B&AH nº 26 [2012] by Marcos Pérez de Tudela Gely

Revista EspaĂąola de HerpetologĂa

Journal of the Spanish Herpetological Society (AHE) Volume 26 (2012) http://bah.herpetologica.es bah@herpetologica.org

BASIC & APPLIED HERPETOLOGY REVISTA ESPAÑOLA DE HERPETOLOGÍA

Spanish Herpetological Society (AHE) President: Juan Manuel Pleguezuelos Gómez Vice-President: Jaime Bosch Pérez General Secretary: Miguel Ángel Carretero Fernández Vice-General Secretary: José Antonio Mateo Miras Vocals: Enrique Ayllón López (Management) César Ayres Fernández (Conservation) Francisco Javier Diego Rasilla (Web page and promotion) Andrés Egea Serrano (Editor, Boletín de la AHE) Gustavo A. Llorente Cabrera (Atlas) Adolfo Marco Llorente (Marine turtles) Albert Montori Faura (Atlas) Manuel E. Ortiz Santaliestra (Editor, Basic & Applied Herpetology) Ana Perera Leg (Editor, Basic & Applied Herpetology) Alex Richter Boix (Editor, Boletín de la AHE) Xavier Santos Santiró (Editor, Boletín de la AHE & Treasurer) Daniel Villero Pi (Atlas)

Basic & Applied Herpetology (Editors) Manuel E. Ortiz Santaliestra (Amphibians) Instituto de Investigación en Recursos Cinegéticos (IREC). CSIC-UCLM-JCCM Ronda de Toledo, s/n 13071 Ciudad Real (Spain) manuele.ortiz@uclm.es

Ana Perera Leg (Reptiles) CIBIO-Universidade do Porto Campus Agrário de Vairão Rua Padre Armando Quintas-Castro 4485-661 Vairão (Portugal) perera@cibio.up.pt

Asociación Herpetológica Española Museo Nacional de Ciencias Naturales C/ José Gutiérrez Abascal, 2 28006 Madrid http://www.herpetologica.es

ISSN 2255 - 1468 Impresión: igrafic. Url: www.igrafic.com

Depósito Legal: M-38882-2012 Maquetación: Marcos Pérez de Tudela. Url: www.marcos-pdt.com

BASIC & APPLIED HERPETOLOGy REVISTA ESPAÑOLA DE HERPETOLOGÍA

CONTENTS Volume 26 (2012)

Reviews GUEST CONTRIBUTION: Herpetofauna and roads: a review V.J. Colino-Rabanal, M. Lizana

Pag. 5

Research papers Quantification of road mortality for amphibians and reptiles in Hoces del Alto Ebro y Rudrón Natural Park in 2005 F. Martínez-Freiría, J.C. Brito

Natural fluctuations in a stream dwelling newt as a result of extreme rainfall: a 21-year survey of a Calotriton asper population A. Montori, A. Richter-Boix, M. Franch, X. Santos, N. Garriga, G.A. Llorente

Distribution review, habitat suitability and conservation of the endangered and endemic Moroccan spadefoot toad (Pelobates varaldii) P. de Pous, W. Beukema, D. Dingemans, D. Donaire, P. Geniez, E.H. El Mouden

Ecological aspects and effects of forestry management on a population of Hermann’s tortoise (Testudo hermanni hermanni) in Catalonia (Spain) M. Casamitjana, J.C. Loaiza, N. Simon, P. Frigola

Measuring body temperatures in small lacertids: Infrared vs. contact thermometers M.A. Carretero

Synopsis of the helminth communities of the lacertid lizards from the Balearic and Canary Islands V. Roca, F. Jorge, M.A. Carretero

107

Cover illustration: Female Lataste's viper (Vipera latastei), Sedano Valley, Burgos, Spain (see article by Martínez-Freiría & Brito in this volume). Author: Fernando Martínez-Freiría.

BASIC & APPLIED HERPETOLOGy REVISTA ESPAÑOLA DE HERPETOLOGÍA

CONTENIDOS Volumen 26 (2012)

Revisiones ARTíCULO INVITADO: Herpetofauna y carreteras: revisión V.J. Colino-Rabanal, M. Lizana

Pag. 5

Artículos de investigación Cuantificación de la mortalidad de anfibios y reptiles en las carreteras del Parque Natural de las Hoces del Alto Ebro y Rudrón en 2005 F. Martínez-Freiría, J.C. Brito

Fluctuaciones naturales en un tritón de arroyo como resultado de la lluvia extrema: seguimiento de una población de Calotriton asper durante 21 años A. Montori, A. Richter-Boix, M. Franch, X. Santos, N. Garriga, G.A. Llorente

Revisión de la distribución, idoneidad del hábitat y conservación del amenazado endemismo sapo de espuelas marroquí (Pelobates varaldii) P. de Pous, W. Beukema, D. Dingemans, D. Donaire, P. Geniez, E.H. El Mouden

Aspectos ecológicos y efectos del manejo forestal en una población de tortuga mediterránea (Testudo hermanni hermanni) en Cataluña (España) M. Casamitjana, J.C. Loaiza, N. Simon, P. Frigola

Notas adicionales sobre la dieta de Japalura swinhonis (Agamidae) en el suroeste de Taiwán, con comentarios acerca de su solapamiento trófico con la especie simpátrica Anolis sagrei (Polychrotidae) G. Norval, S.-C. Huang, J.-J. Mao, S.R. Goldberg, K. Slater

Medida de la temperatura corporal en pequeños lacértidos: Termómetros de infrarrojos vs. termómetros de contacto M.A. Carretero

Sinopsis de las comunidades helmínticas de los lagartos de las Islas Baleares y Canarias V. Roca, F. Jorge, M.A. Carretero

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Ilustración de portada: Hembra de víbora hocicuda (Vipera latastei), Valle de Sedano, Burgos, España (véase artículo de Martínez-Freiría & Brito en este volumen). Autor: Fernando Martínez-Freiría.

Guest contribution

Basic and Applied Herpetology 26 (2012): 5-31

Herpetofauna and roads: a review Víctor J. Colino-Rabanal*, Miguel Lizana Area of Zoology, Department of Animal Biology, Ecology, Soil Science, Parasitology, and Agrochemistry, University of Salamanca, Salamanca, Spain * Correspondence: Department of Animal Biology, Ecology, Soil Science, Parasitology, and Agrochemistry, Campus Miguel de Unamuno 37071 Salamanca, Spain. E-mail: vcolino@usal.es

Received: 14 October 2012; received in revised form: 4 December 2012; accepted: 18 December 2012.

Roads and traffic are tightly related to some of the mains threats for biodiversity. Road network affects wildlife populations due to, among other effects, partial occupation and transformation of landscape, alteration of surrounding habitat, dispersal of physicochemical pollutants, fragmentation and loss of connectivity, or direct road-kills. Because of their ecological characteristics, amphibians and reptiles are very exposed to road effects. In this article we review the relationship between these faunal groups and the road network. Amphibians exhibit high road-kill rates that can condition viability of some populations, and are vulnerable to pollution of road margins. Reptiles also suffer casualties because of road-kills when they move to paved roads for thermoregulation. Roads, especially those with high traffic load, act as barriers that difficult movements and contribute to population isolation in both groups. However, road impacts do not have equal intensity over space and time, and consequently some spatio-temporal patterns can be defined. Not all species show the same degree of exposure to road impacts, which depend on specific ecological requirements in each case. Mobile species are generally more vulnerable than sedentary ones. There are also intra-specific differences as a function of gender and age. All these considerations must be taken into account when designing and implementing the corresponding mitigation measures necessary to reduce the negative effects of roads on herpetofauna populations. Key words: fragmentation; mitigation measures; road ecology; road-effect zone; road-kills; road pollution. Herpetofauna y carreteras: revisión. Las carreteras y el tráfico rodado están estrechamente relacionados con varias de las principales amenazas para la biodiversidad. La red viaria impacta en las poblaciones de fauna silvestre, entre otros, mediante la ocupación y transformación de parte del territorio, la alteración del hábitat circundante, la dispersión de contaminantes físico-químicos, la fragmentación y pérdida de conectividad, o la mortalidad directa por atropello. Debido a sus características ecológicas, tanto los anfibios como los reptiles presentan una gran exposición a los efectos de las carreteras. En este artículo hacemos una revisión de la relación entre estos grupos faunísticos y la red viaria. Los anfibios experimentan elevadas tasas de atropello que pueden condicionar la viabilidad de algunas poblaciones y son vulnerables a la contaminación ligada a los márgenes de las carreteras. Los reptiles también sufren bajas por atropello al acudir a la calzada a termorregular. Para ambos grupos, las carreteras, sobre todo las de alta capacidad, constituyen barreras que dificultan sus movimientos y aíslan sus poblaciones. Sin embargo, los impactos no revisten la misma intensidad ni en el espacio ni en el tiempo, siendo posible definir una serie de patrones espacio-temporales. No todas las especies presentan el mismo grado de exposición a los impactos de las carreteras sino que éste depende de los requerimientos ecológicos específicos de cada una de ellas. En general, las especies más móviles son las más vulnerables. También hay diferencias intraespecíficas en función del sexo y la edad. Todas estas consideraciones deben ser tenidas en cuenta a la hora de diseñar e implementar las medidas de mitigación correspondientes para tratar de reducir los efectos negativos de las carreteras sobre las poblaciones de herpetofauna. Key words: atropellos; contaminación; ecología de carreteras; fragmentación; medidas de mitigación; zona efectiva de la carretera.

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The increment of both transport networks and social concern about environmental issues has led to an increase in the interest for studying the relationships between wildlife and roads. This has attracted the attention of many researchers, resulting in the emergence of a new discipline known as road ecology, whose main goal is the study of the interactions between organisms and the environment linked to road networks and traffic (FORMAN et al., 2003). Roads impact wildlife by means of causing habitat loss, population fragmentation, direct mortality, changes in animal behaviour, physical and chemical alterations of the environment, and dispersal of exotic species (MALO et al., 2004; JAARSMA et al., 2006; FAHRIG & RyTWINSky, 2009). Furthermore, the construction of new roads facilitates the use and modification of adjacent habitats by humans (FORMAN & ALEXANDER, 1998; TROMBULAk & FRISSELL, 2000), which indirectly contributes to additional impacts on wildlife. Amphibians and reptiles show certain ecological characteristics that make them highly vulnerable to roads impacts. For example, they show a low vagility in comparison to other vertebrates, being especially susceptible to habitat fragmentation by linear infrastructures. Moreover, in reptiles, roads constitute an important heat source for thermoregulation. In the case of amphibians, many species show complex life cycles, usually involving periodical migrations among the various complementing habitats in order to complete their annual cycle. Mortality rates during migrations associated with direct road-kills are in some cases high enough to cause effects at the population level. Moreover, their permeable skin, with osmoregulatory and respiratory functions, makes them sensitive to road pollution. These direct and

indirect effects of roads are among the main causes contributing to the global amphibian decline (BLAUSTEIN & WAkE, 1990; HOULAHAN et al., 2000; COLLINS & STORFER, 2003; NySTRöM et al., 2007). In this review, we summarize the progress made in the study of the effects of roads on the herpetofauna at multiple scales. Furthermore, we also analyse the mitigation measures proposed to reduce these road impacts. HERPETOFAUNA IN ROAD MORTALITy STUDIES The first studies in road ecology at the beginning of the 20th century alerted about the direct mortality of wildlife caused by vehicles on roads. The first specific studies about herpetofauna focused mostly on snake mortality (BUGBEE, 1945; FITCH, 1949; CAMPBELL, 1956), but also on amphibians (CARPENTER & DELzELL, 1951). HODSON (1960) made regular counts in a 3.2 km road stretch near Northamptonshire (England) and found 683 vertebrates of 42 species killed on the road. The species with the highest mortality rate was the European common frog (Rana temporaria) with 191 casualties. Most of the studies have been carried out in Europe (e.g. Belgium: BALLASINA, 1989; Germany: PODLOUCky, 1989; the Netherlands: zUIDERWIJk, 1989; Switzerland: RySER & GROSSENBACHER, 1989; Portugal: BRITO & ÁLVARES, 2004; France: LESBARRèRES et al., 2006, Poland: BRzEzIńSkI et al., 2012), North America (PALIS, 1994; ASHLEy & ROBINSON, 1996; RAy et al., 2006) and Australia (SEABROOk & DETTMAN, 1996; HOSkIN & GOOSEM, 2010), but in recent times, new studies have been conducted in South America (CAIRO & zALBA, 2007;

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HARTMANN et al., 2011; COELHO et al., 2012; QUINTERO-ÁNGEL et al., 2012) and Asia (SESHADRI et al., 2009; zHANG et al., 2010; GU et al., 2011; TOk et al., 2011). If the herpetofauna is especially vulnerable to road effects, road surveys will be expected to obtain a high rate of casualties in comparison to other vertebrate groups. Table 1 shows the results of several studies that have quantified road mortality for all vertebrate groups in different countries, and herpetofauna, especially amphibians, appears in most cases as the group with the highest road-kill rates. The percentage distribution among the different groups varies

both with the type of landscape present in the study area and the methodology followed in the road surveys. Thus, in some cases the number of amphibian and reptile casualties is low due to the low suitability of the prospected areas for these groups, or because the census technique underestimates the real number of casualties of these small-sized animals (MONTORI et al., 2003). Nevertheless, GRyz & kRAUzE (2008) found in a two-year monitoring study of a local road across Poland’s Biebrza River Valley that 90.7% of all the reported casualties were amphibians, especially anurans like the common toad (Bufo bufo), the moor frog (Rana

Table 1: Percentage of amphibian and reptile casualties in road surveys aimed to study road mortality for all vertebrate groups. Specific surveys for a certain group were not included in the table. Study

GONzÁLEz-PRIETO et al. (1993)

Amphibians Reptiles (%) (%)

Survey methodology

Traffic volume (vehicles/day)

Road stretch parallel to Miño River, Ourense n/a (Spain). Oak forest with some eucalyptus and pine spots, cultivated areas, scrublands and sub-urban habitat. ASHLEy & ROBINSON 93.8 2.7 Walking and bicycling. Three times Wetlands in the Long Point Causeway, ~ 3000 per week during spring-summer. Lake Erie (Canada). (1996) Vehicle and walking. Methodologies Bow River Valley, along the Trans-Canada 23.3 > 14000 CLEVENGER (1999) adapted to different specimen sizes. highway corridor in Banff National Park. Department of Vendée (France). 29.2 1.1 Vehicle. 19320 LODé (2000) ~ 14000 CLEVENGER et al. (2003) Vehicle. From April to November. Central Rocky Mountains (Canada). 7.1 5.9 Vehicle. New South Wales (Australia). Coastal 5000-20000 TAyLOR & GOLDINGAy 0.4 lowlands and volcanic plateau. Open pastures, (2004) eucalypt forest, plantations, scattered areas of rainforest. Road of access to Carrascal de la Font Roja n/a DÁVILA BLANES et al. 48.0 Vehicle. Daily surveys. 9.0 Natural Park (SE Spain). (2007) Indiana (USA). Wetlands and ditches su1900-6300 GLISTA et al. (2008) 1.3 Vehicle. Twice per week. 93.3 rrounded by agricultural lands and hardwood. GRyz & kRAUzE (2008) Biebrza River Valley (Poland). River, oxbow ~ 850 2.0 Walking. 3-7 days per month. 90.7 lakes, meadows and flooded pastures. HOBDAy & MINSTRELL n/a 0.8 Vehicle. Five regions, one survey Tasmania (Australia). Native and regenerated forest, woodland, grasslands and farming (2008) per season. regions. GEROW et al. (2010) 2200-6000 41.7 Vehicle and walking. Once per week. Saguaro National Park in south-eastern 41.5 Arizona (USA). CARVALHO & MIRA Mediterranean agro-silvo-pastoral system 3000-7000 24.1-66.1 4.4-6.5 Vehicle. Every two weeks. (Portugal). (2011) Road stretches across Catalonia (Spain). 126-10466 GARRIGA et al. 12.2 Vehicle. Three times in spring 62.0 (2012) and three times in autumn. 89.2

5.0

Walking. Weekly surveys.

Location and landscape description

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arvalis) and other Rana spp. that were killed during their migration to the breeding ponds. GLISTA et al. (2008) reported a percentage of amphibians even higher (95%) for four road surveys conducted in Indiana, USA. Unfortunately, despite these evidences, there are not too many studies addressing long-term trends in herpetofauna road-kills. The direction of these trends is difficult to foresee and may vary among areas. In some parts, road-kills have increased in recent decades in line with the road network expansion and the increasing number of vehicles and displacements (CARVALHO & MIRA, 2011). On the other hand, road-kills in other areas have decreased as a consequence of parallel population declines associated with the own road impacts (BROCkIE et al., 2009). Considerable variations among years are common due to population fluctuations related to environmental conditions or stochastic phenomena (COOkE, 1995). Thus, long-term studies are needed to detect accurate trends. As mentioned above, amphibian road-kills can reach very high rates. GOLDINGAy & TAyLOR (2006) estimated a mortality of more than 40 000 frogs per year in a 4-km stretch road in north-eastern New South Wales, Australia. Among reptiles, snakes are also commonly killed on roads due to their use as thermoregulation sites (ROSEN & LOWE, 1994). Freshwater turtles are another of the most vulnerable reptile groups to road mortality (GOODMAN et al., 1994; HAXTON, 2000). Data about lizard road-kills are scarce, but this group does not usually seem to suffer an elevated rate of road mortality; for example, in a study about the impact of off-highway recreation in southern California desert lands, GRANT & DOHERTy (2009) found no direct

mortality of the flat-tailed horned lizard (Phrynosoma mcallii) caused by off-road vehicles, although they caused other disturbances and indirect impacts that reduced habitat quality for the species. However, some saurian species can reach high road-kill rates on a local scale. In a road near Barcelona (Spain), the common wall gecko (Tarentola mauritanica) accounted for 20.2% of all herpetofauna road-kills, being the second most killed species (MONTORI et al., 2003). Nevertheless, the comparison of road mortality rates among all these studies is very complex and the results are difficult to interpret due to differences in study areas, population abundance, species richness, types of road, traffic densities, and species surveyed (JAkOB et al., 2003; PINOWSkI, 2005; GLISTA et al., 2008; SILLERO, 2008; ELzANOWSkI et al., 2009). Just as with other issues in conservation, social perception of wildlife road-kills varies in relation to species size. In general little public awareness is given to small size faunal groups such as amphibians and reptiles. However, they are suffering numerous casualties on the roads that can compromise population viability at a local scale (LANGTON, 1989; FAHRIG et al., 1995). This situation is even more puzzling considering that the implementation of effective solutions would be economically viable, at least for the main hotspots of herpetofauna road-kills. METHODOLOGIES TO STUDy ROAD MORTALITy Road mortality is generally recorded by direct counting of both dead and alive specimens on roads (VAN GELDER, 1973). However, when detecting all casualties by counting is

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not viable because, for example, a high extension of territory is aimed to be covered, a combination of counting and road fencing can be used (GIBBS & SHRIVER, 2005). In these cases there is always a discrepancy between the actual and the quantified numbers. Apart from speciesâ&#x20AC;&#x2122; anatomical and ecological characteristics, detection probability varies with the number of surveys, methodology used, and experience and skills of the staff involved in the census. Other important factors to take into account are the presence of scavengers that can remove the corpses, the topography and type of vegetation on the edges of the road, the weather and the time of the day (ANTWORTH et al., 2005). Thus, it is estimated that the actual number of road-kills can be up to 12-16 times higher than the estimated by a daily census (SLATER, 2002). Amphibian carcasses remain just a few hours on the road (SANTOS et al., 2011). Given that most road-kills occur at night, many carcasses may have disappeared in the morning, especially in roads with elevated traffic volumes. Furthermore, because of their small size and their fragility they are difficult to detect. Some species like the common fire salamander (Salamandra salamandra) may remain for longer on the road due to their tough skin and unpalatability (SANTOS et al., 2011). The season is also important to get accurate estimations; for example, to adequately study the road mortality problem on amphibian populations, sampling should be conducted preferably during migration days. The method of sampling depends on the goals of the study. Quantifying mortality and assessing potential population consequences require a method that allows the detection of as many individuals as possible so that the closer

we get to the real number, the better. These studies require walking surveys because during vehicle surveys many corpses remain undetected, and detection probabilities vary with the species. Some herpetofaunal studies consider that the maximum vehicle speed allowing for detection of all road-killed specimens is 20 km/h (SANTOS et al., 2007). However, if the aim is to locate hotspots where mitigation measures have to be implemented, both walking and vehicle surveys are valid. While walking surveys increase detection probabilities, vehicle surveys allow covering greater distances in less time, which is important considering the short time of residence of the corpses on the road (LANGEN et al., 2007). ROAD-kILL SPATIO-TEMPORAL DISTRIBUTION Herpetofauna road-kills are not randomly distributed in either space or time, but are concentrated in certain road segments during certain periods of time throughout the year. As amphibians and reptiles perform relatively localized displacements, it is possible to identify the variables that explain the spatial distribution of the hotspots, as well as the places where priority to the installation of mitigation measures should be given. The availability of suitable habitats (closeness to ponds and rivers, natural vegetation, absence of anthropogenic disturbance) is a factor commonly identified in road-kill spatial models (SILLERO, 2008). LANGEN et al. (2009) found that, for amphibian and reptile species in the north of New york State, road-mortality hotspots were located within a distance of 100 m from a wetland, especially in elevated road sections that presented wetlands on both

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sides. SANTOS et al. (2007), working in the road network of Catalonia (NE Spain), identified the presence of streams crossing the road, little steep roadsides, and semi-natural vegetation as the main explanatory factors of B. bufo road-kill locations. ORłOWSkI et al. (2008) found that the proportion of habitat occupied by forests and the presence of ponds were the main determinants of the distribution of amphibian road-kills in south-western Poland. Not only the presence but also the size of water bodies in the road vicinity was also relevant (ORłOWSkI, 2007). In other cases, wet grasslands instead of forest are the landscape types most associated with amphibian road-kills (GU et al., 2011). For freshwater turtles in New york State, hotspots were located at causeways that were larger than 200 m length, in close proximity to water, and with high forest coverage (LANGEN et al., 2012). It is important to take into account that the spatial distribution of road-kills is correlated with species local abundance (ORłOWSkI, 2007), which in turn depends on habitat features and varies throughout the year and among years. Apart from landscape variables, road parameters are also important to locate herpetofauna road-kills. For example, the probability of crossing highways is practically null (HELS & BUCHWALD, 2001, but see CARRETERO & ROSELL, 2000). Also, road-kill rate can be lower in main roads than in secondary ones, which could be explained by the previous population decline caused by roads with high traffic intensity (ORłOWSkI, 2007). In general terms, it is supposed that the number of roadkills increases with traffic volume. Due to their small sizes and relatively low speed of movements, amphibians and reptiles suffer

high number of casualties even with very low traffic volumes. Thus, VAN GELDER (1973) estimated that a traffic frequency of 10 vehicles per hour caused the loss of 30% B. bufo females that tried to cross the road to and from a breeding pond in the Netherlands. Furthermore, roads with low traffic volume can be more attractive for basking and feeding, as demonstrated by LEBBORONI & CORTI (2006) in lizards from central Italy, thus involving a higher risk of road-kill than roads with high traffic density. Nevertheless, the relationship between road-kills and traffic volumes varies with the species considered. MAzEROLLE (2004a) found that increased traffic intensities elevated the number of casualties in the American toad (Anaxyrus americanus) but decreased it in the spring peeper (Pseudacris crucifer). For frogs of the genus Lithobates the maximum number of road-kills was obtained at medium traffic volumes. For a given species, it is possible to calculate the probability of an individual being road-killed by considering the speed, the angle of intersection, and the traffic intensity. However, as explained below, some animals show specific behaviours in relation to traffic that reduce accuracy of this kind of estimates. At regional scale, habitat variables are better indicators to define areas with high road mortality rates than traffic volume, since they determine local distribution of the species (ORłOWSkI et al., 2008). For this reason, it is essential to understand the different spatial scales at which threat processes operate. In this sense, road segment and population scales are more suitable than regional and species distribution scales to identify the best location for implementation of the mitigation measures (BEAUDRy et al., 2008).

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The temporal distribution of herpetofauna road-kills is related to speciesâ&#x20AC;&#x2122; activity patterns (BERNARDINO & DALRyMPLE, 1992; BONNET et al., 1999). Inter-specific differences can also be partially explained by differences in speciesâ&#x20AC;&#x2122; ecological requirements. For amphibians, high risk exists during breeding migrations, which show high inter-specific variation in terms of distance (MONTORI et al., 2003; kOVAR et al., 2009). In the Lozoya Valley (Madrid, Spain) the peaks of road mortality for the natterjack toad (Bufo calamita) and B. bufo occurred between the last two weeks of March and the first two weeks of May, during the seasonal migrations to the breeding sites (SCV, 2003). As migration distances increase, the probability of crossing roads also increases. Moreover, the breeding habitat requirements condition roadkill locations. Thus, species that can breed in any accumulation of shallow water will have less defined hotspots than those species that breed in ponds or streams. In turtles and snakes, breeding season also concentrates most road-kills (BONNET et al., 1999; CURETON & DEATON, 2012). Furthermore, daily variations in temperature and precipitation may also influence road mortality rates, as observed by SHEPARD et al. (2008a) in these two groups of reptiles, whose frequency of road-kills is positively correlated with minimum daily temperatures. The behaviour in response to roads and traffic is another important factor determining the spatio-temporal distribution of road-kills, not only at the inter-specific but also at the intra-specific level. Many individuals remain immobile when sensing the arrival of a vehicle, which involves an increase in the crossing time, and consequently in the probability of being killed (MAzEROLLE et

al., 2005). For example, northern leopard frogs (Lithobates pipiens) move at a slower speed and following more tortuous paths as they approach roads. Furthermore, movement speed is also reduced as traffic volume grows, two synergistic factors that reduce the probability of cross success in this species (BOUCHARD et al., 2009). In the case of snakes and other reptiles as the snapping turtle (Chelydra serpentina), the number of road-kills increases as a consequence of the intentional action of the drivers. In a study conducted in Canada, up to 2.5% of the drivers positively selected to hit reptiles (ASHLEy et al., 2007). In other Australian study, 25% of drivers reported that they intentionally ran over invasive cane toads (Rhinella marina). However, field experiments did not confirm this behaviour, but found a rate of collisions not different from random (BECkMANN & SHINE, 2012). IMPACTS OF ROADS ON HERPETOFAUNA The effects of road pollution The impact of the emission of pollutants by vehicles and the use of certain chemicals in road maintenance (herbicides, de-icing salts) on herpetofauna (Fig. 1) has been studied almost exclusively in amphibians, and most examples point to a negative effects on their populations. For example, the proximity to the road correlates with an increased probability of suffering skeletal malformations and a smaller body size in larval wood frogs (Lithobates sylvaticus) (REEVES et al., 2008). Although the causes are not entirely clear, an increased risk of injury caused by predation due to a small size, a change in the composi-

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Figure 1: Main road effects on herpetofauna populations. Construction of new roads can imply the loss of high quality habitats for herpetofauna. Moreover, amphibians and reptiles can be road-killed when trying to cross. Roads difficult their movements by increasing fragmentation and exerting a barrier effect; however, road ditches sometimes act as corridors. Traffic flow also impacts on populations through dispersal of chemical pollutants or other types of pollution as noise or lighting. Some practices like the use of de-icing salt in winter maintenance can also impact amphibian or reptile populations. All these road impacts extend tens or hundreds of meters away from the road edge delimiting the road-effect zone.

tion of predator community, or the presence of chemical contaminants from vehicle traffic could explain this phenomenon (REEVES et al., 2008). Also, amphibian larvae near roads with high traffic volume bioaccumulate lead from fossil fuels at doses that may have physiological and reproductive effects (BIRDSALL et al., 1986). However, this problem has been reduced after removal of this heavy metal from fuels. A widely extended maintenance practice is the use of de-icing salt to prevent the formation of ice and snow on the road. This affects large regions at high latitudes and mountainous areas where frost and snow precipitation are common during winter. Winds or the water runoff can carry salt tens to hundreds of meters away from the roadside, increasing salinity in rivers and ponds nearby. High concen-

trations of salt in the water act as an environmental stressor affecting aquatic fauna with low tolerance to salinity. On amphibians, high salt concentrations can induce death by dehydration. Increased water salinity in road surroundings produced a reduction in survival rate, metamorphosis time, activity, and weight, as well as an increase in the number of malformations in larval L. sylvaticus (SANzO & HECNAR, 2006). For this species and for the spotted salamander (Ambystoma maculatum), kARRAkER et al. (2008) also found a reduction in the survival rate of embryos and larvae along with an up to 50% decrease in the number of egg masses in the vicinity of roads. They concluded that these effects could lead to local extinctions, especially of A. maculatum populations, given its high sensitivity to salinity. Laboratory experiments conducted in micro-

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cosms with water collected from basins receiving runoff from high-capacity roads showed total mortality of exposed L. sylvaticus larvae, while A. americanus larvae did not suffer lethal effects (SNODGRASS et al., 2008). In this sense, because of the inter-specific differences in sensitivity, the salt used on roads could act as a stressor capable of changing the structure of amphibian communities (COLLINS & RUSSELL, 2009). The impact of traffic noise and light High levels of environmental noise caused by traffic pose a great challenge for species that use acoustic communications (Fig. 1). Thus, traffic noise reduces the ability of female Copeâ&#x20AC;&#x2122;s gray treefrogs (Hyla chrysoscelis) to detect male calls (BEE & SWANSON, 2007). In an attempt to avoid masking by high levels of environmental background sound, terrestrial animals can introduce changes in the features of their acoustic signals. This capacity could be a key factor for reproductive success in noisy environments. In amphibians, some well-documented cases for this phenomenon show that the increases in signal frequencies in order to avoid masking (i.e. whistling tree frog, Litoria ewingii) are less marked than those identified for birds, but also sufficient to improve communication capacity (PARRIS et al., 2009). Another species, the Neotropical treefrog (Dendropsophus triangulum), is capable to increase its song rate (kAISER & HAMMERS, 2009). In an experiment exposing anurans inhabiting far away from roads to traffic noise recordings, CUNNINGTON & FAHRIG (2010) showed that animals immediately altered their vocalization characteristics in a similar way as indivi-

duals living permanently in locations with high traffic noise. This plasticity is essential to maintain acoustic communication in environments with traffic noise. However, LENGAGNE (2008) did not detect the ability to modify the frequencies or temporal structure of the call in the European tree frog (Hyla arborea), which suggests that those species capable to acclimatize to traffic noise may have a competitive advantage respect to less plastic species. In this regard, SUN & NARINS (2005) found that in a community exposed to environmental noise, individuals of the two-striped grass frog (Hylarana taipehensis) took advantage of this factor by increasing its song rate while the other three studies species decreased it. Light is another element associated with roads and traffic that may impact herpetofaunal populations. As many species are nocturnal, car lights may lessen their ability to prevent being killed by vehicles, since such lights saturate their retinas leaving them blind and disoriented for a few seconds. Light pollution also affects the ability of some species to detect and capture prey (BUCHANAN, 1993). Habitat fragmentation and loss of connectivity Roads act as barriers and/or filters, causing habitat fragmentation and population isolation (Fig. 1). Fragmentation is among the largest threats for amphibian and reptile populations (BECkER et al., 2007). Several studies have demonstrated the huge fragmentation effect caused by roads on herpetofaunal populations. For instance, in an experiment carried out in six paved roads in Virginia and West Virginia (USA) with translocated individuals of redback salamander

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(Plethodon cinereus) the rate of return for those who had to cross a road was about a 50% lower than for those who just had to pass through the forest (MARSH et al., 2005). In this species, genetic differences are visible in those populations bisected by a major highway but not in those separated by secondary roads (MARSH et al., 2008). As expected, road fragmentation does not affect equally to all species. Short-term impact of fragmentation correlates positively with speciesâ&#x20AC;&#x2122; dispersal ability, although those species with low dispersal abilities may be equally affected over long time periods (CUSHMAN, 2006). Most frog species in a wooded area showed no rejection to the presence of roads and forest roads, and even some species positively selected areas close to roads at certain stages of development. The opposite happened to urodeles as salamanders, whose populations were bigger in remote areas than in zones close to roads (DEMAyNADIER & HUNTER, 2000). For forest salamanders, the edge effect of forest roads due to reduced moisture and vegetation cover is comparable to recently cleared areas (MARSH & BECkMAN, 2004). Roads also alter selection patterns of the spawning sites in urodeles, especially in those species with little tolerance to alteration, as CHAMBERS (2008) observed for the Jefferson salamander (Ambystoma jeffersonianum) or the marbled salamander (Ambystoma opacum), whereas the more tolerant eastern newt (Notophthalmus viridescens) suffered less severe alterations of spawning site due to road impacts. Fire-roads can also act as barriers to herpetofauna movements, although there is important variation among species. CARTHEW et al. (2009) found that for most species of terres-

trial fauna in south-eastern South Australia, including the bold-striped cool-skink (Bassiana duperreyi), fire-roads were not an obstacle. However, for the painted spadefoot toad (Neobatrachus pictus) no crosses of the fire-road were detected. Snake responses to habitat fragmentation by roads can vary substantially depending on the speciesâ&#x20AC;&#x2122; ecological and anatomical features, the smaller being generally more reluctant to cross (ANDREWS & GIBBONS, 2005). The massasauga (Sistrurus catenatus) shows a clear rejection to traverse roads. In Manitoba (Canada) the red-sided garter snake (Thamnophis sirtalis parietalis) avoided gravel roads (SHINE et al., 2004). The same results were found for the eastern box turtle (Terrapene carolina) and the western box turtle (Terrapene ornata). If this rejection was heritable, road-kills would decrease over time because those individuals who tended not to cross the roads would be selected, which in turn would increase the degree of isolation between populations (SHEPARD et al., 2008b). The interruption of seasonal migrations caused by roads reduced genetic diversity and increased genetic differentiation in timber rattlesnakes (Crotalus horridus), even in a short period of time (CLARk et al., 2010). Some species can use roads as corridors (Fig. 1). To move between favorable habitats, green frogs (Lithobates clamitans) frequently used, with high survival rates, road drainage ditches inside a matrix of peat fields where they rarely ventured (MAzEROLLE, 2004b). Furthermore, road ditches are suitable insolation areas in forested ecosystems. Roads can also constitute a route of access to habitats with high vegetation cover for termophile species (HEDEEN & HEDEEN, 1999).

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However, the fact that roads can act as corridors may also have negative effects because, similarly, they may favour the spread of invasive species. Road surroundings are habitats with a substantial level of alteration that are favourable to invasive species. For example, the progressive spread of R. marina in Australia has been especially fast in areas with high road density and high connectivity (URBAN et al., 2008), which is an expected outcome considering that this invasive toad is more abundant in the roads or tracks than in the surrounding habitats (SEABROOk & DETTMAN, 1996). Effects of roads on herpetofauna populations A good review of the effects of roads on herpetofauna abundance (and also for the rest of the vertebrate groups) can be found in FAHRIG & RyTWINSkI (2009). They found in the literature 22 amphibians and six reptiles for which road effects had been evaluated. The responses were negative for more than 70% of amphibians and 80% of reptiles, although the results were sometimes contradictory. Population declines of herpetofauna species have been attributed, at least partially, to road mortality (FAHRIG et al., 1995; GIBBS & SHRIVER, 2002, 2005; MARCHAND & LITVAITIS, 2004; PUky, 2006). FAHRIG et al. (1995) or SUTHERLAND et al. (2010) showed that amphibian density was lower in roads with high traffic volumes than in roads with little traffic. Moreover, the ratio between dead and alive individuals was higher as traffic intensity increased. This fact supports the hypothesis that roadkills can cause the decline of amphibian populations near roads, especially in those

roads with high traffic volumes. These studies were carried out with anurans but the results are similar for urodeles (SEMLITSCH et al., 2007). Mortality rates can be higher in new roads and tend to diminish through the time, which would be related to the fact that road-kills contribute to progressively decrease population size, thus causing population declines (CARRETERO & ROSELL, 2000). In general, amphibian species richness decreases in presence of roads (FINDLAy & HOULAHAN, 1997; GARRIGA et al., 2012). It was estimated for A. maculatum in several breeding pools in New york State that under the average displacement performed during migration to breeding sites, one to threequarters of the population could be affected by road mortality. In those populations, road-kills involved the addition of more than 10% to natural mortality, which would jeopardize their long-term viability (GIBBS & SHRIVER, 2005). Thresholds of both road density and traffic volume over which the implementation of mitigation measures is justified can be defined using life tables and migration distances of each species (GIBBS & SHRIVER, 2005). However, differences in behaviour may make some species more susceptible than others to road impact. Thus, the level of vulnerability depends on the mobility of each species. A research carried out in the Ottawa-Carleton region (Ontario, Canada) revealed that, while a mobile species like L. pipiens showed a negative correlation between their population abundances and traffic density on nearby roads, in another species with shorter displacements like L. clamitans there was no relationship between population size and traffic intensity (CARR & FAHRIG, 2001).

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Reptile populations are also vulnerable to roads impacts. A study in Ontario, Canada, with radio-marked individuals of eastern rat snake (Pantherophis obsoletus) showed that the species did not avoid the road and, although the probability of collision was low (less than 1%), mortality was big enough to compromise the long-term population viability (ROW et al., 2007). For snakes, mobility is also a key factor to explain the inter-specific differences found in population vulnerability to road impacts (BONNET et al., 1999). Thus, forager snakes show higher vulnerability to roads than sedentary ones (MEEk, 2009). A model including road network, traffic volume and mobility of two species estimated an annual road mortality rate about three times higher for mobile species than for sedentary ones (14-21 % for the plain-bellied watersnake, Nerodia erythrogaster, vs. 3-5% in the Lake Erie water snake, Nerodia sipedon) (ROE et al., 2006). It is necessary to point out that road mortality can affect populations as long as these populations are limited by mechanisms independent from population density, in which case road-kills have an additive effect. Moreover, the distribution of casualties in relation to sex and age is also very important to assess the real impact of road-kills. For example, the mortality of females of oviparous species on their egg-laying migrations is likely to be more damaging to population viability than the same mortality rate affecting males or neonates (BONNET et al., 1999). These sex- or agerelated differences in magnitude of road impacts can modify the demographic structure of herpetofauna populations. In northern Portugal, males of Lataste’s viper (Vipera latastei) and Seoane’s viper (Vipera seoanei) were more frequently found road-killed than females, and

the same happened to adults relative to immature individuals. Moreover, a peak in the number of road-kills was found during the spring (BRITO & ÁLVARES, 2004), and the same happened for populations of those two species in northern Spain (MARTíNEz-FREIRíA & BRITO, 2012). On the other hand, MONTORI et al. (2003) suggested that for most snake species in Catalonia, including the horseshoe whip snake (Hemorrhois hippocrepis), the Montpellier snake (Malpolon monspessulanus), the ladder snake (Rhinechis scalaris) and the southern smooth snake (Coronella girondica), immature individuals would be more frequently road-killed than adults, being the peak of road mortality coincident with juvenile dispersal in late summer and fall. However, because of the low detectability of small individuals the real impact of road-kills on immature specimens would be underestimated. Male-biased road mortality has also been reported for other snake species like S. catenatus, with a peak at the end of the summer (SHEPARD et al., 2008a), or C. horridus, with males showing a road mortality rate 13 times higher than females (ALDRIDGE & BROWN, 1995). This bias could be explained by the high mobility of males, especially during the mating season when they increase their movements in search of females. Common green iguana (Iguana iguana) males suffer more road-kills than females, and populations near roads show a female-biased sex ratio of the adult population (RODDA, 1990). For certain freshwater turtle species, roadkills constitute a serious threat for their populations (GIBBS & SHRIVER, 2002). A welldocumented phenomenon in these species is the disproportionately high number of females road-killed during their displacements to

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overland nesting locations, an effect that is not so pronounced in semi-aquatic and terrestrial turtles. As a result, freshwater turtle populations near roads are biased towards males (STEEN & GIBBS, 2004; STEEN et al., 2006) and this bias has increased linearly during last decades (GIBBS & STEEN, 2005). Moreover, the edges of the road may become attractive places to establish nests, which attracts females to the vicinity of the road (ARESCO, 2005a). Thus, in many aquatic turtles the nesting period coincides with the maximum rate of road-kills (BEAUDRy et al., 2010). On the contrary, for other species like the diamondback terrapin (Malaclemys terrapin) there is no relationship between the proximity to roads and variations in individual density or sex ratio (GROSSE et al., 2011). Road impacts do not act only at a local scale. On a regional scale, the presence of roads is one of the main elements that explain the presence of various species of salamanders, but with different results depending on the specific tolerance to altered environments: the tolerant species are benefited while the abundance of the sensitive ones decreases (WARD et al., 2008). Road density is negatively correlated to the presence of R. arvalis whose probability of occurrence in habitats adjacent to roads is halved compared to what happens in non-fragmented areas. This fact shows that habitat fragmentation is one of the factors that explain the species’ spatial distribution patterns (VOS & CHARDON, 1998). The viability of a common spadefoot (Pelobates fuscus) metapopulation isolated by roads is quickly compromised by a slight decrease of individuals dispersing from source populations (HELS & NACHMAN, 2002). In fact, the magnitude of the impact of the road network on

forest anurans can be as high as the impact of habitat loss, as demonstrated by EIGENBROD et al. (2008), who observed in Ontario (Canada) that the species richness and abundance of three of the six species studied were more correlated to traffic density than to the absence of forest. As a consequence of the fragmentation caused by linear infrastructures there is a reduction in genetic exchange (VOS et al., 2001). Fragmented populations are very vulnerable to inbreeding processes. Low heterozygosity values were found in agile frog (Rana dalmatina) populations living in ponds near roads. The likely cause of this genetic homogenization is the reduction of the number of adult individuals, either by road-kills, noise or pollution (LESBARRèRES et al., 2003). Beyond lower allelic richness, populations fragmented by roads have a higher degree of genetic differentiation than populations from non-fragmented habitats (LESBARRèRES et al., 2006). Road-effect zone for herpetofauna The combined effect of all road impacts on herpetofauna populations delimits a “road-effect zone” (Fig. 1) that can be defined as the area in which ecological effects extend outward from a road (FORMAN & ALEXANDER, 1998). In the Galapagos Islands, lava lizards (Microlophus albemarlensis) have a lower density in the upper 100 m from the road edge, and show a probability of having experienced tail loss up to 30 times higher than individuals from other populations (TANNER & PERRy, 2007). The effect on the Tenerife lizard (Gallotia galloti) varies in relation to the adjacent habitat; while roads can play their usual role as barriers or source of

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mortality, in those cases where roads cross habitats unsuitable for the species, like the laurel forest, they may act as corridors (DELGADO GARCíA et al., 2007). Lungless salamanders avoid even the forest roads created for the timber industry, especially the oldest ones, within a band of about 35 m (SEMLITSCH et al., 2007). BOARMAN & SAzAkI (2006) found that the population density of desert tortoises (Gopherus agassizii) increased with the distance to the road edge, being especially small within a band of about 800 m around the road. VON SECkENDORFF HOFF & MARLOW (2002) extended this road-effect zone to a 4000 m bandwidth. The decrease in population density may not have linear relationship with the distance to the road and it is possible to detect thresholds; for example, EIGENBROD et al. (2009) identified in Ontario a threshold in population abundance at 250-1000 m from road edge for more than half of the studied species, as well as for the species richness itself. Finally, off-road vehicles may add an adverse effect on populations inhabiting road proximities (BURy & LUCkENBACH, 2002). Mitigation measures to reduce road impacts on herpetofauna Various mitigation measures with different efficacies and costs have been proposed in order to solve, or at least reduce the impact of road-kills and habitat fragmentation on amphibians and reptiles. The construction of fences or walls that impede access to the road has been effective in reducing mortality of amphibians (RySER & GROSSENBACHER, 1989; DODD et al., 2004). The installation of a temporary fence reduced the mortality in

freshwater turtles by more than 99%, effectively preventing the crossing attempts, which had only a 2% of chance of success (ARESCO, 2005b). However, although the installation of fences decreases the mortality rate due to collisions, it also reduces the connectivity with the other side of the road, making the migration to breeding sites difficult and enhancing population isolation. In order to promote habitat defragmentation, the adaptation of drainages and the construction of wildlife crossings designed for herpetofauna have been proposed. The effectiveness of this measure is supported by the fact that amphibian migration is facilitated by tunnels constructed for drainage under roads (HARTEL et al., 2009). The design and construction features that increase the effectiveness of herpetofauna passages vary among species, although there is considerable flexibility that facilitates decision-making to find a compromise considering the species present in the area. WOLTz et al. (2008) found that all the studied species – two freshwater turtles, C. serpentina and the painted turtle (Chrysemys picta), and two anurans (L. clamitans and L. pipiens) – preferred tunnels with diameter widths above 50 cm and sand or gravel firm, together with a fence of about 60 cm. The species showed a strong dislike for structures with diameters below 30 cm. In this sense, longer tunnels require larger diameters (PUky, 2003). LESBARRèRES et al. (2004) found that while B. bufo and the edible frog (Pelophylax esculentus) used the tunnels, R. dalmatina rejected them. All species preferred soil beds rather than bare concrete and corrugated steel. A viaduct construction together with the installation of a fence designed for amphibians and reptiles drastically reduced mortality in a road that run between two

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wetlands (SCOCCIANTI, 2006). The accessibility must be guaranteed in order to keep the tunnels or other kind of herpetofauna crossings functional (PUky, 2003). The location of herpetofauna passages should be selected following the knowledge acquired in the studies about the spatio-temporal distribution of road-kills as well as about the herpetofauna use of the space (PATRICk et al., 2010). Decision-making must take into account those routes positively selected by animals and the amplitude of their movement ranges (VAN GELDER et al., 1986; HARTEL et al., 2009). Several methodologies have been used to quantify the effectiveness of the mitigation measures: track plates (not specific for herpetofauna) (e.g. MATA et al., 2005), funnel traps or pitfall traps at tunnel exits (DODD et al., 2004) or cameras to monitor tunnel use (PAGNUCCO et al., 2011). Additionally, the structures associated with main roads, like ditches for drainage of rain water, seemingly harmless, can become real traps for amphibians during their migration. zHANG et al. (2010) described this problem for juveniles and sub-adults of Asian common toad (Duttaphrynus melanostictus). A design with sloping side walls with a maximum tilt of 66°, with a firm edge of stone and gravel with a cement grout and vegetation would minimize the impact of these structures on amphibian populations. Another suggested measure to reduce road impacts on fauna is the maintenance of clear, unvegetated margins to allow drivers for seeing animals before they begin to cross the road (ROSELL & VELASCO, 1999). While this is a measure basically though for large mammals, it can indirectly benefit reptiles; especially in forested areas, clear margins would constitute an area of insolation, out of the

zone of collision risk, for most reptiles that tend to use roads for thermoregulation. During road construction, habitat destruction can affect negatively to herpetofauna populations. GUyOT & CLOBERT (1997) proposed to minimize the impact on a population of Hermann’s tortoise (Testudo hermanni) by means of capturing and maintaining them in fenced spaces during construction of the road, and then releasing them after completion of the works. It has also been proposed that, when aquatic habitats used by amphibians for breeding are eliminated as part of the road construction process, the creation of replacement ponds close to where the destroyed ones were located would minimize the impact for populations (LESBARRèRES et al., 2010). Cost-benefit analysis is a possible approach to facilitate decision-making about where to locate mitigation measures. This approach follows economic criteria and tends to maximize the return in measure investments. SHWIFF et al. (2007), following the principles of ecological economics, quantified the economic losses associated with the herpetofauna road-kills in a Florida wilderness. Contrarily to these estimations including, for example, large ungulates, where it is possible to assess vehicle or human damages related to collisions, in the case of herpetofauna the authors selected the penalty that would be imposed by law in the state of Florida for the collection of an individual of the species found killed on roads in order to calculate the economic value of the losses. Most mitigation measures are usually thought and designed for their implementation at a local scale, following the recommendations of the environmental impact assess-

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ments. However, this scale fails to ensure adequate consideration of the potentially serious cumulative, indirect and synergistic ecological effects of roads. An adequate plan to minimize road impacts should start at the regional scale by spatial planning and strategic environmental assessment of the road network (TREWEEk et al., 1998), including the conservation of the herpetofauna species as another goal to achieve. For example, certain species such as the common snake-necked turtle (Chelodina longicollis) use different types of water points as a function of hydroperiods and seasonal movements. Therefore, their conservation requires not only the protection of water points and a land area around them, but should consider the diversity of habitats they use and ensure connectivity between them (ROE & GEORGES, 2007). The first step could be the identification of areas with high herpetofauna diversity. These would be the priority areas to install mitigation measures in the existing roads, and at the same time, they would be the areas protected from new infrastructure projects (BENAyAS et al., 2006). The same treatment should be applied to protected areas, where road-kill rates can be elevated due to the reception of visitors (GARRIGA et al., 2012). CONCLUSION According to the results of dozens of studies about the relationship between herpetofauna and roads, the increasing vehicular traffic is widely suspected to compromise herpetofauna conservation and to play a role in population declines. Nevertheless, the research about the road impacts on the herpetofauna is fragmented, limited for comparisons among zones, with methodological problems,

and inconclusive (ELzANOWSkI et al., 2009). Moreover, research results may sometimes be unexpected. For example, highway water ponds may surprisingly contribute to increase amphibian biodiversity in altered landscapes (LE VIOL et al., 2012). As we stated, the results of the road effects on herpetofauna populations can be contradictory (see FAHRIG & RyTWINSkI, 2009). We should deepen in the origins of these contradictions and the role that the differences in methodology play on them. An important effort to integrate all the information available is required. To increase our understanding about this topic we should exploit any technique or methodology available. Particularly promising are the advances in genetic analyses that allow the detection of small differences between populations that have partially lost their connectivity. At this respect, genetic results reveal that traffic intensity reduction could be a good solution for certain species (e.g. the palmate newt, Lissotriton helveticus) but not for others for which even the secondary roads with low traffic volumes act as barriers (e.g. the common midwife toad, Alytes obstetricans) (GARCIA-GONzALEz et al., 2012). Therefore, DNA analysis becomes a powerful tool to facilitate road planners and environmental managers the decision-making about the most effective mitigation measures. This decision is not always easy since the species present can respond differently to the same mitigation measure. Moreover, population consequences also vary among species both quantitatively and qualitatively. In this context it becomes necessary to develop mechanisms to reach a consensus solution. Road effects have been quantified only for a few number of herpetofauna species. More research on unstudied species and in different

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habitats is required, especially in those areas comprising both high biodiversity levels and an increasing expansion of the road networks (e.g. rainforests, HOSkIN & GOOSEM, 2010). Moreover, other possible road effects or mechanisms of affection should be considered. Thus, the loss of older individuals by road-related mortality could lead long-lived species to a depression in population reproduction rates (kARRAkER & GIBBS, 2011). Much more effort should focus on the increase of public awareness about this conservation issue. Indirect impacts of roads are not visible and, unlike with large mammals, direct effects such as road mortality on herpetofauna often go unnoticed for the society. It is necessary to help people take consciousness of the problem by implementing educational initiatives for transport and spatial planners and designers, drivers, environmental managers, and other people involved in transport and conservation issues, as well as for the public in general. The aim should be to achieve the routine incorporation of mitigation measures at a local scale and an adequate spatial planning to minimize road impacts on herpetofauna populations at a regional scale. Acknowledgement We are grateful to the editors of Basic and Applied Herpetology for giving us the opportunity to publish this review. Special thank to Dr. Manuel E. Ortiz Santaliestra for his collaboration and useful comments. Dr. Albert Montori and another anonymous reviewer contributed to improve the article with their suggestions. This study was partially funded by a predoctoral fellowship by the Junta de Castilla y Leon and the European Social Fund.

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frog Rana arvalis. Journal of Applied Ecology 35: 44-56. VOS, C.C.; ANTONISSE-DE JONG, A.G.; GOEDHART, P.W. & SMULDERS, M.J.M. (2001). Genetic similarity as a measure for connectivity between fragmented populations of the moor frog (Rana arvalis). Heredity 86: 598-608. WARD, R.L.; ANDERSON, J.T. & PETTy, J.T. (2008). Effects of road crossings on stream and streamside salamanders. Journal of Wildlife Management 72: 760-771. WOLTz, H.W.; GIBBS, J.P. & DUCEy, P.k. (2008). Road crossing structures for amphibians and reptiles: Informing design through behavioral analysis. Biological Conservation 141: 2745-2750. zHANG, z.-X.; yANG, H.-J.; yANG, H.-J.; LI, y.-X. & WANG, T.-H. (2010). The impact of roadside ditches on juvenile and subadult Bufo melanostictus migration. Ecological Engineering 36: 1242-1250. zUIDERWIJk, A. (1989). Amphibian and reptile tunnels in the Netherlands, In T.E.S. Langton (ed.) Amphibians and Roads: Proceedings of the Toad Tunnel Conference, Rendsburg, Federal Republic of Germany, 78 January 1989. ACO Polymer Products, Shefford, Bedfordshire, Uk, pp. 67-74.

Basic and Applied Herpetology 26 (2012): 33-42

Quantification of road mortality for amphibians and reptiles in Hoces del Alto Ebro y Rudrón Natural Park in 2005 Fernando Martínez-Freiría*, José C. Brito CIBIO, Centro de Investigação em Biodiversidade e Recursos Genéticos, Instituto de Ciências Agrárias de Vairão, Vairão, Portugal. * Correspondence: CIBIO, Centro de Investigação em Biodiversidade e Recursos Genéticos, Instituto de Ciências Agrárias de Vairão, R. Padre Armando Quintas, 4485-661 Vairão, Portugal. Phone: +351 252660400, Fax: +351 252661780, E-mail: fmartinez-freiria@cibio.up.pt

Received: 3 August 2011; received in revised form: 5 January 2012; accepted: 12 January 2012.

Roads are one the most important human agents of transformation, producing direct non natural, negative effects in wildlife. This work quantified road mortality on amphibian and reptile species in the Hoces del Alto Ebro y Rudrón Natural Park (north of Spain). In 2005, two types of roads (seven secondary and one main road) were sampled by car in order to detect road-killed specimens. Geographical Information Systems (GIS) and G-tests were used for analysing data, and mortality indexes (MI, number of specimens / 100 km sampled) were used as descriptors of the mortality risk on wild species. A total of 291 specimens was recorded, 115 amphibians belonging to four species and 176 reptiles belonging to 13 species. Bufo bufo represented more than 88% of the amphibians with MI peaks in spring and autumn. Natrix maura, Vipera aspis and V. latastei were the most frequently found road-killed reptiles (54.5%), presenting the two viper species MI peaks in spring. The number of road-kills was significantly higher in secondary roads than in the main one and also significantly high in well-preserved habitats. Three sections of high mortality were identified, all located in secondary roads that go through the Natural Park, enhancing the importance of habitat fragmentation as a major threat in biodiversity conservation. Management actions to reduce and/or eliminate the intensity of road mortality should be addressed in the Natural Park’s management plan and detailed studies should be performed to evaluate the effectiveness of installing traffic signs, road barriers and/or under-road passages. Key words: amphibians; GIS; Natural Park; northern Spain; reptiles; road-kills. Cuantificación de la mortalidad de anfibios y reptiles en las carreteras del Parque Natural de las Hoces del Alto Ebro y Rudrón en 2005. Las carreteras son uno de los agentes antrópicos de transformación ambiental más importantes, produciendo efectos negativos no naturales en fauna silvestre. En este trabajo cuantificamos la mortalidad en las carreteras de las especies de anfibios y reptiles presentes en el Parque Natural de las Hoces del Alto Ebro y Rudrón (norte de España). En 2005, se muestrearon en coche dos tipos de carreteras (siete secundarias y una principal) para detectar especímenes atropellados. Usamos Sistemas de Información Geográfica (GIS) y test G para analizar los datos, empleando índices de mortalidad (MI, número de individuos atropellados / 100 km muestreados) como descriptores del riesgo de mortalidad para las especies silvestres. Se registraron un total de 291 especímenes atropellados, 115 anfibios pertenecientes a cuatro especies y 176 reptiles pertenecientes a 13 especies. Bufo bufo representó más del 88% de los anfibios atropellados, con picos de mortalidad en primavera y otoño. Natrix maura, Vipera aspis y V. latastei fueron los reptiles atropellados con mayor frecuencia (54.5%), presentando las dos especies de víboras picos de mortalidad en primavera. El número de atropellos fue significativamente más alto en las carreteras secundarias que en la principal, así como en los hábitats mejor conservados. Se identificaron tres tramos con una alta mortalidad, todos ellos localizados en carreteras secundarias que discurren por el Parque Natural, lo que resalta la importancia de la fragmentación del hábitat como gran amenaza en la conservación de la biodiversidad. El plan de manejo del Parque Natural debería incluir acciones para reducir y/o eliminar los atropellos, al tiempo que se deberían realizar estudios detallados que evalúen la efectividad de la instalación de señales de tráfico, barreras y/o pasos subterráneos en las carreteras. Key words: anfibios; atropellos; GIS; norte de España; Parque Natural; reptiles.

MARTÍNEZ-FREIRÍA & BRITO

Landscape fragmentation is a major threat to the biodiversity conservation (SAUNDERS et al., 1991) and roads are one of the most important human agents of transformation (FORMAN et al., 2003). Roads produce direct non natural, negative effects on wildlife, such as increase of mortality due to collision with vehicles and also modification of animal behaviour (for revisions see TROMBULAk & FRISSELL, 2000; FAHRIG & RyTWINSky, 2009). As consequence, roads act as barriers to animal movement and reduce population connectivity, diminishing gene flow and limiting population dynamics, thus promoting inbreeding and loss of genetic diversity (FERRERAS, 2001; MARSH et al., 2005; ROW et al., 2007). Road-killing is one of the most important sources of unnatural mortality which have significant effects on populations of threatened and endangered species, including mammals, birds, reptiles, amphibians and invertebrates (TROMBULAk & FRISSELL, 2000; FAHRIG & RyTWINSky, 2009). Among vertebrates, amphibians and reptiles present physiological, ecological and behavioural traits that make them very vulnerable to roads (ANDREWS et al., 2008). They move slowly, have high dependency on particular habitats and exhibit seasonal movement patterns during which they are very vulnerable to road traffic. Thus, they are frequently found road-killed and many studies have quantified and evaluated the potential impact of road mortality in populations of both groups (e.g. BONNET et al., 1999; HELS & BUCHWALD, 2001; BRITO & ÁLVARES, 2004; GLISTA et al., 2007; ROW et al., 2007). The aims of this study are to quantify road mortality of amphibians and reptiles and identify road sections with higher mortality in the

protected Hoces del Alto Ebro y Rudrón Natural Park (Northern Spain). Protected areas should preserve better communities of amphibians and reptiles (e.g. in terms of densities) than unprotected areas, since they are keeping high quality habitats. However, protected areas attract many vehicles in determinate seasons (e.g. tourism) and could present higher intensities of road mortality than unprotected areas. The present study is expected to contribute to the Natural Park’s management plan by reducing the intensity of road mortality on amphibians and reptiles. MATERIALS AND METHODS Study area The study area is located in the high course of the Ebro river in northern Spain (latitude: 42º 37.7’ - 42º 58.7’ N; longitude: 3º 37.3’ 3º 58.5’ W). It partially includes the recently protected Hoces del Alto Ebro y Rudrón Natural Park (JCyL, 2008), in north-western Burgos province, and the adjacent Valderredible valley, in south-eastern Santander province. It is a transition among Eurosiberian and Mediterranean regions (RIVAS-MARTíNEz, 1987) and its landscape mainly consists of calcareous plateaus excavated by the Ebro river and its tributary, the Rudrón river, forming canyons and steep valleys. Climate is subhumid Mediterranean with Central European tendency (FONT TULLOT, 1983). Amphibian and reptile species The study area has a relative high diversity of herpetofauna, with 13 and 16 species of amphibians and reptiles, respectively

AMPHIBIAN AND REPTILE ROAD-KILLS IN A NATURAL PARK

(PLEGUEzUELOS et al., 2002). Moreover, given its condition of transition among bioclimatic regions, several species are at the border of their distributional range and/or in contact with species of the same genus (SILLERO et al., 2009). The latter is the case of the Iberian vipers: the study area is the only known contact zone among the three species (Vipera aspis, V. latastei and V. seoanei), where V. aspis and V. latastei meet in sympatry and hybridize (for more details see MARTíNEzFREIRíA et al., 2008, 2009, 2010), enhancing the importance of this area in biogeographical and ecological studies. Road sampling A total of 82.11 km of roads was equally sampled by car with an average monthly frequency of 5.12 (range = 3-6) samplings. Road sampling was performed from March to October of 2005 at a driving speed lower than 40 km/h. Two types of roads were sam-

pled: one main road with 24.16 km of length and with high level of traffic circulation, and seven secondary roads with 57.84 km of length in total and with low level of traffic circulation (Table 1, Fig. 1). Road-kills were identified to specific level using keys based on morphology (e.g. SALVADOR, 1998) and genetic markers (only for vipers, see MARTíNEz-FREIRíA et al., 2009), and their locations were recorded with a GPS (European Datum 1950). Road-kill data analyses Road-kill data were introduced in a georreferenced database and represented in ArcMap 9.3 GIS software (ESRI, Redlands, California, USA). GIS was used for two purposes: i) to extract habitat types for each road-kill by intersecting the road-kill location with Corine Land Cover 2006 raster data version 15 at a resolution of 100 m2 (EEA, 2011), and ii) to visualize and detect road sections of high mortality by

Table 1: Road length (km), number (N) and percentage (%) of road-kills, and mortality index (MI) by roads sampled and in road sections with high mortality. Road

Length (km)

N (%)

Main roads

N-623

24.16

35 (12.03)

144.87

Secondary roads

BU-3345 BU-3342 BU-3465 CA-757 CA-274 CA-275 BU-613 Total

16.40 9.46 7.00 5.50 3.82 6.51 9.26 57.95

98 (33.68) 71 (24.40) 5 (1.72) 1 (0.34) 6 (2.06) 10 (3.44) 65 (22.34) 256 (87.97)

597.56 750.53 71.43 18.18 157.07 153.61 701.94 441.76

High mortality road sections

Section 1 Section 2 Section 3 Total

4.36 6.72 9.26 20.35

65 (22.34) 71 (24.40) 65 (22.34) 201 (69.07)

1490.82 1056.55 701.94 987.72

MARTĂ?NEZ-FREIRĂ?A & BRITO

measuring the degree of clustering of road-kill data using the Average Nearest Neighbor (ANN), Getis-Ord General G (GOG G) and kernel density functions from ArcToolbox (ESRI, Redlands, California, USA). G-tests were used to compare road-kill data by groups (i.e. amphibians and reptiles), subgroups (i.e. anurans, urodeles, lizards and snakes), species, months (from March to October), road types (i.e. main or secondary), roads (eight roads, see Table 1) and habitat types (eight classes including urban areas, bare areas, cultivated areas, grasslands, bushy areas, evergreen and deciduous forests, and river beds). Contingency tables, using G-test statistics, were also performed to test the relationship among groups, sub-groups and species with months, road types, roads and habitats. Statistical procedures were developed on R-software (R DEVELOPMENT CORE TEAM, 2010).

Mortality indexes (MI, number of specimens/100 km of sampled road) were used as descriptors of the mortality risk by month, road type and road. RESULTS A total of 291 specimens were recorded (115 amphibians and 176 reptiles) from four amphibian and 13 reptiles species (Table 2). Hybrids between V. aspis and V. latastei were also found road-killed. G-tests resulted in significant differences for all the variables when analysed separately but only for some pairs of variables when their relationship was analysed (Table 3). Reptiles were more commonly road-killed than amphibians, being anurans and snakes the most frequently road-killed sub-groups (Table 2). Among the most frequently road-killed spe-

Figure 1: Sampled roads with the distribution of road-killed species (left). Maps on the right side depict particular road sections where high mortality was detected (S1: section 1, S2: section 2, S3: section 3).

AMPHIBIAN AND REPTILE ROAD-KILLS IN A NATURAL PARK

Table 2: Number of road-kills (N), percentage of the total (% total), percentage within each group (% group) and mortality index (MI) by species and group. Taxa

% total

% group

Salamandra salamandra Triturus marmoratus Bufo bufo Pelophylax perezi Total Urodela Total Anura Total Amphibia

11 1 101 2 12 103 115

3.78 0.34 34.71 0.69 4.12 35.40 39.52

8.80 0.80 80.80 1.60 9.60 82.40 100.00

13.40 1.22 123.01 2.44 14.62 125.45 140.07

Anguis fragilis Chalcides striatus Lacerta bilineata Lacerta schreiberi Timon lepidus Podarcis hispanica Coronella austriaca Coronella girondica Natrix maura Natrix natrix Vipera aspis Vipera latastei Vipera seoanei Hybrid vipers Total Sauria Total Colubridae Total Viperidae Total Ophidia Total Reptiles

11 4 7 2 8 2 8 10 25 8 46 25 2 18 34 51 91 142 176

3.78 1.37 2.41 0.69 2.75 0.69 2.75 3.44 8.59 2.75 15.81 8.59 0.69 6.19 11.68 17.53 31.27 48.80 60.48

6.25 2.27 3.98 1.14 4.55 1.14 4.55 5.68 14.20 4.55 26.14 14.20 1.14 10.23 19.32 28.98 51.70 80.68 100.00

13.40 4.87 8.53 2.44 9.74 2.44 9.74 12.18 30.45 9.74 56.02 30.45 2.44 21.92 41.41 62.11 110.83 172.94 214.35

cies one amphibian, Bufo bufo, and three snakes, Natrix maura, V. aspis and V. latastei, accounted for more than 67% of all road-killed specimens; this value is even higher (73%) if hybrid vipers are added (Table 2, Fig. 1). The number of road-killed animals was significantly different among months (Table 3), being April the month with the largest number of road-killed animals (N = 86; 29.6%). Moreover, significant differences in the frequency of road-killed specimens were found when analysing the relationships between months and groups, sub-groups or species

(Table 3). The visualization of Fig. 2 allows the detection of monthly MI peaks in spring for B. bufo, V. aspis, V. latastei and for hybrid vipers, as well as for B. bufo and V. aspis in late summer and autumn. The number of road-killed specimens was significantly higher in secondary roads than in the main road (Tables 1, 3). Significantly different frequencies were found among the studied roads when considering all the roadkills together and also when comparing by groups, sub-groups and species (Tables 1, 3). Three secondary roads mostly running

MARTÍNEZ-FREIRÍA & BRITO

Table 3: G-test of goodness of fit for road-kill data when testing for differences as a function of each variable or the relationship between two variables. Variables

d.f.

Group Sub-group Species Month Road type Road Habitat Group x Month Group x Road type Group x Road Group x Habitat Sub-group x Month Sub-group x Road type Sub-group x Road Sub-group x Habitat Species x Month Species x Road type Species x Road Species x Habitat

1 3 17 7 1 7 7 7 1 7 7 21 3 21 21 119 17 119 119

12.882 166.589 385.732 111.318 193.555 105.508 297.474 75.578 0.846 21.974 29.532 92.012 3.498 71.596 42.586 198.525 24.366 288.040 168.129

< 0.001 < 0.001 < 0.001 < 0.001 < 0.001 < 0.001 < 0.001 < 0.001 0.358 0.003 < 0.001 < 0.001 0.321 < 0.001 0.004 < 0.001 0.110 < 0.001 0.002

within the protected area presented more than 80% of all road-killed specimens (Table 1, Fig. 1). GIS cluster analyses resulted in a high degree of clustering for all road-kill data (ANN ratio = 0.22; z = -25.70; P < 0.001), mainly when roads were used as factors (GOG G = 0; z = 4.36; P = 0.01). kernel density function allowed the identification of three sections within the three secondary roads with high mortality, having the 69.1% of all road-kills and the highest values of MI (Table 1, Fig. 1): 1) a section of BU-3345 presented 56% and 55.6% of all road-killed V. latastei and hybrid vipers, respectively; 2) a section of road BU-3342 presented 24.8%, 58.7% and 27.8% of all road-killed B. bufo, V. aspis and hybrid vipers, respectively;

and 3) a section of road BU-613 presented 37.6% and 19.6% of all road-killed B. bufo and V. aspis, respectively. The number of road-killed animals was significantly different among the different types of habitats (Table 3), being more frequent in bushy (N = 117; 40.2%) and cultivated areas (N = 64; 21.9%), and in river beds (N = 48; 16.5%). Also, significant differences were found for the number of roadkilled groups, sub-groups and species by habitat (Table 3). Bufo bufo was found most frequently road-killed in bushy areas (N = 37; 36.6%), N. maura in cultivated (N = 7; 28.0%) and bushy areas (N = 8; 32.0%), and in river beds (N = 7; 28.0%), V. aspis in bushy areas (N = 17; 36.9%), evergreen (N = 11; 23.9%) and deciduous forests (N = 8; 17.4%), V. latastei in cultivated (N = 14; 56.0%) and bushy areas (N = 10; 40.0%) and hybrid vipers in bushy areas (N = 11; 61.1%). DISCUSSION The present single-year estimation of road mortality is the first systematic evaluation of road-killing on amphibians and reptiles in the Hoces del alto Ebro y Rudrón Natural Park. Opportunistic road sampling series conducted in the area during 2006 and 2007 reported similar results regarding most frequent roadkilling species and sections of roads with high mortality (authors’ unpublished data). Road-traffic circulation during 2005 induced high mortality over the amphibians and reptiles in the protected area, especially in B. bufo, N. maura, V. aspis, V. latastei and hybrid vipers. Anurans and snakes have been reported to be the groups with higher vulnerability to road mortality within amphibians and

AMPHIBIAN AND REPTILE ROAD-KILLS IN A NATURAL PARK

Figure 2: Monthly variation of the mortality index for the most frequently roadkilled species and groups.

reptiles, respectively (TROMBULAk & FRISSELL, 2000). The observed monthly pattern of road mortality for these groups seems to be related to seasonal movements from a particular habitat to another (ANDREWS et al., 2008). Bufo bufo had a bimodal road mortality pattern probably related to the annual migration to reproduction sites (HELS & BUCHWALD, 2001). Vipera aspis, V. latastei and hybrid vipers exhibited a unimodal road mortality pattern, with a mortality peak in spring, probably related to dispersal activities after hibernation and mating season (BONNET et al., 1999; BRITO & ÁLVARES, 2004; ROW et al., 2007; MARTíNEzFREIRíA et al., 2010). Vipers were the most road-killed snakes and reptiles. The high viper densities in the study area (MARTíNEz-FREIRíA et al., 2010), the use of roads for thermoregulation and the slow locomotion are probably the factors that explain this pattern (BONNET et al., 1999; BRITO & ÁLVARES, 2004). Road-kills were not randomly spatiallydistributed throughout the roads of the study area. In fact, this study identified road-kill aggregations mainly in three sections of

secondary roads in the protected area, which represented about the 25% of all sampled roads. Landscape variables, including topography and habitat composition, would be expected to play an important role in determining road-kill aggregations but also road variables such as intensity of traffic circulation (FERRERAS, 2001; CLEVENGER et al., 2003). In our study, secondary roads presented lower intensity of traffic circulation than main roads but presented higher levels of road-kills. Due to low permanence of amphibian and reptile carcasses on the roads (SANTOS et al., 2011), road-sampling effort (i.e. monthly road sampling frequency) performed in the current study could be obscuring the detection of more accurate patterns of road mortality in main roads. However, among the secondary roads, with similar intensity of traffic circulation, those going through the protected area presented higher number of road-kills than roads located outside. This suggests that landscape could be playing an important role in the observed pattern of road mortality in secondary roads.

MARTÍNEZ-FREIRÍA & BRITO

The topography of the protected area is quite different from the surroundings (steep valleys and canyons vs. flat areas) and road-kills were more common in well-preserved habitats (e.g. road-kills in bushy areas and river beds represented about 56% of all road-kills) than in the other habitats. The only exception appears in cultivated areas, which also presented high road-kill numbers (21.95% of all road-kills); however, in the protected area, cultivated lands have been abandoned since the 1970s and currently present herbaceous and bushy vegetation, favouring high densities of amphibians and reptiles (e.g. MARTINEz-FREIRíA et al., 2010). Therefore, the three sections of secondary roads with high mortality are probably acting as major agents of habitat fragmentation and barriers to wildlife movements, since they go along of the canyons, close to rivers, and divide patches of well-preserved habitats. Road-sampling effort should be increased in future studies to obtain more accurate patterns of road mortality (SANTOS et al., 2011), and also management strategies to reduce the intensity of road mortality should be addressed in the Natural Park’s management plan (e.g. HELS & BUCHWALD, 2001; BRITO & ÁLVARES, 2004; GLISTA et al., 2007; ROW et al., 2007; WOLTz et al., 2008). These strategies should consider: i) installation of signals and construction of road barriers and under-road passages to prevent road mortality; ii) translocation of living animals and removal of carcasses found on roads to prevent further mortality; iii) studies quantifying population structure and connectivity using molecular tools to identify potential roadbarriers to gene flow; and iv) local environmental education campaigns promoted by the Natural Park administration.

Acknowledgement Authors want to thank to Asociación sociocultural Hoces del Alto Ebro y Rudrón (Burgos, Spain), friends who helped in the interminable hours of road-sampling and P. Tarroso for his support in the analyses. FMF and JCB are supported by FCT (SFRH/BPD/69857/2010 and Programme Ciência 2007, respectively). REFERENCES ANDREWS, k.M.; GIBBONS, J.W. & JOCHIMSEN D.M. (2008). Ecological effects of roads on amphibians and reptiles: a literature review, In J.C. Mitchell, R.E. Jung Brown & B. Bartholomew (eds.) Urban Herpetology. Series: Herpetological Conservation, vol. 3. Society for the Study of Amphibians and Reptiles, Salt Lake City, Utah, USA, pp. 121-143. BONNET, X.; NAULLEAU, G. & SHINE, R. (1999). The dangers of leaving home: dispersal and mortality in snakes. Biological Conservation 89: 39-50. BRITO, J.C. & ÁLVARES, F. (2004). Patterns of road mortality in V. latastei and V. seoanei from northern Portugal. AmphibiaReptilia 25: 459-465. CLEVENGER, A.P.; CHRUSzCz, B. & GUNSON, k. (2003). Spatial patterns and factors influencing small vertebrate fauna road-kill aggregations. Biological Conservation 109: 15-26. EEA (2011). Corine Land Cover 2006 raster data - version 15 (08/2011). European Environment Agency, Copenhagen, Denmark. Avalaible at http://www.eea. europa.eu/data-and-maps/data/corineland-cover-2006-raster-1. Retrieved on 12/12/2011.

AMPHIBIAN AND REPTILE ROAD-KILLS IN A NATURAL PARK

FAHRIG, L. & RyTWINSky, T. (2009). Effects of roads on animal abundance: an empirical review and synthesis. Ecology and Society 14: 21. FERRERAS, P. (2001). Landscape structure and asymmetrical inter-patch connectivity in a metapopulation of the endangered Iberian lynx. Biological Conservation 100: 125-136. FONT TULLOT, I. (1983). Climatología de España y Portugal. Instituto Nacional de Meteorología, Madrid, Spain. FORMAN, R.T.T.; SPERLING, D.; BISSONETTE, J.A.; CLEVENGER, A.P.; CUTSHALL, C.D.; DALE, V.H.; FAHRIG, L.; FRANCE, R.; GOLDMAN, C.R.; HEANUE, k.; JONES, J.A.; SWANSON, F.J.; TURRENTINE, T. & WINTER, T.C. (2003). Road Ecology: Science and Solutions. Island Press, Washington, DC, USA. GLISTA, D.J.; DEVAULT, T.L. & DEWOODy, J.A. (2007). Vertebrate road mortality predominantly impacts amphibians. Herpetological Conservation and Biology 3: 77-87. HELS, T. & BUCHWALD, E. (2001). The effect of road kills on amphibian populations. Biological Conservation 99: 331-340. JCyL (2008). Red de Espacios Naturales. Junta de Castilla y León, Valladolid, Spain. Available at http://www.jcyl.es/web/jcyl/ MedioAmbiente/es/Plantilla100/1131977 537178/_/_/_. Retrieved on 03/12/2008. MARSH, D.M.; MILAM, G.S.; GORHAM, N.P. & BECkMAN, N.G. (2005). Forest roads as partial barriers to terrestrial salamander movement. Conservation Biology 19: 2004-2008. MARTíNEz-FREIRíA, F.; SILLERO, N.; LIzANA, M. & BRITO, J.C. (2008). GIS-based

niche models identify environmental correlates sustaining a contact zone between three species of European vipers. Diversity and Distributions 14: 452-461. MARTíNEz-FREIRíA, F.; SANTOS, X.; PLEGUEzUELOS, J.M.; LIzANA, M. & BRITO, J.C. (2009). Geographical patterns of morphological variation and environmental correlates in contact zones: a multi-scale approach using two Mediterranean vipers (Serpentes). Journal of Zoological Systematics and Evolutionary Research 47: 357-367. MARTíNEz-FREIRíA, F.; LIzANA, M.; AMARAL, J.P. & BRITO, J.C. (2010). Spatial and temporal segregation allows coexistence in a hybrid zone among two Mediterranean vipers (Vipera aspis and V. latastei). Amphibia-Reptilia 31: 195-212. PLEGUEzUELOS, J.M.; MÁRQUEz, R. & LIzANA, M. (2002). Atlas y Libro Rojo de los Anfibios y Reptiles de España. Dirección General de Conservación de la Naturaleza - Asociación Herpetológica Española, Madrid, Spain. R DEVELOPMENT CORE TEAM (2010). R: A Language and Environment for Statistical Computing. R Foundation for Statistical Computing, Vienna, Austria. Available at http://www.r-project.org/. Retrieved on 12/12/2011. RIVAS-MARTíNEz, S. (1987). Memoria del Mapa de Series de Vegetación de España. Ministerio de Agricultura, Pesca y Alimentación, ICONA, Madrid, Spain. ROW, J.R.; BLOUIN-DEMERS, G. & WEATHERHEAD, P.J. (2007). Demographic effects of road mortality in black ratsnakes (Elaphe obsoleta). Biological Conservation 137: 117-124.

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SALVADOR, A. (1998). Reptiles. Series: Fauna Ibérica, vol. 10 (M.A. Ramos, coord.). Museo Nacional de Ciencias Naturales, CSIC, Madrid, Spain. SANTOS, S.M.; CARVALHO, F. & MIRA, A. (2011). How long do the dead survive on the road? Carcass persistence probability and implications for road-kill monitoring surveys. PLoS ONE 6: e25383. SAUNDERS, D.A.; HOBBS, R.J. & MARGULES, C.R. (1991). Biological consequences of ecosystem fragmentation: a review. Conservation Biology 5: 18-32.

SILLERO, N.; BRITO, J.C.; SkIDMORE, A.k. & TOXOPEUS, A.G. (2009). Biogeographical patterns derived from remote sensing variables: the amphibians and reptiles of the Iberian Peninsula. Amphibia-Reptilia 30: 185-206. TROMBULAk, S.C. & FRISSELL, C.A. (2000). Review of ecological effects of roads on terrestrial and aquatic communities. Conservation Biology 14: 18-30. WOLTz, H.W.; GIBBS, J.P. & DUCEy, P.k. (2008). Road crossing structures for amphibians and reptiles: Informing design through behavioral analysis. Biological Conservation 141: 2745-2750.

Basic and Applied Herpetology 26 (2012): 43-56

Natural fluctuations in a stream dwelling newt as a result of extreme rainfall: a 21-year survey of a Calotriton asper population Albert Montori*, Alex Richter-Boix, Marc Franch, Xavier Santos, Núria Garriga, Gustavo A. Llorente Departament de Biologia Animal (Vertebrats), Facultat de Biologia, Universitat de Barcelona, Barcelona, Spain. * Correspondence: Departament de Biologia Animal (Vertebrats), Facultat de Biologia, Universitat de Barcelona, Avinguda Diagonal 645, 08028 Barcelona, Spain. Phone: +34 93 402 1456, Fax: +34 93 403 5740, E-mail: amontori@ub.edu

Received: 23 January 2012; received in revised form: 18 July 2012; accepted: 20 July 2012.

Amphibians living in streams are often subjected to spates. These waterfloods are the main cause of organism drift and mortality, and these effects can be confounded with population decline. Discrimination between population decline and natural fluctuations in unpredictable habitats requires the study of population dynamics through monitoring and long data series. We conducted a 21-year demographic field survey of a newt (Calotriton asper) population in the eastern Pyrenees. Our results indicate that the adult population showed high fluctuations in response to heavy rainfall. Maximal rainfall in 24 h (higher than 50 l / m2) caused population decreases as a result of catastrophic drift. The larval population also decreased after heavy rainfall. The data from this survey show that the population recovered three years after catastrophic drift. Subadult C. asper show terrestrial activity and are not affected by waterfloods. Interannual dynamics revealed that the stock of subadults allowed for rapid population recovery after catastrophic drift episodes. Flooding produced higher mortality when it occurred during winter than during the active cycle of newts. This long-term study provides new insights into the survival strategies displayed by newts in response to extreme stream environments. Key words: Calotriton; capture-recapture; natural fluctuations; long-term survey; Salamandridae. Fluctuaciones naturales en un tritón de arroyo como resultado de la lluvia extrema: seguimiento de una población de Calotriton asper durante 21 años. Los anfibios que viven en arroyos están sujetos a menudo al impacto de las crecidas. Estas inundaciones son la causa principal de arrastre y muerte de los organismos, pudiendo confundirse estos efectos con declives poblacionales. Discriminar entre declives y fluctuaciones naturales requiere del estudio de las dinámicas poblacionales mediante la monitorización y las series de datos prolongadas. Realizamos un estudio de campo durante 21 años de una población del tritón pirenaico (Calotriton asper) en el Pirineo oriental. Nuestros resultados indican que la población de adultos mostró fluctuaciones marcadas como respuesta a las precipitaciones intensas. La precipitación máxima registrada en un periodo de 24 horas (superior a los 50 l / m2) causó un descenso de la población como consecuencia de una avenida catastrófica. La población de larvas también se redujo tras las lluvias intensas. Los datos del presente estudio muestran que la población se recuperó tres años después de la catástrofe. Los subadultos de C. asper presentan actividad terrestre y no se ven afectados por las inundaciones. Las dinámicas interanuales revelaron que la reserva de subadultos permitió la rápida recuperación de la población tras las crecidas catastróficas. Las inundaciones causaron más muertes cuando sucedieron en invierno que cuando lo hicieron durante el periodo de actividad de los tritones. Este estudio a largo plazo ayuda a comprender las estrategias de supervivencia que presentan los tritones en respuesta a ambientes extremos. Key words: Calotriton; captura-recaptura; fluctuaciones naturales; Salamandridae; seguimiento a largo plazo.

MONTORI ET AL.

Organisms inhabiting running waters are frequently exposed to strong changes in hydrological conditions (RESH et al., 1988), downstream drift being one of the most important factors modifying their population structure (LANCASTER et al., 1990; LANCASTER & HILDREW, 1993) and a major cause of individual mortality (THIESMEIER & SCHUHMACHER, 1990; BARRETT et al., 2010). These catastrophic perturbations are expected to produce high mortality rates in natural populations; however, many species and communities living in these habitats seek shelter in benthic refuges, such as stones, debris dams, woody debris or fissures in the rocky riverbed, and thus typically recover quickly from strong hydraulic discharge (BILBy & LIkENS, 1980; LANCASTER et al., 1990; LANCASTER & HILDREW, 1993). It is expected that organisms from habitats exposed to strong temporal alterations will show high population growth rates and are thus resilient in those constantly altered environments (MAy et al., 1974). The relationship between population growth rate and size is crucial to an understanding of population dynamics and the speed at which populations return to equilibrium (near population carrying capacity) after displacement by an external perturbation (MAy et al., 1974; SALVIDIO, 2011). Amphibians living in running waters exposed to strong changes in hydrological conditions are good candidates to examine temporal population trends and how they recover after perturbations. In fact, amphibians exhibit strong year-to-year population fluctuations (PECHMANN et al., 1991). Given the current trend in population decline and species extinction worldwide (BLAUSTEIN et

al., 1994; HOULAHAN et al., 2000; STUART et al., 2004; BEEBEE & GRIFFITHS, 2005), it is sometimes difficult to distinguish population decline from natural fluctuations. Thus deeper knowledge of amphibian demography in natural habitats over time and the analysis of demographic time-series data may allow for detecting true population trends (BLAUSTEIN et al., 1994; REED & BLAUSTEIN, 1995) and ultimate factors that determine amphibian population dynamics (MEyER et al., 1998). Furthermore, BIEk et al. (2002) consider that inventory and monitoring efforts should be complemented by demographic studies in order to apply quantitative analyses to a wide range of species and life-history groups. Accordingly, long-term demographic monitoring is required to analyze temporal fluctuations and how populations recover after natural perturbations. We have examined temporal variation in population trends in the Pyrenean newt Calotriton asper, a species highly adapted to running-water habitats. It presents a flattened body, rough skin, lung reduction and horny nails. This newt inhabits clear oxygen-rich mountain streams throughout the Pyrenean Range (MONTORI & HERRERO, 2004), where spates and hydrological fluctuations are common. It is also found in mountain lakes and subterranean water bodies. The habitats occupied by C. asper make it vulnerable to water perturbations. Newt population size was estimated in a Pyrenean stream during a 21-year fieldwork study (from 1982 to 2002). In November 1982, an unusually high rainfall affected the stream and decreased the newt population as a result of drift caused by the spate (MONTORI, 1988; MONTORI et al., 2008). We collected

DECLINE OR NATURAL FLUCTUATIONS IN CALOTRITON ASPER

demographic data over time to determine the natural dynamics of the population, obtain information on the speed at which the population returns to equilibrium after perturbations, and establish the environmental variables that govern the growth and demographic stochasticity of C. asper populations. In summary, this long-term study is aimed to examine how climatic events modify population size and to identify newt recovery strategies in response to catastrophic events. MATERIALS AND METHODS

and then reaches the Segre River. The stream has riffles, waterfalls, pools and canyons, and receives little direct sunlight during the day because of its northern orientation and vertical surrounding hillsides. The average stream depth and width are 30 cm and 2.5 m, respectively. The water temperature varies from 0.2°C in January to 15.4°C in July. Newts occupy a 4-km stretch of the stream at an altitude between 1150 and 1600 m (MONTORI, 1988). This torrent is considered a characteristic habitat of the Pyrenean newt (CLERGUE-GAzEAU & MARTINEz-RICA, 1978; MONTORI, 1988).

Study area

Newt sampling

The study was carried out since August 1982 till September 2002 in the Pi brook, a stream located in the Cadí range in the Spanish PrePyrenean Mountain Range (Fig. 1). This is one of a series of chains in the Spanish Pre-Pyrenean Mountains that run mostly in parallel to the main Pyrenean axis (Fig. 1). Located in the “Parc Natural del Cadí-Moixeró”, the Pi brook (42° 19' 43.17" N; 1° 45' 18.68" E) descends through the Pi Valley from 1700 to 1100 m above sea level,

Newt sampling was conducted during a 21-year field work (1982-2002, see MONTORI, 1988). Newts were collected by visual observation of active individuals in open water and also by turning rocks. During the first year (August 1982-August 1983), we estimated population size in a 1500-m stretch of stream using capture-markrecapture techniques. We caught newts along the stretch on a monthly basis over that year.

50 km

Figure 1: Localization of study stream (Black square).

MONTORI ET AL.

This intensive method was used to accurately measure population size of C. asper in the study area. To maximize marking efficiency, an intensive initial capture was conducted in August 1982. Each newt was marked by toe clipping. No more than one toe per limb (mostly two or three in total) was clipped. Toeclipping is a standard technique used in amphibian field research. Several studies have found no negative effects of this method (FUNk et al., 2005) and alternative techniques may affect individual survival (e.g. SCHLAEPFER, 1998). MCCARTHy & PARRIS (2004) demonstrated that individual return rates in several frog species were affected only when more than three toes were clipped. Hence, removing two or three toes does not affect newt viability (MONTORI, 1988; MONTORI et al., 2008). After marking the animals, we released them to the same capture or recapture zone. During that period we used beryllium nitrate 0.05 N in order to inhibit toe regeneration (MONTORI, 1988). However, we detected the first regeneration episode in the second study year. For this reason, we stopped the capture-mark-recapture method in 1983 and used transect censuses to estimate annual variation in newt population size thereafter to the end of the study in 2002. From 1983 to 2002 we surveyed a 150-m stretch within the former 1500-m stretch in order to examine temporal variation in population size over 21 years of fieldwork sampling. This 150-m stretch was subdivided in three 50-m segments in order to obtain three different population size estimations for the 1500-m stream, allowing this procedure to calculate the standard error of mean. These estimations were calculated by comparing the observed abundance of newts in September 1982 with

values obtained for each 50-m segment sampled from 1983 to 2002. These three 50-m stretches were chosen on the basis of three conditions: 1) the presence of running water throughout the year, 2) high density of newts, and 3) high density of newts in adjacent segments. To compare inter-annual fluctuations, capture effort was constant along years (i.e. two researchers, same start and end time and zone, and similar hydrological and weather conditions). When meteorological or hydrological conditions changed during the sampling, this survey was excluded and sampling was repeated in the following days. Estimation of annual population size and statistical analyses After the first year of capture-mark-recapture fieldwork in August 1982-August 1983, we applied two statistical methods to estimate population size: 1) after the first and second capture-recapture visits in August and September 1982, population size and 95% confidence interval were estimated by Chapmanâ&#x20AC;&#x2122;s modification of the LincolnPetersen method (CHAPMAN, 1951) for closed populations. Given the low mobility of the species (MONTORI et al., 2008) and the short period between the first and second visits (14 days) we assumed that newt population was closed; 2) for the rest of the year up to August 1983 we used the Jolly-Seber method (JOLLy, 1965; SEBER, 1965) to estimate monthly variation of the number of newts. The inter-annual population growth rate (R) was defined as R = loge (Nt ) - loge (Nt - 1), where Nt is the number of individuals at time t (BERRyMAN, 1999; SIBLy & HONE, 2002). This demographic parameter provides insight

DECLINE OR NATURAL FLUCTUATIONS IN CALOTRITON ASPER

into whether the population remains constant near its carrying capacity (R = 0), increases (R > 0) or decreases (R < 0) in abundance over time (SIBLy & HONE, 2002). From September 1983 to September 2002, adult newts and larvae were counted during each annual sampling. Inter-annual balance (BPOP) was estimated as the number of adults in year i minus the number of adults in year i - 1. Mortality was considered higher than recruitment when BPOP < 0 (population decreases). This assumption assumes that recruitment was offset by natural and waterflow-related mortality, although it is not possible to discern between the proportion of mortality caused by each factor. A stepwise multiple regression was used to model which climate variables were the best predictors of the number of adult newts and larvae found per year, as well as the interannual balance of adult population. For these analyses, the numbers of adults and larvae were log-transformed to linearize regressions. The following meteorological variables were recorded as they were expected to affect stream hydrology and consequently to influence inter-annual C. asper population size variation from 1982 to 2002: annual rainfall, daily thermal oscillation, annual snow, number of days with temperature higher than 30ºC, number of days with temperature lower than -10ºC, number of rainy days, maximal rainfall in 24 h, maximal temperature, minimal temperature, and average temperature. For a census of year i, abiotic variables considered for the analysis were those recorded in the previous year (i - 1). High scores of the “maximal rainfall in 24 h” variable were used as indicator of unusual or extreme rainfall and waterflow occurrence.

Autocorrelation function (ACF) was used to evaluate periodicity and a partial rate correlation function (ACFP) to detect the feedback dimension in the time series of number of adult newts from 1982 to 2002, number of larvae and inter-annual balance. An analysis of the variance (ANOVA) was used to compare the mortality of newts related to biological period in which extreme rainfalls occurred (wintering or activity period) using size population balance as a mortality estimator. Wintering in newts occurred approximately between October 15th and March 15th. Analyses were performed with the Statistica software package (StatSoft, Tulsa, Oklahoma, USA). RESULTS First-year sampling (August 1982-August 1983) During the first study period, 1476 adults were captured, marked and released at the place of capture, and 251 newts were recaptured (including multiple recaptures of the same individual). In the study stretch (1.5 km), for the first capture-recapture period (August 1982September 1982) the estimated population size (mean ± SE) was 4998 ± 948 newts as estimated by the Jolly-Seber method, and 3673 newts (95% confidence interval 3166-4336) as estimated by the Chapman’s method (Table 1). Although there was a wide divergence between the two methods, both gave a good indication of the abundance of newts in the stretch sampled. The unusual rainfall episode in November 1982 (nearly 155 l / m2 in one day) produced a severe flood, and the structure of the stream changed drastically, with a dramatic decrease in newt population (Fig. 2). The estimated survival from September 1982 to April 1983 was 0.263.

MONTORI ET AL.

Table 1: Population size estimations by means of Jolly-Seber and Chapmanâ&#x20AC;&#x2122;s modification of the LincolnPetersen method in the 1.5 km of brook. ni: number of newts captured at time i. ri: number of newly marked newts released at time i. mi: number of newts recaptured at time i. yi: number of newts marked at time i and recaptured at time i + 1. Ni: estimated adult population size. Fi: survival rate from time i - 1 to time i. Bt: net number of individuals entering the population between samplings. SE: standard error. CI: confidence interval. NC: not calculable because of data structure.

Jolly-Seber estimation

Date

Aug 82 Sep 82 Apr 83 May 83 Jun 83 Jul 83 Aug 83

724 546 75 88 80 95 119

Date of last capture Aug 82 Sep 82 Apr 83 May 83 Jun 83 Jul 83

724 438 108 59 11 62 17 51 18 67 15 75 13

5 5 7 6 10

4 2 3 6

2 3 4

1 7

Chapmanâ&#x20AC;&#x2122;s modification of the Lincoln-Petersen method

After the flood, the number of newts decreased and successive monthly surveys (between March and August) did not show a quick population recovery to the number of newts caught in September 1982 (Table 1), even though the capture effort was constant throughout the study

Ni Mean SE

95% CI

183 108 41 4998 948 16 12 1561 472 26 9 975 237 29 7 905 249 28 14 1091 310 44 2317 1327 3673

Mean

0.263 0.754 1 0.923 NC 0.004

0.069 0.206 0.301 0.037

389 -146 18 329 NC -88

3166-4336

period. From July to August 1983, the number of individuals increased as a result of the incorporation of terrestrial subadults into the torrent upon reaching sexual maturity (immature newts of this species have a terrestrial period until sexual maturity, Fig. 3).

Figure 2: Population size of Calotriton asper at the Pi stream (solid line, rhombuses) estimated by transects from 1982 to 2002 and maximal rainfall in 24 h (broken line, squares) occurred before the censuses. R = - 0.539, P < 0.01171, overall standard error of the estimation: 1006.5.

DECLINE OR NATURAL FLUCTUATIONS IN CALOTRITON ASPER

Figure 3: Biological cycle of Calotriton asper at the Pi stream in their first five years of life. H: hatching. W: wintering. M: metamorphosis. SM: sexual maturity.

1982-2002 sampling The annual surveys from 1982 to 2002 also showed a significant decrease in the number of individuals after the sampling in 1982 as a result of the spate that happened in November 1982 (Table 2). In 1984, the number of newts

increased as recently matured newts returned to the aquatic environment (Fig. 3). As in 1982, another episode of extreme rainfall followed by high newt mortality occurred in 1994. After both perturbations, the newt population showed rapid recovery: three years after the 1982 and 1994 floods (Fig. 2). The mean estimated

Table 2: Data used in the study. %M: percentage of males. LV: number of larvae. NPOP: estimated population size. SE: standard error. BPOP: balance of population size between years. RMAX24: maximal rainfall in 24 h (l / m2). DTO: daily thermal oscillation (ºC). GS: snow grams. NDT+30: number of days with temperatures above 30ºC. NDT-10: number of days with temperatures below -10ºC. NRD: number of days with rain. RTOT: total annual rainfall (l / m2). TMAX: maximal year temperature (ºC). TMIN: minimal year temperature (ºC). TAVG: annual average temperature (ºC). year %M LV NPOPa SE BPOP RMAX24 DTO GS NDT+30 NDT-10 NRD RTOT TMAX TMIN TAVG 1982 1983 1984 1985 1986 1987 1988 1989 1990 1991 1992 1993 1994 1995 1996 1997 1998 1999 2000 2001 2002 a

43.8 44.0 50.8 45.0 38.5 46.9 55.1 60.6 51.5 66.4 51.3 51.0 49.7 57.3 56.7 51.4 49.7 52.5 54.4 51.5 51.5

1 4 3 15 26 36 11 19 24 4 18 46 55 1 7 20 45 44 60 54 10

3673 793 2172 1905 2611 3437 3244 3104 3284 3184 3750 3204 3474 1866 3811 3283 4317 5036 5250 5385 4868

125 341 207 210 136 129 53 142 74 280 263 422 172 684 215 278 537 324 779 829

-2844 1379 -266 706 826 -193 -140 180 -100 566 -546 270 -1608 1945 -528 1034 719 214 135 -517

72.0 155.0 44.0 35.0 38.0 42.0 39.0 17.0 47.0 64.0 55.0 24.0 17.0 88.0 37.0 58.0 28.8 53.0 37.2 28.4 39.8

11 15 9 15 12 14 10 6 9 10 5 11 18 12 12 8 10 11 8 10 9

251 98 117 102 154 125 119 106 137 136 176 106 177 105 192 131 105 125 62 70 58

4 2 0 0 2 2 0 0 0 0 0 1 1 1 0 0 0 0 0 4 1

0 5 0 4 1 4 4 0 0 0 0 0 0 1 2 0 0 6 0 5 0

140 109 131 122 139 148 146 132 150 131 151 128 140 147 164 147 127 143 140 122 152

1203 766 935 766 781 1100 789 877 1089 770 1308 774 1012 951 1373 1039 778 1105 947 840 822

32 31 29 29 31 32 29 30 29 30 28 31 39 31 30 29 30 29 29 32 30

-12 -19 -14 -20 -16 -15 -17 -12 -15 -16 -11 -11 -14 -12 -14 -13 -12 -16 -14 -10 -6

8.1 9.2 7.4 8.5 8.4 8.6 8.5 9.1 8.8 8.2 8.2 8.0 9.5 8.5 7.6 9.3 8.8 8.4 8.9 8.9 8.4

Estimation of 1982 made by the Chapman’s method (see Table 1). Estimations from 1983 to 2002 made by contrast with the 1982 value.

MONTORI ET AL.

inter-annual population growth rate (R) calculated after the 21-year surveys was close to zero (0.0063), thereby indicating that the newt population remained near its carrying capacity. The ACF used to evaluate periodicity and the ACFP did not detect significant periodicity in annual population estimated from the 1982 to 1992 series.

variables: number of adults, number of larvae and inter-annual population balance (Table 3). Besides, a set of climatic variables were included in the final stepwise models as predictors of the three population parameters (see Table 3 for variables included in the final steptwise models). The intercept of the regression between the population balance and the maximal annual rainfall in 24 h was close to 50 l / m2 in such period (Fig. 4). We consider that rainfall over this amount may produce a negative population balance. Mortality was higher when extreme rainfall (higher than 50 l / m2) occurred during the wintering period than when it happened during newt activity period (ANOVA: F1,18 = 11.801; P = 0.0029) (Fig. 5).

Abiotic factors determining population size The explained variances of the stepwise multiple regression analyses were higher than 60% and showed that the variable â&#x20AC;&#x153;maximal rainfall in 24 hâ&#x20AC;? was negatively correlated with the three dependent

Table 3: Multiple regression summary for three dependent variables (Log of number of adults, Log of number of larvae, inter-annual population balance). Bold values are significant. All variables contributing to increase the R2 of the model are displayed, regardless of their statistical significance. Variable names are summarized in Table 2. Adults Variable

Beta B (SE) (SE)

Model adjustment

Selected variables

Intercept

3.85 (0.20) RMAX24 -0.62 0.00 (0.14) (0.00) 0.35 0.02 TMIN (0.17) (0.01) 0.23 0.00 RTOT (0.14) (0.00) NDT-10 0.25 0.02 (0.16) (0.01) 0.25 0.05 DTO (0.27) (0.05) R R2 Adjusted R2 Std. Error F d.f. P d.f. = 10

0.8537 0.7287 0.6383 0.11340 8.059 5.15 0.00073

Larvae ta

Variable

Beta B (SE) (SE)

19.16 <0.001 Intercept -4.30 0.001 RMAX24 -0.70 (0.14) 2.08 0.059 TAVG 0.49 (0.14) 1.70 0.112 NDT-10 0.25 (0.14) 1.55 0.140

Population balance ta

-2.08 -1.95 (1.07) -0.01 -4.87 (0.00) 0.44 3.48 (0.13) 0.06 1.75 (0.03)

Variable

Beta (SE)

B (SE)

0.068 Intercept

3556.01 (2300.79) <0.001 RMAX24 -0.70 -23.85 (0.13) (4.34) 0.003 RTOT 2.20 0.39 (0.12) (0.69) -0.32 -623.25 0.10 TAVG (0.12) (235.59) -0.20 -64.74 TMIN (0.13) (41.03)

0.91 0.394

R 0.8190 R2 0.6708 Adjusted R2 0.6127 Std. Error 0.29545 F 11.545 d.f. 3.17 P 0.00023

R R2 Adjusted R2 Std. Error F d.f. P

0.8902 0.7924 0.7370 0.53295 14.314 4.15 0.00005

1.55 0.138 -5.50 <0.001 3.20 0.006 -2.65 0.018 -1.58 0.143

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DISCUSSION The two methods to estimate population size during the first sampling year showed considerable divergence in newt population in September 1982 (4997 and 3673 from Jolly-Seber and Chapman methods, respectively). However, both estimators coincided in indicating the high abundance of newts in the 1.5-km stretch of brook. Our results indicate that maximal rainfall in 24 h (over 50 l / m2) is the main factor influencing newt dynamics. Despite this high newt density, periodic floods in the stream due to extreme rainfall produced marked inter-annual fluctuations in population size. The rapid population recovery detected after floods suggests that this species is highly adapted to survive in running-water habitats affected by periodic and severe floods. Demographic effects of extreme rainfall episodes producing catastrophic drift and high population decline, mainly of larvae, has

Figure 4: Linear regression adjust (Âą 95% confidence interval) of maximal rainfall in 24 h with (a) inter-annual population balance, (b) estimated population size, and (c) number of larvae. See table 2 for abbreviators.

Figure 5: Inter-annual population balance when rainfall occurs during wintering or during the active cycle. Boxes indicate standard error and bars indicate standard deviation of the mean.

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been observed previously (M端LLER, 1954; PETRANkA & SIH, 1986; BARRET et al., 2010). For example, PETRANkA & SIH (1986) analysed the survival rate of Ambystoma texanum larvae in relation to various biotic and abiotic factors and reported that up to 90% of young larvae do not survive to periods of high waterflow. We detected a negative relation between abundance of C. asper larvae in the Pi Valley and maximal rainfall in 24 h. These waterflows cause high larval mortality, thereby reducing population recruitment as they determined the number of adult newts present in the stream in subsequent years. Similar findings have been reported by BARRETT et al. (2010) who concluded that the frequency and magnitude of waterflows, which then lead to decreased larval density, were the best descriptors to explain stream salamander decline in urban areas. Moreover, experimental data support that spate frequency (which was highly correlated with spate magnitude) was the best predictor of salamander density across an urban-rural stream gradient (BARRETT, 2009; BARRETT et al., 2010). These examples suggest that salamanders cannot maintain their position in the stream during high flows and that recolonization after subsidence of intensive flows is low. In addition, THIESMEIER (1992) demonstrated that stream dwelling species like C. asper or the fire salamander Salamandra salamandra drifted less than pond species (e.g. Triturus spp). THIESMEIER & SCHUHMACHER (1990) observed a catastrophic drift of larval S. salamandra. Drift of the youngest larvae was determined by the behaviour of spawning females and by stream conditions, whereas the largest larvae

drift occured when their retreats became too small to provide an adequate food supply. However, the C. asper population is not food-limited in the Pi Valley stream (MONTORI, 1991, 1992). BAUMGARTNER et al., (1999) used field data to argue that larval S. salamandra preferred lower current speeds within a given stream, as they found fewer specimens in streams with higher mean stream discharge. Other authors have documented species that are susceptible to spates; however, those studies were performed largely in the context of hydrological changes that occur with increasing stream order (DUDGEON, 1993; BAUMGARTNER et al., 1999; LEIPELT, 2005). For example, LEIPELT (2005) used artificial streams to evaluate the response of four species of Odonata to a high-flow stream environment. In that study, the two species most susceptible to drift were found in lower order streams that were less prone to spates. This author interpreted these findings as evidence that hydrological factors shape species distribution and survivorship. Collectively, these studies support the notion that spate frequency and / or magnitude influence the abundance and distribution of stream organisms. Extreme rainfall produces changes in the bed structure and in the bank morphology of streams. However, in a previous study (MONTORI et al., 2008) we demonstrated that the distribution pattern of newts along the stream was not influenced by these structural changes associated with spates. Thus the question remains as to how C. asper maintains or recovers population levels in habitats frequently affected by severe waterfloods. During the 1982-2002 sampling, we recorded two episodes of extreme rainfall, in

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1982 and 1994. After these episodes, increased waterflow produced high mortality; however, the population recovered rapidly, despite being a species with a small clutch size if compared to other European newt species (CLERGUE-GAzEAU, 1971; MONTORI, 1988, 1992; MONTORI & HERRERO, 2004). In fact, population quickly recovered values similar to those observed before the flood. This observation is attributed to the fact that subadults are almost exclusively terrestrial until sexual maturity (MONTORI & HERRERO, 2004). Hence, subadult C. asper would act as a population stock to facilitate population recovery in a few years after waterfloods. As spates are common in stream environments, this recovery strategy would allow this species to overcome these catastrophic episodes. The findings of our 19822002 sampling are consistent with this explanation: the number of newts severely decreased the first year after each catastrophic drift, whereas in the second year the return of sexually mature subadults produced a parallel increase in the population. The decrease in numbers observed in the third year after the flood can be explained by the absence of recruitment as those larvae that should have reached sexual maturity died during the drift caused by the spate. This pattern was observed in the two occasions when extreme rainfall was registered. The hypothesis of a population recover by means of upstream dispersal can be rejected from the results of a previous study demonstrating that no upstream or downstream movements occur in this population (MONTORI et al., 2008). These data verify for the first time the relevance of the terrestrial life period of newts for the recovery of populations in unpredicta-

ble aquatic environments frequently affected by extreme rainfall. In this regard, it is particularly important to determine whether the increases in extreme rainfall caused by climatic change predicted by some authors (BATES et al., 2008) would produce higher vulnerability and risk of extinction for C. asper stream populations. Interestingly, mortality was higher when extreme rainfall occurred during the overwinter period than during the activity period. This result indicates that during the active cycle, C. asper has developed behavioural strategies to avoid waterflow drift. In fact, MONTORI (2008) reported that Pyrenean newts predict waterflows, although the exact mechanism by which they do this is unknown. According to the field observations made by CLERGUE-GAzEAU & MARTíNEz-RICA (1978), in running waters, when a spate starts, newts leave refuges and move to banks where water velocity is lower. In periods of high waterflow, there are observations of newts leaving the water (MONTORI, 2008). These behavioural strategies are not displayed in winter as newts are buried in stream banks. In conclusion, our findings further demonstrate the adaptability of Pyrenean newts to extreme environments, such as mountain streams where climatic conditions plus periodic waterflows may preclude the dynamic of amphibian species and the presence of richer amphibian communities. Acknowledgement The authors thank Marc Grau, Juanjo Francesch, Lluís Serra-Cobo, Antonia Calle and Jordi Serra-Cobo for field assistance. We thank the Parc Natural del Cadí-Moixeró for

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allowing us to conduct this study. Permission for newt capture was granted by the Departament de Medi Ambient de la Generalitat de Catalunya. REFERENCES BARRETT, k. (2009). Stream-Breeding Amphibian Responses to Land Use Disturbances. Ph.D. dissertation, Auburn University, Auburn, Alabama, USA. BARRETT, k.; HELMS, B.S.; GUyER, C. & SCHOONOVER, J.E. (2010). Linking process to pattern: Causes of stream-breeding amphibian decline in urbanized watersheds. Biological Conservation 143: 1998-2005. BATES, B.C.; kUNDzEWICz, z.W.; WU, S. & PALUTIkOF, J.P. (2008). Climate Change and Water. Series: IPCC Technical Paper, vol. IV. Intergovernmental Panel on Climate Change Secretariat, Geneva, Switzerland. BAUMGARTNER, N.; WARINGER, A. & WARINGER, J. (1999). Hydraulic microdistribution patterns of larval fire salamanders (Salamandra salamandra salamandra) in the Weidlingbach near Vienna, Austria. Freshwater Biology 41: 31-41. BEEBEE, T.J.C. & GRIFFITHS, R.A. (2005). The amphibian decline crisis: A watershed for conservation biology? Biological Conservation 125: 271-285. BERRyMAN, A.A. (1999). Principles of Population Dynamics and Their Application. Stanley Thornes, Cheltenham, Uk. BIEk, R.; FUNk, W.C.; MAXELL, B.A. & MILLS, L.S. (2002). What is missing in amphibian decline research: insights from ecological sensitivity analysis. Conservation Biology 16: 728-734.

BILBy, R.E. & LIkENS, G.E. (1980). Importance of organic debris dams in the structure and function of stream ecosystems. Ecology 61: 1107-1113. BLAUSTEIN, A.R.; WAkE, D.B. & SOUSA, W.P. (1994). Amphibian declines: judging stability, persistence, and susceptibility of populations to local and global extinctions. Conservation Biology 8: 60-71. CHAPMAN, D.G. (1951). Some properties of the hypergeometric distribution with applications to zoological sample censuses. University of California Publications in Statistics 1: 131-160. CLERGUE-GAzEAU, M. (1971). L'euprocte pyrénéen. Conséquence de la vie cavernicole sur son développement et sa reproduction. Annales de Spéléologie 26: 825-960. CLERGUE-GAzEAU, M. & MARTINEz-RICA, J.P. (1978). Les différents biotopes de l’urodèle pyrénéen, Euproctus asper. Bulletin de la Société d’Histoire Naturelle de Toulouse 114: 461-471. DUDGEON, D. (1993). The effects of spateinduced disturbance, predation, and environmental complexity on macroinvertebrates in a tropical stream. Freshwater Biology 30: 189-197. FUNk, W.C.; DONNELLy, M.A. & LIPS, k.R. (2005). Alternative views of amphibian toe-clipping. Nature 433: 193. HOULAHAN, J.E.; FINDLAy, C.S.; SCHMIDT, B.R.; MEyER, A.H. & kUzMIN, S.L. (2000). Quantitative evidence for global amphibian population declines. Nature 404: 752-755. JOLLy, G.M. (1965). Explicit estimates from capture-recapture data with both death and immigration-stochastic model. Biometrika 52: 225-247.

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LANCASTER, J.; HILDREW, A.G. & TOWNSEND, C.R. (1990). Stream flow and predation effects on the spatial dynamics of benthic invertebrates. Hydrobiologia 203: 177-190. LANCASTER, J. & HILDREW, A.G. (1993). Characterizing in-stream flow refugia. Canadian Journal of Fisheries and Aquatic Sciences 50: 1663-1675. LEIPELT, k.G. (2005). Behavioural differences in response to current: implications for the longitudinal distribution of stream odonates. Archiv für Hydrobiologie 163: 81-100. MAy, R.M.; CONWAy, G.R.; HASSELL, M.P. & SOUTHWOOD, T.R.E. (1974). Time delays, density-dependence and single-species oscillations. Journal of Animal Ecology 43: 747-770. MCCARTHy, M.A. & PARRIS, k.M. (2004). Clarifying the effect of toe clipping on frogs with Bayesian statistics. Journal of Applied Ecology 41: 780-786. MEyER, A.H.; SCHMIDT, B.R. & GROSSENBACHER, k. (1998). Analysis of three amphibian populations with quarter-century long time-series. Proceedings of the Royal Society B 265: 523-528. MONTORI, A. (1988). Estudio Sobre la Biología y Ecología del Tritón Pirenaico Euproctus asper (Dugès 1852) en la Cerdanya. Ph.D. dissertation, University of Barcelona, Barcelona, Spain. MONTORI, A. (1991). Alimentación de los adultos de Euproctus asper (Dugès 1852) en la montaña media del prepirineo catalán (España). Revista Española de Herpetología 5: 23-36. MONTORI, A. (1992). Alimentación de las larvas de tritón pirenaico, Euproctus asper, en el prepirineo de la Cerdaña, España. Amphibia-Reptilia 13: 157-167.

MONTORI, A. (2008). El tritó pirinenc (1ª part). El Picot Negre, Revista Informativa del Parc Natural del Cadí Moixeró 11: 9-11. MONTORI, A. & HERRERO, P. (2004). Caudata, In M. García-París, A. Montori & P. Herrero (eds.) Amphibia. Lissamphibia. Series: Fauna Ibérica, vol. 24 (M.A. Ramos, coord.). Museo Nacional de Ciencias Naturales, Madrid, Spain. MONTORI, A.; LLORENTE, G.A. & RICHTERBOIX, A. (2008). Habitat features affecting the small-scale distribution and longitudinal migration patterns of Calotriton asper in a Pre-Pyrenean population. Amphibia-Reptilia 29: 371-381. MüLLER, k. (1954). Investigations on the organic drift in north Swedish streams. Reports of the Institute of Fresh-water Research, Drottningholm 35: 133-148. J.H.k.; SCOTT, D.E.; PECHMANN, SEMLITSCH, R.D.; CALDWELL, J.P.; VITT, L.J. & GIBBONS, J.W. (1991). Declining amphibian populations: the problem of separating human impacts from natural fluctuations. Science 253: 892-895. PETRANkA, J.W. & SIH, A. (1986). Environmental instability, competition, and density-dependent regulation of a salamander in streams. Ecology 67: 729-736. REED, J.M. & BLAUSTEIN, A.R. (1995). Assessment of “nondeclining” amphibian populations using power analysis. Conservation Biology 9: 1299-1300. RESH, V.H.; BROWN, A.V.; COVICH, A.P.; GURTz, M.E.; LI, H.W.; MINSHALL, G.W.; REICE, S.R.; SHELDON, A.L.; WALLACE, J.B. & WISSMAR, R.C. (1988). The role of disturbance in stream ecology. North American Benthological Society 7: 433-455.

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SALVIDIO, S. (2011). Stability and annual return rates in amphibian populations. Amphibia-Reptilia 32: 119-124. SCHLAEPFER, M.A. (1998). Use of a fluorescent marking technique on small terrestrial anurans. Herpetological Review 29: 25-26. SEBER, G.A.F. (1965). A note on the multiplerecapture census. Biometrika 52: 249-259. SIBLy, R.M. & HONE, J. (2002). Population growth rate and its determinants: an overview. Philosophical Transactions of the Royal Society B 357: 1153-1170. STUART, S.N.; CHANSON, J.S.; COX, N.A.; yOUNG, B.E.; RODRIGUES, A.S.L.; FISCHMAN, D.L. & WALLER, R.W.

(2004). Status and trends of amphibian declines and extinctions worldwide. Science 306: 1783-1786. THIESMEIER, B. (1992). Comparative experiments on larval drift of five European urodelan species in a water channel – preliminary results, In z. korsós & I. kiss (eds.) Proceedings of the Sixth Ordinary General Meeting SEH Budapest 1991. Societas Europaea Herpetologica, Budapest, Hungary, pp. 439-442. THIESMEIER, B. & SCHUHMACHER, H. (1990). Causes of larval drift of the fire salamander, Salamandra salamandra terrestris, and its effects on population dynamics. Oecologia 82: 259-263.

Basic and Applied Herpetology 26 (2012): 57-71

Distribution review, habitat suitability and conservation of the endangered and endemic Moroccan spadefoot toad (Pelobates varaldii) Philip de Pous1, 2, 3, *, Wouter Beukema3, 4, Diederik Dingemans3, David Donaire5, Philippe Geniez6, El Hassan El Mouden7 Escola Tècnica Superior d'Enginyeria Agrària, Departament de Producció Animal (Fauna Silvestre), University of Lleida, Lleida, Spain. Institute of Evolutionary Biology (CSIC-UPF), Barcelona, Spain. 3 Society for the Preservation of Herpetological Diversity, Den Haag, the Netherlands. 4 ITC, Faculty of Geo-Information Science and Earth Observation, University of Twente, Enschede, the Netherlands. 5 Calle Mar Egeo 7, 11407 Jerez de la Frontera, Cádiz, Spain. 6 Biogéographie et Ecologie des Vertébrés, Ecole Pratique des Hautes Etudes, UMR 5175, CEFE-CNRS, Montpellier, France. 7 Université Cadi Ayyad, Faculté des Sciences Semlalia, Laboratoire Biodiversité et Dynamique des Ecosystèmes, Marrakech, Morocco. 1 2

* Correspondence: Escola Tècnica Superior d'Enginyeria Agrària, Departament de Producció Animal (Fauna Silvestre), University of Lleida, Lleida, Spain. E-mail: philipdepous@gmail.com

Received: 12 August 2011; received in revised form: 21 February 2012; accepted: 23 April 2012.

The Moroccan spadefoot toad (Pelobates varaldii) has received little scientific attention since its discovery. Currently, P. varaldii is listed as Endangered on the IUCN Red List due to a multitude of threats, while its distribution is partially unknown and fragmented. The current study addresses distribution, threats and the potential niche using ecological niche modelling, while emphasizing conservation strategies and immediate actions. The distribution of P. varaldii can be divided into four disjunct areas, at least two of which consist of small populations. The largest threats to P. varaldii include the transformation of habitat and breeding ponds into agricultural and industrial areas, the pollution of breeding ponds due to extensive livestock pasturing and the possible expansion of Procambarus sp. into the Mamora cork oak forest. Additional threats constitute a reduction of gene flow and loss of genetic variability as a result of habitat fragmentation. The ecological niche models (ENMs) of P. varaldii revealed fundamental environmental conditions along parts of the northeast Moroccan Atlantic and Mediterranean coastline. The species mainly inhabits well vegetated areas upon Quaternary soils at low altitudes. Proposed conservation actions include the development of a biannual monitoring program, identification and designation of protected areas within the distribution and development of a management plan for Mamora forest. Key words: Anura; conservation; habitat suitability; Maxent; Morocco; NDVI. Revisión de la distribución, idoneidad del hábitat y conservación del amenazado endemismo sapo de espuelas marroquí (Pelobates varaldii). Desde su descubrimiento, el sapo de espuelas marroquí (Pelobates varaldii) ha recibido poca atención por parte de la comunidad científica. Como consecuencia de multitud de amenazas, P. varaldii está actualmente catalogado como En Peligro de Extinción en la lista roja de la UICN, siendo su distribución muy fragmentada y no completamente conocida. El presente estudio analiza su distribución, amenazas y nicho potencial utilizando el modelado de nicho ecológico al tiempo que se enfatiza en posibles estrategias de conservación y acciones inmediatas. La distribución de P. varaldii puede dividirse en cuatro áreas disyuntas, al menos dos de las cuales corresponden a poblaciones de pequeño tamaño. Las mayores amenazas para P. varaldii incluyen la transformación de su hábitat y medios de reproducción en áreas agrícolas e industriales, la contaminación de sus charcas de reproducción como consecuencia de la ganadería extensiva y la posible expansión de Procambarus sp. hacia el alcornocal de Mamora. La reducción del flujo genético y pérdida de variabilidad genética como consecuencia de la fragmentación del hábitat constituyen amenazas adi-

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cionales. Los Modelos de Nicho Ecológico (ENMs) de P. varaldii revelaron determinadas condiciones ambientales fundamentales a lo largo de algunos sectores de las costas atlántica y mediterránea del nordeste de Marruecos. La especie habita principalmente en áreas con vegetación ubicadas sobre sustratos cuaternarios a baja altitud. Las acciones de conservación propuestas incluyen el desarrollo de un programa de seguimiento bianual, la identificación y designación de áreas protegidas dentro del rango de distribución y el desarrollo de un plan de gestión del alcornocal de Mamora. Key words: Anura; conservación; idoneidad del hábitat; Maxent; Marruecos; NDVI.

The Moroccan spadefoot toad (Pelobates varaldii) was first mentioned in the literature by PELLEGRIN (1925) under the denomination Pelobates cultripes (Cuvier, 1829) before it was described as a separate species by PASTEUR & BONS (1959), based on morphological and osteological characters. Subsequently, the species received little scientific attention until BUSACk et al. (1985) confirmed differentiation between P. varaldii and P. cultripes based on electrophoretical and immunological analyses, and suggested the species to have diverged since the Early Pliocene. In recent years, GARCíA-PARíS et al. (2003) and VEITH et al. (2006) found a wellsupported sister taxon relationship between P. varaldii and P. cultripes, which, according to their results, likely diverged after the collapse of the Gibraltar land bridge during the end of the Messinian salinity crisis, approximately 5.3 millions of years ago (kRIJGSMAN et al., 1999). The distribution of P. varaldii was long considered to be confined to a rectangle delimited by the towns Mehdia, Sidi Slimane, khemisset and Rabat, until DORDA DORDA (1984) recorded the species 20 km south of Larache, thus expanding the range of the species about 70 km northwards. A few years later, DESTRE et al. (1989) significantly increased the range of the species southwards to Bir Jdid, the mouth of the Oued Nefifikh and Oualidia, as well as northwards into the town of Larache. BONS & GENIEz (1996) provided an overview of all the

known localities up to that point and provided the first detailed distribution map (51 unique localities). The confirmation of the species near Oualidia, with the finding of live specimens in 2001 by CROCHET & GENIEz (2003) and in 2007 (J.M. Thirion, A. Portheault, P. Evrard, personal communication) not only confirmed the records of DESTRE et al. (1989), but also definitely increased the range of P. varaldii 190 km southwards. More recently, GUzMÁN et al. (2007) expanded the range of the species several kilometres eastwards with records just south of Ouezzane, while LAPEñA et al. (2011) reported the species between Asilah and Tangier, extending the distribution around 50 km northwards. Despite these recent range extensions, P. varaldii is currently listed as Endangered (B2ab(iii)) on the IUCN Red List because its area of occupancy is probably less than 500 km2, the distribution is severely fragmented, and there is a continuing decline in the extent and quality of its habitat in Morocco (SALVADOR et al., 2004; PLEGUEzUELOS et al., 2010). Additionally, the species does not occur in any protected area (DE POUS et al., 2011). The main threats to P. varaldii include the transformation of habitat and breeding ponds into agricultural and industrial areas, intensification of livestock pasturing, including the pollution of stagnant waters with livestock droppings (e.g. kNUTSON et al., 2004), while arable agriculture may be lea-

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ding to the loss or disturbance of the sandy substrate, which the species is strongly associated with. Populations of P. varaldii are now often restricted to the direct vicinity of temporary ponds, and those remaining in permanent water bodies are being eliminated through the presence of the predatory mosquito fish Gambusia holbrooki (SALVADOR et al., 2004; authors' personal observation). The Evolutionary Distinctive and Globally Endangered (EDGE) program of the zoological Society of London recently listed P. varaldii on place 36 of their EDGE global amphibian top 100 (ISAAC et al., 2007), making a review of the species distribution, conservation status and range wide threats highly preferred. In this paper we aim at: (1) reviewing the complete distribution of P. varaldii, (2) identifying threats, (3) analysing habitat suitability and (4) emphasizing conservation strategies. MATERIALS AND METHODS

Additionally, adults were searched for during night time surveys (visual encounter surveys and road surveys). GPS coordinates (WGS 1984) of each locality were taken and descriptions of habitat, possible threats and population status were noted. A total of 220 distribution records were assembled from literature and fieldwork. The distribution records went through a process of filtering such as removing duplicate records within unique grid cells in ENMtools 1.3 (WARREN et al., 2010) and thinning of dense clusters using a kernel density grid as implemented in the Java program OccurrenceThinner v.1.0.2 (VERBRUGGEN, 2012). After filtering, a total of 43 distribution records were used for ENM. Most of these records (N = 41) were collected in the field using GPS devices while additional records (N = 2) were derived from GUzMÁN et al. (2007) and LAPEñA et al. (2011).

Species occurrence records

Ecological niche modelling

From December 2008 until April 2009 and from January until the end of February 2010, the entire distribution range of P. varaldii was visited to collect data on distribution and threats. Suitable areas characterized by presence of temporary ponds were selected with the use of satellite images (Google Earth®), preliminary ecological niche models (ENMs) and maps of the area (1:50 000). Larvae were sampled by standardized dip netting of temporary ponds using a 55 x 70 cm dip net with a mesh width of 3 mm. Dip netting is a standard technique used to sample amphibian assemblages in lentic habitats (SHAFFER et al., 1994) and sampling of larvae is in the case of such a rare, nocturnal and fossorial species more appropriate.

Initially, a total of 19 BioClim variables and altitude were downloaded from the WorldClim database version 1.4 (HIJMANS et al., 2005a) to form the climatic dataset (HIJMANS et al., 2005b) at a resolution of 30 arc seconds (nearly 1 x 1 km). Additionally, a categorical normalized difference vegetation index (NDVI) and superficial geology were obtained from BEUkEMA et al. (2010) and downloaded from PERSITS et al. (1997), respectively. Collinearity of the variables was measured with Pearson’s correlation coefficient in ENMtools 1.3. A total of 11 variables, all of which had a correlation degree lower than 0.75 (Pearson coefficient) were retained. Selection of predictor variables was

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based on ecological understanding of the species (e.g. rainfall in the breeding season). The models were generated by the presence/background algorithm Maxent, version 3.3.3k (PHILLIPS et al., 2006; PHILLIPS & DUDík, 2008). It has been shown that Maxent produces high quality predictions that are often more successful when evaluated and compared with other predictive models (HERNANDEz et al., 2006; JIMéNEz-VALVERDE et al., 2008; GIOVANELLI et al., 2010). Additionally, Maxent has a successful prediction power even when using low sample sizes (PEARSON et al., 2007; WISz et al., 2008). This algorithm uses environmental parameters in combination with geographical coordinates in order to predict the distribution of the species of interest. Maximum entropy is achieved by the constraint that the expected value for each variable under the estimated distribution has to match its empirical average – the mean value of a random set of coordinates within the distribution (PHILLIPS et al., 2006) –. In other words, the model minimiFigure 1: Overview map of the kingdom of Morocco with an identification of altitude (darker colours indicating higher altitude) and the main topographic and geographic features.

zes the relative entropy between two probability densities (one from the presence data and one from the landscape) defined in covariate space (ELITH et al., 2010). The model output displays the relative occurrence probability of a species within the grid cells of the study area. Maxent was used with default settings (Convergence threshold = 0.00001, maximum number of iterations = 500 and bj = 1) while partitioning the geographical records between training and test samples (75% and 25% respectively). This technique has been proven to achieve high predictive accuracy (PHILLIPS & DUDík, 2008). Several studies have recently addressed the importance of selecting pseudo-absence or background locations in ENM (e.g. VANDERWAL et al., 2009; ANDERSON & RAzA, 2010). Moreover, some of these studies reported that using very large areas for model calibration, especially if the species is absent from these areas, can result in serious ramifications for predictions and performance of ENMs (VANDERWAL et al., 2009; ANDERSON & RAzA, 2010). We therefore

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followed the background approach of WEBBER et al. (2011) and THOMPSON et al. (2011) and downloaded kรถppen-Geiger polygons from the CliMond version 1.1 database (kRITICOS et al., 2012). Subsequently, models were projected onto a larger area (Fig. 1). Models were created using three different combinations of environmental variables: (A) bioclimatic variables + NDVI + geology, (B) bioclimatic variables + NDVI, and (C) bioclimatic variables only. The average of ten pseudo-replicated models with randomly selected test samples was used to produce habitat suitability models, which were plotted in logistic format. The final models were reclassified in ArcGIS 10 (ESRI, Redlands, California, USA) into binary presence-absence maps based on the lowest presence threshold (LPT). All models were tested with receiver operating characteristics (ROC) curve plots, which plot the truepositive rate against the false-positive rate. The average area under the curve (AUC) of the ROC plot of ten models was taken as a measu-

re of the overall fit of each model. Due to the fact that Maxent operates only with presence records, the AUC is calculated using pseudoabsences chosen at random from the study area (PHILLIPS et al., 2006). The AUC values range between 0.5 (highly unsuitable) and 1.0 (highly suitable) and display the probability that a randomly chosen presence site will be ranked above a randomly chosen absence site (PHILLIPS et al., 2006). Models with AUC values above 0.75 are considered useful (ELITH, 2002). RESULTS Distribution and conservation status The distribution of P. varaldii can be divided into four main areas (Fig. 2): (1) The Mamora cork oak (Quercus suber) forest and surroundings, (2) the area around Larache, including the Loukkos basin and southwards along the road to ksar-el-kbir, as well as inside a royal hunting reserve northeast of Figure 2: Distribution area of Pelobates varaldii in north-western Morocco with occupied UTM grids (10 x 10 km) from literature (grey points) and current study (black points).

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Figure 3: Detail of the distribution of Pelobates varaldii in the Mamora cork oak forest and surroundings (north-eastern Morocco). Black triangles indicate both newly discovered and literature records.

Moulay Bousselham, (3) the area between Asilah and Tangier, and (4) the Doukkala region between Rabat and Oualidia. (1) During the 2008-2010 surveys in the Mamora cork oak forest more than one hundred and fifty temporary ponds distributed throughout the entire forest have been checked for presence of amphibian larvae. Pelobates varaldii larvae were detected in over 95% of the ponds, making it the most common amphibian species in the forest (Fig. 3). Many of the available breeding ponds (30%) show human modifications such as deepening and ploughing, and surrounding habitats show high levels (55%) of disturbance. Most of the ponds (> 80%) show clear disturbance as a result of livestock pasturing. (2) The 2008-2010 surveys detected and confirmed the presence of P. varaldii south of Larache on the way to ksar-el-kbir,

approximately 20 km southeast of Larache, as well as in a Royal hunting reserve, between Moulay Bousselham and ksar-el-kbir. Threats in this part of the distribution are comparable to Mamora forest, but include: (i) the predation by G. holbrooki in some permanent water bodies where P. varaldii breeds; (ii) the recent (2007-2010) transformation of ponds into agricultural fields; and (iii) the discovery of the exotic crayfish species Procambarus (Ortmannicus) lophotus (J. Cooper, personal communication) in a P. varaldii breeding pond (35.078269째, -6.069084째) in 2007. (3) A single locality south of Tangier airport was detected on 2 February 2009. The locality consists of a single breeding pond inside a Pinus pinea forest fragment and larvae density estimates in 2009 (catch per unit effort, CPUE) indicate a low population size. In January 2010, the

DISTRIBUTION AND CONSERVATION OF PELOBATES VARALDII

Table 1: Response values percent contribution and permutation importance (in %) of the climatic and categorical predictor variables for the ecological niche models of Pelobates varaldii from north-western Morocco (see text for the description of variables). Variable A Altitude NDVI Geology BIO3 BIO4 BIO5 BIO9 BIO10 BIO15 BIO16 BIO17

38.9 32.9 9.7 1.3 5.1 0.2 0 0 2.4 9.1 0.4

Percent contribution B C 41.9 38.8 0.5 5.4 0.2 0 0.2 3.1 9.6 0.1

47.4 3 16.3 2 0.1 0.4 2.4 27.9 0.6

locality was revisited and only very few, mostly malformed larvae could be caught. (4) During the 2008-2010 surveys, no new populations could be detected in the region between Rabat and Oualidia, except for some populations just south of Mamora forest. Several expeditions into the Ben Slimane cork oak forest, with many temporary ponds present, did not result in a single observation of P. varaldii. In 2009, the Oualidia region was extensively surveyed for eight days, in which we searched for both tadpole presence and adults, but no specimens could be observed. On 16 February 2010, during rainy weather conditions, we confirmed the presence of the species in this region, with the finding of twelve individuals on the road, about 16 km west of Oualidia. Ecological niche modelling The final set of environmental predictor variables for the ENMs consisted of: altitude, isothermality (BIO3), temperature seasona-

Permutation importance A B C 24.5 23 2.2 1.2 26.2 1.6 0 1.8 11.2 2.2 5.6

33.6 20.6 0.3 19 1.9 0 8.2 9 5.5 2

10 3.1 38 9.1 0 2.7 5.9 29.2 2

lity (BIO4), maximum temperature of the warmest month (BIO5), mean temperature of the driest quarter (BIO9), mean temperature of the warmest quarter (BIO10), precipitation seasonality (BIO15), precipitation of the wettest quarter (BIO16), precipitation of the driest quarter (BIO17), NDVI and superficial geology. Maxent produced models of high predictive accuracy, according to the average test AUC of the three models (A = 0.969 ± 0.013, B = 0.973 ± 0.015, C = 0.964 ± 0.013). The main predictor variables (percent contribution and permutation importance) of P. varaldii can be found in Table 1. Visual interpretation of the response curves revealed that the probability of presence of the species decreases with increasing altitude, whereas it increases with higher precipitation of the wettest quarter. The species has a high probability of presence in eight NVDI classes (75 classes in the study area), all of which have a high yearround green biomass and occurs mainly on sandy Quaternary and karstic Tertiary

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DISCUSSION Distribution and conservation status

Figure 4: Habitat suitability maps (potential niche models) of Pelobates varaldii created using (A) climate variables + NDVI + geology, (B) climate variables + NDVI, and (C) climate variables only. Models are above the average lowest presence threshold.

soil deposits (35 classes in the study region). The models of P. varaldii revealed a small and fragmented potential niche along parts of the northeast Moroccan coastline until Essaouira, as well as at some areas along the Mediterranean coast to the Algerian border except for model C (Fig. 4).

The Mamora forest forms the core distribution area of P. varaldii and harbours the largest populations. Despite that P. varaldii is still widely distributed throughout the forest, substantial habitat modifications such as overgrazing, logging, exotic tree plantations, transformation to agricultural land and industrial development are severely threatening the biodiversity of this unique habitat (e.g. CHERkAOUI et al., 2009). The Mamora forest is highly fragmented (CHERkAOUI et al., 2009) and the original cork oak forest has shown a major decrease over the last centuries (HOOkER & BALL, 1878; OLIVER, 1986; NAFAA & WATFEH, 2000). Populations of P. varaldii are now often confined to the direct surroundings of temporary ponds and populations are being fragmented due to unfavourable habitat and agricultural fields, leading to potential loss of gene flow and genetic variability. Gambusia holbrooki was only detected in permanent streams and water bodies. Since P. varaldii mostly breeds in temporary ponds, elimination through these predatory fishes therefore seems unlikely in the Mamora region. The distribution between Mamora forest and the localities in the area around Larache is not well understood. The ENMs revealed substantial areas of suitable habitat in this region but fieldwork and literature did not confirm the presence of P. varaldii in this area. Literature records are limited to EL HAMOUMI & HIMMI (2010) who found P. varaldii in the Loukkos basin, and GUzMĂ N et al. (2007) who recorded the species just south of Ouezzane.

DISTRIBUTION AND CONSERVATION OF PELOBATES VARALDII

During our surveys we visited the latter area on two different occasions but we could not confirm the presence of P. varaldii in this area. The newly discovered population south of Tangier airport is under severe threat by the development of a nearby beach resort and immediate actions in the form of habitat protection are needed to prevent this population from local extinction. The population seems to be confined to a single breeding pond in a small forest fragment and CPUE analysis indicates a low population size (unpublished data). The recent discovery of the species 6 km southwards of the Tangier airport locality (LAPEñA et al., 2011) might indicate a wider distribution in this area and this requires further investigation. The area between Rabat and Oualidia has been extensively explored by herpetologists during the past decades (BONS & GENIEz, 1996). Despite this, very few records of P. varaldii from this region exist. The cork oak forest surrounding Ben Slimane harbours substantial amounts of original habitat and over 600 temporary ponds are present in this region covering a total surface area of 1994 ha (RHAzI et al., 2009). The multiple surveys into this area surprisingly did not produce any records of P. varaldii presence despite extensive dip netting of many temporary ponds. The presumed absence in this area is likely a result of unsuitable microhabitat and the lack of sandy areas was confirmed during surveys. The forest surrounding Ben Slimane is characterized by thick bushy undergrowth, and the additional presence of many small streams makes this area unsuitable for P. varaldii. The distribution in the area between Rabat and Oualidia remains not well understood. Niche models indicate few suitable areas in this region and century long herpeto-

logical explorations have produced very few sightings (BONS & GENIEz, 1996). Although most of these records came from bone remains in the rejection pellets of the barn owl (Tyto alba) and the African marsh owl (Asio capensis), this method can be considered a reliable prospection mode for P. varaldii because the radius of operation of T. alba rarely exceeds 4 km (MICHELAT & GIRAUDOUX, 1991). The Oualidia population is confined to a very small region of suitable habitat with several temporary ponds present. Despite extensive sampling of these ponds, no larvae could be detected in both 2009 and 2010. It remains unclear if the Oualidia population faces additional threats. yUS RAMOS & CABO HERNÁNDEz (1986) reported the presence of P. cultripes near the Spanish enclave Melilla in north-eastern Morocco, but despite extensive herpetological exploration in this area, this occurrence was never confirmed. Several suitable areas are present in the region (Fig. 4) and therefore Pelobates sp. might indeed have inhabited the area. The origin of this possible and presumed extinct population remains unclear and can constitute both an introduction of P. cultripes from Spain, a relict population of P. varaldii following a range expansion during the glacial period or even a result of multiple colonization routes into Morocco from Spain. The area surrounding Melilla has seen major industrial development in recent years and the shrinkage of suitable lowland habitat as a consequence might have driven this population to local extinction. Another plausible scenario is that yUS RAMOS & CABO HERNÁNDEz (1986) misidentified other amphibians with P. cultripes, but due to the clear morphological differences this seems unlikely.

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Several currently identified threats have also been identified by previous authors (see introduction), whereas additional future threats to P. varaldii might be the recent arrival of the chytrid fungus (Batrachochytrium dendrobatidis) in Morocco (EL MOUDEN et al., 2011), the reduction of gene flow and loss of genetic variability as a result of habitat fragmentation (e.g. DIXO et al., 2009), the possible effects of climate change (DRIOUECH et al., 2010) and the expansion of the recently detected invasive crayfish Procambarus (Ortmannicus) lophotus (e.g. CRUz et al., 2006). Ecological niche modelling Century long herpetological explorations in most parts of Morocco, but especially the Tingitana Peninsula, Melilla region and central Atlantic coastal areas have never documented the presence of P. varaldii in these areas and therefore it is highly unlikely that these areas are inhabited by P. varaldii populations. TARkHNISHVILI et al. (2009) found Pelobates syriacus presence significantly associated with integrative normalized difference vegetation index (INDVI). This result indicates the importance of vegetation cover productivity on the spatial distribution of Pelobates species and stresses the need to identify the distribution of suitable NDVI classes within the distribution of P. varaldii. As the NDVI values categorically increase towards that of the highest year-round green biomass and P. varaldii only occurs amongst the higher classes. This can be explained by the preference of the species for forested habitats. The associated NDVI classes partially match the known distribution (Fig. 4), but additionally extend eastwards into the Rif Mountains, parts of the Middle Atlas

Mountains and the region around Melilla. Suitable NDVI classes also occur in the Doukkala region, the area around Essaouira and Agadir, as well as in parts of the High Atlas. However, a major gap in suitable NDVI appears in the area between Casablanca and Oualidia (Fig. 4). Many authors have shown that Pelobates species are dependant of arenaceous substrates (e.g. SCHLEICH et al., 1996; SCALI & GENTILLI, 2003), which are mostly present along the Atlantic coastal plain in the northwest of Morocco. Despite that P. varaldii also occurs on other geology classes within the study region, most of these populations are restricted to the sandy microhabitat patches available. As an example, the area around Oualidia is characterized by saturated karsts (FAkIR et al., 2002) but small suitable sandy areas harbouring P. varaldii populations exist there. Morocco is considered to be the foremost country in the Mediterranean basin for its richness in temporary ponds (GRILLAS et al., 2004). Despite the fact that these temporary ponds are distributed throughout the country, they show a remarkable diversity as a result of a range of climatic, geological and geomorphological conditions (GRILLAS et al., 2004). The coastal Atlantic plains are renowned for their high numbers of temporary ponds, which are characterized by soils that are either hydromorphic over a sandstone or schist substrate (Ben Slimane forest), or sandy over an impermeable clay layer (Mamora forest) (GRILLAS et al., 2004). The larvae of Pelobates species have both the largest size and longest larval period known among anurans (BUCHHOLz & HAyES, 2002) and can reach lengths of 130 mm with larval periods that can last for over four months depending on condi-

DISTRIBUTION AND CONSERVATION OF PELOBATES VARALDII

tions (e.g. SzékELy et al., 2010). The unique combination of geological and geomorphological conditions of the soil and the climate in the coastal Atlantic plain creates temporary ponds that have the ability to keep water levels for a period long enough to allow P. varaldii larvae for completing their larval development. Moreover, P. varaldii is known to inhabit direct surroundings of breeding waters and mainly breeds in temporary ponds (author's personal observation), therefore making the spatial distribution of suitable breeding waters a key indicator for identifying suitable habitat for the species. The temporary ponds of the type present in the coastal Atlantic plains stretch from around Tangier southwards to the area around Essaouira (GRILLAS et al., 2004) and this corresponds to the potential niche models of P. varaldii (BONS & GENIEz, 1996). Conclusion and recommendations Overall, the ENMs seem to have performed well in predicting the distribution of P. varaldii. The Mamora forest forms the core distribution area and harbours the largest and healthiest population of P. varaldii, whereas several smaller and possibly fragmented populations exist in the area around Larache, Tangier, and Oualidia. The distribution between Rabat and Oualidia has still not been fully resolved but potential niche models and fieldwork indicates that this area mostly lacks suitable habitat. The largest threats to P. varaldii consist of a reduction of gene flow and loss of genetic variability as a result of habitat fragmentation, the transformation of habitat and breeding ponds into agricultural and industrial areas, the pollution of breeding ponds due to extensive livestock pasturing and the

possible expansion of exotic crayfish species into the Mamora forest. Moreover, the report of B. dendrobatidis presence in Morocco and the infection of P. varaldii larvae near Larache (EL MOUDEN et al., 2011) could constitute a major future threat. Proposed conservation actions include: (1) the development of a biannual monitoring program to assess the long term population status, (2) the identification of areas to be proposed as protected areas, and (3) the development of a management plan for Mamora forest to reduce negative effects of agriculture, logging and livestock on breeding ponds and habitat. Immediate conservation actions in the form of habitat protection are needed to protect the small population south of Tangier airport. It is highly recommended to monitor the status and expansion of exotic crayfish and to establish a program to extirpate this devastating invasive species. Acknowledgement This research was conducted in collaboration with the University of Applied Science Van Hall-Larenstein and the EDGE program (zoological Society of London). M. Weterings, P-A. Crochet, J. Bosch and J. Cooper are thanked for their help and data sharing. DD and EHEM were partially supported by a 2004 DAPTF seed-grant. PdP was supported by a 2010 conservation grant of the Societas Europaea Herpetologica. Fieldwork in Morocco was conducted under permit decisions 84°HCEFLCD/DLCDPN/ DPRN/CFF (2008-2009) and 10°HCEFLCD/ DLCDPN/DPRN/CFF (2010) issued by Haut Commissariat aux Eaux et Forêts et à la Lutte Contre la Désertification.

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Basic and Applied Herpetology 26 (2012): 73-86

Aspectos ecológicos y efectos del manejo forestal en una población de tortuga mediterránea (Testudo hermanni hermanni) en Cataluña (España) Maria Casamitjana1, Juan Carlos Loaiza2, 3, *, Núria Simon1, Pere Frigola4 Ingeniero forestal, consultor particular, Universidad de Lleida, España. Escuela de Biociencias, Universidad Nacional de Colombia, Medellín, Colombia. 3 Centre Tecnològic Forestal de Catalunya, Solsona, España. 4 Departament d’Agricultura, Ramaderia, Pesca, Alimentació i Medi Natural, Generalitat de Catalunya, España. 1 2

* Correspondencia: Universidad Nacional de Colombia, sede Medellin, Escuela de Biociencias, Bloque 14, oficina 208, Medellín, Colombia. E-mail: jc.loaiza@ctfc.es

Recibido: 20 noviembre 2011; revisión recibida: 16 agosto 2012; aceptado: 22 agosto 2012.

El objeto de esta investigación es la evaluación de los efectos de los trabajos forestales en Testudo hermanni hermanni, así como establecer parámetros ecológicos útiles para el manejo de su hábitat. El estudio se realizó en el noroeste de España, en la sierra de la Albera, donde se encuentra una de las últimas poblaciones naturales. Mediante radioseguimiento semanal se estudiaron 21 tortugas en un periodo de seis meses comprendido entre julio de 2009 y enero de 2010, realizándose trabajos forestales entre octubre de 2009 y enero de 2010. El impacto de los trabajos y daños asociados fue estimado usando dos metodologías: el seguimiento de los desplazamientos de las tortugas monitorizadas y la utilización de modelos de yeso colocados en ubicaciones y condiciones similares a los individuos vivos. Los resultados del radioseguimiento muestran que la presencia de tortugas en zonas abiertas (pastos y alcornocal disperso) representa el 30% de las localizaciones, mientras que el 70% se ubica bajo bosque denso y matorral alto. El análisis del efecto de los trabajos forestales indicó que en el 45% de los casos las tortugas reaccionaron a los trabajos forestales incluso en periodo de hibernación. Los resultados de la simulación del impacto de los trabajos forestales sobre modelos de yeso mostraron que los trabajos realizados manualmente (desbrozadoras) ocasionaron heridas graves en un 4% de las tortugas, mientras que los daños realizados por trabajos mecanizados con tractor de cadenas en zonas planas representaron un 22% de mortalidad y un 6% de heridas graves. Key words: dominio vital; tortuga mediterránea; trabajos forestales; radioseguimiento. Ecological aspects and effects of forestry management on a population of Hermann’s tortoise (Testudo hermanni hermanni) in Catalonia (Spain). The aim of this research was to evaluate the effects of forestry works on Testudo hermanni hermanni, and to establish ecological parameters useful for habitat management. This study took place in the northeast of Spain, in the Albera Mountains, where one of the last natural populations of Hermann’s tortoises lives. Twenty-one tortoises were weekly monitored using radiotracking during six months, between July 2009 and January 2010. Between October 2009 and January 2010, forestry works took place at the study site. The impact of the forestry works and the associated damages were estimated using two approaches: the displacements of the monitored tortoises and the use of plaster models resembling tortoises, which were placed in similar locations and conditions to live tortoises. Radiotracking results showed that 30% of the tortoises were found on open areas (pastures and sparse cork oak forest), whereas 70% corresponded to dense forests and heath scrubland locations. The analysis of the effect of forestry works showed that tortoises were affected by the works in 45% of the cases, even during hibernation period. The simulation results of the impact of the forestry works based on plaster models showed that 4% of the models were seriously damaged when using manual works (brush cutters and clearing saws), whereas mechanized works (chain tractor) were responsible for an estimation of 22% mortality and 6% serious damages. Key words: forestry works; home range; Mediterranean tortoise; radiotracking.

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Testudo hermanni hermanni es una especie ligada a ambientes mediterráneos, con presencia en Europa del este y de forma aislada en los países del oeste (España, Francia, Italia, Islas Baleares, Córcega y Cerdeña (BERTOLERO, 2010)). Debido a la presión humana (cambio de usos del suelo, incendios forestales, capturas ilegales de individuos adultos y manejo forestal) y a la depredación, esta especie ha desaparecido en la mayoría de las regiones de Europa, siendo considerada en peligro de extinción (MADEC, 1996; GUyOT & CLOBERT, 1997; BUDó et al., 2003). Estudios previos en Cataluña describen varios parámetros de esta especie como densidades, sex-ratios, parámetros reproductivos, efectos de la temperatura del suelo y otros aspectos ecológicos (BERTOLERO, 2002; BUDó et al., 2003; Franch & Oller, datos no publicados), pero los estudios de sus dominios vitales son muy escasos (BERTOLERO, 2010), teniendo como única referencia estudios realizados en otras latitudes (CALzOLAI & CHELAzzI, 1991). Otro aspecto que no ha sido analizado es la influencia directa de los trabajos forestales y el efecto en el dominio vital. Debido a las capacidades miméticas de esta especie y a que gran parte del año permanece enterrada hibernando, la detección y seguimiento son complicados, tal y como sucede en la mayoría de los herpetos (MAzEROLLE et al., 2007), con las dificultades que esto conlleva a la gestión de los espacios donde se encuentra presente. En Cataluña el sucesivo abandono de los campos y bosques en los últimos años ha generado una vegetación muy cerrada con continuidad vertical, sin ningún tipo de gestión, incrementando la biomasa forestal seca y el riesgo de incendios, con el consecuente efecto negativo en las

poblaciones de vertebrados en estos ambientes (MOREIRA & RUSSO, 2007). Los bosques mediterráneos precisan estrategias de manejo forestal con el doble propósito de promover la conservación de la biodiversidad de especies, y disminuir la ocurrencia de incendios forestales y la degradación de los hábitats (SCARASCIA et al., 2000). Debido al incremento de los costes de contratación de mano de obra en las últimas décadas, se ha extendido el uso de tractores de cadenas para el manejo forestal, siendo precisa una evaluación de estas prácticas y de su efecto en la tortuga mediterránea. Este estudio se realizó en la Albera, la sierra más oriental de los Pirineos, donde se encuentra una de las últimas poblaciones nativas de T. h. hermanni en Cataluña (LLORENTE et al., 2002; VILARDELL et al., 2008). El estudio tiene dos objetivos, el primero es describir parámetros ecológicos del dominio vital y estrategias de uso del espacio de T. h. hermanni en diferentes periodos de actividad para discutir la idoneidad del periodo de la realización de los trabajos. El segundo objetivo es evaluar el impacto de los trabajos forestales en las poblaciones de T. h. hermanni así como las tasas de mortalidad asociadas a éstos. MATERIALES y MéTODOS Zona de estudio y tratamientos forestales El estudio se llevó a cabo en la Sierra de la Albera, concretamente en el Valle de la Balmeta, zona protegida mediante las figuras PEIN (Espacio de Interés Natural) y Red Natura 2000, concretamente en una finca gestionada por la administración de Medio Ambiente (GE-3002) conocida como Mas

MANEJO FORESTAL CON TORTUGA MEDITERRÁNEA

Guanter, del municipio de Llançà, en el Pirineo Oriental del noroeste de España (Fig. 1). Mas Guanter tiene un área de 578,57 ha, de las cuales 359,07 ha son arboladas y 219,5 ha son matorrales y pastos distribuidos en forma de mosaico. La localidad de estudio se ubica en la parte norte del valle en un área de 111,3 ha. El clima y la vegetación son típicamente mediterráneos, con una precipitación media anual de 563 mm (según fuentes del Paraje Natural de Interés Nacional la Albera), una temperatura media estival de 22,5ºC, y una temperatura media en invierno de 6,4ºC. El relieve del valle tiene pendientes pronunciadas, presentando el 85% de la zona de

estudio pendientes superiores al 30%, con una altitud que varía entre los 104 y los 605 m sobre el nivel del mar. El seguimiento de las tortugas se realizó entre julio de 2009 y enero de 2010. En el período comprendido entre octubre de 2009 y enero de 2010, se estudió el efecto de los trabajos forestales en las poblaciones monitorizadas aprovechando la realización de trabajos forestales en la finca de gestión pública donde se ubica la zona de estudio. Para ello, se delimitaron zonas con distintos tratamientos en función de la gestión realizada (Fig. 1). Los tratamientos estudiados fueron: desbrozado total, microparcelación, mantenimiento de

Figura 1: Localización de la zona de estudio. En el mapa se detallan los tratamientos forestales realizados, la localización de las trampas de muestreo y la situación de las parcelas de muestreo de simulación de daños donde se colocaron modelos de tortugas de yeso para evaluar el potencial impacto de los trabajos forestales sobre T. h. hermanni.

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cortafuegos y mantenimiento de bosques de ribera. En la Tabla 1 se detalla la superficie afectada, junto con la maquinaria precisa, el tipo de restos producidos y la finalidad del trabajo forestal realizado. Dado que las tortugas usan los restos producidos en los trabajos (triturado y troceado), en este estudio se monitorizó el uso que hacían los individuos de los distintos restos. La trituración de restos implica un número superior de impactos de la máquina desbrozadora contra el suelo que en el troceado, y por tanto, un impacto potencialmente mayor sobre las tortugas.

Antes de empezar los trabajos forestales, los obreros recibieron formación sobre cómo actuar para reducir el riesgo de dañar tortugas con la maquinaria forestal manual, mediante la realización de un corte previo de la vegetación a 15 cm del suelo, para examinar la posible presencia de tortugas, y poder proceder a la trituración y troceado de restos con el menor impacto posible. Técnicas de captura y radioseguimiento El proceso de monitorización se realizó mediante radioseguimiento para establecer el efecto a corto plazo de los trabajos forestales así

Tabla 1: Tratamientos efectuados en la zona de estudio. Para cada tratamiento se detalla la superficie afectada, la maquinaria utilizada, el tratamiento de restos realizados y la finalidad. Tipo de tratamiento realizado sobre vegetación arbustiva

Superficie (ha)

Maquinaria

Restos

Finalidad

Desbrozado total

30*

Desbrozadora manual

Troceado de 50 cm con desbrozadora

Mejora del arbolado. Prevención incendios. Tratamiento comúnmente realizado en silvicultura

Microparcelación Desbroce de arbustos solo en los 3 m de radio alrededor de los alcornoques

30*

Desbrozadora manual

Troceado de 50 cm con desbrozadora

Mejora del arbolado y también la creación de zonas abiertas para las tortugas

Cortafuegos Desbroce total de sotobosque

3,9*

Tractor de cadenas

Trituración de restos

Prevención de incendios (cimas de cordilleras)

Bosque de ribera Desbroce total de sotobosque

1,3*

Desbrozadora manual

No actuación

* Poda manual del arbolado

Trituración de restos Prevención de avenidas con desbrozadora, de agua y prevención manteniendo zonas de propagación de incendios por los intactas cada 30 m. fondos del valle -

Comparación de comportamiento de las tortugas

MANEJO FORESTAL CON TORTUGA MEDITERRÁNEA

como establecer parámetros ecológicos útiles para el manejo de su hábitat, hasta ahora inexistentes para esta especie. El radioseguimiento se realizó entre julio de 2009 y enero de 2010, y los trabajos forestales tuvieron lugar entre octubre de 2009 y Enero de 2010. Esto permitió obtener información sobre los desplazamientos de las tortugas marcadas antes y después de la realización de los tratamientos forestales. Durante el mes de julio de 2009 se capturaron 21 ejemplares de T. h. hermanni en la zona de estudio, a mano o usando trampas de gravedad (Fig. 2). La mayoría de las capturas se realizaron durante las primeras horas del día o al anochecer (periodos de máxima actividad en verano), en ambientes húmedos como bajo zarzamoras y vegetación de ribera. Los individuos fueron sexados, medidos, pesados y marcados mediante marcas en las placas del caparazón (CAGLE, 1939). El peso medio de las hembras estudiadas fue de 647 ± 150,13 gr (media ± desviación estándar, SD) con unas dimensiones promedio de caparazón de 14,83 ±

Figura 2: Diseño de las trampas de gravedad utilizadas para capturar tortugas en el fondo de valle durante el periodo estival, constituidas por recipientes plásticos enterrados totalmente en el suelo y cubiertas con vegetación (moras y dientes de león sobre hojarasca y un entramado de tallos largos) para atraer a las tortugas.

1,24 cm de longitud y 11,5 ± 0,79 cm de ancho. En los machos el peso medio fue de 447,4 ± 184,1 gr con una longitud media de 13,56 ± 2,22 cm y un ancho de 9,95 ± 1,33 cm. Además se anotaron anomalías en los individuos, el hábitat y la actividad desarrollada en el momento de la captura (ver más abajo las variables consideradas). A cada tortuga se le colocó un emisor (DL10, Ayama Segutel, Pantertronic, s.l., Mataró, España) asociado a una señal de radio para ser localizada periódicamente, gracias a un receptor VHF de 21 canales, de 216 MHz, de 10 CV con un margen de 100 dB, el cual permite obtener una gran precisión en la búsqueda. Los emisores instalados tenían una durabilidad de 6 a 8 meses, con batería de botón, con un pulso de 20 mS y 40 pulsaciones por minuto. Los transmisores fueron adheridos al caparazón mediante resina Demotec (Demotec 90, Ankapodol, Nideran, Alemania) colocándolos entre la segunda y tercera placa costal en machos y en la parte central del caparazón en hembras, para interferir en la menor medida posible en la actividad de las tortugas. Una vez colocados, las tortugas fueron liberadas en su localización inicial. El radioseguimiento consistió en una localización semanal (aunque si los trabajadores forestales se aproximaban a las tortugas se realizaba un seguimiento diario). En cada localización, se anotó el tipo de vegetación usada como refugio (herbáceas, restos de vegetación muerta, Erica arborea) así como el hábitat (brezal, alcornocal, campo, vegetación ribera) y el grado de apertura de la vegetación, la actividad (hibernación, desplazamiento, alimentación, reproducción) y las coordenadas UTM. Cabe resaltar que tres de las tortugas monitorizadas presentaban miembros amputados, lo que sirvió para evaluar la capacidad de desplazamiento de individuos discapacitados, comunes en la zona estudiada, ante traba-

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jos forestales. Para analizar la actividad de las tortugas se calculó el desplazamiento acumulado y el desplazamiento medio diario por cada estación del año estudiado (otoño y verano). Caracterización del dominio vital Para estimar el dominio vital previo a los trabajos forestales se analizaron los puntos obtenidos durante el radioseguimiento utilizando polígonos kernel fijos. Este método se aproxima más a la realidad que otras metodologías como los centros de actividad y otras alternativas no paramétricas (POWELL, 2000; kERNOHAN et al., 2001). Los datos fueron analizados mediante el software ArcView (ESRI, California, USA) y las extensiones Animal Movement Analysis (Alaska Biological Science Center, Alaska, USA) (HOOGE & EICHENLAUB, 1997) y “Home Range Extension” HRE (Center for Northern Forest Ecosystem Research, Ontario, Canada) (RODGERS & CARR, 2002), muy eficiente en la determinación de contornos de probabilidad de dominios vitales (MITCHELL, 2006). Se calcularon los polígonos kernel de cada uno de los individuos estudiados, así como de todas las observaciones conjuntas, analizando la zona núcleo (kernel 50%) y la zona representativa (kernel 95%) que indican el 50 y el 95% del área de campeo del animal, respectivamente, y que corresponden a la probabilidad de encontrar un individuo en estas superficies. Se estudió el dominio vital en función del sexo y de la estación del año, así como los parámetros de las zonas de hibernación elegidas. De acuerdo con los polígonos obtenidos se describió el dominio vital (vegetación, pendiente, hábitat) de la tortuga para establecer los parámetros necesarios para su conservación (LONGEPIERRE et al., 2001). Para la caracterización de la vegetación del domi-

nio vital y confirmación de las delimitaciones realizadas en el campo, se realizó un mapa de vegetación de la zona de estudio mediante fotointerpretación. Los datos obtenidos se analizaron mediante una t de Student o G-test en función de las variables a analizar (continuas o categóricas, respectivamente (McDonald, 2009)), usando el software SPSS (version 18.0, SPSS Inc, Chicago, USA). Estudio de los daños directos causados por trabajos forestales Con el objetivo de analizar el potencial efecto de la maquinaria utilizada en los trabajos forestales sobre la población de T. h. hermanni estudiada, se diseñó un molde de 15 cm de longitud y 11 cm de ancho, con dimensiones y forma similares a una tortuga, con el que se fabricaron modelos en yeso. éstos se distribuyeron de forma aleatoria en tres sub-parcelas de muestreo de 200 m2 por cada tratamiento forestal (manual con troceado, manual con trituración de restos y mecanizado). En el caso de trabajos manuales se estudió otro parámetro, la vegetación cortada, considerando tres tipos de vegetación con distinta resistencia al corte (zarzamora, brezal (E. arborea) y vegetación herbácea (en su mayoría Brachypodium retusum)). En total se colocaron nueve modelos por tipo de vegetación y parcela de muestreo, sumando un total de 81 modelos por cada tratamiento manual (manual con troceado restos y manual con triturado de restos). En el caso de trabajos mecanizados se estimó el factor relieve, seleccionando tres parcelas en ubicaciones planas y tres parcelas en pendiente. En cada parcela de muestreo se colocaron nueve modelos de yeso, en lugares similares a los observados en tortugas en condiciones de actividad. El número total de

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Tabla 2: Superficie del dominio vital (representativo y núcleo) de acuerdo con el sexo. Para cada zona se muestra la media y desviación estándar (SD). Dominio vital representativo (ha) (95% kernel) Media SD Machos (N = 10) Hembras ( N = 11) Total individuos

5,7 3,24 4,35

modelos de yeso usados para analizar el efecto de la maquinaria fue de 54, sumando un total de 216. Cuando los trabajos finalizaron, se recogieron los modelos y se calculó el porcentaje de daños con distintos grados de afectación (no tocada, daños leves, daños graves, muerte) de acuerdo con la profundidad de las marcas realizadas por la maquinaria. Debido a que hay dos posibles factores estimados como influyentes en los resultados del impacto de la maquinaria en las tortugas, se planteó un diseño estadístico en bloques aleatorios, considerando la maquinaria utilizada como factor principal y la vegetación como factor secundario en el caso de los trabajos manuales, y la pendiente en el caso de los trabajos mecanizados. Se partió de la hipótesis nula de que los distintos tratamientos afectan de igual manera a las tortugas. Los datos se analizaron mediante un análisis de frecuencias (G-test), mediante el software SPSS (version 18.0, SPSS Inc, Chicago, USA). RESULTADOS Dominio vital De acuerdo con los datos registrados (376 localizaciones totales, siendo 316 en fase activa y 60 en fase de hibernación) se obtuvo un dominio vital total de 70,89 ha (95% de las

3,42 2,36 3,07

Área núcleo (ha) (50% kernel) Media SD 1,24 0,66 0,92

0,83 0,51 0,71

localizaciones) y una zona núcleo (50% localizaciones) de 17,4 ha. Comparando el dominio vital de los individuos estudiados según el sexo se puede observar como éste es ligeramente superior en el caso de los machos (Tabla 2), aunque de forma no significativa (t = 1,902; P = 0,073). El área del dominio vital en verano y otoño (Fig. 3) es de 36,83 y 33,00 ha (95% probabilidad) y 9,82 y 7,80 ha (50% probabilidad), respectivamente. Estadísticamente la diferencia estacional de la superficie del dominio vital no es significativa (t = 1,893; P = 0,051). En verano, los individuos presentan mayores desplazamientos, debido a una mayor actividad y dominios vitales más extensos. El solapamiento del dominio vital con los mapas de vegetación mostró que el dominio vital posee el 50% de la superficie en zonas con vegetación abierta (Tabla 3). Asimismo, si comparamos dichos porcentajes con la distribución de hábitats en la zona de estudio, se observa que los brezales cerrados fueron utilizados con menor intensidad que lo esperado. Durante verano y otoño la vegetación utilizada para resguardarse durante la noche o en las horas de máximo calor durante el día fue la siguiente: 47% bajo zarzamora, 28% bajo brezal 25% bajo B. retusum y restos de vegetación muerta (Fig. 4). Se observa una diferencia estacional (verano, otoño, invierno) en el uso de la vegetación del dominio vital. En verano, cuando las

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Figura 3: Dominios vitales de las 21 tortugas monitorizadas mediante radioseguimiento en las estaciones de verano y otoño.

temperaturas son altas (26,7ºC de tempera- 25% bajo R. ulmifolius. El 13,9% de los inditura media) y la pluviometría es baja (70 mm viduos en hibernación fueron encontrados en 3 meses), las tortugas monitorizadas eligie- bajo restos de vegetación desbrozada (Fig. 4). ron las zonas de fondo de valle (zona Tabla 3: Porcentaje de presencia de T. h. hermanni en los sin agua y con poca insolación) con hábitats analizados en la zona de estudio (111,3 ha), así presencia de zarzamora para escon- como en el dominio vital representativo (95% kernel; ver derse en el 53,8% de los casos, mien- texto para más detalles). tras que en el 30,3% de los casos se % zona estudio % Dominio vital escondían bajo brezo y en el 15,9% Hábitat en zonas abiertas con B. retusum o Brezal abierto 29,7 33,7 herbáceas, donde la insolación es Brezal cerrado 27,6 18,2 15,6 19,9 mayor. En otoño se sitúan en las lade- Alcornocal disperso con sotobosque bajo ras, en zonas abiertas. Las zonas de Campo 0,8 1,3 hibernación utilizadas (a partir de Vegetación de ribera y 7,6 7,4 noviembre), son bajo B. retusum y zarzamoras 18,8 19,5 herbáceas en un 47,1% de los casos, Alcornocal denso con sotobosque cerrado en un 19,4% bajo E. arborea, y en un

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Figura 4: Tipos de vegetación refugio. En el periodo activo corresponde a la vegetación usada como refugio durante las noches o en las horas más calurosas. En el periodo inactivo (o de hibernación) corresponde a la vegetación bajo la cual las tortugas se semienterraron.

Finalmente, la diferencia del uso de la vegetación como refugio es significativa entre machos y hembras en otoño (G = 11,670; P = 0,0014) e invierno (G = 44,197; P < 0,001), mientras que en verano no hubo diferencias (G = 0,702; P = 0,35), siendo R. ulmifolius el refugio empleado mayoritariamente. La caracterización de la pendiente del dominio vital muestra que la mayoría de la actividad se desarrolla en pendientes entre el 30 y el 50%, representando un 43,7% de la superficie del dominio vital, siendo las tortugas capaces de subir antiguas terrazas antrópicas de piedra seca. Las zonas de hibernación se caracterizan por el uso mayoritario de terrazas en las zonas donde se realizaron trabajos manuales de desbroce total de sotobosque con troceado de restos. Machos y hembras realizaron desplazamientos similares en las pendientes (t = 0,752; P = 0,461). Por otro lado, aunque con un resultado marginal, las tortugas amputadas no presentaron una inferioridad significativa en cuanto a la capacidad de subir pendientes (t = 2,007; P = 0,057). Desplazamientos Los desplazamientos observados se pueden consultar en la Tabla 4. Se observó un incre-

mento de la actividad y desplazamiento de los individuos con las lluvias estivales (Fig. 5). Asimismo, se constató un aumento de la actividad de las tortugas monitorizadas frente a la proximidad de los trabajos forestales, algunas de las cuales realizaron cortos desplazamientos en dirección opuesta a los trabajos. En relación a las tortugas con miembros amputados, no se observaron diferencias entre los desplazamientos totales realizados por éstas y por las tortugas no lesionadas (t = 1,566; P = 0,13). Impacto de los trabajos forestales Durante la realización de los trabajos forestales, los operarios manuales encontraron 15 tortugas vivas e hirieron levemente una en la tercera placa costal, sin afectar a su supervivencia (seguimiento en libertad durante un año posterior). Además se encontraron ocho caparazones de tortuga con marcas de depredación, hecho no relacionado con este estudio. Los resultados obtenidos de los modelos (Tabla 5) muestran que los trabajos mecanizados pueden llegar a causar un 37% de mortalidad (moldes de yeso destruidos debido al atropello por la maquinaria o cortes en la parte dorsal central donde se ubica la colum-

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Tabla 4: Desplazamientos acumulado y medio diario (media y desviación estándar SD) observados en función del el sexo y la estación del año. El desplazamiento global representa la media del desplazamiento total de machos y hembras. Desplazamiento acumulado (m)

Desplazamiento medio diario (m / día)

Machos

Hembras

Machos

Hembras

Verano

Media SD

642 334,6

477,8 354,7

14,1 7,4

10,1 7,7

Otoño

Media SD

142,9 61,3

132,8 67,2

3,7 1,9

3,6 3

Invierno

Media SD

73 97,4

31,9 24,4

2,3 3,7

1 0,7

Total

Media SD

857,9 324,9

642,4 408,1

8,3 12,4

Desplazamiento global

Media SD

752,6 370,4

na de las tortugas) en terrenos planos, mientras que este porcentaje disminuye en relieves irregulares (19%), ya que la cadena de la maquinaria pasa muy por encima de los individuos. Los trabajos manuales con desbrozadora mostraron que los individuos encuentran mayor protección bajo E. arborea, alcanzando un 5% de daños leves en trituración y un 4% de daños graves en troceado de restos. En relación al uso de herbáceas y zarzamora como protección, la simulación muestra un 10% y un 11% de daños leves bajo troceado, respectivamente, mientras que en los trabajos de trituración se estimó un 7% de daños leves bajo herbáceas. Realizando un análisis de frecuencias (G-test), se encontró que la diferencia de daños entre los distintos tratamientos (manuales o mecanizados) es altamente significativa (G = 62,810; P < 0,001). El efecto del relieve en el tratamiento mecanizado es altamente significativo en cuanto a los daños ocasionados (G = 27,652; P < 0,001). La vegetación existente en el caso de tratamientos manuales también comporta efectos signi-

7 14 7,7 13,2

ficativamente distintos en las tortugas simuladas (G = 7,861; P = 0,009). Finalmente, no se encontraron diferencias significativas en el efecto del tipo de trabajo manual realizado (troceado versus triturado) en B. retusum (G = 2,126; P = 0,345). En cambio, los resultados fueron significativos en E. arborea (G = 7,868; P = 0,0097). DISCUSIóN Los resultados obtenidos en este estudio han permitido vislumbrar algunos aspectos que son importantes para la gestión del territorio en presencia de tortuga mediterránea. Entre los parámetros estudiados se confirmó que el rango altitudinal del dominio vital de las tortugas se mantuvo entre 148 y 299 m sobre el nivel del mar, tal y como indicaban los resultados encontrados por LLORENTE et al. (1995), en los que se observó un rango entre el nivel del mar y los 400 metros de altitud. Por otro lado, no se encontraron diferencias significativas en el tamaño del dominio vital

MANEJO FORESTAL CON TORTUGA MEDITERRÁNEA

Figura 5: Relación entre temperatura, pluviometría y desplazamiento de las tortugas monitorizadas durante el periodo de estudio. Unidades del eje vertical expresadas entre paréntesis en la leyenda.

por sexo ni por estación, aunque sí se encontraron diferencias en el uso de la vegetación refugio por sexos y por estación. Se confirmó que dentro del dominio vital de las tortugas los espacios abiertos tienen una gran importancia, representando un 30% del dominio vital, tal y como observó BERTOLERO (2010), siendo también importante la presencia de zonas de vegetación más cerrada donde poder resguardarse. Por lo tanto, se recomienda la realización de trabajos forestales para mantener zonas abiertas, y al mismo tiempo espaciados en el tiempo, que permitan mantener una vegetación en mosaico. Este hecho per-

mitirá mantener la vegetación refugio necesaria en cada estación del año, especialmente en verano, cuando las tortugas hacen un mayor uso de zonas más cerradas (53,8% de los casos bajo zarzamora y 30,8% bajo brezo). Considerando estos resultados, sería preferible la realización de trabajos forestales en invierno, debido a que las plantas desbrozadas tienen tiempo suficiente para regenerarse hasta el verano entrante. Asimismo, durante el periodo de hibernación los restos de desbroce pueden ser utilizados como refugio. Para la realización de los trabajos forestales en este periodo es importante destacar que la

Tabla 5: Grado de afectación o daños provocados por el impacto directo de los distintos tratamientos forestales realizados sobre 216 tortugas simuladas en yeso. Tratamiento forestal No tocadas

% de tortugas Daños leves Daños graves

Muerte

Trabajos manuales (corte de restos de 50 cm, con desbrozadora)

B. retusum E. arborea R. ulmifolius Media

85 95 90 90

11 4 10 9

4 1 0 1

0 0 0 0

Trabajos manuales (con trituración de restos con desbrozadora)

B. retusum E. arborea Media

89 96 92

7 0 4

4 4 4

0 0 0

Pendiente zonas planas Media

67 56 61

15 7 11

1 11 6

17 26 22

Trabajos mecanizados (tractor de cadenas)

CASAMITJANA ET AL.

hibernación de T. h. hermanni es discontinua y que por lo tanto se puede ver interrumpida por varios factores como el incremento puntual de la temperatura en pleno invierno o la proximidad de trabajos forestales (BERTOLERO, 2010). Por lo tanto, los trabajos realizados en invierno, en zonas con tortugas y en periodo de hibernación tienen que ser realizados cuidadosamente. La realización de trabajos en otras estaciones, como en la primavera, que coincide con el inicio del periodo reproductor, y por lo tanto con un periodo de máxima actividad, puede resultar inadecuada debido al impacto sonoro y a la destrucción del hábitat idóneo para el periodo estival. Por otro lado, tal y como se ha confirmado en nuestro estudio, el mayor riesgo de daños se observa en terrenos planos y en el empleo de maquinaria forestal pesada. Por esa razón, en las zonas por debajo de 300 m de altitud se recomienda la realización de trabajos manuales de forma exclusiva, evitando tractores de cadenas. Además, para minimizar al máximo el impacto de los trabajos forestales se recomienda que éstos sean realizados por trabajadores sensibilizados con la presencia de tortugas. El resultado relativo al impacto de los trabajos manuales (4%) es inferior a la estimación de mortalidad debida a incendios forestales, estudiada mediante técnicas de capturarecaptura por varios autores (70% y 57% de impacto según HAILEy, 2000 y COUTURIER et al., 2011, respectivamente). Los resultados de nuestro estudio indican un impacto variable del tipo de maquinaria y tratamiento de restos (troceado o triturado) en función del tipo de vegetación, de manera que aunque en especies leñosas no hay diferencias, en el sotobosque herbáceo sí que se observan diferencias. Por lo tanto, se recomienda un especial cuida-

do en la realización de tratamiento de restos por trituración en sotobosque herbáceo. En este estudio solo se han considerado los daños que pueden provocar mortalidad directa. Sin embargo, otros factores como la infección de heridas leves, no han sido considerados. Los resultados obtenidos del análisis de individuos con miembros amputados muestran que éstos no diferían significativamente del resto en la capacidad de desplazamiento ni en la capacidad de subir pendientes, aunque los dominios vitales eran más reducidos que en los individuos sanos. A pesar de que la probabilidad de muerte directa a causa de los trabajos forestales en los individuos con miembros amputados se estima, por lo tanto, igual que en los individuos sin lesiones, los primeros parecen más susceptibles a la depredación, debido a los dominios vitales más reducidos. Esta afirmación concuerda con los resultados observados por ESQUE et al. (2010) en tortugas del desierto (Gopherus agassizii), donde las hembras, con dominio vital inferior, sufrieron una mayor mortalidad por depredación. Finalmente, en un futuro será interesante ampliar el estudio a otras estaciones del año donde la actividad de las tortugas es mayor, especialmente durante la primavera y verano, y así cerrar el ciclo anual de la especie. También se recomienda realizar un estudio de depredación sobre T. h. hermanni para evaluar el impacto relativo de los trabajos forestales en relación al impacto por depredación. Agradecimiento Esta investigación fue encargada por la Dirección General de Medio Ambiente y Política Forestal. Agradecemos la colabora-

MANEJO FORESTAL CON TORTUGA MEDITERRÁNEA

ción de TRAGSA S.L. y Miguel Galán, por facilitar la investigación. El autor principal quiere agradecer la colaboración del Centre de Reproducció de Tortugues de l’Albera por su apoyo durante el transcurso del estudio. REFERENCIAS BERTOLERO, A. (2002). Biología de la tortuga mediterránea aplicada a su conservación. Tesis doctoral, Universitat de Barcelona. BERTOLERO, A. (2010). Tortuga mediterránea – Testudo hermanni, In A. Salvador & A. Marco (eds.) Enciclopedia Virtual de los Vertebrados Españoles. Museo Nacional de Ciencias Naturales, Madrid. Disponible en http://www.vertebradosibericos.org/reptiles/tesher.html. Consultado el 15/06/2012. BUDó, J.; CAPALLERAS, X.; MASCORT, R. & FèLIX, J. (2003). Estudi de la depredació de postes de tortuga mediterrània (Testudo hermanni hermanni) a la serra de l'Albera (Pirineu oriental, Catalunya). Butlletí de la Societat Catalana d’Herpetologia 16: 20-23. CAGLE, F.R. (1939). A system of marking turtles for future identification. Copeia 3: 170-172. CALzOLAI, R. & CHELAzzI, G. (1991). Habitat use in a central Italy population of Testudo hermanni Gmelin (Reptilia Testudinidae). Ethology Ecology and Evolution 3: 153-166. COUTURIER, T.; CHEyLAN, M.; GUéRETTE, E. & BESNARD, A. (2011). Impacts of a wildfire on the mortality rate and small-scale movements of a Hermann's tortoise Testudo hermanni hermanni population in southeastern France. Amphibia-Reptilia 32: 541-545. ESQUE, T.C.; NUSSEAR, k.E.; DRAkE, k.k.; WALDE, A.D.; BERRy, k.H.; AVERILLMURRAy, R.C.; WOODMAN, A.P.; BOARMAN,

W.I.; MEDICA, P.A.; MACk, J. & HEATON, J.S. (2010). Effects of subsidized predators, resource variability, and human population density on desert tortoise populations in the Mojave Desert, USA. Endangered Species Research 12: 167-177. GUyOT, G. & CLOBERT, J. (1997). Conservation measures for a population of Hermann’s tortoise Testudo hermanni in southern France bisected by a major highway. Biological Conservation 79: 251-256. HAILEy, A. (2000). The effects of fire and mechanical habitat destruction on survival of the tortoise Testudo hermanni in northern Greece. Biological Conservation 92: 321-333. HOOGE, P.N. & EICHENLAUB, B. (1997). Animal Movement Extension to Arcview. ver. 1.1. Alaska Science Center Biological Science Office, U.S. Geological Survey, Anchorage, Alaska, USA. kERNOHAN, B.J.; GITzEN, R.A. & MILLSPAUGH, J.J. (2001). Analysis of animal space use and movements, In J.J. Millspaugh & J.M. Marzluff (eds.) Radio Tracking Animal Populations. Academic Press, San Diego, pp. 126-166. LLORENTE, G.A.; MONTORI, A.; SANTOS, X. & CARRETERO, M.A. (1995). Atlas dels amfibis i rèptils de Catalunya i Andorra. Editorial El Brau, Figueres, Spain. LLORENTE, G.A.; MONTORI, A.; CARRETERO, M.A. & SANTOS, X. (2002). Testudo hermanni (Gmelin, 1789) Tortuga mediterránea, In J.M. Pleguezuelos, R. Márquez & M. Lizana (eds.) Atlas y Libro Rojo de los Anfibios y Reptiles de España. Dirección General de Conservación de la Naturaleza - Asociación Herpetológica Española, Madrid, pp. 151-153. LONGEPIERRE, S.; HAILEy, A. & GERNOT, C. (2001). Home range area in the tortoise

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Testudo hermanni in relation to habitat complexity: implications for conservation of biodiversity. Biodiversity and Conservation 10: 1131-1140. MADEC, D. (1996). La predation dans le processus de conservation de la Tortue d'Hermann, Testudo hermanni, In B. Devaux (ed.) Proceedings of the International Congress of Chelonian Conservation, SOPTOM editorial, Gonfaron, France, pp. 181-183. MAzEROLLE, M.J.; BAILEy, L.L.; kENDALL, W.L.; ROyLE, J.A.; CONVERSE, S.J. & NICHOLS, J.D. (2007). Making great leaps forward: accounting for detectability in herpetological field studies. Journal of Herpetology 41: 672-689. MCDONALD, J.H. (2009). Handbook of Biological Statistics, 2nd ed. Sparky House Publishing, Baltimore, Maryland, USA. MITCHELL, B.R. (2006). Comparison of programs for fixed kernel home range analysis. Remotely Wild. Disponible en http://www.wildlife.org/ wg/gis/newsletter/jun06/hrcompar.htm. Consultado el 15/06/2012.

MOREIRA, F. & RUSSO, D. (2007). Modelling the impact of agricultural abandonment and wildfires on vertebrate diversity in Mediterranean Europe. Landscape Ecology 22: 1461-1476. POWELL, R.A. (2000). Animal home ranges and territories and home range estimators, In L. Boitani & T.k. Fuller (eds.) Research Techniques in Animal Ecology. Columbia University Press, New york, pp. 65-110. RODGERS, A.R. & CARR, A.P. (2002). HRE: The Home Range Extension for ArcView™ (Beta Test Version 0.9, July 1998).User’s Manual. Center for Northern Forest Ecosystem Research, Thunder Bay, Ontario, Canada. SCARASCIA, G.; OSWALD, H.; PIUSSI, P. & RADOGLOU, k. (2000). Forests of the Mediterranean region: gaps in knowledge and research needs. Forest Ecology and Management 132: 97-109. VILARDELL, A.; CAPALLERAS, X.; BUDó, J.; MOLIST, F. & PONS, P. (2008). Test of the efficacy of two chemical repellents in the control of Hermann’s tortoise nest predation. European Journal of Wildlife Research 54: 745-748.

Basic and Applied Herpetology 26 (2012): 87-97

Additional notes on the diet of Japalura swinhonis (Agamidae) from southwestern Taiwan, with comments about its dietary overlap with the sympatric Anolis sagrei (Polychrotidae) Gerrut Norval1,*, Shao-Chang Huang2, Jean-Jay Mao3, Stephen R. Goldberg4, Kerry Slater1 Applied Behavioural Ecology & Ecosystem Research Unit, Department of Environmental Sciences, University of South Africa, Republic of South Africa. 2 Queensland Brain Institute, The University of Queensland, St. Lucia, Queensland, Australia. 3 Department of Forestry & Natural Resources, National Ilan University, Ilan, Taiwan, R.O.C. 4 Department of Biology, Whittier College, Whittier, California, USA. 1

* Correspondence: Global Village Organization, 577 Chong San Road, Chiayi City, 600, Taiwan, R.O.C. Phone: +886 5 286 9310, Fax: +886 5 227 9967, E-mail: gnorval@gmail.com

Received: 9 September 2012; received in revised form: 28 October 2012; accepted: 31 October 2012.

Japalura swinhonis is an endemic agamid lizard in Taiwan, and although its diet has been examined in northern Taiwan and Orchid Island, it has not been investigated in other parts of its range. Investigating the diet of a species from different parts of its range is crucial due to temporal and spatial variations in it. This study examined the dietary items of 47 J. swinhonis from Santzepu and yunlin, southwestern Taiwan. We also reviewed the diet of J. swinhonis and compared it with that of Anolis sagrei from Santzepu, where these species are sympatric in anthropogenically created habitats such as Areca catechu plantations and fruit orchards. The diet of J. swinhonis from Santzepu was dominated by hymenopterans, followed by coleopterans, lepidopterans and trichopterans, while that of the J. swinhonis from yunlin was dominated by isopterans, followed by hymenopterans, lepidopterans and coleopterans. The diet of A. sagrei from Santzepu was mainly dominated by hymenopterans, lepidopterans, araneids, hemipterans, coleopterans, dipterans, isopterans and orthopterans, in that order of frequency. From the results of this study it is evident that in areas where J. swinhonis and A. sagrei are sympatric there is a substantial dietary niche overlap, and competition for prey is very likely. Key words: arboreal; competition; invasive species; sit-and-wait forager. Notas adicionales sobre la dieta de Japalura swinhonis (Agamidae) en el suroeste de Taiwán, con comentarios acerca de su solapamiento trófico con la especie simpátrica Anolis sagrei (Polychrotidae). Japalura swinhonis es un agámido endémico en Taiwán, y aunque su dieta se ha examinado en el norte de Taiwán y la isla de Orchid, no se ha estudiado en otras partes de su rango de distribución. Investigar la dieta de una especie en diferentes partes de su rango geográfico es crucial debido a la existencia de variaciones temporales y espaciales. Este estudio examina la dieta de 47 J. swinhonis de Santzepu y yunlin, en el suroeste de Taiwan. También revisamos la dieta de J. swinhonis y la comparamos con la de Anolis sagrei en Santzepu, donde estas especies aparecen en simpatría en hábitats creados por el hombre, tales como las plantaciones de Areca catechu y frutales. La dieta de J. swinhonis de Santzepu estuvo dominada por himenópteros, seguidos de coleópteros, lepidópteros y tricópteros, mientras que la de J. swinhonis de yunlin estuvo dominada por isópteros, seguidos de himenópteros, lepidópteros y coleópteros. La dieta de A. sagrei de Santzepu consistió principalmente de himenópteros, lepidópteros, araneidos, hemípteros, coleópteros, dípteros, isópteros y ortópteros, en este orden de frecuencia. A partir de los resultados de este estudio parece evidente que en áreas en las que J. swinhonis y A. sagrei aparecen en simpatría existe un solapamiento sustancial de la dieta, con una probable competencia por las presas. Key words: arbórea; competencia; especies invasoras; predador sit-and-wait.

NORVAL ET AL.

A crucial part of the natural history of an animal is its diet, because not only does it reveal the source of the animal’s energy for growth, maintenance, and/or reproduction (DUNHAM et al., 1989; zUG et al., 2001), but it also provides information on the ecological roles of the animal. Diets are thus often used in intraspecific (e.g. ROCHA & ANJOS, 2007; BULTé et al., 2008) and interspecific (e.g. ORTEGA-RUBIO et al., 1995; VIEIRA & PORT, 2007) niche-overlap studies. Since there may be temporal and spatial variations in the diet of a species (e.g. RODRíGUEz et al., 2008; HAWLENA & PéREz-MELLADO, 2009; GOODyEAR & PIANkA, 2011), information on diet from different localities is useful in elucidating the ecological niche of an animal. The Swinhoe’s tree lizard (Japalura swinhonis Günther, 1864) is an endemic agamid lizard in Taiwan and occurs throughout the island, and on the offshore Orchid Island at elevations below 1500 m (OTA, 1991). It is diurnal and occurs in various habitat types, ranging from forests to areas extensively altered by anthropogenic activities, provided there is sufficient sunlight (kUO et al., 2007). Dietary descriptions have only been made for J. swinhonis from northern Taiwan (kUO et al., 2007) and Orchid Island (HUANG, 2007). There is thus still a need for dietary descriptions from other parts of its range. The brown anole (Anolis sagrei Duméril & Bibron, 1837), also known as Norops sagrei (köHLER, 2000; LEE, 2000), is an exotic invasive lizard species in Taiwan (NORVAL et al., 2002, 2009; CHANG, 2007). It is a diurnal trunk-ground species that favours a variety of sunny habitat types and areas disturbed by anthropogenic activities (SCHWARTz & HENDERSON, 1991). Anolis sagrei is an aggres-

sive competitor that has been shown to displace other species of lizards from their habitats (SALzBURG, 1984; TOkARz & BECk, 1987; LOSOS et al., 1993; LOSOS & SPILLER, 1999). Ongoing research on this species in Taiwan indicates that in areas disturbed by anthropogenic activities, it is increasingly becoming part of local ecosystems, both as predator (HUANG et al., 2008a; NORVAL et al., 2010) and prey (NORVAL et al., 2004, 2007, 2011; NORVAL & MAO, 2008; CHIU et al., 2011). This study aims to identify the dietary items of J. swinhonis, using individuals that were collected on an ad hoc basis from southwestern Taiwan, to contribute to the overall understanding of the diet of this species. Furthermore, the current knowledge on the diet of J. swinhonis is reviewed and compared with that of A. sagrei from Taiwan. MATERIALS AND METHODS On the 28th of June, 2002, a J. swinhonis male was found dead on a road (DOR) in Santzepu, Sheishan District, Chiayi County, southwestern Taiwan (23º 25' 46" N, 120º 28' 55" E; datum WGS84). During July and September 2004, and on the 5th of October, 2007, live J. swinhonis were also collected from a small betel nut palm (Areca catechu L.) plantation in Santzepu (23º 25' 42" N, 120º 29' 06" E; datum WGS84), where J. swinhonis is sympatric with A. sagrei. Between the 24th of May and the 1st of June, 2008, additional specimens were collected from two localities (23º 36' 34" N, 120º 34' 13" E and 23º 35' 44" N, 120º 35' 47" E) in yunlin County, southwestern Taiwan. The snout-vent length (SVL) and tail length (TL) were measured to the nearest mm with a transparent plastic ruler, and the ani-

ADDITIONAL NOTES ON THE DIET OF JAPALURA SWINHONIS

Photos Gerrut Norval

mals were weighed (body mass) to the nearest 0.1 g with a digital scale. The DOR lizard was dissected by making a mid-ventral incision and the stomach was removed and slit longitudinally, after which the stomach contents were removed. In order to determine the diet of the living J. swinhonis in a non-lethal manner in the field, the mouth of each collected individual was held open by placing a hard plastic tube (diameter = 10 mm; length = 30 mm) in the oral cavity of the lizard when it opened its mouth as a threat display in response to handling. The stomach contents were then obtained by inserting a no. 4 catheter, lubricated with sunflower cooking oil, into the throat

Figure 1: Stomach flushing methodology. (top) Equipment used for performing stomach flushing. (bottom) A Japalura swinhonis male undergoing stomach flushing.

and stomach of the lizard, and then flushing the stomach contents out by holding the lizard facing downwards over a plastic container while injecting 15 ml clean drinking water through the catheter into the stomach (Fig. 1). A similar method was employed by kUO et al. (2007), and in a preliminary study we found that typically more than 80% (often nearly 100%) of the stomach contents were flushed from the stomachs of the lizards (S.C. Huang & G. Norval, unpublished data), so we considered the method reliable enough for our investigation. After the lizards were stomach-flushed they were released back into the wild at the same locality where they had been collected. In the laboratory, the stomach contents were spread in a petri dish and examined under a dissection microscope, and all the prey items were identified to the order level (samples from the year 2004) and, if possible, to the family level (samples from years other than 2004). The dietary descriptions for J. swinhonis from Orchid Island (HUANG, 2007), and northern (kUO et al., 2007) and southwestern (this study) Taiwan were incorporated for comparison with the dietary descriptions for A. sagrei from Santzepu, southwestern Taiwan (HUANG et al., 2008a; NORVAL et al., 2010). Because not all the dietary items were identified to the family level, the order level was used in the analysis as the resource category. We used the software Ecological Methodology Version 6.1.1. (Exeter software, Setauket, New york, USA) to analyze and compare the diets of J. swinhonis and A. sagrei as follows. The diet diversity index of each species was measured with the Shannon-Wiener function, using the original (HÂ´) and the natural logarithm-transformed (N1) indexes; however, since HUANG (2007) and kUO et al. (2007) did

NORVAL ET AL.

not indicate the numbers of prey items, the results from their studies were not used for the diet diversity index calculations. Niche breadths of J. swinhonis and A. sagrei were estimated using the Levin’s measure of niche breadth (B) and the Levin’s standardized niche breadth (BA) (kREBS, 1999). Only studies indicating the numbers of lizards that preyed on each prey type (i.e. frequency of occurrence: F) were used for these analyses, being therefore excluded the results of HUANG et al. (2008a) and part of the results described herein. The dietary niche overlap was measured with the Pianka’s measure, and the percentage overlap (Schoener Overlap Index) was used to examine the extent of the dietary niche overlap (kREBS, 1999). Again, due to a lack of the required information, the results of the studies by HUANG (2007) and kUO et al. (2007) were excluded from the analysis of dietary niche overlap. RESULTS Thirty-nine J. swinhonis (20 males and 19 females) were collected from Santzepu (Table 1), with SVL ranging from 51 to 77 mm

(mean ± SD = 65.6 ± 7.0), TL from 123 to 196 mm (mean ± SD = 160.5 ± 21.0), and body mass from 3.9 to 13 g (mean ± SD = 7.7 ± 2.6). Eight J. swinhonis (seven males and one female) were collected from yunlin (Table 1), with SVL ranging from 68 to 78 mm (mean ± SD = 73.0 ± 3.2), TL from 168 to 202 mm (mean ± SD = 183.4 ± 9.8), and body mass from 8.7 to 14.9 g (mean ± SD = 11.0 ± 1.9). For description and analysis, we pooled all the dietary data of both sexes from each locality. We recorded 1277 prey items, belonging to 13 orders from three classes, from the Santzepu individuals, and 47 prey items, belonging to six orders from two classes, from the yunlin ones (Table 2). In the stomachs of the J. swinhonis from Santzepu, the most numerous prey items were hymenopterans, more than 90% of which were ants (Formicidae), followed by coleopterans, lepidopterans and trichopterans, respectively (Table 2). For the J. swinhonis from yunlin, the most numerous prey items were termites (Isoptera), followed by hymenopterans (all Formicidae), lepidopterans (all caterpillars) and coleopterans, respectively (Table 2). Cumulatively, the most numerous prey items in the diet of J. swinhonis from southwestern

Table 1: Range and mean (± SD) snout-vent length (SVL), tail length (TL) and body mass of the Japalura swinhonis from Santzepu and yunlin used in this study. Locality

Collection period

Santzepu

2002 2004

Sex ♂ ♂ ♀

2007 yunlin

2008

♂ ♂ ♀

SVL (mm)

TL (mm)

Mass (g)

1 18

72 58-77 (70.4 ± 4.6) 51-73 (61.4 ± 5.6) 54

178 148-196 (177.8 ± 15.0) 129-167 (146.0 ± 6.0) 123

8.7 5.9-13 (9.7 ± 2.2) 3.9-9.3 (6.0 ± 1.2) 4.2

68-78 (73.4 ± 3.2) 70

168-202 (184.3 ± 10.2) 177

8.7-14.9 (11.1 ± 2.0) 10.1

19 1 7 1

ADDITIONAL NOTES ON THE DIET OF JAPALURA SWINHONIS

Table 2: Percentage of each type of dietary item (% items) in Japalura swinhonis from two sites in southwestern Taiwan, and cumulated values considering both sites altogether. For the yunlin population, the occurrence frequency (F) of each prey is also shown.

Class

Insecta

Arachnida Crustacea Magnoliopsida Unknown

Order

Blattaria Coleoptera Dermaptera Diptera Hemiptera Hymenoptera Isoptera Lepidoptera Orthoptera Phasmatodea Trichoptera Aranea Isopoda Asterales

Santzepu % items

yunlin % items F

Cumulated % items

(N = 1277)

(N = 47)

(N = 8)

(N = 1324)

0.31 1.88 0.08 0.63 0.71 93.50 0.08 1.18 0.08 0.08 1.10 0.31 0.00 0.08 0.00

0.00 4.26 0.00 0.00 0.00 14.89 59.57 12.77 2.13 0.00 0.00 0.00 2.13 0.00 4.26

0 2 0 0 0 3 4 5 1 0 0 0 1 0 2

0.30 3.25 0.08 0.60 0.68 90.71 2.19 1.59 0.15 0.08 1.06 0.30 0.08 0.08 0.15

Taiwan were hymenopterans (primarily Formicidae), followed by coleopterans, isopterans, lepidopterans and trichopterans, in that order of frequency (Table 2), while the prey items from the remaining orders made up ca. 3%. The diet of A. sagrei from Taiwan (based on the studies done in Santzepu), consisted mostly of hymenopterans (primarily Formicidae), followed by lepidopterans, araneids, hemipterans, coleopterans, dipterans, isopterans and orthopterans, respectively (Table 3). Prey items from the remaining orders made up ca. 5% of the diet of A. sagrei from Taiwan. The Shannon-Wiener function resulted in H´ = 0.7135 (N1 = 0.6099) for J. swinhonis and H´ = 2.6238 (N1 = 0.1622) for A. sagrei. Niche breadth measures were B = 0.6644 and BA = 0.0129 for J. swinhonis, and B = 1.6149 and BA = 0.0237 for A. sagrei. The Pianka’s

measure and the percentage dietary niche overlap between J. swinhonis and A. sagrei were 0.945 and 65.7%, respectively. DISCUSSION Based on what is understood about the diet of J. swinhonis (HUANG, 2007; kUO et al., 2007), this species can be regarded as a dietary generalist, preying upon a variety of prey types although, according to the calculated Shannon-Wiener function and Levin’s measures, depending to a large extent on large numbers of a relatively few prey types. The occurrence in the diet of J. swinhonis of prey primarily active, mobile, moving on the surface, and visually conspicuous (e.g. ants, termites, and lepidopterans and their larvae) suggests that this species is an opportunistic sit-and-wait (ambush) foraging species (H UEy & PIANkA , 1981;

NORVAL ET AL.

Table 3: Percentage of each type of dietary item (% items) in Anolis sagrei from Santzepu, southwestern Taiwan, according to the studies of HUANG et al. (2008a) and NORVAL et al. (2010), and cumulated values considering both studies altogether. Using the data from NORVAL et al. (2010), the occurrence frequency (F) of each prey is calculated.

Class

Order

Insecta

Blattaria Coleoptera Collembola Dermaptera Diptera Hemiptera Homoptera Hymenoptera Isoptera Lepidoptera Mantodea Neuroptera Odonata Orthoptera Plecoptera Psocoptera Thysanoptera Trichoptera Scolopendromorpha Spirobolida Acarina Aranea Isopoda Stylommatophora Squamata

Chilopoda Diplopoda Arachnida Crustacea Gastropoda Reptilia Unknown

HUANG et al. (2008a) % items

NORVAL et al. (2010) % items F

Cumulated % items

(N = 2355)

(N = 2984) (N = 502)

(N = 5339)

0.72 5.99 0.17 0.04 3.65 4.76 2.72 49.51 1.66 20.72 0.04 0.76 0.00 0.47 0.00 0.89 0.09 0.43 0.00 0.00 0.00 7.39 0.00 0.00 0.00 0.00

PIANkA & VITT, 2003). Prey such as millipedes (Diplopoda) are known to produce toxins, and it is suggested that actively foraging lizards can presumably detect the toxic compounds released by these prey and therefore avoid them, whereas sit-and-wait foragers generally do not (VITT & COOPER, 1986). Seeing as J. swinhonis does feed on millipedes, this further supports the suggestion that this species utilizes a sit-and-wait foraging strategy.

0.50 6.94 0.34 0.24 7.86 4.26 2.01 49.93 3.02 6.10 0.00 0.24 0.27 2.68 0.24 0.03 0.10 0.07 1.58 0.17 0.03 9.99 0.91 1.07 0.17 1.27

13 113 7 6 101 100 50 260 25 140 0 7 6 67 7 1 3 1 40 5 1 166 16 30 5 25

0.60 6.52 0.26 0.15 6.01 4.48 2.32 49.75 2.42 12.55 0.02 0.47 0.15 1.70 0.13 0.41 0.09 0.23 0.88 0.09 0.02 8.84 0.51 0.60 0.09 0.71

Most of the prey items of the J. swinhonis from southwestern Taiwan belonged to orders described as prey of J. swinhonis in the studies done in other parts of Taiwan. Dietary studies from northern Taiwan (kUO et al., 2007) and Orchid Island (HUANG, 2007) found that hymenopterans (primarily Formicidae) were the most frequently preyed upon items, and although these prey were not as large as some of the other common prey orders, either volumetrically or by mass, they still tended to have

ADDITIONAL NOTES ON THE DIET OF JAPALURA SWINHONIS

a high index of relative importance (IRI; Table 4). As in those studies, hymenopterans comprised a large part of the diet of J. swinhonis from southwestern Taiwan. HUANG (2007) did not record isopterans in the diet of J. swinhonis from Orchid Island, while kUO et al. (2007) infrequently recorded isopterans in the diet of J. swinhonis from northern Taiwan. Although isopterans were the most numerous dietary items in the diet of J. swinhonis samples from yunlin, it is highly unlikely that this is due to a regional dietary variation. Rather, the isopterans recorded were winged sexuals, which would illustrate the opportunistic feeding behaviour of J. swinhonis, and how these lizards make use of a temporarily available dietary resource. None of the J. swinhonis dietary studies exa-

mined the seasonal utilization of prey or examined the diet of these lizards systematically for a whole year, and thus there is not enough information to comment on the seasonality of prey. However, seasonal differences could exist and deserve further empirical studies. It should also be noted that even though the diet of J. swinhonis consist primarily of numerous, relatively common, small prey, such as ants and lepidopteran larvae, these lizards may also eat relatively large prey (Fig. 2). Since J. swinhonis usually perches a short distance above the ground on tree trunks and other objects, but forages readily on the ground (JUN-yI & kAU-HUNG, 1982; kUO et al., 2007), it can be described as a trunkground species (WILLIAMS, 1972). This means that the exotic invasive lizard, A. sagrei, occu-

Table 4: Occurrence frequency (F), mean percent weight (W) or mean percent volume (V), and index of relative importance (IRI) of each type of dietary item in male and female Japalura swinhonis from northern Taiwan (data retrieved from kUO et al., 2007) and in adult J. swinhonis from Orchid Island (data retrieved from HUANG, 2007). The IRI for the Orchid Island population is calculated following BJORNDAL et al. (1997) with data from the original publication. The IRI for the northern Taiwan population is retrieved from the original publication.

Class

Order

Northern Taiwan Males (N = 29) Females (N = 27) F

Insecta

Coleoptera Hemiptera Homoptera Hymenoptera Isoptera Lepidoptera Orthoptera Psocoptera Insect larvae Chilopoda Scolopendromorpha Diplopoda Spirobolida Arachnida Aranea Crustacea Isopoda Gastropoda Stylommatophora

6 5 5 24 2 15 4 3 3 2 10 2 3

IRI

5.65 2.52 6.92 2.57 11.06 4.11 28.45 50.7 1.80 0.27 32.17 35.83 3.04 0.9 0 0

10 7 4 25 0 15 6 2

1.53 0 3.18 0.77 1.30

0.34 0 2.36 0.11 0.29

3 2 15 5 3

Orchid Island (N = 20)

IRI

7.15 5.23 5.75 2.94 5.62 1.64 22.45 41.04 0 0 38.55 42.29 0.54 0.24 0 0

11.79

2.62

1.51 4.02 3.13 1.35 8.06

0.33 0.59 3.43 0.49 1.77

2 16

11.85 5.26 7.31 25.97

55.05 61.11

3 1

1.97 7.25

1.31 1.61

4.79

2.13

NORVAL ET AL.

Photos Gerrut Norval

Figure 2: Japalura swinhonis feeding on large prey. (top) A male in Santzepu, Sheishan District, Chiayi County, southwestern Taiwan, preying on a relatively large moth (Sphingidae). (bottom) A female from the same area consuming a stag beetle (Lucanidae) grub.

pies the same niche as J. swinhonis, and both species have the same foraging mode. The Levin’s measures also indicate that the niche breadth of A. sagrei is broader than that of J. swinhonis. From the results of this study it is also clear that in areas where these two lizard species are sympatric there is a substantial dietary niche overlap, and therefore competition for prey is very likely. However, J. swinhonis males and females have SVLs that range from 22 to 82 mm (N = 104; mean ± SD = 70.5 ± 8.4) and 22 to 74 mm (N = 59; mean ± SD = 58.2 ± 13.9) respectively (G. Norval, personal observation; from a sample of animals larger than what was used for the present study), and are thus substantially larger than A. sagrei from populations in Taiwan, where males and

females have SVLs that range from 16 to 64 mm (N = 522; mean ± SD = 46.2 ± 9.1) and 17 to 48 mm (N = 538; mean ± SD = 38.2 ± 5.5), respectively (G. Norval, personal observation). This means that adults of J. swinhonis should be able to feed on larger prey than A. sagrei. It is also worth pointing out that in interactions in the wild, A. sagrei usually gives way to J. swinhonis (G. Norval, personal observation). So, even though there is a substantial dietary overlap between these species, it is unlikely to have a negative effect on J. swinhonis. Lister (1976) found that in the presence of competition from sympatric anole species, A. sagrei tends to occupy lower perches. However, in the presence of terrestrial predators, A. sagrei tends to be more arboreal (SCHOENER et al., 2002; LOSOS, 2009). Both the competition for prey and the microhabitat shift to higher perches result in reduced foraging opportunities and a subsequent reduction in body sizes of these lizards (LISTER, 1976; SCHOENER & SCHOENER, 1978; SCHOENER et al., 2002; LOSOS, 2009). So, competition between J. swinhonis and A. sagrei, especially in habitats where terrestrial predators such as Eutropis longicaudata occur, could reduce the ability of A. sagrei to compete with other sympatric saurian species in Taiwan. It must be noted that, through predation, A. sagrei impacts arthropod populations (SPILLER & SCHOENER, 1994; SCHOENER & SPILLER, 1996; SCHOENER et al., 2002; HUANG et al., 2008a,b). Because sympatric J. swinhonis and A. sagrei prey on many of the same types of prey, it can be expected that they can thus exert substantial pressure on arthropods. How such combined pressures would affect arthropod communities warrants empirical studies.

ADDITIONAL NOTES ON THE DIET OF JAPALURA SWINHONIS

There does not seem to be an apparent decline in J. swinhonis or A. sagrei in areas where these lizards are sympatric (G. Norval, personal observation). In highly disturbed and open habitats, such as in A. catechu plantations, A. sagrei tends to be more abundant than J. swinhonis (HUANG et al., 2008a). However, in less disturbed areas, such as parks with large trees, even though A. sagrei is present, J. swinhonis tends to be more abundant (G. Norval, personal observation). Thus, the differences in the densities of these lizards most likely depend on the habitat structure. Japalura swinhonis is more shady-habitat tolerant than A. sagrei (HUANG et al., 2008a), which does not occur in closed habitats (LOSOS et al., 1993), so the re-establishment of large areas of broadleaf forests in disturbed lowland areas of Taiwan will contribute to the conservation of J. swinhonis and other native forest species. Such areas will also function as reservoirs of species like J. swinhonis that can compete with A. sagrei, as well as being barriers for its spread. Acknowledgement The authors thank Jin-Hsiang Wu for his assistance with the collection of lizards from yunlin County. The research presented here adhered to the Guidelines for the Use of Live Amphibians and Reptiles in Field Research (SSAR, ASIH, and HL), and the legal requirements of Taiwan, R.O.C. Since A. sagrei is an exotic invasive species in Taiwan, which the authorities wishes to exterminate, and J. swinhonis is not listed as a protected species, and because the collection was not done within a National Park or other conservation area, no collection permit or other documentation was required.

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ting on arthropod diversity in southern Taiwan. Zoological Science 25: 1121-1129. HUANG, S.-C.; NORVAL, G. & TSO, I.-M. (2008b). Predation by an exotic lizard, Anolis sagrei, alters the ant community structure in betelnut palm plantations in southern Taiwan. Ecological Entomology 33: 569-576. HUANG, W.-S. (2007). Ecology and reproductive patterns of the agamid lizard Japalura swinhonis on an East Asian island, with comments on the small clutch sizes of island lizards. Zoological Science 24: 181-188. HUEy, R.B. & PIANkA, E.R. (1981). Ecological consequences of foraging mode. Ecology 62: 991-999. JUN-yI, L. & kAU-HUNG, L. (1982). Population ecology of the lizard Japalura swinhonis formosensis (Sauria: Agamidae) in Taiwan. Copeia 1982: 425-434. köHLER, G. (2000). Reptilien und Amphibien Mittelamerikas. Band 1: Krokodile Schildkröten Echsen. Herpeton, Offenbach, Germany. kREBS, C.J. (1999). Ecological Methodology, 2nd ed. Addison Wesley Longman, Menlo Park, California, USA. kUO, C.-y.; LIN, y.-S. & LIN, y.k. (2007). Resource use and morphology of two sympatric Japalura lizards (Iguania: Agamidae). Journal of Herpetology 41: 713-723. LEE, J.C. (2000). A Field Guide to the Amphibians and Reptiles of the Maya World. The Lowlands of Mexico, Northern Guatemala, and Belize. Cornell University Press, Ithaca, New york, USA. LISTER, B.C. (1976). The nature of niche expansion in West Indian Anolis lizards I: ecological consequences of reduced competition. Evolution 30: 659-676. LOSOS, J.B. (2009). Lizards in an Evolutionary

Tree. Ecology and Adaptive Radiation of Anoles. University of California Press, Berkeley, California, USA. LOSOS, J.B. & SPILLER, D.A. (1999). Differential colonization success and asymmetrical interactions between two lizard species. Ecology 80: 252-258. LOSOS, J.B.; MARkS, J.C. & SCHOENER, T.W. (1993). Habitat use and ecological interaction of an introduced and a native species of Anolis lizard on Grand Cayman, with a review of the outcomes of anole introductions. Oecologia 95: 525-532. NORVAL, G. & MAO, J.-J. (2008). An instance of arboricolous predation by a mountain wolf snake (Lycodon ruhstrati ruhstrati [Fischer, 1886]) on a brown anole (Norops sagrei Duméril & Bibron, 1837). Sauria 30: 59-62. NORVAL, G.; MAO, J.-J.; CHU, H.-P. & CHEN, L.-C. (2002). A new record of an introduced species, the brown anole (Anolis sagrei) (Duméril & Bibron, 1837), in Taiwan. Zoological Studies 41: 332-336. NORVAL, G.; MAO, J.-J. & CHU, H.-P. (2004). Mabuya longicaudata (long-tailed skink). Predation. Herpetological Review 35: 393-394. NORVAL, G.; HUANG, S.-C. & MAO, J.-J. (2007). Mountain wolf snake (Lycodon r. ruhstrati) predation on an exotic lizard, Anolis sagrei, in Chiayi County, Taiwan. Herpetological Bulletin 101: 13-17. NORVAL, G.; MAO, J.-J.; BURSEy, C.R. & GOLDBERG, S.R. (2009). A deformed hind limb of an invasive free-living brown anole (Anolis sagrei Duméril & Bibron, 1837) from Hualien City, Taiwan. Herpetology Notes 2: 219-221. NORVAL, G.; HSIAO, W.-F.; HUANG, S.-C. & CHEN, C.-k. (2010). The diet of an intro-

ADDITIONAL NOTES ON THE DIET OF JAPALURA SWINHONIS

duced lizard species, the brown anole (Anolis sagrei), in Chiayi County, Taiwan. Russian Journal of Herpetology 17: 131-138. NORVAL, G.; CHIU, P.-k.; CHU, H.-P. & MAO, J.-J. (2011). An instance of predation on a brown anole (Anolis sagrei Duméril & Bibron, 1837) by a Malay night heron (Gorsachius melanolophus Swinhoe, 1865). Herpetology Notes 4: 5-7. ORTEGA-RUBIO, A.; GONzÁLEz-ROMERO, A. & BARBAULT, R. (1995). Food analysis and resource partitioning in a lizard guild of the Sonoran Desert, Mexico. Journal of Arid Environments 29: 367-382. OTA, H. (1991). Taxonomic redefinition of Japalura swinhonis Günther (Agamidae: Squamata), with a description of a new subspecies of J. polygonata from Taiwan. Herpetologica 47: 280-294. PIANkA, E.R. & VITT, L.J. (2003). Lizards: Windows to the Evolution of Diversity. University of California Press, Berkeley, California, USA. ROCHA, C.F.D. & ANJOS, L.A. (2007). Feeding ecology of a nocturnal invasive alien lizard species, Hemidactylus mabouia Moreau de Jonnès, 1818 (Gekkonidae), living in an outcrop rocky area in southeastern Brazil. Brazilian Journal of Biology 67: 485-491. RODRíGUEz, A.; NOGALES, M.; RUMEU, B. & RODRíGUEz, B. (2008). Temporal and spatial variation in the diet of the endemic lizard Gallotia galloti in an insular Mediterranean scrubland. Journal of Herpetology 42: 213-222. SALzBURG, M.A. (1984). Anolis sagrei and Anolis cristatellus in southern Florida: a case study in interspecific competition. Ecology 65: 14-19. SCHOENER, T.W. & SCHOENER, A. (1978). Estimating and interpreting body-size

growth in some Anolis lizards. Copeia 1978: 390-405. SCHOENER, T.W. & SPILLER, D.A. (1996). Devastation of prey diversity by experimentally introduced predators in the field. Nature 381: 691-694. SCHOENER, T.W.; SPILLER, D.A. & LOSOS, J.B. (2002). Predation on a common Anolis lizard: can the food-web effects of a devastating predator be reversed? Ecological Monographs 72: 383-407. SCHWARTz, A. & HENDERSON, R.W. (1991). Amphibians and Reptiles of the West Indies: Descriptions, Distributions, and Natural History. University Press of Florida, Gainesville, Florida, USA. SPILLER, D.A. & SCHOENER, T.W. (1994). Effects of top and intermediate predators in a terrestrial food web. Ecology 75: 182-196. TOkARz, R.R. & BECk, JR., J.W. (1987). Behaviour of the suspected lizard competitors Anolis sagrei and Anolis carolinensis: an experimental test for behavioural interference. Animal Behaviour 35: 722-734. VIEIRA, E.M. & PORT, D. (2007). Niche overlap and resource partitioning between two sympatric fox species in southern Brazil. Journal of Zoology 272: 57-63. VITT, L.J. & COOPER, JR., W.E. (1986). Foraging and diet of a diurnal predator (Eumeces laticeps) feeding on hidden prey. Journal of Herpetology 20: 408-415. WILLIAMS, E.E. (1972). The origin of faunas. Evolution of lizard congeners in a complex island fauna: a trial analysis. Evolutionary Biology 6: 47-89. zUG, G.R.; VITT, L.J. & CALDWELL, J.P. (2001). Herpetology. An Introductory Biology of Amphibians and Reptiles, 2nd ed. Academic Press, San Diego, California, USA.

Basic and Applied Herpetology 26 (2012): 99-105

Measuring body temperatures in small lacertids: Infrared vs. contact thermometers Miguel A. Carretero CIBIO, Centro de Investigação em Biodiversidade e Recursos Genéticos, Universidade do Porto, Vairão, Portugal. Correspondence: CIBIO, Centro de Investigação em Biodiversidade e Recursos Genéticos, Universidade do Porto, Campus Agrário de Vairão, 4485-661 Vairão, Portugal. E-mail: carretero@cibio.up.pt

Received: 28 March 2012; received in revised form: 2 August 2012; accepted: 3 August 2012.

Infrared thermometers (IRT) are gaining popularity in herpetological thermal ecology due to their several advantages compared to contact thermometers (CT). To evaluate their accuracy in small lacertids, lab parallel measurements using IRT and CT are compared for a set of 52 adult lizards belonging to four different Podarcis forms, including males, pregnant and non-pregnant females, exposed to a photothermal gradient. Skin temperature was measured with an IRT and cloacal temperature with a CT at 10 time intervals, completing 520 paired measurements. Models of the relations were constructed using standardised major axis (SMA) regression. As expected, IRT and CT measurements were significantly correlated but determination coefficients were only moderate, IRT values being systematically higher. Moreover, the SMA regression lines deviated from slope 1 and intercept 0 in all cases, revealing a nonisometric bias; IRT tended to give progressively higher readings than CT for higher temperatures. Results provide methodological insights for further studies on thermal ecology of lacertids. Key words: contact thermometers; infrared thermometers; Lacertidae; measurement bias; thermal ecology. Medida de la temperatura corporal en pequeños lacértidos: Termómetros de infrarrojos vs. termómetros de contacto. El uso de termómetros de infrarrojos (IRT) se está popularizando en ecología térmica de anfibios y reptiles debido a una serie de ventajas respecto de los termómetros de contacto (CT). Con objeto de evaluar su exactitud en pequeños lacértidos, se comparan medidas paralelas tomadas en laboratorio con IRT y CT para un total de 52 lagartijas adultas pertenecientes a cuatro formas de Podarcis, incluyendo machos, hembras grávidas y hembras no grávidas, expuestas a un gradiente térmico. En 10 intervalos temporales, se midió la temperatura de la piel con un IRT y la cloacal con un CT, completándose un total de 520 registros dobles. Se construyeron modelos de regresión entre ambas variables mediante la regresión estandarizada del eje mayor (SMA). Aunque, como era de esperar, las medidas de IRT y CT se correlacionaron significativamente, los coeficientes de determinación fueron sólo moderados, siendo los valores de IRT sistemáticamente más elevados. Además, las rectas de regresión SMA se separaron de la pendiente 1 y del intercepto 0 en todos los casos, indicando un sesgo no isométrico, de modo que IRT tendió arrojar valores cada vez más altos que CT para temperaturas más elevadas. De estos resultados se derivan consecuencias metodológicas que deben tenerse en cuenta en futuros estudios de ecología térmica en lacértidos. Key words: ecología térmica; Lacertidae; termómetros de contacto; termómetros de infrarrojos; sesgo de medida.

Being most functions in ectotherms temperature-dependent, accurate determination of body temperature is crucial to interpret their biological processes. Thus, numerous studies record body temperatures to analyse

the proximate mechanisms linking temperature to physiology, life history, and behaviour of organisms (ANGILLETTA, 2010). Many of them use lizards as models and several have been conducted with lacertids (CASTILLA et

100

CARRETERO

al., 1999). However, obtaining reliable lizard temperatures may be problematic. The standard method consists of inserting a contact thermometer (CT), usually with a thermocouple probe but formerly with a fluid column, in the cloaca. However, this procedure forces the researcher to capture and handle the lizard, which may increase its metabolic rate, degree of stress (MOORE et al., 1991; LANGkINDE & SHINE, 2006) and induce subsequent behavioural changes due to the disturbance, all potentially diminishing the representativeness of the temperatures recorded. Of the alternative methods proposed, implanted transmitters equally require lizard capture and surgery, and are usually more expensive and less suitable for small species (HARE et al., 2007). Because of that, infrared technology is gaining popularity, since reduced handling potentially provides more biological relevance to the measurements, minimises stress, prevents disease transmission and decreases reading time (HARE et al., 2007). More practical than the bulky and expensive infrared cameras, the portable infrared thermometers (IRT), essentially a pistol-shaped handle associated to a sensor and a laser pointer, are now widely used in lizard thermal ecology. Due to their low invasiveness, IRT offer advantages when working with small lizards. In fact, some of the early studies were carried out on lacertids (JONES & AVERy, 1989; TOSINI et al., 1995), and these already stressed the necessity of a proper evaluation and calibration. Here is important to distinguish, following TAyLOR (1999), between precision (the degree to which repeated measurements show the same results) and accuracy (the degree of closeness of measurements). Even if IRT are highly precise, their accuracy cannot be

simply assumed, but tested to prevent or correct systematic biases (ALFORD & RAWLEy, 2007; HARE et al., 2007; ROWLEy & ALFORD, 2007). Here, parallel measurements using both types of thermometers are compared to determine the accuracy of IRT in small lacertids and the eventual pattern of bias if IRT readings are not accurate. It is assumed that CT provides a reliable measure of the lizard’s core body temperature regardless it could result from lizard disturbance induced by the researcher. MATERIALS AND METHODS A set of 52 adult lizards belonging to four different Iberian Podarcis forms were used as models for the experimental tests, namely, three representatives of the P. hispanica species complex (CARRETERO, 2008; kALIONTzOPOULOU et al., 2011, 2012), P. hispanica Galera type (N = 10, Galera, Granada province, 37.744429° N, 2.549562° W), P. vaucheri Southern Spain (N = 9, Chiclana de la Frontera, Cadiz province, 36.366937° N, 6.179339° W) and P. hispanica type 1B (N = 19, Alba de Tormes, Salamanca province, 40.825812° N, 5.515328° W), as well as P. muralis (N = 14, Tanes, Asturias province, 43.205417° N, 5.400220° W). Lizards of each locality were collected by noosing (GARCíA-MUñOz & SILLERO, 2010) and brought to the laboratory. These samples, including males, pregnant and non-pregnant females, were considered as representative of the small lacertid lizards. Snout-vent lengths (SVL, mean ± SE), measured with a digital calliper to the nearest 0.01 mm, were 45.21 ± 0.71 for P. hispanica Galera type, 58.53 ± 0.85 for P. vaucheri S Spain, 55.77 ± 0.10 for P. hispanica type 1B and 55.98 ± 0.91 for P. muralis, respectively.

INFRARED VS. CONTACT THERMOMETERS IN LACERTIDS

Lizards were kept in individual terraria with food (Tenebrio molitor larvae and grasshoppers) and water provided ad libitum and under a natural regime of light and temperature. After an acclimatisation period of less than one week, they were individually exposed to a thermal gradient (~20-50ºC, 0.3 × 0.4 × 1.0 m length experimental terrarium, see VERíSSIMO & CARRETERO, 2009) induced by a 100 W infrared reflector bulb fixed 15 cm above a sand substrate, maintaining the external natural photoperiod. The bulb was switched on with the lizards inside the terrarium one hour before the first measurement and voluntary temperatures were then recorded every hour at ten time intervals, between 9:00 and 18:00 (GMT). Skin temperature was first measured with an IRT (Fluke® 68, precision 0.1ºC, accuracy according to the manufacturer ±1%) directing the laser pointer to the centre of the lizard’s back in line with the body axis (HARE et al., 2007), at a distance of approximately 20 cm. According to the manufacturer specifications, this should restrict the area of infrared measurement to 2 cm. After no more than 10 seconds since the IRT reading, the lizard was collected and the cloacal temperature was measured with a CT (Hibok® 18, precision 0.1ºC, accuracy according to the manufacturer ±0.2%) associated to a k-thermocouple probe. Time between capture and CT reading did not exceed 10 seconds, which was sufficient for the reading to stabilise. After the experiments, lizards were again supplied with water and food ad libitum and then released in the capture sites. Data were not transformed since distributions did not deviate from normality (Lilliefors tests, P > 0.05), were homoscedas-

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tic (univariate Levene tests and multivariate Box M, P > 0.05) and variances and means were uncorrelated. Tests were addressed to compare and relate the temperature measurements of IRT and CT, but not lizard populations or classes. However, since measurements were repeated for the same individual, analysis of variance (ANOVA) for repeated measures was performed to compare their means. Sphericity assumption was tested through Mauchley’s tests. Pearson’s correlations between both temperature readings were calculated. However, because both measurements carried an error, models of temperature relationships were constructed using standardised major axis regression (SMA) using the (S)MATR software (v.2 FALSTER et al., 2006). The remaining statistical tests were performed in Statistica 10 (STATSOFT, 2011). RESULTS In total, 520 paired measurements were recorded. Temperatures (Table 1) varied between 23.60 and 45.40ºC for the IRT and between 24.00 and 37.90ºC for the CT. Results of the ANOVA for repeated measures considering the 10 time intervals (Table 2) indicated thermometer type as the main factor of variation for temperature readings, which did not interact with the remaining factors. The IRT values tended to be systematically higher than the CT and this effect persisted for all Podarcis forms and time intervals (Fig. 1). This result remained even when only the temperatures recorded in the first interval were considered (Table 2). Although the measurements of IRT and CT were significantly correlated, determination coefficients were only moderate. This was

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Table 1: Descriptive statistics of the temperature measurements by thermometer type and Podarcis form.

Podarcis forms P. hispanica Galera P. vaucheri S Iberia P. hispanica type 1B P. muralis Pooled

Skin temperature IRT (ºC) Mean ± SE Range

N 100 90 190 140 520

33.38 ± 0.37 32.70 ± 0.51 33.67 ± 0.28 33.71 ± 0.22 33.46 ± 0.16

Table 2: Results of the ANOVA for repeated measures on the temperature measurements considering the thermometer type (thermometer), the Podarcis form (form) and the time interval (time).

24.80 - 40.50 23.60 - 31.88 24.30 - 45.40 25.70 - 40.10 23.60 - 45.40

Cloacal temperature CT (ºC) Mean ± SE Range 31.65 ± 0.27 31.88 ± 0.34 31.87 ± 0.19 31.80 ± 0.14 31.81 ± 0.11

24.00 - 35.60 25.60 - 37.90 24.80 - 37.80 27.40 - 36.70 24.00 - 37.90

ferences were detected across populations or time intervals (test for common slopes = 48.401, P = 0.448). DISCUSSION

ANOVA (10 time intervals) form thermometer thermometer*form time time*form thermometer*time thermometer*time*form ANOVA (1st time interval) form thermometer thermometer*form

d. f.

0.71 98.83 1.16 1.41 1.41 1.14 0.67

3 1 3 9 27 9 27

0.56 < 10-6 0.35 0.18 0.09 0.34 0.89

d. f.

1.44 18.22 1.40

3 1 3

0.25 < 10-5 0.26

true both for the pooled data (Fig. 2), for the four Podarcis populations separately (R2 = 0.487, y = 1.473 x – 13.275 in P. hispanica Galera type; R2 = 0.442, y = 1.351 x – 9.384 in P. vaucheri S Spain; R2 = 0.308, y = 1.589 x – 16.807 in P. hispanica type 1B; and R2 = 0.668, y = 1.502 x – 15.187 in P. muralis) and for the different time intervals (not displayed). In all cases, the SMA regression lines deviated significantly from slope = 1 and intercept = 0 (Fig. 2, F = 162.59, P < 0.01 and T = -9.149, P < 0.01, respectively, for the pooled data), while no slope dif-

As expected, temperatures recorded with both types of thermometers were obviously related. However, such relation was not close, with IRT measurements deviating significantly from internal temperatures recorded with CT. Specifically, IRT values tended to be higher than CT ones, regardless other circumstances. More importantly, such bias was not constant (same slope, different intercept), but increased with the temperature (higher slope, different intercept). Remarkably, this bias cannot be attributed to the disturbance due to lizard manipulation, since it remained similar during the first temperature readings before lizards were ever captured inside the terrarium. In other words, regarding CT, IRT measurements were precise but not accurate (sensu TAyLOR, 1999). What caused this bias? Biophysical models (STEVENSON, 1985; FEI et al., 2012) allow the discard of substantial differences of temperature between the skin and core body in small lizards (<100 g). Two recent studies comparing parallel IRT and CT measures reported dissimilar results for amphibian and lizards. While

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Figure 1: Variation of the temperature measurements using infrared (IRT) and contact (CT) thermometers according to the Podarcis form and time interval.

Figure 2: Comparison of the infrared (IRT) and contact (CT) temperature measurements pooled for the whole lizard sample.

correspondence was almost complete in bulkyshaped anurans (ROWLEy & ALFORD, 2007), it was strongly dependent on body mass and orientation in slender lizards (HARE et al., 2007). This suggests that, regardless the cautions taken with thermometer orientation and distance, the small, elongated body of these

lacertids provided an insufficient (and probably variable) skin area for the infrared sensor, which would have also measured the background (substrate) infrared radiation (HARE et al., 2007). Further experiments with a broader spectrum of lizard sizes and species should determine to what extent the recorded bias depends on body mass, shape and heat source not only in lacertids but also in other families. In principle, large, bulky and tigmothermic lizards would be expected to display lower biases (HARE et al., 2007), although phylogenetic influences (via physiology) cannot be discarded. Accounting and correcting the bias between IRT and CT may reveal more difficult than simply applying the general equation extracted from the linear regression. On one hand, the relation between both readings was not strong, despite temperatures taken in terrarium minimised measurement error in terms of distance to

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the lizard, reading area and time. Usual conditions in field work are predicted to make nothing but increasing the bias. Certainly, the bias could be minimised if the IRT could be approached to a minimal distance of the skin, but this would imply the disturbance or even the lizard capture (VASCONCELOS et al., 2012), hence, their invasiveness approaching that of CT. On the other hand, even if no IRT-CT slope differences were detected between groups, variation in body temperatures across species, classes, seasons and time intervals has commonly been reported in both lab and field conditions (CASTILLA et al., 1999; CARRETERO et al., 2005, 2006; VERíSSIMO & CARRETERO, 2009). Thus, according to the pattern observed, shifts are expected to be higher in the most thermophile groups, suggesting that IRT would accentuate eventual inter- and intraspecific differences. In conclusion, despite their low invasiveness, infrared thermometers provide biased data regarding internal body temperatures records, compared to contact thermometers. This prevents their uncritical use, at least in small lacertids. Interspecific comparisons between species, sexes, size classes and time intervals with different body temperatures expected are to be evaluated with caution and mixing temperature data coming from IRT and CT is not recommended. Acknowledgement Supported by the projects PTDC/BIABEC/101256/2008 and PTDC/BIABEC/101256/2008 funded by Fundação para a Ciência e a Tecnologia (FCT, Portugal). Collecting permits provided by Junta de Andalucía, Junta de Castilla-León and Principado de Asturias. Experiments

following the guidelines of University of Porto (Portugal). Several colleagues of CIBIO and N. Sillero helped in the field work. Two anonymous reviewers contributed with their comment to the improvement of an earlier draft of the manuscript. REFERENCES ALFORD, R.A. & ROWLEy, J.J.L. (2007). Comment on Papers by Hare et al. and Rowley and Alford. Herpetological Review 38: 316. ANGILLETTA, M.J. JR. (2010). Thermal adaptation. Oxford University Press, Oxford. CASTILLA, A.M.; VAN DAMME, R. & BAUWENS, D. (1999). Field body temperatures, mechanisms of thermoregulation and evolution of thermal characteristics in lacertid lizards. Natura Croatica 8: 253-274. CARRETERO, M.A. (2008). An integrated assessment of the specific status in a group with complex systematics: the Iberomaghrebian lizard genus Podarcis (Squamata, Lacertidae). Integrative Zoology 4: 247-266. CARRETERO, M.A.; ROIG, J.M. & LLORENTE, G.A. (2005). Variation in preferred body temperature in an oviparous population of Lacerta (Zootoca) vivipara. Herpetological Journal 15: 51-55. CARRETERO, M.A.; MARCOS, E. & DE PRADO, P. (2006). Intraspecific variation of preferred temperatures in the NE form of Podarcis hispanica, In C. Corti; P. Lo Cascio & M. Biaggini (eds.) Mainland and Insular Lacertid Lizards: a Mediterranean Perspective. Firenze University Press, Florence, pp. 55-64. FALSTER, D.S., WARTON, D.I. & WRIGHT, I.J. (2006). SMATR: standardised major axis tests and routines, Version 2.0. Available at:

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http://www.bio.mq.edu.au/ecology/SMAT R/. Retrieved on 02/24/2012. FEI, T.; SkIDMORE, A.k.; VENUS, B.; WANG, T.; SCHLERF, M.; TOXOPEUS, B.; VAN OVERJIJk, S.; BIAN, M. & LUI, y. (2012). A body temperature model for lizards as estimated from the thermal environment. Journal of Thermal Biology 37: 56-64. GARCíA-MUñOz, E. & SILLERO, N. (2010). Two new types of noose for capturing herps. Acta Herpetologica 5: 259-263. HARE, J.R.; WHITWORTH, E. & CREE, A. (2007). Correct orientation of a handheld infrared thermometer is important for accurate measurements of body temperatures in small lizards and Tuatara. Herpetological Review 38: 311-315. JONES, S.M. & AVERy, R.A. (1989). The use of a pyroelectric vidicon infra-red camera to monitor the body temperatures of small terrestrial vertebrates. Functional Ecology 3: 373-377. kALIONTzOPOULOU, A.; PINHO, C.; HARRIS, D.J. & CARRETERO, M.A. (2011). When cryptic diversity blurs the picture: a cautionary tale from Iberian and North African Podarcis wall lizards. Biological Journal of the Linnean Society 103: 779-800. kALIONTzOPOULOU, A.; CARRETERO, M.A. & LLORENTE, G.A. (2012). Morphology of the Podarcis wall lizards (Squamata: Lacertidae) from the Iberian Peninsula and North Africa: patterns of variation in a putative cryptic species complex. Zoological Journal of Linnean Society 164: 173-193. LANGkINDE, T. & SHINE, R. (2006). How much stress do researchers inflict on their study animals? A case study using a scincid lizard, Eulamprus heatwolei. Journal of Experimental Biology 209: 1035-1043.

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M OORE, M.C.; THOMPSON, C.W. & MARLER, C.A. (1991). Reciprocal changes in corticosterone and testosterone levels following acute and chronic handling stress in the Tree Lizard, Urosaurus ornatus. General and Comparative Endocrinology 81: 217-226. ROWLEy, J.J.L. & ALFORD, R.A. (2007). Non-contact infrared thermometers can accurately measure amphibian body temperatures. Herpetological Review 38: 308-311. STEVENSON, R.D. (1985). Body size and limits to the daily range of body temperature in terrestrial ectotherms. American Naturalist 125: 102-117. STATSOFT, INC. (2011). STATISTICA (data analysis software system), version 10. Available at http://www.statsoft.com. TAyLOR, J.R. (1999). An Introduction to Error Analysis: The Study of Uncertainties in Physical Measurements. University Science Books, Sausalito, California, USA. TOSINI, G., JONES, S. & AVERy, R.A. (1995). Infra-red irradiance and set point temperatures in behaviourally-thermoregulating lacertid lizards. Journal of Thermal Biology 20: 497-503. VASCONCELOS, R.; SANTOS, X. & CARRETERO, M.A. (2012). High temperatures constrain microhabitat selection and activity patterns by the insular Cape Verde wall gecko. Journal of Arid Environments 81: 18-25. VERíSSIMO, C.V. & CARRETERO, M.A. (2009). Preferred temperatures of Podarcis vaucheri from Morocco: intraspecific variation and interspecific comparisons. Amphibia-Reptilia 30: 17-23.

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Basic and Applied Herpetology 26 (2012): 107-116

Synopsis of the helminth communities of the lacertid lizards from the Balearic and Canary Islands Vicente Roca1,2,*, Fátima Jorge2, Miguel Ángel Carretero2 1 2

Departament de Zoologia, Facultat de Ciències Biològiques, Universitat de València, Spain. CIBIO-UP, Centro de Investigação em Biodiversidade e Recursos Genéticos, Vairão, Portugal.

* Correspondence: Departament de Zoologia, Facultat de Ciències Biològiques, Universitat de València. Dr. Moliner 50, 46100 Burjassot, Spain. Phone: +34 963544606, Fax: +34 963543049, E-mail: Vicente.roca@uv.es

Received: 16 October 2012; received in revised form: 15 November 2012; accepted: 18 November 2012.

Helminth communities of reptiles have usually been considered as depauperate and isolationist, with low abundance and species richness when compared to other vertebrates. Nevertheless there are some insular reptile populations in which this general rule is not fulfilled. In this study, we compare helminth faunas from two groups of lizards living in two different archipelagos and having different feeding habits. Lacertid lizards from Canary Islands, belonging to the endemic genus Gallotia, showed by contrast with other lacertids, a tendency to high consumption of plant matter and to rich and diverse helminth communities. Differences were found even between the lizards living in different islands, being the “giant lizard” G. stehlini the most herbivorous and G. atlantica the most carnivorous. Podarcis spp. from Balearic Islands showed lower tendency to herbivory which was not mirrored in their helminth communities. The composition and structure of helminth communities of lacertid lizards from both archipelagos are related to the conditions of insularity and the phylogeny of the hosts. Key words: Balearic Islands; Canary Islands; helminthes; lizards. Sinopsis de las comunidades helmínticas de los lagartos de las Islas Baleares y Canarias. Las comunidades helmínticas de los reptiles han sido consideradas generalmente como depauperadas y aislacionistas, con escasa riqueza y abundancia de especies comparadas con otros vertebrados. Sin embargo esta norma general no se cumple en algunas poblaciones de reptiles insulares. En el presente estudio se comparan las helmintofaunas de dos grupos de lagartos que habitan dos archipiélagos diferentes, y que exhiben distintas estrategias de alimentación. Los lagartos de las Islas Canarias, pertenecientes al género Gallotia, mostraron, en contraste con otros lacértidos, una tendencia hacia un alto consumo de materia vegetal y al establecimiento de ricas y diversas comunidades helmínticas. Se encontraron diferencias entre lagartos que habitan diferentes islas, siendo G. stehlini el más herbívoro y G. atlantica el más carnívoro. Las lagartijas Podarcis spp. de las Islas Baleares mostraron menor tendencia hacia el herbivorismo, tendencia que no se ve reflejada en sus comunidades helmínticas. La composición y estructura de las comunidades helmínticas de los lagartos de ambos archipiélagos se conforman de acuerdo con las condiciones impuestas por la insularidad y por la filogenia de los hospedadores. Key words: Islas Baleares; Islas Canarias; helmintos; lagartos.

Studies on the community ecology of parasites of European reptiles, and particularly lacertids, have increased in the last years. In this context, islands have received less attention, although they are very interesting areas since

they show peculiar conditions that may influence the characteristics of the populations of both parasites and hosts. From the point of view of the lizard hosts, pressure by terrestrial predator is lower for insular lizards than for

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their continental relatives (NOVOSOLOV et al., in press). This results in: (i) high lizard densities increasing intraspecific competition (including cannibalism on juveniles) and favouring conservative reproductive strategies (CARRETERO, 2006); (ii) lower, unpredictable prey availability promoting alternative trophic strategies, including kleptoparasitism and herbivory (PéREz-MELLADO & CORTI, 1993; VAN DAMME, 1999; CARRETERO, 2004); (iii) a trend to larger body size and more accentuated sexual dimorphism (MEIRI, 2007; NOVOSOLOV et al., in press). Last but not least, the abovementioned processes display substantial phylogenetic signal and depend on the time of evolution in insular conditions (CARRETERO, 2006). All these aspects are potentially important for the formation of parasite communities, namely in terms of promoting reproductive isolation between islands, increasing probability of infestation by conspecifics while decreasing that from heterospecifics, creating new infestation pathways and providing different usually more complex host environments. The island syndrome has also been reported in nematode parasites, with loss of genetic diversity and a niche enlargement (NIEBERDING et al., 2006). The founder host may reach islands with only a subset of their parasite fauna, resulting in a decrease in species richness when compared with the continental relatives. This loss of richness has been detected also in helminth species of small mammals of Mediterranean islands, which was correlated with the area of the island and distance from mainland and decrease of host specificity (MAS-COMA et al., 2000; GOüy DE BELLOCQ et al., 2002). In the two last decades our laboratory has carried out studies about helminth

parasites from Mediterranean and Atlantic islands (i.e. ROCA et al., 1987; ROCA, 1993, 1999; ROCA & HORNERO, 1994; MARTIN & ROCA, 2004a,b; CARRETERO et al., 2006; ROCA et al., 2006, 2009; JORGE et al., 2012). We focused on the archipelagos of the Balearic and Canary Islands both administratively belonging to Spain but located at different geographic areas, namely in the Mediterranean Sea and the Atlantic Ocean, respectively (Fig. 1a), which have undergone quite divergent paleogeographic histories (see below). Lacertid lizards inhabiting each archipelago are also different. The Balearic Islands are occupied by two endemic species of the genus Podarcis Wagler, 1830, widely ranging the Mediterranean Basin, belonging to the Palearctic clade of Lacertini (ARNOLD et al., 2007). By contrast, lacertids living in the Canary Islands are represented by species of the endemic genus Gallotia Boulenger, 1916, a lineage phylogenetically distant from most other lacertids including Lacerta or Podarcis (HARRIS et al., 1998; ARNOLD et al., 2007; PAVLICEV & MAyER, 2009). Abovementioned characteristics of the history of both archipelagos and their lizards, and also reproductive isolation conducting to new host species and subspecies (MACA-MEyER et al., 2003; COX et al., 2010), and so new possibilities for new parasite species (and subspecies), could give rise to the current parasite faunas of these hosts. Canary Islands are volcanic. They erupted from the sea and are considered as “oceanic islands”, that is, never being connected to the African continent. They are located off the north west coast of Africa (Fig. 1a), at 27º37’-29º24’ N, 13º37’-8º10’ W, and

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Figure 1: Localization of the Balearic and Canary Islands. (a) General localization. (b) Canary Islands. (c) Balearic Islands.

comprises seven main islands and a number of peripheral islets (Fig. 1b) having a total surface of 7493.65 km2 and being their maximum elevation 3718 m above sea level (Teide Volcano in Tenerife island). They form the biogeographical region of Macaronesia together with Cape Verde, Madeira, Azores and Selvagens archipelagos (Fig. 1a). Balearic Islands are “continental islands” that were part of the continent before reaching to be islands. In fact they constituted a part of the Iberian Peninsula known as “Balearic headland”. Successive fractures and isolations gave rise to the actual conformation of this archipelago (CAVAzzA & WEzEL, 2003). They are located east of the Iberian Peninsula (Fig. 1a), at 40º05’-38º38’ N, 4º19’-1º09’ E, and comprises five main islands and a number of peripheral islets (Fig. 1c) having a total surface of 5014 km2, ranging in elevation from the sea level to 1445 m above sea level. In this study we synthesize the results obtained in the parasitological analysis of lizards of both archipelagos and we compare their parasite faunas in the light of the biotic and abiotic features of the hosts.

MATERIALS AND METHODS Gallotia lizards from the Canary Islands were captured in the main islands (Lanzarote, Fuerteventura, Gran Canaria, Tenerife, La Palma, La Gomera and El Hierro) during several periods between 1994 and 1997. From 1987 to 1989, Podarcis lilfordi were captured in some peripheral islets surrounding Mallorca and Menorca islands (Gymnesic islands) and Podarcis pityusensis were captured in Pityusic islands (Ibiza and Formentera) and some peripheral islets. The number of specimens of sampled lizards is detailed in Table 1. All lizards were captured by hand and were killed with an overdose of chloroform. The body cavity, digestive tract, heart, lungs, and liver were removed, opened, and placed in Ringer’s solution for examination. Helminthes were removed, washed in distilled water, fixed, and mounted according to standards techniques. Parasites were identified, when possible, to species, and the number and location of individuals of each parasite species were recorded. The use of descriptive ecological terms follows BUSH et al. (1997). Brillouin’s index

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Table 1: Host species and number of individuals analyzed in each archipelago. Island

Lizard Species

Balearic Islands Gymnesian (islets surrounding Mallorca and Menorca)* Pityusic (islets surrounding Ibiza and Formentera)* Canary Islands Lanzarote Fuerteventura Gran Canaria Tenerife La Palma La Gomera El Hierro

Number of inspected hosts

Podarcis lilfordi Podarcis pityusensis

386 564

Gallotia atlantica Gallotia atlantica Gallotia stehlini Gallotia galloti Gallotia galloti Gallotia caesaris Gallotia caesaris

70 42 33 27 27 21 318

*For details see ROCA & HORNERO (1994).

of diversity was used for calculating diversity according to MAGURRAN (2004). RESULTS Seventeen helminth species (two Trematoda, four Cestoda, 10 Nematoda and one Acanthocephala) were found in the Balearic hosts. Twenty four (one Trematoda, five Cestoda, 17 Nematoda and one Acanthocephala) were recorded in Canarian hosts. Table 2 shows the presence/absence of helminth species recorded in the hosts. Global prevalence of infection and overall diversity parameters for each host are given in Fig. 2 and Table 3, respectively. In all the hosts, Pharyngodonidae nematodes (species of the genera Skrjabinodon, Spauligodon, Parapharyngodon, Thelandros, Tachygonetria and Alaeuris) were the main component of their helminthes infracommunities. The helminthes found as larval forms were located in the body cavity of the hosts, whereas those found as adults were located at different sites (Table 2).

DISCUSSION Helminth faunas of lizards from both archipelagos were globally similar in including widespread helminth species, as the nematodes Skrjabinodon medinae and Parapharyngodon spp., the cestodes Nematotaenia tarentolae and Oochoristica spp., and several cestode and nematode larval forms. Nevertheless, some differences are also found between both helminth faunas, namely regarding the endemisms. Pseudoparadistomum yaizaensis is an endemic genus and species parasitizing only Gallotia atlantica from Lanzarote Island (ROCA, 2003). This finding was interesting because (i) no dicrocoeliids had been found from other reptiles in Macaronesia; and (ii) another member of the same family was found in the Balearic Islands, Paradistomun mutabile which is distributed in the European Mediterranean Basin (ROCA & HORNERO, 1994; ROCA, 1995). Since geographical separation of gene stocks is important in speciation phenomena, we could suggest an origin of P. yaizaensis from P. mutabile, considering that there may have been interchanges of helmin-

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Table 2: Presence/absence of the recorded helminth species in each host. P.l. Podarcis lilfordi; P.p. P. pityusensis; G.a. Gallotia atlantica; G.s. G. stehlini; G.g. G. galloti and G.c. G. caesaris. Helminth species Digenea Paradistomum mutabile Pseudoparadistomum yaizaensis Brachylaima sp. Cestoda Oochoristica agamae Oochoristica gallica Nematotaenia tarentolae Diplopylidium acanthotetra (larvae) Dipylidium sp. (larvae) Mesocestoides sp. (larvae) Nematoda Skrjabinodon medinae Spauligodon cabrerae Spauligodon atlanticus Spauligodon sp.* Parapharyngodon micipsae Parapharyngodon echinatus Parapharyngodon bulbosus Thelandros galloti Thelandros filiformis Thelandros tinerfensis Alaeuris numidica Tachygonetria dentata Tachygonetria macrolaimus Tachygonetria conica Tachygonetria numidica Skrjabinelazia hoffmanni Skrjabinelazia pyrenaica Strongyloides ophiusensis Abbreviata sp. Acuaria sp. (larvae) Spirurida gen sp. (larvae) Acanthocephala Centrorhynchus sp. (larvae)

Site

P.l.

P.p.

G.a.

G.s.

G.g.

G.c.

Gall bladder Small intestine Small intestine Small intestine Small intestine Small intestine Body cavity Body cavity Body cavity Caecum Caecum Caecum Caecum Caecum Caecum Caecum Caecum Caecum Caecum Caecum Caecum Caecum Caecum Caecum Small intestine Stomach Small intestine Small intestine Body cavity Body cavity Body cavity

*Unpublished data.

thes between both groups of lacertid lizards, Podarcis and Gallotia, as suggested by the composition of both helminth faunas. Other helminth species or subspecies endemic to the Canary Islands, as Spauligodon atlanticus, Spauligodon sp, Thelandros galloti, T. filiformis, T. tinerfensis and Alaeuris numidica canariensis

were Gallotia specialists (sensu EDWARDS & BUSH, 1989) since they have been found in several Gallotia lizards and not in other hosts. Balearic lizards also harboured only two endemic helminth species, one of them, Strongyloides ophiusensis, being found in Podarcis pityusensis but not in P. lilfordi. One of

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Figure 2: Overall prevalence of infection of the searched lizard hosts.

the most interesting differences among the endemisms is related to the Spauligodon species. In the Gallotia hosts, two different not related but resembling Spauligodon species are present, S. atlanticus parasite of G. atlantica and Spauligodon sp. found in the Gallotia species from the western islands (JORGE et al., 2011). Their colonization patterns are still unclear, and could have resulted from the host only colonizing the islands with a different subset of the two species, or extinctions or one of this species resulting from a host switch from another species. In the Podarcis present in the Balearic Islands, only one species has been found, S. cabrerae, probably resulting from colonization by descent.

Family Pharyngodonidae is considered as good indicator of the diet of reptiles (ROCA, 1999). In the different genera of this family parasitizing reptiles, two lines can be distinguished based mainly on the disposition of the genital papillae (PETTER, 1966). One of these lineages evolved in carnivorous reptiles (genera Pharyngodon, Spauligodon, Skrjabinodon and Parapharyngodon) and the other one in herbivorous reptiles (genera Alaeuris, Mehdiella, Tachygonetria and Thelandros) (PETTER & QUENTIN, 1976; ROCA, 1999). Both groups of Balearic and Canarian lizards harboured species belonging to the lineage of carnivorous reptiles, but

Table 3: Overall diversity parameters (mean ± SD, with the range in parentheses) from the searched hosts. Host Balearic Islands Podarcis lilfordi Podarcis pityusensis Canary Islands Gallotia atlantica Gallotia stehlini Gallotia galloti Gallotia caesaris

No. of helminth species/host

No. of helminthes/host

Brillouin’s diversity index

0.93 ± 0.74 (0 - 4) 1.35 ± 1.02 (0 - 5)

7.45 ± 12.84 (0 - 110) 16.44 ± 31.22 (0 - 420)

0.108 ± 0.200 (0 - 0.815) 0.242 ± 0.292 (0 - 1.211)

1.0 ± 0.8 (0 - 4) 4.9 ± 1.6 (2 - 7) 1.72 ± 0.18 (0 - 4) 1.4 ± 1.0 (0 - 5)

16.3 ± 21.2 (0 - 156) 468.2 ± 644.6 (31 - 2734) 54.4 ± 14.28 (0 - 373) 47.3 ± 68.5 (0 - 350)

0.13 ± 0.24 (0 - 0.89) 1.2 ± 0.4 (0.2 - 1.7) 0.55 ± 0.09 (0-1.86) 0.3 ± 0.4 (0 - 1.0)

HELMINTHES OF LIZARDS FROM BALEARIC AND CANARY ISLANDS

only Canarian ones had eight helminth species typical of herbivorous reptiles. This suggested a strong tendency to herbivory of these hosts. Diversity of helminth communities has also been used to estimate the tendency of reptiles to eat plant matter. Herbivorous reptiles tend to contain richer and more diverse communities than carnivorous ones (ROCA, 1999). According to this, Gallotia stehlini and G. galloti from the Canary Islands showed major tendency to herbivory than other species from the archipelago, because their helminth communities were highly rich and diverse, so we may consider them as rich helminth communities (AHO, 1990). This was in accordance to the presence of Pharyngodonidae typical of herbivorous reptiles (MARTIN & ROCA, 2004a,b). Diet analysis of gut contents (PéREzMELLADO & CORTI, 1993; BROWN & PéREzMELLADO, 1994; ROCA, 1999; CARRETERO et al., 2001; VALIDO & NOGALES, 2003; ROCA et al., 2005; CARRETERO et al., 2006) allowed these Balearic and Canarian lizard species to be located in a continuum of the food type strategy, in extremes are strict carnivory and herbivory. Near of the end of herbivory we found G. stehlini, the “Giant lizard” from Gran Canaria Island which eats more than 94% of plan matter. Gallotia galloti from Tenerife and G. caesaris from El Hierro and La Gomera islands were omnivorous with marked tendency to herbivory (between 70% and 85% of plant matter). Remarkably, Gallotia lizards consumed stems and leaves of the plants, which are of low profitability and require long fermentation periods for symbiotic flora to act (CARRETERO, 2004). By contrast, both Balearic lizards P. pityusensis and P. lilfordi (PéREz-MELLADO & CORTI,

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1993; CARRETERO et al., 2001), and also the populations of G. atlantica from Lanzarote and Fuerteventura islands (authors’ unpublished data), eat mainly arthropods and a little amount of plant matter, usually the most profitable parts such as fruits, seeds, flowers or pollen, and so we found them near of the end of carnivory. Results from the analysis of helminthes (MARTIN & ROCA, 2004a,b; MARTIN & ROCA, 2005; ROCA et al., 2005) placed these lizard species in the same carnivory-herbivory continuum, according to their helminth community features mentioned above. We found a full coincidence of the position of G. stehlini in the continuum, due to its helminthological characteristics (high diversity, presence of many Pharyngodonidae of the lineage of herbivorous reptiles). It is the biggest and the most herbivorous lizard, with the most diverse helminth community. High coincidence was also found in the case of G. galloti showing relatively high diversity and presence of some peculiar species of Pharyngodonidae, as happens with G. caesaris. Low diversity and absence of pharyngodonids typical of herbivorous reptiles placed G. atlantica and both Podarcis at the end of the carnivory, in accordance with the results of their feeding habits. The tendency to a restricted herbivory (consumption of the most energetic parts of plants) observed in Balearic lizards (PéREzMELLADO & CORTI, 1993) is not reflected in their helminth communities. Thus, ROCA (1999) suggested that they are primarily carnivorous reptiles and we suggest the same for G. atlantica. It has been proposed that such differences between both lizard genera are related to the different times of evolution under insular conditions (CARRETERO,

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2004). In particular, the absence or the impoverishment of terrestrial predatory species releasing lizards for devoting more time for eating (PéREz-MELLADO & CORTI, 1993) together with prolonged lizard basking at high temperatures (MÁRQUEz et al., 1997) and the retention of prey items in a compartmented rectal caecum (HERREL et al., 2004) would not only enhance prey digestion (zIMMERMAN & TRACy, 1989), but also provide a favourable ambient for distinctive, more complex helminth communities. In accordance with the reasons mentioned above, we can conclude the following: (i) the composition and structure of helminth communities of these reptile hosts are related to their phylogeny and their feeding habits; (ii) the tendency to the herbivory is much more marked in Canary lacertid lizards than in Balearic ones. Acknowledgement Permits for collecting living specimens were granted by the Viceconsejería de Medio Ambiente of the Canarian Government, and the Conselleria de Agricultura i Pesca of the Balearic Islands Government. The authors carried out the work in accordance with the relevant regulations concerning the capture and handling of animals. REFERENCES ARNOLD, E.N.; ARRIBAS, O.J. & CARRANzA, S. (2007). Systematics of the Palaearctic and Oriental lizard tribe Lacertini (Squamata: Lacertidae: Lacertinae), with descriptions of eight new genera. Zootaxa 1430: 1-8. BROWN, R.P. & PéREz-MELLADO, V. (1994). Ecological energetics and food acquisition in

dense Menorcan islet populations of the lizard Podarcis lilfordi. Functional Ecology 8: 427-434. BUSH, A.O.; LAFERTy, k.D.; LOFT, J.M. & SHOSTAk, A.W. (1997). Parasitology meets ecology on its own terms: Margolis et al. revisited. Journal of Parasitology 83: 575-583. CARRETERO, M.A. (2004). From set menu to a la carte. Linking issues in trophic ecology of Mediterranean lacertids. Italian Journal of Zoology 74: 121-133. CARRETERO, M.A. (2006). Reproductive cycles in Mediterranean lacertids: plasticity and constraints, In C. Corti, P. Lo Cascio & M. Biaggini (eds.) Mainland and Insular Lizards. A Mediterranean Perspective. Firenze University Press, Florence, Italy, pp. 33-54. CARRETERO, M.A.; LLORENTE, G.A.; SANTOS, X. & MONTORI, A. (2001). The diet of an introduced population of Podarcis pityusensis. Is herbivory fixed?, In L. Vicente & E.G. Crespo (eds.) Mediterranean Basin Lacertid Lizards. A Biological Approach. ICN, Lisbon, Portugal, pp. 113-124. CARRETERO, M.A.; ROCA, V.; MARTIN, J.E.; LLORENTE, G.A.; MONTORI, A.; SANTOS, X. & MATEOS, J. (2006). Diet and helminth parasites in the Gran Canaria giant lizard Gallotia stehlini. Revista Española de Herpetología 20: 105-117. CAVAzzA, W. & WEzEL, F.C. (2003). The Mediterranean region, a geological primer. Episodes 26: 160-168. COX, S.C.; CARRANzA, S. & BROWN, R.P. (2010). Divergence times and colonization of the Canary Islands by Gallotia lizards. Molecular Phylogenetics and Evolution 56: 747-757. EDWARDS, D.D. & BUSH, A.O. (1989). Helminth communities in avocets:

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importance of compound community. Journal of Parasitology 98: 439-445. GOüy DE BELLOCQ, J.; MORAND, S. & FELIU, C. (2002). Patterns of parasite species richness of Western Palaearctic micromammals: island effects. Ecography 25: 173-183. HARRIS, D.J.; ARNOLD, E.N. & THOMAS, R.H. (1998). Relationships of lacertid lizards (Reptilia: Lacertidae) estimated from mitochondrial DNA sequences and morphology. Proceedings of the Royal Society of London B 265: 1939-1948. HERREL, A.; VANHOOyDONCk, B. & VAN DAMME, R. (2004). Omnivory in lacertid lizards: adaptive evolution or constrait?. Journal of Evolutionary Biology 17: 974-984. JORGE, F.; ROCA, V.; PERERA, A.; HARRIS, D.J. & CARRETERO, M.A. (2011). A phylogenetic assessment of the colonisation patterns in Spauligodon atlanticus Astasio-Arbiza et al., 1987 (Nematoda: Oxyurida: Pharyngodonidae), a parasite of lizards of the genus Gallotia Boulenger: no simple answers. Systematic Parasitology 80: 53-66. JORGE, F.; CARRETERO, M.A.; PERERA, A.; HARRIS, D.J. & ROCA, V. (2012). A new species of Spauligodon (Nematoda: Oxyurida: Pharyngodonidae) in geckos from Sao Nicolau island (Cape Verde) and its phylogenetic assessment. Journal of Parasitology 98: 160-166. MACA-MEyER, N.; CARRANzA, S.; RANDO, J.C.; ARNOLD, E.N. & CABRERA, V.N. (2003). Status and relationships of the extinct giant Canary Island lizard Gallotia goliath (Reptilia: Lacertidae), assessed using ancient mtDNA from its mummified remains. Biological Journal of the Linnean Society 80: 659-670. MAGURRAN, A.E. (2004). Measuring Biological Diversity. Blackwell Publishing,

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Malden, Massachusetts, USA. MÁRQUEz, R.; CEJUDO, D. & PéREz-MELLADO, V. (1997). Selected body temperatures of four lacertid lizards from the Canary Islands. Herpetological Journal 7: 122-124. MARTIN, J.E. & ROCA, V. (2004a). Helminth infracommunities of Gallotia caesaris caesaris and Gallotia caesaris gomerae (Sauria: Lacertidae) from the Canary Islands (Eastern Atlantic). Journal of Parasitology 90: 266-270. MARTIN, J.E. & ROCA, V. (2004b). Helminth infracommunities of a population of the Gran Canaria giant lizard Gallotia stehlini. Journal of Helminthology 78: 319-322. MARTIN, J.E. & ROCA, V. (2005). Helminths of the Atlantic lizard Gallotia atlantica (Reptilia, Lacertidae), in the Canary Islands (Eastern Atlantic): Composition and structure of component communities. Acta Parasitologica 50: 85-89. MAS-COMA, S.; ESTEBAN, J.G.; FUENTES, M.V.; BARGUES, M.D.; VALERO, M.A. & GALÁNPUCHADES, M.T. (2000). Helminth parasites of small mammals (insectivores and rodents) on the Pityusic island of Eivissa (Balearic Archipelago). Research and Reviews in Parasitology 60: 41-49. MEIRI, S. (2007). Size evolution in island lizards. Global Ecology and Biogeography 16: 702-708. NIEBERDING, C.; MORAND, S.; LIBOIS, R. & MICHAUX, J.R. (2006). Parasites and the island syndrome: the colonization of the western Mediterranean islands by Heligmosomoides polygyrus (Dujardin, 1845). Journal of Biogeography 33: 1212-1222. NOVOSOLOV, M.; RAIA, P. & MEIRI, S. (in press). The island syndrome in lizards. Global Ecology and Biogeography DOI: 10.1111/j.1466-8238.2012.00791.x.

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PAVLICEV, M. & MAyER, W. (2009). Fast radiation of the subfamily Lacertinae (Reptilia: Lacertidae): History or methodical artefact? Molecular Phylogenetics and Evolution 52: 727-734. PéREz-MELLADO, V. & CORTI, C. (1993). Dietary adaptations and herbivory in lacertid lizards of the genus Podarcis from western Mediterranean islands (Reptilia: Sauria). Bonner Zoologische Beiträge 44: 193-220. PETTER, A.J. (1966). équilibre des espèces dans les populations de nématodes parasites du colon des tortues terrestres. Mémoires du Muséum Nationale d’Histoire Naturelle Série A Zoologie 39: 1-252. PETTER, A.J. & QUENTIN, J.C. (1976). Keys to Genera of the Oxyuroidea. Series: C.I.H. keys to the Nematode Parasites of Vertebrates, vol. 4 (R.C. Anderson, A.G. Chabaud & S. Willmott, eds.). CAB International, Farhan Royal, London, Uk. ROCA, V. (1993). Helmintofauna dels rèptils, In J.A. Alcover, E. Ballesteros & J.J. Fornós (eds.) Història Natural de l’Arxipèlag de Cabrera. CSIC-Editorial Moll, Palma de Mallorca, Spain, pp. 273-292. ROCA, V. (1995). An approach to the knowledge of the helminth infracommunities of Mediterranean insular lizards (Podarcis spp.), In G.A. LLorente, A. Montori, X. Santos & M.A. Carretero (eds.) Scientia Herpetologica. Asociación Herpetológica Española, Barcelona, Spain, pp. 285-292. ROCA, V. (1999). Relación entre las faunas parásitas de reptiles y su tipo de alimentación. Revista Española de Herpetología 13: 101-121. ROCA, V. (2003). A new genus of Dicrocoeliidae (Digenea) from the lizard Gallotia atlantica (Sauria: Lacertidae)

from the Canary Islands (Spain). Journal of Natural History 37: 1401-1406. ROCA, V. & HORNERO, M.J. (1994). Helminth infracommunities of Podarcis pityusensis and Podarcis lilfordi (Sauria: Lacertidae) from the Balearic Islands (western Mediterranean Basin). Canadian Journal of Zoology 72: 658-664. ROCA, V.; GARCíA-ADELL, G.; LóPEz, E. & zAPATERO-RAMOS, L.M. (1987). Algunas formas adultas y larvarias de platelmintos de reptiles de las islas Canarias. Revista Ibérica de Parasitología 47: 263-270. ROCA, V.; CARRETERO, M.A.; LLORENTE, G.A.; MONTORI, A. & MARTIN, J.E. (2005). Helminth communities of two lizard populations (Lacertidae) from Canary Islands (Spain): Host-diet parasite relationships. Amphibia-Reptilia 26: 535-542. ROCA, V.; LO CASCIO, P. & MARTIN, J.E. (2006). Gastrointestinal parasites in saurians from some central Mediterranean islands. Boletín de la Asociación Herpetológica Española 17: 54-58. ROCA, V.; FOUFOPOULOS, J.; VALAkOS, E. & PAFILIS, P. (2009). Parasitic infracommunities of the Aegean wall lizard Podarcis erhardii (Lacertidae, Sauria): isolation and impoverishment in small island populations. Amphibia-Reptilia 30: 493-503. VALIDO, A. & NOGALES, M. (2003). Digestive ecology of two omnivorous Canarian lizard species (Gallotia, Lacertidae). AmphibiaReptilia 24: 331-344. VAN DAMME, R. (1999). Evolution of herbivory in lacertid lizards: effects of insularity and body size. Journal of Herpetology 33: 663-674. zIMMERMAN, L.C. & TRACy, C.R. (1989). Interactions between the environment and ectothermy and herbivory in reptiles. Physiological Zoology 62: 374-409.

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REFERENCES For references in the text give full surnames of the first author followed by the publication year and separated by a comma (Pleguezuelos, 1997). For papers with two authors, use the term “&” to separate surnames (Semlitsch & Bodie, 2003). Papers with three or more authors will be quoted with the surname of the first author followed by ‘et al.’ (note italics) (Stuart et al., 2004). To distinguish between two papers by the same author(s) in the same year use lower-case letters (a,b) after the year, without space, arranged in alphabetical order as the references are quoted in the text (Harris et al., 2004a,b). List multiple citations in chronological order, using alphabetical order for citations within the same year. Separate citations with semicolons (Tyler, 1991; Wake, 1991; Blaustein et al., 1994a,b; Stuart et al., 2004). If the citation is part of the sentence, move the surname(s) of the author(s) out of the brackets and delete the comma. “As pointed by Pleguezuelos (1997)” “Blaustein et al. (1994a) reviewed the situation of amphibians” The reference list should include all and only the references mentioned in the text, tables and figures. Cite references in the reference list in alphabetical order according to the authors' surnames. Multiple citations for the same author should be organized as follows: single citations first (in chronological order), two-author citations second (in alphabetical order), three or more authors third (in chronological order). Spell out (i.e. do not abbreviate) the names of all journals. The references should conform to the following formats:

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Articles in periodicals: • Stuart, S.N.; Chanson, J.S.; Cox, N.A.; young, B.E.; Rodrigues, A.S.L.; Fishman, D.L. & Waller, R.W. (2004). Status and trends of amphibian declines and extinctions worldwide. Science 306: 1783-1786. • Wiens, J.J. & Penkrot, T.A. (2002). Delimiting species using DNA and morphological variation and discordant species limits in spiny lizards (Sceloporus). Systematic Biology 51: 69-91. Books: • Dodd, Jr., C.k. (ed.) (2009). Amphibian Ecology and Conservation. A Handbook of Techniques. Oxford University Press, Oxford. • Vitt, L.J. & Caldwell, J.P. (2009). Herpetology: An Introductory Biology of Amphibians and Reptiles, 3rd ed. Academic Press, Burlington, Massachusetts. Book chapters: • king, R.B. (2009). Population and conservation genetics, In S.J. Mullin & R.A. Seigel (eds.) Snakes: Ecology and Conservation. Cornell University Press, Ithaca, New york, pp. 78-122. Web pages (authors are recommended to keep the use of web pages to a minimum; use peerreviewed literature instead when possible): • IUCN (2010). The IUCN Red List of Threatened Species, v. 2010.3. International Union for Nature Conservation and Natural Resources, Gland, Switzerland. Available at http://www.iucnredlist.org/. Retrieved on 10/31/2010.

SUPPORTING MATERIAL Authors can submit with their manuscripts supporting material related to the work (additional tables and figures, detailed protocols, data logs, audio and video recordings, etc.). The supporting material will be uploaded to the online site of B&AH with a reference code that will be used to quote such material in the final version of the article. In the initial version of the manuscript, supporting material should be quoted as “SM” followed by a number according to the same format as for tables and figures. Supporting material must be submitted as independent files. Name each file with the code used in the initial version of the manuscript (SM1, SM2, etc.). Authors may also refer to supporting material available from a different online site (e.g. GenBank, MorphoBank), in which case the exact access reference will be indicated in the final version of the article.

BIOETHICAL CONSIDERATIONS Because right animal use and care is an area of major concern to the AHE, authors must guarantee that all animals used for research purposes are treated ethically and in accordance with the laws and regulations established by governmental authorities and bioethics committees of each institution. Therefore, authors are recommended to state in the Acknowledgement section that they have followed the corresponding regulation and legislation on animal care. Authors should cite in this section the information regarding collection permits and experimental protocols approved by bioethics or animal care committees. Editors might request from authors as much information as they consider necessary to confirm the fulfilment of such premises. Failure to comply with these bioethical principles will suppose immediate rejection of the article, regardless of the reviewers’ recommendation. Las normas de publicación en castellano están disponibles para su consulta en la página web de Basic and Applied Herpetology (http://bah.herpetologica.es/)

BASIC & APPLIED HERPETOLOGy REVISTA ESPAÑOLA DE HERPETOLOGÍA

On behalf of the Spanish Herpetological Society, the editorial board of Basic and Applied Herpetology wants to acknowledge the work of the following experts who have worked as manuscript reviewers for the elaboration of the present volume (in alphabetical order): Josabel Belliure (Universidad de Alcalá, Spain) Charles R. Bursey (Pennsylvania State University, USA) Thibaut Couturier (CEFE-CNRS, France) Soumia Fahd (Université Abdelmalek Essaddi, Morocco) Carles Feliu (Universitat de Barcelona, Spain) Núria Garriga (Universitat de Barcelona, Spain) Eva Graciá (Universidad Miguel Hernández, Spain) Olivier Guillaume (Station d’Ecologie Expérimentale-CNRS, France) Fernado Martínez-Freiría (CIBIO-Universidade de Porto, Portugal) íñigo Martínez-Solano (IREC-CSIC-UCLM-JCCM, Spain) Albert Montori (Universitat de Barcelona, Spain) Dennis Rödder (zoologisches Forschungmuseum Alexander koenig, Germany) Paulo Sá-Sousa (Universidade de évora, Portugal) Jarmo Saarikivi (University of Helsinki, Finland) Gregorio Sánchez-Montes (Universidad de Navarra, Spain) Thomas W. Schoener (University of California Davis, USA) Neftalí Sillero (Universidade de Porto, Portugal)

AHE

B&AH nº 26 [2012]

Articles inside

B&AH nº 26 [2012]