment, Food and Rural Affairs and the Foreign and Commonwealth Office: 1–253. Available online at: http://www.seaturtle.org/mtrg/ projects/tcot/

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Groombridge, B. & Luxmoore, R., 1989. The green turtle and hawksbill (Reptilia: Cheloniidae) world status, exploitation and trade. Lausanne, Switzerland. Hirth, H. F., 1997. Synopsis of the biological data on the Green turtle Chelonia mydas (Linnaeus 1758). Biological Report 97(1), Fish and Wildlife Services, U.S. Department of Interior. John, C. T., 1984. The national report for the country of Montserrat. In: Proceedings of the Western Atlantic Turtle Symposium: 3.332– 3.328, volume 3, appendix 7, the national reports (P. Bacon, F. Berry, K. Bjorndal, H. Hirth, L. Ogren & M. Weber, Eds.). Univ. of Miami Press, Miami. Limpus, C. J., Miller, J. D., Baker, V. & McLachlan, E., 1983. The hawksbill turtle, Eretmochelys imbricata (L.), in the North–Eastern Australia: the Campbell Island Rookery. Australian Wildlife Research, 10: 185–197. Meylan, A. B., 1983. Marine turtles of the Leeward Islands, Lesser Antilles. Atoll Research Bulletin, 278: 1–43. – 1999. Status of the hawksbill turtle (Eretmochelys imbricata) in the Caribbean region. Chelonian Conservation & Biology, 3: 177. Pritchard, P. C. H. & Mortimer, J. A., 1999. Taxonomy, External Morphology, and Species Identification. In: Research and Management Techniques for the Conservation of Sea Turtles: 21–38 (K. L. Eckert, K. A. Bjornda, F. A. Abreu– Grobois & M. Donnelly, Eds.). IUCN/SSC Marine Turtle Specialist Group Publication No. 4. Richardson, J. I., Bell, R. & Richardson, T. H., 1999. Population Ecology and Demographic Implications Drawn From an 11–Year Study of Nesting Hawksbill Turtles, Eretmochelys imbricata, at Jumby Bay, Long Island, Antigua, West Indies. Chelonian Conservation and Biology, 3: 244–250. Rogers, C., 1990. Responses of coral reefs and reef organisms to sedimentation. Marine Ecology Progress Series, 62: 185–202. Scott, N. & Horrocks, J. A., 1993. WIDECAST Sea turtle Recovery Action Plan for St. Vincent and the Grenadines, CEP Technical Report No. 27. In: UNEP Caribbean Environment Programme: 1–80 (K. L. Eckert, Ed.). Kingston, Jamaica. Sybesma, J., 1992. WIDECAST Sea turtle Recovery Action Plan for the Netherlands Antilles, CEP Technical Report No 11. In: UNEP Caribbean Environment Programme: 1–63 (K. L. Eckert, Ed.). Kingston, Jamaica.

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Conservation implications of wild animal biomass extractions in Northeast India Hilaluddin, R. Kaul & D. Ghose

Hilaluddin, Kaul, R. & Ghose, D., 2005. Conservation implications of wild animal biomass extractions in Northeast India. Animal Biodiversity and Conservation, 28.2: 169–179. Abstract Conservation implications of wild animal biomass extractions in Northeast India.— We investigated the patterns of wild meat extraction and consumption by indigenous communities in Northeast India. Our respondents hunted at least 134 species of wild animals over the previous year in the villages surveyed and continued to harvest and use wild meat as their cash income increased. These indigenous communities of Northeast India showed an average of 32 to 59% dependency on the forestry sector. Wild meat contributed significantly (up to 25%) to their economies, suggesting previous assessments of dependence on the forestry sector should be reviewed. All sections of the society exploited wild meat equally. As education seems to play a role in reducing wild meat extractions, increased awareness in conservation of natural resources should be promoted . Key words: Wild meat consumption, Wild meat trade, Dependency, Northeast India. Resumen Repercusiones en la conservación debidas a las extracciones de biomasa animal salvaje en el nordeste de la India.— Investigamos los patrones de extracción y consumo de carne de caza por parte de las comunidades indígenas del nordeste de la India. En la aldea estudiada, los sujetos interrogados habían cazado al menos 134 especies de animales salvajes durante el año anterior, y continuaron cazando y utilizando la carne de caza cuando sus ingresos aumentaron. Estas comunidades indígenas del nordeste de la India dependían en promedio del 32 al 59% del sector forestal. La carne de caza contribuía significativamente (hasta un 25%) a sus economías, lo que sugiere que deberían revisarse las evaluaciones previas sobre la dependencia del sector forestal. Todas las capas sociales explotaban la carne de caza de igual forma. Dado que parece que la educación juega un papel significativo en la reducción de las extracciones de carne de caza, debería promoverse una mayor concienciación de la conservación de los recursos naturales. Palabras clave: Consumo de carne de caza, Comercio de carne de caza, Dependencia, Nordeste de la India. (Received: 11 VIII 04; Conditional acceptance: 25 I 05; Final acceptance: 3 V 05) Hilaluddin, Rahul Kaul & Dipankar Ghose, World Pheasant Association, South Asia Field Office, J–7/21, DLF Phase II, Gurgaon–122 002. Corresponding author: Hilaluddin. E–mail. hilaluddin@yahoo.com

ISSN: 1578–665X

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Introduction In Northeast India, people hunt wild animals for several reasons and therefore rural people are heavily dependent on wild meat (Hilaluddin, 2005a, 2005b). However, game may often be over–hunted and may have caused local extinctions of several species such as the Green peafowl Pavo muticus in Southeast Asia (McGowan et al., 1998). The problem can only be tackled by looking at the wider economic and institutional context within which hunting occurs, from household economies to trade (Abernethy et al., 2003). However, quantitative data on the amount of wild meat harvest, its consumption and trade in Southeast Asia in general and Northeast India in particular, are lacking. Therefore, there is an urgent need to quantify the intensity of wild meat extractions and assess impacts of such extractions on wild animal populations. There has also been little work to determine the contribution of the forestry sector to the life of local people (Bahuguna, 1993). Economic benefits accruing to local economies from the forests have seldom been estimated (Bahuguna, 2000) and in most cases are incomplete. The economic value of animal biomass may have been significant but it was often ignored in earlier assessments which mainly pertained to timber and non–timber forest products (fuelwood, fodder, fruits, seeds, medicinal derivatives of plants, etc.). People’s dependence on wild meat, in particular, remains unknown despite harvesting of roughly 23,500 tonnes annually in Sarawak (Bennett, 2002), 67,000–1,64,000 tonnes in the Brazilian Amazon (Robinson & Redford, 1991; Peres, 2000) and 1 million to 3.4 million tonnes in Central Africa (Wilkie & Carpenter, 1999; Fa et al., 2002). There is also a need to assess the benefits derived from the wild meat in order to demonstrate the tangible contribution of the forestry in general and wild meat in particular to the society. This is also essential to understand the significance of wild meat in the local economy —both for cash and subsistence needs— and local cultural beliefs (Abernethy et al., 2003). The economic theory of "Income and Consumption" (Kuznets, 1955), which is now used world–wide in most natural resources conservation action plans, suggests that consumptions of a commodity go up with an increase in household income if it has no substitutes or is considered superior to substitutes. Otherwise, the use of goods falls with rising income, showing inverted "U shape" patterns. Kuznets’ model of consumption may not be universally applicable to all goods, however, even if they are inferior, especially in regions of the world where people have developed a taste for a few specific goods for reasons other than economic. His model may thus vary across the nature of goods and areas. Therefore, there is a need to investigate the validity of Kuznets (1955) model in consumption of important forest products such as wild meat before its incorporation into a conservation action plan.

We undertook a survey in Northeast India to assess whether the extraction of wild meat by Angami, Apatani, Mizo and Nishi communities was a conservation problem in the region. Specifically, we sought to determine whether consumption of wild meat was linked to people’s income. In order to answer this question we studied the prevalence of wild meat extraction and consumption, the species hunted and differences in hunting patterns of indigenous communities, the linkage between wild animal hunting and trade, the role of wild meat in local economy, and the impact of education, age and profession of a person on wild meat extraction. We also collected information on other forest products harvested by a household in order to calculate income of that household from the forestry. The amounts of all forest products extracted by a household are quantified and their quantities are converted into monetary values based on their prevalent spot prices for estimating a household income from forestry (Malhotra et al., 1991; Hedge et al., 1996). According to Bahuguna (1993, 2000), the income of a household must be calculated by summing incomes of that household from all sources viz. agriculture (labour and crops), forestry (forest products and forest management activities) and other employment opportunities (self and government employment). Methods The survey included three methods: A general village level survey, a household level survey, and finally a market survey. Animal extraction data were collected by way of a detailed set of questionnaires and were not independently measured amounts. The qualitative and quantitative information both at village/hamlet and household level on the animal extraction patterns was gathered following a combination of PRA (Sankaran et al., 2000) and RRA (Sethi & Hilaluddin, 2001) methods. We collected information on the animal species and their number(s) killed by a household during the previous year. The respondents were shown pictures of animal species for the purpose. A total of 25 villages were surveyed, representing four communities (Angamis 6; Apatanis 5; Nishis 8; Mizos 6). The villages were from the interior and exterior forest blocks among the settlements of the studied communities, thereby covering most of their habitation ranges. Generally, one interview with a group of villagers was conducted at the village/hamlet level. During this interview we sought wide–ranging information about the resource use patterns (those interested in the questionnaire and the list of species hunted with their numbers will be sent the information upon request to the author) Such interactions were usually a good introduction to the purpose of our surveys, and subsequent data collection at the village level became easier (Hilaluddin et al., in press b).

Animal Biodiversity and Conservation 28.2 (2005)

2 4 1 3 India

Fig. 1. Locations of surveyed villlages in Northeast India: 1. Angami; 2. Apatani; 3. Mizo; 4. Nishi. Fig. 1. Localización de las aldeas del nordeste de la India estudiadas: 1. Angami; 2. Apatani; 3. Mizo; 4. Nishi.

After the village level focal group interviews, we were able to focus on individuals involved in some level of forest produce gathering. A total of 134 household level interviews (Angamis 33; Apatanis 33; Nishis 30; Mizos 38) were conducted in Aizwal, Kohima and Lower Subansiri districts of Northeast India (fig. 1). Following these interviews, we divided familiy units into hunting or non–hunting households. We further classified them into business, farming and service communities. We selected respondents for household level interviews following random sampling techniques (Sutherland, 1996). Sampling efforts covered at least five percent for each category of households (hunting and non–hunting) in each surveyed village and town for collecting data on range of animals extracted by the household. We also gathered information on agriculture crops and wild plants (timber, firewood, fodder, bamboo, medicinal plants and other NTFPs) products, with their prevailing spot price, gathered by the household during the previous year. In addition, a household’s income from other avenues (agriculture labour, forestry labour and other employment opportunities) was also quantified. We also collected data on age, education status and size of the respondent’s family. Educational level was assessed from the number of school years (1–15) he/she had passed from a recognized institution. The respondents in the household level interviews were mainly selected randomly but sometimes on the advice of our guide who hailed from the village. If both a man and a woman from the household were present, we interviewed the man because only male members, within the indigenous communities studied, hunt wild animals.

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We also conducted wild meat trade surveys for a period of 15 days each in the local markets of Kohima city (Nagaland) and Hapoli town (Arunachal Pradesh). The main purpose of this survey was to establish whether there was trade of wild meat in urban centers and also whether these markets connected to the remote areas of our survey sites. We recorded species being sold in the markets with their numbers and price. Data on the intensity of hunting within a village was calculated from the estimated number of animals killed by each household/annum for each species. Crude wildmeat amount extracted by a household for each species was calculated using the average body weight of adult individuals. Mean body masses of animals were taken from the literature (Prater, 1971; Ali & Ripley, 1987) with the exception on fishes. Information on the quantities of extraction of fishes and other forest products by a household were directly gathered in per unit measurement in the field. We calculated a household’s income from forestry by converting quantities of wild animal and plant species extracted by that household into monetary values based on their prevalent spot prices. We also included the income of that household from forest management activities such as forestry labour, nursery, and forest watch and ward activities. The gross annual incomes of households were calculated by summing their incomes from various income sectors viz. agriculture (crops and labour), forestry (plants, animals and employment through forest management activities) and other employment opportunities (self and government employment). We investigated the relationship between wild meat extraction and consumption rates of Angami, Apatani, Mizo and Nishi communities using Independent sample t test because these communities harvest wild meat both for self–use and for sale. The impact of socio–economic variables, specifically age, educational status, and incomes derived from cash avenues on wildmeat extraction rates, were calculated using Pearson’s correlation coefficient. The differences in mean values of wild meat extracted by people in different occupations were investigated using the Kruskal–Wallis test. The monetary significance of wildmeat extractions to local economies of Angami, Apatani, Mizo and Nishi communities were examined using One Way ANOVA and therefore the null hypothesis "the variations in mean values of dependency sources were statistically non–significant" was tested. This was used to infer whether dependencies of a household on each source contributed significantly to the local economy. We also investigated the impact of cash income from agriculture, forestry (other than wildmeat) and other employment opportunities on wildmeat consumption using Pearson’s Product correlation. This was used to examine the impact of cash income on wildmeat use by a household. We compared gross annual incomes and dependencies of surveyed households on various income sources

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across their respective villages using Kruskal–Wallis. All the data, wherever appropriate, were normalized and statistical procedures were applied following Sokal & Rohlf (1995). Results Socio–economic profile of respondents Out of 134 respondents, Angami and Apatani indigenous communities each represented 25%, 22% were from the Nishi community and the rest were Mizo. Most of our respondents were literate. Their age, education status and family size are presented in table 1. Wild meat survey Our respondents in the villages surveyed extracted at least 137 wild animal species, including 50 mammals and one reptile, during the previous year. Apatani household extracted on average of 282 kg wild meat annually (table 2), mainly from mammals (85%). Birds formed 5.98% of Apatani’s extraction. The most relevant group among birds was galliformes. Other animals (mainly fish) represented a significant part of the Apatani diet and constituted 8.4% of the extracted meat by weight. An Angami household extracted about 457 kg of wild meat annually, which was mainly (89%) mammals. Birds too formed a substantial component (9.4%). A Mizo household extracted a mean of approximately 278 kg wild meat annually, of which 89% came from mammals. Birds constituted 3.4% and the majority were galliformes. Other animals formed 7.8% of the Mizos’ total wild meat extracted. Amongst Nishis, mammals constituted 69% of the total wild meat extracted (average approximately 545 kg) annually. Other animals formed a substantial component (17.5%) and birds formed about 13.4%. Wild meat market survey We observed a total of 773 dead animals (233 mammals and 540 birds) in the markets of Kohima over 15 full days of observation and recorded 53 wild animal species (15 mammals and 38 birds). Similarly, we examined a total of 601 dead wild animals (418 mammals and 183 birds) in the markets of Hapoli, and recorded 19 wild animal species (10 mammals and 9 birds). A total of 118.62 kg of wild meat (80.39% from mammals and the rest from birds) was available at Hapoli and 154.33 kg of wild meat (73.92% from mammals and the rest from birds) at Kohima. All animal meat came from adjoining rural areas. Wild meat and socio—economic variables We investigated the relationship between wild meat extraction and socio–economic variables (table 3). A significant relationship emerged only amongst

Angami and Mizo communities. Angamis with a higher income from sources other than wild meat tended to harvest more wild meat. The extraction of wild meat amongst Mizos declined the higher the education level. Extraction of wild meat showed no statistically difference in relation to occupation (Kruskal–Wallis, n.s.) An analysis using Pearson correlation coefficient was performed to determine the effect of income on wild meat consumption. With the exception of Mizo community (fig. 1), significant positive correlations were observed between gross cash income and amount of wild meat consumed by Angami, Apatani and Nishi communities (fig. 2). Incomes and dependencies We estimated average annual gross incomes of the indigenous communities included in the study (table 4). Incomes were interpreted as accruals on the basis of cash values of the forest and agriculture– based goods obtained by a household in addition to incomes from other employment opportunities (e.g. self–employment i.e. business, and government employment i.e. state and federal government funded employment in various public departments). Bulk of average income (approximately 25%) to a Nishi household is derived from wild meat, which is conspicuously higher than their incomes from agriculture and other employment vocations (self and government employment). Similarly, Angami, Apatani and Mizo households derived average 14– 16% incomes from wild meat. Gross annual incomes of the study communities from various income sources (crops, agricultural work, wild plant products, forest management activities, wild meat, self–employment and government employment) and now they state more possibilities) showed significant differences (Angami: F 6 224 = 6.47, P < 0.001; Apatani: F6 224 = 11.6, P < 0.001; Mizo: F6 259 = 2.71, P < 0.01; Nishi: F6 203 = 3.92, P < 0.001, One way ANOVA). Similarly, these sources of income varied significantly among the four communities (Angami: F 6 224 = 16.4, P < 0.001; Apatani F 6 224 = 18.3, P < 0.001; Mizo F6 259 = 7.66, P < 0.001; Nishi F6 203 = 15.53, P < 0.001, One way ANOVA). However, incomes and dependencies of these communities on various sources across their respective villages did not show significant variations (Kruskal– Wallis test, n.s.). Discussion A large number of mammals and birds are hunted in Northeast India (Hilaluddin, 2005a, 2005b) and many of these are of concern to conservation (Birdlife International, 2000; IUCN, 2003). In the villages surveyed, the hunted animals included 20 species considered as threatened on the Red Data List (IUCN, 2003); four Endangered (Aceros nipalensis, Bubalus bubalus, Elephas maximus and

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Table 1. Socio–economic profile of the respondents' communities. Tabla 1. Perfil socioeconómico de los sujetos de las comunidades interrogados.

Age (years) Tribe

Mean

S.E.

Education (class) Mean

S.E.

Family size Mean

S.E.

% Literacy rate Literate Illiterate

Angami (N = 33)

84.6

15.4

Apatani (N = 33)

57.6

42.4

Nishi (N = 30)

53.3

46.7

Mizo (N = 38)

94.7

5.3

Table 2. Wild meat extraction patterns (in kg/household/annum, mean ± CI) of sampled indigenous communities in Northeast India. (Data at 95% confidence level.) Tabla 2. Patrones de extracción de carne de caza (en kg/familia/año, media ± CI) de las comunidades indígenas muestreadas del nordeste de la India. (Datos con un nivel de confianza del 95%.)

Mode of Income

Angami (N = 33)

Apatani (N = 33)

Nishi (N = 30)

Mizo (N = 38)

Wildmeat

651.7 ± 349.1 457.5 ± 211.8

282.1 ± 138.4 239.1 ± 92.0

545.9 ± 186.3 496.2 ± 151.5

277.7 ± 140.8 188.8 ± 71.7

Mammals

564.3 ± 291.8 408.2 ± 192.7

241.4 ± 105.5 208.5 ± 84.9

377.3 ± 132.0 346.2 ± 109.6

246.5 ± 132.8 172.0 ± 70.7

Herbivores

9.91 ± 201.98 322.64 ± 147.97

184.2 ± 88.92 164.56 ± 69.65

277.66 ± 117.6 255.76 ± 92.8

206.5 ± 110.11 151.68 ± 66.01

Carnivores

164.42 ± 95.68 85.53 ± 50.53

57.17 ± 23.76 43.93 ± 21.91

99.64 ± 35.9 90.49 ± 31.08

40.02 ± 24.82 20.30 ± 9.15

Birds

51.5 ± 31.1 43.2 ± 22.9

16.9 ± 9.1 16.0 ± 7.8

73.1 ± 48.1 54.5 ± 19.4

9.5 ± 6.0 6.2 ± 2.7

Galliformes

23.38 ± 18.9 15.62 ± 8.14

11.9 ± 8.74 11.19 ± 7.32

29.73 ± 13. 34 26.66 ± 10.43

4.83 ± 4.42 3.00 ± 1.44

Other birds

28.16 ± 14.72 27.60 ± 19.53

4.86 ± 2.43 4.81 ± 2.06

43.35 ± 43.03 27.8 ± 14.51

4.67 ± 2.26 3.17 ± 1.68

35.9± 42.7 6.1 ± 5.7

23.8 ± 23.6 14.6 ± 13.3

95.5 ± 74.7 95.5 ± 74.7

21.7 ± 12.3 10.6 ± 3.2

Other animals

Table 3. Pearson’s correlation coefficients between socio–economic factors and wild meat harvest of the sampled indigenous communities of Northeast India: A. Age; E. Education; I. Income; * Denotes level of significance (P < 0.05). Tabla 3. Coeficientes de correlación de Pearson entre los factores socioeconómicos y la extracción de carne de caza de las comunidades indígenas muestreadas del nordeste de la India: A. Edad; E. Educación; I. Ingresos; * Indica nivel de significación (P < 0,05).

Angami (N = 33) Product Wild meat

–0.29 0.04 0.44*

Apatani (N = 33) A

0.11 –0.27 0.26

Nishi (N = 30) A

–0.23 –0.07 –0.09

Mizo (N = 38) A

0.08 –0.40* –0.07

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3.5 Wildmeat consumption

3.0

r = 0.22 N = 38 P = n.s.

2.5 2.0 1.5 1.0 0.5 0.0 0

6 Wildmeat consumption

r = 0.51 N = 30 P < 0.01

4 3 2 1 0 0

2 US $ income

Fig 2. Relationship between wildmeat consumption (kg/year) and incomes: A. Mizo community; B. Nishi community. (All values normalized.) Fig. 2. Relaci贸n entre el consumo de carne de caza (kg/a帽o) y los ingresos: A. Comunidad Mizo; B. Comunidad Nishi. (Valores normalizados.)

Panthera tigris), eight Vulnerable (Capricornis sumatraensis, Macaca assamensis, Manis pantadactyla, Neofelis nebulosa, Panthera pardus, Presbytes pileatus, Selenarctos thibetanus and Tragopan blythii) and rest Lower Risk: near threatened (Columba punicea, Cuon alpinus, Felis bengalensis, F. viverrina, Hylopetes alboniger, Nemorhaedus goral, Nycticebus coucang and Prionodon pericolor). In India, under the Wild Life Protection Act 1972, it is illegal to kill any wild life (Anon, 2003). Our interactions with respondents revealed that almost half were aware of this law and the penalties for violation . We therefore feel that in some cases our respondents may have revealed lower figures of animals than those actually hunted and the conservation problem may be graver than reported here.

The loss of species to hunting warrants urgent attention in Northeast India because forests here are already much reduced in area and are increasingly fragmented as a result of shifting cultivation (FSI, 2003). This implies that populations of species endemic to this habitat type are not only at risk of loss of habitat and populations becoming isolated from each other, but also from easier access to hunters. The loss of relatively small numbers of individuals, especially species that are included in the Red List, may have a disproportionate impact on small and isolated populations. Another important issue that warrants attention is the wild meat consumption patterns of the surveyed indigenous communities. Our findings (fig. 2) are contrary to the Kuznets model (1955) of income and consumption of goods. Our models indicate that households in Northeast India continue to use wild

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5 Wildmeat consumption

r = 0.73 N = 33 P < 0.001

3 2 1 0 –1

3.5 Wildmeat consumption

3.0

r = 0.97 N = 33 P < 0.001

2.5 2.0 1.5 1.0 0.5 0.0 0.0

0.5

1.0

1.5

2.0 2.5 3.0 US $ income

3.5

4.0

Fig. 2. Relationship between wildmeat consumption (kg/year) and incomes: C. Angami community; D. Apatani community. (All values normalized.) Fig. 2. Relación entre el consumo de carne de caza (kg/año) y los ingresos: C. Comunidad Angami; B. Comunidad Apatani. (Valores normalizados.)

meat and may even increase their consumption of wild meat as their income increases. The models show that a rise in income may have the unexpected and undesirable side effect of promoting consumption–driven increases in hunting pressure. Given the open access nature of wild meat and its demand in the region, a rise in income levels could easily enhance the demand for wild meat and consequently induce over–harvesting of the species both in the short and the long term. The long term is not relevant for the species that are most threatened by hunting, for which extinction within a decade is a real possibility (Nelleman & Newton, 2002). Therefore, strong intervention is required where there is a need to reduce hunting levels. It is essential to understand not only the impact of hunting on wild populations but also the reason why certain species are hunted (Kaul et al., 2004).

Firstly, hunting has a religious and cultural significance to many communities in Northeast India (Hilaluddin, 2005a). For example, the religious rituals of the Apatani community include generous offerings of smoked Funambulus palmarus, F. pennanti, Hylopetes alboniger and Dremomys lokriah. The Apatani community also sacrifices Macaca assamensis to propitiate their deity during their annual spring festival, "Morum". The festival’s feasting includes a voluminous amount of Muntiacus muntjak and Sus scrofa meat. Barbets, specifically Megaliama virens, are often served to entertain special family guests. Nishi priests decorate their headgear with Selenarctos thibetanus skins and a pair of hornbill tail feathers. Furthermroe, Nishis prize the skin of Presbytes pileatus for making sheaths for their traditional daggers, "Davs". Other community members adorn

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Table 4. Gross annual incomes (in US $) of sampled indigenous communities of Northeast India from wild meat and other income avenues. (Data at 95% confidence level.) Tabla 4. Ingresos anuales brutos (en $ americanos) de las comunidades indĂgenas muestreadas del nordeste de la India, provinentes de la carne de caza y otras fuentes. (Datos con un nivel de confianza del 95%.)

Angami (N = 33) Mode of Income

Mean

Apatani (N = 33)

Mean

Nishi (N = 30) Mean

Mizo (N = 38) Mean

Agriculture

2,164.9 656.6

11,505.7 396.9

853.5 333.9

1,434.8 1,052.7

Agriculture crop

2,105.9 656.8

11,459.7 401.3

754.5 316.3

1,361.4 1,052.6

Agriculture labor

73.1

73.4

95.9

Forestry

2,365.6 902.6

11,343.0 300.8

2,149.6 775.4

1,216.8

559.2

Plant

1,088.6 330.7

711.1 172.3

1,241.4 739.3

545.1

463.9

Timber

736.3 319.9

172.1 113.1

194.4 168.2

66.8

129.8

Gross NTFPs

352.2 136.1

Bamboo Fuelwood Other NTFPs Forest management

59.0

50.5

445.9

29.7

539.0

91.8

98.4

176.1

57.8

231.9 114.2

324.7 38.1

68.4 51.8

30.3

98.9

1,047.9 688.2

478.4

452.8

9.07

293.4

257.7

66.7

487.4 161.2

183.9

65.1

20.9

539.1 538.4

1.1

0.9

20.5

136.5 155.1

91.2 136.4

109.9

109.0

1,277.1 847.5

495.4 219.8

817.0 275.9

561.7

294.9

Mammals

875.3 539.9

368.7 169.3

517.2 173.5

420.0

245.9

Birds

324.2 249.1

Animal

Other animals

77.7

88.6

Other employment opportunity 1,636.7 965.2 Government employment Self employment Gross Income

927.2 388.4

86.9

50.1

156.6

64.6

46.8

30.9

39.8

45.6

143.2 124.3

94.9

65.5

1,219.5 472.4

962.4 615.8

833.8 367.8

538.8 313.2

2,171.5 1,193.6 835.9

357.1

709.5 779.5

385.6 366.5

423.6 465.4

1,335.5 1,066.5

6,163.91,679.8

4,068.2 679.9

3,965.51,230.3

4,823.1 2,016.1

their caps with hornbill beaks, specifically Aceros nipalensis, and a pair of Dicrurus paradiseus and/or D. remifer tail feathers. Several species are also popular among locals for their role in traditional medicines in local beliefs (Hilaluddin, 2005b). Amongst Mizos, flesh of Macaca assamensis is associated with relieving delivery pains and is also believed to aid the development of the infant while inside the motherâ&#x20AC;&#x2122;s womb; bats are supposed to cure asthama; the gall bladder of Selenarctos thibetanus heals jaundice; and the liver of Hylobates hoolock kills malarial parasites. Angamis consume Upapa epops to alleviate male impotency. Secondly, it appears that the primary objective is to secure an animal for consumption or sale. The opportunity cost for the extraction of a wild animal which is relatively more common than others should be less than that of less common ones. Thus, the most abundant wild animals are expected to be harvested more intensively than the less abundant ones. However, the opportunity cost also depends

on body size of the target quarry, and therefore the quantity of meat rather than the quality generally dictates direct preferences. Unfortunately, abundance estimates for most animal species are lacking for Northeast India in general and our study area in particular, making it difficult to determine whether offtake is adversely affecting wild populations (Hilaluddin et al., in press a). This requires investigation. Thirdly, wild meat in our surveyed areas is also harvested for trade. It appears that families living in comparatively remote areas have poor access to markets and where substitutes are not available, people mainly rely on wild meat for protein. However, those who have migrated to cities and towns for a better living have not lost the "taste" for wild meat. In such areas, wild meat constitutes a "superior good" and people pay 1 to 5 times the domesticated animal meat. The markets in the towns seem to be fed directly from the remote areas where people may kill wild animals mainly to cater for the

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Table 5. % Dependencies of sampled indigenous communities of Northeast India on wild meat and other income sources. (Data of 95% confidence level.) Tabla 5. Dependencias de la carne de caza y de otras fuentes de ingresos de las comunidades indĂgenas muestreadas del nordeste de la India. (Datos con un nivel de confianza del 95%.)

Angami (N = 33)

Apatani (N = 33)

Nishi (N = 30)

Mizo (N = 38)

Mode of Income

Mean

Agriculture

39.4

9.3

37.0

5.8

22.98

5.8

29.8

9.0

Agriculture crop

38.0

9.2

35.2

5.4

20.04

5.6

28.0

8.8

Agriculture labor

1.3

1.7

1.2

2.94

2.0

1.8

2.2

Forestry

37.7

7.2

37.4

7.1

59.07

7.6

32.4

8.1

Plant

23.5

7.9

19.2

3.5

31.66

5.8

12.4

3.9

Timber

16.9

7.6

3.8

1.9

4.91

2.8

1.3

2.4

Gross NTFPs

6.6

2.5

15.4

2.9

26.74

5.6

11.2

3.5

Bamboo

0.7

0.6

4.7

1.7

0.63

0.2

3.3

2.4

Fuelwood

4.8

2.2

9.5

2.0

16.34

4.4

7.8

2.7

Other NTFPs

1.0

0.6

1.2

0.7

9.78

3.7

0.1

3.5

3.6

2.21

3.7

3.5

4.5

Animal

14.1

5.8

14.7

5.8

25.2

7.4

16.4

6.4

Mammals

10.3

4.4

10.7

4.5

15.79

4.6

11.9

5.5

Birds

3.1

1.4

2.9

1.6

5.81

2.7

1.1

0.5

Other animals

0.7

0.6

1.1

3.6

2.8

3.4

2.4

Other employment opportunity

23.1

8.5

25.6

8.7

17.95

7.2

37.8

9.6

Government employment

16.4

7.9

18.4

7.7

12.43

5.3

21.0

9.3

6.7

4.9

7.2

5.8

5.52

6.1

16.8

6.6

Forest management

Self employment

demand of urban areas and thereby ensuring a supply of wild meat even if city dwellers do not have the time or opportunity to hunt regularly. Therefore, wild animal hunting in our study area clearly demonstrates a direct link between level of harvest and economic growth of those involved in wildlife trade. Thus, there is likely to be an increase in the wild meat extraction intensities of commercial hunters with an increase in urban populations. This increased appetite for goods should further stress the need to exploit animal resources in remote forest areas. One such example is found in the link between demand for tropical hard timber in the international market and overâ&#x20AC;&#x201C;exploitation of forests in Southeast Asia and the Amazon basin (Kolk, 1996; Dauvergne, 1997; Barker, 1998). The majority of our respondents belonged to an economic stratum well above the poverty line (annual income above Indian Rupees 11,000/household or 244.44 US $). Taking the other income sources into account (fodder and shifting cultivation), these figures would further add up to a significant annual income.

Amongst Nishis, a large proportion of rural income is derived from the forestry. Amongst Mizos, the bulk of income is derived from other employment opportunities (self and government employment) and the agriculture sectors followed by the forestry. Such patterns are contrary to the general economy and employment of rural India which is largely agriculture based (Sethi & Hilaluddin, 2001). However, the rural economy of Angamis seems to conform the general agriculture based economic pattern of rural India. The rural economy of the Apatanis shows equal dependence on agriculture and forestry sectors. Amongst the Nishis, wild meat occupies an important place in village economy. Such an economic pattern is similar to the rural economy of Ghana where wild meat makes a significant contribution to both the household food supply and as cash income (Dei, 1989). Our respondents seemed to be highly dependent upon wild meat for both their kind and cash values. Our estimated annual incomes and dependencies of Angami, Apatani, Mizo and Nishi communities on the forestry are not directly comparable with

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income levels and dependencies on the forestry sector reported earlier in India or elsewhere in the world. This is because in their estimates, previous workers (e.g. Malhotra et al., 1991; Bahuguna, 1993, 2000; Hedge et al., 1996; Sethi & Hilaluddin, 2001) have overlooked incomes derived from wild meat contributing significantly to local economies. We thus feel previous estimates are incomplete and that previous appraisals should be revised Our analysis on occupation status vis–à–vis wildmeat extraction suggests that all sections of the society: be they custodian of the law or farmer or businessman, remove animal biomass equally. It also seems that an increase in education among the Mizo decreased the amount of wild meat extraction. With a higher level of education, people have access to better jobs, and this in turn presumably leaves them with little time to hunt. However, in certain communities such as the Angami, increased cash incomes from vocations other than wild meat resulted in higher extraction of wild meat. It is likely that an improvement in financial status of a household also increases the desire to consume more. Therefore, policies linking poverty alleviation programs with the conservation of natural resources should be drafted with utmost care. Policies linking extraction of wild meat to alleviate poverty with conservation of natural resources require major review. Acknowledgements Our sincere thanks are due to Mr. James Goodhart, who provided financial assistance for fieldwork. Drs. Claudia Ruttee, John Carroll, Michael Conroy, Peter Garson, Philip McGowan, Francesc Uribe, Francisca Castro, Ghazala Shabuddin and Indrani Chowdhary commented on draft manuscript and made several useful suggestions. We are grateful for their efforts and concern! We also thank our respondents for their tremendous hospitality during fieldwork and also for sharing their views openly during interviews. References Abernethy, K., Baker, M., Banett, L. E., Boclmer, R., Brashanes, J., Cowlishaw, G., Elkan, P., Eves, H., Fa, J. E., Milner–Gulland, E. J., Peres, C., Roberts, C, Robinson, J. G., Rowcliffe, M. & Wilkie, D., 2003. Wild meat: the bigger picture. Trends in Ecology and Evolution, 18: 351–357. Ali, S. & Ripley, S. D., 1987. A Handbook of Birds of India and Pakistan: compact edition. Oxford University Press. New Delhi. Anon, 2003. Indian Wildlife Protection Act 1972 (as amended 2002). Natraj Publishers. DehraDun, Uttranchal, India. Bahuguna, V. K., 1993. Forestry in Ecodevelopment. Indian Institute of Forest Management. Bhopal. – 2000. Forests in the economy of rural poor: an estimation of the dependency level. Ambio, 29:

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126–129. Barker, C., 1998. Forest resource scarcity and social conflicts in Indonesia. Environment, 40: 4–37. Bennett, E. L., 2002. Is there a link between wild meat and food security? Conservation Biology, 16: 590–592. Birdlife International, 2000. Threatened Birds of the World. Clark, Birdlife International, Cambridge, U.K. Dauvergene, P., 1997. Shadows in the Forest: Japan and the Politics of timber in Southeast Asia. MIT Press. Boston, U.S.A. Dei, G. J. S., 1989. Hunting and gathering in a Ghanaian rain forest community. Ecology, Food and Nutrition, 22: 225–243. Fa, J. E., Peres, C. A. & Meeuwig, J., 2002. Bushmeat exploitation in tropical forests: an International comparison. Conservation Biology, 16: 232–237. FSI, 2003. The State of Forest 2003. Forest Survey of India. DehraDun, Uttranchal, India. Hedge, R., Suryaprakash, S., Achoth, L. & Bawa K. S., 1996. Extraction of non–timber forest products in the forests of Biligiri Rangan Hills, India: contribution to rural income. Economic Botany, 50: 243–251. Hilaluddin, 2005a. Illicit staple. Down to Earth, 13: 50–52. – 2005b. Wild meat boom and wildlife bust in Northeast. Kashmir Times dated 22 January. Hilaluddin, Kaul, R. & Ghose, D. (In press a). Galliformes extraction and use by indigenous people of Northeast India. In: proceedings of 3rd International Galliformes Symposium (R. A. Fuller & S. Browne Eds.). World Pheasant Association–International, U.K. Hilaluddin, Kaul, R., Pradhan, S. & Taylor, J. (In press b). Conservation significance of wildmeat exploitation and use in the north Bengal Himalaya, India. Conservation & Society. IUCN, 2003. 2003 Red List of Threatened Species. IUCN, Gland, Switzerland. Kaul, R., Hilaluddin, Jandrotia, J. S. & McGowan, P. J. K., 2004. Hunting of large mammals and pheasants in the western Indian Himalaya. Oryx, 9: 426–431. Kolk, A., 1996. Forests in International Environmental Politics: International Organizations, NGOs and the Brazilian Amazon. International Utrecht. Kuznets, S., 1955. Economic growth and income inequity. American Economic Review, 445: 1–28. Malhotra, K. C., Deb, D., Datta, M., Vasulu, T. S., Yadav, G. & Adhikari, M., 1991. Role of non– timber forest produces in village economy: a household survey in Jamboi Range, Midnapore, West Bengal. Institute of Socio–biological Research and Development, Calcutta, India, Unpublished Report. 192 pp. McGowan, P. J. K., Duckworth, W., Xianji, W., Balen, S. W., Xiaojun, Y., Kahn, M., Yatim, H., Thanga, L., Setiawan, I. & Kaul, R., 1998. A review of the status of Green peafowl Pavo muticus and recommendations for future action.

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Bird Conservation International, 8: 331–341. Nelleman, C. & Newton, A., 2002. The Great Apes: The Road Ahead. UNEP, Gland, Switzerland. Peres, C. A., 2000. Effects of subsistence hunting on vertebrate community structure in Amazonian forests. Conservation Biology, 14: 240–253. Prater, S. H., 1971. The Book of Indian Animals. Oxford Univ. Press, New Delhi. Robinson, J. G. & Redford, K. H., 1991. New Tropical Wildlife Use and Conservation. Chicago University Press, Chicago. Sankaran, V., Hilaluddin, R. K. & Bandhopadhyay, P., 2000. The process of participatory mapping for biodiversity conservation. In: Setting Biodiversity Priorities for India: 34–56 (S. Singh, A. R. K. Shastri, R. Mehta & V. Uppal, Eds.). WWF–India, New Delhi.

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Sethi, P. & Hilaluddin, 2001. Structuring financial empowerment for localized development with Joint Forest Management (JFM): examples from Madhya Pradesh, India. Sustainable Development, 9: 87–102. Sokal, R. R. & Rohlf, F. J., 1995. Biometry: Principals and Practice of Statistics in Biological Research. W.H. Freeman and Company. New York, U.S.A. Sutherland, W. J., 1996. Why census? In Ecological census techniques: a handbook: 1–9 (W. J. Sutherland, Eds.). Cambridge Univ. Press. Cambridge, U.K. Wilkie, D. S. & Carpenter, J. F., 1999. Bushmeat hunting in the Congo Basin: an assessment of impacts and options for mitigations. Biodiversity Conservation, 8, 929–955.

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

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Land snail diversity in a threatened limestone district near Istanbul, Turkey A. Örstan, T. A. Pearce & F. Welter–Schultes

Örstan, A., Pearce, T. A. & Welter–Schultes, F., 2005. Land snail diversity in a threatened limestone district near Istanbul, Turkey. Animal Biodiversity and Conservation, 28.2: 181–188. Abstract Land snail diversity in a threatened limestone district near Istanbul,Turkey.— The limestone meadows located to the north–northwest of Istanbul, Turkey, are in danger of being overrun by the rapidly expanding city. Past surveys showed that these habitats harbor rare plant species, including endemics to Turkey. To further evaluate the conservation value of these habitats, especially in terms of the often neglected invertebrates, one limestone area to the north of Küçükçekmece Lake and surrounding Sazlidere Dam was surveyed for land snails. Our findings strengthen the case for the protection of these unique habitats. Twenty–four species of land snails were identified in the survey area. Of these, 21 are native to Turkey, including three whose type location is Istanbul. In addition, two species that are at or near the limits of their ranges are considered to represent peripheral populations that may be especially worth conserving. Although the area surrounding Sazlidere Dam is under protection, the other limestone habitats are severely threatened by ongoing development. Key words: Biodiversity, Conservation, Istanbul, Pulmonata, Prosobranchia. Resumen Diversidad de los caracoles terrestres en una zona caliza amenazada cercana a Estambul, Turquía.— Las praderas calcáreas situadas al NNO de Estambul están en peligro de ser rápidamente invadidas por la ciudad en expansión. Estudios anteriores demostraron que estos hábitats albergan especies vegetales raras, incluyendo algunos endemismos turcos. Con objeto de seguir evaluando el valor conservativo de dichos hábitats, en especial en cuanto a los invertebrados, a menudo ignorados, se han estudiado los caracoles terrestres de una zona calcárea al norte del lago Küçükçekmece y alrededor de la presa Sazlidere. Nuestros descubrimientos enfatizan la necesidad de una política de protección de estos hábitats únicos. En el área estudiada se identificaron 24 especies de caracoles terrestres. De ellas, 21 son nativas de Turquía, incluyendo tres cuya localización tipo es Estambul. Además, se considera que dos especies que se hallan en o cerca de los límites de su zona de distribución representan poblaciones periféricas especialmente merecedoras de conservación. A pesar de que la zona que rodea a la presa Sazlidere está protegida, el resto de los hábitats calcáreos está muy amenazado por el creciente desarrollo. Palabras clave: Biodiversidad, Conservación, Estambul, Pulmonata, Prosobranchia. (Received: 4 V 05; Conditional acceptance: 21VI 05; Final acceptance: 8 VII 05) Aydin Örstan, Timothy A. Pearce, Section of Mollusks, Carnegie Museum of Natural History, 4400 Forbes Ave., Pittsburgh, PA, 15213, U.S.A.– Francisco Welter–Schultes, Zoologisches Inst., Berliner Str. 28, D–37073, Goettingen, Germany.

ISSN: 1578–665X

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Introduction Unprotected wildlife habitats located near expanding residential or industrial centers are subject to rapid and permanent destruction. Especially in developing countries, unique habitats near growing cities may be destroyed before they are properly surveyed and measures implemented for their protection. One case in point is the city of Istanbul, whose population has grown from about 800,000 in 1927 to 10 million in 2000 (Istanbul Metropolitan Municipality, 2005). Prior to the 20th century, the then much smaller city of Istanbul was surrounded by villages separated from each other by more or less degraded, but nevertheless uninhabited and undeveloped land that included agricultural fields and orchards (Anonymous, 1844). Such areas must have provided habitats for at least some of the native wildlife. However, since the beginning of the 20th century, the rapidly expanding Istanbul has absorbed most of the villages, turning them into districts within the city and, in the process, has all but decimated the wildlife habitats. Özhatay et al. (2003) recently brought attention to several threatened unique habitats (designated as Important Plant Areas) surrounding Istanbul that are rich in rare and endemic plant species. One of these Important Plant Areas is the limestone meadows to the north–northwest of Istanbul. Özhatay et al. (2003) singled out three remaining fragments of these meadows in the region extending west from the vicinity of the town of GaziosmanpaÕa north of Istanbul to the north of the Küçükçekmece Lake and identified 19 rare plant taxa, including seven that are endemic to Turkey, growing on these meadows. The westernmost of these fragments is located along Sazli Creek (Sazli Dere) that empties into Küçükçekmece Lake (fig. 1). To supply drinking water for Istanbul, Sazlidere Dam was recently built over this creek, partly flooding the creek’s broad valley. Istanbul and its environs are the type locations of about 10 species of land snails that were described mostly in the 19th century (Schütt, 2001). Unfortunately, because of the ongoing loss of land to development, it has now become difficult to find habitats in and around the city that still maintain the original native land snail fauna. Our attention was, therefore, attracted to the Sazli Creek area not only because of the presence of limestone, which generally supports a high diversity of land snails, but also because of the threatened status of this habitat type in the Istanbul area. Furthermore, the area had not been properly surveyed for land snails. Therefore, to further evaluate the conservation value of the limestone meadows, especially in terms of the diversity of their fauna of native land snails, we conducted a land snail survey of the Sazlidere Dam area. Materials and methods The study area, located west of the city of Istanbul, extended from the north of Küçükçekmece Lake up the broad valley of Sazli Creek to the low limestone

hills surrounding Sazlidere Dam (fig. 1). The area was not forested, but consisted of meadows and grassy hills with limestone outcrops. The survey was conducted on two days (26 VI and 8 VII 2004). Twelve stations, scattered along an approximately 12.5–km long transect, were designated in the field (fig. 1), but during the analysis of the results three of the stations (C4, C5 and C6) that were located within less than 100 m of each other were treated as one. The UTM coordinates (for zone 35) and elevation of each station were measured with a GPS receiver with an accuracy of about 10 m. The following list gives the description and coordinates of each station (fig. 1): C1. Limestone cliff above road to Ôamlar Village; UTM E646269 m, N4548620 m; elevation 20 m. C2. Northeast bank of Sazli Creek; UTM E645861 m, N4548704 m; elevation 0 m; C3. Rocky hill northeast of Sazli Creek; UTM E645575 m, N4550187 m; elevation 70 m; C4–C6. Rocky hill south of road to Sazlidere Dam; UTM E645578 m, N4550370 m; elevation 55 m; C7. Limestone rocks near unpaved road, north of Ôamlar Village; UTM E646246 m, N4554293 m; elevation 100 m; C8. Limestone rocks on hillside above Sazlidere Dam lake; UTM E645082 m, N4553385 m; elevation 30 m; C9. Limestone rocks along grassy field; UTM E643946 m, N4553427 m; elevation 25 m; C10. Limestone rocks on hillside above unpaved road, north of old limestone quarry. UTM E641977 m, N4555331 m; elevation 25 m; C11. Grassy field along road to Hadimköy, west of roadway across Sazlidere Dam lake; UTM E638674 m, N4557811 m; elevation 25 m; C12. Limestone rocks on hillside facing a residential district, north of Küçükçekmece Lake; UTM E646015 m, N4547613 m; elevation 65 m. Surface collections were done at each station by two or three people. In addition, soil samples were taken from six stations, sieved and sorted for small shells. The identifications of Oxyloma elegans, Monacha ocellata, Monacha solidior, Xerolenta obvia, Xeropicta krynickii and Cernuella virgata were confirmed by dissection. Sixty–one lots (537 specimens), including at least one lot of every land snail species found in the study area (excluding Eobania vermiculata), have been deposited with the Carnegie Museum of Natural History, Pittsburgh, PA, U.S.A. (CM 70300–70357, 70762–70764). Additional lots are in the collection of the first author. Reference samples of Albinaria caerulea were obtained from the Field Museum of Natural History (FM), Chicago, U.S.A. Results We found 24 species of land snails in the survey area, representing 12 families (table 1). In addition, slug shells were collected at stations C1 and C9, but these could be identified only to the family level (Reuse, 1983). A live, dormant Eobania vermiculata was seen at station C1 but not taken. Fourteen species (58% of total) were found only as empty shells (table 1) and

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Animal Biodiversity and Conservation 28.2 (2005)

Black Sea Survey area

Istanbul Sazlidere Dam

Sea of Marmara

Sazli Creek

Istanbul Küçükçekmece Lake

N 5 km

Sea of Marmara

Fig. 1. The survey site showing the locations of the collection stations. The inset shows the location of the survey site in relation to the metropolitan Istanbul (shaded area). The background satellite photograph (file name: ISS008–E–21753.jpg, taken on 16 IV 2004) was downloaded from http:// eol.jsc.nasa.gov (image courtesy of the Image Analysis Laboratory, NASA Johnson Space Center). Fig. 1. Lugar del estudio mostrando la situación de las estaciones de recolección. El recuadro muestra la localización del lugar de estudio en relación con el área metropolitana de Estambul (área sombreada). La fotografía de satélite del fondo (nombre del archivo: ISS008–E–21753.jpg, tomada el 16 IV 2004) se bajó de http://eol.jsc.nasa.gov (por cortesía del Image Analysis Laboratory, NASA Johnson Space Center).

two species (8% of total) only as live specimens. We consider 21 of the land snail species to be native to the survey area. Cochlicella acuta, Cernuella virgata and Eobania vermiculata are non–native species that have been introduced to Turkey. Additional notes on some of the species Pomatias elegans The earliest record of P. elegans from around Istanbul dates to the 19th century (Sturany, 1894). This species strictly requires calcareous substrates (Boycott, 1934) and we found it quite abundantly at some of our stations. Oxyloma elegans Two individuals were found crawling on plants growing in Sazli Creek at station C2. Oxyloma elegans is found throughout Europe (Hecker, 1965; Kerney & Cameron, 1979) and has been recorded in Turkey before (Schütt, 2001). However, our record of this species appears to be the first for the Istanbul area. Oxyloma elegans cannot always be reliably separated from O. sarsii (Esmark, 1886) by shell char-

acteristics (Kerney & Cameron, 1979). Therefore, the previous records of O. elegans and O. sarsii from Turkey that were not confirmed by dissection may not be reliable. Pupilla cf. sterrii The Pupilla shells from stations C1 and C12 were finely striated and their apertures had three teeth: a parietal, a palatal and a deep columellar that was visible when the aperture was turned slightly sideways (fig. 2). The sample of nine adult shells from C1 had a mean length of 2.83 mm and a mean diameter of 1.57 mm. To identify these specimens, we considered two species: P. triplicata (Studer, 1820) and P. sterrii. Although P. sterrii is stated to have usually two teeth, a parietal and a palatal (Kerney & Cameron, 1979; Falkner, 1990), we identified our specimens tentatively as P. sterrii rather than P. triplicata, because the microsculpture of our shells agreed better with that of P. sterrii and the dimensions of our specimens were closer to those of P. sterrii and slightly above the ranges, especially for diameters, of those of P. triplicata (Germain, 1930; Kerney & Cameron, 1979).

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Table 1. The land snail species collected in the survey area and the stations where each was found: * Species that were found only as empty shells. (For information on stations see the text, Material and methods, and fig. 1.) Tabla 1. Especies de caracoles terrestres recogidas en la zona de estudio y estaciones donde se hallaron. Los asteriscos indican las especies de las que sólo se encontraron conchas vacías. (Para información sobre las estaciones ver el texto, Material and methods, y la fig. 1.)

Snail species

Stations

Pomatias elegans* (Müller, 1774)

C3, C4–C6, C7, C8, C9, C12

Oxyloma elegans (Risso, 1826)

Truncatellina cylindrica* (Férussac, 1807)

C4–C6

Granopupa granum* (Draparnaud, 1801)

C1, C4–C6, C8, C12

Pupilla cf. sterrii* (Voith, 1838)

C1, C12

Pleurodiscus balmei* (Potiez and Michaud, 1838)

C1, C9

Chondrus tournefortianus* (Férussac, 1821)

C3, C8

Multidentula ovularis* (Olivier, 1801)

C1, C3, C4–C6, C8

Mastus carneolus (Mousson, 1863)

C1, C3, C4–C6, C7, C8, C9, C10, C11, C12

Oxychilus hydatinus* (Rossmässler, 1838)

C1, C3, C4–C6

Cecilioides acicula* (Müller, 1774)

C4–C6, C10

Albinaria caerulea (Deshayes, 1835)

C1, C3, C4–C6, C7, C8, C12

Bulgarica denticulata* (Olivier, 1801)

C1, C3, C4–C6, C8, C9, C12

Cochlicella acuta* (Müller, 1774)

C1, C7, C10

Trochoidea pyramidata (Draparnaud, 1805)

C1, C3, C4–C6, C7, C8, C9, C10, C12

Monacha claustralis* (Mousson, 1859)

C10, C11

Monacha ocellata (Roth, 1839)

C1, C8, C9, C11

Monacha solidior (Mousson, 1863)

C3, C7, C8, C9

Xerolenta obvia (Menke, 1828)

C1, C3, C4–C6, C8, C9, C10, C12

Xeropicta krynickii (Krynicki, 1833)

C1, C2, C3, C4–C6, C7, C8, C9, C10, C11, C12

Cernuella virgata (Da Costa, 1778)

C1, C2, C3, C7, C8, C9, C10, C11, C12

Eobania vermiculata (Müller, 1774)

Helix lucorum* Linnaeus, 1758

C1, C8, C9

Helix pomacella* Mousson, 1854

C4–C6, C11

Limacidae

C1, C9

Multidentula ovularis This species was not previously recorded from the province of Istanbul (Forcart, 1940; Schütt, 2001). We found it in abundance at our station C3 and less abundantly at three other stations. Albinaria caerulea This strictly calciphilic species was abundant at several of our stations. Several subspecies of A. caerulea are distributed along the coastal regions of southwestern Turkey (Örstan, 2001), on the Greek islands near mainland Turkey (Nordsieck, 1977; Zilch, 1977) and in Attiki, Greece (Giokas & Mylonas, 2002). In addition, Schütt (2001) gave a record of

A. caerulea from Çatalca, about 18 km west of our survey area. We compared our specimens with samples of A. caerulea caerulea from the island of Chios (FM 206645), A. caerulea milleri (Pfeiffer, 1850) from the island of Delos (FM 206815), and A. caerulea maculata (Rossmassler, 1836) and A. caerulea calcarea (Boettger, 1878) from the vicinity of Ephesus, Turkey (Örstan, private collection). Using these comparisons, we determined that our specimens were conchologically closest to the nominal subspecies. We note that our specimens were also identical to the Field Museum lot of A. caerulea caerulea (FM 161499) with the collection location given broadly as "Thracien, Istanbul".

Animal Biodiversity and Conservation 28.2 (2005)

Fig. 2. A specimen of Pupilla cf. sterrii (2.7 x 1.5 mm) from station C1. Arrows point at the teeth. Fig. 2. Ejemplar de Pupilla cf. sterrii (2,7 x 1,5 mm) de la estación C1. Las flechas señalan los dientes.

Discussion We are unaware of any previously published survey of the land snails of the Sazlidere Dam area. However, we found published records for four species of land snails from the vicinity: C. acuta from Küçükçekmece (Sturany, 1902), M. carneolus from YeÕilköy (previously San Stefano) and Florya, districts southeast of Küçükçekmece Lake (Sturany, 1902; Gittenberger, 1967) and O. hydatinus and A. caerulea from Çatalca west of our survey area (Riedel, 1995; Schütt, 2001). We also found these four species in our survey (table 1). Sturany’s 1902 record of C. acuta from Küçükçekmece indicates that this introduced species has been in the area for more than 100 years. In the survey area we saw A. caerulea aestivating attached to limestone rocks. There is evidence that Albinaria species that aestivate on rocks have occasionally been transported to areas outside of their native ranges by humans on rocks intended for buildings or other purposes (Welter–Schultes, 1998). Therefore, we considered the possibility that A. caerulea was introduced to our survey area on rocks that were brought from elsewhere, perhaps southwestern Turkey. However, because of the relatively slow dispersal rate (~2 m/year) of Albinaria species (Schilthuizen & Lombaerts, 1994), unintentional introductions by humans usually result in localized distributions of the introduced species

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(Welter–Schultes, 1998). In comparison, the distance between the two farthest stations where we found A. caerulea, about 6.7 km, indicates that the distribution of the species in our survey area was not localized. We also note that Schütt’s (2001) record of A. caerulea from Çatalca, west of our survey area, suggests that the range of this species extends over an even wider area. Moreover, there is no evidence that large quantities of calcareous rocks were transported into the Sazli Creek area in the past (such materials are rarely used in modern buildings); the limestone quarry near station C10 that was in operation until recently indicates that limestone was actually exported from the area. Therefore, these arguments lead us to conclude that A. caerulea is native to the survey area. In addition to the previously published records listed above, Chalcolithic fossils of Helix pomatia Linnaeus, 1758 (from a layer ~6880 radiocarbon years B.P.) were reported from the Yarimburgaz Cave within our survey area (Meriç et al., 1991). However, because H. pomatia is not an extant species in Turkey (Schütt, 2001; Yildirim et al., 2004), we suggest that the specimens Meriç et al. (1991) reported as H. pomatia are probably the conchologically similar H. lucorum, which we found at stations C1, C8 and C9. Nevertheless, we note that the present day range of H. pomatia extends from northern Europe through the Balkan Peninsula down to Macedonia (Falkner, 1990) and that during the Chalcolithic period the species may have lived as far south as our survey area or may have been taken there by humans. We found only empty shells of 14 species (58% of total) and only live specimens of two species (8% of total). These results, obtained during the dry season in the Istanbul area, are comparable to the results Rundell & Cowie (2003) obtained in Hawaiian dry forests, where 40 to 47% of species were collected dead only and 0 to 7% live only. As Rundell & Cowie (2003) pointed out, if a survey produces only empty shells of a species, this result may imply that the species is either very rare or extinct at that location. However, since most snail species hide deep in the soil or under rocks during the dry season, we believe that we would have found live specimens of most, if not all, of the recorded species if we had collected during a rainy period, or if we had searched more intensely. We consider our results as constituting a baseline and we believe that only by conducting follow–up surveys of the area in the future will it be possible to accurately monitor any changes in the extant land snail fauna. Özhatay et al. (2003) based their arguments for the conservation value of the last remaining limestone meadows in the Istanbul area on their botanical richness. The results of our survey add four additional justifications for the protection of these habitats. First, the majority of the land snail species found in the survey area (21 out of 24) are native to Turkey. We believe that the protection of the few remaining undeveloped areas in and around Istan-

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Örstan et al.

Fig. 3. Residential development that is gradually overrunning the limestone hills. The photograph was taken on 26 VI 2004 from station C6 (the rocks in the foreground) looking west across the broad valley of Sazli Creek. Fig. 3. Desarrollo urbano que gradualmente va invadiendo las colinas calizas. La fotografia se tomó el 26 VI 2004 desde la estación C6 (rocas del primer plano) mirando hacia el oeste a través del ancho valle del riachuelo Sazli.

bul is urgently necessary to provide habitats for the snail species that are native to the region. Second, Istanbul is the type location of three of the species that we found: M. carneolus, M. ocellata and H. pomacella. For taxonomical reasons, we think that it is very important to protect the type locations of animal and plant taxa. If in the future specimens are needed for critical taxonomical comparisons, for example for genetic or anatomical analyses, the most suitable place to get specimens that are likely to be the same species as the one that was described from the type location would be the type location itself. Although the Sazlidere Dam area was not specified as the type location of any of these species, we believe that the area is close enough to Istanbul to be considered within the type location. Therefore, considering that the original type locations within Istanbul are likely to have been destroyed by now, the protection of the nearby areas with the same species would ensure the survival of these taxonomically important populations. Third, some of the species we found, for example, P. elegans and A. caerulea, are strict calciphiles and would not survive on noncalcerous substrates. The strict dependence of these species on calcareous rocks and soil underlines the need to protect their limestone habitats. Fourth, in our survey area P. elegans and A. caerulea may be at or near the limits of their distribution ranges. The range of P. elegans extends from England and an isolated spot in western

Ireland to northwestern Turkey (Kerney & Cameron, 1979; Örstan, 2005). The Istanbul area may be at or close to the southeastern limit of the range of this species (Örstan, 2005). As for A. caerulea, all of its other known populations are from southwestern Turkey (Örstan, 2001), the adjacent Greek islands (Nordsieck, 1977; Zilch, 1977) and southern Greece (Giokas & Mylonas, 2002), so the Istanbul area certainly represents the northernmost limit of its range. Since peripheral populations are often genetically and morphologically divergent from central populations, and since genetically divergent populations are valuable as potential sites of future speciation events (Mayr, 1970; Lesica & Allendorf, 1995), peripheral populations are important candidates for conservation (Lesica & Allendorf, 1995). At least until further studies have been carried out to evaluate the degree of genetic divergence from central populations of the peripheral land snail colonies around Istanbul, the habitats of peripheral land snail colonies should be protected. Özhatay et al. (2003) classified the conservation needs of these limestone meadows as "very urgent". Our observations during the survey support the determination of Özhatay et al. (2003) that the limestone meadows north–northwest of Istanbul are under imminent threat from the expansion of the nearby residential neighborhoods. For example, our station C1 was on a cliff below a densely populated hilltop, while C12 was less than 100 m from recently constructed apartment buildings. The overtaking of the latter station by further develop-

Animal Biodiversity and Conservation 28.2 (2005)

ment is probably only a matter of time. A photograph taken from station C6 (fig. 3) shows the extent of encroaching development and illustrates the general threat these habitats are facing. On the other hand, the presence of a dam within our survey area that was built to supply water for the ever–growing population of Istanbul paradoxically offers some protection to the surrounding land (Özhatay et al., 2003). The regulations of the municipal agency that administers the dam, Istanbul Su ve Kanalizasyon Idaresi (ISKI; Istanbul Water and Sewer Administration), prohibit all development, other than those associated with water purification, within 1000 m of water reservoirs if the water collection basin in question extends at least that far (ISKI, 2004). Therefore, if these regulations are enforced as intended, the presently undeveloped land within 1,000 m of the dam lake may be considered to be under protection for the time being. However, the same regulations do allow for residential buildings outside the 1,000–m limit, which means that our stations C1, C2, C3 and C12 and the surrounding areas located up to about 4 km away from Sazlidere Dam may be lost in the future unless protected. Turkey currently has a number of national parks and various types of nature conservation areas (Kaya & Raynal, 2001; Guclu & Karahan, 2004). The protection provided by such areas to all wildlife notwithstanding, the establishment of parks and conservation areas is usually justified in terms of the protection they will offer to mostly large mammals, birds and plants (Yilmaz, 1998; Kaya & Raynal, 2001; Guclu & Karahan, 2004), while the conservation needs of invertebrates are almost never taken into consideration. The land snail faunas in many countries are increasingly being threatened with extinction (Lydeard et al., 2004). Turkey has a rich land snail fauna with many endemic species (Schütt, 2001) that, in our opinion, deserve no less protection than any other animal or plant group. We hope that our results regarding these threatened limestone meadows will bring attention to the conservation needs of the native land snails in particular and invertebrates in general. Acknowledgements We thank Teri Varnali for driving us to and around the survey area, Bernhard Hausdorf for his help with some of the identifications and Zeki Yildirim for comments that improved the manuscript. References Anonymous, 1844. Map of Constantinople, Stambool. In: Maps of the Society for the Diffusion of Useful Knowledge, vol. 1. Chapman and Hall, London. Available from http://www.davidrumsey.com

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Boycott, A. E., 1934. The habitats of land Mollusca in Britain. Journal of Ecology, 22: 1–38. Falkner, G., 1990. Binnenmollusken. In: Weichtiere. Europäische Meeres– und Binnenmollusken (R. Fechter & G. Falkner, Eds.). Mosaik, München. Forcart, L., 1940. Monographie der türkischen Enidae. Verhandlungen der Naturforschenden Gesellschaft, 51: 106–263. Germain, L., 1930. Mollusques Terrestres et Fluviatiles. Faune de France, vol. 21. Lechevalier, Paris. Giokas, S. & Mylonas, M., 2002. Spatial distribution, density and life history in four Albinaria species (Gastropoda, Pulmonata, Clausiliidae). Malacologia, 44: 33–46. Gittenberger, E., 1967. Die Enidae (Gastropoda, Pulmonata) gesammelt von der niederländischen biologischen Expedition in die Türkei in 1959. Zoologische Mededelingen, 42: 125–141. Guclu, K. & Karahan, F., 2004. A review: the history of conservation programs and development of national parks concept in Turkey. Biodiversity and Conservation, 13: 1373–1390. Hecker, U., 1965. Zur Kenntnis der mitteleuropaeischen Bernsteinschnecken (Succineidae). I. Archive für Molluskenkunde, 94: 1–45. Istanbul Metropolitan Municipality, 2005. The population of Istanbul according to the census years. Available from http://www.ibb.gov.tr/IBB/DocLib/ pdf/bilgihizmetleri/yayinlar/istatistikler/demografi/ t211.pdf Istanbul Su ve Kanalizasyon Idaresi (ISKI), 2004. Içmesuyu Havzalari Koruma ve Kontrol Yönetmeligi [Drinking Water Basins Protection and Control Regulations]. Available from http://www.iski.gov.tr/dosya/ yonetmelikler/ICHKOYNT.pdf Kaya, Z. & Raynal, D. J., 2001. Biodiversity and conservation of Turkish forests. Biological Conservation, 97: 131–141. Kerney, M. P. & Cameron, R. A. D., 1979. A Field Guide to the Land Snails of Britain and North– west Europe. Collins, London. Lesica, P. & Allendorf, F. W., 1995. When are peripheral populations valuable for conservation? Conservation Biology, 9: 753–760. Lydeard, C., Cowie, R. H., Ponder, W. F., Bogan, A. E., Bouchet, P., Clark, S. A., Cummings, K. S., Frest, T. J., Gargominy, O., Herbert, D. G., Hershler, R., Perez, K. E., Roth, B., Seddon, M., Strong, E. E. & Thompson, F. G., 2004. The global decline of nonmarine mollusks. BioScience, 54: 321–330. Mayr, E., 1970. Populations, Species, and Evolution. Harvard Univ. Press, Cambridge. Meriç, E., Sakinç, M., Özdogan, M. & Açkurt, F., 1991. Mollusc shells found at the Yarimburgaz Cave. Journal of Islamic Academy of Sciences, 4: 6–9. Nordsieck, H., 1977. Taxonomische Revision des Genus Albinaria. Archive für Molluskenkunde, 107: 285–307.

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Örstan, A., 2001. A preliminary survey of Albinaria populations around Kusadasi Bay, Turkey. Triton, No. 4: 42–44. – 2005. The status of Pomatias elegans in Istanbul, Turkey. Tentacle, No. 13: 8–9. Available from http://www.hawaii.edu/cowielab/ Tentacle.htm Özhatay, N., Byfield, A. & Atay, S., 2003. Türkiye’nin Önemli Bitki Alanlari [Important Plant Areas in Turkey]. WWF Türkiye, Istanbul. Reuse, C., 1983. On the taxonomic significance of the internal shell in the identification of European slugs of the families Limacidae and Milacidae (Gastropoda, Pulmonata). Biologisch Jaarboek Dodonaea, 51: 180–200. Riedel, A., 1995. Zonitidae sensu lato (Gastropoda, Stylommatophora) der türkei. Übersicht der Arten. Fragmenta Faunistica, 38: 1–86. Rundell, R. J. & Cowie, R. H., 2003. Preservation of species diversity and abundances in Pacific island land snail death assemblages. Journal of Conchology, 38: 155–163. Schilthuizen, M. & Lombaerts, M., 1994. Population structure and levels of gene flow in the Mediterranean land snail Albinaria corrugata (Pulmonata: Clausiliidae). Evolution, 48: 577– 586.

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Schütt, H., 2001. Die türkischen Landschnecken. Acta Biologica Benrodis, Supplementband 4: 1–549. Sturany, R., 1894. Zur Molluskenfauna der europäischen Türkei. Annalen des K. K. Naturhistorischen Hofmuseums Wien, 9: 369–390. – 1902. Beitrag zur Kenntnis der kleinasiatischen Molluskenfauna. Sitzungsberichte der Mathematisch–Naturwissenschaftlichen Classe der Kaiserlichen Akademie der Wissenschaften Wien, 111: 123–140. Welter–Schultes, F., 1998. Human–dispersed land snails in Crete, with special reference to Albinaria (Gastropoda: Clausiliidae). Biologica Gallo– hellenica, 24: 83–106. Yildirim, M. Z., Kebapçi, Ü. & GümüÕ, B. A., 2004. Edible Snails (Terrestrial) of Turkey. Turkish Journal of Zoology, 28: 329–335. Yilmaz, K. T., 1998. Ecological diversity of the Eastern Mediterranean region of Turkey and its conservation. Biodiversity and Conservation, 7: 87–96. Zilch, A., 1977. Die Typen und Typoide des Natur– Museums Senckenberg, 57: Mollusca: Clausiliidae. Archive für Molluskenkunde, 107: 309–363.

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Principles and interest of GOF tests for multistate capture–recapture models R. Pradel, O. Gimenez & J.–D. Lebreton

Pradel, R., Gimenez, O. & Lebreton, J.–D., 2005. Principles and interest of GOF tests for multistate capture–recapture models. Animal Biodiversity and Conservation, 28.2: 189–204. Abstract Principles and interest of GOF tests for multistate capture–recapture models.— Optimal goodness–of–fit procedures for multistate models are new. Drawing a parallel with the corresponding single–state procedures, we present their singularities and show how the overall test can be decomposed into interpretable components. All theoretical developments are illustrated with an application to the now classical study of movements of Canada geese between wintering sites. Through this application, we exemplify how the interpretable components give insight into the data, leading eventually to the choice of an appropriate general model but also sometimes to the invalidation of the multistate models as a whole. The method for computing a corrective overdispersion factor is then mentioned. We also take the opportunity to try to demystify some statistical notions like that of Minimal Sufficient Statistics by introducing them intuitively. We conclude that these tests should be considered an important part of the analysis itself, contributing in ways that the parametric modelling cannot always do to the understanding of the data. Key words: Memory, Transients, Trap–dependence, Test WBWA, Contingency tables partitioning, Test M. Resumen Principios e interés de los test Bondad de Ajuste (GOF) para los modelos de captura–recaptura multiestado.— Los procedimientos óptimos de bondad de ajuste, aplicados a los modelos multiestado, son nuevos. Trazando un paralelismo con los correspondientes procesos de uniestado, presentamos sus particularidades y mostramos como el test general puede descomponerse en componentes susceptibles de ser interpretados. Todos los desarrollos teóricos están ilustrados con una aplicación del ya clásico estudio de los desplazamientos de la barnacla canadiense entre sus lugares de invernada. Mediante esta aplicación, presentamos un ejemplo de cómo los componentes susceptibles de ser interpretados nos proporcionan una idea de los datos que nos pueden llevar a la elección de un modelo general apropiado, pero también a veces a la invalidación de los modelos de multiestados en su conjunto. Se menciona entonces el método para calcular un factor de corrección de la sobredispersión. Aprovechamos esta ocasión para intentar también desmitificar algunas nociones estadísticas, como las Estadísticas Suficientes Mínimas, introduciéndolas intuitivamente. La conclusión es que estas pruebas deberían considerarse una parte importante del propio análisis, contribuyendo a la comprensión de los datos, de un modo que el modelaje paramétrico no siempre consigue. Palabras clave: Memoria, Transeúntes, Dependencia del trampeo, Test WBWA, Partición de tablas de contingencia, Test M. (Received: 8 VII 05; Final acceptance: 19 VII 05) Roger Pradel, Olivier Gimenez & Jean–Dominique Lebreton, CEFE, CNRS, 1919 Route de Mende, 34293 Montpellier cedex 5, France.

ISSN: 1578–665X

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Introduction Multistate capture–recapture models are very appealing for studying a variety of biological questions such as dispersal where states are geographical sites (Hestbeck et al., 1991), trade–off between reproductive status and survival where states are breeder vs. non–breeder (Nichols et al., 1994), rate of accession to reproduction where states are experienced vs. inexperienced breeders (Pradel & Lebreton, 1999), etc. Furthermore, different types of demographic information, such as live recaptures and recoveries of dead individuals by the public can be analyzed simultaneously using adequate multistate models (Lebreton et al., 1999). A general review of the biological relevance of multistate capture–recapture models can be found in Lebreton & Pradel (2002). In multistate capture–recapture models (Arnason, 1972, 1973; Hestbeck et al., 1991), marked individuals can move among a finite number of states, or die, between discrete time occasions. Survivors are detected ("encountered") in each state, not exhaustively at each occasion. Based on parameters which are the transition, survival and encounter probabilities, the probability of an individual encounter history —conditional on the date and state of first encounter, marking and release— can be calculated in a way similar to that used for the classical one–state Cormack–Jolly–Seber (CJS) model. Under the assumption of independence between individuals, the likelihood for a particular data set is then obtained as the product of the probabilities for each individual encounter history. The rationale of model selection, based on the AIC, assumes that the set of models considered encompasses a model that fits the data (Burnham & Anderson, 1998). If not, the deviance will tend to be inflated, favoring the incorrect selection of overparametrized models and thus leading to erroneous biological conclusions. Moreover, the precision of the final estimates will also be biased if some lack–of–fit or overdispersion is ignored. The consequences of lack–of–fit are thus too deleterious to be ignored. Yet, difficulties with goodness– of–fit issues have been recurrent in the application of capture–recapture methodology. In a survey of the literature, Begon (1983) concluded that fewer than 11% of the applications of the Jolly–Seber model addressed in a quantitative way or discussed the assumptions inherent in the model. This state of fact was the consequence of the absence at that time of any general goodness–of– fit procedure. The simplest approach, which consists of comparing observed and expected numbers of animals with a particular encounter history, was hampered by the large number of encounter histories (in the one–site case, with 10 occasions, there are more than 1,000 different encounter histories), and as a consequence by the very low expected numbers (the resulting sparseness makes ² distributions for quadratic X² statistics or for the deviance quite inadequate (McCullagh & Nelder,

Pradel et al.

1989)). Today, bootstrap procedures may be a way around distributional problems. However, another weakness of the omnibus approach of comparing expected and observed numbers is that it lacks power against specific alternatives and that it is not informative when it rejects. Specialized tests have been built to address frequent causes of departure. Examples are the Leslie–Carothers test of equal catchability (Carothers, 1971), the Brownie–Robson test of marking–induced deaths (Robson, 1969; Brownie & Robson, 1983), which has later been shown to test also for the presence of transients (Pradel et al., 1997), and, in the context of multistate models, a test of memory (Pradel et al., 2003). However, the relationships between the particular tests will remain unknown until a careful study of the likelihood is carried out. Only such a study can provide the basis for a sound partitioning of the information. A major step in this direction was the development of optimal goodness–of–fit procedures for the CJS model (Pollock et al., 1985). The global test, organized into several interpretable components and based on adequately pooled tables, was implemented in the RELEASE programme (Burnham et al., 1987). Since then, several specialized tests have been shown to be components of this general test (the test for the presence of transients (Pradel et al., 1997): that for trap– dependence (Pradel, 1993)) and a slightly different version of the general test is now proposed in program U–CARE (Choquet et al., 2005). Recently, Pradel et al. (2003) have developed a similar approach for the multistate model called JMV (Brownie et al., 1993), a model which generalizes the Arnason–Schwarz model by allowing encounter probabilities to vary by site occupied at the previous occasion. The AS model, regarded as the reference model for multistate capture–recapture, has not yet received a specific treatment. The purpose of this paper is to review the principles on which the goodness–of–fit tests for CJS and JMV are based, underlining their similarities and differences, and to examine how alternatives of interest can be embedded within the general tests. This paper is intended for the biologist with some experience of capture–recapture analysis but no deep statistical training. Thus, we assume that the reader knows what the CJS and the AS models are. On the other hand, we have tried to use everyday words in place of statistical terms. For instance, we seek to introduce notions like minimal sufficient statistics from a practical angle. Most of the paper is illustrated with one example, that of the Canada goose data originally analyzed by Hestbeck et al. (1991). We proceed by steps. First, we present and discuss the features of the goodness–of–fit test of the simpler CJS model and specialized tests embedded within it. For the specialized tests, we examine some statistics particularly suitable to address the alternatives of interest. The second section presents and discusses the goodness–of–fit test of the JMV model, drawing a parallel —as far

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Table1. m–array for the Canada goose data (Hestbeck et al., 1991) pooled over sites: i. Occasion of release; Ri. Number released at i; mij. Number reencountered at j among those released at i; ri. Number ever reencountered among those released at i; mj. Total number reencountered at occasion j. Tabla 1. Serie m para los datos de la barnacla canadiense (Hestbeck et al., 1991) una vez reunidos los de diversas localidades: i. Ocasión de liberación; Ri. Número de liberaciones en la ocasión i; m ij . Número de reencuentros siguientes en la ocasión j con una liberación dada en la ocasión i; ri. Número que se han vuelto a ver con una liberación dada en la ocasión i; mj. Número de reencontrados en la ocasión j.

Table 2. A fake m–array with the same margins as that of the Canada goose data. (For abbreviations see table 1.) Tabla 2. Una serie m simulada, con los mismos márgenes que la de los datos de la barnacla canadiense. (Para las abreviaturas ver tabla 1.) mij's i

3,494 1,138

7,098

7,603

6,804

5,170

459

42 1,722

2,241

745

154 3,174

2,580

629 3,249

2,205

402 2,607 1,472

1,138 2,250 3,073 3,054 2,699

mij’s i

1 3,494 1,138 2 7,098 3 7,603

309

159

42 1,722

1,941

734 2,180

345 154 3,174 740 329 3,249

4 6,804

1,905 702 2,607

5 5,170

1,472

1,138 2,250 3,073 3,054 2,699

as is possible— with the goodness–of–fit test of the CJS model. Finally, the last section is devoted to proposals for the improvement of the present situation and tries to identify future directions of research. The material presented in this paper has been implemented in program U–CARE, and is freely available at http://ftp.cefe.cnrs.fr/biom/Soft–CR/. Likelihood–based goodness–of–fit test for the single–site CJS model A perfect segregation of information between "estimation of parameters" and "test of assumptions" For the sake of illustration, let us consider the observations of 28,849 Canada Geese (Branta canadensis) banded with individually–coded neck bands and re–observed at three locations: the mid– Atlantic (New York, Pennsylvania, New Jersey), the Chesapeake (Delaware, Maryland, Virginia), and the Carolinas (North and South Carolina) (Hestbeck et al., 1991). Ignoring the locations for the moment, the data can be summarized in what is called an m– array (table 1). At the beginning of each row is the number of geese released on each occasion, followed by the numbers of them reencountered for the

first time on each subsequent occasion. The m– array is an interesting summary because it turns out that any set of encounter histories that produces the same m–array yields the same maximum likelihood estimates (MLE) of survival and encounter probabilities under the Cormack–Jolly–Seber (CJS) model. For this reason, the m–array is said to be a sufficient statistic for the CJS model. Actually, even the margins of the m–array, i.e. the total number of reencounters per occasion mj’s and the numbers ever seen again among those released at every occasion ri, are sufficient (Burnham et al., 1987). This is in fact the maximum reduction possible and these margins are thus logically called minimal sufficient statistics (MSS). Table 2 presents a different m–array with the same margins as the Canada Geese data set. Thus, this m–array leads to the same MLE’s under the CJS model. However, of two data sets that lead to the same estimates one may respect the model assumptions while the other may not. For instance, of the 3,494 individual geese released at occasion 1, we know for sure that 309 + 159 + 64 + 42 = 574, which had not been encountered at occasion 2 but were encountered later, were still alive at occasion 3. At the same time, 1,941 + 734 + 345 + 154 = 3,174 of the 7,098 geese released at occasion 2 were also alive. The two groups had experienced distinct encounter histories up to occasion 3 but under the assumptions of the CJS model, this should be irrelevant as regards their future; for instance, each of them should have an equal chance of being encountered at occasion 3. This may be tested using a contingency table: Seen at 3 Seen later Last seen at 1

309

265

Last seen at 2

1,941

1,233

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As it turns out, this table is slightly unbalanced but far less than the corresponding table from the fake m–array: Seen at 3

Seen later

Last seen at 1

565

Last seen at 2

2,241

933

Thus, while knowing the MSS suffices to estimate the parameters, details of the data i.e. the encounter histories, are needed to test the model assumptions. One may wonder, on the other hand, whether something can be learned from the MSS about the respect of model assumptions. The answer to this question depends on the model. In general, there is indeed something to learn from the examination of the MSS but not in the case of the CJS model. The CJS model has indeed a peculiarity: its number of MSS is exactly equal to its number of parameters. For instance, the Canada goose study spans 6 years; thus there are 5 m’s and 5 r’s, and hence a total of 10 margins. However, the sum of the m’s and that of the r’s are both equal to the total number of animals in the data set. Therefore, there are only 9 minimal sufficient statistics (one margin can be spared). The CJS model with 6 occasions has 5 survival and 5 encounter parameters, 10 parameters in total; but again, at the last time step, only the product of the last survival by the last encounter is estimable, and hence there are only 9 true parameters in total. It can be shown that every time that the number of individual statistics making up the MSS is exactly equal to the number of parameters in the model — as is true of the CJS model— there is nothing to learn from the MSS with respect to model assumptions. The likelihood can always be factorized into two terms: one, the probability of the encounter histories given the MSS, and the other, the probability of the MSS given the parameters. Pr (data; )= Pr (data / MSS) Pr (MSS; )

(1)

In the case of the CJS model, (1) corresponds to a perfect separation of the information. Pr (data / MSS) serves to check the model assumptions, and Pr (MSS; ) serves solely to estimate the parameters. The construction of an optimal goodness–of–fit test is thus based on the sole first part, Pr (data / MSS). The CJS model makes several assumptions. Based on the encounter histories of otherwise similar individuals, not all are verifiable: for instance, the assumption that the marked animals are representative or that the band codes are not misread. In fact, based on the study of the part Pr (data / MSS) of the likelihood, it can be shown that the verifiable assumptions come down to essentially one thing: all animals present at any given time are assumed to behave the same. Pollock et al. (1985) have further shown that this, in turn, can be divided into two (conditionally) independent main

points to be checked: 1) all animals released together have the same expected future whatever their past encounter history and 2) all animals alive at the same date that will be seen again do not differ in the timing of their reencounters whether they are currently encountered or not. The first point leads to what is known as TEST 3; the second to TEST 2, which is also known as the Jolly–Balser test (Balser, 1984). This is actually not the only way of breaking down the general test (Pollock et al. do propose another form of their goodness–of–fit test) but it is the most commonly used and the one we will consider. Starting from this decomposition, it becomes possible to see how tests of specific hypotheses articulate with the general test and among themselves. This has not been done systematically and to our knowledge, the Leslie– Carothers test of unequal catchability (Carothers, 1971) for instance has never been related to the optimal GOF test of CJS. There are already at least two specific tests which have been fully incorporated into the GOF test of the CJS model and to which alternative models have been attached. We examine them now in turn. A test of transience TEST 3 theoretically compares, at each occasion, the future history of encountered individuals with respect to their previous encounter history. In practical implementations, the comparison is limited to newly marked and previously marked individuals. That these two categories should have similar expectations implies an equal chance of being seen again. It is thus possible to distinguish two steps in TEST 3: first, the check that newly and already– marked animals have an equal chance of being seen again and then, for those seen again, the check that the spread of next reencounters over time is similar in the two categories. (This corresponds in practice to the partitioning of contingency tables, a very classical statistical technique.) The first subcomponent (table 3) has been suggested many times and has been known since at least 1969 (Robson, 1969). It has received an interpretation as a test for an effect of marking on immediate survival i.e. in the period immediately following release (Brownie & Robson, 1983). It has also been shown to be the adequate test to detect the presence of transients, animals that are passing through the study site en route to other locations (Pradel et al. 1997). This test is called the Brownie–Robson test or TEST 3.SR. An interesting point is that it is the test of comparison of CJS —or ( t, pt) in the notation of Lebreton et al. (1992)— with the more general model that provides for 2 age–classes in survival ( a2*t, pt). As a consequence, a GOF test for model ( a2*t, pt) is readily available from the GOF test of CJS by ignoring subcomponent TEST 3.SR. If the alternative of interest is the presence of transients, the direction of departure is predictable. In this case a directional test is appropriate. One such test can be computed by taking the square

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Table 3. TEST 3.SR. This subcomponent of the CJS goodness–of–fit test is also a specific test of transience. The signs indicate the expected difference between observed and expected values if there are transients caught in the samples: Sl. Seen later; Nsa. Never seen again; Nsb. Never seen before; Sb. Seen before. Tabla 3. TEST 3.SR. Este subcomponente del test de bondad de ajuste CJS es también un test específico de divagancia. Los signos indican la diferencia esperada entre los valores observados y esperados, si en las muestras existen transeúntes: Sl. Visto posteriormente; Nsa. No se ha vuelto a ver; Nsb. Nunca visto con antelación; Sb. Visto con antelación.

Nsa

Nsb ("new" or "newly marked")

–

Sb ("old" or "already marked")

–

Table 4. Results of TEST 3.SR for the Canada goose data. The test can be calculated at each of the 4 intermediate occasions. The table gives the Pearson chi–square statistics (X2) and the corresponding P–value (P) as well as the signed square root (z) of the Pearson chi–square statistic which is normally N(0,1) distributed. z is positive when there is an excess of never seen again among the newly marked. Tabla 4. Resultados del TEST 3.SR para los datos de la barnacla canadiense. Este test puede calcularse en cada una de las cuatro ocasiones intermedias. La tabla proporciona la ji–cuadrado de Pearson (X 2 ) y su correspondiente valor P (P), así como la raíz cuadrada provista de signo (z) de la ji– cuadrado de Pearson, que presenta una distribución normal N(0,1). z es positiva cuando existe un exceso de individuos nunca vistos antes entre los que acaban de ser marcados.

Occasion

root of the Pearson chi–square statistics and giving it a conventional sign (see table 3). For the Canada geese (table 4), the overall test is highly significant ( 2(4) = 54.24; P < 10–10). A more specific and thus more powerful overall test of transience can be based on the statistic

1 p

1.11

1.24

0.27

5.16

26.58

0.00

3.79

14.33

0.00

3.48

12.09

0.00

∑z

(for p components)

i =1

which is standardized normal N(0,1) under H0 (no transience or age effect) and will tend to be positive under H1. Here z = 6.766 is highly significant. A test of short–term trap–dependence The other test of a specific alternative that has been fully incorporated in the general GOF test of CJS is directed at detecting immediate trap–dependence on encounter probability, meaning that animals that are encountered at occasion i have a different, higher (in case of trap–happiness) or lower (in case of trap–shyness), probability of encounter than the rest of the population at the next occasion i+1 (table 5). The tables built in Section "A perfect segregation of information between 'estimation of parameters' and 'test of assumptions'" are examples of this test. As mentioned earlier, TEST 2 compares the future of animals alive at the same occasion which are then encountered or not encountered. Just as TEST 3.SR was obtained as a subcomponent of TEST 3, the test of trap–dependence, called 2.CT, is obtained as a subcomponent of TEST 2. (The complement of TEST 2.CT in TEST 2 investigates the timing of next encounters

Table 5. TEST 2.CT. This subcomponent of the CJS goodness–of–fit test is also a specific test of immediate trap–dependence. The signs indicate the expected difference between observed and expected values in case of trap– happiness. They should be reversed for trap– shyness: Ei+1. Encountered at i+1; El. Encountered later; NEi. Not encountered at i; Ei. Encountered at i. Tabla 5. TEST 2.CT. Este subcomponente del test de bondad de ajuste CJS es también un test específico de la dependencia del trampeo inmediato. Los signos indican la diferencia esperada entre los valores observados y esperados en el caso de animales habituados a la trampa. Deberían ser contrarios en el caso de individuos no habituados a la trampa: Ei+1. Encontrado en i+1; El. Encontrado más tarde; NEi. No encontrado en i; Ei. Encontrado en i.

Ei+1

NEi

–

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Table 6. Results of TEST 2.CT for the Canada goose data. This test can be calculated at each intermediate occasions but the penultimate. The table gives the Pearson chi–square statistics (X2) and the corresponding P–value (P) as well as the signed square root (z) of the Pearson chi–square statistics which is normally N(0,1) distributed. z is negative when there is an excess of "encountered at i+1" among the "encountered at i". Tabla 6. Resultados del TEST 2.CT para los datos de la barnacla canadiense. Este test puede calcularse en cada una de las ocasiones intermedias exceptuando la penúltima. La tabla proporciona la ji–cuadrado de Pearson (X2) y su correspondiente valor P (P), así como la raíz cuadrada provista de signo (z) de la ji– cuadrado de Pearson, que presenta una distribución normal N(0,1). z es negativa cuando existe un exceso de individuos "encontrados en i+1" entre los "encontrados en i".

Component

-3.29

10.86

0.00

-5.02

25.16

0.00

-3.13

9.80

0.00

of the animals not encountered at i+1.) The alternative model to CJS here is the generalization, noted ( t,pt*m), allowing for a different encounter probability of animals just released. A GOF test for this model can be obtained by leaving out subcomponent 2.CT from the GOF test of CJS. Unlike for transients, the direction of departure can be in any direction with trap–dependence. Yet, we expect the effect to be consistent over occasions. Thus the signed z statistic remains useful when combining the TESTs 2.CT of the different occasions: the evidence for a trap effect accumulates with tables repeatedly unbalanced in the same direction (table 6). For the geese, there is overwhelming evidence that encounter probability is much higher for a goose encountered at the previous occasion. Both the omnibus chi–squared statistics and the directional test are highly significant (X2 = 45.8212, P < 10–9; z = –6.6061, P < 10–10). However, we have up to now ignored the site of observation. If, as is likely, the effort of observation is unequal and the geese tend to be faithful to the same site from year to year, a goose that frequents the site with high observation pressure will tend to be reobserved consistently more often, leading to a spurious trap effect. To get around this problem, we now turn our attention toward multisite (also multistate) models, more specifically, the JMV model.

Likelihood–based goodness–of–fit test for the multistate model JMV In multisite protocols, the individuals are sampled over K occasions and s sites. In the example of the Canada goose, there are 3 main areas in the Atlantic flyway, which we will now distinguish. The data can again be summarized in a multisite or multistate m– array (Brownie et al., 1993) (table 7), a generalization of the m–array for one–site data. The comparison of table 7 with table 1 should make clear how the multistate m–array is built. Therefore, we introduce here another approach to the m–array. Each encounter history can be split into several pieces, from the first release to the next reencounter, from the subsequent release to the next reencounter and so on until the end of the study period. For instance, the capture history 302300 over 6 occasions may be seen as made of the three pieces: 302000, 002300 and 000300. Each time that the individual is reencountered (the first two pieces), it is treated as if removed from the data set; this insures that only one individual remains present at the same time in the data set. Each piece is then treated as if coming from an independent individual. The m–array is essentially the tally of these pieces arranged by rows according to the occasion and state of release, and by columns according to the occasion and state of next reencounter (plus a "never–reencountered" column). Obviously, for a model that assumes that the fate of an individual is not affected by its past capture history, the information retained is sufficient. However, because of the loss of information accompanying the construction of the m–array, some assumptions can no longer be checked; for instance, whether some individuals are encountered significantly more often than others. This explains why the objectives of checking the model assumptions and that of estimating the parameters tend to use the complementary part of the total information. An imperfect segregation of information between "estimation" and "test of assumptions" The basic assumptions inherent in the JMV model are similar to those of the CJS model except that differences between individuals in the different states are now acknowledged. Again, the fate of the individuals that are in the same state at the same time does not depend on their past. A consequence is that the multistate m–array is a sufficient statistic. Moreover, it can be shown that, unlike the one–site m–array, the multistate m–array is minimally sufficient. Now, the number of sufficient statistics, i.e. the number of independent cells in the multistate m–array, is K (K – 1) s² / 2. This is greater than the number of identifiable parameters as soon as K > 3: there are indeed (K – 1) s² transition probabilities, plus (K – 1) s2 encounter probabilities minus s2 because the encounter probabilities of the last occasion are not estimable separately from the last transitions; a total of (2K – 3) s2 true parameters. The JMV model does not therefore have the nice properties of the CJS model. For instance, it is

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Animal Biodiversity and Conservation 28.2 (2005)

Table 7. Multisite m–array for the Canada goose data. Sites are North Atlantic (1), the Chesapeake (2) and the Carolinas (3). Only the first 2 and the last occasions of release are shown: mijr s. Number of next reencounter at occasion j in site s given release at occasion i in site r; i. Release occasion; r. Release sites; Rir. Number released. Tabla 7. Serie m multilocalidad para los datos de la barnacla canadiense. Las localidades son el Atlántico Norte (1), la región de Chesapeake (2) y las dos Carolinas (3). Sólo se muestran las dos primeras y la última ocasión de liberación: mijr s. Número del siguiente reencuentro en la ocasión j en la localidad s dada la liberación en la ocasión i en la localidad r; i. Ocasión de liberación; r. Localidades de liberación; Rir. Número de liberaciones. mijrs 2

Rir

785

239

2,086

85 615

36 158

623

2,082

3,918

1,698

… …

49 67

491 134

149

159 869

63 335

41 164

14 101 158

14 11

…

1,291

271

2,887

137 654 18

992

18 105 16

no longer possible to factorize the likelihood into a term used solely for parameter estimation and another for assessing the goodness of fit; the term Pr(MSS; ) of formula (1) retains some information about the respect of model assumptions and has to be examined when assessing the fit of the JMV model. There is, however, an analogy with the CJS model which still holds. The verifiable assumptions come down again to one very similar thing: all animals present at any given time at the same site behave the same. And this is again equivalent to the verification of two (conditionally) independent points: 1) animals released together have the same expected future whatever their past encounter history and 2) animals present at the same site at the same date that are eventually reencountered do not differ in the timing of their reencounters whether they are currently encountered or not. Thus, apart for the precision of a common site, the exact same two main components are retrieved. Past encounter history should not matter (TEST 3G) The first main component of the GOF test of the JMV model, called TEST 3G, examines the effect of the past capture history on the future of animals captured and released at the same time on the same site (Pradel et al., 2003). It is thus the equivalent of TEST 3 of which it is a generalization. Again, there are many possible past capture histo-

… … 99

ries and the practical implementation of this test as found in program U–CARE version 2.0 (Choquet et al., 2003) considers only a limited number of situations: the newly caught animals are on the first row while the previously caught ones are dispatched over the subsequent rows according to their site of most recent encounter (see table 8); the columns correspond to the particulars (time and site) of the next encounter if any. As can be seen in table 8, even with a large data set like that of the Canada geese, empty cells easily occur and some sort of pooling is needed. The results in table 9 were obtained with U–CARE version 2.0 which has an automatic pooling algorithm built in. They show that the Canada geese caught together differ strongly depending on their past (over all TEST 3G: X2(103) = 749.27; P < 10–14). A close examination of the individual contingency tables like that of table 8, especially the comparison of expected and observed numbers in each cell, might prove useful in understanding the reasons for the departure. However, the breaking up of TEST 3G into meaningful subcomponents is a better option. A generalized test of transience A first subcomponent can be built to test for the presence of transients in each sample defined by a site and an occasion (table 10). This straightfor-

Pradel et al.

196

Table 8. Component 3G (2,1) of the JMV goodness–of–fit test applied to the Canada goose data. This component is based on the animals caught at occasion 2 on site 1. They are dispatched according to the site of most recent encounter in rows and the particulars (time and location) of the next encounter in columns. The "–" sign is used for animals that are caught for the first time (first row) or that will never be encountered again (last column). Tabla 8. Componente 3G (2,1) del test de bondad de ajuste JMV aplicado a los datos de la barnacla canadiense. Este componente se basa en los animales capturados en la ocasión 2 en la localidad 1. Se han distribuido en filas según el lugar de encuentro más reciente y en columnas según las circunstancias (tiempo y localidad) del siguiente hallazgo. El signo "–" se utiliza para los animales capturados por primera vez (primera fila) o que nunca serán vueltos a encontrar (última columna).

Time (j) and location (v) of first reencounter Location

at time 1

–

3 1

–

390 124

122

920

128

Table 9. Results of TEST 3G for the Canada goose data. The table gives the Pearson chi– square statistic ( 2) and the corresponding P–value (P) as well as the number of degrees of freedom after pooling (df): Oc. Occasion; S. Site. Tabla 9. Resultados del TEST 3G para los datos de la barnacla canadiense. Las tablas presentan la ji–cuadrado de Pearson (X 2) y su correspondiente valor P (P), así como el número de grados de libertad (df) tras la reunión: Oc. Ocasión; S. Localidad.

40.85

0.000

6.73

0.566

17.57

0.007

115.35

0.000

72.64

0.000

57.64

0.000

89.19

0.000

94.49

0.000

50.57

0.000

62.33

0.000

53.75

0.000

88.16

0.000

Table 10. A suitable partitioning of TEST 3G isolates TEST 3G.SR. This subcomponent is exactly similar to TEST 3.SR but involves a further stratification by site (and not only by date). It is also a specific test of transience. The signs indicate the expected difference between observed and expected values when transients are caught in the samples: Sl. Seen later, Nsa. Never seen again; Nsb. Never seen before; Sb. Seen before. Tabla 10. Una partición adecuada del TEST 3G aísla al TEST 3G.SR. Este subcomponente tiene una similitud exacta con el TEST 3.SR, pero presenta una mayor estratificación respecto a la localidad (y no sólo a la fecha). También constituye un test específico de la divagancia. Los signos indican la diferencia esperada entre los valores observados y esperados cuando en las muestras se capturan transeúntes: Sl. Visto posteriormente; Nsa. No se ha vuelto a ver; Nsb. Nunca visto con antelación; Sb. Visto con antelación.

Nsa

Nsb ("new" or "newly marked")

–

Sb ("old" or "already marked")

–

197

Animal Biodiversity and Conservation 28.2 (2005)

Table 11. Results of TEST 3G.SR for the Canada goose data: Oc. Occasion; S. Site. Tabla 11. Resultados del TEST 3G.SR para los datos de la barnacla canadiense: Oc. Ocasión; S. Localidad.

0.004

0.95

0.000

0.99

8.130

0.00

11.394

0.00

2.708

0.10

33.459

0.00

10.608

0.00

0.353

0.55

10.168

0.00

11.013

0.00

0.129

0.72

29.785

0.00

Table 12. TEST WBWA, this subcomponent of TEST 3G tests for a memory effect. It compares the site of the most recent observation in row ("Where Before") to the site of the next observation in column ("Where After"). The "+" signs indicate where the observed values should exceed the expected values when animals tend to return to previously visited sites: Nss. Next seen on site; Lss. Last seen on site. Tabla 12. TEST WBWA, este subcomponente TEST 3G comprueba el efecto de la memoria. Compara la localidad de la observación más reciente en la fila ("Where Before"), con la localidad de la siguiente observación en la columna (“Where After”). El signo "+" indica dónde los valores observados deberían exceder a los valores esperados, cuando los animales tienden a volver a las localidades previamente visitadas: Nss. Siguiente avistamiento en la localidad; Lss. Último avistamiento en la localidad.

Nss 1 Nss 2 Lss 1 Lss 2

ward generalization of TEST 3.SR, called TEST 3G.SR, provides a GOF test of a generalized JMV model, a model with two age classes on survival. This generalization of the classical JMV model is denoted Fa2*from*t, from*to*t, pfrom*to*t in the notation of Choquet et al. (2004): F is survival, transition and p encounter probability ; from is the site of departure, to the site of arrival. The goodness–of–fit test of this model is obtained by leaving out subcomponent 3G.SR when calculating the overall GOF test. Applied to the Canada geese, TEST 3G.SR reveals an interesting feature (table 11). Although globally significant (X2(12) = 117.753; P < 10–13), the test is not significant when restricted to site 2 only (X2(4) = 3.19; P = 0.53). Thus, there seem to be no transients in the central Chesapeake region! A directional z statistic could be calculated in the same manner as with TEST 3.SR. A test of memory Animals may make decisions of movement based on the knowledge of previously visited sites. Hestbeck et al. (1992) have identified this phenomenon in the election of wintering sites by Canada geese. This "memory effect", which is probably common in many long–lived species, is a violation of the assumption of the JMV model of the sort that TEST 3G examines: it leads to different behaviour for animals belonging to the same sample depending on which sites they had visited previously. This memory effect is detectable by the specific test of

…

Nss s

+ +

…

Lss s

memory, called WBWA, proposed by Pradel et al. (2003). We will show in the next section that TEST WBWA presented in table 12 is a subcomponent of TEST 3G. Applied to the Canada geese (table 13), TEST WBWA confirms the very strong role of memory in the movements of these birds (X2(20) = 472.86; P < 10–14): the overdispersion factor calculated for this test alone is 472.86/20 = 23.6, much higher than that of the overall TEST 3G (7.27) or that of the test for transience (9.81). In order to more specifically target the departures which are expected under the memory effect along the diagonal, an alternative statistic to the Pearson Chi–square can be used. One such possibility is Cohen’s kappa (Cohen, 1960), which has a standardized normal N(0,1) distribution. The individual kappa tests can be combined in the same manner as the z tests of section 1.2 to get an overall test of memory. They are added and their sum is divided by the square root of the number of components p (Gimenez, 2003):

1 p

∑κ i =1

Pradel et al.

198

Table 13. Results of TEST WBWA for the Canada goose data. The table gives the Pearson chi–square statistic (X 2) and the corresponding P–value (P) as well as the number of degrees of freedom after pooling (df): Oc. Occasion; S. Site. Tabla 13. Resultados del TEST WBWA para los datos de la barnacla canadiense. Las tablas presentan la ji–cuadrado de Pearson (X2) y su correspondiente valor P (P), así como el número de grados de libertad (df) tras la reunión: Oc. Ocasión; S. Localidad.

X2 19.59

P 0.000

Table 14. Directional test of memory applied to the Canada goose data. This test is distributed as N(0,1) and looks at a consistent excess (or lack) on the diagonal. Tabla 14. Test direccional de memoria aplicado a los datos de la barnacla canadiense. Este test se distribuye como N(0,1) y demuestra un exceso (o una falta) consistente en la diagonal.

3.87

0.00

4.33

0.00

4.33

0.00

7.22

0.00

1.81

0.04

4.22

0.00

6.36

0.00

37.87

0.000

8.59

4.49

0.034

5.98

0.00

6.47

0.00

80.59

0.000

1.01

0.21

4.44

0.00

98.76

0.000

0.81

0.369

27.71

0.000

53.69

0.000

25.29

0.000

43.66

0.000

50.93

0.000

29.48

0.000

This directional test of memory confirms the strong positive correlation between the previous and the next sites of observation of the Canada geese ( = 16.92; P < 10–13). There is as yet no simple alternative model associated to TEST WBWA (like the 2–age model associated to TEST 3G.SR), but the model that accounts for the location at i–1 in the transitions (Brownie et al., 1993) will probably treat most of the "memory effect". Unfortunately, this model cannot be fitted in the framework of multistate models for the full data as it belongs to a more general family of capture–recapture models (Pradel, 2005). The full decomposition of TEST 3G TEST 3G.SR and TEST WBWA are two independent subcomponents of TEST 3G but they do not make up for TEST 3G alone. We illustrate here how the original table of TEST 3G is partitioned to isolate these specific tests with the example of the Canada geese encountered at occasion 2 and site 1 (table 8). This procedure of partitioning is very general with contingency tables (Everitt, 1977). In a first step, table 8 is replaced with two tables. This step consists in setting aside the previously

captured geese (first 3 rows) in a separate table and then confronting them all pooled together against the newly captured geese in a second table. (see 2 contingency tables of step 1 below) Then, within each one of these two new tables, the never–seen–again geese (last column) are set aside leading to four new tables, one of which is the component of TEST 3.SR relative to occasion 2 and site 1. In this step, the timing of first reencounters is compared among the different rows in a first table, and then the total of reencounters is compared with the number of never–seen–again animals among the same rows in a second table. (see 4 contingency tables of step 2 below) Eventually, the first of the four previous tables, which summarizes the first reencounters of the previously encountered individuals, is replaced with four tables: one contains the reencounters made at site 1, one those made at site 2, one those made at site 3, and the last one contrasts the number of reencounters at each site depending on the site of most recent encounter (in rows). This last one is the component of TEST WBWA relative to occasion 2 and site 1. Of the 7 tables obtained at this stage (the last three tables of step 2 remain unchanged in the last step), only two belong to the specific tests described in the previous sections. (see 7 contingency tables of step 3 below)

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Animal Biodiversity and Conservation 28.2 (2005)

Step1. Two contingency tables. Paso 1. Dos tablas de contingencia.

128

390

124

122

920

101

184

The remaining tables constitute together TEST 3G.Sm of U–CARE version 2.2. To summarize, TEST 3G is made up of 3 subcomponents: TEST 3G.SR, which tests specifically for transients; TEST WBWA, which aims at detecting a memory effect, and the complementary composite TEST 3G.Sm. To be caught or not should have no effect (TEST M) The second main component of the JMV model GOF test, called TEST M, contrasts the animals

not caught at a given occasion —yet known to be alive— to those caught at the same occasion. Again, the JMV assumptions imply that there should be no difference between two animals when one is caught and the other is not. However, the exact location of the animals not encountered remains unknown. This is the most far–reaching difference with the one–site context and the reason why the multistate JMV model has not all the nice properties of the single–site CJS model (see Section "An imperfect segrega-

Step 2. Four contingency tables. Paso 2. Cuatro tablas de contingencia.

111

128

390

124

122

101

(TEST 3G.SR) 814

920

164

184

Pradel et al.

200

Step 3. Seven contingency tables. Paso 3. Siete tablas de contingencia.

(TEST WBWA)

102

111

128

(TEST 3G.SR)

814

920

164

184

390

124

122

101

Table 15. Component of TEST M relative to date 2 for the Canada geese. The first three rows correspond to the geese not observed at occasion 2 that were released at sites 1, 2 and 3 respectively at date 1. The last three rows are for the geese observed at date 2 on the three sites in the same order. The columns correspond to the particulars (site within date) of the next reencounter. Tabla 15. Componente del TEST M, relativa a la fecha 2, para la barnacla canadiense. Las tres primeras filas corresponden a los gansos no observados en la ocasi贸n 2, que fueron liberados en la fecha 1 en las localidades 1, 2 y 3, respectivamente. Las tres 煤ltimas filas corresponden a los gansos observados en la fecha 2 en las tres localidades en el mismo orden. Las columnas corresponden a los datos (localidad en una fecha determinada) del siguiente reencuentro.

Time (j) and location (v) of first reencounter j= v=

3 1

6 2

last seen at site 1

158

491

134

149

159

869

63 335

164

currently seen at site

101 158

201

Animal Biodiversity and Conservation 28.2 (2005)

Table 16. Results of TEST M for the Canada geese. This test cannot be computed at the first occasion or less than 2 occasions before the end of the study: Oc. Occasion. Tabla 16. Resultados del TEST M para la barnacla canadiense. Este test no puede calcularse a la primera ocasión, o a menos de dos ocasiones antes del final del estudio.

24.119

0.044

36.037

0.007

23.098

0.006

tion of information between 'estimation' and 'test of assumptions'"). As regards the tests, because of the uncertain location of the not–encountered individuals, homogeneity tests of contingency tables are now replaced with more complex tests of mixtures. Let us consider the table retained in U–CARE version 2.0 for the component of TEST M relative to date 2 for the Canada geese (table 15). The first three rows correspond to the geese that were last released at date 1 at sites 1, 2 and 3 respectively; the last three rows to the geese currently released at the same sites. The columns correspond to the timing and place of the next reencounter. In this table, the first three rows do not play the same role as the last ones; they should be approximate linear combinations of the last ones. The rationale for this is as follows: the animals not observed at date 2 may have moved since they were last released; hence, their current location can be any one of the three sites. These animals are thus a mixture of animals in the different sites in unknown proportions. In accordance with the model assumptions, those at site 1 (resp. at site 2 and 3), i.e. on rows 1 (resp. 2 and 3), should behave like those caught and released at site 1 (resp. at site 2 and 3), i.e. on rows 4 (resp. 5 and 6). The results concerning the Canada geese (table 16) are significant (overall test: 2(41) = 83.254; P < 10–3), although not as strong as those from TEST 3G and its subcomponents. It is difficult to know the reason for departure simply by examining a complex table like table 15, even if the expected numbers were given. A suitable partitioning is again the key to a better understanding. A test of short–term trap–dependence Drawing a parallel with the CJS GOF test, a test for trap–dependence can be considered. The immediate question that arises is what trap–dependence means when there are several sites. Once an animal has been caught, is it expected

Table 17. TEST M.ITEC: this subcomponent of TEST M tests for trap–dependence. In case of local trap–happiness, observed numbers should exceed expected numbers where there are "++" signs. If trap–dependence is present even if the animal has moved to another site (or more presumably state), the excesses will show also where there are "+" signs: Sns1. Seen next occasion at site 1; Sns2. Seen next occasion at site 2; Sls1. Sen later at site 1; Sls2. Seen later at site 2; Lss1. Last seen at site 1; Lss2. Last seen at site 2. Cs1. Currently seen at site 1; Cs2. Currently seen at site 2. Tabla 17. TEST M.ITEC: este subcomponente de los TEST M está destinado a demostrar la dependencia a la trampa. En el caso de una habituación a la trampa local, las cifras observadas deberían exceder a las esperadas en los lugares en los que existen signos "++". Si la dependencia a la trampa existe incluso si el animal se ha desplazado a otro lugar (o más probablemente a otro estado), los excesos se pondrán también de manifiesto donde existan signos "+": Sns1. Visto la siguiente ocasión en la localidad 1; Sns2. Visto la siguiente ocasión en la localidad 2; Sls1. Visto más tarde en la localidad 1; Sls2. Visto más tarde en la localidad 2; Lss1. Visto la última vez en la localidad 1; Lss2. Visto la última vez en la localidad 2. Cs1. Visto habiatualmente en la localidad 1; Cs2. Viso habitualmente en la localidad 2.

Sns1 Sns2

…

Sls1

Sls2

...

Lss1 Lss2 … Cs1

Cs2

…

to change its behaviour the next time only if it remains at the same site, or should it change even if it moves to a different site? Presumably, the first option is more reasonable. However, when dealing with states instead of sites, the second option may be better: the animal will be faced with the same trap whatever its state. The table for testing for an immediate trap response is the same in both cases (table 17). It is the region of the table where departure is expected tath differs. In the second case, the whole lower left quarter of the table should exhibit high (resp. low) numbers observed in case of trap–happi-

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202

Table 18. Component of TEST M.ITEC for the Canada geese relative to date 3. There is evidence of local trap–happiness with the number in bold greater than expected: Sdi–sj. Seen at day i and site j. Tabla 18. Componente del TEST M.ITEC para la barnacla canadiense, relativa a la fecha 3. Existen pruebas de una adicción a la trampa local, siendo las cifras en negrita mayores de lo esperado: Sdi– sj. Visto el día i en la localidad j.

Sd4–s2

Sd4–s3

Sd>4–s1

Sd>4–s2

Last seen at site 1

Sd4–s1 162

Sd>4–s3 4

Last seen at site 2

427

292

Last seen at site 3

Currently at site 1

564

200

202

162

Currently at site 2

125

1,017

471

Currently at site 3

178

ness (resp. trap–shyness); in the first case, only the diagonal in the same quarter is expected to be affected. The table corresponding to date 3 for the Canada geese is given in table 18. There is evidence of local trap–happiness in the geese with a systematic excess on the diagonal of the lower left quarter of the table (table 18) and a globally highly significant test (X2(27) = 68.177; P < 10–4) (table 19). Discussion Goodness–of–fit testing is not the most popular part of a capture–recapture analysis, probably because it is neither automatic nor very appealing. Although, some automatic procedures, such as the bootstrap procedure built in MARK (White, 2001), are available, they have their limitations (White, 2002) and above all do not suggest what may be wrong when a model is rejected. On the other hand, optimal goodness–of–fit procedures exist only for a very

Table 19. Results of TEST M.ITEC for the Canada geese: Oc. Occasion.

limited number of models, and have long been entirely missing for the multistate models. However, we believe that such procedures can be made more user–friendly and interpretable than they currently are, and that they have a great potential in helping understand capture–recapture data. There is certainly a lot of work yet to be done in this direction, but we have tried to show in this paper that there is already a lot to be learned from them. We have shown in particular that the goodness–of–fit multistate test of the JMV model as proposed by Pradel et al. (2003) can be partitioned in subcomponents directly related to some frequent violations of the assumptions (transience, trap–dependence, memory). Some

Table 20. Overdispersion factor calculated for different components (straight police) and subcomponents (italics) of the JMV and CJS goodness–of–fit tests. Tabla 20. Factor de sobredispersión calculado para diferentes components (redondilla) y subcomponents (cursiva) de los test de bondad de ajuste JMV (primeras dos columnas) y CJS (últimas dos columnas).

Tabla 19. Resultados del TEST M.ITEC para la barnacla canadiense: Oc. Ocasión.

JMV 3G

3GSR

CJS 7.25 9.8

14.242

0.114

WBWA

30.837

0.000

2.03

23.098

0.006

M.ITEC

2.53

13.5

3SR

2.CT

23.64

203

Animal Biodiversity and Conservation 28.2 (2005)

of these effects can be treated more (transience) or less (trap–dependence) easily by fitting appropriate multistate models; others (memory) call for entirely different models. The fit of the alternative model to be used in case of transience, namely the generalized JMV model with two age classes on survival, is itself exactly testable by deducting TEST 3.SR from the overall goodness–of–fit test of JMV. Thus, this is a new model for which an optimal goodness–of–fit test is available. It would be interesting to examine whether a model with an optimal goodness–of–fit test could be identified as well in case of trap– dependence. The memory effect is a more difficult challenge. If this effect is strong, like for the Canada geese, all multistate models are invalidated. However, if it is weak, it can be kept out of the structural part of the model provided an overdispersion factor (the ratio of the X2 statistic to its number of degrees of freedom) calculated from the goodness–of–fit tests is used in the analysis. An overdispersion factor can be used more generally whenever there is no obvious structural explanation for a lack of fit. Suitably partitioned goodness–of–fit tests are thus a more general tool than initially apparent for a correct assessment of the situation. Beyond their purely technical usage, partitioned goodness–of–fit tests can serve to unveil some biological information. For instance, the intensity of transit is likely related to dispersion (Perret et al., 2003; Cam et al., 2004); heterogeneity of capture (a test of which has yet to be incorporated within a general goodness–of–fit test) may be a reflection of the intensity of social structuration; the role of memory helps understand how the organism apprehends its environment. This potential has yet to be fully exploited. The analysis of the Canada goose data set that we have used throughout this paper yields examples of the insight gained from the different components and subcomponents of the goodness–of–fit tests. A simple way to rank the relative strength of different effects is to calculate an overdispersion factor per components or subcomponents of the CJS and the JMV goodness–of–fit tests (table 20). A first remark is that the corresponding subcomponents for transients, 3.SR and 3G.SR, and particularly trap–dependence, 2.CT and M.ITEC, present higher overdispersion coefficients in the one–site than in the multisite context. Obviously, taking account of the location has removed part of the heterogeneity. This is not surprising as encounter probabilities tend to be higher at some sites and at the same time the geese exhibit a high fidelity to their wintering sites; hence, the same individual geese tend to be consistently reencountered. The examination of the subcomponents of the multisite TEST 3G reveals in turn that memory is by far the most important cause of departure confirming the need for specific generalized models (Hestbeck et al., 1991; Brownie et al., 1993). Going through the occasion– and site–specific tables, we have also gained along the way new insights into the data: transit seems to affect only the peripheral sites and

trap dependence is more precisely local trap happiness. All this information has been obtained without fitting a single model so that, at the onset of modelling, we know for instance that a model with transients on the two peripheral sites is appropriate. The risk of overfitting, which must be kept in mind, is limited here by the consistency of the effects through several occasions. Another safeguard is provided by the use of even more specialized tests more precisely targeting the alternative of interest. The z–score tests of transience and the Cohen’s kappa test of memory are two examples, but more can be developed, in particular for the detection of trap–dependence. Although the Arnason–Schwarz model (AS) is generally considered as the reference for multistate analyses, we have not examined it specifically. This is because there is currently no specific goodness– of–fit test for it. The best approach is to treat the AS model as a particularization of the JMV model. After assessing the fit of JMV, JMV and AS can be fitted using program M–SURGE (Choquet et al., 2004) and the two models compared with the AIC criterion (possibly modified to incorporate an overdispersion factor). However, there is no more a priori reason to fit the AS model than any other multistate model. At present, the most urgent need is the study of the statistical properties of the new tests, notably the tests of mixture. For instance, in presence of sparse data, there is no equivalent to the Fisher’s exact test. Another very promising extension is the use of the multistate tests with recovery data. This is possible because recoveries can be presented as multistate data with two states: ‘alive’ and ‘dead’ (Lebreton et al., 1999). However, as the state ‘dead’ is absorbing, the tests have first to be modified accordingly. There are more generally various potential original applications of the non–parametric tests presented in this paper (see for instance Gauthier et al. (2001) for seasonal trap–dependence). We believe that these tests should no longer be considered only as the necessary routine first step of a capture–recapture analysis but also as an important part of the analysis itself, contributing in ways that the parametric modelling cannot always do to the understanding of the data. References Arnason, A. N., 1972. Parameter estimates from mark–recapture experiments on two populations subject to migration and death. Researches on Population Ecology, 13: 97–113. – 1973. The estimation of population size, migration rates and survival in a stratified population. Researches on Population Ecology, 15: 1–8. Balser, J. P., 1984. confidence interval estimation and tests for temporary outmigration in tag– recapture studies. Doctoral dissertation, Cornell Univ., Ithaca, New York. Begon, M., 1983. Abuses of mathematical techniques in ecology: applications of Jolly’s cap-

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geese. Ecology, 72: 523–533. Lebreton, J.–D., Alméras, T. & Pradel, R., 1999. Competing events, mixture of information and multistrata recapture models. Bird Study, 46: 39–46. Lebreton, J.–D., Burnham, K. P., Clobert, J. & Anderson, D. R., 1992. Modeling survival and testing biological hypotheses using marked animals: A unified approach with case studies. Ecological Monographs, 62: 67–118. Lebreton, J.–D. & Pradel, R., 2002. Multistate recapture models: modelling incomplete individual histories. Journal of Applied Statistics, 29: 353– 369. McCullagh, P. & Nelder, J. A., 1989. Generalized linear models, second edition. Chapman and Hall, New York, USA. Nichols, J. D., Hines, J. E., Pollock, K. H., Hinz, R. L. & Link, W. A., 1994. Estimating breeding proportions and testing hypotheses about costs of reproduction with capture–recapture data. Ecology, 75: 2052–2065. Perret, N., Pradel, R., Miaud, C., Grolet, O. & Joly, P., 2003. Transience, dispersal, and survival rates in newt patchy populations. Journal of Animal Ecology, 72: 567–575. Pollock, K. H., Hines, J. E. & Nichols, J. D., 1985. Goodness–of–fit tests for open capture–recapture models. Biometrics, 41: 399–410. Pradel, R., 1993. Flexibility in Survival analysis from recapture data: Handling trap–dependence. Pages 29–37 in Lebreton & North, editors. Marked individuals in the study of bird population. Birkhaüser Verlag, Basel, Switzerland. Pradel, R., Hines, J. E., Lebreton, J.–D. & Nichols, J. D., 1997. Capture–recapture survival models taking account of transients. Biometrics, 53: 60–72. Pradel, R. & Lebreton, J.–D., 1999. Comparison of different approaches to the study of local recruitment of breeders. Bird Study, 46: 74–81. Pradel, R., Wintrebert, C. M. A. & Gimenez, O., 2003. A proposal for a goodness–of–fit test to the Arnason–Schwarz multisite capture–recapture model. Biometrics, 59: 43–53. Robson, D. S., 1969. Mark–recapture methods of population estimation. BU–168, Cornell Univ. White, G. C., 2001. Advanced features of Program Mark. Pages 368–377 in Field, Warren & Sievert, editors. Wildlife, Land, and People: Priorities for the 21st Century. Proceedings of the Second International Wildlife Management Congress. The Wildlife Society, Bethesda, Maryland, USA. – 2002. Discussion comments on: the use of auxiliary variables in capture–recapture modelling. An overview. Journal of Applied Statistics, 29: 103–106.

Animal Biodiversity and Conservation 28.2 (2005)

Animal Biodiversity and Conservation

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Animal Biodiversity and Conservation (abans Miscel·lània Zoològica) és una revista interdisciplinària publicada, des de 1958, pel Museu de Zoologia de Bar celona. 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.bcn.es/ABC, de manera que permet una difusió mundial dels seus articles. Tots els manuscrits són revisats per l'editor execu tiu, 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.

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Animal Biodiversity and Conservation 28.2 (2005)

Animal Biodiversity and Conservation Animal Biodiversity and Conservation (antes Miscel·lània Zoològica) es una revista inter disciplinar, 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 genéti ca) 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 re gistrada en todas las bases de datos importantes y además está disponible gratuitamente en internet en http://www.bcn.es/ABC, 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 garan tizar su calidad. El proceso de revisión es rápido y constructivo, y se realiza vía correo electrónico siem pre 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 propie dad de la revista. Ésta se reserva los derechos de autor, y ninguna parte del trabajo podrá ser reprodu cida sin citar su procedencia.

Normas de publicación Los trabajos se enviarán preferentemente de forma electrónica (abc@mail.bcn.es). El formato preferido es un documento Rich Text Format (RTF) o DOC, que incluya las figuras (TIF). Si se opta por la versión im presa, 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 necesa rios 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 de berán enviar a la Redacción una copia impresa de la versión final junto con un disquete del manuscrito preparado con un procesador de textos e indicando el programa utilizado (preferiblemente Word). Las pruebas de imprenta enviadas a los autores deberán ISSN: 1578–665X

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remitirse corregidas al Consejo Editor en el plazo máximo de 10 días. Los gastos debidos a modifica ciones sustanciales en las pruebas de imprenta, intro ducidas 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 ningu no, 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 consen timiento del editor. Nombre del autor o autores. Abstract en inglés de 12 líneas mecanografiadas (860 espacios como máximo) y que exprese la esencia del manuscrito (introducción, material, métodos, resulta dos y discusión). Se evitarán las especulaciones y las citas bibliográficas. Irá encabezado 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. © 2005 Museu de Ciències Naturals

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 ante cedentes del tema tratado, así como de los objetivos del trabajo. Material y métodos. Incluirá la información referente a las especies estudiadas, aparatos utilizados, me todología de estudio y análisis de los datos y zona de estudio. Resultados. En esta sección se presentarán úni camente los datos obtenidos que no hayan sido publicados previamente. Discusión. Se discutirán los resultados y se compara rán 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 spe cies 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 conserva tion 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 vari ation 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)..." 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 ajustarse a 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. Si las fotografías son en color, el coste de su publicación irá a cargo de los autores. 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.

Animal Biodiversity and Conservation 28.2 (2005)

Animal Biodiversity and Conservation

Manuscripts

Animal Biodiversity and Conservation (formerly Miscel·lània Zoològica) is an interdisciplinary journal which has been published by the Zoological Mu seum 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. Animal Biodiversity and Conservation is registered in all principal data bases and is freely available online at http://www.bcn.es/ABC, thus assuring world–wide access to articles published therein. All manuscripts are screened by the Executive Edi tor, 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 prop erty of the journal, which reserves copyright, and no published material may be reproduced without quoting its origin.

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 con tributions 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 writ ten 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.

Information for authors Electronic submission of papers is encouraged (abc@mail.bcn.es). The preferred format is a do cument 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 on 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 especify that the authors follow current norms on the protection of animal species and that they have obtained all relevant 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. 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. The first author will receive 50 reprints free of charge and an electronic version of the article in PDF format.

ISSN: 1578–665X

Formatting of articles Title. The title must be concise but as informative as possible. Part numbers (I, II, III, etc.) 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. 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 bibliogra phy 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 spe cies 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 conserva tion 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 vari ation 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 Wemmer (1998)...", "...has been defined by Robinson & Redford (1991)...", "...the prospections that have been carried out (Begon et al., 1999)..." 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, num bered consecutively in Arabic numerals (1, 2, 3, etc.) and with reference in the text. Glossy print photographs, if essential, may be included. Colour photographs may be published but its publication will be charged to authors. 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 Refe rences) 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.

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Animal Biodiversity and Conservation 28.2 (2005)

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Animal Biodiversity and Conservation 28.2 (2005)

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Animal Biodiversity and Conservation 28.2 (2005)

Arxius de Miscel·lània Zoològica vol. 2 (2004) 2004 Museu de Ciències Naturals de la Ciutadella ISSN: 1698–0476

Índex/Índice/Contents Fuentes, M. V., Sáez, S., Trelis, M., Muñoz–Antoli, C. & Esteban, J. G., 2004. The helminth community of Apodemus sylvaticus (Rodentia, Muridae) in the Sierra de Gredos (Spain): Arxius de Miscel·lània Zoològica, 2: 1–6. Abstract The helminth community of Apodemus sylvaticus (Rodentia, Muridae) in the Sierra de Gredos (Spain).— The Spanish mountain range of Gredos was included in the studies conducted on the Iberian peninsula to investigate helminth fauna of small mammals. The helminth community of the wood mouse, Apodemus sylvaticus (Rodentia, Muridae), was analysed. Qualitatively, 13 helminth species were detected: Plagiorchis sp. I and Plagiorchis sp. II (Trematoda); Taenia parva larvae, T. martis larvae, T. taeniaeformis larvae, Rodentolepis straminea and R. fraterna (Cestoda); and Trichuris muris, Heligmosomoides polygyrus, Syphacia stroma, S. frederici, Aspiculuris tetraptera and Rictularia proni (Nematoda). Quantitatively, the highest prevalence (65.0%) and the mean abundance (36.9%) of H. polygyrus stand out. In comparison with the other mountain ranges studied, analysis of the global results demonstrates that the helminth fauna of the host species studied is diverse despite the adverse climatic conditions. This could be related to both the particular ecological characteristics and the appropriate state of preservation of this ecosystem. Key words: Helminths, Apodemus sylvaticus, Rodentia, Muridae, Sierra de Gredos, Spain. Bros, V., 2004. Mol·luscs terrestres i d’aigua dolça de la serra de Collserola (Barcelona, NE península Ibèrica). Arxius de Miscel·lània Zoològica, 2: 7–44. Abstract Land and freshwater molluscs of the Collserola mountains (Barcelona, NE Iberian peninsula).— A malacological survey was made taking into account all the 91 1 x 1 km UTM grid squares that cover the area of the Collserola mountains, Barcelona (north-eastern Iberian peninsula). From 108 sampled localities, 1,261 records of molluscs were obtained. A total of 73 species were identified: seven freshwater species, 11 slugs, and 55 land molluscs. The first discussions on the results and a draft of the initial conclusions are shown. Information on dominant gastropod habitat preference in the ecosystems of Collserola is also provided. Key words: Mollusca, Continental molluscs, Collserola Park, Faunistics, Monitoring, Bioindicators.

All works are licensed under a Creative Commons Attribution–NonCommercial 3.0 License Web: http://www.bcn.es/arxiusMZ

Animal Biodiversity and Conservation 28.2 (2005)

Agradecimiento a los asesores Our grateful thanks to the referees

El Editor ejecutivo, los Editores, el Consejo editor y el Consejo asesor quieren agradecer a todos los asesores su incalculable ayuda en la revisión de los artículos sometidos a Animal Biodiversity and Conservation durante el período 2003 a 2005: The Executive Editor, the Editors, the Editorial Board and the Advisory Board wish to thank all the referees for their invaluable help in reviewing articles submitted to Animal Biodiversity and Conservation for the period 2003 to 2005: Allen, C. Alonso, J. C. Arnason, N. Atlegrim, O. Baillie, S. Bairlein, F. Baker, A. S. Balke, M. Bareth, C. Bellés, X. Bewster–Wingard, G. L. Biström, O. Borges, S. Brooks, S. J. Brown, C. R. Burnham, K. P. Cam, E. Cameron, R. A. D. Cárdenas Talaverón, A. M. Carrascal, L. M. Carroll, J. P. Casals, F. Castro, F. Cobo, F. Conroy, M. J. Cooch, E. Cooper, W. E. Cuervo, J. J. Dávalos, L. M. Dhondt, A. Díaz, M. Doherty, P. Eckert, K. L. Estrada, A. Fagan, W. Ferguson L. M. Fernández–Haeger, J. Fernández–Juricic, E.

Ferrer, M. Fiedler, K. Francis, C. M. Fresneda, X. García–Barros, E. Gómez, B. J. González, J. Grim, E. Grossman, G. D. Hines, J. Hopkins, B. Horrocks, H. Illera Cobo, J. C. Jordana, R. Juan, C. Kitahara, M. Klompen, H. Konopacka, A. Krell, F.–T. Laayouni, H. Lebreton, J.–D. Lee, D. C. Lhonoré, J. Machado, A. MacCall, A. MacKenzie, D. Manfrín, L. Maranhao, P. Marco, A. Márquez, F. Martikainen, P. Martin, B. Mendes, L. F. Metcalfe, N. Micó Balaguer, E. Miller, K. B. Mínguez, E. Mischis, C. C. de

Montreuil, O. Munguira, L. M. Negre, B. Nichols, J. D. Nogales, M. Nomakuchi, S. Novoa, F. Ortuño Hernández, V. Palmer Vidal, M. Pérez–Enciso, M. Pérez–Tris, J. Pollock, K. H. Pons, J. Pons, P. Prat Baella, F. Pretus, J. M. Real, R. Rosser, A. Santos Maroño, M. Schmitz García, M. F. Schwarz, C. J. Seoaane Pinilla, J. Serrano Marino, J. Simón Benito, J. C. Teisaire, E. S. Tella, J. L. Templado, J. Thibaud, J. M. Thomson, D. L. Valido, A. Van Veller, M. G. P. Varga, Z. Watanabe, M. White, G. C. Wiklund, C. Yildirim, Z. Zardoya, R. Zhang, Z. G.

Les cites o els abstracts dels articles d’Animal Biodiversity and Conservation es resenyen a / Las citas o los abstracts de los artículos de Animal Biodiversity and Conservation se mencionan en / Animal Biodiversity and Conservation 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, Directory of Open Acces Journals (DOAJ), Ecological Abstracts, Ecology Abstracts, Entomology Abstracts, Environmental Abstracts, Environmental Periodical Bibliography, Genetic Abstracts, Geographical Abstracts, índex de Sumaris Electrònics del Consorci de Biblioteques de Catalunya, 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, Red de Revistas Científicas Españolas (REVICIEN), Referatirnyi Zhurnal, Science Abstracts, Serials Directory, Ulrich’s International Periodical Directory, Zoological Records.

Índex / Índice / Contents Animal Biodiversity and Conservation 28.2 (2005) ISSN 1578–665X

101–119 Carrascal, L. M. & Palomino, D. Preferencias de hábitat, densidad y diversidad de las comunidades de aves en Tenerife (Islas Canarias) 121–130 Waite, T. A., Vucetich, J., Saurer, T., Kroninger, M., Vaughn, E., Field, K. & Ibargüen, S. Minimizing extinction risk through genetic rescue 131–136 Martínez–Abraín, A., Oro, D., Belenguer, R., Ferrís, V. & Velasco, V. Long–term changes in species richness in a small Mediterranean archipelago bird– breeding community 137–147 Komonen, A. & Kouki, J. Occurrence and abundance of fungus– dwelling beetles (Ciidae) in boreal forests and clearcuts: habitat associations at two spatial scales 149–157 Simón Benito, J. C., Espantaleón, D. & García– Barros, E. Stachorutes cabagnerensis n. sp., Collembola (Neanuridae) from Central Spain, and a preliminary approach to phylogeny of genus

159–168 Martin, C. S. Jeffers, J., Godley, B. J. The status of marine turtles in Montserrat (Eastern Caribbean) 169–179 Hilaluddin, Kaul, R. & Ghose, D. Conservation implications of wild animal biomass extractions in Northeast India 181–188 Örstan, A., Pearce, T. A., Welter–Schultes, F. Land snail diversity in a threatened limestone district near Istanbul, Turkey 189–204 Pradel, R., Gimenez, O. & Lebreton, J.–D. Principles and interest of GOF tests for multistates capture–recapture models 205–206 A la memoria de Xavier Domingo–Roura (1964–2005) 207–208 In memoriam: Xavier Domingo– Roure (1964–2005) IX Abstracts del volumen 2 (2004) de Arxius de Miscel·lània Zoològica Abstracts of volume 2 (2004) of Arxius de Miscel·lània Zoològica