ABC 41.2 (2018) by Museu Ciències Naturals de Barcelona MCNB

en J. Hatchwell, Univ. of Sheffield, UK

Dibuix de la coberta / Ddibujo de la portada / Drawing of the cover: Phyllopteryx taeniolatus, cavallet de mar d’algues, dragón marino común, weedy seadragon (Jordi Domènech)

Secretaria de Redacció / Secretaría de Redacción / Editorial Office Museu de Ciències Naturals de Barcelona Passeig Picasso s/n. 08003 Barcelona, Spain Tel. +34–93–3196912 Fax +34–93–3104999 E–mail abc@bcn.cat Secretària de Redacció / Secretaria de Redacción / Managing Editor Montserrat Ferrer Assistència Tècnica / Asistencia Técnica / Technical Assistance Eulàlia Garcia Anna Omedes Francesc Uribe Assessorament lingüístic / Asesoramiento lingüístico / Linguistic advisers Carolyn Newey Pilar Nuñez

Animal Biodiversity and Conservation 41.2, 2018 © 2018 Museu de Ciències Naturals de Barcelona, Consorci format per l'Ajuntament de Barcelona i la Generalitat de Catalunya Autoedició: Montserrat Ferrer Fotomecànica i impressió: CEVAGRAF SCCL ISSN: 1578–665 X eISSN: 2014–928 X Dipòsit legal: B. 5357–2013

Animal Biodiversity and Conservation es publica amb el suport de / Animal Biodiversity and Conservation se publica con el apoyo de / Animal Biodiversity and Conservation is published with the support of: Asociación Española de Ecología Terrestre – AEET Sociedad Española de Etología y Ecología Evolutiva – SEEEE Sociedad Española de Biología Evolutiva – SESBE Disponible gratuitament a internet / Disponible gratuitamente en internet / Freely available online at: www.abc.museucienciesjournals.cat

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Editor en cap / Editor responsable / Editor in Chief Joan Carles Senar Museu de Ciències Naturals de Barcelona, Barcelona, Spain Editors temàtics / Editores temáticos / Thematic Editors Ecologia / Ecología / Ecology: Mario Díaz (Asociación Española de Ecología Terrestre – AEET) Comportament / Comportamiento / Behaviour: Adolfo Cordero (Sociedad Española de Etología y Ecología Evolutiva – SEEEE) Biologia Evolutiva / Biología Evolutiva / Evolutionary Biology: Santiago Merino (Sociedad Española de Biología Evolutiva – SESBE) Editors / Editores / Editors Pere Abelló Institut de Ciències del Mar CMIMA–CSIC, Barcelona, Spain Javier Alba–Tercedor Universidad de Granada, Granada, Spain Russell Alpizar–Jara University of Évora, Évora, Portugal Marco Apollonio Università degli Studi di Sassari, Sassari, Italy Miquel Arnedo Universitat de Barcelona, Barcelona, Spain Xavier Bellés Institut de Biología Evolutiva UPF–CSIC, Barcelona, Spain Salvador Carranza Institut de Biologia Evolutiva UPF–CSIC, Barcelona, Spain Luís Mª Carrascal Museo Nacional de Ciencias Naturales–CSIC, Madrid, Spain Pablo Castillo, Institute for Sustainable Agriculture–CSIC, Córdoba, Spain Adolfo Cordero Universidad de Vigo, Vigo, Spain Mario Díaz Museo Nacional de Ciencias Naturales–CSIC, Madrid, Spain Darío Díaz Cosín Univ. Complutense de Madrid, Madrid, Spain José A. Donazar Estación Biológica de Doñana–CSIC, Sevilla, Spain Arnaud Faille Museum National histoire naturelle, Paris, France Jordi Figuerola Estación Biológica de Doñana–CSIC, Sevilla, Spain Gonzalo Giribet Museum of Comparative Zoology, Harvard Univ., Cambridge, USA Susana González Universidad de la República–UdelaR, Montivideo, Uruguay Sidney F. Gouveia Universidad Federal de Sergipe, Sergipe, Brasil Gary D. Grossman University of Georgia, Athens, USA Ben J. Hatchwell University of Sheffield, Sheffield, UK Joaquín Hortal Museo Nacional de Ciencias Naturales-CSIC, Madrid, Spain Jacob Höglund Uppsala University, Uppsala, Sweden Damià Jaume IMEDEA–CSIC, Universitat de les Illes Balears, Esporles, Spain Miguel A. Jiménez–Clavero Centro de Investigación en Sanidad Animal–INIA, Madrid, Spain Jennifer A. Leonard Estación Biológica de Doñana-CSIC, Sevilla, Spain Jordi Lleonart Institut de Ciències del Mar CMIMA–CSIC, Barcelona, Spain Josep Lloret Universitat de Girona, Girona, Spain Jorge M. Lobo Museo Nacional de Ciencias Naturales–CSIC, Madrid, Spain Pablo J. López–González Universidad de Sevilla, Sevilla, Spain Jose Martin Museo Nacional de Ciencias Naturales–CSIC, Madrid, Spain Santiago Merino Museo Nacional de Ciencias Naturales–CSIC, Madrid, Spain Juan J. Negro Estación Biológica de Doñana–CSIC, Sevilla, Spain Vicente M. Ortuño Universidad de Alcalá de Henares, Alcalá de Henares, Spain Miquel Palmer IMEDEA–CSIC, Universitaat de les Illes Balears, Esporles, Spain Per Jakob Palsbøll University of Groningen, Groningen, The Netherlands Reyes Peña Universidad de Jaén, Jaén, Spain Javier Perez–Barberia Estación Biológica de Doñana–CSIC, Sevilla, Spain Juan M. Pleguezuelos Universidad de Granada, Granada, Spain Oscar Ramírez Institut de Biologia Evolutiva UPF–CSIC, Barcelona, Spain Montserrat Ramón Institut de Ciències del Mar CMIMA–CSIC, Barcelona, Spain Ignacio Ribera Institut de Biología Evolutiva UPF–CSIC, Barcelona, Spain Diego San Mauro Universidad Complutense de Madrid, Madrid, Spain Ramón C. Soriguer Estación Biológica de Doñana–CSIC, Sevilla, Spain Constantí Stefanescu Museu de Ciències Naturals de Granollers, Granollers, Spain Diederik Strubbe University of Antwerp, Antwerp, Belgium Miguel Tejedo Madueño Estación Biológica de Doñana–CSIC, Sevilla, Spain José L. Tellería Universidad Complutense de Madrid, Madrid, Spain Simone Tenan MUSE–Museo delle Scienze, Trento, Italy Francesc Uribe Museu de Ciències Naturals de Barcelona, Barcelona, Spain José Ramón Verdú CIBIO, Universidad de Alicante, Alicante, Spain Carles Vilà Estación Biológica de Doñana–CSIC, Sevilla, Spain Rafael Villafuerte Inst.ituto de Estudios Sociales Avanzados (IESA–CSIC), Cordoba, Spain Rafael Zardoya Museo Nacional de Ciencias Naturales–CSIC, Madrid, Spain

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Understanding nutrient landscapes for giant pandas in the Qinling Mountains, China: the relationships between bamboo mineral content and giant panda habitat selection during migration Q. Huang, X. Liu, Y. Li, J. Kraus, M. Songer Huang, Q., Liu, X., Li, Y., Kraus, J., Songer, M., 2018. Understanding nutrient landscapes for giant pandas in the Qinling Mountains, China: the relationships between bamboo mineral content and giant panda habitat selection during migration. Animal Biodiversity and Conservation, 41.2: 195–208. Abstract Understanding nutrient landscapes for giant pandas in the Qinling Mountains, China: the relationships between bamboo mineral content and giant panda habitat selection during migration. Bamboo comprises over 99 % of the diet of giant pandas (Ailuropoda melanoleuca). Giant pandas face a complex nutrient landscape. They eat more than one species of bamboo and various parts of the plant, and they move seasonally to find optimal forage. Though the seasonal habitat preferences of giant pandas have long been known, the spatial and temporal nutrient gradient of bamboo between seasonal habitats remains unclear. Few studies detail the nutrient content of bamboo in relation to the seasonal habitat selection of giant pandas in the wild. In this study, we collected bamboo samples from 57 plots considering four factors (seasons, elevations, species, and plant parts). We evaluated the effect of these factors on the contents of seven bamboo mineral elements (Cu, Zn, Fe, Mn, K, Ca, and Mg) and used a non–parametric ensemble tree model to model giant pandas' presence and absence based on bamboo mineral content. Our results showed strong correlations between pairs of mineral contents (up to r = 0.69) with specific mineral elements such as Mn, consistently showing great importance in the models for differentiating the habitat selection. We also observed significant variation in mineral concentrations between seasons, bamboo species, and plant parts. Our results suggest that the studied bamboo mineral content strongly associates giant pandas' habitat preferences. Our research may be useful for the development of conservation and reserve management strategies by providing guidelines to increase giant pandas' opportunities to obtain sufficient nutrient within the Qinling region. Key words: Giant pandas, Habitat selection, Machine learning, Mineral elements, Nutrient content, Qinling Mountains Resumen Comprender la distribución de nutrientes en el territorio del panda gigante en las montañas Qinling, en China: las relaciones entre el contenido de minerales del bambú y la selección del hábitat del panda gigante durante la migración. A pesar de que el bambú constituye más del 99 % de la dieta del panda gigante (Ailuropoda melanoleuca), esta especie se enfrentan a un complejo patrón de disponibilidad de nutrientes, ya que consumen más de una especie de bambú y varias partes de la planta, y se desplazan de forma estacional para encontrar el alimento óptimo. Si bien las preferencias estacionales del hábitat del panda gigante se conocen desde hace tiempo, el gradiente espacial y temporal de los nutrientes del bambú entre los hábitats estacionales sigue siendo poco claro. Son pocos los estudios en los que se describe con detalle el contenido de nutrientes del bambú en relación con la selección estacional del hábitat del panda gigante en el medio silvestre. En este estudio, recogimos muestras de bambú de 57 parcelas teniendo en cuenta cuatro factores (estación del año, altura, especie y parte de la planta). Se evaluó el efecto de estos factores en el contenido de siete elementos minerales del bambú (Cu, Zn, Fe, Mn, K, Ca y Mg) y se utilizó un modelo de árbol no paramétrico para determinar la presencia y ausencia del panda gigante en función del contenido de minerales del bambú. Nuestros resultados mostraron que existen fuertes correlaciones entre pares de contenidos de minerales (hasta r = 0,69) y que algunos elementos minerales, como el Mn, son siempre un factor importante en la selección del hábitat. También observamos una variación significativa en las concentraciones de minerales entre estaciones, especies de bambú y partes de la planta. Los resultados sugieren que el contenido de minerales del bambú estudiado está fuertemente relacionado con ISSN: 1578–665 X eISSN: 2014–928 X

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las preferencias de hábitat de los pandas gigantes. Nuestra investigación puede ayudar a elaborar estrategias de conservación y gestión de reservas al ofrecer pautas que ayuden a aumentar la posibilidad de que el panda gigante pueda obtener los nutrientes que necesita en la región de Qinling. Palabras clave: Pandas gigantes, Selección del hábitat, Aprendizaje automático, Elementos minerales, Contenido de nutrientes, Montañas Qinling Received: 28 II 17; Conditional acceptance: 17 V 17; Final acceptance: 12 IX 17 Qiongyu Huang, Jacob Kraus, Melissa Songer, Conservation Ecology Center, Smithsonian Conservation Biology Institute, Front Royal, Virginia 22630, USA.– Xuehua Liu, Yajun Li, School of Environment, Tsinghua Univ., Beijing 100084, China. Corresponding author: Xuehua Liu. E–mail: xuehua–hjx@tsinghua.edu.cn

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Introduction Bamboo comprises over 99 % of the diet of wild giant pandas (Ailuroplda melanoleuca) (Sheldon, 1937; Milton, 1979; Dierenfeld et al., 1982; Tuanmu et al., 2012). Though giant pandas forage primarily on vegetation, their gut anatomy resembles that of a carnivore, with a simple stomach and short gastrointestinal tract that greatly limits its ability to digest fiber. The poor nutrition content of bamboo, which is low in protein and high in fiber, requires giant pandas to consume large quantities of bamboo to meet their dietary requirements (Milton, 1979; Dierenfeld et al., 1982; Tuanmu et al., 2012). Climate change and shifts in the phenology and physiological conditions of bamboo may exacerbate the challenges of finding suitable habitat to meet giant pandas dietary requirements (Liu et al., 1999; Li and Manfred, 2001; Hunter et al., 2003; Songer et al., 2012; Hull et al., 2014). Given the dependence of giant pandas on bamboo, a deeper understanding of the nutrient landscape faced by giant pandas is critical to conservation efforts (Mainka et al., 1989; Reid et al., 1989; Reid and Hu, 1991; Taylor et al., 1991). Previous studies highlighted a variety of factors (seasons, elevations, species, and bamboo plant parts) affecting the mineral content of bamboo (Fu et al., 1990; Liu, 2008; Sun et al., 2010; Wang et al., 2010, 2013). Most studies, however, focused on the effect of only one or two of these factors and presented only the empirical measurements of the chemical composition of bamboo (Taylor and Qin, 1993; Li et al., 2007; Wang et al., 2009; Wu et al., 2009). A comprehensive analysis of the relationship between multiple covariables and bamboo nutrients is needed (Liu, 2001; Liu et al., 2002, 2005; Finley et al., 2011; Hull et al., 2011). Giant panda’s habitat selection in the Qinling Mountains has been extensively documented. Giant pandas migrate in late spring from their home range at low elevations to their summer home range at higher elevations and return to their low elevation range in autumn (Schaller et al., 1989; Liu et al., 2002, Lu et al., 2007; Qi et al., 2011, 2012). It has been hypothesized that this migration pattern evolved to facilitate access to either abundant or more nutritious forage (Zhang et al., 2006, 2014; Sims et al., 2007; Viña et al., 2010; Wang et al., 2010). However, little work has been done to understand the relationships among the nutrient composition of bamboo and giant pandas habitat preferences. In our study, we measured the nutrient content of bamboo across seasons and elevation gradients. We collected samples from three elevation strata in spring, summer and autumn. Each season, giant pandas are present in one of the strata and absent in the other two. This enabled us to examine the nutrient landscape of giant pandas at various locations and seasons and to explore the statistical association between bamboo nutrient content and giant pandas’ habitat choices. Furthermore, it allowed us to compare the relative contribution of individual mineral elements to giant pandas' habitat selection. We first tested the correlation between seven mineral elements. We then examined the effect of the four factors (season, elevation, species, and plant

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part) on the mean concentrations of each mineral. Finally, we modeled the presence and absence of giant pandas using the mineral element contents with a non–parametric ensemble tree model. For each season we built one model and evaluated the importance of predictive variables to the model. We sought to answer the following research questions: (1) How do each of the seven mineral elements change with season, elevation, part, and species? (2) What are the relative associations between the seven mineral elements across landscapes and giant pandas’ seasonal selection of habitat? Material and methods Foping Nature Reserve (FNR) in Shaanxi Province was founded in 1978 in the center of a giant panda reserve network in the Qinling Mountain range. FNR (32º 32' – 33º 43' N, 107º 41' – 107º 56' E) covers an area of 293 km2 and is one of earliest reserves established for giant pandas. The annual temperature at FNR was 11.5 ºC (mean min = –3 ºC in January, mean max = 28 ºC in July) with elevation ranging between 1000 m and 2900 m. FNR is home to three bamboo species: arrow bamboo (Fargesia qinlingensis) dominating at higher elevation (1,700–2,900 m), wooden bamboo (Bashania fargesii) mostly found at lower elevation (1,000–1,900 m), and dragon–head bamboo (Fargesia racocephala) restricted to a small area in the southeast corner of the FNR (1,000– 1,800 m). Arrow bamboo is the giant pandas’ only food resource in summer, while they forage primarily on wooden bamboo and dragon–head bamboo in other seasons (State Forestry Administration of the People’s Republic of China, 2006; Zhang et al., 2014). Guanyinshan Nature Reserve (GNR) was founded in 2002 to protect additional giant panda habitat. GNR (33º 35' – 33º 45' N, 107º 51' – 108º 01' E) covers an area of 135 km2, and is located on the south slope of the Qinling Mountainsdirectly adjacent to the eastern boundary of FNR. The average annual temperature at GNR is 11.5 ºC with elevation ranging between 1,150 to 2,574 m. Arrow bamboo (located > 1,800 m) and dragon–head bamboo (located below 1,800 m and close to water) are common in the reserve. Because the GNR area was previously part of a state–owned forest enterprise, logging activities were a common source of disturbance. After the establishment of the Natural Forest Conservation Program in 1998, logging activities were banned (Liu et al., 2008). Giant pandas were recently spotted several times in dragon–head bamboo forest during spring and autumn in this reserve (Liu et al., 2013). Sample processing We collected bamboo samples from 57 plots across the FNR and GNR in April, July, and October 2013 (fig. 1). We chose these sampling periods because they are in the middle of spring (March to May), summer (June to August), and autumn (September to November). We divided the elevation into three strata: low elevation

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stratum (1,500–1,750 m) with wooden and dragon–head bamboo, transition elevation stratum (1,750–2,200 m) with arrow and wooden bamboo, and high elevation stratum (2,200–2,700 m) with only arrow bamboo. We took wooden bamboo, arrow bamboo and dragon–head bamboo samples in Sanguanmiao, Guangtoushan, and Xigou where there were large forests of the three bamboo species. Each season we sampled one transect (arrow bamboo) at high elevation, two transects (arrow and wooden bamboo) at the transitional elevation, and two transects (wooden and dragon–head bamboo) at the low elevation. Each transect contained three or four plots. In each plot (1 x 1 m) we took samples (200–300 g, wet) of shoots, leaves, branches, one–year culms, and culms older than one year. Each sample was washed with distilled water, ground to 80 g mesh by a high–speed grinder, and dried to a constant weight at 60 ºC. We analyzed seven mineral elements (Cu, Zn, Fe, Mn, K, Ca, and Mg) by the standard atomic absorption spectrophotometry procedure (Skoog and West, 1980). Data analysis We tested the correlation between the seven elements using Pearson correlation and the mean content difference between levels of each factor (seasons, elevations, parts, and species) by non–parametric Kruskal–Wallis test (significance level at α = 0.05). During spring and winter, giant pandas are present only at low elevations and absent at transitional and high elevations. In late spring, giant pandas leave low elevation areas and migrate rapidly through the transitional region to reach high elevation areas where they stay until the end of summer (Liu, 2001; Nie et al., 2014). To examine the relationships between mineral elements and spatial distribution of giant pandas, we summarized 35 variables based on the mean nutrient concentration in all bamboo parts (5 parts x 7 elements). We used a Random Forest (RF) approach to model the presence and absence of giant pandas using these 35 predictive variables. We implemented the Random Forest package (Liaw and Wiener, 2002) in R software (R Development, 2008) to build the models. RF is a non–parametric ensemble tree model for classification that operates by constructing a multitude of decision trees. RF can accommodate large number of input variables without overfitting and it does not require a specific distribution for predictive variables (Breiman, 2001; Svetnik et al., 2003; Prasad et al., 2006). The RF model is particularly well suited for modeling environmental co–variables which are often correlated (Breiman, 2001; Svetnik et al., 2003; Prasad et al., 2006). The algorithm of RF starts with the selection of bootstrapped samples from the original data. There are approximately 63 % of the data in each bootstrap sample. Each decision tree grows with one bootstrap sample and a randomly selected subset of variables. Observations outside the bootstrap sample, called 'out–of–bag' observations, are used as testing data to examine the prediction error. One of the advantages of the RF model is that it provides variable importance by measuring how much model performance (specifically measured by the Gini index) declines if the variable is randomly permuted. The greater the increase

of prediction error, the greater the variable importance in the model. The final prediction is obtained by aggregating over the ensemble trees (Breiman, 2001; Cutler et al., 2007; Biau, 2012). For a small number of plots, we did not collect one– year culms or shoots because the two parts were not observed year–round. In order for RF to incorporate variables with missing data, we used multiple imputation by chained equation (MICE) (Buuren and Groothuis–Oudshoorn, 2011) to create plausible values for the missing data (Van Buuren and Oudshoorn, 1999). Compared with other methods, such as single imputation, MICE allows for the imputation of multivariate data based on the distribution of observed data, without the need to specify a joint distribution of predictor variables (White et al., 2011). It can also handle different variable types since each variable is imputed using its own imputation model, with good quality prediction and less biased estimates (Ambler et al., 2007). Normally, for a given missing data point, a multiple imputation method generates 3–5 imputations and missing values are replaced by the average of the multiple imputed values (Hui et al., 2004). In our research, we generated the imputed values using the MICE package in R software. To take into account the varying availability of bamboo shoots over different seasons, we used the realized nutrient concentration of each plant part, that is, the product of measured or imputed nutrient values and availability of the specific part over the season for the predictive variables in the RF model. For species in which bamboo shoots are only available for a short period (spring for wooden bamboo and summer for arrow and dragon–head bamboo), we set the availability as 1 for the seasons when shoots were present and 0 for the other seasons. For the other species, the availability of leaves, branches, culm and one–year culm were set to 1 for all seasons. We next compared the mean difference of each mineral value between giant panda presence and absence for each predictor using the Wilcoxon rank sum test (significance level at α = 0.05). Finally, we built RF models for spring (SPR), summer (SUM), and autumn (AUT) to classify the presence and absence of giant pandas based on the nutrient concentrations determined for each element. To determine how the predictors associated with a suitable habitat for giant pandas in all three seasons, we built a combined model (COM), including all mineral data. Results Correlation of seven mineral elements Most elements had a correlation value under 0.4 (table 1). Ca and Mg showed the strongest correlation (r = 0.69), followed by Cu:Zn (r = 0.64), Mn:Mg (r = 0.46), and Mn:Ca (r = 0.43) (fig. 1s) and lowest correlation was between Cu and Mg (r = 0.02). Mean mineral difference test We tested the mean difference between different levels for each factor (season, elevation, plant part, and

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Elevation (m) High 2,883 Low 1,076 Foping Nature Reserve Guanyinshan Nature Reserve Bamboo samples River

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Fig. 1. Schematic map of the study area and sampling points. Fig. 1. Mapa esquemático de la zona de estudio y los puntos de muestreo.

species) (see supplementary material). No significant differences were found between three elevation ranges for any mineral element (all p ˃ 0.05). However, concentrations between the three seasons differed (all p ≤ 0.05) for all mineral elements except K. The highest concentrations of Cu and Zn were in spring, followed by those in autumn and summer (fig. 2, table 1s), while Fe had a higher value in summer than spring and autumn (table 1s). For Mn, Ca, and Mg, samples collected in autumn were at a highest concentration followed by those in spring and summer. Within species, only Mn significantly differed between the three bamboo species (p = 0.028), with wooden bamboo showing the highest level, followed by dragon–head and arrow bamboo (fig. 2). Finally, all mineral elements demonstrated a significant difference between the five plant parts (all p ≤ 0.05). The highest concentration of K was found in shoots; Mn, Ca and Mg were in highest concentration in leaves; Cu, Zn and Fe were at their highest levels in branches (table 1s). In SPR, Zn found in branches, Ca found in shoots, and Mn found in one–year culms and in branches were all in higher concentration in habitats where giant pandas were present than in habitats where they were absent (fig. 3). In SUM, the Fe, Ca and Mg content of shoots was higher in habitats where giant pandas were present. However, Cu, K, Ca, and Mg concentrations found in leaves were significantly higher in habitats where giant pandas were absent (fig. 3). In AUT, Mn concentration found in leaves and one–year culms was higher in habitats where giant pandas were present, while Mg concentration found in branches was higher in habitats where giant pandas were absent (fig. 3) (all p ≤ 0.05). Classification models Among the four models, the lowest error rate occurred in SUM (10.5 %), while the highest error rate was 31.6 % in AUT. COM and SPR had 20.4 % and 18.6 % error rates, respectively. In COM, Mn found in

one–year culms had the largest decrease of the Gini index, followed by Mn in culms, Ca in shoots, and Mn in leaves (fig. 4). In spring, Mn found in one–year culms showed the highest variable importance, followed by Ca in shoots, and Ca and Mn in branches. However, Mg and Ca found in shoots and K and Mg found in leaves were the most important predictors in summer. Mn and K found in one–year culms, Mg found in branches, and Mn found in culms had the highest importance measure in AUT (fig. 4). Discussion Our results show strong associations between giant pandas’ habitat selection and bamboo mineral content. The results show contrasting dietary consequences as giant pandas migrate to different habitats at different times of the year (fig. 3). Our results suggest that the nutrient composition in the giant pandas' forage

Table 1. Correlation of seven bamboo mineral elements. Tabla 1. Correlación entre siete elementos minerales del bambú.

Cu Zn Fe Mn K Ca

Zn 0.64 Fe 0.36 0.19 Mn 0.26 0.34 0.11 K 0.19 0.07 0.13 0.11 Ca 0.36 0.36 0.06 0.43 0.29

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20 0

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SUM Season

AUT

1,000 Fe

1,000

SPR

Arrow

500 0

Dragon–head Wooden Species

Branch Culm Leaf 1–Cl Shoot Species

16,000 7,500

12,000

5,000

8,000

2,500

4,000 Branch Culm Leaf 1–Cl Parts

Shoot

Branch Culm Leaf 1–Cl Shoot Parts

Fig. 2. The effect of different factors on mineral content (mg/kg). The box plots show the distribution of six mineral content values differentiated by levels of three factors (seasons, bamboo species, and bamboo parts). We only show here the elements with the most significant differences between levels: 1–Cl, one year culm. Fig. 2. Efecto de diferentes factores en el contenido de minerales (mg/kg). Los diagramas de caja muestran la distribución de los valores del contenido de seis minerales según la categoría de tres factores (estación, especie de bambú y partes del bambú). Solo mostramos aquí los elementos con las diferencias más significativas entre categorías: 1–Cl, cañas de un año.

bamboo is an important co–variant as they migrate seasonally, and such nutrient differences can be used to differentiate giant pandas' seasonal habitat selection. Our study supports the hypothesis that the seasonal migration pattern might have evolved to facilitate access to specific composition of nutrients from forage. One of the important hypotheses previously proposed to explain giant pandas' movement is that shoots sprouting in the Qinling Mountains in late spring and early summer drives their seasonal movements from low to high elevation (Pan et al., 1988). Evidently, dragon–head and arrow bamboos produce shoots only during summer and wooden bamboo shoots are available only in spring (Nie et al., 2014). Additionally, for almost all bamboo species, due to temperature

differences, the shoots sprout first at lower elevation and shift to higher elevation sequentially. This coincides closely with giant pandas' migrations to higher elevations. However, this hypothesis is mostly based on observed shoot availability and it is unclear how mineral content in bamboo shoots associates with such patterns. By incorporating the availability of shoots with the measured mineral content, our study allowed us to quantitatively test for associations between panda habitat choice and various minerals in shoots. Based on our classification model, we found that the concentrations of Fe, Ca, and Mg in shoots effectively predict differences in presence and absence at habitats during the summer. The highest concentrations of K, Ca, and Mg in summer were found in bamboo

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Fig. 3. The contrast of nutrient concentration (mg/kg) between the seasonal habitat where giant pandas are present and the habitat where they are absent. The concentration of nine nutrients in spring (top row), summer (middle row), and autumn (bottom row) and from different bamboo parts are compared. Fig. 3. Diferencia de concentración de nutrientes (mg/kg) entre el hábitat estacional en el que el panda gigante está presente y el hábitat en el que está ausente. Se compara la concentración de nueve nutrientes en primavera (fila superior), verano (fila media) y otoño (fila inferior) y en diferentes partes del bambú.

leaves in areas where giant pandas were absent (fig. 3). It is possible that shoots are preferred because they have the lowest concentration of fiber and cost less energy to obtain than other parts of the plant. In our results, Mn was the most influential predictor of giant pandas' presence not only in COM but also in SPR and AUT. Lack of Mn can lead to many physiological problems in animals, e.g., decrease of sperm, fecundity decline, premature birth and lactation cessation (Kemmerer et al., 1931; Plumlee et al., 1956; Hurley and Doane, 1989). The physiological function of Mn could contribute to its importance, especially when giant pandas are in estrus in spring and pregnant in summer (Schaller et al., 1989; Zhu et al., 2001). Similarly, Ca is mainly stored in bones, and

lack of Ca could cause arrested development, bone deformities, and miscarriages or stillbirths (Hightshoe et al., 1991; Bhanderi et al., 2014). Animals will also experience face muscle twitching or convulsions if they cannot get enough Mg (Kaneko et al., 2008; Chandra et al., 2013). The physiological function of Ca and Mg might have contributed to the observations that Ca and Mg were more concentrated in the presence areas in summer and they were ranked high in our models. Ultimately, if giant panda migration is evolutionarily driven by nutrient gradients, our models suggest that Ca, Mg, and Mn play an important role in the process. Further research is required to investigate the specific biochemical interactions of the mineral elements in panda physiology.

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COM SPR Mn.one year culm Mn.culm Ca.shoot Mn.leaf Zn.branch Fe.culm Mg.shoot Mn.branch Cu.one year culm Ca.culm

Mn.one year culm Ca.shoot Ca.branch Mn.branch Mg.one year culm Zn.branch Fe.culm Fe.leaf K.leaf Mn.culm 1.0

1.5

2.0

2.5

0.2

0.4

0.6

0.8

SUM AUT Mg.shoot Mg.one year culm Ca.shoot Mg.branch K.leaf Mn.culm Mg.leaf K.one year culm K.one year culm Mn.leaf Fe.shoot Ca.branch Ca.leaf Mg.culm K.culm Cu.culm K.shoot Cu.one year culm Cu.leaf Ca.leaf 0.2 0.3 0.4 0.5 0.6 0.7 0.5 1.0 1.5 Mean decrease of Gini Mean decrease of Gini Fig. 4. Relative importance of variables in four Random Forest models: COM, combined model; SPR, spring model; SUM, summer model; and AUT, autumn model. Variable importance plots show the mean decrease of Gini index of the top 10 influential predictors in four RF models: the black symbols indicate a significant difference between habitats where giant pandas were present and absent and the grey symbols indicate no significant difference; triangles indicate the mean of the presence habitat is higher than that of the absence habitat and the opposite is represented by circles. Fig. 4. Importancia relativa de las variables en cuatro modelos de bosques aleatorios: COM, modelo combinado; SPR, modelo primavera, SUM, modelo verano; AUT, modelo otoño. En los gráficos de la importancia variable se muestra la reducción media del índice de Gini de los 10 factores de predicción más influyentes en cuatro modelos de bosques aleatorios: los símbolos negros indican que hay una diferencia significativa entre los hábitats donde el panda gigante estaba presente y ausente, y los símbolos grises indican que no hay diferencia significativa entre ambos; los triángulos indican que el promedio del hábitat donde estaba presente es superior al del hábitat donde no lo estaba; lo contrario se representa con círculos.

Our comparison of bamboo samples between presence and absence habitats provides a new approach to identifying potentially important dietary factors underlying panda movements. Using this mean comparison method, we found that predictors such as K and Mg found in leaves in SUM and Mg found in branches in AUT with high variable importance were significantly lower in the elevations where giant pandas were present. We speculate that despite the low concentration of these mineral elements in the presence habitat of giant pandas, the level may still meet their requirements for that mineral, and giant pandas may prefer to acquire them from other parts of bamboo. On the other hand, other predictors such as Ca found in branches in SPR, K found in one–year culms in SUM, and K found in one–year culms in AUT did not show significant differences between presence and absence habitats, but they ranked high in their im-

portance for classifying the presence of giant pandas. This result suggests small but consistent differences in content between presence and absence habitats. It could also indicate that non–linear relationships exist between these habitats (fig 4). Further studies are needed to examine those relationships and how they affect giant panda activities (Swaisgood et al., 2009). The RF model can perform well using predictors with correlations and still provide an accurate variable importance ranking (Breiman, 2001; Biau, 2012; Freeman et al., 2012). Normally in linear model, the proper way to deal with correlated variables having redundant information is to remove one or multiple correlated variables. In our case, all elements are included because they have specific physiological functions. For example, Ca and Mg had the highest correlation and they demonstrated similar importance rankings in each model. Our model highlights the advantage

Animal Biodiversity and Conservation 41.2 (2018)

of a non–parametric RF model which often deals with a large number of correlated nutrient variables. It is worth noting that this study examined the total concentration of key nutrients.It does not consider the fiber–bonded mineral element. As giant pandas’ digestive system has lower efficiency absorbing fiber–bonded minerals than non–fiber–bonded minerals (Dierenfeld et al., 1982), the portion of fiber– bonded mineral element in bamboo is an important factor that affect the amount of nutrient utilized by giant panda (according to analysis done on leaves of Phyllostachys aureosulcata, 25 % of the total mineral are fiber bound and not available to giant pandas (unpublished data and personal communication with Dr. Michael Power). However, measurements of total mineral content are a necessary first step to understanding the patterns of giant panda habitat selection with the nutrient landscape. Knowing what minerals are concentrated where within the landscape and how the concentration of these elements changes both spatially and temporally improves our understanding of potential drivers of giant panda migration and habitat preferences. This knowledge also provides insights for improving conservation planning and management. With the increasing giant panda density in FNR there has been an increasing number of observations of giant pandas dispersing from west to east across the reserve boundary into GNR to colonize unoccupied habitat (Hu et al., 2010; Liu et al., 2013). In order to create a suitable environment, local governmental and conservation organizations have been planting bamboo in GNR (World Wide Fund for Nature, 2006). Our results will help to provide guidelines for determining the bamboo species and the locations for the planting practices. According to our results, wooden bamboo is recommended for giant panda habitat restoration in low elevations because of its highest concentration of nutrients. GNR does not have any large distribution of wooden bamboo due to its logging history. We recommend planting wooden bamboo in GNR, especially in low elevational areas, because it should support giant panda dispersal from FNR in the west to GNR in the east. To achieve better results, bamboo should be planted every few years to minimize the impact of periodic flowering on the panda population (Kawamura, 1927; Chai et al., 2006). In 2011 and 2013 our team started planting wooden bamboo in GNR. However, bamboo contains the lowest nutrient levels in summer, and only arrow bamboo is distributed in high elevation. Ensuring adequate arrow bamboo in high elevational areas is critical to giant panda conservation because an abundant food supply can compensate for lower nutrient content. Since the establishment of the reserves, economic activities have been banned within them. However, some wild giant pandas still live in the area outside the protected areas where they have to compete with local people harvesting bamboo shoots. We strongly recommend that future conservation policies are enacted to develop a sustainable bamboo shoot harvesting plan to ensure a sufficient supply of shoots for giant pandas both inside and outside reserves.

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Acknowledgements This research received support from the National Natural Science Foundation 'Influences of Forest Landscape Pattern on Animal Diversity and Behavior Characteristics under Disturbance in the Qinling Mountains' (41271194) and David M. Rubenstein. We are grateful to all those involved in fieldwork, Pengfeng Wu, Xiaodong Jia and Hongxing Li. We also want to thank Dr. Michael Power for giving us valuable feedback on the manuscript, as well as providing data on bamboo fiber content. References Ambler, G., Omar, R. Z., Royston, P., 2007. A comparison of imputation techniques for handling missing predictor values in a risk model with a binary outcome. Statistical Methods in Medical Research, 16(3): 277–298. Bhanderi, B. M., Garg, M. R., Sherasia, P. L., 2014. Mineral status of feeds, fodder and dairy animals in Jalgaon district of Maharashtra state. Scholars Journal of Agriculture and Veterinary Sciences, 1(4A): 222–226. Biau, G., 2012. Analysis of a random forests model. The Journal of Machine Learning Research, 98888(1): 1063–1095. Breiman, L., 2001. Random forests. Machine learning, 45(1): 5–32. Buuren, S., Groothuis–Oudshoorn, K., 2011. MICE: Multivariate imputation by chained equations in R. Journal of Statistical Software, 45(3): 1–67. Chai, Z., Qin, Y., Hua, X., Wang, Z., Wang, Q., 2006. Advance of studies on bamboo flowering causes. Journal of Zhejiang Forestry Science and Technology, 26: 53 [In Chinese]. Chandra, A. K., Sengupta, P., Goswami, H., Sarkar, M., 2013. Effects of dietary magnesium on testicular histology, steroidogenesis, spermatogenesis and oxidative stress markers in adult rats. Indian Journal of Exprimental Biology, 51: 37–47. Cutler, D. R., Edwards, T. C., Beard, K. H., Cutler, A., Hess, K. T., Gibson, J., Lawler, J. J., 2007. Random forest for classification in ecology. Ecology, 88(11): 2783–2792. Dierenfeld, E. S., Hintz, H. F., Robertson, J. B., 1982. Utilization of bamboo by the giant panda. J. Nutr., 112: 636–641. Finley, T. G., Sikes, R. S., Parsons, J. L., Rude, B. J., Bissell, H. A., Ouellette, J. R., 2011. Energy digestibility of giant pandas on bamboo–only and on supplemented diets. Zoo Biology, 30(2): 121–133. Freeman, E. A., Moisen, G. G., Frescino, T. S., 2012. Evaluating effectiveness of down–sampling for stratified designs and unbalanced prevalence in Random Forest models of tree species distributions in Nevada. Ecological Modelling, 233: 1–10. Fu, Q., Wen, Y., Wang, A., Huang, J., Peng, H., Wang, J., 1990. Study on the trace element content in bamboo in Wolong Natural Reserve. Journal of Bamboo Research, 4: 9. Hightshoe, R. B., Cochran, R. C., Corah, L. R., Ki-

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Supplementary material

Table 1s. Mean concentration (mg/kg) and p–value of each factor for seven mineral elements: * p– value < 0.05. (Season: SPRm, spring mean; SUMm, summer mean; AUTm, autumn mean. Elevations: H, high; L, low; T, transition. Species: Wb, wooden bamboo; Dhb, dragon head bamboo; Ab, arrow bamboo. Part: Sh, shoots; Lv, leaves; Br, branches; Cl, culms; 1–Cl, one–year culms). Tabla 1s. Concentración media (mg/kg ) y valor de p de cada factor para siete elementos minerales. * p < 0,05. (Para las abreviaturas, véase arriba). Season

Elevation

p–value SPRm

SUMm

AUTm p–value H

Cu * 17.44 9.05 10.04 12.66 12.11 10.94 Zn * 30.95 14.51 22.16 21.61 22.84 20.67 Fe * 232.30 353.36 208.90 320.76 253.46 263.04 Mn * 230.64 132.11 320.00 184.37 291.26 178.06 K 8,158.59 7,620.37 8,116.10 8,305.18 7,878.96 7,826.25 Ca * 3,858.59 1,824.48 4,345.19 3,044.40 3,092.10 3,492.78 Mg * 863.79

823.77 2,524.31

Species

p–value

1,322.32 1,313.33 1,549.34

Part Wb

Dhb

p–value Sh

1–Cl

Cu 12.01 10.29 12.05

* 14.47 14.04 15.55 7.02 8.55

Zn 23.67 18.95 20.79

* 20.11 23.84 27.05 17.70 15.43

Fe 239.87 255.50 309.34

* 218.71 302.38 369.33 183.22 211.64

Mn * 295.29 176.48 167.46

* 61.29 381.24 208.93 178.06 78.16

K 7,968.46 7,793.12 7,971.04

* 10,740.36 9,258.57 7,225.72 6,618.28 8,743.27

Ca 3,537.71 2,579.12 3,232.01

* 1,869.84 5,430.65 3,536.32 1,797.73 1,826.67

Mg 1,567.96 1,140.42 1,366.30

* 925.52 2,113.83 1,406.93 1,078.93 871.35

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207

Table 2s. Mean concentration (mg/kg) and p–value of each RF model for every predictor: * p–value < 0.05; – not applicable: Prm, presence mean, Abm, absence mean: 1–Cl, one–year culms. Tabla 2s. Concentración media (mg/kg) y valor de p de cada modelo de bosque aleatorio para cada factor de predicción: * p < 0,05; – no aplicable. (Para las abreviaturas, véase arriba).

COM p–value Prm

SPR Abm p–value Prm

SUM Abm p–value Prm

AUT Abm p–value Prm

Abm

Cu.shoots 5.93 3.52 11.13 4.40 7.62 5.66 – –

–

Zn.shoots 6.91 3.96 12.75 4.81 9.41 6.44 – –

–

Fe.shoots 106.74 52.70 117.86 54.83

–

* 329.84 91.03

– –

Mn.shoots 22.17 17.76 36.20 16.61 41.15 31.72 – –

–

K.shoots 4,937.97 3,073.36 7,431.86 3,446.54 10,640.83 5,168.47 – –

–

Ca.shoots

–

* 1,437.86 251.45

* 2,812.09 400.03

* 1,586.33 356.46

–

Mg.shoots 600.14 174.94 998.56 319.45 * 1,070.83 224.86 – –

–

Cu.leaves 16.92 12.94 25.68 26.27 * 14.96 8.04 9.00 9.48 Zn.leaves 25.78 23.02 32.28 34.95 18.06 15.96 22.60 23.47 Fe.leaves 305.75 300.25 324.79 410.69 433.23 295.27 232.08 224.05 Mn.leaves * 592.32 283.11 674.53 392.80 227.92 159.75 666.29 365.32 K.leaves

* 8,476.56 9611.15 9,603.23 10779.71

* 6,117.00 9,833.75 8,361.14 8,437.92

Ca.leaves 5,636.73 5,350.61 5,983.42 6311.95 * 1,852.00 3,494.22 6,912.07 7,104.79 Mg.leaves 2,142.04 2,099.35 1,602.67 1,752.38 * 630.33 1,362.13 3,329.29 3,342.54 Cu.branches 17.60 14.76 23.39 24.56 9.32 9.35 15.36 14.61 Zn.branches * 34.48 23.79 * 41.89 29.79 15.70 16.86 35.13 28.53 Fe.branches 330.54 390.18 297.52 369.50 394.43 448.89 336.19 327.41 Mn.branches * 274.64 182.96 * 226.52 102.81 190.71 169.64 358.73 260.84 K.branches 7,546.93 7,065.38 7,022.11 6,538.24 7,473.33 6,542.34 8,103.29 8,158.13 Ca.branches 3,764.07 3,439.22 3,765.60 3,994.51 1,798.17 1,718.81 4,605.07 5,316.63 Mg.branches 1,392.61 1,401.47 562.77 679.73 909.17

827.88

* 2,429.64 2,707.58

Cu.culms 7.38 7.01 6.50 5.17 10.11 8.56 7.10 6.31 Zn.culms 19.90 16.93 24.57 27.08 16.99 12.74 16.48 14.91 Fe.culms 180.19 203.41 125.11 86.27 454.73 335.24 117.61 115.50 Mn.culms

* 313.86 113.18 229.96 92.54 167.20 98.58

* 460.62 148.13

K.culms 6,945.27 6,322.02 5,876.52 6,607.14 7,396.67 5,083.78 7,820.57 7,759.17 Ca.culms

* 2,341.86 1,506.31 2,603.95 1,512.95 2,512.33 776.66 2,006.71 2,474.21

Mg.culms 1,117.65 966.87 369.86 419.84 1,072.33 469.06 1,884.86 2,040.88 Cu.1–Cl 7.99 8.78 7.60 8.13 6.98 8.29 8.81 9.91 Zn.1–Cl 16.64 15.73 17.25 21.54 13.93 11.95 17.18 16.42 Fe.1–Cl 217.71 192.58 144.66 93.12 437.88 277.61 196.40 153.80 Mn.1–Cl

* 140.56 57.56

* 74.57 29.79 148.38 66.70

* 203.19 66.21

K.1–Cl 8,590.33 8,883.44 8,885.35 9,611.32 7,991.67 9,078.19 8,551.87 8,077.85 Ca.1–Cl 2,340.55 1,755.04 2,839.74 2,407.89 3,261.33 1,295.57 1,446.73 1,878.01 Mg.1–Cl 1,008.96 896.07 537.42 659.70 996.33 527.02 1,485.91 1,565.41

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Mineral content (mg/kg)

3,000

2,000

208

1,000 r = 0.69 0

2,500

5,000 Ca

7,500

3,000

7,500

2,000

5,000

1,000 0

r = 0.64

0 10

30 Cu

2,500 r = 0.46 0

500

1,000

r = 0.43 0

500

1,000

Fig. 1s. Correlation of mineral elements. Scatter plots show the four pairs of the mineral elements with the strongest correlation. Fig. 1s. Correlaciรณn de elementos minerales. En los grรกficos de dispersiรณn se muestran los cuatro pares de elementos minerales con la mayor correlaciรณn.

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Effects of natural and artificial light on the nocturnal behaviour of the wall gecko B. Martín, H. Pérez, M. Ferrer

Martín, B., Pérez, H., Ferrer, M., 2018. Effects of natural and artificial light on the nocturnal behaviour of the wall gecko. Animal Biodiversity and Conservation, 41.2: 209–215. Abstract Effects of natural and artificial light on the nocturnal behaviour of the wall gecko. In the present study, we evaluated the effects of nocturnal light level (i.e. lunar phase and artificial lighting) on the activity of wall geckos (Tarentola mauritanica) of different ages in an anthropic environment. Data on individual behaviour were collected by direct observation and later examined by means of generalized linear mixed model (GLMM) analysis. The presence of moonlight increased the number of active wall geckos. Artificial lighting reduced the effect of moonlight on the number of active geckos but not on their individual activity. A greater number of adult geckos was found around artificial light as large individuals monopolized the best foraging sites. The ability to use artificially–lit human habitats, particularly on new moon nights, can benefit the foraging activity of nocturnal lizard species such as the wall gecko. Key words: Moon phase, Artificial lighting, Competition, Reptile, Farmhouse Resumen Efectos de la iluminación natural y artificial en el comportamiento nocturno de la salamanquesa común. En este trabajo evaluamos los efectos del grado de iluminación nocturna (fase lunar e iluminación artificial) en la actividad de ejemplares de salamanquesa común (Tarentola mauritanica) de diferentes edades en un ambiente humanizado. Los datos de comportamiento individual se recogieron mediante observación directa y posteriormente se examinaron mediante un análisis lineal generalizado mixto (GLMM, por su sigla en inglés). La presencia de luna aumentó el número de salamanquesas activas. La iluminación artificial redujo el efecto de la luna en el número de salamanquesas activas, pero no en su actividad individual. Además, se observaron más salamanquesas adultas alrededor de fuentes de luz artificial debido a que los individuos de mayor tamaño monopolizaban los mejores lugares de alimentación. La capacidad de utilizar hábitats humanizados iluminados artificialmente, especialmente durante las noches de luna nueva, puede beneficiar la actividad de alimentación en especies de reptiles nocturnos como las salamanquesas. Palabras clave: Fase lunar, Iluminación artificial, Competición, Reptiles, Cortijo Received: 21 IV 17; Conditional acceptance: 12 VI 17; Final acceptance: 15 IX 17 Beatriz Martín, Héctor Pérez, Fundación Migres, International Bird Migration Center (CIMA), ctra. N340 km 85, Tarifa, E–11380 Cádiz, Spain.– Miguel Ferrer, Applied Ecology Group, Doñana Biological Station– CSIC, Seville, Spain. Corresponding author: B. Martín. E–mail: bmartin@fundacionmigres.org

ISSN: 1578–665 X eISSN: 2014–928 X

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Introduction Several vertebrate taxa have been shown to adjust their nocturnal activity and behaviour in terrestrial habitats in relation to the levels of moonlight (e.g. Longland, Price, 1991; Kotler et al., 1991, 2010; Clarke et al., 1996, Lillywhite and Brischoux, 2012). Since higher levels of light increase predation risks, moonlight involves a trade–off between activity and safety (Kotler et al., 2010). Therefore, night–time activity of many nocturnal prey animals is reduced during full–moon nights (Perry and Fisher, 2006). However, in contrast to prey species, the effect of light conditions at night on predators has been less studied (but see Skutelsky, 1996; Mukherjee et al., 2009). Divergent responses to moonlight among species of nocturnal reptiles such as lizards seem to arise from differential activity patterns of both prey and predators of these animals in relation to the level of illumination (Reichmann, 1998). The use of artificial outdoor lighting has increased greatly over the last century, changing illumination levels at night in most of the world’s urban areas and adjacent habitats (Cinzano et al., 2001; Hölker et al., 2010). Reptile species that appear to be at highest risk of being affected by artificial lighting are generally those that occur around human dwellings and nearby roads because they are likely affected by skyglow or glare (Mazerolle et al., 2005; Rich and Longcore, 2006). However, reptile taxa occurring in or close to urban areas have rarely been studied in this context (Perry et al., 2008). Although studies on the effects of lighting on reptiles are scarce, increased prey availability around artificial night lighting has been documented in some lizard species (Perry et al., 2008). The wall gecko (Tarentola mauritanica) is typically associated with vertical surfaces and is frequently observed in anthropogenic landscapes (Luiselli and Capizzi, 1999). It is mainly nocturnal, and frequently uses areas around artificial light sources for foraging. In the present study, we tested the effects of nocturnal light level (i.e. lunar phase and artificial lighting) on wall gecko activity in an anthropic environment. Since the wall gecko is a visual predator, higher natural or artificial light levels can be expected to enhance its nocturnal foraging activity. We hypothesized that larger geckos would monopolize the best sites in terms of prey availability and would therefore be more abundant in artificially lit areas. Material and methods Study species The wall gecko (Fam. Gekkonidae) is a common and widespread reptile species in North Africa, where its ranges extends from Egypt west to Morocco and the northwest Western Sahara, and in the Mediterranean, where it is found from Greece to the the Iberian peninsula, usually in warm, dry coastal areas (IUCN, 2015; Arnold and Ovenden, 2002; Gasc et al., 1997). The wall gecko is mainly nocturnal but it

shows significant diurnal activity in both foraging and thermoregulatory behaviour (Arnold and Ovenden, 2002). It is a medium–sized gecko (45–85 mm, snout–vent length –SVL–; Carretero, 2008) that preys on arthropods. Its diet and its hunting strategy vary with habitat (Arnold and Ovenden, 2002). In urban habitats, wall geckos can be found on walls, ruins and houses (Luiselli and Capizzi, 1999) where they follow a typical sit–and–wait foraging strategy (Hódar et al., 2006). The diet of these geckos is generally similar regarding age and sex (Gil et al., 1994), but geckos can be territorial and aggressive towards conspecifics (Carretero, 2008). Study area Our study was conducted at a farmhouse and its adjacent farm buildings in Tarifa (fig. 1), southern Spain (36º 02' 27.37' N, 5º 37' 04.80'' W). Sampling geckos The survey was carried out at night three times per week between 18 July–1 August and 5–27 September 2013, by an observer wearing a headlamp. The distance between the geckos and the observer did not induce any response in the former. Walls were inspected every 30 minutes over 2 h periods after sunset. No trend was found in the number of geckos during these 2 h periods. Each individual was recorded with respect to two fixed point sources of artificial light (40 W incandescent lamps; fig. 1) defining an area of 3 m around each lamp as influenced by light (distinguishing between 'area with light' and 'dark area'). We visually estimated the size of individuals (< 6 cm; between 6 and 12 cm; between 12 and 16 cm; > 16 cm). Intra–observer variability regarding gecko size estimates was minimized through training sessions prior to sampling. Geckos smaller than 6 cm were classified as juveniles (following Atzori et al., 2007; Lisičić, 2012). We did not distinguish between sexes as it is difficult to do by direct observation (Atzori et al., 2007; Zuffi et al., 2011). We also recorded the proportion of geckos that were active at the time of the census. For each night, we assigned a 'percent moon fullness' (0 % no moon; 100 % full moon). There was no precipitation during the study period, nocturnal temperature was warm (20 ºC on average), and temperature variation was low (between 22 and 19 ºC). Therefore, no surveys were conducted on rainy or cool nights (< 17 ºC). However, the weather was very windy in the study area. Because insect activity is markedly affected by wind speed, surveys were not conducted on extremely windy nights (> 4, Beaufort scale). Moreover, we recorded wind speed per night on the Beaufort scale from the Windguru website to control its effects on the foraging activity of geckos (http://www.windguru.cz/es/). Data analysis To avoid potential biases caused by behavioral effects of capturing the individuals (i.e. observer avoidance),

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Fig. 1. Study area. Sketch of the farmhouse and its adjacent farm building. Black lines, sampled walls; asterisks, fixed point sources of artificial light (40 W incandescent lamps); grey dots, gecko locations in dark areas; black dots, gecko locations in areas with light. (Real length of the maximum rectangle: about 50 x 30 m). Fig. 1. Zona de estudio. Esquema de la finca y su edificio agrícola adyacente. Líneas negras, paredes muestreadas; asteriscos, fuentes fijas de luz artificial (bombillas incandescentes de 40 W); puntos grises, localización de salamanquesas en zonas oscuras; puntos negros, localización de salamanquesas en zonas iluminadas. (Longitud real del rectángulo máximo: aproximadamente 50 x 30 m).

we did not mark the geckos. However, we estimated the total size of the study population by recording the particular location and size of individuals observed at each sample (every 30 minutes) per night; in this way we obtained a snapshot of the geckos every half an hour. Based on the size and location of each particular individual in each sample, we were able to individualize a minimum number of different geckos over the night. At each sample, we distinguished two types of individuals: (1) 'marked' (i.e. geckos observed in one or more prior samples on a particular night); and (2) 'unmarked' (geckos not observed before on a particular night). From this record of individual geckos, we derived a minimum population size using mark–recapture data analysis techniques. Specifically, we applied the Schumacher–Eschmeyer method (Schumacher and Eschmeyer, 1943), a refinement of the Schnabel method (Schnabel, 1938), recommended for use when departures from randomness are probable (Mares et al., 1981). These methods are extensions of the Petersen method to a series of samples and they are useful for closed populations (Krebs, 1999). According to the Schumacher and Eschmeyer method, there is a linear relationship between the proportion of marked individuals in each

sample and the total number of marked animals, thus the total number of animals (N) could be estimated by: m

SM C S MR 2

t=1 m

t t t=1 where, Mt is the number of individuals previously ‘marked’ (before time t); Ct is the total number of individuals caught at time t; Rt is the number of individuals already 'marked' when catching in time t; and m is the total number of samples. Following Schumacher and Eschmeyer (1943), for each night we plotted the number of individuals previously 'marked' (Mt) in the x–axis, and the proportion of 'marked' individuals in the t–th sample (Rt/Ct) in the y–axis. The plotted points should lie on a straight line of slope 1/N passing through the origin (fig. 2), thus allowing estimation of the population size from this regression slope (1/N). We measured population size estimates per night and averaged these results for the total sample of nights. We can not rule out the possibility that the number of geckos of the same size at the same location on different nights was underestimated. However, it is unlikely that different individuals of the same size

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Table 1. Estimates of the gecko population size (N) in the study area (Schumacher and Eschmeyer, 1943). Regression slope (1/N): 'marked' (Mt) in the x–axis, and the proportion of 'marked' individuals in the t–th sample (Rt/Ct) in the y–axis (see Material and methods and figure 2). Estimates per sample night, averaged population size and confidence intervals (CI) for the total sample of nights. Tabla 1. Estimaciones del tamaño de la población de salamanquesas (N) en la zona de estudio (Schumacher y Eschmeyer, 1943). Curva de regresión (1/N): ejemplares marcados (Mt) en el eje de abscisas y proporción de ejemplares marcados en la muestra t (Rt/Ct) en el eje de ordenadas (véase el apartado Material and methods y la figura 2). Estimaciones por muestra nocturna, promedio del tamaño de la población e intervalos de confianza (CI) para la muestra total de noches. Slope

Slope

14.99

0.04

26.88

0.33

3.00

0.10

10.50

0.07

14.43

0.08

12.99

0.06

17.30

0.05

18.38

0.08

12.00

0.11

9.35

0.26

3.86

0.10

9.59

0.06

16.64

0.06

17.61

0.21

4.75

0.03

32.15

0.17

5.91

0.07 14.25

0.07

0.08 11.90

of artificial lighting and moon phase as a predictor. In order to avoid pseudoreplication among samples, we included a random intercept shared by the samples within the same date. We further assessed whether the activity was affected by the moon light cycle using a GLMM binomial (logit link) model structure. We included the same predictors as in the model above along with age: juveniles (< 6 cm) and adults (> 6 cm). In order to avoid pseudoreplication among samples, we also included a random intercept shared by the samples that had the same value for 'individual' on the same date. Additionally, we analyzed differences in the proportion of adults (> 6cm) and juveniles (< 6 cm) in light and dark areas by means of a x2–test. All analyses were performed using R 2.13.0 (R Core Development Core Team, 2011). Results We observed 372 geckos (34 % in areas with light and 66 % in dark areas, respectively) over 19 nights between 18th July and 27th September 2013. According to estimates derived using the Schumacher and Eschmeyer method (1943), the total gecko population in the study ranged between 10 and 17 individuals (95 % confidence interval based on the total sample of nights; table 1). Areas with light had a larger proportion of adult geckos whereas dark areas had a higher proportion of juveniles (x2 = 9.354, df = 1, p–value = 0.002; fig. 3). Table 2 shows backward stepwise procedure results and significant variables. Individual activity and the number of active geckos increased in moonlight. The larger the average size of the geckos, the fewer the individuals seen per night. Juveniles showed more frequent activity than adults. Activity decreased in relation to wind speed. Artificial lighting reduced the

CIs

–95 % –95 % Mean

13.50

10.18

16.82 Rt/Ct

occurred at the same location on the same night in view of territoriality and interspecific aggression (Salvador, 2002) and the gecko's sit–and–wait foraging strategy (Hódar et al., 2006). We analysed the effects of changes in nocturnal light levels on the gecko activity using generalized linear mixed models (GLMM). Regarding the fixed effects, we assessed whether the number of geckos observed per night was conditioned on moon cycle (measured as percent moon fullness), artificial lighting and wind speed using a Poisson (log link) model structure. Because the effect of the level of natural light (i.e. percent moon fullness) on gecko activity may be altered by artificial lighting, we also included the interaction term between the presence

1/N

Mt Fig. 2. Schumacher and Eschmeyer method (1943). Fig. 2. Método de Schumacher y Eschmeyer (1943).

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Table 2. GLMM model results of: A, the number of geckos observed per night (Poisson–log link); B, individual gecko foraging activity (binomial–logit link): S.E., standard error. Tabla 2. Resultados del modelo mixto lineal generalizado de: A, número de salamanquesas observadas por noche (Poisson–logaritmo como función de enlace); B, actividad individual de alimentación de las salamanquesas (binomial–función de enlace logit): S.E., error estándar. A

Estimate

S.E.

z–value

Pr(>|z|)

(Intercept)

Level of effect

2.989

0.262

11.411

< 0.0001

% Moon

0.515

0.196

2.632

0.0085

Gecko size

–0.051

0.021

–2.425

0.0153

Artificial lighting

–0.053

0.214

–0.250

0.8027

Wind speed

Light

–0.121

0.051

–2.347

0.0190

% Moon: artificial lighting

Light

–0.800

0.317

–2.519

0.0118

Level of effect

Estimate

S.E.

z–value

Pr(>|z|)

(Intercept)

–1.419

0.482

–2.943

0.0033

% Moon

2.530

0.548

4.616

< 0.0001

Age

–0.817

0.284

–2.879

0.0040

Artificial lighting

0.035

0.724

0.049

0.9610

Light

Wind speed

–0.341

0.166

–2.056

0.0398

% Moon: artificial lighting

–0.656

0.939

–0.698

0.4850

Light

Observed/expected

2 Dark Light

0 Juveniles

Adults

Fig. 3. Ratio between observed and expected frequencies of juvenile and adult wall geckos in areas with light and in dark areas (x2 = 9.354, df = 1, P < 0.01). Fig. 3. Proporción entre las frecuencias observadas y esperadas de salamanquesas comunes juveniles y adultas en zonas iluminadas y zonas oscuras (x2 = 9,354, gl = 1, P < 0,01).

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effect of moonlight on the number of geckos observed per night but not on individual gecko activity. Discussion Wall geckos in our study modified their nocturnal activity in relation to levels of light. Individual activity and the number of active geckos were higher on full moon nights. This contrasts with many nocturnal animals whose activity decreases on full–moon nights to reduce predation risks (Perry and Fisher, 2006). In anthropic environments, low levels of natural predators of reptiles allow wall gecko activity to be mainly conditioned by prey (Gil et al., 1994). Geckos take advantage of moonlightto increase their ability to detect their prey (Mills, 1986). Artificial night lighting can also enhance their visual detection. Species that are normally active during the night, including other gecko species, have frequently been documented around night lights (Perry et al., 2008). Artificial lighting affects geckos directly by altering their behavior, but it also affects them indirectly by altering the behaviour of their prey. Entomologists have long known that invertebrates are attracted to artificial lighting at night, increasing the availability and predictability of food around such light sources for species like bats (Rydell, 1992), amphibians and reptiles (Garber, 1978; Rich and Longcore, 2006). Activity of many invertebrates increases around the new moon (Bowden and Church, 1973) due to less activity of visual predators (Perry and Fisher, 2006). According to our results, artificial lighting during naturally dark periods allows larger number geckos to forage under new moon nights when invertebrate attraction to artificial light is greater. Activity levels differed between juveniles and adult wall geckos, with higher rates in juveniles than in those in adult individuals. However, adult geckos were more often encountered around artificial lighting, whereas juveniles were more frequently observed in dark areas, which may be linked to territoriality (Salvador, 2002; Carretero, 2008). Differences in spatial distribution of age classes around artificial lighting seem to be related to avoidance of aggression by small individuals (Hitchcock and McBrayer, 2006) and the monopolization of the best sites for foraging by large individuals. The ability to use human habitats with artificial night lighting, particularly on new moon nights, can benefit the foraging activity of nocturnal lizard species such as the wall gecko. Acknowledgements We are grateful to the master students (Aroa Hernández —Banco Santander grants, Claudia L. Cara and Josué Cortés —CIECEM) for help with data collection, and also the Board of the Migres Foundation. This study was funded by the grants of ENDESA and Red Eléctrica de España (REE) and is part of the 'Plan para la Conservación de la Biodiversidad' (Plan for Biodiversity Conservation)

of ENDESA. Special thanks to Purificación Macías Soto, Andrés Araujo and Luz Iglesias, the owner and the keepers respectively, for allowing the use of their farmhouse as our study area. Special thanks to John O'Keeffe and Rashi Sharma for their kind revision of the language of the manuscript. We would like to thank the editor and two anonymous referees for providing us with comments and suggestions that greatly helped in improving the manuscript. References Arnold, N., Ovenden, D., 2002. A Field Guide to the Reptiles and Amphibians of Britain and Europe. London, HarperCollins Publishers. Atzori, A., Berti, F., Cencetti, T., Fornasiero, S., Tamburini, M., Zuffi, M. A. L., 2007. Advances in methodologies of sexing and marking less dimorphic gekkonid lizards: the study case of the Moorish gecko, Tarentola mauritanica. Amphibia–Reptilia, 28: 449–454. Bowden, J., Church, B., 1973. The influence of moonlight on catches of insects in light–traps in Africa. Part II. The effect of moon phase on light–trap catches. Bulletion of Entomological Research, 63: 129–142. Carretero, M. A., 2008. Preferred temperatures of Tarentola mauritanica in spring. Acta Herpetologica, 3: 57–64. Cinzano, P., Falchi, F., Elvidge, C., 2001. The first world atlas of the artificial night sky brightness. Monthly Notices of the Royal Astronomical Society, 328: 689–707. Clarke, J. A., Chopko, J. T., Mackessy, S. P., 1996. The effect of moonlight on activity patterns of adult and juvenile Prairie rattlesnakes (Crotalus viridis viridis). Journal of Herpetology, 30: 192–197. Garber, S., 1978. Opportunistic feeding behaviour of Anolis cristatellus (Iguanidae: Reptilia) in Puerto Rico. Transactions of the Kansas Academy of Sciences, 81: 79–80. Gasc, J., Cabela, A., Crnobrnja-Isailovic, J., Dolmen, D., Grossenbacher, K., Haffner, P., Lescure, J., Martens, H., Martínez-Rica, J., Maurin, H., Oliveira, M., Sofianidou, T., Veith, M., Zuiderwijk, A., 1997. Atlas of Amphibians and Reptiles in Europe. Societas Europaea Herpetologica and Muséum national d’Histoire naturelle, Paris. Gil, M., Guerrero, F., Pérez–Mellado, V., 1994. Seasonal variation in diet composition and prey selection in the Mediterranean gecko Tarentola mauritanica. Israel Journal of Zoology, 40: 61–74. Hitchcock, M., McBrayer, L., 2006. Thermoregulation in nocturnal ecthotherms: seasonal and intraspecific variation in the Mediterranean gecko (Hemidactylus turcicus). Journal of Herpetology, 40: 185–195. Hódar, J. A., Pleguezuelos, J. M., Villafranca, C., Fernández–Cardenete, J. R., 2006. Foraging mode of the Moorish gecko Tarentola mauritanica in an arid environment: Inferences from abiotic setting, prey availability and dietary composition. Journal

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Prediction of Iberian lynx road–mortality in southern Spain: a new approach using the MaxEnt algorithm G. Garrote, J. Fernández–López, G. López, G. Ruiz, M. A. Simón

Garrote, G., Fernández–López, J., López, G., Ruiz, G., Simón, M. A., 2018. Prediction of Iberian lynx road– mortality in southern Spain: a new approach using the MaxEnt algorithm. Animal Biodiversity and Conservation, 41.2: 217–225. Abstract Prediction of Iberian lynx road–mortality in southern Spain: a new approach using the MaxEnt algorithm. In recent years, the Iberian lynx (Lynx pardinus) has experienced a significant increase in the size of its population and in its distribution. The species currently occupies areas in which it had been extinct for decades and new road mortality black spots have been identified. Its conservation requires an intensive risk assessment of road–deaths in potential future distribution areas. Using the MaxEnt algorithm we aimed to identify the roads where there is a greater risk of road collision for the Iberian lynx. More than 1,150 stretches of road were evaluated in Andalusia (southern Spain). Both road–related and habitat variables were included in the model. A total of 1,395 km of the 7,384 km evaluated (18.9 %) were classified as high risk road. Our results could help plan future conservation strategies. To our knowledge, this is the first time that the MaxEnt algorithm has been used to provide spatially–explicit predictions about wildlife road mortality. Key words: Road mortality, Iberian lynx, MaxEnt, Linear data Resumen Predicción de la mortalidad en la carretera del lince ibérico en el sur de España: un nuevo método utilizando el algoritmo MaxEnt. En los últimos años, el tamaño de población del lince ibérico (Lynx pardinus) y su área de distribución han aumentado de forma significativa. Actualmente, la especie habita zonas en las que había estado extinto durante décadas y en las que se han identificado nuevos puntos negros de mortalidad en carretera. La conservación de esta especie requiere que se haga una evaluación exhaustiva del riesgo de muerte en carretera en su posible distribución futura. En este estudio se emplea el algoritmo MaxEnt para identificar las carreteras donde hay mayor riesgo de atropellar un lince ibérico. Más de 1.150 tramos de carreteras fueron evaluados en Andalucía (sur de España). En el modelo se utilizaron variables relacionadas con las carreteras y el hábitat circundante. En total, 1.395 km de los 7.384 km evaluados (el 18,9 %) se han calificado como de alto riesgo. Nuestros resultados podrían ayudar a planificar futuras estrategias de conservación. Hasta donde conocemos, esta es la primera vez que se utiliza el algoritmo MaxEnt para predecir de forma espacialmente explícita la mortalidad de fauna silvestre en carreteras. Palabras clave: Mortalidad en carretera, Lince ibérico, MaxEnt, Datos lineales Received: 31 III 17; Conditional acceptance: 01 VI 17; Final acceptance: 23 IX 17 G. Garrote, G. López, G. Ruiz, Agencia de Medio Ambiente y Agua de Andalucía, c/ Johan Gutenberg s/n., Isla de la Cartuja, 41092 Seville, Spain.– J. Fernández–López, Real Jardín Botánico de Madrid–CSIC, Plaza de Murillo 2, 28014 Madrid, Spain; Depto. de Zoología y Antropología Física, Fac. de Ciencias Biológicas, Univ. Complutense de Madrid, 28040 Madrid, Spain.– M. A. Simón, Consejería de Medio Ambiente de la Junta de Andalucía, c/ Doctor Eduardo García–Trivino López 15, 23009 Jaén, Spain. Corresponding author: Germán Garrote. E–mail: gergarrote@gmail.com ISSN: 1578–665 X eISSN: 2014–928 X

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Introduction Over the past decade, a set of tools for modelling species distributions has been developed for use in wildlife management and conservation (Elith and Leathwick, 2009). The objectives of these approaches are currently expanding and they are now being used to model a set of processes that are closely tied to biological interactions and conservation biology (Yañez–Arenas et al., 2014). The use of these statistical approaches–predict risks for wildlife helps focus the economic efforts required by conservation plans (Mateo–Tomás et al., 2012). In this context, routinely collected road–kill information can be used in species conservation to mitigate the impact of mortality on roads, one of the most important anthropogenic impacts on wildlife communities in a human–dominated context (Forman and Alexander, 1998; Forman, 2003; Sáenz–de–Santa–María and Tellería, 2015; D’Amico et al., 2015). The Iberian lynx is the most endangered felid species in the world, listed as Endangered by the IUCN (IUCN, 2015). In 2002, only two populations remained —one in Andújar–Cardeña, in Eastern Sierra Morena, and one in Doñana, Southern Spain)— totalling fewer than 100 individuals (Guzmán et al., 2004). Since then, however, the Iberian lynx has undergone a significant increase in population size due to conservation measures developed as part of conservation projects for the species (Simón et al., 2012). By 2016, the number of Iberian lynx in the wild was 483 (Junta de Andalucía, 2016), almost five times more than just over a decade before. One of the main goals of these conservation projects was to decrease mortality caused by anthropogenic pressure, especially through poaching and road accidents. Road–kills are considered a chief cause of death in the Iberian lynx in the Doñana area, accounting for 16.7 % of the total annual mortality rate (AMR) in the 1980s (Ferreras et al., 1992). During the first decade of the 21st century, conservation programs carried out numerous prevention measures to prevent road kills in the distribution area of the Iberian lynx, particularly in the Doñana area. A recent mortality study based on radio–tracking data found that road–kills had a minor impact on the Iberian lynx population in 2006–2011 (López et al., 2014), with a percentage of the total AMR of 6.6 % in Sierra Morena and 8.6 % in Doñana. In the same period, the average annual number of lynx found dead due to road–kill (tracked and non–radio tracked) was 3.16 SE:1.83 (Junta de Andalucía, 2016). However, road–kills associated with the expansion of the species have decreased notably since 2012 (2012–2015 annual lynx road–killed: 13 SE:6.05; Junta de Andalucía, 2016). Both natural colonisation and reintroductions have led to the species being present today in areas in which measures to prevent road–kills have not yet been implemented and where new road–kill black spots have been identified. Within this scenario, managers need tools that allow them to detect potentially dangerous areas for the lynx before it reaches their areas in order to plan and put conservation measures into practice.

Despite the historical importance of road kills in the conservation of the Iberian lynx, the factors involved have not been identified. Existing studies have been limited to estimating mortality rates (Ferreras et al., 1992; López et al., 2014). Previous studies have found that road–kills involving carnivores depend on population density, species biology, habitat and landscape structure, and road and traffic characteristics (Clevenger et al., 2003; Malo et al., 2004, Grilo et al., 2009). Road–related features that have shown to affect the spatial distribution of casualties include vehicle speed (Jaarsma et al., 2006), traffic volume (Clarke et al., 1998), roadside topography (Clevenger et al., 2003), adjacent vegetative cover (Ramp et al., 2005) and type of nearby passages (Clevenger et al., 2003; Malo et al., 2004). To date, most studies describing casualty patterns have analysed only the role of proximate causes of road–kill (e.g., local road features), so their management recommendations are difficult to translate to other localities (D'amico et al., 2015). In the present we aimed to identify the roads with the highest risk of road collision for the Iberian lynx. We included road kill data the Andalusian road network and used both road–related feature and land cover features to build a risk map for this critically endangered felid. It uses the MaxEnt algorithm (Phillips et al., 2006; Phillips and Dudík, 2008), which is especially efficient in dealing with presence–only data and a small sample size (Wisz et al., 2008). These aspects make MaxEnt appropriate for modelling road–kills, because absence data cannot be recorded with confidence in many cases, and the only data available are often government reports. To our knowledge, this is the first time that the MaxEnt algorithm has been used to provide spatially–explicit predictions about wildlife road mortality using linear data, i.e. non–raster data. Material and methods A total of 99 Iberian lynx road fatalities were recorded by the Andalusian Regional Ministry of the Environment from 1979 to 2013. From these data, we used only the records of deaths on monitorized stretches, i.e. roads with available data about use intensity and vehicle speed. A total of 46 records from 2001 to 2013 were selected for analysis. The predictors we used were four road variables taken from the Basic Spatial Data of Andalusia database (BSDA) (table 1), namely, (i) vehicle speed (SPEED), (ii) intensity of road use (INTENS), (iii) road hierarchy (HIERAR) and (iv) road type (TYPE). We also included as a predictor a categorical land cover variable taken the CORINE Land Cover Map (COVER) because many collisions can be linked to the attraction exercised by a specific habitat (Barrientos and Bolonio, 2009; D’Amico et al., 2015) We used the MaxEnt algorithm (Phillips et al., 2006; Phillips and Dudík, 2008) to model the distribution of Iberian lynx fatalities on the road network. We used a methodology similar to that described by Elith et al. (2011), concerning the distribution of the freshwater

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Table 1. Description of variables included as predictors in the analysis. Tabla 1. Descripción de las variables incluidas como factores predictores en el análisis. Road–related variables SPEED

Continuous. Average speed (km/h) of the total vehicles between 2008 and 2011

INTENS

Continuous. Average intensity of use (veh/day) of the total of vehicles between

2010 and 2011

HIERAR

Categorical, 10 levels. Titularity and dependences of each road (see spatial

reference for Andalusia Database, SRAD)

TYPE

Categorical, 5 levels. Physical characteristics of the road, number of road lanes, etc.

(see spatial reference for Andalusia Database, SRAD)

Land cover variables COVER

Categorical, 11 levels. Obtained from CORINE Land Cover 2006: non–irrigated

arable land; permanently irrigated land; paddy/rice fields; vineyards; olive grove;

complex cultivation patterns/farms; agro–forestry areas; natural grasslands/pasture;

coniferous forest; sclerophyllous shrub; transitional woodland–shrub

fish Gadopsis bispinosus in a water basin (GIS–vector river data). As background data (282 samples for this data set) we used each stretch of road in the entire area in which Iberian lynx populations are stable (fig. 1). All these stretches, that range 0.1 km to 33 km long, are defined as road fragments between kilometer points with homogeneous road features: use intensity, speed, and type of road. Next, we extracted variables for all roads stretches on which an Iberian lynx had been killed and built a second matrix to use as the presence dataset (46 records). These two datasets were managed as section–based (non–gridded) data and included the MaxEnt algorithm as a samples–with–data (SWD) format. We ran 10 replicates of the MaxEnt algorithm following the advice of Merow et al. (2013). We used raw output to avoid post–processing assumptions (Merow et al., 2013); we chose regularization multiplier b = 1 by default since we had no prior information that could be used to optimize these parameters; we removed product and hinge features, avoiding interactions between variables and increasing the interpretability of our models (Merow et al., 2013). The convergence threshold was set at 0.00001. We ran 10 replicates, using as training–data 70 % of the presences and the remaining 30 % for Maxent intrinsic AUC test validation (Fielding and Bell, 1997). To widen the scope of our results to cover areas of conservation interest (near current Iberian lynx distribution areas and target areas for reintroduction programmes, etc.), we projected the model on all roads in Andalusia, using a third dataset that included variables for all available Andalusian roads, a total of 7,398 km including 1,175 stretches. In order to estimate the transferability of our model areas with stable

populations of Iberian lynx to the whole of Andalusia, we explored the clamping scores provided by MaxEnt (see Randin et al., 2006; Barrientos and de Dios Miranda, 2012). We used the Equal test sensitivity and specificity to establish a threshold to categorize roads with a high–low risk of lynx kill to enable a comparative analysis between road stretches. We used QGIS 2.2.0 (QGIS Development Team, 2014) to plot the projected risk of Iberian lynx road– kills provided by MaxEnt onto a geographical platform to obtain a risk map for the evaluated roads and a hazard index for the Iberian lynx on these roads. Results The MaxEnt intrinsic validation test obtained an AUC under ROC score of 0–804 (SD ± 0.056) for the mean of 10 replicates. Clamping scores for model projection ranged 0–0.06, with a mean of 0.00021. Thus, we consider that the values of the background variables provided were robust enough to interpolate the model for the whole of the Andalusian road network. The most important variable for Iberian lynx road– kills was COVER, followed by HIERAR and SPEED according to the jackknife plot of testing gain for variable importance provided by MaxEnt (Phillips et al., 2006) (fig. 2). INTENS and TYPE had the lowest scores. When analysing COVER variables, we found that roads with scrub land cover had a greater probability of Iberian lynx road–kills. Within the hierarchy variable (HIERAR), the most dangerous roads were highways and other main roads that link large urban areas. The SPEED variable had a positive relationship with Iberian lynx deaths at speeds over 90 km/h, in

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–6

–4

–2

38º 0'

36º 0' Road stretches analyzed High risk Low risk Other roads Fig. 1. Risk map for the 1,175 roads evaluated. Fig. 1. Mapa de riesgo de las 1.175 carreteras evaluadas.

agreement with HIERAR effects. The INTENS variable had little importance in our model and its effect was negative, contrary to our initial expectations (fig. 2). The road–kill risk provided by our model ranged minimum risk of 0 to maximum risk of 1 with a mean of 0.222. Threshold between high to low kill risks was established at 0.233 following equal test specificity and sensitivity criterion. A total of 1,395 km out of the 7,384 km evaluated were classified as high risk road. Discussion Our model identified the stretches of the road network at a regional level where road collisions for the Iberian lynx were most likely. An internal validation test provided by MaxEnt obtained an AUC score = 0.804 (SD ± 0.056). A model is usually considered useful if AUC scores are higher than 0.75 (Phillips and Dudik, 2008; Lobo et al., 2007). Our model therefore performed well in predicting Iberian lynx road–kills. To increase performance, further variables such as fine–scale habitat type surrounding a road or the presence of mitigation measures on each stretches could be included in the model (Barrientos and Bolonio, 2009). Our model shows that the main factor affecting road–kill records was the type of land cover around the road, this being the key variable (Malo et al., 2004). Scrub and coniferous forest were identified as the most dangerous land covers. This pattern could be related to two main processes on a large scale: (1) habitat fragmentation caused by linear infrastruc-

tures and (2) the increase of the edge effect in areas occupied by terrestrial mammals (Tellería et al., 2011). Since Iberian lynx mainly set up their territories in Mediterranean scrub areas with differing types of forest coverage (Palomares, 2001; Fernandez et al., 2006), this fragmentation and the edge effect could lead–an increase in mortality due to the penetration of dangers non–core areas (animal–vehicle collisions in our case) (Palomares, 2001). On a local scale, these kinds of habitats attract more potential prey. Moreover, high rabbit (the Iberian lynx's staple prey item) densities are recorded near some roads (Garrote, obs. pers.) due to (1) the prohibition of shooting near roads and (2) the suitability of embankments for the building of warrens (Barrientos and Bolonio, 2009). These high rabbit densities increase the probability of incursion by lynxes onto roads and therefore the likelihood of lynx road–kills. Mitigation measures such as clearing scrub alongside roads or ensuring that new roads do not run through this type of habitat could help to reduce lynx to vehicle collisions. Other habitats that seem to favour Iberian lynx road–kills include paddy fields, farms and vineyards. In this agricultural matrix, roadside ditches are a typical refuge for rabbits because they are suitable places to build their warrens (Calvete et al., 2004; Gea–Izquierdo et al., 2006). These concentrations attract carnivores such as the Iberian lynx, thereby increasing the probabilities of road–kills. Similar results were obtained by Barrientos and Bolonio (2009), who suggested that road–kills in dry agricultural landscapes are linked to the presence of a greater number of rabbit warrens in road verges and greater traffic flow and speed.

0.10

Vehicle intensity

0.06

Road hierarchy 0.8 0.6 0.4

Hw access

Local

Inter local

Highways

0.2

Trans. shrub

Toad type

0.2 Scler. shrub

0.18

Pasture

Vehicle speed

0.4

Conif.forests

0.42

0.6

Farms

Road hierarchy

0.8

Paddy

0.60

Road–kill probability

Land cover type

Irrigated

Jacknife score

Road–kill probability

Variable

221

Road–kill probability

Animal Biodiversity and Conservation 41.2 (2018)

Vehicle speed 0.85 0.80 0.75 0.70 0.65 0

60 80 100 120 km/h

Fig. 2. A, contributions of the variables on Iberian lynx road–mortality. Effect of the three most important variables: B, land cover type; C, road hierarchy; D, vehicle speed. Fig. 2. A, contribución de las variables en la mortalidad en carreteras del lince ibérico. Efecto de las tres variables más importantes: B, tipo de cobertura terrestre; C, jerarquía de carreteras; D, velocidad del vehículo.

The two other most important variables in our model, HIERAR and SPEED, are intrinsically related. SPEED can be related to lynx mortality due to the fact that driver reaction time decreases with increasing speed (Barrientos and Bolonio, 2009). HIERAR is related to the socioeconomic characteristics (hierarchy levels are mainly related to population connections, and local or state ownership, etc.) and the physical characteristics of the road (width, hard–shoulder size, etc.) (Barrientos and Bolonio, 2009). Therefore, the better the roads, the greater possibility of reaching higher speeds. In agreement with the results obtained for other mammals (Malo et al., 2004; Clark et al., 1998), speeds over 90 km/h were positively related to the probability of the deaths of Iberian lynx. Mitigation actions in main roads such as (1) increasing the effectiveness of fencing, (2) ensuring the proper maintenance of existing fences, and (3) building of road overpasses or underpasses for animals should be contemplated for the most dangerous roads. These mitigation measures have proven to be the most effective in reducing wildlife mortality on roads (Rytwinski et al., 2016).

From a conservation point of view, the most important contribution of the model described here, is its ability to identify the potentially dangerous stretches of roads on a large scale. This information may be incorporated into the first stages of selecting reintroduction areas for the Iberian lynx. The protocol to select Iberian lynx reintroduction areas includes both the evaluation of road–kill risk and the identification of especially dangerous road stretches (Junta de Andalucía, 2012). Fine–scale analysis/detection of black spots by means of fieldwork can become laborious, expensive and slow if the road network is large. Initially identifying the dangerous stretches using the model may help to guide later work in more detail, concentrating efforts in the areas of greater danger, and thereby optimising resources. At the 5th Iberian Lynx International Conservation Seminar (Garrote, 2016), a peninsula–wide habitat suitability model was presented to detect potential areas for Iberian lynx reintroduction. It would be very useful to include the results of this study in the model for selecting reintroduction areas and their subsequent priority order or categorisation. Similar

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exercises should be carried out for the different CCAA and Portugal, with the necessary adaptations depending on the type of data available as they are not homogeneous throughout the peninsula. The use of the MaxEnt algorithm in species distribution models has been extensively tested in many conservation plans. It has been broadly used to design reserve networks (Meller et al., 2014) and topredict the invasion ability of exotic species (Peterson and Vieglais, 2001). This study describes an approach using this algorithm to conduct a wide–range analysis based on digital information provided by local administrations. Predictive models for animal–vehicle collisions can generate high performance risk maps to be used to channel investment for improving the safety of both drivers and animals. In summary, the Iberian lynx is categorized as Endangered (IUCN, 2015), but its range is currently expanding. This increase, however, implies the species is present near many more kilometers of roads, so the probability of road kill is higher. For this reason, predicting the most dangerous roads for the Iberian lynx should be a priority in conservation planning reintroduction initiatives. Acknowledgements The study was supported by the LIFE Project 10NAT/ ES/570 'Recovery of the historical distribution of the Iberian lynx (Lynx pardinus) in Spain and Portugal'. We thank all the staff of the LIFE projects involved in collecting road–kill data and the two anonymous reviewers for their constructive comments. References Barrientos, R., Bolonio, L., 2009. The presence of rabbits adjacent–roads increases polecat road mortality. Biodiversity and Conservation, 18: 405–418. Barrientos, R., de Dios Miranda, J., 2012. Can we explain regional abundance and road–kill patterns with variables derived localscale roadkill models? Evaluating transferability with the European polecat. Diversity and Distribution, 18: 635–647. Calvete, C., Estrada, R., Angulo, E., Cabezas–Ruiz, S., 2004. Habitat factors related–wild rabbit conservation in an agricultural landscape. Landscape Ecology, 19(5): 531–542. Clarke, G. P., White, P. C., Harris, S., 1998. Effects of roads on badger Meles meles populations in south–west England. Biological Conservation, 86(2): 117–124. Clevenger, A. P., Chruszcz, B., Gunson, K. E., 2003. Spatial patterns and factors influencing small vertebrate fauna vehicle–kill aggregations. Biological Conservation, 109: 15–26. D’Amico, M., Román, J., de los Reyes, L., Revilla, E., 2015. Vertebrate road–kill patterns in Mediterranean habitats: who, when and where. Biological Conservation, 191: 234–242. Elith, J., Graham, C. H., 2009. Do they? How do they?

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Proosdij, A. S., Sosef, M. S., Wieringa, J. J., Raes, N., 2015. Minimum required number of specimen records–develop accurate species distribution models. Ecography, 39: 542–552. QGIS Development Team., 2014. QGIS Geographic Information System. Open Source Geospatial Foundation Project. https://www.qgis.org/en/site/ [Accessed on 25 June 2017]. Randin, C. H., Dirnböck, T., Dullinger, S., Zimmermann, N. E., Zappa, M., Guisan, A., 2006. Are niche–based species distribution models transferable in space? Journal of Biogeography, 33: 1689–1703. Ramp, D., Caldwell, J., Edwards, K. A., Warton, D., Croft, D. B., 2005. Modelling of wildlife fatality hotpsots along the Snowy Mountain Highway in New South Wales, Australia. Biological Conservation, 126: 474–490. Rytwinski, T., Soanes, K., Jaeger, J. A. G., Fahrig, L., Findlay, C. S., Houlahan, J., van der Ree, R., Van der Grift, E. A., 2016. How Effective Is Road Mitigation at Reducing Road–Kill? A Meta–Analysis. PLOS One, 11: e0166941, doi:10.1371/journal. pone.0166941. Sáenz–de–Santa–María, A., Tellería, J. L., 2015. Wildlife–vehicle collisions in Spain. European Journal of Wildlife Research, 61: 399–406. Simón, M. A., Gil–Sánchez, J. M., Ruiz, G., Garrote, G., Mccain, E. B., Fernandez, L., López–Parra, M., Rojas, E., Arenas–Rojas, R., Del Rey, T., García– Tardío, M, López G., 2012. Reverse of the decline of the endangered Iberian lynx. Conservation Biology, 26: 731–736. Tellería, J. L., Díaz, J. A., Pérez–Tris, J., De Juana, E., De la Hera, I., Iraeta, P., Salvador, A., Santos, T., 2011. Barrier effects on vertebrate distribution caused by a motorway crossing through fragmented forest landscape. Animal Biodiversity and Conservation, 34(2): 331–340. Wisz, M. S., Hijmans, R. J., Li, J., Peterson, A. T., Graham, C. H., Guisan A., 2008. Effects of sample size on the performance of species distribution models. Diversity and Distribution. 14: 763–773. Yañez–Arenas, C., Peterson, A. T., Mokondoko, P., Rojas–Soto, O., Martínez–Meyer, E., 2014. The Use of Ecological Niche Modeling–Infer Potential Risk Areas of Snakebite in the Mexican State of Veracruz. PLOS One, 9: e100957.

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Appendix 1. Road stretches with high road mortality risk. Apéndice 1. Tramos de carreteras analizados con alto riesgo de mortalidad por atropello. ID

From – to

A–456

Fuentes de Andalucía – Lora del Río

ID From – to A–8126 Morón de la Frontera – Algodonales

A–4132 Órgiva – Trevélez

A–486

San Juan del Puerto – Bonares

A–4127 A–348 – Bérchules

A–5052 El Rompido – Punta Umbría

A–2233 Conil de la Frontera – Barbate

A–4105 Access – La Peza and Lopera

A–395

Granada – Sierra Nevada

HU–4402 Villanueva de los Castillejos –

A–455

Cazalla de La Sierra – Lora del Río

A–475

Calañas – Puebla de Guzmán

HU–6400 El Granado – Puente del Chanza

A–389

Arcos de la Frontera – Medina

(Pomarao)

Sanlúcar de Guadiana

Sidonia

HU–7104 N–435 – Cabezas Rubias

A–499

Old A–382 Old road Jerez de la Frontera

Ayamonte – Puebla de Guzmán

A–7378 A–2300 – A–374 by Arroyo

– Arcos de la Frontera

Montecorto

A–476

El Castillo de las Guardas – Minas

A–352

de Riotinto

A–4129 A–4132 – Capileira by Bubión

A–363

Morón de la Frontera – Olvera

A–421

Villafranca de Córdoba – Villanueva

A–4133 Vélez de Benaudalla – Motril

de Córdoba

A–2229 A–2230 – Vejer de la Frontera

Cuevas del Almanzora a Garrucha

A–2226 Benalup – Casas Viejas – A–381

A–402

A–4132 Órgiva – Trevélez

A–4050 N–323 – Almuñécar

Moraleda de Zafayona – Viñuela

A–4132 Órgiva – Trevélez

A–366

A–4102 A–92 – Alcudia de Guadix

A–3176 Villaharta – Puerto del Caballón by

A–337

Cherín – La Calahorra

Obejo

A–369

Ronda – Gaucín

A–496

Valverde del Camino – Cabezas

A–348

Lanjarón – Almería by Ugíjar

Rubias

A–314

Vejer de la Frontera – Barbate

A–493

A–349

Tabernas – Olula del Río by Macael

Camino

A–395

Granada – Sierra Nevada

A–373

Villamartín – Algatocín

A–497

Huelva – Punta Umbría

A–355

Casapalma – Marbella

A–479

Aracena – Campofrío

A–7200 A–92 – Estación de Salinas by

A–422

Alcaracejos – Belalcázar by

Archidona

Hinojosa del Duque

A–483

Ronda – Coín

La Palma del Condado – Valverdel

Bollullos del Condado –

A–7206 Cómpeta – N–340 by Algarrobo

Matalascañas

A–332

Cuevas de Almanzora – San Juan

A–480

Chipiona – Jerez de la Frontera

de los Terreros

A–369

Ronda – Gaucín

Alhaurín el Grande – Fuengirola

A–310

Puente Génave – Siles

A–387

A–7282 Access – Antequera

A–319

Peal de Becerro – Hornos por Cazorla

A–324

A–355

Casapalma – Marbella

La Cerradura – Huelma

A–92N Guadix – Límite de Región de Murcia

A–2325 N–340 – Punta Paloma

A–7207 Canillas de Albaida – Torrox – Costa

A–494

San Juan del Puerto – Matalascañas

by Cómpeta

por Mazagón

A–497

Huelva – Punta Umbría

A–1100 N–340a – A–334 por Uleila del Campo

A–405

Gaucín – San Roque

A–395

Granada–Sierra Nevada

Órgiva – Vélez de Benaudalla

A–315

Torreperogil – Baza por Pozo Alcón

A–346

Animal Biodiversity and Conservation 41.2 (2018)

225

Appendix 1. (Cont.)

From – to

Old A–483 Bollullos del Condado – Matalascañas

N–435

Huelva – Zafra

N–340

N–630

Ruta de la Plata (Sevilla – Gijon)

N–433 A–66 Ruta la Plata – Rosal de la

A–7

Almeria – Murcia

Frontera

A–7

Almeria – Murcia

MA–20 Malaga – Almeria

N–340a Almeria – Murcia

N–630 Ruta de la Plata (Sevilla – Gijon)

N–340a Almeria – Murcia

N–435

Huelva – Zafra

N–435

Huelva – Zafra

N–435

Huelva – Zafra

AP–7

Algeciras – Málaga (peaje)

A–66 Ruta de la Plata – Rosal de la

AP–7

Algeciras – Málaga (peaje)

N–433

From – to Malaga – Almeria

Frontera

N–340a Almeria – Murcia

A–49

Sevilla – Huelva – Ayamonte

N–433

A–44

Linares – Motril

Frontera

A–4

Sevilla – Córdoba – Madrid

A–7

Almeria – Murcia

N–435

Huelva – Zafra

A–7

Algeciras – Málaga

A–45

Cordoba – Malaga

A–7

Algeciras – Málaga

A–48

Cádiz – Algeciras

N–340a Almeria – Murcia

A–66 Ruta de la Plata – Rosal de la

CA–34 Algeciras – Málaga

A–44

Linares – Motril

N–340

Cádiz – Algeciras

N–340

Malaga – Almeria

N–435

Huelva – Zafra

N–420

A–4 Montoro – Toledo

AP–7

Algeciras – Málaga (peaje)

N–340a Almeria – Murcia

AP–7

Algeciras – Málaga (peaje)

N–331

Cordoba – Malaga

N–442 N–442

A–49

Sevilla – Huelva – Ayamonte

N–435

Huelva – Zafra

No name

N–433

A–66 Ruta de la Plata – Rosal de la

No name

Frontera

N–340a Almeria– Murcia

226

Garrote et al.

Animal Biodiversity and Conservation 41.2 (2018)

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Raptor nest–site use in relation to the proximity of coalbed–methane development J. D. Carlisle, L. E. Sanders, A. D. Chalfoun, K. G. Gerow

Carlisle, J. D., Sanders, L. E., Chalfoun, A. D., Gerow, K. G., 2018. Raptor nest–site use in relation to the proximity of coalbed–methane development. Animal Biodiversity and Conservation, 41.2: 227–243. Abstract Raptor nest–site use in relation to the proximity of coalbed–methane development. Energy development such as coalbed–methane (CBM) extraction is a major land use with largely unknown consequences for many animal species. Some raptor species may be especially vulnerable to habitat changes due to energy development given their ecological requirements and population trajectories. Using 12,977 observations of 3,074 nests of 12 raptor species across nine years (2003–2011) in the Powder River Basin, Wyoming, USA, we evaluated relationships between raptor nest–site use and CBM development. Our objectives were to determine temporal trends in nest–use rates, and whether nest–site use was related to the proximity of CBM development. Across the study area, nest–use rates varied across species and years in a non–linear fashion. We developed a novel randomization test to assess differences in use between nests at developed and undeveloped sites, while controlling for annual variation in nest–site use. Red–tailed hawks (Buteo jamaicensis), burrowing owls (Athene cunicularia), and long–eared owls (Asio otus) used nests in undeveloped areas more than nests in developed areas (i.e. nests near CBM development). Differences between development groups were equivocal for the remaining nine species; however, we caution that we likely had lower statistical power to detect differences for rarer species. Our findings suggest potential avoidance of nesting in areas near CBM development by some species and reveal that CBM effects may be fairly consistent across distances between 400–2,415 m from wells. Future work should consider habitat preferences and fitness outcomes, and control for other key factors such as local prey availability, raptor densities, and weather. Key words: Raptors, Coalbed methane, Energy development, Nest–site use, Wildlife conservation, Randomization test Resumen Uso de los sitios de nidificación en rapaces en relación con la proximidad de yacimientos de metano en capas de carbón. Las actividades de generación de energía como la extracción de metano en capas de carbón es un uso importante de la tierra que tiene consecuencias prácticamente desconocidas para numerosas especies de animales. Algunas especies de rapaces, dadas de sus necesidades y su evolución demográfica, pueden ser especialmente vulnerables a los cambios en el hábitat provocados por la generación de energía. Utilizando 12.977 observaciones de 3.074 nidos de 12 especies de rapaces durante nueve años (2003–2011) en la cuenca del río Powder, en Wyoming, EE.UU., evaluamos la relación entre el uso de los sitios de nidificación de las rapaces y la extracción de metano en capas de carbón. Nuestros objetivos fueron determinar tendencias temporales en los índices de utilización de nidos y si el uso de los sitios de nidificación estaba relacionado con la proximidad a yacimientos de metano en capas de carbón. En la zona de estudio, los índices de utilización de los nidos variaron en función de la especie y de los años de forma no lineal. Elaboramos una nueva prueba de aleatorización para evaluar las diferencias de uso entre los nidos en zonas extractivas y zonas no extractivas, a la vez que se controlaba la variación anual del uso de los sitios de nidificación. El ratonero de cola roja (Buteo jamaicensis), la lechuza madriguera (Athene cunicularia) y el búho chico (Asio otus) utilizaron nidos en zonas no extractivas más que en zonas extractivas (es decir, nidos cercanos a yacimientos de metano en capas de carbón). Las diferencias entre los grupos fueron ambiguas para las otras nueve especies; no obstante, hemos de advertir que probablemente teníamos menor potencia estadística para detectar diferencias en especies menos ISSN: 1578–665 X eISSN: 2014–928 X

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Carlisle et al.

frecuentes. Nuestros resultados sugieren que algunas especies podrían evitar nidificar en zonas cercanas a los yacimientos de metano en capas de carbón y revelan que los efectos de esta sustancia pueden ser relativamente constantes en un radio de entre 400 y 2.415 m de los pozos. En estudios futuros se deberían analizar a las preferencias de hábitat y a la eficacia biológica, y controlar otros factores decisivos como la disponibilidad de presas en el ámbito local, la densidad de rapaces y las condiciones meteorológicas. Palabras clave: Rapaces, Metano en capas de carbón, Generación de energía, Uso de sitios de nidificación, Conservación de la fauna silvestre, Prueba de aleatorización Received: 24 II 17; Conditional acceptance: 21 VII 17; Final acceptance: 28 IX 17 Jason D. Carlisle, Program in Ecology, Wyoming Cooperative Fish and Wildlife Research Unit, Dept. of Zoology and Physiology, Univ. of Wyoming, Laramie, Wyoming, USA.– Lindsey E. Sanders, Wyoming Cooperative Fish and Wildlife Research Unit, Dept. of Zoology and Physiology, Univ. of Wyoming, Laramie, Wyoming, USA.– Anna D. Chalfoun, U. S. Geological Survey Wyoming Cooperative Fish and Wildlife Research Unit, Dept. of Zoology and Physiology, Univ. of Wyoming, Laramie, Wyoming, USA.– Ken G. Gerow, Dept. of Statistics, Univ. of Wyoming, Laramie, Wyoming, USA. Corresponding author: Jason D. Carlisle. E–mail: jcarlisle@west–inc.com

Animal Biodiversity and Conservation 41.2 (2018)

Introduction Human–induced habitat changes can alter the ability of landscapes to support wildlife populations (Munns, 2006) via behavioral avoidance of disturbed areas (Frid and Dill, 2002; Blumstein, 2006; Sih, 2013), loss of critical habitat elements (Coristine and Kerr, 2011), and/or fitness consequences (i.e. decreased survival and/or reproduction; Acevedo–Whitehouse and Duffus, 2009; Kociolek et al., 2011). Energy development (i.e. oil, natural gas, coal, solar and wind) continues to increase as a global land use and a ubiquitous form of human–induced habitat change (U.S. Energy Information Administration, 2013; Allred et al., 2015; Jones et al., 2015), with largely unknown consequences for wildlife (Gilbert and Chalfoun, 2011; Garvin et al., 2011; Northrup and Wittemyer, 2013;). Coalbed natural gas, also known as coalbed methane (hereafter, CBM), has emerged in recent decades as an alternative source of natural gas extracted from coal beds (U.S. Geological Survey, 2000). CBM development can influence wildlife habitat, demographic rates, and population persistence (Walker et al., 2007; Doherty et al., 2008; Buchanan et al., 2014). Identifying patterns of habitat use over time in relation to human activity such as CBM development is a critical first step in assessing development–related effects on wildlife (Kennedy et al., 2014). Birds of prey, including members of the orders Accipitriformes (e.g., hawks, eagles, harriers, and vultures), Falconiformes (e.g., falcons), and Strigiformes (e.g., owls), (hereafter, raptors) play important roles in their ecological communities, can serve as indicators of biodiversity and/or environmental degradation, and often have a high profile in the public eye and in conservation strategies (Sergio et al., 2005; Bart et al., 2006; Burgas et al., 2014; Donázar et al., 2016). Concomitant with expanded energy development worldwide, raptors have garnered increased conservation attention as a result of regional population declines (Woffinden and Murphy, 1989; Kochert and Steenhof, 2002) and demonstrated sensitivity to habitat change (Krüger, 2002; Brown et al., 2014; Coates et al., 2014) and human activities (Suter and Joness, 1981; Martínez–Abraín et al., 2010). Furthermore, federal laws and international treaties applicable in many parts of the world have been established to protect raptor species (e.g., Migratory Bird Treaty Act, Bald and Golden Eagle Protection Act, Endangered Species Act, Birds Directive, and Berne Convention; Romin and Muck, 2002; Stroud, 2003), making the assessment and monitoring of impacts from anthropogenic activities a primary concern for many government agencies. For birds, including raptors, the nesting period is a critical life history stage in terms of fitness and population viability (Stahl and Oli, 2006). Human activities such as energy development that encroach upon historic nesting habitats may elicit behavioral avoidance (Krüger, 2002; Coates et al., 2014; Johnston et al., 2014) or indirectly affect nesting success (Hethcoat and Chalfoun, 2015a). Because raptors tend to exhibit high fidelity to nesting areas (Newton, 1979; Millsap et al., 2015) and often reuse the same nests, changes in

229

nest–site use over time could signal the existence of one or both of the aforementioned effects and warrant further examination. Monitoring activities that identify potential changes in raptor nesting habitat use with respect to energy development will be particularly important for management prescriptions geared towards maintaining sustainable raptor populations. Our main objective was to determine whether the proximity of CBM development influenced raptor nest–use rates. Determining the risks that wildlife face from anthropogenic stressors such as energy development is often difficult because individual species may react to stressors differently, the appropriate scale of analysis can be unclear, and available datasets are typically limited (Munns, 2006). Here, we leverage a large, long–term dataset that overcomes these challenges by including multiple species, spatially–explicit data analyzed at multiple spatial scales, and several thousand observations spanning nearly a decade. Our specific objectives were to (1) determine temporal trends in raptor nest–use rates, by species; (2) examine whether nest–use rates were related to the proximity of the nest to CBM development; and (3) investigate whether effects were consistent across multiple distance–to–development thresholds. Material and methods Study area Data for our study were collected in the Powder River Basin (PRB), one of the region's key raptor areas (Olendorff and Kochert, 1992). The PRB includes ~29,800 km2, overlapping Campbell, Johnson, and Sheridan counties in northeast Wyoming, USA (44.2º N, 106.15º W; fig. 1). The PRB (excluding the Bighorn Mountains on the PRB’s western border) is characterized by a mixture of lowlands, rugged badlands, and steep buttes ranging in elevation from ~1,000–1,600 m (Knight, 1994). The climate is semi– arid, with average annual precipitation of ~25–40 cm (Knight, 1994). The dominant land cover types (Homer et al., 2015) within the study area were grasslands (54.8 % of land area), shrublands (30.2 %), and forests (10.3 %, primarily concentrated in the Bighorn Mountains). Land cover types associated with anthropogenic activities were relatively uncommon, with 1.4 % of the land area classified as planted/cultivated agricultural lands, and 0.8 % of the area classified as developed (Homer et al., 2015). Approximately 83,800 people lived in the PRB in 2010, with a human population density of 2.81 persons/km2 (U. S. Census Bureau, 2017). The majority of the human population in the PRB (61.0 %) was concentrated in only three cities: Gillette, Sheridan, and Buffalo (U. S. Census Bureau, 2017). Dominant human land uses within the study area included energy development (i.e., oil and natural gas extraction, coal mining), agriculture, and livestock ranching/grazing. From 2003–2011, Wyoming was responsible for 20–30 % of the CBM production within the United States (U.S. Energy Information Administration, 2014),

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with the PRB being one of the most productive CBM areas nationwide (U.S. Geological Survey, 2000). Much of the land surface and/or subsurface (i.e. mineral rights) within the PRB was managed by the U.S. Bureau of Land Management (BLM) through its Buffalo Field Office. The BLM required energy developers requesting permits for new CBM wells to survey for nesting raptors within 805 m (0.5 miles) of proposed development sites for the first five years following proposal submission (Powder River Basin Wildlife Taskforce, 2005). Through the duration of the study, well construction and other surface disturbances were restricted within 805 m of occupied raptor nests between 1 February and 31 July to avoid disturbance to breeding raptors. Most raptor nests in this study were either located on private land, or landlocked by private land and difficult for the public to access. Field methods and data preparation Surveys were conducted by private contractors in accordance with BLM–suggested protocols (Powder River Basin Wildlife Taskforce, 2005). Field personnel visually surveyed for new nests near proposed development sites each year from 2003 to 2011 (Andersen, 2007). Once nests were located, they were surveyed at least once between 15 April and 15 June during the year of discovery for signs of nesting activity within the nesting season (e.g., presence of adult, eggs or juveniles in nest, evidence of depredated nest contents, or evidence of abandoned eggs). Most unoccupied nests were visited twice each year to verify their activity status. Surveys were conducted by viewing the nest from the ground with optics, taking care not to disturb adults or juveniles in the nest (Steenhof and Newton, 2007). If a proposed well site had an active raptor nest very close by, such that biologists felt it could be influenced by daily activities associated with the well, the proposed well was often relocated (B. Ostheimer, personal communication). Only a portion (~70%) of nests documented in previous years were surveyed in any given subsequent year through 2011. Not all planned CBM development took place; therefore, our sample included nests that had active CBM wells nearby and those that did not, thereby facilitating a comparison between nest–use rates at developed and undeveloped sites. The construction and decommission of CBM wells occurred throughout the duration of our study; therefore, nests changed development group assignments over the course of the study. To categorize nests into undeveloped and development groups annually, we obtained point locations of CBM wells (Wyoming Oil and Gas Conservation Commission, 2012). Each CBM well had a recorded date of construction; however, reliable information regarding the lifespan and removal of individual CBM wells was unavailable. We therefore assumed a 10–year lifespan for each CBM well in our study, as a conservative estimate of the production lifespan of individual CBM wells (International Energy Agency, 2012; De Bruin et al., 2013; Riazi and Gupta, 2016) and defined 'active wells' as those constructed within the last 10 years. After a CBM well is decommissioned,

very little human disturbance takes place at the well pad (B. Ostheimer, personal communication). We calculated the distance from each nest to the nearest active CBM well annually using the sp (Pebesma and Bivand, 2005), rgdal (Bivand et al., 2016), and rgeos (Bivand and Rundel, 2016) packages in Program R (R Core Team, 2016). We considered nests ≤ 805 m from the nearest active CBM well to be nests at developed sites, and those > 805 m from the nearest active well to be nests at undeveloped sites. The 805–m threshold is reflective of the survey methods used to locate nests and thus maintained conformity between data collection and analytical methods. We assigned each nest site to one raptor species and excluded all nests from analysis that were used by more than one species during our study period. Statistical analysis We treated each nest as an independent sampling unit, after an examination of sample variograms (Pebesma, 2004) suggested a lack of spatial autocorrelation in nest–use across a range of nest proximities for the most abundant species, red–tailed hawk (Buteo jamaicensis). We constructed a nine–year nest–use history for each nest spanning 2003–2011, classifying each nest as occupied, unoccupied, or not surveyed for each year based on survey data. Nests in the dataset were not consistently surveyed multiple times within a year, which limited our ability to assess detection probability and utilize multi–season occupancy modeling, a common method used to estimate yearly occupancy rates corrected for imperfect detection (MacKenzie et al., 2006). Instead, we calculated an annual nest–use rate as the number of nests observed in use divided by the number of nests surveyed for each year. We conducted all analyses separately for each raptor species using Program R (R Core Team, 2016). Overall trends in nest–use rates We first summarized all nest data without regard to nearby CBM development. We calculated the yearly proportion of nests in use by species across the nine–year study to examine long–term trends and interspecific differences in nest–use rates. We generated 95% confidence intervals (CIs) for the yearly proportion of nests in use using a bootstrapping routine with 1,000 iterations, where nests in each group (e.g., all bald eagle, Haliaeetus leucocephalus, nests surveyed in 2003) were resampled with replacement (Manly, 1997; Carlisle and Albeke, 2016). No CIs appear in figures when there was no variation in nest use within the sample. We also summarized each species’ nine annual nest–use rates into an overall mean nest–use rate using a weighted mean, where weights were proportional to the number of nests surveyed in that year. This had the effect of treating each observation of a nest as a datum, rather than the observed proportion in a year as such. For instance, in a two–year study with 40 of 100 sites used in year one and both of two sites used in year two, an unweighted average of the two proportions would

Animal Biodiversity and Conservation 41.2 (2018)

231

Raptor nests

CBM wells

Sheridan

Buffalo

Gillette

Zoomed extent

0 25 50 km

Buffalo

Gillette

Coalbed methane well Major city

WY 4

Raptor nest (developed)

8 km

Raptor nest (undeveloped)

Major road

Fig. 1. Map of the Powder River Basin, Wyoming study area. Raptor nests of 12 species are shown (n = 3,074) as well as coalbed methane wells (n = 28,786). The inset map shows the proximity and distribution of nests at undeveloped sites (white circles) and developed sites (black circles) relative to wells (grey exes) within a zoomed extent. Where wells are highly dense, exes overlap each other and appear as grey polygons. Fig. 1. Mapa de la zona de estudio en la cuenca del río Powder, en Wyoming. Se indican los nidos de 12 especies de rapaces (n = 3.074) y los pozos de metano en capas de carbón (n = 28.786). En el mapa ampliado se muestran la proximidad de los nidos a las zonas no extractivas (círculos blancos) y a las zonas extractivas (círculos negros) y su distribución en relación con los pozos (cruces grises). Donde los pozos están densamente distribuidos, las cruces se superponen entre sí y parecen polígonos grises.

yield an estimate of (0.4 + 1.0) / 2 = 0.7. Weighting by the number surveyed each year effectively reflected that there were 102 observations, of which 42 (41.2 %) were occupied. Nest use relative to CBM development We employed a control–impact statistical design to test for differences in nest use between nests at undeveloped and developed sites. For each species, we calculated this difference separately within each year, then summarized the annual differences into one overall nine–year measure. Because many nests did not include observations both before and after the construction of a nearby well, we lacked the temporal replication required to utilize a paired design (e.g., before–after–control–impact; Gotelli and Ellison, 2004). For each species, we compared the yearly proportion of nests in use between nests at developed and undeveloped sites across the nine–year study to examine whether nest–use rates were related to the

proximity of CBM development. To do so, we constructed a nine–year development–group history for each nest spanning the same period and assigning each nest to the developed or undeveloped group for each year based on CBM well locations and assumptions about well lifespan and an initial distance threshold of 805 m. Therefore, each nest was represented in our dataset as a nine–year nest–use history, a nine–year development–group history, and a species. To summarize the CBM effect, we first calculated the difference in nest use between development groups (undeveloped–developed; therefore, positive values indicated higher nest use at undeveloped sites relative to developed sites) for each year, and then summarized those differences into one overall species–level measure using a weighted mean of the yearly differences. Because the number of nests within each development group varied from year to year, we weighted each year’s contribution to the overall mean difference, where weights were proportional to the inverse of the variance of the difference

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between development group sample sizes. This was analogous to weighted least squares regression (Ramsey and Shafer, 2013), since sample variances are at least approximately inversely proportional to sample sizes. Thus for year i, the weight (w) was calculated using the following equation, where subscripts d and u stand for nests at developed and undeveloped sites, respectively: wi =

nid niu nid + niu

i=1

1 nid niu 1 = V nid + niu V =

nid niu

nid + niu

i=1

nid niu

nid + niu nid niu

i=1

nid + niu

nid niu

nid + niu

We developed a randomization test to assess whether rates of nest use differed between development groups for each species. A randomization test is a non–parametric method for null hypothesis testing that was well suited to our study because the test can be adapted to accommodate non–standard test statistics, does not rely on parametric assumptions, and does not require random sampling of the population (Manly, 1997). The procedure of our randomization test mimicked the design and implementation of the study and employed a control–impact statistical design that treated each nest as a sampling unit, with each species analyzed separately. The null hypothesis was that development (i.e., being near a CBM well) had no effect on nest–use rates. We used 1,000 iterations of the following steps to simulate the repetition of the study where the null hypothesis was true, thus generating the null distribution against which to test our observed statistic (nine–year average difference in nest use between nests at undeveloped and developed sites). The principal mechanism for each iteration of the randomization test was reassigning each nest to a random development group, implying that if the null hypothesis were true, the assignment to development groups of each nest was meaningless. We first selected all nests that were surveyed in the first year of the study and randomized their development group assignments, maintaining the original, relative balance between the numbers of nests within each development group within that year. We then repeated that process for each year of the study. Because nests moved between development groups during the study period as CBM wells were built and decommissioned, we did not require nests in the randomization procedure to retain their development assignments across years. After randomizing development group assignments, we recalculated the statistic of interest (nine–year average difference in nest use between nests at undeveloped and developed sites) as previously described. All randomization tests were two–tailed and used P < 0.05 as the criterion to reject the null hypothesis.

Table 1. Sample size (n) of nests for each species of raptor documented nesting near planned areas of coalbed methane development in the Powder River Basin, Wyoming between 2003–2011: * species with fewer than 15 nests (excluded from analysis due to small sample size). Tabla 1. Tamaño de la muestra (n) de nidos para cada especie de rapaz que se haya documentado nidificando cerca de yacimientos de metano en capas de carbón en la cuenca del río Powder, en Wyoming, entre los años 2003 y 2011: * especies con menos de 15 nidos (se excluyeron del análisis debido al reducido tamaño de la muestra). Common name

Scientific name

Turkey vulture*

Cathartes aura 2

Bald eagle

Haliaeetus

leucocephalus 28 Northern harrier

Circus cyaneus 28

Sharp–shinned hawk* Accipiter striatus

Cooper’s hawk

Accipiter cooperii

Swainson’s hawk

Buteo swainsoni 90

Red–tailed hawk

Buteo jamaicensis 1,046

Ferruginous hawk

Buteo regalis 933

Golden eagle

Aquila chrysaetos 283

Barn owl*

Tyto alba 1

Great horned owl

Bubo virginianus 286

Burrowing owl

Athene cunicularia 139

Long–eared owl

Asio otus 87

Short–eared owl*

Asio flammeus 9

American kestrel

Falco sparverius 108

Merlin*

Falco columbarius 3

Peregrine falcon*

Falco peregrinus 4

Prairie falcon

Falco mexicanus 30

Alternative distance thresholds The 805–m threshold dividing nests at developed from undeveloped sites was dictated largely by the BLM’s protocols for nest searching, and most nest observations (73.5% when pooled across years and species) were from nests within 805 m of a CBM well (table 1s). The dataset did, however, include some nests farther from CBM development (likely because planned developments did not all take place), especially for abundant species. We therefore repeated the analysis using three additional distance thresholds to test whether our comparisons of nest use rates by development group were sensitive to the distance used to classify nests as either at a

Animal Biodiversity and Conservation 41.2 (2018)

Bald eagle

233

Northern harrier

Cooper’s hawk

Swainson’s hawk

Great horned owl

1.0

Red–tailed hawk

Ferruginous hawk

Golden eagle

Burrowing owl

Long–eared owl

American kestrel

2003 2004 2005 2006 2007 2008 2009 2010 2011

0.0

2003 2004 2005 2006 2007 2008 2009 2010 2011

Proportion of nests in use

0.5

1.0 0.5 0.0 Prairie falcon

1.0 0.5

Year

2003 2004 2005 2006 2007 2008 2009 2010 2011

0.0

Fig. 2. The annual proportion (± 95 % CI) of raptor nests in use from 2003–2011 in the Powder River Basin, Wyoming, USA. Fig. 2. Proporción anual (± 95 % IC) de nidos de rapaces en uso entre 2003 y 2011 en la cuenca del río Powder, en Wyoming, EE.UU.

developed or undeveloped site. These additional distance criteria were approximately half, twice, and three times as large as the original distance criteria (400 m, 1,610 and 2,415 m, respectively). Because most nests monitored were near CBM development, not all species had nests far enough from CBM wells to have nests in the 'undeveloped' category when the larger distance criteria were applied. Therefore, we restricted this portion of the analysis to the four species for which there were nests in both development groups at even the 2,415 m distance criteria (i.e. red–tailed hawk; ferruginous hawk, Buteo regalis; golden eagle, Aquila chrysaetos; and great horned owl, Bubo virginianus). Results We analyzed 12,977 observations from 3,074 raptor nests of 12 species conducted across nine years (table 1). The four most prevalent species (red–tailed hawk, ferruginous hawk, golden eagle, and great horned owl) accounted for the vast majority (82.9 %, n = 2,548) of nests. When pooled across years and the

four most abundant species, 54.4 % of observations were from nests within 400 m of a CBM well, 88.8% within 1,610 m, and 94.4% within 2,415 m. Based on our assumption of well lifespan, there were ~14,000 active wells in the PRB when the study began in 2003. Wells were constructed at a fairly constant rate between 2003 and 2009, when the number of active wells peaked at ~26,000. Well counts then declined to ~21,000 active wells in 2011 (fig. 1s). Annual trends in nest–use rates The mean proportion of nests in use varied from year to year. Trends in nest use were non–liner and cyclical in appearance, especially for those species with larger sample sizes (fig. 2). Trends were fairly consistent across species; for most, the peak of nest use was in 2005 or 2006, and the lowest between 2009 and 2011 (fig. 2). The average proportion of nests in use varied across species, as did the magnitude of changes in use from year to year (fig. 2). Bald eagles had the highest overall average use (63.6 %), whereas ferruginous hawks had the lowest (8.2 %). All other species averages ranged from 19.5–42.6 % (table 2s).

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Bald eagle

Northern harrier

Cooper’s hawk

Swainson’s hawk

Red–tailed hawk

Ferruginous hawk

Golden eagle

Great horned owl

Burrowing owl

Long–eared owl

American kestrel

2003 2004 2005 2006 2007 2008 2009 2010 2011

Developed

2003 2004 2005 2006 2007 2008 2009 2010 2011

Undeveloped

1.0

Proportion of nests in use

0.5 0.0 1.0 0.5 0.0 Prairie falcon

1.0 0.5

2003 2004 2005 2006 2007 2008 2009 2010 2011

Year

2003 2004 2005 2006 2007 2008 2009 2010 2011

0.0

Fig. 3. The annual proportion (± 95 % CI) of raptor nests in use, by development group, from 2003–2011 in the Powder River Basin, Wyoming, USA. Nests at developed sites were those ≤ 805 m from the nearest active coalbed methane well, and nests at undeveloped sites were those > 805 m from the nearest active coalbed methane well. Fig. 3. Proporción anual (± 95 % IC) de nidos de rapaces en uso, por grupo de extracción, entre 2003 y 2011 en la cuenca del río Powder, en Wyoming, EE.UU. Los nidos en zonas extractivas se encontraban a ≤ 805 m del pozo activo de metano en capas de carbón más cercano, mientras que los nidos en zonas no extractivas se encontraban a > 805 m del pozo activo de metano en capas de carbón más cercano.

Nest use relative to CBM development Trends in nest use were similar between nests at undeveloped and developed sites for most species (fig. 3). Nests at undeveloped sites had higher use than nests at developed sites for red–tailed hawks (effect size = 5.1 %, P < 0.01), burrowing owls (Athene cunicularia, 11.5 %, P = 0.02), and long–eared owls (Asio otus, 9.5 %, P = 0.02; fig. 4). Of the remaining nine species, differences in nest use between development groups were equivocal at the α = 0.05 level (fig. 4). We likely had minimal statistical power to detect any effect of CBM development on nest use for rarer species, however, which had smaller sample sizes. Alternative distance thresholds Differences in nest–use rates between nests at undeveloped and developed sites did not vary across

the four different distance–to–development thresholds for ferruginous hawk, golden eagle, and great horned owl (fig. 5). For red–tailed hawks, the direction of the difference was the same across threshold distances (i.e., higher use at undeveloped sites relative to developed sites), and the effect size generally increased with the distance threshold. For all species, there was substantial overlap among the 95 % CIs for the effect sizes across scales (fig. 5). Discussion One growing form of human–induced habitat change with largely unknown consequences to breeding birds is extraction for energy resources (Northrup and Wittemyer, 2013; Donázar et al., 2016). Raptor species may be particularly vulnerable to energy development due to their often large area requirements (Watson

Species

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Bald eagle Northern harrier Cooper’s hawk Swainson’s hawk Red–tailed hawk Ferruginous hawk Golden eagle Great horned owl Burrowing owl Long–eared owl American kestrel Prairie falcon –0.3

–0.2

–0.1 0.0 0.1 0.2 0.3 Difference in proportion of nests in use (undeveloped–developed)

Fig. 4. The nine–year average difference (± 95 % CI) in the annual proportion of raptor nests in use from 2003–2011 in the Powder River Basin, Wyoming, USA. The difference was calculated as undeveloped – developed, so positive values indicate higher use at undeveloped sites relative to developed sites. Nests at developed sites were those ≤ 805 m from the nearest active coalbed methane well, and nests at undeveloped sites were those > 805 m from the nearest active coalbed methane well. Fig. 4. Diferencia media de los nueve años (± 95 % IC) en la proporción anual de nidos de rapaces en uso entre 2003 y 2011 en la cuenca del río Powder, en Wyoming, EE.UU. La diferencia se calculó como zonas no extractivas – zonas extractivas, de tal forma que los valores positivos indican un mayor uso en las primeras en comparación con las últimas. Los nidos en zonas extractivas se encontraban a ≤ 805 m del pozo activo de metano en capas de carbón más cercano, mientras que los nidos en zonas no extractivas se encontraban a > 805 m del pozo activo de metano en capas de carbón más cercano.

et al., 2014; Crandall et al., 2015), relatively slow life histories (Bennett and Owens, 2002), and demonstrated sensitivity to human disturbance and habitat alteration (White and Thurow, 1985; Kostrzewa, 1996; Krüger, 2002; Martinez–Abraín et al., 2010; Brown et al., 2014; Coates et al., 2014). We leveraged a large dataset to evaluate temporal patterns of nest–site use by 12 species of raptors in relation to the proximity of CBM development in Wyoming, USA. The mean rate of nest use varied annually in an apparently non–linear manner for all 12 species. Three species (i.e. red–tailed hawk, burrowing owl, and long–eared owl) were significantly more likely to use nests away from CBM wells. The red–tailed hawk result was somewhat surprising given that this species is considered to be one of the more disturbance–tolerant raptor species (Berry et al., 1998; Hobbs et al., 2006; Coates et al., 2014; Duerr et al., 2015). Similarly, burrowing owls tend not to significantly alter behaviors in relation to human disturbance or land use type (Plumpton and Lutz, 1993; Chipman et al., 2008). Different types of, or distances to, human disturbance, however, may elicit varying responses by wildlife, and to our knowledge, no other published study has examined the responses of red–tailed hawks or burrowing owls to coalbed–methane development. Long–eared owls

are known to be sensitive to development–related habitat loss and to human disturbance at nest and roost sites (Marks et al., 1994). Additionally, long–eared owls tend to prefer nesting areas containing fewer paved roads (Martínez and Zuberogoitia, 2004). Our work further confirms the sensitivity of long–eared owls to human activity and extends those activity types to include CBM development. Collectively, our results suggest unique responses of raptor species to energy development (also see Smith et al., 2010), and taken in the context of previous work, that behavioral responses within species can vary across different environmental contexts. The suite of species examined in our study displayed wide diversity in their traits, including morphological (e.g., body size), behavioral (e.g., foraging strategies and activity periods), and natural history (e.g., home range size, specific nesting requirements). Differing species–specific responses of raptors nesting in relation to nearby CBM development were therefore not surprising (also see Martinez–Abrain et al., 2010). One potential reason for the weak relationship between nest–site use and CBM development for many species is that suitable nest sites can be limiting (Newton, 1998) and the construction of the large nests often built by raptors

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Distance threshold (m):

400

805

1,610

2,415

Species

Red–tailed hawk Ferruginous hawk Golden eagle Great horned owl –0.15

–0.10

–0.05 0.00 0.05 0.10 0.15 Difference in proportion of nests in use (undeveloped–developed)

Fig. 5. The nine–year average difference (± 95 % CI) in the annual proportion of raptor nests in use from 2003–2011 in the Powder River Basin, Wyoming, USA across multiple scales of analysis (i.e., the minimum distance a nest at an undeveloped site could be from an active coalbed methane well). The difference was calculated as undeveloped – developed, so positive values indicate higher use at undeveloped sites relative to developed sites. Nests at developed sites were those ≤ the specified distance from the nearest active coalbed methane well, and nests at undeveloped sites were those > the specified distance from the nearest active coalbed methane well. Fig. 5. Diferencia media de los nueve años (± 95 % IC) en la proporción anual de nidos de rapaces en uso entre 2003 y 2011 en la cuenca del río Powder, en Wyoming, EE.UU. en múltiples escalas de análisis (esto es, la distancia mínima de un pozo activo de metano en capas de carbón). La diferencia se calculó como zonas no extractivas – zonas extractivas, de tal forma que los valores positivos indican un mayor uso en las primeras en comparación con las últimas. Los nidos en zonas extractivas se encontraban a ≤ la distancia especificada del pozo activo de metano en capas de carbón más cercano, mientras que los nidos en zonas no extractivas se encontraban a > la distancia especificada del pozo activo de metano en capas de carbón más cercano.

is a significant energetic investment (Moller and Nielsen, 2015). Birds would therefore have to weigh the potential costs of remaining faithful to nest sites in disturbed areas versus the costs of locating or building a new nest structure. If the actual fitness costs of nest–site use in developed areas are low, parent birds would likely remain site–faithful to those territories and nests, but assessing fitness measures (i.e. survival and reproductive success) was outside the scope of this study. A critical next step for future work is to evaluate actual habitat preferences and fitness outcomes in relation to energy development infrastructure (Hethcoat and Chalfoun, 2015b). Otherwise, one cannot discern whether nest–site use of birds in areas of energy development is an adaptive response versus an ecological trap (Robertson and Hutto, 2006). One of the major impetuses for this study was to determine whether the 805–m buffers implemented by the BLM were biologically meaningful in terms of raptor responses and sufficient for protective measures. Our analysis of distance thresholds entailed approximately halving, doubling,

and tripling the 805–m radius, which did not result in significant differences in the direction or size of the observed effect for any of the four species examined. These results suggest that the 805–m radius may be sufficient for limiting avoidance of areas with nearby CBM development by raptors for nesting, but more targeted study of actual nesting productivity in relation to CBM should be conducted for confirmation. If the temporal patterns in nest–site use in our study reflected actual population trends (Sergio and Newton, 2003; Kennedy et al., 2014), there were likely key factors affecting local raptor populations independent of potential energy–development effects. Raptor occupancy, site fidelity, and population trends are known to be associated with local prey availability (Smith et al., 1981; MacLaren, et al. 1988; Woffinden and Murphy, 1989; Kostrzewa, 1996; Kochert and Steenhof, 2002; Sergio et al., 2006; Millsap et al., 2015) which varies temporally and spatially (Fedy and Doherty, 2011; Simes et al., 2015). Raptor prey deficits can be exacerbated by exogenous factors such as

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drought (Ranta et al., 1999); and 2004 was a relatively dry year (U.S. Geological Survey, 2017), which may have precipitated subsequent predator responses with a one to two–year time lag (e.g., Lehikoinen et al., 2011). Future work designed to assess the influence of energy development on raptor populations should therefore account for other key population drivers such as food availability and weather (Steenhof et al., 1997). Some rodent species, especially generalists such as deer mice (Peromyscus maniculatus), can increase in abundance around energy development (Hethcoat and Chalfoun, 2015b), which may actually benefit some raptor species. Finally, the lowest rates of nest–site use were generally observed during the last several years of the study (2009–2011, fig. 2). Therefore, continued monitoring of raptor populations in this region would likely be of particular interest to conservation practitioners and may shed light on the population cycling suggested by our results and which has been documented for some raptor prey in Wyoming (Fedy and Doherty, 2011). We acknowledge some important caveats and limitations of our study. Some raptor species, such as golden eagles and ferruginous hawks, maintain several potential nest sites within their territory among which they can rotate in different years (Kochert and Steenhof, 2002; Smith et al., 2010; Millsap et al., 2015). Surveyors in our study monitored nest sites and not entire nesting territories, which means that our nest–use rates were likely consistently lower than actual territory–use rates for species with multiple nests per territory. Additionally, two of the three species (red–tailed hawk and burrowing owl) for which we observed a negative association between nest–use rates and CBM development were those with some of the highest sample sizes of nests. Sample size may therefore have played a role in our ability to detect effects for the rarer species. Density dependence can have strong effects on bird populations, and fluctuations in abundance or nest–site use over time and space can be indicative of density–dependent regulation (Newton, 1998). The breeding–season abundance and distribution of raptors in particular can be limited by a lack of suitable nest sites, especially in open habitats with few natural structures (e.g., cliffs and trees; Steenhof et al., 1993; Newton, 1998). Indeed, raptors have colonized previously unsuitable areas after artificial structures suitable for nesting (such as those associated with energy development and transmission lines) were installed (Steenhof et al., 1993; Newton, 1998). The colonization of newly suitable areas can therefore be a manifestation of density dependence, as individuals escape intra– or inter–specific competition for resources (Newton, 1998). We observed fluctuations in nest–use rates through time, but we lacked information regarding key factors such as available nest substrates, prey densities, and densities of co–existing raptors, with which to assess potential density–dependent responses (Newton, 1998). In summary, we documented non–linear annual patterns of nest–site use for 12 species of raptors in the Powder River Basin, Wyoming, USA, an arid basin

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heavily influenced by coalbed–methane development. Three species displayed potential avoidance of nesting within 805 m of CBM development, though the mechanisms for these patterns remain unclear. Future work that clarifies actual nesting habitat preferences and outcomes in relation to energy development, while simultaneously accounting for local prey availability, densities of competitors, and weather, would be highly beneficial. Disentangling the specific attributes of energy extraction (e.g., habitat change, noise, movement, artificial light) that elicit wildlife responses would also be a particularly fruitful line of inquiry. Acknowledgments We thank the employees of the BLM, other agencies, and private firms who collected and compiled these data, with special thanks to B. Ostheimer and D. Stafford. We thank S. Albeke for R programming advice. This study and manuscript were improved by thoughtful comments from B. Ostheimer, P. Stouffer, A. Davis, and several anonymous reviewers. Any use of trade, product or firm names is for descriptive purposes only and does not imply endorsement by the U. S. Government. References Acevedo–Whitehouse, K., Duffus, A. L. J., 2009. Effects of environmental change on wildlife health. Philosophical Transactions of the Royal Society B, 364: 3429–3438. Allred, B. W., Smith, W. K., Twidwell, D., Haggerty, J. H., Running, S. W., Naugle, D. E., Fuhlendorf, S. D., 2015. Ecosystem services lost to oil and gas in North America. Science, 348: 401–402. Andersen, D. E., 2007. Survey techniques. In: Raptor Research and Management Techniques: 89–100 (D. M. Bird, K. L. Bildstein, Eds.). Hancock House Publishers, Blaine, WA, USA. Bart, J., Fuller, M., Jacobs, R., Kitchell, K., McCluskey, C., Mills, D., Suring, L., Whittington, D., 2006. Balancing energy development and raptor conservation in the western United States: Workshop summary report. Available at: http://northern–ecologic.com/publications/25.pdf [Accessed on 18 Sept. 2017]. Bennett, P. M., Owens, I. P. F., 2002. Evolutionary ecology of birds: life histories, mating systems and extinction. Oxford University Press, Oxford. Berry, M. E., Bock, C. E., Haire, S. L., 1998. Abundance of diurnal raptors on open space grasslands in an urbanized landscape. The Condor, 100: 601–608. Bivand, R. S., Keitt, T., Rowlingson, B., 2016. rgdal: bindings for the geospatial data abstraction library. R package version 1.1–8. Available at: https:// cran.r–project.org/package=rgdal [Accessed on 18 Sept. 2017]. Bivand, R. S., Rundel, C. W., 2016. rgeos: Interface to geometry engine–open source (GEOS). R package

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Supplementary material

Coalbed methane wells

25,000 20,000 15,000 10,000 5,000

2011

2010

2009

Year

2008

2007

2006

2005

2004

2003

Fig. 1s. The number of active coalbed methane wells from 2003–2011 in the Powder River Basin, Wyoming. These data assume a 10–year lifespan for each well. Fig. 1s. Número de pozos activos de metano en capas de carbón entre 2003 y 2011 en la cuenca del río Powder, en Wyoming. Estos datos suponen una vida útil de 10 años para cada pozo.

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Table 1s. Sample sizes for each species–year–development group combination. Developed nests were those ≤ 805 m from the nearest active coalbed methane well, and undeveloped nests were those > 805 m from the nearest active coalbed methane well: Y, year; Ud, undeveloped; D, developed; T, total. Tabla 1s. Tamaños de muestra para cada combinación del grupo especie–año–actividad extractiva. Los nidos en zonas extractivas se encontraban a ≤ 805 m del pozo activo de metano en capas de carbón más cercano, mientras que los nidos en zonas no extractivas se encontraban a > 805 m del pozo activo de metano en capas de carbón más cercano. (Para las abreviaturas, véase arriba). Common name

Common name

Bald eagle

2003

Red–tailed hawk

2003

Bald eagle

2004

Red–tailed hawk

2004 173 147 320

45 66

Bald eagle

2005

Red–tailed hawk

2005 128 186 314

Bald eagle

2006

Red–tailed hawk

2006 182 295 477

Bald eagle

2007

Red–tailed hawk

2007 194 402 596

Bald eagle

2008

Red–tailed hawk

2008 126 470 596

Bald eagle

2009

Red–tailed hawk

2009 135 527 662

Bald eagle

2010

Red–tailed hawk

2010 154 594 748

Bald eagle

2011

Red–tailed hawk

2011 137 645 782

Northern harrier

2003

Ferruginous hawk

2003

204 272

Northern harrier

2004

Ferruginous hawk

2004

270 356

Northern harrier

2005

Ferruginous hawk

2005

267 313

Northern harrier

2006

Ferruginous hawk

2006

305 378

Northern harrier

2007

18 22

Ferruginous hawk

2007 115 403 518

Northern harrier

2008

16 18

Ferruginous hawk

2008

Northern harrier

2009

16 20

Ferruginous hawk

2009 119 391 510

Northern harrier

2010

16 21

Ferruginous hawk

2010 107 370 477

Northern harrier

2011

18 21

Ferruginous hawk

2011 101 420 521

Cooper's hawk

2003

Golden eagle

2003

14 21

Cooper's hawk

2004

Golden eagle

2004

36 82

Cooper's hawk

2005

Golden eagle

2005

50 86

Cooper's hawk

2006

Golden eagle

2006

90 147

Cooper's hawk

2007

Golden eagle

2007

110 172

Cooper's hawk

2008

Golden eagle

2008

124 170

Cooper's hawk

2009

Golden eagle

2009

140 185

Cooper's hawk

2010

Golden eagle

2010

158 210

Cooper's hawk

2011

Golden eagle

2011

162 207

Swainson's hawk

2003

17 20

Great horned owl

2003

Swainson's hawk

2004

30 41

Great horned owl

2004

27 56

Swainson's hawk

2005

31 33

Great horned owl

2005

41 68

Swainson's hawk

2006

38 40

Great horned owl

2006

68 128

Swainson's hawk

2007

44 47

Great horned owl

2007

93 153

Swainson's hawk

2008

30 31

Great horned owl

2008

116 177

Swainson's hawk

2009

33 37

Great horned owl

2009

165 214

Swainson's hawk

2010

29 31

Great horned owl

2010

171 224

Swainson's hawk

2011

24 32

Great horned owl

2011

190 232

317 409

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Table 1s. (Cont.)

Common name

Burrowing owl

2003

American kestrel

2003

Burrowing owl

2004

American kestrel

2004

Burrowing owl

2005

21 37

American kestrel

2005

16 24

Burrowing owl

2006

32 48

American kestrel

2006

34 46

Burrowing owl

2007

49 60

American kestrel

2007

40 61

Burrowing owl

2008

42 54

American kestrel

2008

47 64

Burrowing owl

2009

71 82

American kestrel

2009

55 77

Burrowing owl

2010

77 97

American kestrel

2010

57 80

Burrowing owl

2011

78 99

American kestrel

2011

70 86

Long–eared owl

2003

Prairie falcon

2003

Long–eared owl

2004

Prairie falcon

2004

Long–eared owl

2005

12 21

Prairie falcon

2005

Long–eared owl

2006

28 50

Prairie falcon

2006

Long–eared owl

2007

35 52

Prairie falcon

2007

10 19

Long–eared owl

2008

41 59

Prairie falcon

2008

10 18

Long–eared owl

2009

49 69

Prairie falcon

2009

12 23

Long–eared owl

2010

54 68

Prairie falcon

2010

12 23

Long–eared owl

2011

62 81

Prairie falcon

2011

16 28

Table 2s. The weighted mean proportion of nests in use for 12 raptor species nesting in the Powder River Basin, Wyoming between 2003–2011. Tabla 2s. Proporción media ponderada de nidos en uso para 12 especies de rapaces que anidan en la cuenca del río Powder, en Wyoming, entre 2003 y 2011. Common name

Scientific name

Nests in use

Bald eagle

Haliaeetus leucocephalus

63.6 %

Northern harrier

Circus cyaneus

19.5 %

Cooper’s hawk

Accipiter cooperii

27.1 %

Swainson’s hawk

Buteo swainsoni

42.6 %

Red-tailed hawk

Buteo jamaicensis

33.2 %

Ferruginous hawk

Buteo regalis

8.2 %

Golden eagle

Aquila chrysaetos

33.7 %

Great horned owl

Bubo virginianus

28.6 %

Burrowing owl

Athene cunicularia

27.9 %

Long–eared owl

Asio otus

26.2 %

American kestrel

Falco sparverius

32.7 %

Prairie falcon

Falco mexicanus

39.3 %

244

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Efficient vs. structured biodiversity inventories: reptiles in a Mexican dry scrubland as a case study S. Trujillo–Caballero, J. A. González–Oreja

Trujillo–Caballero, S., González–Oreja, J. A., 2018. Efficient vs. structured biodiversity inventories: reptiles in a Mexican dry scrubland as a case study. Animal Biodiversity and Conservation, 41.2: 245–256. Abstract Efficient vs. structured biodiversity inventories: reptiles in a Mexican dry scrubland as a case study. Many sampling methods allow the study of species richness and diversity in biological communities, but it is not known whether a single method can determine both the number and diversity of species in an unbiased and efficient way. Here we assess whether the least biased and most efficient method to determine reptile species richness in a Mexican dry scrubland is also the best method to estimate species diversity. The local assemblage was composed of 10 species, with the Mexican mud turtle (Kinosterton integrum) and the Jalapa spiny lizard (Sceloropus jalapae) being the dominant ones. Microhabitat surveys (MHS) were the most accurate and the most efficient method to estimate species richness, but they over–estimated species diversity (+67.1 %) as much as the other sampling methods, i.e., transect surveys and pitfall–trap stations, under–estimated it (–59 %). Our study shows that the best sampling method to determine the number of species in local assemblages may not be the best method to study species diversity. Although combining different sampling methods can increase the project costs in terms of time, effort and money, the use of structured inventories is recommended for the analysis of species diversity. Key words: Biodiversity knowledge, Number of species, Number of equiprobable species, Sampling methods, Sampling effort, Pitman efficiency Resumen Comparación entre inventarios de biodiversidad eficientes y estructurados: los reptiles de un matorral xerófilo de México como ejemplo. Se han propuesto muchos métodos para estudiar la riqueza y la diversidad de especies en comunidades biológicas, pero se desconoce si existe alguno que pueda determinar tanto el número como la diversidad de especies sin sesgo y de forma eficiente. En este estudio evaluamos si el método menos sesgado y más eficiente para determinar la riqueza de reptiles en un matorral xerófilo de México es también el mejor para estimar la diversidad de especies. La comunidad local estaba compuesta por 10 especies, de las que la tortuga de pecho quebrado y pata rugosa (Kinosterton integrum) y la lagartija escamosa jalapeña (Sceloropus jalapae) eran las dominantes. Los muestreos en microhábitats (MHS) fueron el método más exacto y eficiente para estimar la riqueza de especies, pero sobrestimaron (+67,1 %) la diversidad de especies tanto como la subestimaron (–59 %) los demás métodos (i.e., los itinerarios y las estaciones de trampas de caída). Nuestro estudio muestra que el mejor método de muestreo para determinar el número de especies en las comunidades locales tal vez no sea el mejor para estudiar la diversidad de especies. Aunque combinar diferentes métodos de muestreo puede aumentar el tiempo, el esfuerzo y los costos económicos relacionados con los proyectos, recomendamos el uso de inventarios estructurados para analizar la diversidad de especies. Palabras clave: Conocimiento de la biodiversidad, Número de especies, Número de especies equiprobables, Métodos de muestreo, Esfuerzo de muestreo, Eficiencia de Pitman Received: 17 V 17; Conditional acceptance: 19 X 17; Final acceptance: 23 X 17 Saúl Trujillo–Caballero, José Antonio González–Oreja, Fac. de Biología. BUAP, Benemérita Univ. Autónoma de Puebla, Edificio 112–A, Ciudad Universitaria, 72570 Puebla, México. Corresponding author: José Antonio González–Oreja. E–mail: jgonzorj@hotmail.com ISSN: 1578–665 X eISSN: 2014–928 X

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Introduction Historically, the concept of biodiversity has been nearly equivalent to the number of species in a local assemblage (Sarkar, 2002; Tucker, 2005; Leitner and Turner, 2013). Knowing species richness is a legitimate objective in ecology and conservation biology (Colwell and Coddington, 1994; Gotelli and Colwell, 2001; Leitner and Turner, 2013), but equality between biodiversity and species richness is inaccurate (Swingland, 2013). In addition to the total number of species, the measurement of biodiversity should include information about the relative importance (equitability) of the species in the studied assemblage (Gotelli and Colwell, 2011). Fortunately, the toolbox of the modern biodiversity student includes many measures that consider species equitability (Magurran, 2005; Jost and González–Oreja, 2012; Gotelli and Chao, 2013). Many factors can influence the measurement of species richness and diversity (Hill et al., 2005; Magurran, 2005), such as sampling effort (Colwell and Coddington, 1994; Gotelli and Colwell, 2001; Gotelli and Chao, 2013) and the effects that the sampling method has on the fraction of biodiversity sampled (Hill et al., 2005; Sutherland, 2006; Eekhout, 2010; McDiarmid et al., 2012). To increase the set of truly sampled species and to obtain a more accurate estimation of the total richness, biodiversity students can combine (i.e., add up) data obtained with different sampling methods. This practice is known as a structured inventory (Gotelli and Ellison, 2013) and it has frequently been applied in biodiversity studies (for instance, see King and Porter, 2005; Coddington et al., 2009; Gotelli et al., 2011, or Castro et al., 2017, for arthropods; Jenkins et al., 2003; Hutchens and DePerno, 2009; Sung et al., 2011; Foster, 2012;, or Carpio et al., 2015, for amphibians and reptiles; or Pech Canche et al., 2011, for bats). Nevertheless, because of systematic biases in favor of or against certain activity periods, animal behaviors or body sizes (Hutchens and DePerno, 2009), the different sampling methods applied in a structured inventory can differ in their probability to capture different species (Yoccoz et al., 2001). As a result, differences may arise in the relative abundance of the registered species (Longino et al., 2002; Coddington et al., 2009) and lead to contrasting estimates of species diversity obtained with varied sampling methods. Reptiles are ectothermic animals, and the study of reptile biodiversity is not only strongly influenced by their biology, physiology and behavior (Willmer et al., 2005; McDiarmid et al., 2012; Vitt and Caldwell, 2014) but also by the environmental variables which constrain them (i.e., temperature or humidity; Latham et al., 2005). All these factors can involve different capture probabilities for the same species with different sampling methods (Hutchens and DePerno, 2009; Rodda, 2012). Since the election of a good sampling method can be a critical factor in a biodiversity study (Cardoso, 2009), we asked the following question: can the best method to es-

timate species richness simultaneously be the best method to estimate species diversity? To the best of our knowledge, this is the first time this question has been posed in the biodiversity literature. In this paper, using our data on the reptile assemblage in a Mexican dry scrubland, we (i) compared the performance (in terms of bias and efficiency) of three sampling methods frequently used in reptile biodiversity studies (Eekhout, 2010; McDiarmid et al., 2012): microhabitat surveys, transect surveys, and pitfall–trap stations; and (ii) analyzed whether the most accurate and most efficient (i.e., the best) method to estimate the number of species in our local reptile assemblage was, at the same time, the best method to study a measure of diversity that considers species equitability. Material and methods Study area Field work was conducted at Tecali de Herrera (Puebla, Mexico: 18º 48' 24'' ‒ 18º 57' 54'' N, 97º 57' 54'' ‒ 98º 05' 42'' W), at an altitude of ca. 2,000. The climate is temperate (mesothermic) and subhumid (average values for 1951–2010: annual temperature, ca. 17 ºC; total annual precipitation, ca. 600 mm), and rain is mostly recorded during summer (SMN, 2015). Since insolation is high and humidity is low, evapotranspiration can reach elevated values (mean total annual evaporation for the period 1951–2010: 1,890 mm). Because of anthropic changes in land use, most of the original vegetation has disappeared (SEGOB, 2006) and the landscape is currently covered by an open, dry scrubland over diverse soil types and geological substrates (Rzedowski, 1988; Saldaña Munive, 2011). Several cacti (e.g., Echinocactus sp., Mammillaria sp.), tree–like cacti (Stenocereus sp.), and other succulent plants (like Agave salmiana, A. stricta, and Yucca periculosa) were dominant in this landscape. Finally, because of the extreme climatic conditions, crop fields were scarce, and several kilometers away from where we completed our field work. Field work From June 2005 to April 2006, we completed monthly reptile inventories using the following methods (Latham et al., 2005; Blomberg and Shine, 2006): (1) microhabitat surveys, (2) transect surveys, and (3) pitfall–trap stations. Microhabitat surveys (MHS) are the simplest method to capture small reptiles (Blomberg and Shine, 2006). We actively searched for reptiles in sites (i.e., microhabitats) that are usually preferred as refuges by the considered species (McDiarmid et al., 2012), such as under rocks, fallen logs, metal sheets and the bark of tree trunks, or between cracks. All specimens were captured by hand. The total number of MHS was 23, distributed throughout the whole study period (no less than once per month; table 1s in supplementary material).

Animal Biodiversity and Conservation 41.2 (2018)

Transect surveys (TS) consisted of two itineraries (1 km each) across the study area. These itineraries intersected several small, temporal or permanent creeks. Transect surveys were sampled at low pace, first between 8 and 9 a.m., and then between 4 and 5 p.m., 23 times in total (table 1s in supplementary material). We recorded all the species observed along each TS, as well as other data on their activity. Pitfall–trap stations (PFS) were modified from the arrangement recommended by Crosswhite et al. (1999). Each PFS consisted of four pitfall–traps (45 cm deep, with the bottom perforated to reduce the risk of drowning by rain water accumulation), plus six double–ended funnel–traps covered with a cylindrical screen (Ø = 35 cm), all connected by three drift fences (10 m long, buried in the ground to a depth of 10 cm, and with additional 50 cm above the ground). PFS, with the pitfall–traps partially covered with leaves to reduce the risk of death due to the high insolation, were open for 24 h. To limit the stress to trapped reptiles, we checked traps twice a day (morning and midday). Throughout the study, we used appropiate tools to collect and handle reptiles (e.g. rubber bands, snake hooks, nets), and identified all captured individuals by using the keys of Casas and McCoy (1979) and Flores Villela et al. (1995). We completed only 13 PFS, all from June to October 2005, because our PFS suddenly 'disappeared' from the study area. Therefore, we split the analyses into two time periods: (1) data from June to October 2005 (the initial period) allowed us to compare, with low sampling efforts, MHS, TS and PFS; (2) data from June 2005 to April 2006 (all our field work) allowed us to compare only MHS and TS, but with higher sampling efforts. Measuring reptile biodiverisity With the data obtained from all the three sampling methods from the first period, or only from MHS and TS throughout all our field work (table 1s in supplementary material), we computed the following descriptors of reptile biodiversity (Magurran, 2005). Richness (S): number of species, computed with individual–based rarefaction or extrapolation techniques, after standardization of all the methods compared to a common number of individuals captured (Gotelli and Colwell, 2001). Modern individual–based richness estimation techniques through extrapolation allowed to compute the expected number of species that could be obtained with a total abundance above the number really observed (Colwell et al., 2012; Gotelli and Chao, 2013). However, these techniques offer credible results if the total number of individuals is below 3× (or, even better, 2×) the actual total abundance (Col–well et al., 2012). Therefore, to assess the quality of the inventories in the first period, we compared the observed richness per method with the one computed by extrapolation to 164 individuals; and, in all our field work, up to 224 individuals (i.e., 2× the corresponding total abundance: n = 82 individuals for the first period, and n = 112 throughout all our field work). To compare

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richness values obtained with different numbers of individuals, we used the more traditional rarefaction technique (Gotelli and Colwell, 2001). We present values computed by rarefaction from n individuals as Sn, and those by extrapolation as S*n; for example, S35 is the number of species expected in a sample of 35 individuals, computed by rarefaction, whereas S*246 is the number of species expected in a sample of 246 individuals, computed by extrapolation. True species diversity (1D): number of equiprobable species (Jost, 2006; Jost and González–Oreja, 2012; Moreno et al., 2011), computed with the total species abundances from each sampling method (i.e., MHS vs. TS vs. PFS), or with the sum of the abundances from several, pooled methods (vg., MHS + TS): qD = [Spiq]1/(1 – q), where pi is the proportion of individuals of species i, and q is the order of the diversity measure. Since we used q = 1, the previous formula was equivalent to exp[– Spi × ln(pi)] (Jost, 2006; see Cruz Elizalde and Ramírez Bautista, 2012 for the same approach in a reptile biodiversity study). We also standardized sampling effort to a common number of individuals, which allowed us to compare true diversity measures by using modern resampling techniques (Colwell, 2013). We present expected diversity measures computed this way from a sample of n individuals as 1Dn; for instance, 1D35 is the mean number of equiprobable species expected in a sample of 35 individuals, computed by resampling methods. We performed all richness and diversity calculations with EstimateS vers. 9.1.0 (Colwell, 2013). To obtain smooth curves for the richness and diversity estimators, we always used n = 100 repetitions, with no replacement between indviduals. Assessing performance (bias and efficiency): we followed Walther and Moore (2005) and Zar (2010) to assess the bias of both richness and true species diversity measures obtained with a given sampling effort j (Ej) as the difference between that measure and a reference value, accepted as real (A). For the richnes estimates, we used the following expression (fig. 1A): PARj (percent of actual richness) = 100 * SMEj + 100, where SMEj (scaled mean error) = MEj/A, and MEj (mean error) = Ej – A. For the diversity estimates, instead of PAR, the corresponding term was PADj (percent of actual diversity), and all the rest were equivalent. Finally, we considered the asymptote from the smooth richness accumulation curve obtained by pooling all the sampling methods as the best available estimation of true richness (A); in a parallel way, we considered the diversity accumulated with all the sampling methods as the best available estimation of true diversity (Walther and Moore, 2005; Ellison et al., 2007). We computed the Pitman efficiency index (DasGupta, 2005; Zar, 2010) and considered that a sampling method, M1, was more efficient than a second method, M2, if M1 could achieve the same response value (y) as M2 with a lower sampling effort (i.e., if n1 < n2; fig. 1B). We present the efficiency of M1 relative to M2 as the ratio between the sample sizes needed to achieve y with both methods (Zacks, 2005): n2/n1. If this ratio > 1, then M1 was more

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Accumulated richness (or diversity)

ME(1)30 > 0

ME(2)30 < 0

6 5 4

3 2 1 0

25 Accumulated richness (or diversity)

20 25 Individuals

15 y

10 5 0

n1 0

10 15 Individuals

Fig. 1. A, an example of how to measure bias in species richness (or diversity) estimates from two sampling methods, M1 and M2. ME(1)30 and ME(2)30 are the mean errors in the richness (or diversity) estimates by M1 and M2 with a common sampling effort; in this example, 30 individuals. a is the reference value, accepted as real, for richness (or diversity); in this example, a = 7. B, an example of how to measure the relative efficiency of two sampling methods, M1 and M2. The common response to attain by the two methods, y, is the minimum richness (or diversity) accumulated with one of the two methods compared; in this example, y = 15. The corresponding sampling efforts to attain the y value with the two methods are n1 and n2. Fig. 1. A, un ejemplo de cómo cuantificar el sesgo en las estimaciones de la riqueza (o diversidad) de especies calculadas con dos métodos de muestreo: M1 y M2. ME(1)30 y ME(2)30 son los errores medios de las estimaciones de la riqueza (o diversidad) M1 y M2 con un esfuerzo de muestreo común, que en este ejemplo es de 30 individuos. a es el valor de referencia, aceptado como real, de la riqueza (o diversidad); en este ejemplo, a = 7. B, un ejemplo de como cuantificar la eficiencia relativa de dos métodos de muestreo: M1 y M2. La respuesta común de ambos métodos, y, es la riqueza (o diversidad) mínima acumulada con uno de los dos métodos comparados; en este ejemplo, y = 15. Los esfuerzos de muestreo correspondientes para obtener el valor de y con los dos métodos son n1 y n2.

efficient than M2. We assessed the relative efficiency of our sampling methods to describe both species richness and true species diversity; the value of y used in each evaluation was the minimum of the total accumulated richness, or total accumulated diversity, obtained with any of the sampling methods being compared (fig. 1B).

Results Short description of the reptile assemblage During the first study period (from June to October 2005) we did not find any animals (i.e., we obtained only negative results) in two of 15 MHS, in three of

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Table 1. Frequency and abundance of the reptile fauna in the study area, by time periods (initial period, from June 2005 to October 2005; and total period, from June 2005 to April 2006) and sampling methods: MHS, microhabitat surveys; TS, transect surveys; PFS, pitfall–trap stations. Frequency (f) is the number of sampling units where each species was recorded, and abundance (N) is the number of individuals actually recorded: – absences. Tabla 1. Frecuencia y abundancia de la herpetofauna en la zona de estudio, por períodos (período inicial, de junio de 2005 a octubre de 2005; y período total, de junio de 2005 a abril de 2006) y métodos de muestreo: MHS, muestreos en microhábitats; TS, itinerarios; PFS, estaciones de trampas de caída. La frecuencia (f) es el número de unidades de muestreo en las que se registra cada especie y la abundancia (N) es el número de individuos registrado: – ausencias.

Time period Initial MHS

Total

TS f

PFS N

MHS N

TS f

Anolis quercorum

1 2 – – – – 2 4 – –

Aspidoscelis sacki

Kinosternon integrum

5 11 – – – – 5 11 – –

Masticophis mentovarius 1 1 – – – – 1 1 – – Phyrnosoma braconnieri 2 2 – – – – 2 2 – – Salvadora intermedia 3 3 – – – – 3 3 1 1 Sceloporus horridus

1 1 1 1 1 1 2 2 3 3

Sceloporus jalapae

Tantilla bocourti

1 2 – – – – 2 3 – –

Thamnophis cyrtopsis

5 5

10 –

15 TS, and in eight of 13 PFS (supplementary material). The mean number of captured species was 2.1 (range: 1‒4) per MHS with positive results; 1.6 (range: 1‒3) per positive TS, and 1.2 (range: 1‒2) per positive PFS. During this period, we captured 36 individuals with MHS, 35 with TS, and 11 with PFS. The two most frequently captured species in the MHS (table 1) were the Mexican mud turtle (Kinosterton integrum) and the Black–necked gartersnake (Thamnophis cyrtopsis); they were also the two most abundantly captured species, toghether with the Jalapa spiny lizard (Sceloropus jalapae). The most frequently captured and the most abundant species in the TS was S. jalapae followed by Sack’s giant whiptail lizard (Aspidoscelis sacki). In the PFS, S. jalapae was the most frequently captured species and A. sacki was the most abundant. Throughout our field work we did not capture any individuals in six out of 23 MHS (supplementary material); in the positive MHS, the mean number of captured species was 1.9 (range: 1‒4). With all the MHS, we captured 47 specimens of the same species captured in the first period with this method (table 1); S. jalapae, K. integrum and T. cyrtopsis were the most frequent, whereas K. integrum and S. jalapae were the most abundant. In five out of 23 TS we did

27 –

3 –

7 5

10 6

18 –

51 –

not find any individuals (supplementary material); the mean number of species in the positive TS was 1.6 (range: 1‒4). With all the TS, we found 65 specimens of the same species captured in the first period with this method, plus the Oaxacan patchnose snake (Salvadora intermedia) (table 1); S. jalapae and A. sacki were the most frequent and abundant species. Richness During the first study period, the accumulated richness curve obtained with the sum of abundance data pooled from the three sampling methods was located above the corresponding curve for PFS and below that for MHS; the curve with the smallest slope was that for TS (fig. 2A). The only sampling method that captured all the species recorded in this study was MHS (S36 = 10). After theoretically doubling the total pooled sampling size, expected richness remained constant (S*164 = 10.2; 95 % confidence interval, based on the unconditional estimator of standard deviation: 9.1‒11.2; fig. 2A). For all our field work, richness results were like those we presented above. Although using TS we captured a total of 65 individuals (almost 2× the corresponding total abundance from the first period),

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250

A Accumulated richness

MHS [36,10]

MHS + TS + PFS [82,10]

4 2

PFS [11,3] TS [35,3]

14 Accumulated richness

[164,10.2]

0 0

60 80 Individuals

100

120

140

160

12 MHS [47,10]

MHS + TS [112,10]

[224,10.4]

8 6 4

TS [65,4]

2 0

100 Individuals

150

200

Fig. 2. Smooth, accumulated species richness curves (y–axis: number of species) vs. increasing sampling effort (x–axis: number of accumulated individuals) for: A, initial period (when the three methods were operating); B, all our field work (only for MHS and TS). In each panel, the dotted line shows the extrapolation up to 2× the corresponding total abundance. The total abundance and richness values for each period×method are shown in brackets: MHS, microhabitat surveys; TS, transect surveys; PFS, pitfall–trap stations. Fig. 2. Curvas continuas de la riqueza de especies acumulada (eje de las y: número de especies) en relación con un esfuerzo de muestreo creciente (eje de las x: número de individuos acumulados) para: A, el período inicial (cuando se aplicaban los tres métodos); B, todo nuestro trabajo de campo (solo para los MHS y los TS). En cada gráfico, la línea discontinua muestra la extrapolación hasta el doble de la abundancia total correspondiente. Los valores totales de abundancia y riqueza de cada período×método se muestran entre corchetes: MHS, muestreos en microhábitats; TS, itinerarios; PFS, estaciones de trampas de caída.

the accumulated total richness with this method S65 = 4 (i.e., just one above the three from the first period), and the accumulated richness curve seemed to finally reach the asymptote (fig. 2B). The accumulated richness with all the MHS did not increase from the value obtained in the first period (S47 = 10). The total richness accumulated with the sum of these two sampling methods did not increase either (S112 = 10). After extrapolating this pooled curve up

to 2× the total number of individuals, the increase in expected richness was negligible (S*224 = 10.4; 95 % confidence interval: 8.4‒12.5; fig. 2B). True species diversity After pooling data from the three sampling methods, the accumulated species diversity curve (1D35 = 5.1 and 1D82 = 5.5; fig. 3A) was located between tho-

Accumulated species diversity

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Accumulated species diversity

251

9 8

MHS [36,7.7]

7 6 5

MHS + TS + PFS [82,5.5]

4 3

PFS [11,2.4]

TS [35,1.9]

1 0

30 40 50 Individuals

MHS [47,8.0]

8 7 6 5

MHS + TS [112,4.8]

4 3

TS [65,2.0]

2 1 0

40 60 Individuals

100

Fig. 3. Smooth, accumulated species diversity curves (y–axis: number of equiprobable species) vs. increasing sampling effort (x–axis: number of accumulated individuals) for: A, initial period (when the three methods were operating); B, all our field work (only for MHS and TS). The total abundance and richness values for each period × method are shown in brackets: MHS, microhabitat surveys; TS, transect surveys; PFS, pitfall–trap stations. Fig. 3. Curvas continuas de la diversidad de especies acumulada (eje de las y: número de especies equiprobables) en relación con un esfuerzo de muestreo creciente (eje de las x: número de individuos acumulados) para: A, el período inicial (cuando se aplicaban los tres métodos); B, todo nuestro trabajo de campo (solo para los MHS y los TS). Los valores totales de abundancia y riqueza de cada período × método se muestran entre corchetes: MHS, muestreos en microhábitats; TS, itinerarios; PFS, estaciones de trampas de caída.

se for MHS (1D36 = 7.7) and PFS (1D11 = 2.4); the first curve to reach the asymptote was that from TS (1D35 = 1.9; fig. 3A). For all our field work, the MHS diversity curve seemed to approach the asymptote (1D47 = 8.0; fig. 3B), and the TS curve did not change from the first period (1D65 = 2.0), even though the total number of individuals did increase (see above). The total species diversity computed with the sum of abundances from all our field work with these two methods (1D82 = 4.7; 1D112 = 4.8)

was slightly inferior to that previously obtained with the individuals captured only during the first study period (i.e., when the three sampling methods were functioning: 1D82 = 5.5; figs. 3A, 3B). Bias and efficienciy Regarding species richness, MHS was the least biased method: with high sampling efforts, bias was negligible (e.g., for the first study period:

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Table 2. Bias (A) and relative efficiency (B) of microhabitat surveys (MHS), transect surveys (TS), and pitfall–trap stations (PFS). Bias was computed as the percentage of actual richness, or diversity, obtained with n individuals. Efficiency was computed as the ratio between the numbers of individuals needed to achieve the given value (richness or diversity) with the two methods compared. As for the initial period 'Sum' was MHS + TS + PFS, but it was MHS + TS for the total: – results not defined. Tabla 2. Sesgo (A) y eficiencia relativa (B) de los muestreos en microhábitats (MHS), los itinerarios (TS) y las estaciones de trampas de caída (PFS). El sesgo se calculó como porcentaje de la riqueza real, o la diversidad, obtenida con n individuos. La eficiencia se calculó como la proporción entre el número de individuos necesarios para lograr un valor dado (riqueza o diversidad) con los dos métodos comparados. Con respecto al período inicial "Sum" fue MHS + TS + PFS, pero para el total fue MHS + TS: – resultados no definidos. A Initial

Total

Richness

Diversity

n = 11

n = 35

n = 11

n = 35

MHS

–36.4

–0.6

1.1

39.3

–77.4

–70.0

–67.8

–65.9

–70

–

–57.0

–

PFS

Richness

Diversity

n = 47 n = 65

–

67.1

–76.3 –60.0 –

–

–59.2 –58.8

–

B Initial Richness

Total Diversity

S = 3

S = 10

MHS vs. TS

8.8

–

10.5

–

MHS vs. PFS

2.8

MHS vs. (Sum) 1.3

Richness

D = 2.0 1D = 2.4 1D = 5.5

–

S = 4 S = 10 10.8

–

Diversity D = 2.0 1D = 4.8

7.0

–

2.0

3.7

–

2.3

1.5

1.3

7.5

1.5

2.4

1.0

12.6

–

TS vs. PFS

0.3

–

0.1

TS vs. (Sum)

0.1

–

0.1

–

0.1

–

0.1

–

PFS vs. (Sum)

0.5

–

0.8

0.4

–

PAR35 = ‒0.6 %) or non–existent (i.e., for the whole field work, with 47 individuals; table 2). The other two sampling methods presented negative biases (i.e., PAR < 0), with absolute values above 60 % (table 2). As for species diversity, MHS displayed positive bias (e.g. for the first study period: PAD35 = +39.3 %), whereas the other two sampling methods showed negative biases. For all our field work, the absolute value of the bias corresponding to the diversity measure from MHS (PAD47 = +67.1) was slightly above that from TS (PAD47 = ‒59.2 %; table 2). In all the comparisons between sampling methods, regardless of the richness or diversity values used in the comparisons, MHS was always the most efficient method to estimate both the number of species and the true species diversity (table 2). To achieve the maximum richness obtained during the first study period with TS or PFS (i.e., S = 3), MHS was between 1.3 (vs. all the methods combined) and

8.8 (vs. TS) times more efficient. More significantly, to achieve the total number of species recorded in our study (i.e., S = 10), MHS was 2.3 times more efficient than all methods combined. TS was the least efficient method to estimate both species richness and diversity (table 2). Discussion As biodiversity studies can not usually determine total richness in a local assemblage through exhaustive enumeration of all the species (but see González–Oreja et al., 2010),extrapolation (Colwell and Coddington, 1994; Gotelli and Colwell, 2001) or other techniques (Gotelli and Chao, 2013) are often needed. Our results show that the reptile assemblage in the study area was composed of 10 species; or, sensu Longino et al. (2002), that the set of species

Animal Biodiversity and Conservation 41.2 (2018)

really sampled by the methods used was composed of only 10 species. Moreover, all 10 species were recorded with only one sampling technique (i.e., microhabitat surveys), and with a small sampling effort. Our results also show that MHS was the best method (i.e., the least biased and the most efficient method) to determine the number of species. In biodiversity studies with other animal groups, various authors have observed that sampling methods such as the MHS we used (e.g., the hand collection of ants and spiders) can not only be applied in settings where environmental conditions can exclude the use of other sampling methods but they can also be more efficient (King and Porter, 2005; Gotelli et al., 2011). Notwithstanding, sampling methods such as MHS and hand–collecting are influenced by the previous experience of the field worker (Longino et al., 2002; Ellison et al., 2007; Gotelli et al., 2011; see, also, Cardoso, 2009), which makes their standardization difficult (Blomberg and Shine, 2006; see also Mehrabi et al., 2014). For instance, expert field herpetologists could direct their attention, deliberately or not, to those microhabitats where rare species can be recorded. Inter–observer bias can be ruled out in our study as all the field work was completed by one observer (i.e., the first author of this study). Like many other authors (e.g.Peterson et al., 2004; Hutchens and DePerno, 2009; Fernández Badillo and Goyenechea–Mayer Goyenechea, 2010; Percino Daniel et al., 2013; Hidalgo Penninger, 2014), we also used transect surveys and pitfall–trap stations. On one hand, TS were strongly biased, even with large sampling efforts, and were the least efficient method. A possible reason for this negative result is that, during the TS, the observer was detected by the certain reptile species first, and not the other way round. This could help to explain the low accumulated total richness obtained with this method. On the other hand, the accumulation of species by PFS did not approach the asymptote even with the highest number of captured individuals. It would thus be interesting to replicate our study and evaluate the performance of this method with larger sampling efforts. Consequently, at least to determine the number of reptile species in the dry scrubland we studied, intensive searching for the target species in those sites that are frequently used as refuges would suffice. However, in line with the limitation previously observed regarding estimation of the total richness in the study area (sensu Longino et al., 2002), it can not be ruled out that other sampling methods could have detected new reptile species. There is usually no sampling method that allows all the species in a local assemblage to be captured (because of the detection bias; Yoccoz et al., 2001), and different sets of species can be under–represented or over–sampled in the samples obtained by contrasting methods (Gotelli and Colwell, 2001). Still, the number of species in our reptile assemblage is similar to the species richness other authors have documented in thorny scrublands (six or seven species: Fernández Badillo and Goyenechea–Mayer Goyenechea, 2010) and dry shrublands (nine species: Vite Silva et al., 2010) from other Mexican regions (see also Ramírez Bautista et al., 2010).

253

Now, we will answer the question we raised concerning whether the best method to determine the number of species can simultaneously be the best method to estimate the true diversity of the local assemblage. This was clearly not the case in our study. If the species diversity accumulated by pooling all the sampling methods (i.e., the structured inventory) were the best available estimation of the actual diversity (as has been considered by previous authors: Walther and Moore, 2005; Ellison et al., 2007), then the best method to estimate species richness (i.e., MHS) was also a biased method (as were the other sampling techniques) that overestimated the actual diversity value as much as other methods underestimated it. Our finding does not support the suggestion by Steiner et al. (2005; cited by Gotelli et al., 2011) that, to compare between sites or habitat diversity measures in animal assemblages (in their case, ants), it can be better to use a single collecting method. Presumably, a structured inventory (even with an unbalanced design; Cardoso, 2009) will expose a more accurate image of the studied assemblage; therefore, the best sampling method to study species richness may not be the best method to study species diversity. Although combining different collecting methods may increase the budgets of time, field–work and money needed for the study (Gotelli et al., 2011), not to mention the need to consider the specific biases of each method (Gotelli et al., 2011), the structured inventory can be the most convenient way to study species diversity. Other studies with structured inventories usually observe that some sampling methods can find sets of unique species that are not recorded using other methods (King and Porter, 2005; Hutchens and DePerno, 2009; Gotelli et al., 2011). In our case study, MHS (which was also a biased method to estimate species diversity) could have overestimated the abundance of those species linked to bodies of water, like K. integrum and T. cyrtopsis. Moreover, because of its arboreal habits, the abundance of the Oaxacan oak anole (Anolis quercorum) could have been underestimated by the same method. If this were true, it would be appropriate to combine diverse sampling methods at the same time and obtain the species inventory with the highest possible diversity. Otherwise, estimates of species diversity could be misleading. Finally, our study brings into question the utility of comparing estimates of species diversity obtained with different methods. Whereas it is true that the best method to estimate the number of species in a local assemblage cannot yield a higher value than the true richness, and therefore comparisons of species richness obtained with contrasting sampling methods can result in reasonable values, this is not true for the estimation of species diversity. In fact, even the best method to estimate species diversity can over– or under–estimate the true value. Therefore, not knowing the corresponding bias (or the accuracy) of the sampling methods involved in the comparison can render useless comparisons of diversity estimates between methods.

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Supplementary material

Table 1s. Abundance (number of individuals) of all the reptile species recorded during field work in the study area (Tecali de Herrera, Puebla, Mexico), by sampling date (from June 6th, 2005 to April 6th, 2006) and sampling methods (microhabitat surveys, transect surveys and pitfall–trap stations).

April 6, 2006

March 11, 2006

February 27, 2006

January 29, 2006

January 28, 2006

December 10, 2005

November 26, 2005

November 12, 2005

October 9, 2005

October 8, 2005

September 11, 2005

September 10, 2005

September 1, 2005

August 28, 2005

August 27, 2005

June 25, 2005

June 24, 2005

June 23, 2005

June 18, 2005

June 12, 2005

June 11, 2005

June 10, 2005

June 9, 2005

Tabla 1s. Abundancia (número de individuos) de todas las especies de reptiles registradas durante el trabajo de campo en el área de estudio (Tecali de Herrera, Puebla, México), por fecha de muestreo (del 6 de junio de 2005 al 6 de abril de 2006) y métodos de muestreo (muestreos en microhábitats, itinerarios y estaciones de trampas de caída).

Microhabitat surveys Anolis quercorum 0 0 0 0 0 0 0 0 0 2 0 0 0 0 2 0 0 0 0 0 0 0 0 Aspidoscelis sacki 0 0 0 0 0 0 0 0 0 0 1 0 0 3 0 0 0 0 0 0 0 1 0 Kinosternon integrum 1 0 0 0 0 6 1 0 0 0 0 1 2 0 0 0 0 0 0 0 0 0 0 Masticophis mentovarius 0 0 0 0 0 0 0 0 0 0 0 0 0 0 1 0 0 0 0 0 0 0 0 Phyrnosoma braconnieri 1 0 0 0 1 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 Salvadora intermedia 0 0 0 0 1 0 0 1 0 0 0 0 0 0 1 0 0 0 0 0 0 0 0 Sceloporus horridus 0 0 0 0 0 0 0 0 0 0 0 1 0 0 0 0 0 0 0 0 0 0 1 Sceloporus jalapae 0 0 0 0 0 0 0 0 1 0 1 1 0 3 0 0 0 0 0 0 2 1 1 Tantilla bocourti 0 0 0 0 0 0 0 0 0 0 0 0 0 0 2 0 0 0 0 1 0 0 0 Thamnophis cyrtopsis 0 0 2 0 0 1 0 0 1 0 1 0 0 1 0 0 0 0 0 0 0 0 0 Transect surveys Anolis quercorum 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 Aspidoscelis sacki 1 0 1 0 2 0 0 0 1 0 0 1 0 0 2 0 1 0 0 0 1 0 0 Kinosternon integrum 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 Masticophis mentovarius 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 Phyrnosoma braconnieri 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 Salvadora intermedia 0 0 0 0 1 0 0 0 0 0 0 0 0 0 0 0 1 0 0 0 0 0 0 Sceloporus horridus 0 0 0 0 1 0 0 0 0 0 0 0 0 0 0 0 1 0 0 0 1 0 0 Sceloporus jalapae 4 0 2 0 4 3 2 0 3 2 3 3 2 2 2 0 2 4 3 0 3 4 3 Tantilla bocourti 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 Thamnophis cyrtopsis 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 0 Pitfall–trap stations surveys Anolis quercorum 0 0 0 0 0 0 0 0 – – 0 0 0 0 0 – – – – – – – – Aspidoscelis sacki 0 0 2 0 0 0 0 5 – – 0 0 0 0 0 – – – – – – – – Kinosternon integrum 0 0 0 0 0 0 0 0 – – 0 0 0 0 0 – – – – – – – – Masticophis mentovarius 0 0 0 0 0 0 0 0 – – 0 0 0 0 0 – – – – – – – – Phyrnosoma braconnieri 0 0 0 0 0 0 0 0 – – 0 0 0 0 0 – – – – – – – – Salvadora intermedia 0 0 0 0 0 0 0 0 – – 0 0 0 0 0 – – – – – – – – Sceloporus horridus Sceloporus jalapae Tantilla bocourti Thamnophis cyrtopsis

0 0 0 0 0 0 0 0 – – 0 1 0 0 0 – – – – – – – – 0 1 1 1 0 0 0 0 – – 0 0 0 0 0 – – – – – – – – 0 0 0 0 0 0 0 0 – – 0 0 0 0 0 – – – – – – – – 0 0 0 0 0 0 0 0 – – 0 0 0 0 0 – – – – – – – –

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Jaguar (Panthera onca) and puma (Puma concolor) diets in Quintana Roo, Mexico D. M. Ávila–Nájera, F. Palomares, C. Chávez, B. Tigar, G. D. Mendoza

Ávila–Nájera, D. M., Palomares, F., Chávez, C., Tigar, B., Mendoza, G. D., 2018. Jaguar (Panthera onca) and puma (Puma concolor) diets in Quintana Roo, Mexico. Animal Biodiversity and Conservation, 41.2: 257–266. Abstract Jaguar (Panthera onca) and puma (Puma concolor) diets in Quintana Roo, Mexico. A study was carried out for two years in Northwest Quintana Roo, México, using scat analysis to determine the diet and prey preferences of pumas and jaguars. Cat species and gender were determined using molecular techniques (rapid classificatory protocol: polymerise chain reaction, RCP–PCR), and prey abundance was estimated from camera trapping. The scats contained remains from 16 wild mammal species, but there was no evidence of livestock or other taxa. The diet breadths of jaguar (0.32) and puma (0.29) indicated a high degree of prey specialization, which combined with their dietary overlap (Pianka index 0.77) suggested competition between them. However, both felids showed a preference for red brocket deer Mazama temama, and frequently consumed collared peccaries Pecari tajacu. The importance of such large ungulates in the felids' diets is similar to the expected patterns of wild meat consumption in rural areas of the Northern Yucatan Peninsula. Therefore, future conservation management plan initiatives should involve local rural communities in the management of sustainable hunting, considering these ungulates are also the felid prey species. Key words: Diet breadth, Diet overlap, Felines, Human–felid conflict, Subsistence hunting, Wild meat Resumen La dieta del jaguar (Panthera onca) y del puma (Puma concolor) en Quintana Roo, en México. El estudio se realizó durante dos años en el noroeste de Quintana Roo, en México y se utilizó el análisis de excrementos para determinar la dieta y las preferencias de presas del puma y del jaguar. Se utilizaron técnicas moleculares para identificar la especie de félido y el sexo (protocolo de clasificación rápida: reacción en cadena de la polimerasa, RCP–PCR), y se estimó la abundancia de presas mediante el método de trampeo fotográfico. Los excrementos contenían restos de 16 especies de mamíferos salvajes, pero no se encontraron restos de ganado ni de otros taxones. La amplitud de la dieta del jaguar (0,32) y del puma (0,29) indica que son especies con un alto grado de especialización, lo cual, junto con el traslape de las dietas (índice de Pianka = 0,77) sugiere que ambos félidos compiten entre sí. Asimismo, ambos mostraron preferencia por el venado temazate, Mazama temama, y frecuentemente consumieron pecarí de collar, Pecari tajacu. La importancia de la presencia de este tipo de ungulados en la dieta de los félidos se corresponde con la pauta esperada de consumo de carne de caza en las zonas rurales del norte de la península de Yucatán. Por lo tanto, las futuras iniciativas encaminadas a planificar la conservación de ambos félidos deberían hacer partícipes a las comunidades rurales en la gestión de la cacería sustentable, considerando que estos ungulados también son presas de los félidos. Palabras clave: Amplitud de dieta, Traslape de dieta, Félidos, Conflicto humano–félido, Caza de subsistencia, Carne de caza Received: 05 IV 17; Conditional acceptance: 22 VII 17; Final acceptance: 30 X 17

ISSN: 1578–665 X eISSN: 2014–928 X

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D. M. Ávila–Nájera, G. D. Mendoza, Depto. de Producción Agrícola y Animal, Univ. Autónoma Metropolitana, Unidad Xochimilco, Calzada del Hueso 1100, Col. Villa Quietud, D. F. 04960, México.– F. Palomares, Estación Biológica de Doñana–CSIC, Avda. María Luisa s/n., 41013 Sevilla, España.– C. Chávez, Depto. de Ciencias Ambientales, CBS Univ. Autónoma Metropolitana, Unidad Lerma, Hidalgo Pte. 46, Col. La Estación Lerma, Estado de México 52006, México.– B. Tigar, School of Forensic and Applied Sciences, Univ. of Central Lancashire, Preston, PR1 2HE, UK. Corresponding author: German David Mendoza. E–mail: gmendoza@correo.xoc.uam.mx

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Introduction Jaguars (Panthera onca) and pumas (Puma concolor) are two large felids that occur sympatrically across much of the Americas. The distribution of jaguar populations ranges from northern Mexico to Argentina, but has declined in recent years and they are currently thought to occupy only about 46 % of their historic range (Sanderson et al., 2002). The IUCN considers the jaguar to be a Near Threatened species and it is listed in Appendix I of CITES (Caso et al., 2008). In contrast, pumas have a much wider geographical distribution and tolerate a wider range of climate types than jaguars, and occur from Canada throughout parts of the USA, Central and South America, including the southern tip of Chile (Sunquist and Sunquist, 2002). Pumas are listed as being species of Least Concern by the IUCN (Nielsen et al., 2015) and are included in Appendix II of CITES, although they no longer occur in some regions where they were previously common (Nowell and Jackson, 1996). However, the rarer Eastern and Central American subspecies of puma (P. c. coryi, P. c. costaricensis and P. c. cougar) are listed separately in Appendix I of CITES (Nowell and Jackson, 1996). In general, global populations of large felids continue to decline due to habitat loss and fragmentation, frequently exacerbated by the impact of increased human activity and the risk of conflict and persecution by hunters and livestock farmers (Loveridge et al., 2010; Foster et al., 2014). In the Yucatan Peninsula, socioeconomic development has caused large scale land–use changes including deforestation and habitat fragmentation (Cespedes–Flores and Moreno–Sánchez, 2010) which have been accompanied by increased hunting of wild game species (Naranjo et al., 2010). In tropical Mexico, up to 70 % of the meat consumed by rural communities originates from hunting, mainly large species of ungulate such as tapirs Tapirus bairdii, white–tailed deer, Odocoileus virginianus, and collared and white–lipped peccaries (Pecari tajacu and Tayassu pecari) (Marmolejo, 2000), which are also consumed by large predators. The jaguar and puma are obligate carnivores and where their distributions overlap in Central and Latin America (Sunquist and Sunquist, 2002) they both prey opportunistically on mammals (Oliveira, 2002; Scognamillo et al., 2003; Novack et al., 2005; Weckel et al., 2006). In the Southern Yucatan Peninsula, Mexico, both these felids mainly consume large prey like collared and white–lipped peccaries, red brocket deer (Mazama temama) and white–tailed deer (Chávez et al., 2007). Despite the potential for interspecific competition for food, these similarly– sized felids are able to coexist in many parts of their range through differences in their prey–use, including specialization according to the size, species, age and total biomass of prey consumed, combined with differences in their spatial and temporal habitat–use (Taber et al., 1997; Chávez, 2010). Therefore, the prey preferences and diet breadths of the two cats can vary according to the local availability and abundance of prey (Núñez et al., 2000; Hernández–Saint Martín et al., 2015).

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The Yucatan Peninsula has the largest jaguar population in Mexico (Chávez et al., 2007) and is one of six proposed priority areas for its conservation (Rodríguez–Soto et al., 2011). However, little is known about the diet or likely competition for prey between these two felids and humans in the Northern Yucatan Peninsula (Ávila–Gómez, 2003). Therefore, the objectives of the present study were: (1) to determine the diet and prey consumption patterns of both cat species; (2) to estimate their trophic niche widths and the degree of prey specialization; and (3) to measure the amount of dietary overlap between them. The results were compared with published data on local hunting practices to explore the potential competition between the cats and local rural communities, and are considered in terms their implications for the sustainable management of large felids and their prey species in a region where socio–economic development continues to make profound changes to rural lifestyles (Santos–Fita et al., 2012). Material and methods Study site The study took place in the Eden Ecological Reserve (Eden) and surrounding Lazaro Cardenas municipality, Quintana Roo, Mexico (21° 36' 00'' – 20° 34' 00'' N and 87° 06' 00'' – 87° 45' 00'' W). This 3077 ha private reserve is part of the Yalahau biological conservation region (Gómez–Pompa et al., 2011). The vegetation is dominated by medium–stature tropical forest and secondary forest (Schultz, 2003), and the reserve is surrounded by a landscape mosaic of secondary forest and managed habitats, including indigenous milpa cultivation (slash and burn) and rural villages. Methods Faecal pellet collection and camera trapping occurred during May to July 2011 and August to September 2012. The camera trap locations were selected using the CENJAGUAR (Chávez et al., 2007), which requires at least nine adjacent 9 km2 study plots each containing two or three camera stations, with at least one station per plot having two cameras directly facing each other. Cameras were placed 1.5–3 km apart along a series of forest trails, firebreaks and dirt roads of differing lengths (8–16 km) and their locations are shown in figure 1 and described in Ávila–Nájera et al. (2015). In 2011, there were 22 camera stations operating continuously for 82 days, with 24 cameras operating over 72 days in 2012. Scats were collected daily by systematically searching along each dirt road, firebreak and forest trail where the cameras were located. Scats were stored in plastic bags and divided into two. One half underwent a rapid classificatory protocol–PCR (RCP–PCR) to assign a species (jaguar or puma) and gender to the scat (Roques et al., 2011). This method consisted of a single–tube multiplex RCP–PCR yielding species– specific banding patterns on an agarose gel, which

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N W

E S

Acahual Agricultural Savanna Medium forest Camera station Sink hole Fig. 1. Map of Mexico showing the current distribution of jaguars (Panthera onca) in grey dots and pumas (Puma concolor) in dark lines. The two boxes show the study site (the Eden Ecological Reserve, Quintana Roo) with vegetation types and camera site locations for 2011 (left) and 2012 (right). Fig. 1. Mapa de México en el que se muestra la distribución actual del jaguar (Panthera onca), en puntos grises, y la del puma (Puma concolor), en líneas oscuras. Los dos cuadros muestran el área de estudio (Reserva Ecológica El Edén, en Quintana Roo) con los tipos de vegetación y los sitios de muestreo fotográfico de 2011 (izquierda) y 2012 (derecha).

ensures the unambiguous identification of jaguars and pumas from other felid species. For sex determination, we used Pilgrim et al.'s (2005) method based on the differences in size between the RCP–PCR products amplified from the male Y–chromosome copy (AMELY) and the X–chromosome gene (AMELX), and optimised for faecal samples from Neotropical felid species such as jaguar, puma, ocelot and margay, as described by Palomares et al. (2012). The other half of each scat was washed with water and oven dried at 45 °C; the remains of all traces of hair, bones and teeth were removed and identified by comparison with Mexican reference material, as described by Monroy–Vilchis and Rubio–Rodríguez (2003). The relative consumption of each prey species was estimated from the frequency and percentage frequency of occurrence, and the percentage of times that remains of each species were recovered from scats. The relative amount of biomass consumption (RBC) of each prey species and the number of organisms consumed (NOC) were calculated for both felids using Ackerman et al.'s (1984) conversion for puma: RBC = (AF*Y) / ∑/FA*Y

where AF is the absolute frequency of prey in the scats and Y is the weight of food consumed to generate a scat for each prey species and: NOC = (RBC/p) / ∑ (RBC/p) where p is the mean live prey weight (kg) according to Ceballos and Oliva (2005), but excluding long–tailed weasels, Mustela frenata, which were the only species below the 2 kg threshold for this equation (Ackerman et al., 1984). Dietary diversity (diet breadth) was calculated using Levins' index (Levins, 1968) and the overlap between the diet of jaguars and pumas was estimated using Pianka's index (Pianka, 1973). The overlap between the potential prey based on species identified in the camera traps, and actual prey species recovered in the scats was estimated using Sorensen's similarity coefficient (Ss) (Krebs, 1999). The significance of the overall niche overlap between the cats was tested by comparing our observed values with values obtained by randomizing the original matrices following 1,000 iterations with the ra3 algorithm, using the EcoSim–R package in R (Gotelli and Entsminger, 2001; Winemiller and Pianka, 1990).

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Relative abundance

12 10 8 6 4 2 0

Mt Nn Prey species

Fig. 2. Annual relative abundance of felid prey from camera trapping at the Eden Ecological Reserve, Mexico, for 10 species photographed (black bars show records for 2011 and white bars show records for 2012): Cp, Cuniculus paca; Dn, Dasypus novemcinctus; D, Didelphys sp; Lw, Leopardus wiedii; Mt, Mazama temama; Nn, Nasua narica; Ov, Odocoileus virginianus; Pt, Pecari tajacu; Pl, Procyon lotor; Uc, Urocyon cinereoargenteus. Fig. 2. Abundancia relativa anual de presas de félidos por medio de trampeo fotográfico en la Reserva Ecológica El Edén, en México, para 10 especies fotografiadas (las barras negras y blancas muestran registros correspondientes a 2011 y 2012, respectivamente). (Para las abreviaturas de las especies presa, véase arriba).

The relative abundance of prey was derived from the number of independent records of each species photographed in camera–traps per sampling effort (Monroy–Vilchis et al., 2011). An independent record was considered to have occurred when (1) photographs of an individual animal were more than 30 min apart, (2) different individuals of the same species could be distinguished in consecutive photos, (3) several individuals could be identified in the same photo and (4) a new event was recorded after three hrs if it was not possible to identify different individuals of the same species in consecutive photos. Each predator's preference for a prey species was calculated using Ivlev's electivity index (E) (Strauss, 1979) on a scale from –1 to +1, where –1 indicates rejected or inaccessible prey, +1 indicates actively selected prey, and zero indicates prey that were consumed according to their relative abundance. Finally, the biomass and estimated number of prey consumed by both felids were compared with published data on patterns of wild meat hunting from Quintana Roo to assess potential competition between felids and humans. Results We found a total of 49 scats, of which 23 were from jaguars and 26 were from pumas. Of the jaguar scats, 20 were from males and the other three could not be assigned a gender by RCP–PCR, whilst for pumas, 13 scats were assigned as males and nine as females,

with four puma scats that could not be assigned. We found remains from 16 mammal species in the scats from both felids, with no evidence of bird, reptile or livestock remains. We detected diet breadths of 0.32 for jaguar and 0.29 for puma, and the dietary overlap between them was 0.77 with a mean similarity index of 0.50 and a variance of 0.02 at P (observed ≥ expected) < 0.04. In the jaguar scats we found remains from 15 species, with up to five prey per scat. Their most frequently occurring prey were the large ungulates M. temama and P. tajacu (in > 18 % scats), followed by smaller mammals, kinkajous Potos flavus and nine–banded armadillos Dasypus novemcinctus (in > 8 % scats) (table 1). In the puma scats we found remains from 11 species, with up to three prey per scat. Their most frequently occurring prey were P. tajacu (in > 37 % scats) followed by O. virginianus and coatis Nasua narica (in > 11 % of scats), and Geoffrey’s spider monkeys Ateles geoffroyi, M. temama and D. novemcinctus (in > 8 % scats) (table 1). The differences between the diets included opossum Didelphys sp. remains in puma but not jaguar scats, whilst striped hog–nosed skunks Conepactus semiestratus, Central American agoutis Dasyprocta punctata, long–tailed weasels, northern tamanduas Tamandua mexicana and gray foxes Urocyon cinereoargenteus were found in jaguar but not puma scats. The estimated biomass and number of prey consumed suggest that nearly half the biomass of jaguar diets came from two ungulates, M. temama and P. tajacu, although their most numerous prey were

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Ivlev’s electivity index

262

1 0.8 0.6 0.4 0.2 0 –0.2 –0.4 –0.6 –0.8 –1

0.9

0.35

0.72

0.69

0.61 0.46

0.29 0.02

–0.04

0.25 0.14

–0.04

–1 –0.53

0.02

–0.54

–0.66

–0.85

–1

Mt Nn Lw Prey species

Fig. 3. Prey selection by jaguar (black bars) and puma (white bars) in the Eden Ecological Reserve, Quintana Roo, Mexico, according to Ivlev’s electivity index (E) based on prey remains in scats (n = 23 for jaguar and n = 26 for puma). (For the abbreviations of prey species, see figure 2). Fig. 3. Selección de presas por el jaguar (barras negras) y el puma (barras blancas) en la Reserva Ecológica El Edén, en Quintana Roo, México, según el índice de selectividad de Ivlev (E) basado en los restos de presas encontrados en los excrementos (n = 23 y n = 26 para el jaguar y el puma, respectivamente). (Para las abreviaturas de las especies presa, véase la figura 2).

small mammals, D. novemcinctus, P. flavus and U. cinereoargenteus (mean live body weights all > 2 and < 4.8 kg, table 1). P. tajacu contributed the highest amount of biomass to puma diets (> 36 %) followed by the two deer species, O. virgineanus (17.3 %) and M. temama (11.8 %), whilst their most numerous prey were P. tajacu and N. narica, followed by D. novemcinctus and A. geoffroyi (table 1). Ten of the 16 species recovered from the scats were also recorded in the camera traps (fig. 2) with a high overlap between the animal diversity in camera traps and that of the jaguar (Ss = 0.64) and puma (Ss = 0.60) diets. We also recovered prey items from scats which were not captured in the camera traps, including the arboreal species A. geoffroyi, P. flavus and T. mexicana, and smaller mammals like C. semiestratus, spotted pacas Cunilicus paca and Northern raccoons Procyon lotor. Ivlev’s electivity indices suggested a degree of prey preference and avoidance by the cats (fig. 2), with M. temama preferred by both felids although rarely photographed. O. virginianus appeared to be avoided or inaccessible to jaguars but not pumas, whilst P. tajacu was frequently photographed and consumed by both felids (fig. 3, table 1). N. narica were photographed frequently but were either avoided or inaccessible to jaguars and to a lesser extent pumas, Didelphys spp. were avoided by or were inaccessible to jaguars, whilst U. cinereoargenteus were avoided by or were inaccessible to pumas and, to a lesser extent, jaguars. Finally, C. paca and margays Leopardus wiedii appeared to be consumed according to their availability by jaguars, whilst pumas showed a slight preference for C. paca.

Discussion Despite some evidence of dietary overlap, jaguars and pumas can coexist at Eden due to differences in their prey preferences, their indiviudal niche breadths, and the relative amount of biomass of each prey species they consume. The dietary overlap of the felids found in this study (0.37) was similar to that found in regions such as Campeche (Mexico), Costa Rica and Peru (0.26–0.39). However, intermediate (Brazil and Abra–Tanchipa, Mexico, 0.49–0.57) and high dietary overlaps have been reported elsewhere, including other parts of Mexico, Jalisco, Brazil, and Paraguay (0.78–0.84) (Oliveira, 2002). Diet breadths at Eden were low (both ≤ 0.32) and are similar to those of other studies in Mexico (Gómez and Monroy–Vilchis, 2013; Hernández–Saint Martin et al., 2015), with both felids consuming relatively few species, which is typical of animals with specialist diets (≤ 0.6, Krebs, 1999). Jaguars preyed upon a slightly higher number of species than pumas, with four prey species recovered from jaguar but not puma scats, and one species recovered in puma but not jaguar scats. Ivlev's indices suggested that both felids showed preferences for and avoidance of particular prey species, including their high consumption of M. temama which was rarely recorded in the camera traps, and of P. tajacu which was frequently photographed. These two ungulates contributed about half of the jaguars' dietary biomass, whilst P. tajacu and O. virginianus together contributed more than half of the pumas' prey biomass, including more than a third of this from P. tajacu alone. O. virginianus was consumed by pumas in proportion

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Table 1. Frequency and relative consumption of prey species by jaguar (Po, Panthera onca) and puma (Pc, Puma concolor) in the Eden Ecological Reserve, Quintana Roo, Mexico, estimated from their remains in scats (n = 26 and 23 respectively). Tabla 1. Frecuencia y consumo relativo de especies de presa por el jaguar (Po, Panthera onca) y el puma (Pc, Puma concolor) en la Reserva Ecológica El Edén, en Quintana Roo, México, estimados a partir de excrementos (n = 26 y 23, respectivamente).

Frequency of occurence

Number consumed

13 11.5

6.1 8.6

5.5

Conepatus semiestratus 3 0 8.7 0

4.1 0

3.5 0

4.5 0

4.1 2.9

4 2.6

1.4 2.1

8.7 3.9

Biomass consumed

Percentage of prey

Ateles geoffroyi

Frequency in the scats

4.1 12.1

Cuniculus paca

3 1

Dasyprocta punctata

3 3 8.7 11.5 4.1 8.6 3.5 6.8 4.8 15

Dasypus novemcinctus 4 0 17.4 0

6.7 0

Leopardus wiedii

1 1 4.3 3.9

Mazama temama

Mustela frenata

3 0 13 0 6.1 0 NA 0 NA 0

Nasua narica

Odocoileus virginianus

2 4

Potos flavus

4 1 17.4 3.9

0 2.9

7.2 0

0 1 3

0 3.9

8.2 0

Didelphys sp.

0 2.2

0 9.3

2 2.9 1.8 2.3

2 5.9

39.1 11.5 18.4 8.6 27.9 11.8 2.5 2.5 4.3 15.4

2 11.4

4.3 15.4

2 11.4 3.4 17.3 0.3 3 8.2 2.9

1.8 9.1

1.7 19.6

7.1 2.2

8.6 6.3 4.5 4.4

Procyon lotor

3 1

13 3.9

6.1 2.9

5.5 2.3

Tamandua mexicana

2 0 8.7 0

4.1 0

3.6 0 3.2 0

Pecari tajacu

18.4 37.1 19.9 36.4

39.1 50

Urocyon cinereoargenteus 3 0 13 0 6.1 0 5.3 0

to its high relative abundance in the camera traps, but it appeared to be avoided by or inaccessible to jaguars. This may suggest differences in the felids’ use of prey species and habitat, since O. virginianus tolerates open terrain, including pasture and areas under cultivation which have expanded in the Yucatan Peninsula where they are associated with its recent population increases (Fitos–Santos et al., 2012). The absence of livestock remains in scats at Eden is significant, and because an abundant supply of wild prey is thought to reduce the incidence of felid attacks on livestock (Amit et al., 2013) our results suggest there are sufficient natural prey to support both felids, despite Eden’s small size and the occasional presence of livestock in the reserve. The frequency of prey in the diet and the relative biomass of each prey species consumed by the two predators varies widely across their range (Oliveira, 2002). At Eden, small mammals (< 10 kg) contributed 35–41 % and large mammals contributed 59–65 % of the felids' dietary biomass, in contrast to the Southern Yucatan Peninsula where four large prey species, M. temama and P. tajacu, O. virginianus and T. pecari,

4.7 20 6 0

contributed 86–95 % of the dietary biomass (Chávez et al., 2007). This emphasizes the need for accurate local data on prey preferences and availability especially where felids occur in close proximity to human populations, since hunting for wild–meat could create conflict and competition. The hunting rates reported for human populations in the Northern Yucatan (Fitos–Santa et al., 2012) and the prey consumption patterns of felids in this study suggest that ungulates, armadillos and coatis are major dietary components of both humans and felids. Further evidence for potential competition between felids and humans is that across the whole of Southern Mexico the ungulates are the most commonly used animals for food, medicine and decoration (Contreras–Moreno et al., 2012; Naranjo et al., 2010; Retana–Guiascón et al., 2011; Tejeda–Cruz et al., 2014; Toledo et al., 2008). There are reports of some hunters taking up to 4,900 kg wild–meat yr–1 (Ojasti, 2000; Pug–Gil and Guiascón, 2012), and in Chiapas State the main prey species of felids reported here are widely hunted for human consumption, with over 450 O. virginianus, M. temama, P. tajacu and D. punctata, plus many D. novemcinctus and N.

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narica taken by one community in a single year (Ávila–Gómez, 2003). Contemporary hunters, including those in the Northern Yucatan, are less dependent on wild–meat and usually target larger game such as deer and peccaries (Santos–Fita et al., 2012), which are also the preferred prey of felids (this study and Chávez et al., 2007; Núñez et al., 2000). Therefore, the relatively high number of small prey recorded for felids in Eden may be a response to competition with hunters for large prey and/or their avoidance of areas frequented by humans. There are some limitations to this study. First, most jaguar scats were confirmed as originating from male cats, probably reflecting an inherent collecting bias because female jaguars rarely use open tracks (Palomares et al., 2012). However, there are few viable options for finding scats from wild felids in the natural vegetation prevalent at Eden. The mean body mass of males of both felids species is higher than for females, and most scats at Eden were from male felids that appeared to hunt small prey compared with studies from other parts of the Yucatan Peninsula (Chávez et al., 2007), which could indicate that they were smaller individuals, less likely to hunt larger prey. In addition, some species recovered in the scats were under–recorded by the camera traps, including arboreal or small mammals (A. geoffroyi, P. flavus and T. mexicana, C. semiestratus, C. paca and P. lotor), which is likely to increase their electivity indices. We did not study seasonal differences in diet and prey availability between dry and rainy seasons because previous experience during heavy rains resulted in camera malfunctions and scats being washed away. Other limitations in this study include the low number of scats collected, although this is consistent with estimated population densities of up to 3.6 jaguar and 5.2 puma for Eden (Ávila–Nájera et al., 2015). Despite their small size, Eden and similar reserves may play a disproportionate role in maintaining the overall populations of large felids because these animals require large territories and safe access to sufficient prey, and regularly move across both protected and unprotected areas. In the Northern Yucatan (Santos–Fita et al., 2012) and other parts of Latin America, felid predation of wild–meat species has been used to justify their persecution, even though there is no evidence to confirm that they reduce the population density of their natural prey (Foster et al., 2014). At Eden we found no evidence that they consume livestock. However, as their main prey species are those also favored as wild–meat, long–term conservation management plans of the endangered jaguar can only be achieved by co–managing the sustainable harvesting of wild–meat in the Northern Yucatan Peninsula, in close collaboration with rural communities (Rodríguez–Soto et al., 2011). Acknowledgements We thank Marco Lazcano and Kathy Cabrero for contributing to field research at the Eden Ecological Reserve and the Tropical Research Center, Universidad

Veracruzana, and Erik J. Torres, Juan Castillo, Alejandro Pacheco, Brady Hollinsgworth and Global Vision International volunteers for help with logistics and field work. We also thank Joaquin Arroyo and Aurelio for advice and access to the reference collection at the Archaeozoology Laboratory of the National Institute of Anthropology and History, and also Magdalena Crosby for permit to works in the Animal Nutrition Laboratory in Colegio de Postgraduados. Thanks to Alicia Morales Villarreal for information data. DNA analysis was carried out at the Laboratorio de Ecología Molecular in the Estación Biológica de Doñana, Spain (LEM–EBD– CSIC). The collection permit is SGPA/DGVS/03167/12 of SEMARNAT. We also thank the editor for helpful suggestions that improved the article. References Ackerman, B. B., Lindzey, F. G., Hemker, T. P., 1984. Cougar food habits in southern Utah. Journal of Wildlife Management, 48(1): 147–155. Amit, R., Gordillo–Chávez, E., Bone R., 2013. Jaguar and puma attacks on livestock in Costa Rica. Human–Wildlife Interactions, 7(1): 77–84. Ávila–Gómez, G., 2003. Manejo de Fauna Silvestre en bosques tropicales por ejidos forestales de Quintana Roo. Master thesis, Colegio de Posgraduados. Ávila–Nájera, D. M., Chávez, C., Lazcano–Barrero, M. A., Pérez–Elizalde, S., Alcántara, J. L., 2015. Estimación poblacional y conservación de felinos (Carnivora: Felidae) en el norte de Quintana Roo, México. Revista de Biologia Tropical, 63(3): 799–813. Caso, A., López–González, C., Payan, E., Eizirik, E., de Oliveira, T., Leite–Pitman, R., Kelly, M., Valderrama, C., 2008. Panthera onca. In: The IUCN Red List of Threatened Species 2008. Ceballos, G., Oliva, G., 2005. Los mamíferos silvestres de México. Conabio/Fondo de Cultura Económica, D.F. México. Céspedes–Flores, S., Moreno–Sánchez, E., 2010. Estimación del valor de la pérdida de recursos forestales y su relación con la reforestación en las entidades federativas de México. Investigación ambiental, 2(2): 5–13. Chávez, C., 2010. Ecología y conservación del jaguar (Panthera onca) y puma (Puma concolor) en la región de Calakmul y sus implicaciones para la conservación de la Península de Yucatán. PhD thesis, Universidad de Granada. Chávez, C., Zarza, H., Ceballos, G., Amín, M., 2007. Ecología poblacional del jaguar y sus implicaciones para la conservación en la Península de Yucatán, Análisis de viabilidad de poblaciones y hábitat del jaguar en México. In: Conservación y Manejo del Jaguar en México estudios de caso y perspectivas de conservación: 101–110 (G. Ceballos, C. Chávez, R. List, H. Zarza, Eds.). Alianza WWF/Telcel, CONABIO, CONANP, Ecociencias S.C. Distrito Federal. Contreras–Moreno, F. M., De la Cruz–Félix, K., Bello–Gutiérrez, B., 2012. Uso Patrones de Cacería y Preferencia de Presas en dos sitios del Parque

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Bio–ecology of Potamon algeriense (Herbst, 1785) (Crustacea, Decapoda) in eastern Morocco A. F. Taybi, Y. Mabrouki, A. Berrahou, A. A. El Abd

Taybi, A. F., Mabrouki, Y., Berrahou, A., El Abd, A. A., 2018. Bio–ecology of Potamon algeriense (Herbst, 1785) (Crustacea, Decapoda) in eastern Morocco. Animal Biodiversity and Conservation, 41.2: 267–274. Abstract Bio–ecology of Potamon algeriense (Herbst, 1785) (Crustacea, Decapoda) in eastern Morocco. To contribute to the knowledge of Potamon algeriense bio–ecology in eastern Morocco and the Moulouya watershed, a total of 90 stations were surveyed between 2013 and 2016. Of these, only 14 stations (out of the 90 surveyed) were positive concerning the occurrence of this species, which was found to be limited to the middle and lower watercourses of the Moulouya watershed, and specifically to Oued Za, Zegzel and the lower watercourse of the sub–basin Mouth of the Moulouya. The results of the statistical analyses showed that the main factors influencing the distribution and abundance of Potamon algeriense were temperature, BOD5, and conductivity. Key words: Anthropogenic activities, Freshwater crab, eastern Morocco, Moulouya Resumen Bioecología de Potamon algeriense (Herbst, 1785) (Crustacea, Decapoda) en Marruecos oriental. Para contribuir al conocimiento de la bioecología de Potamon algeriense en Marruecos oriental y la cuenca hidrográfica del río Muluya, se estudiaron 90 estaciones entre 2013 y 2016, de las que solo 14 albergaban a esta especie, que se observó se encontraba limitada a los cursos de agua medio e inferior de la cuenca hidrográfica del río Muluya y más concretamente a las zonas de Oued Za, Zegzel y el curso inferior de la subcuenca de la desembocadura del río Muluya. Los resultados de los análisis estadísticos mostraron que los principales factores que influyen en la distribución y la abundancia de Potamon algeriense fueron la temperatura, la BOD5 y la conductividad. Palabras clave: Actividades antropogénicas, Cangrejo de agua dulce, Marruecos oriental, Muluya Received: 10 VII 17; Conditional acceptance: 14 X 17; Final acceptance: 16 XI 17 Abdelkhaleq Fouzi Taybi, Youness Mabrouki, Ali Berrahou, Abdelaziz Ait El Abd, Lab. Sciences de l’eau, l’environnement et du Développement Durable, Dépt. de Biologie, Fac. des Sciences, Univ. Mohamed Premier, B. P. 524, 60000 Oujda, Maroc. Corresponding author: Youness Mabrouki. E–mail: younes_mab@hotmail.fr; y.mabrouki@ump.ac.ma

ISSN: 1578–665 X eISSN: 2014–928 X

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Introduction The Potamoidea Ortmann, 1896 superfamily includes two families, namely, Potamidae Ortmann, 1896 and Potamonautidae Bott, 1970. In North Africa, each one is represented by one genus, Potamon Ortmann, 1896, and Potamonautes MacLeay, 1837, respectively (Cumberlidge, 1999, 2009; Brandis et al., 2000). The family Potamidae comprises 95 genera and more than 505 species that are distributed throughout the southern Palaearctic and eastern zoogeographic regions (Cumberlidge et al., 2008, Yeo et al., 2008). The representative of this family in North Africa and the Maghreb is Potamon algeriense (Herbst, 1785). Bott (1967) established this taxon as the subspecies Potamon fluviatilis algeriensis, and Pretzmann (1976) treated it as the subspecies Potamon fluviatilis berghetripsorum. This species was later revised to Potamon (Eutelphusa) algeriense by Brandis et al. (2000), and it is treated here as Potamon algeriense following the opinion of Cumberlidge (1998). The species is included in the subfamily Potaminae Ortmann, 1896, whose members are located around the Mediterranean, the Middle East and the Himalayas (Cumberlidge, 2010). In fact, P. algeriense represents the most western extension of this subfamily in North Africa. It is found in the temperate rivers of the Maghreb and in seasonal dry–water bodies in arid climates where crabs tend to be semi–terrestrial and live in burrows (Bott, 1967; Brandis et al., 2000). The species is found only in the northwestern African region, specifically in Morocco, Algeria and Tunisia (Cumberlidge, 2009). In Morocco, the species has been recorded from the Rif in the watershed of the Oued Laou near Chefchaouen, from the Middle Atlas in the Oued Oum Rbia basin (Aymerich, 2002), and from Eastern Morocco (García et al., 2010). The species seems to inhabit almost any type of freshwater habitat, including springs and their outfalls, streams, and mountain rivers at altitudes ranging between 9 and 875 m. They are usually found under stones or dead logs, in full water, or in the riparian area. The burrows are dug in the muddy soil and can be 30–40 cm deep with water at the bottom. Younger specimens are also present in the water. Adults can be found far from the water, in dried river beds and even in the forest (Beni Snassen) at night. In seasonal rivers such as Oued Cherraa, the species tends to be semi–terrestrial, and during drought, it digs galleries near watercourses (Bott, 1967; Brandis et al., 2000). P. algeriense is an omnivore: its diet consists of invertebrates, earthworms, crustaceans (including small individuals of the same species), cadavers, and detritus of any kind. Adult individuals occupy a more favorable area for availability of food than juveniles who rely on small prey such as batrachian eggs, fish and vegetable waste. Despite the wide distribution of P. algeriense in the Maghreb, its distribution is discontinuous and fragmented. The decrease in its populations in Morocco is worrying. Indeed, this species has not been seen for many years in Oued Sebou (Provinces of Fez and Kenitra) (Cumberlidge, 2010). P. algeriense is threatened with

extinction due to anthropogenic activity by industrial and domestic liquid waste (García et al., 2010). The main aim of this manuscript was to contribute to the knowledge of the bio–ecology of P. algeriense in the watershed of Moulouya and Eastern Morocco, to update the information on its geographical distribution in the studied region, and to highlight the species preferences in terms of type of habitat and the physico–chemical quality of the water. Material and methods Study area Morocco is currently divided, according to the new administrative division, into 12 regions including the Oriental Region (fig. 1A), which occupies the entire eastern side of the country and covers an area of 88,681 km2. This area is bounded to the north by the Mediterranean Sea, to the east and south by the Morocco–Algerian border, and to the west by the administrative regions of Tangier–Tetouan, Al Hoceima, Fez Meknes, and Draa–Tafilalt. The Oriental region includes the wilaya of Oujda (Oujda–Angad prefecture) and the provinces of Berkane, Taourirt, Jerada, Nador, Figuig, Driouch, and Guercif. The watershed of the Moulouya (74,000 km2) (fig. 1A), which includes nearly 44,000 km2 to the east of Morocco, covers much of the Oriental region. It is located between parallels 36 and 39 degrees north and the meridians 5.5 and 7 degrees west. With a length of 600 km, the Moulouya is the largest North African river flowing into the Mediterranean. It starts at the junction of the High and Middle Atlas chains, and flows primarily along a southwest–northeast axis. Its main tributaries, Anzegmir Wadi, Melloulou Wadi and Za Wadi, are perennial, while others flow only during floods (3–5 floods on average per year). The river flows through various Mediterranean bioclimatic zones (Berrahou et al., 2001). Surveys The field surveys were carried out between 2013 and 2016. To update the rare data on the bio–ecology of P. algeriense in Moulouya’s watershed and eastern Morocco, we prospected 45 stations along the Moulouya and its respective tributaries, Oued Anzegmir, Oued Melloulou and Oued Za. These surveys were supplemented by data from more than 45 other stations spread throughout eastern Morocco, from Nador, Saida and Béni Snassen in the north, to Figuig in the southeast, and Talessint and Bouanane in the southwest. Sampling for P. algeriense was carried out by a turbid net in aquatic environments, while the search for adults outside the water was carried out by hand or with the aid of clamps. These operations took about one hour of excavation at each station. All captured specimens were released into their natural habitats after being sexed and measured, and after additional information was taken.

Animal Biodiversity and Conservation 41.2 (2018)

2º 0' 0'' W

35º0'0''N Atlantic Ocean

râ

Lo w er

Zegzel

1º 0' 0'' W

Algeria

4º 0' 0'' W 3º 0' 0'' W Mediterranean Sea uy a

5º 0' 0'' W

269

34º 0' 0'' N

lo el

uy a

33º 0' 0'' N

Station prospected Oriental Region Moulouya watershed

32º 0' 0'' N B

120 km

5º 0' 0'' W

4º 0' 0'' W

3º 0' 0'' W

2º 0' 0'' W

1º 0' 0'' W

5º 0' 0'' W

4º 0' 0'' W

3º 0' 0'' W

2º 0' 0'' W

1º 0' 0'' W

Mediterranean Sea 35º 0' 0'' N Atlantic co Ocean roc o M

Algeria

M Lo ou w lo er uy a

râ a Zegzel

34º 0' 0'' N

uy a

Lo w er

ou lo uy a M

33º 0' 0'' N

S14 r râ S11 a S10 Zegzel S15 S16

ou lo

120 km

S13 S12

lo el

20 km

S7 S6

32º 0' 0'' N

Stations Oriental Region Moulouya watershed 5º 0' 0'' W

4º 0' 0'' W

3º 0' 0'' W

S4 S3 S2 S1

2º 0' 0'' W

1º 0' 0'' W

Fig. 1. A, studied localities of the Oriental Region of Morocco and the Moulouya River basin; B, distribution area of Potamon algeriense in the study area. Fig. 1. A, localidades estudiadas de la región oriental de Marruecos y la cuenca hidrográfica del río Muluya; B, área de distribución de Potamon algeriense en la zona de estudio.

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In order to contribute to the study of its diet, and in addition to our harvested field data, we kept an adult couple and two juveniles in conditioned tanks in the laboratory. They were fed on various resources found in their living environment. The animals were released into their natural habitats once the study was completed. Water sampling From each station, two replicates of water samples were taken in polyethylene bottles of 500 ml (the results of which are their averages). The water samples were preserved with 2 ml of concentrated hydrochloric acid (pH = 2). According to standard norms ISO 5667–6 (1990), ISO 5667–2 (1991) et ISO 5667–3 (1994), the water samples were conveyed to a cooler at a low temperature (± 4 °C) to stop metabolic activity of organisms in the water. The following physico–chemical parameters were measured in the laboratory: sulfates (SO42–), biological oxygen demand after five days (BOD5), orthophosphates (PO43–), ammonium (mg N–NH4),and nitrates (mg N–NO3). These parameters are determined according to the standards AFNOR (1997) and Rodier et al. (1996). Conductivity, pH, dissolved oxygen, and temperature were measured (in situ) in the field. Statistical processing of data Statistical analyses were carried out using software R in version 3.3.1. To determine the ecological factors governing the distribution and abundance of P. algeriense, we subjected a matrix of nine abiotic parameters, representing the average of the different descriptors measured in the 16 stations of the lower Moulouya river (annex 1), to a normed principal component analysis. Results and discussion Distribution and habitat types In the Moulouya watershed and in eastern Morocco, the distribution of P. algeriense was limited to the sub–watersheds of Oued Za from S1 to S9 (High Plateaux), the Oued Zegzel–Cherrâa complex from S14 to S16 (Beni Snassen), and Lower Moulouya from the two dams to the pre–mouth area from S10 to S13 (see fig. 1B), where healthy populations were detected and adults reached a maximum carapace length of 50 millimetres. Analysis of the physical–chemical parameters of the different stations sheltering P. algeriense Of the 16 stations selected, P. algeriense occurred in all but two (stations S7 and S13). Annex 1 presents the nine physico–chemical variables measured at each of the 16 stations:

Apart from a slight alkalinity observed in S11, pH values showed neutrality in most stations (annex 1). This is probably due to the presence of carbonates that buffer the waters flowing to the Moulouya Wadi, streaming and infiltrating into the maro–dolomitic and calcareous cover (Taybi, 2016). In the wadi of the Moulouya, the temperature of the water is linked to local conditions (climate, duration of sunshine, flow and altitude) and to seasonality (Taybi, 2016). During the study period, average temperatures do not exceed 23 °C. Given the lack of sampling during winter, our study does not reveal the lowest temperature at which the crab remains active. The conductivity of the Moulouya waters fluctuates in time and space, and generally increases from upstream to downstream (Taybi, 2016). The stations with a positive occurrence of the species studied were characterized by acceptable conductivity values, fluctuating between 305 μs.cm–1 recorded in upstream Oued Zeghzel and about 1,700 μs.cm–1 downstream from Oued Moulouya. The waters of the positive stations for P. algeriense had high to acceptable levels of dissolved oxygen. A maximum of 11.5 mg.l–1 was recorded at station S14, which is a rheocrenous source very rich in macrophytes. The values recorded in major ions (NH4, NH3, SO4 and PO4) and BOD5 in the waters of the studied stations (annex 1) make it possible to place these waters in general in the grid of good quality (ABHM, 2012), with the exception of the station S6 which recorded relatively high values. Indeed, this station located downstream of Oued Za receives discharges of domestic and industrial waste water from peripheral agglomerations (Mabrouki et al., 2016a; Bensaad et al., 2017). The freshwater crab thus appears to be relatively demanding in terms of the physico–chemical quality of the water (fig. 2). Thus, in the Moulouya watershed and in Eastern Morocco, the species is present only in richly oxygenated and weakly mineralized neutral waters, with low concentrations of major ions and BOD5. Concerning temperature, it is difficult to accurately judge its interval in the study area since no sampling was done during the flood season, but the species seems to prefer temperatures cooler than 23 °C. Statistical analysis The principal component analysis of the different mesological parameters of the prospected stations shows that the first two axes, F1 and F2 (fig. 3), hold most of the information since they represent 75.14 % of the total inertia. Examination of the correlations between the axes and the various mesological components studied explains the significance of each axis in the structured distribution of the station cloud and the relation between abundance and environmental variables. The F1 axis (62.75 % of total inertia) was negatively correlated with the pollution parameters (BOD5, r = –0.91) and positively with dissolved oxygen (r = 0.87). The axis F1 therefore expresses a gradient

Animal Biodiversity and Conservation 41.2 (2018)

10 9 8 7 6 5 4 3 2 1 0

25 20 15 10 5 pH

Temp (ºC)

2,000 10 1,800 9 1,600 8 1,400 7 1,200 6 1,000 5 800 4 600 3 400 2 200 1 0 0 Cond (µs.cm–1) Dis_O (Mg.l–1)

1.4

350

1.2

300

250

1 0.8

150

0.4

100

0.2 0

200

0.6

N_NH (mg.l–1)

271

50 NO (mg.l–1)

SO (mg.l–1)

2 1.8 1.6 1.4 1.2 1 0.8 0.6 0.4 0.2 0

12 10 8 6 4 2 0

BOD5 (mg.l–1)

Min

Max

PO (mg.l–1)

Fig. 2. Graphs showing the tolerance interval of Potamon algeriense in relation to the abiotic descriptors studied. Fig. 2. Gráficos del intervalo de tolerancia de Potamon algeriense en relación con los descriptores abióticos estudiados.

of pollution by separating the most polluted station to the left (S7: on Oued Za River) and the most oxygenated stations to the right. The total absence of P. algeriense in station S7 is certainly due to pollution caused by the liquid and industrial discharges that this station receives from the province of Taourirt (Mabrouki et al., 2016a; Bensaad et al., 2017). The F2 axis (12.39 % of the total inertia) respectively translates two environmental gradients, temperature and conductivity, and isolates the station closest to the Mediterranean, S13. The relatively high salinity of this station explains the absence of P. algeriense in S13. The maximum abundance of P. algeriense was recorded in low conductivity waters (S14, S15 and S16). The correlation ratio of the environmental parameters of the environment (fig. 3D) with the abundance of P. algeriense shows that the latter is positively correlated with dissolved oxygen (r = 0.75), and negatively with BOD5 (r = –0.62), conductivity (r = –0.62) and temperature (r = –0.59). The results of the analyses show that P. algeriense prefers oxygenated water, and that high conductivity and pollution are among other factors limiting its abundance.

Conclusion Despite efforts to conserve the aquatic ecosystems of the Moulouya watershed and eastern Morocco, severe degradation continues to accelerate. This deterioration is becoming more and more worrying because of the multiplication of sources of pollution of domestic, industrial, and agricultural origin (Taybi et al., 2016a; Mabrouki et al., 2016a, in press a; Bensaad et al., 2017; Ramdani et al., 2017; Yahya et al., 2017). This anthropogenic activity, aggravated and accentuated by episodes of drought, results in a loss of aquatic biodiversity (Berrahou et al., 2001; Millán et al., 2016; Taybi, 2016; Taybi et al., 2016b; Mabrouki et al., 2016b, 2017b, in press a, in press b). The results of this study show that P. algeriense is restricted to the northeastern part of the study area, the Moulouya watershed, where it prefers oxygenated water with a low to medium conductivity. Our results also strongly suggest that the large reduction in its regional range is caused by anthropogenic pollution. With these unmeasured anthropogenic activities, the regional distribution of many species has declined sharply or disappeared (Taybi, 2016; Mabrouki, 2017; Taybi et al., 2017a, 2017b). However, further surveys

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d = 2 Abun PO NO N_NH BOD5

SO S9

S10

Abun

S14 S15 S16

N_ D_

S12 S3 S2

BOD5

SO Dis_O

S8 S11 S13

Con

Eigenvalues 62.75 % LT = 10

12.30 %

Temp

First factor BOD5 PO SO NO N_NH Dis_O Cond Temp pH Abun –1.0

Correlation with abundance

–0.91

–0.72 –0.64

Temp

BOD5 PO SO NO N_NH Dis_O Cond Temp pH

0.87

0.71 –0.5 0.0 0.5 Correlation ratio

1.0

–1.0

–0.62

–0.62 –0.59

0.75

–0.5 0.0 0.5 Correlation ratio

1.0

Fig. 3. A, biplot of the principal component analysis (PCA) in plane F1–F2 presenting the physical–chemical parameters and the stations; B, correlation circle of the PCA; C, correlation relationship between the physical–chemical parameters and the first axis of the PCA; D, correlation relationship between physical– chemical parameters and abundance. Fig. 3. A, diagrama de dispersión biespacial (biplot) del análisis de componentes principales (ACP) en el plano F1–F2 en el que se presentan los parámetros fisicoquímicos y las estaciones; B, círculo de correlaciones del ACP; C, correlación entre los parámetros fisicoquímicos y el primer eje de ACP; D, correlación entre los parámetros fisicoquímicos y la abundancia.

targeting inaccessible areas could increase the known range occupied by P. algeriense in eastern Morocco and the basin of Moulouya River. Indeed, a new survey carried out in 2017 not far from the town of Guercif (34º 13' N – 3º 19' W) has allowed to expand the distribution of the species in the study area. References ABHM (Agence du Bassin Hydraulique de la Moulouya), 2012. Etat de la qualité des ressources en eau dans le bassin de la Moulouya 2011–2012. ABHM, Oujda.

AFNOR (Association française de Normalisation), 1997. Qualité de l’eau. Recueil des Normes Françaises Environnement, Tomes 1, 2, 3 et 4. AFNOR, Saint Denis. Aymerich, M., 2002. Carnets de Voyages Naturalistes au Maroc. http://geos nature.org/carnet_07_08_2002.html Bensaad, H., Mabrouki, Y., Taybi, A. F., Chafi, A., 2017. Assessment of wastewater discharges from Taourirt City on the water quality of the Oued Za (Eastern Morocco). Journal of Materials and Environmental Science, 8(7): 2365–2371. Berrahou, A., Chavanon, G., Belloulali, A., Richoux, P., 2001. Études sur la Basse Moulouya (Maroc

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oriental): 7 – Les Coléoptères aquatiques de l’oued Za. Bulletins mensuels–Société linnéenne de Lyon, 70(5): 127–131. Bott, R., 1967. Potamidae (Crustacea, Decapoda) aus Afghanistan, Westasien und dem Mittelmeerranum (Eine Revision der Untergattung Potamon s str.). Videnskabelige Meddelelser fra Dansk Naturhistorisk. Foren.: 7–43. Brandis, D., Storch, V., Türkay, M., 2000. Taxonomy and zoogeography of the freshwater crabs of Europe, North Africa and the Middle East. Senckenbergiana biologica, 2: 5–56. Cumberlidge, N., 1998. The African and Madagascan freshwater crabs in the Zoologische Staatssammlung, Munich (Crustacea, Decapoda, Brachyura, Potamoidae). SPIXIANA, 21(3): 193–214. – 1999. The freshwater crabs of West Africa. Family Potamonautidae. In: Faune et Flore Tropicales, 35: 1–382. Institut de recherche pour le développement–IRD (ex–ORSTOM), Paris. – 2009. Freshwater Crabs and Shrimps (Crustacea: Decapoda) of the Nile Basin. In: The Nile. Origin, Environments, Limnology and Human, Use, chapter 27: 547–561 (H. J. Dumont, Ed.). Monographiae Biologicae, 89. Springer, New York. – 2010. The status and distribution of freshwater crabs, In: The Status and Distribution of Freshwater Biodiversity in Northern Africa, 6: 71–78 (N. García, A. Cuttelod, D. Abdul Malak, Eds.). IUCN, Gland, Switzerland, Cambridge, UK, and Spain. Cumberlidge, N., Daniels, S. R., Sternberg, Rv., 2008. A revision of the higher taxonomy of the Afrotropical freshwater crabs (Decapoda: Brachyura) with a discussion of their biogeography. Biological Journal of the Linnean Society, 93: 399–413. García, N., Cuttelod, A., Abdul Malak, D. (Eds.), 2010. The Status and Distribution of Freshwater Biodiversity in Northern Africa. IUCN, Gland, Switzerland, Cambridge, UK, and Malaga, Spain. Mabrouki, Y., 2017. Comparative study of the longitudinal distribution of the benthic invertebrates of the two catchments: Za and Melloulou (tributaries of the Moulouya). PhD Thesis, Univ. Mohamed Premier, Oujda Morocco (in French). Mabrouki, Y., Taybi, A. F., Bensaad, H., Berrahou, A., 2016a. Variabilité spatio–temporelle de la qualité des eaux courantes de l’Oued Za (Maroc Oriental). Journal of Materials and Environmental Science, 7(1): 231–243. Mabrouki, Y., Taybi A. F. & Berrahou, A. (in press a). L’évolution spatio–temporelle de la qualité des eaux courantes de l’Oued Melloulou (Maroc). Revue des sciences de l’eau, 30.2. Mabrouki, Y., Taybi, A. F., Chavanon, G., Vinçon, G., Berrahou, A., 2016b. Contribution à l’étude des plécoptères dans le Maroc Oriental et le bassin versant de la Moulouya et leur distribution en fonction des étages bioclimatiques. Journal of Materials and Environmental Science, 7(6): 2178–2193. Mabrouki, Y., Taybi, A. F., El Alami, M. & Berrahou,

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A., 2017b. New and interesting data on distribution and ecology of Mayflies from Eastern Morocco (Ephemeroptera), Journal of Materials and Environmental Science, 8(8): 2832–2859. Mabrouki, Y., Taybi, A. F., Berrahou, A., Chaabane, K., Legssyer, B. (in press b). Case study of the freshwater shrimp Atyaephyra desmarestii (Millet, 1831) (Crustacea, Decapoda) in the watershed of Moulouya and Eastern Morocco. Vie et Milieu. Millán, A., LMohdi, O., Antonio, Carbonell, J., Taybi, A. F., Dakki, M., 2016. A new species of Aphelocheirus (Hemiptera: Heteroptera: Aphelocheiridae) from Morocco. Zootaxa, 4173(6): 577–582. Pretzmann, G., 1976. Fluviatilis berghetripsorum nov. subspec, eine neue Süsswasserkrabbe aus Marokko. Ann. naturhist. Mus. Wien, 80: 451–452. Ramdani, M., Taybi, A. F., Mabrouki, Y., Haloui, B., El Asri, O., Elmsellem, H., El Khiati, N., Mostareh, M., 2017. The spatial variability of water quality in the Mediterranean of eastern Morocco. Moroccan Journal of Chemistry, 5(2): 227–235. Rodier, J., Bazin, C., Broutin, J. P., Chambon, P., Champsaur, H., Rodi, L., 1996. L’analyse de l’eau, 8ème édition. Edition Dunod, Paris, France. Taybi, A. F., 2016. Hydrobiological study of the Moulouya: Structure of the biodiversity and longitudinal zonation of benthic invertebrates. PhD Thesis, Univ. Mohamed Premier, Oujda, Morocco (in French). Taybi, A. F., Mabrouki, Y., Berrahou, A., Chaabane, K., 2016a. Évolution spatiotemporelle des paramètres physicochimiques de la Moulouya. Journal of Materials and Environmental Science, 7(1): 272–284. Taybi, A. F., Mabrouki, Y., Berrahou, A., Peris–Felipo, F. J., Chaabane, K., 2016b. Contribution à l’étude de la relation «plante–hôte–parasite» entre Elodea canadensis Michx., Hydrellia sp. (Diptera) et Ademon decrescens (Nees, 1811) (Hymenoptera, Opiinae) dans le bassin versant de la Moulouya (Maroc), Journal of Materials and Environmental Science, 7(7): 2445–2452. Taybi, A. F., Mabrouki, Y., Chavanon, G., Millán, A., Berrahou, A., 2017b. New data on aquatic beetles of Morocco (Coleoptera Adephaga: Gyrinidae, Haliplidae and Dytiscidae). Baltic Journal of Coleopterology, 17(1): 83–106. Taybi, A. F., Mabrouki, Y., Ghamizi, M., Berrahou, A., 2017a. The freshwater malacological composition of Moulouya’s watershed and Oriental Morocco. Journal of Materials and Environmental Science, 8(4): 1401–1416. Yahya, H. S. A., Taybi, A. F., Mabrouki, Y., Fahsi, A., Chafi, A., Chafik, Z., 2017. The Metallic pollution in the groundwater of Triffa Plain (Eastern Morocco), Journal of Materials and Environmental Science, 8(9): 3372–3381. Yeo, D. C. J., Ng, P. K. L., Cumberlidge, N., Magalhaes, C., Daniels, S. R., Campos, M., 2008. Global diversity of crabs (Crustacea: Decapoda: Brachyura) in freshwater. In: Freshwater Animal Diversity Assessment. Hydrobiologia, 595: 275–286.

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Annex 1. The physical–chemical parameters of the studied stations (between 2013 and 2016): L, localities; Alt, altitude (in m); Abun, abundance; pH, hydrogen potential; T, water temperature (in ºC); Cond, conductivity; Dis_O, oxygen dissolved in water; NH4 and NH3, ammonia; NO, nitrates; SO, sulfates; PO, orthophosphates; BOD5, biological oxygen demand. Anexo 1. Los parámetros fisicoquímicos de las estaciones estudiadas (entre 2013 y 2016): L, localidades; Alt, altitud (en m); Abun, abundancia; pH, potencial de hidrógeno; T, temperatura del agua (en ºC); Cond, conductividad; O_dis, oxígeno disuelto en agua; NH4 y NH3, amoníaco; NO, nitratos; SO, sulfatos; PO, ortofosfatos; BOD5, demanda biológica de oxígeno.

GPS

37º 7' 5.7'' N

2º 5' 26.8'' W

Alt Abun

Cond

Dis_O

N_NH 0.06

875

7.9 21.2 750

7.1

34º 14' 31.61'' N 785

8.3

7.15 0.025 2.71

0.019 6.15

2º 20' 11.98'' W

34º 13' 36.8'' N

767

7.4 20.3 640

7.11

1.27

0.013 5.05

2º 23' 34.5'' W

34º 14' 21.6'' N

750

7.3 16.5 890

7.71 0.014 3.27

0.01

2º 24' 34.8'' W

34º 12' 23.1'' N

625

7.3

17 1,237

7.4

0.019 2.17

2º 38' 52.3'' W

34º 25' 15.6'' N

370

7.93

21 1,655 5.21 1.266 51.55 317

1.815 13.9

2º 52' 52.9'' W

S7 34º28'44.51''N 295 0 S8 S9

34º 28' 44.51'' N 295

34º 53' 11'' N

2º 39' 45'' W

S11 34º 54' 27.53'' N 35º 3' 5.7'' N 2º 25' 42.4'' W

S13

35º 5' 51.4'' N

7.11

23 1,750 6.75 0.127 2.77

0.029 10.55

7.39 19.1 1,320 7.75 0.029 14.95 300

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8.31 19.5 1,655 7.65 0.095

0.91

8.9 21.4 1,697 8.75 0.095 2.89 219

0.021 0.43

7.7 21.1 1,700 7.95 0.097 9.75

0.775 2.55

7.83 23.3 2,290 8.55 0.015 9.87

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7.5

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2º 23' 19'' W

S14

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2º 25' 36.0'' W

S15 34º 53' 08.3'' N 2º 20' 34.1'' W

S16 34º 50' 20.3'' N

8.3 24 2,280 2.92 2.075 51.75 339 1.985 29.25

2º 38' 8.86'' W

0.027 2.75

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34º 33' 41.09'' N 222

S10

0.01

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S12

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3.15 160

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Evaluation of the expansion of Mantis religiosa (L.) in Poland based on a questionnaire survey D. Zieliński, C. J. Schwarz, R. Ehrmann

Zieliński, D., Schwarz, C. J., Ehrmann, R., 2018. Evaluation of the expansion of Mantis religiosa (L.) in Poland based on a questionnaire survey. Animal Biodiversity and Conservation, 41.2: 275–280. Abstract Evaluation of the expansion of Mantis religiosa (L.) in Poland based on a questionnaire survey. Mantis religiosa (L.) is the only species of praying mantis in Poland. The main habitat where its occurrence is permanent and confirmed is the Sandomierska Basin (SE Poland). Numerous reports suggest a significant dispersion of these insects in Poland. Therefore, we decided to use and collect information to update the knowledge about the spread of the praying mantis in Poland. With the use of Google Maps, we created a map on which respondents were asked to select the location where they encountered a M. religiosa specimen in Poland in recent years (2013–2016). In total, we obtained 159 locations for the European mantis. These findings show the significant spread of this species in all directions from its main habitat in Sandomierska Basin. However, there is a need for more studies on this topic, especially to confirm the existence of reproductive populations in the provided locations and to confirm the existence of two different M. religiosa lineages in Poland. Key words: Praying mantis, Mantis religiosa, Poland, Distribution, Questionnaire survey Resumen Evaluación de la expansión de Mantis religiosa (L.) en Polonia basada en una encuesta. Mantis religiosa (L.) es la única especie de mantis religiosa en Polonia. El hábitat principal en el que se ha confirmado su presencia permanente es la cuenca de Sandomierska (SE de Polonia). Numerosos informes sugieren una dispersión significativa de estos insectos en Polonia. Por lo tanto, decidimos utilizar y recopilar información para actualizar el conocimiento sobre la propagación de M. religiosa en Polonia. Con el uso de Google Maps, se creó un mapa sobre el que se pidió a los encuestados que seleccionaran el lugar donde hubieran encontrado un espécimen de mantis en Polonia en los últimos años (2013–2016). En total, se obtuvieron 159 lugares para la mantis europea. Estos resultados muestran la dispersión significativa de esta especie en todas las direcciones desde su hábitat principal en la cuenca de Sandomierska. Sin embargo, es necesario realizar más estudios sobre este tema, especialmente para confirmar la existencia de poblaciones reproductivas en las localizaciones proporcionadas y de dos diferentes linajes de M. religiosa en Polonia. Palabras clave: Mantis, Mantis religiosa, Polonia, Distribución, Encuesta mediante cuestionario Received: 11 VIII 17; Conditional acceptance: 17 X 17; Final acceptance: 13 XII 17 Damian Zieliński, Dept. of Ethology and Animal Welfare, Fac. of Biology, Animal Sciences and Bioeconomy, Univ. of Life Sciences in Lublin, Akademicka 13, 20–950 Lublin, Poland.– Christian J. Schwarz, Conservation Biology Unit, Dept. of Biology and Biotechnology, Ruhr Univ. Bochum, ND 1/31, D–44780 Bochum, Germany.– Reinhard Ehrmann, Neuer Weg 2, D–76228 Karlsruhe, Germany. Corresponding author: D. Zieliński. E–mail: damian.zielinski@up.lublin.pl

ISSN: 1578–665 X eISSN: 2014–928 X

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Introduction Polish fauna contains only one species of praying mantis, the European Mantis, Mantis religiosa (L.). The 'European' mantis is distributed throughout most of the Old World. Besides Europe, it occurs in Africa, Asia, and most of the Sunda archipelago, crossing 55° N in parts of Russia (Bolshakov et al., 2010). Older records of the species occurring in Australia, Bolivia, and the Caribbean, carried over in the literature for more than a century have been shown by Berg et al. (2011) to be erroneous. However, M. religiosa was introduced in Rochester, New York in 1899 (Berg et al., 2011), and the species is widely distributed across North America today (Gurney, 1950; Cannings, 2007; Berg et al., 2011). The species has also recently shown a significant northward spread in several European countries, such as Germany (Berg et al., 2008, 2011; Stärz et al., 2010; Ehrmann, 2011; Landeck et al., 2013; Linn and Griebeler, 2015; Schwarz et al., 2017), France (Voisin, 2003), the Czech Republic (Piszkiewitz et al., 2000; Hanac and Hudeček, 2001; Tichá, 2005; Janšta et al., 2008; Gruchala, 2010; Holuša et al., 2012; Chobot, 2016), Slovakia (Kočárek et al., 1999; Krištín et al., 2004; Krištín and Hrúz, 2005; Fedor, 2007; Fedor et al., 2010), Latvia (Pupiņš et al., 2012), Russia (Bolshakov et al., 2010; Shcherbakov and Savitsky, 2015), Ukraine (Nagy et al., 2011), Croatia (Romanowski and Romanowski, 2014), Hungary (Nagy and Sziráki, 2002; Nagy and Kisfali, 2007) and Poland (Sępioł, 2005; Buczyńska et al., 2006; Liana, 2007, Bonk and Kajzer, 2009; Bonk et al., 2011; Ćwik et al., 2012; Kozina, 2015; Błoński, 2015; Zieliński and Łazarecki, in press). Historically, the Polish population has been assigned to its own subspecies, Mantis religiosa polonica Bazyluk, 1960, due to differences in morphology and postembryonic development from southern and central European nominative populations (Bazyluk, 1960). Its subspecific status, however, has not been universally accepted (see Berg et al., 2011; Schwarz et al., 2017 for a review). M. religiosa is protected by national law in Poland (Ordinance of the Minister of the Environment of 16 December 2016 on protection of endangered species. Dz. U. from 2016 pos. 2183). Its main area of occurrence is located in the Sandomierska Basin macro–region (SE Poland), especially the Sandomierskie Forest, Janowskie Forests and Lipskie Forest in the southeastern part of Poland (Liana, 2007). However, there are many reports about new sites in other parts of country (Bonk and Kajzer, 2009; Bonk et al., 2011; Ćwik et al., 2012; Kozina, 2015; Zieliński, 2016). Moreover, every year during summer, in early August, accidental findings of praying mantises in different parts of Poland are published on Polish websites. During this time, M. religiosa undergoes its final molt, after which the specimens gain the ability to fly. Volant specimens are attracted by light and often show up inside houses or apartments. Usually, people do not know what kind of insect they are faced with; therefore, the encounter is an interesting event worth posting online (Zieliński et al., 2017).

The aim of this study was to use the knowledge of Polish people about the occurrence of the praying mantis to compile an online distribution map and to evaluate its expansion in Poland. Material and methods We created an editable map of Poland using Google Maps® software. Respondents were asked to pinpoint the location where they found a M. religiosa specimen in Poland in recent years (2013–2016). To achieve this, we sent links to the map via social networking sites and online forums about wildlife and nature to people agreeing to assist in data collecting. Volunteers were also asked to provide information about the mantis they have encountered, with regard to stage (nymph, adult or ootheca), color (green or brown), sex (male or female), and location. Based on the information provided, the distributional records were plotted onto a UTM grid. To emphasize the different metapopulations of M. religiosa in different parts of Poland, their anticipated occurrence was drawn as ellipses. The ellipses were made according to the possible spread of the species on the basis of data provided by respondents and scientific data. Results and discussion Based on the replies from 283 respondents, 159 locations of the European Mantis in Poland were obtained (fig. 1). M. religiosa is a protected species by national law in Poland as well as in Belgium, Germany, Luxembourg, Switzerland, Austria, the Czech Republic, Slovakia, and Hungary (Berg et al., 2011; Schwarz et al., 2017). However, due to its extensive range and higher abundances in climatically more favorable regions, it is listed in the IUCN Red List of Threatened Species in the 'Least Concern' category (Battiston, 2016). Additionally, in Poland this species is listed in the 'Red List of Threatened and Endangered Animals' (Głowaciński, 1992a; 2002) and in two editions of the 'Polish Red Book of Animals' (Głowaciński, 1992b; Głowaciński and Nowacki, 2004), initially with the status CR (critically endangered), later reduced to EN (species of very high risk). Despite having been considered endangered or critically endangered at the northern edge of its distribution for the best part of last century, its recent rapid spread in North America and Central Europe shows that the species is actually a good disperser, able to rapidly invade new localities when environmental conditions are favorable (Pupiņš et al., 2012; Schwarz et al., 2017). Perhaps, if the current expansive trend continues, the conservation status of this species in the Polish law will have to be reconsidered. The main sources of information about the recent distribution of Mantis religiosa in Poland are the paper of Liana (2007) and several smaller papers and popular articles (e.g. Bonk and Kajzer, 2009; Królik, 2010; Bonk et al., 2011; Kozina, 2015; Kozina and Łopucki, 2016) which reported findings of single mantises in

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VU VT

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YS BB

XR YR BA

GA KR

Fig. 1. Map of Poland showing Polish volunteers' pinned locations of Mantis religiosa encounters on a UTM grid of 10 x 10 km squares (created with GNOMON 3.1). Fig. 1. Mapa de Polonia en el que se muestran los lugares en los que los voluntarios polacos indicaron haber encontrado un espécimen de Mantis religiosa en una cuadrícula UTM de 10 x 10 km (creado con GNOMON 3.1).

YS BB

XR ?

YR BA

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? Fig. 2. Current expansion of Mantis religiosa in Poland and possible migration paths: A, long–term occurrence according to Liana (2007); B, possible expansion of populations originating from the population in circle A; C, new populations of Mantis religiosa probably not related to the populations from circle A. Fig. 2. Expansión actual de Mantis religiosa en Polonia y posibles rutas migratorias: A, presencia durante un largo período de tiempo según Liana (2007); B, posible expansion de las poblaciones originarias de la población contenida en el círculo A; C, nuevas poblaciones de Mantis religiosa que probablemente no guardan relación con las poblaciones del círculo A.

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different parts of the country. This indicates a significant spread of the European mantis in Poland. This study underscores this. Liana (2002) designates the Sandomierskie Forests, the Janowskie Forests and the Lipskie Forests, which are located in southeast Poland in the Sandomierska Basin macro–region, as the northern boundary of occurrence of these insects. In the mentioned 'forests' ('Puszcza' in Polish), the occurrence of this insect is permanent, confirmed and documented back to the early 1920s (Liana, 2007). In the 18th century, it even occurred as far north as Warsaw (Berg et al., 2011). We have found that today M. religiosa is widely distributed north of the boundary mentioned by Liana (2002). The subsequent study of Liana (2007) confirmed the presence of M. religiosa in the following regions of Poland: the Sandomierska Basin, the Lubelska Upland, the Małopolska Upland and the Carpathians. Since Liana (2007) summarized entomological data until 2006, a great recent range expansion of M. religiosa in Poland becomes obvious. This research corroborates the data of Liana (2007), but enumerates many more regions in which mantises were encountered (fig. 1). Based on these data, several conclusions about the significant spread of this species in all directions from their distributional core can be drawn (fig. 2). We consider these new distributional localities to be indeed new and not just overlooked older populations for the following reasons. First, in Central Europe, including Poland, orthopterans are a well–researched group because of their suitability for habitat quality assessment, and as such are frequently monitored. Mantis religiosa is also monitored in such studies (e.g. Liana, 2002; Nagy and Sziráki, 2002; Krištín et al. 2004; Krištín and Hrúz, 2005; Nagy and Kisfali, 2007; Nagy et al., 2011; Holuša et al., 2012), and would have been recorded earlier if previously present. Second, the current spread observed in Poland is paralleled by that in neighboring countries (see above). And third, the species would have been previously encountered by laypersons if present in a certain location, even if abundances were lower and encounters rare. Previous records from most of the new localities, many of which are outside the known historical range of the species, are unknown. The historical occurrence of M. religiosa in the Sandomierz Basin is not surprising as the warm climate of the region supports thermophilic taxa like mantises (Walther et al., 2009; Linn and Griebeler, 2016). However, the new reports of mantises in the southern parts of the country are intriguing. Liana (2007) had already mentioned the presence of two metapopulations in Poland, 'the Sandomierz' and 'the Carpathian'. In 1960 Bazyluk reported the subspecies Mantis religiosa polonica for the Sandomierz Basin population. But he also included the Ukrainian and northeast Austrian populations into M. r. polonica. Several authors were skeptical about the distinctive status of M. r. polonica, most notably A. Kaltenbach (1963 in Kočárek et al., 2005), suggesting that the Sandomierz population was only an ecological form characteristic for the northern edge of the range of this species (see also Berg et al., 2011). Królik (2010)

compared his specimens from UTM YS01 with the description of M. r. polonica given by Bazyluk (1960), and found them to be incompatible. He concluded that the specimens he found in Opole were M. r. religiosa, which now occurs abundantly in the Czech Republic (Chobot, 2016) and Slovakia (Fedor et al., 2010). Therefore, the specimens from southern Poland are representatives of morphologically typical M. r. religiosa which may have migrated through the Moravian Gate from Czech Moravia. They would thus represent a subspecies previously not recorded in Poland (fig. 2). Our own results confirm the occurrence of M. religiosa at UTM YS01 indicated by Królik (2010), supplemented by additional populations in the nearest area to the south (UTM BA98, UTM BA89, UTM CA08, UTM CA19). New genetic data do not support the existence of more than one subspecies in Europe. Linn and Griebeler (2015) showed differences between the populations of western and eastern parts of Germany, using four mitochondrial markers. They detected a threefold origin of German haplogroups, two originating from France, and the other from the Czech Republic. In a more extensive study, Vitáček (2016) showed that European populations represent three distinct lineages derived from Pleistocene refugia in S Europe: an Eastern lineage derived from the Black Sea refuge, a Western lineage derived from a Franco–Iberian refuge, and a central lineage which found refuge in southern Italy and the Adriatic region. The Central European (E Germany, Czech Republic, Poland, Slovakia, Austria, S Switzerland) populations all derive from an eastern branch of the central lineage. Vitáček’s (2016) data also show that the Sandomierz population has an impoverished haplotype set and clusters deeply nested among M. r. religiosa central lineage populations. This means that this population does not represent an entity worth of subspecific rank. Consequently, Schwarz et al. (2017) recently synonymized M. r. polonica Bazyluk, 1960 with M. r. religiosa Linnaeus, 1758. Conclusions The widescale spreading of Mantis religiosa in Europe is a natural consequence of global warming and improved conservation measures (Walther et al., 2009; Linn and Griebeler, 2015; Schwarz et al., 2017). Even though the European mantis is the most common and widespread Mantodea species in the world, and despite being the subject of many different studies since historical times, there are many aspects of its biology and distribution which still need to be addressed, in addition to potential threats. The new localities of the European Mantis in Poland obtained through the present study show its recent rapid spread across the country. These data cannot yet be considered as confirmed sites of vital M. religiosa populations because they are based on anonymous responses, not on thorough entomological studies. Nevertheless, they are a serious signal for entomologists to undertake studies to confirm these

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locations and update the data on the prevalence of the European Mantis in Poland. Moreover, there is a need for a refinement of genetic data in order to track the colonization pathways of M. religiosa in Poland and other European countries. Acknowledgements We would like to thank all the respondents who provided data about their encounters with praying mantises in Poland. We also thank Paweł Buczyński (Department of Zoology, Maria Curie Skłodowska University in Lublin) for helping create the UTM map with the localities. References Battiston, R., 2016. Mantis religiosa. The IUCN Red List of Threatened Species 2016: e.T44793247A44798476, doi: 10.2305/IUCN. UK.2016–1.RLTS.T44793247A44798476 Bazyluk, W., 1960. Die geographische Verbreitung und Variabilität von Mantis religiosa (L.) (Mantodea, Mantidae) sowie Beschreibungen neuer Unterarten. Annales Zoologici, 18(15): 231‒272. Berg, M. K., Dueker, C., Keller, M., Krueger, B., Luebcke, N., Luebcke, T., 2008. Die Gottesanbeterin, Mantis religiosa Linnaeus, 1758 (Mantodea: Mantidae), im Freistaat Sachsen. Entomologische Nachrichten und Berichte, 52(2): 93‒98. Berg, M. K., Schwarz, C. J., Mehl, J. E., 2011. Die Gottesanbeterin, Mantis religiosa. Westarp Wissenschaften, Die Neue Brehm–Bücherei Bd. 656, Hohenwardsleben, Germany. Bolshakov, L. V., Shcherbakov, E. O., Mazurov, S. G., Alekseev, S. K., Ryabov, S. A., Ruchin, A. B., 2010. Northernmost records of praying mantis Mantis religiosa (Linnaeus, 1758) (Mantodea: Mantidae) in European Russia. Eversmannia, 23/24: 22‒25. Bonk, M., Kajzer, J., 2009. Increase of number of new localities of the Praying Mantis Mantis religiosa L. in Małopolska Upland. Chrońmy Przyryrodę Ojczystą, 65(3): 189‒194. Bonk, M., Kajzer, J., Szafrański, A., 2011. New records of the praying mantis Mantis religiosa in the Świętokrzyskie Mountains and Mazovia. Kulon, 16: 129‒133. Błoński, W., 2015. Praying mantis Mantis religiosa Linnaeus, 1758 in the Świętokrzyski National Park. Naturalia, 3: 145‒146. Buczyńska, E., Buczyński, P., Pałka, K., 2006. European Mantis (Mantis religiosa L.) (Mantodea: Mantidae) in the Roztocze Upland. Wiadomości Entomologiczne, 25(1): 56‒57. Cannings, R. A., 2007. Recent range expansion of the Praying Mantis, Mantis religiosa Linnaeus (Mantodea: Mantidae), in British Columbia. Journal of the Entomological Society of British Columbia, 104: 73‒80. Chobot, K., 2016. Mapa rozšíření Mantis religiosa (Linnaeus, 1758). In: BioLib, Biological Library, (O.

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D20162183.pdf [Accessed on June 10th, 2017]. Piszkiewitz, J., Beneš, J., Konvička, M., 2000. Record of praying mantis, Mantis religiosa (L.), on the Velký Kosíř Hill, Central Moravia. Časopis Slezskeho Zemského Muzea, Serie A Vědy Přírodní, 49: 58. Pupiņš, M., Kalniņš, M., Pupiņa, A., Jaundaldere, I., 2012. First records of European Mantid Mantis religiosa (Linnaeus, 1758) (Insecta: Dictyoptera, Mantidae) in Latvia. Acta Biologica Universitatis Daugavpiliensis, 12(2): 175‒184. Romanowski, J., Romanowski, M., 2014. Mantids (Mantodea) from Pelješac peninsula, southern Croatia. Entomologia Croatica, 18(1–2): 7‒11. Schwarz, C. J., Keller, M., Berger, D., 2017. Neues zur Gottesanbeterin, Mantis religiosa Linnaeus, 1758 (Mantodea, Mantidae), dem Insekt des Jahres 2017. Entomologische Nachrichten und Berichte, 61(1): 1‒18. Sępioł, B., 2005. The new localisation of Mantis religiosa (Linnaeus 1758) in the north of the Land of the Holy Cross Mountains. Kulon, 10(1/2): 77. Shcherbakov, E. O., Savitsky, V. Y., 2015. New data on the fauna, taxonomy and ecology of praying mantises (Dictyoptera, Mantodea) from Russia. Entomological Review, 95(2): 181‒199. Stärz, C., Buchweitz, M., Fartmann, T., 2010. Feuer – (k)eine Chance für die Gottesanbeterin? Populations und Larvalökologie von Mantis religiosa in Rebböschungen. Arbeiten aus dem Institut für Landschaftsökologie, 19: 21–69. Tichá, K., 2005. Record of Mantis religiosa (Mantodea: Mantidea) in NNM Švařec (Czech Republic, Bohemian–Moravian Highlands). Acta Rerum Naturalium, 1: 155. Vitáček, J., 2016. Šíření kudlanky nábožné (Mantis religiosa) v Evropě. MS Thesis, Karl University Prague. Voisin, J. L., 2003. Atlas des Orthopteres et des Mantides de France. Museum National d‘Histoire Naturelle de Paris, Paris, France. Walther, G. R., Roques, A., Hulme, P. E., Sykes, M. T., Pyšek, P., Kühn, I., Czucz, B., 2009. Alien species in a warmer world: risks and opportunities. Trends in Ecology & Evolution, 24(12): 686‒693. Zieliński, D., 2016. Mantis religiosa – dispersion of species in Poland and in Europe. The new site in the Lublin region. Nauki Przyrodnicze, 3(13): 10‒18. Zieliński, D., Czyżowski, P., Karpiński, M., Goleman, M., Drozd, L., 2017. European mantis Mantis religiosa (Linnaeus, 1758) in Poland: identification and handling. Medycyna Weterynaryjna – Veterinary Medicine–Science and Practice, 73(3): 189‒191. Zieliński, D., Łazarecki, T. (in press). Modliszka zwyczajna (Mantis religiosa L.) z Białegostoku, Suwałk i okolic Białowieskiego Parku Narodowego. Wiadomości Entomologiczne.

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agnitude, composition M and spatiotemporal patterns of vertebrate roadkill at regional scales: a study in southern Spain D. Canal, C. Camacho, B. Martín, M. de Lucas, M. Ferrer

Canal, D., Camacho, C., Martín, B., de Lucas, M., Ferrer, M., 2018. Magnitude, composition and spatiotemporal patterns of vertebrate roadkill at regional scales: a study in southern Spain. Animal Biodiversity and Conservation, 41.2: 281–300. Abstract Magnitude, composition and spatiotemporal patterns of vertebrate roadkill at regional scales: a study in southern Spain. Although roadkill studies on a large scale are challenging, they can provide valuable information to assess the impact of road traffic on animal populations. Over 22 months (between July 2009–June 2010, and April 2011–March 2012) we surveyed 45 road sections of 10 km within a global biodiversity hotspot in Andalusia (87,000 km2), in southern Spain. We divided the region into five ecoregions differing in environmental conditions and landscape characteristics and recorded the relative magnitude, composition and spatiotemporal patterns of vertebrate (birds, mammal, amphibians, and reptiles) mortality. We used roadkill data from monthly surveys of road stretches with different speed limits, traffic volume, road design, and adjacent landscape composition. Roadkills varied over time and were not randomly distributed across ecoregions and road types. Overall, the groups most frequently encountered were mammals (54.4 % of total roadkills) and birds (36.2 %). Mortality rates in these two groups were higher on highways than on national or local roads, whereas those of amphibians (4.6 %) and reptiles (4.3 %) did not differ between road types. Except for mammals, the observed variation in vertebrate roadkills across ecoregions reflects the patterns of species richness previously described in the literature. Roadkills were concentrated over relatively short periods and this pattern was repeated over study periods and for all vertebrate classes. Our findings provide baseline information about road types, time periods and taxa with a higher probability of roadkills across an extensive region. These data represent an essential step towards the future implementation of broad–scale mitigation measures. Key words: Wildlife vehicle collisions, Collision patterns, Collision hotspots, Mitigation measures, Non–natural mortality, Human impact Resumen Magnitud, composición y patrones espaciotemporales de la mortalidad de vertebrados en las carreteras a escala regional. A pesar de que los estudios a gran escala sobre mortalidad de animales en las carreteras son complejos, pueden aportar información valiosa para evaluar la incidencia del tráfico en las poblaciones de animales. Durante 22 meses (entre julio de 2009 y junio de 2010 y entre abril de 2011 y marzo de 2012), muestreamos 45 tramos de carretera de 10 km de longitud distribuidos en una zona con una gran diversidad en la región de Andalucía (87.000 km2), en el sur de España. La región se dividió en cinco ecorregiones con diferentes condiciones ambientales y características del paisaje, y se analizaron la magnitud, la composición y los patrones espaciotemporales de la mortalidad de vertebrados (aves, mamíferos, anfibios y reptiles). Usamos datos de atropellos obtenidos durante muestreos mensuales en tramos de carretera con diferentes límites de velocidad, volumen de tráfico, diseño de la carretera y composición del paisaje adyacente. Los animales atropellados fueron distintos en el tiempo y no se distribuyeron aleatoriamente entre ecorregiones ni entre tipos de carretera. En total, los grupos que se encontraron con mayor frecuencia fueron los mamíferos (el 54,4 % de los atropellos registrados) y las aves (el 36,2 %). La tasa de mortalidad observada en estos dos grupos fue mayor en autopistas que en carreteras nacionales o locales, mientras que la mortalidad de anfibios (el 4,6 %) y de reptiles (el 4,3 %) no presentó diferencias entre tipos de carretera. A excepción de los mamíferos, la variación observada de la mortalidad en las carreteras entre las diferentes ecorregiones refleja los patrones ISSN: 1578–665 X eISSN: 2014–928 X

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de riqueza de especies descritos en las publicaciones científicas. Los atropellos se concentraron en períodos de tiempo relativamente cortos y este patrón se repitió en las dos temporadas de estudio y con respecto a todas las clases de vertebrados. Nuestros resultados proporcionan información de referencia sobre los tipos de carretera, los períodos de tiempo y los taxones con una mayor probabilidad de morir por atropello en una extensa región, lo que supone un paso esencial para la implementación de medidas de mitigación a gran escala. Palabras clave: Colisiones de vehículos con animales, Patrones espaciotemporales de las colisiones, Puntos calientes de colisiones, Medidas de mitigación, Mortalidad no natural, Impacto humano Received: 20 V 17; Conditional acceptance: 17 X 17; Final acceptance: 13 XII 17 David Canal, Manuela de Lucas, Miguel Ferrer, Applied Ecology Group, Doñana Biological Station (EBD–CSIC), Av. Américo Vespucio s/n., 41092 Seville, Spain.– Carlos Camacho, Dept. of Evolutionary Ecology, Doñana Biological Station (EBD–CSIC), Av. Américo Vespucio s/n., 41092 Seville, Spain.– Beatriz Martín, Fundación Migres, Complejo Huerta Grande, crta. N–340 km 96.7, 11390 Pelayo, Algeciras, Spain. Corresponding author: David Canal. E–mail: davidcanal@ebd.csic.es

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Introduction Linked to socio–economic development, the number of roads has increased dramatically in developed countries over the past decades. This rapid growth has led to a conflict between the need to amplify transport routes for socio–economic progress and the environmental impact caused by these infrastructures (e.g. barrier effect, habitat loss and contamination; reviewed in Forman and Alexander, 1998; Coffin, 2007; Torres et al., 2016). Collisions between vehicles and animals are an important consequence of this road–environmental conflict due to their significant socio–economic, environmental and traffic safety impact (Forman and Alexander, 1998; Forman et al., 2003; Langbein and Putman, 2006; Bissonette et al., 2008). Every year, wildlife–vehicle collisions cause an important number of accidents, resulting in animal fatality and human injury or death, and generating a substantial economic burden due to expenses such as medical costs and material damages (Bissonette et al., 2008; Huijser et al., 2009). Although it is difficult to accurately quantify the annual number of road casualties, some studies suggest about 500,000 occur per year in Europe (Bruinderink and Hazebroek, 1996; Bissonette et al., 2008). As these studies focused mostly on large, ungulate mammals, the number of fatalities could increase by several orders of magnitude when small road–killed vertebrates are considered. For example, considering birds only, the annual estimates of roadkill could reach 27 million fatalities in some European countries (Erritzøe et al., 2003). Traffic–related mortality is currently considered a major source of non–natural mortality in wildlife (birds: Erritzøe et al., 2003; mammals: Sáenz–de–Santa–María and Tellería, 2015; amphibians and reptiles: Colino– Rabanal and Lizana, 2012). This type of mortality may have significant effects on animal populations, such as increased inbreeding associated with isolation and reduced population size, population declines, and local extinctions (Forman et al., 2003; Coffin, 2007; Jackson and Fahrig, 2011). Traffic–related mortality may also affect the structure of animal populations because it has a differential incidence between sexes or age classes when associated with phenological events (e.g. dispersal or breeding; Mumme et al., 2000; Madsen et al., 2002; Jackson and Fahrig, 2011; Colino–Rabanal and Lizana, 2012). Traffic–related mortality has thus become an issue of major concern worldwide, as besides its potentially severe impact on natural populations, it may entail significant socio–economic costs (Forman and Alexander, 1998; Forman et al., 2003; Coffin, 2007; Bissonette et al., 2008). In an attempt to reduce roadkill rates, a number of mitigation measures have been designed (Glista et al., 2009). However, these measures may not be universally valid because their effectiveness depends on a wide array of factors (e.g. adjacent landscape features, species involved, and their phenology). We therefore need to understand the composition, magnitude and patterns of road kill on large scales in order to optimize prevention measures (Glista et al., 2009; van der Grift et al., 2013; Rytwinski et al., 2015).

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Here, we investigated the composition, seasonality and spatial patterns of vertebrate (birds, mammals, amphibian and reptiles) roadkill across an extensive Mediterranean region (Andalusia: 87,268 km2 southernmost part of the Iberian Peninsula), located in one of the most important biodiversity hotspots in the world (Myers et al., 2000). Over 22 months, we monitored 45 road sections distributed across the entire region each month, constituting, to our knowledge, one of the largest surveys to date on vertebrate roadkill. Material and methods The study was conducted in the Autonomous Community of Andalusia, southern Spain (fig. 1) over two periods: from July 2009 to June 2010 (with the exception of September and October), and from April 2011 to March 2012. The region harbors a large diversity of species and landscapes, recognized by 30 % of its territory being under some form of national or regional protection status (GIASA et al., 2006). The natural ecosystems in this region are considered highly sensitive to global change drivers, and it is predicted that they will experience dramatic biodiversity changes in the coming decades (Myers et al., 2000). The climate in Andalusia is Mediterranean. It has a marked environmental gradient (annual rainfall varies from 170 mm/year to more than 1,800 mm/year) and a wide elevation range (from sea level to approximately 3,500 m a.s.l.). These gradients display a high degree of spatial and temporal variation in vegetation and landscape composition, including semiarid zones, forests, mountains, and marshlands. To cover such environmental diversity, we surveyed roads within different ecoregions, defined as areas characterized by similar landscape characteristics and environmental conditions (GIASA et al., 2006). The study area included five well–differentiated ecoregions: (1) lowlands and green fields of the Guadalquivir river valley (LGG: annual crops, vineyards, olive groves, mosaic crops and traditional irrigation lands); (2) medium and high mountain areas of the Baetic system (MHMB: olives groves, traditional irrigation lands, Mediterranean scrub, cork oak forests, other forests, grasslands and upland crops); (3) Atlantic and Continental bio–geographic regions of Sierra Morena (ACSM: Mediterranean scrub, holm oak forests and dehesas and upland crops); (4) Atlantic and Mediterranean coastline (AMC: beach–dune systems, Mediterranean meadows, orchards and greenhouse crops); and (5) arid zones in southeastern Andalusia (AZS: sub–desert scrub and extensive steppes; fig. 1). Besides the environmental conditions, we selected road sections according to their physical characteristics (number of lanes, speed limit and traffic volume), aiming to capture a representative picture of the entire road network of Andalusia. We grouped the road network into three categories: (1) type I, including highways characterized by a dual carriageway and 120 km/h speed limit; (2) type II, including all roads belonging to the state, regional and interregional networks except highways (hereafter, national roads); and

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(3) type III, all roads belonging to the complementary regional network and provincial councils (hereafter, local roads). Both national and local roads are characterized by a single carriageway and 90 km/h speed limit. Traffic volume is assumed to decrease gradually from highways to national and local roads. Within each of the five ecoregions, we surveyed nine road sections (three sections of each road category) of 10 km in length each month. In sum, we inspected 9,900 km (4,500 and 5,400 km in the first and second study periods, respectively) divided into 45 road segments (tables 1 in supplementary material). The monthly surveys were carried out by two experienced observers driving a vehicle at low speed (25–30 km/h) along the roadside and with the emergency lights flashing. Surveys were conducted during the whole day. The sampling order of the surveyed sections was set at random from month to month and survey session. When a roadkill was encountered on either the paved road or the road verge, the animal was identified at the species level (whenever possible) and its location was recorded using a GPS. Carcasses were removed to avoid duplicating records during later surveys. Several potential sources of carcass removal (e.g. scavengers, rainfall and runoff) might have biased the spatiotemporal patterns described here (reviewed in Guinard et al., 2012; Teixeira et al., 2013). We are aware that a shorter sampling periodicity is usually required to accurately estimate the accumulated number of road casualties because small animals such as amphibians and reptiles may remain on the roads for short periods of time and roadkills may be clustered in time (Guinard et al., 2012; Teixeira et al., 2013; Santos et al., 2015). This means that the number of roadkills detected do not represent the real numbers for some species, especially the small ones (Teixeira et al., 2013), and direct comparisons between species in the total number of roadkills, therefore, are meaningless. In this regard, despite their potential inaccuracy, we decided to report and interpret with caution the roadkill patterns of amphibians and reptiles since they can be useful for future studies. Note that although the actual magnitude of road mortality in Andalusia cannot be accurately determined, our survey allows us to assess the relative incidence of each road section and to explore roadkill patterns at the (eco) regional level, which is the aim of this study. Indeed, the composition and temporal mortality patterns observed in our study area are similar to those reported by other surveys conducted in the Iberian Peninsula at finer temporal resolution (i.e. weekly or fortnightly; see discussion; Frías, 1999; Grilo et al., 2009; Garriga et al., 2012; D’Amico et al., 2015). Therefore we are confident that our results reflect the true roadkill pattern in the Andalusian road network, potential differences in mortality patterns among ecoregions are assumed to be reliable. Statistical analyses We used chi–square goodness of fit test to test for temporal (monthly) and spatial (ecoregional) differences in roadkill frequency and to investigate roadkill

variations based on the road category and taxonomic group. Exploratory analyses showed that patterns of traffic–related mortality were similar between years both spatially (Wilcoxon paired test within ecoregions: V = 140.5, P = 0.19; or according to the type of road: V = 11, P = 0.43) and temporally (monthly: V = 21.0, p = 0.54). Data from both years were thus pooled in further analyses. Statistical analyses above were carried out in R (R Core Team, 2016). We used the nearest neighborhood distance (NND) method (Gonser et al., 2009) to assess whether roadkills were aggregated or, conversely, independently and homogeneously distributed according to the uniform distribution over the road network. For this purpose, we tested the complete spatial randomness (CSR) hypothesis in terms of the number of roadkills at a given point set, showing, that the shortest distance from one roadkill to another roadkill was less than a parametric shortest path distance (Okabe and Sugihara, 2012). To test CSR, we applied the K function method in SANET 4.1 (Okabe et al., 2013). We identified collision hotspots along the road network by means of a kernel density analysis (Okabe and Sugihara, 2012) using SANET V4.1 (Okabe et al., 2013). The width of the kernel function (i.e. the area of influence of each kernel function, or bandwidth) chosen for the present work was 500 m (Conruyt–Rogeon and Girardet, 2012; Morelle et al., 2013). Estimated densities were then classified using the Jenks’ methods, based on minimization and maximization of variance within and between density classes, respectively (Morelle et al., 2013). Since we aimed to determine the main mortality hotspots for each taxa in the whole study area regardless of the differences in wildlife density between ecoregions, we looked for spatial aggregation in all the roadkills distributed throughout the entire road network. Previous research suggests that Malo et al.'s method (Malo et al., 2004) should be preferred for hotspot identification (Gomes et al., 2009). As an alternative hotspot identification method, we compared the spatial pattern of fatality occurrences with that expected in a random situation, in which the likelihood of collisions for each road segment exhibits a Poisson distribution (Malo et al.'s method). During our surveys, we found that domestic animals (dogs, Canis lupus familiaris and cats, Felis silvestris catus; see table 1) represented 18.8 % of all mammals killed by vehicles. Results of the analyses excluding domestic animals were qualitatively similar to those including all data. For this reason, results from the analyses including the whole dataset are presented here. Results General results A total of 1,535 animals belonging to 102 species were recorded as killed by vehicles during the two study periods (table 1). The class most frequently found dead along roads were mammals (χ23 = 1135.8,

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N 0 30 60

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Road networks Atlantic and Continental biogeographic regions of Sierra Morena (ACSM) Atlantic and Mediterranean coastline (AMC) Lowlands and greenfields of Guadalquivir River (LGG) Median and high mountain areas of the Baetic system (MHMB) Subdesert and arid zones in the southeastern (AZS) Fig. 1. Situation of the road sections and main ecological units (ecoregions) surveyed during the study period. Surveyed points (roadkill and control points) are highlighted in black. Fig. 1. Localizacion de los tramos de carreteras y las principales unidades ecológicas (ecorregiones) muestreadas en el estudio. Los puntos muestreados (puntos de atropello y puntos de control) están resaltados en color negro.

P < 0.001; 54.4 %), followed by birds (36.2 %), amphibians (4.6 %), and reptiles (4.3 %). Only 2.8 % of roadkill could not be identified at species level due to damage and/or poor conservation condition, yet 85 % (n = 51) of these casualties could be identified at class level (0.5 % of carcasses remained undetermined at the class level; table 1). Six of the 102 species found are threatened in Spain, listed as either 'Endangered' (Milvus milvus) or 'Vulnerable' (Calandrella brachydactyla, Streptopelia turtur, Oryctolagus cuniculus, Salamandra salamandra and Mauremys leprosa; table 1). Nine further species are listed as 'Near Threatened' in Spain (Lanius senator, Lanius meridionalis, Asio flammeus, Phylloscopus trochilus, Mustela putorius, Alytes obstetricans and Pleurodeles waltl) or worldwide (Sylvia undata and Timon lepidus), although the latter has not been evaluated in Spain (table 1). The incidence of traffic–related mortality for (near) threatened species was lower than six animals in all cases, with the exception of Oryctolagus cuniculus (n = 298 casualties) and Salamandra salamandra (n = 11). Spatiotemporal patterns of roadkill Overall, the abundance of roadkill was not uniformly distributed across months (x211 = 47.4, P < 0.001; fig. 2 and table 2s in supplementary material). The analyses within each vertebrate class showed a similar pattern, i.e. mortality rates differed between months (birds: x211 = 22.5, P = 0.02; mammals: x211 = 19.9,

P = 0.04; amphibians: x211 = 67.9, P < 0.001 and reptiles: x211 = 53.3, P < 0.001; fig. 2). Spatially, the observed roadkill mortality varied markedly across ecoregions, (x24 = 259.4, P < 0.001; fig. 3 and table 3s in supplementary material). Overall, the ecoregion with the highest number of roadkills was the Guadalquivir Lowlands (29.2 %), followed by the medium and high mountain areas (26.6 %), Atlantic and Mediterranean coastline (23.9 %), Sierra Morena (11.5 %), and the sub–desert and arid zone in southeastern Andalusia (8.8 %). The number of casualties recorded also varied significantly between ecoregions for each taxonomic group (x212 = 198.2, P < 0.001). The highest number of road–killed mammals was recorded in the Guadalquivir Lowlands (42.4 % of all detected mammals), whereas medium and high mountain areas of the Baetic system recorded the highest rate of road–killed birds (28.7 % of all detected birds). Concerning amphibians and reptiles, the most hazardous ecoregions were Sierra Morena (40.8 % of all detected amphibians) and the arid zone in southeastern Andalusia (28.8 % of all detected reptiles; fig. 3 and table 3s in supplementary material). The number of roadkills also varied among road types (x22 = 207.4, P < 0.001), with the highest (50.8 % of the total) vertebrate mortality recorded on highways, followed by national (31.5 %) and local roads (17.7 %; fig. 1s in supplementary material). This pattern was true for mammals and birds (birds: x22 = 84.7, P < 0.001; mammals: x22 = 254.9, P < 0.001), whereas amphibians had the lowest mortality on national roads

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Table 1. Number of individuals, ecoregions and conservation status of the vertebrate species found killed by traffic along Andalusian roads between the 2009–2010 and 2011–2012 surveys. Conservation categories of the species were obtained from: (1) the Red Book of Spanish Vertebrates first published in 1992 (RB1992; Blanco and González, 1992), and then revised (RB2000) in 2002 for amphibians (Pleguezuelos et al., 2002), in 2004 for birds (Madroño et al., 2004), and in 2007 for mammals (Palomo et al., 2007); (2) CNEA, the Spanish National Catalogue of Threatened Species; and (3) IUCN (Version 2015.2). Status categories are: NE, not evaluated; DD, data deficient; K, insufficiently known; LC, least concern; NT, near threatened; VU, vulnerable; EN, endangered; SI, species of special interest (for the abbreviations of ecoregions in Andalusia, see figure 1): * top roadkilled species recorded in D’Amico et al. (2015) (Huelva, South Spain); ¡ top roadkilled species recorded in Garriga et al. (2012) (Catalonia, Northeaster Spain); + top roadkilled species recorded in Frías (1999) (Toledo, Central Spain) (only birds were surveyed); and $ top roadkilled species recorded in Grilo et al. (2009) (Alentejo, Portugal) (only nine species of medium–size carnivores were surveyed). Tabla 1. Número de individuos, ecorregión y estado de conservación de las especies de vertebrados encontradas atropelladas en las carreteras de Andalucía durante los muestreos de 2009–2010 y 2011–2012. Las categorías de conservación de las especies se tomaron de: (1) el Libro Rojo de los Vertebrados de España, publicado por primera vez en 1992 (RB1992; Blanco y González, 1992) y luego revisado (RB2000) en 2002 para anfibios (Pleguezuelos et al., 2002), en 2004 para aves (Madroño et al., 2004) y en 2007 para mamíferos (Palomo et al., 2007); (2) CNEA, el Catálogo Español de Especies Amenazadas y (3) UICN, la Unión Internacional para la Conservación de la Naturaleza (Versión 2015.2). Las categorías de conservación son: NE, no evaluado; DD, datos insuficientes; K, insuficientemente conocida; LC, preocupación menor; NT, casi amenazada; VU, vulnerable; EN, en peligro; SI, especie de especial interés (para las abreviaturas de las ecorregiones en Andalucía, véase figura 1): * especies que sufren más atropellos mortales según D’Amico et al. (2015) (Huelva, sur de España). ¡ especies que sufren más atropellos mortales según Garriga et al. (2012) (Cataluña, noroeste de España); + especiee que sufren más atropellos mortales según Frías (1999) (Toledo, España central) (solo se estudiaron aves); $ especies que sufren más atropellos mortales según Grilo et al. (2009) (Alentejo, Portugal) (solo se estudiaron nueve especies de carnívoros de tamaño medio).

Conservation status Ecoregions RB2000 RB1992 CNEA IUCN Records LGG AMC MHMB ACSM AZS Mammals 835 Oryctolagus cuniculus * VU NE – NT 298 111 45 109 21 12 Rattus norvegicus¡ LC NE – LC 103 25 49 24 1 4 Canis lupus familiaris – – – – 93 30 21 16 7 19 Erinaceus europaeus * LC NE SI LC 76 46 16 5 5 4 Felis silvestris catus – – – – 63 28 15 9 9 2 Lepus sp. – – – – 54 41 3 8 0 2 Lepus europaeus LC NE – LC 46 18 6 16 5 1 Vulpes vulpes *$ LC NE – LC 30 5 2 9 5 9 Lepus granatensis LC NE – LC 15 9 1 4 0 1 Apodemus sylvaticus LC NE – LC 11 4 4 1 2 0 Genetta genetta LC NE – LC 8 4 2 1 0 1 Herpestes ichneumon $ LC K – LC 8 2 3 3 0 0 Pipistrellus pipistrellus LC NE SI LC 7 3 1 3 0 0 Martes foina $ LC NE – LC 6 1 1 4 0 0 Mustela putorius NT K – LC 4 0 2 1 0 1 Eptesicus serotinus LC K SI LC 3 0 0 1 1 1 Meles meles LC K – LC 3 0 1 0 1 1 Eliomys quercinus LC NE – NT 2 0 2 0 0 0 Mustela nivalis LC NE – LC 1 0 0 1 0 0 Undetermined mammals

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Table 1. (Cont.)

Conservation status Ecoregions RB2000 RB1992 CNEA IUCN Records LGG AMC MHMB ACSM AZS Birds 555 ¡+ Passer domesticus NE NE – LC 70 13 30 14 8 5 Sylvia melanocephala DD NE SI LC 48 1 14 16 8 3 + Athene noctua NE NE SI LC 42 16 4 10 6 6 Alectoris rufa+ DD NE – LC 41 10 4 13 13 1 Turdus merula NE NE – LC 30 2 9 16 2 1 Sylvia atricapilla NE NE SI LC 29 1 12 14 1 1 Erithacus rubecula * DD NE SI LC 28 6 5 6 7 4 Columba livia NE NE – LC 12 2 1 4 4 1 Fringilla coelebs NE NE – LC 12 2 1 4 1 4 Galerida cristata NE NE SI LC 12 1 3 5 1 2 Saxicola torquata NE NE SI LC 11 1 4 2 2 2 Tyto alba NE NE SI LC 11 4 3 2 1 0 Caprimulgus ruficollis * NE K SI LC 8 3 1 3 1 0 Miliaria calandra NE NE – LC 8 3 1 2 2 0 Pica pica NE NE – LC 8 0 2 2 3 1 Melanocorypha calandra NE NE SI LC 7 4 1 1 0 1 Turdus philomelos NE NE – LC 7 2 0 0 0 5 Hirundo rustica NE NE SI LC 8 2 2 2 0 0 Alauda arvensis NE NE – LC 6 3 0 1 1 1 Serinus serinus NE NE – LC 6 1 4 1 0 0 Cisticola juncidis NE NE SI LC 5 2 1 1 1 0 Hirundo daurica

Phylloscopus collybita *

Bubo bubo

Alectoris rufa +

–

Calandrella brachydactyla VU

Carduelis carduelis

–

Jynx torquilla

Merops apiaster

Motacilla alba

Passer montanus

–

Picus viridis

Streptopelia decaocto

–

Asio otus

Bubulcus ibis

–

Cuculus canorus

–

Lanius senator

Sylvia conspicillata

Sylvia undata

Upupa epops

Anas platyrhynchos

–

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Table 1. (Cont.)

Conservation status Ecoregions RB2000 RB1992 CNEA IUCN Records LGG AMC MHMB ACSM AZS Cettia cetti

Cyanopica cooki

Gallus gallus domesticus – – – –

2 1 0 1 0 0

Lanius meridionalis

–

Larus argentatus

–

Otus scops

Cyanistes caeruleus

Phoenicurus ochruros

Strix aluco

Sturnus unicolor

–

Sturnus vulgaris

–

Sylvia cantillans

–

Accipiter nisus

Asio flammeus

Carduelis cannabina

–

Carduelis chloris

–

Circus cyaneus

Corvus monedula

–

Coturnix coturnix

–

Delichon urbicum

Falco tinnunculus

Galerida theklae

Gallinula chloropus

–

Garrulus glandarius

–

Larus michahellis Milvus milvus Parus major

–

Petronia petronia

–

Phylloscopus trochilus

–

Streptopelia turtur

–

Undetermined birds Reptiles

Malpolon monspessulanus LC

–

Rhinechis scalaris

Timon lepidus

Mauremys leprosa

Coronella girondica

Psammodromus algirus LC

Hemorrhois hippocrepis ¡*

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Table 1. (Cont.)

Conservation status RB2000 RB1992 CNEA IUCN Records Amphibians

Bufo bufo

Ecoregions LGG AMC MHMB ACSM AZS

–

Salamandra salamandra ¡ VU

–

Pelophylax perezi *

–

Bufo calamita

Alytes obstetricans

Pleurodeles waltl *

Undetermined amphibians

Undetermined (class)

Mammals Season 1 Season 2

50 30

0 Jan Feb Mar Apr May Jun Jul Aug Sep Oct Nov Dec

Season 1 Season 2

Reptiles

Season 1 Season 2

Amphibians

0 Jan Feb Mar Apr May Jun Jul Aug Sep Oct Nov Dec

Season 1 Season 2

Birds

Jan Feb Mar Apr May Jun Jul Aug Sep Oct Nov Dec

Month

Jan Feb Mar Apr May Jun Jul Aug Sep Oct Nov Dec

Month

Fig. 2. Temporal variation in roadkill for each vertebrate group found on the Andalusian roads. Note that no road surveys were conducted in September or October during the first study period. Fig. 2. Variación temporal del número de individuos atropellados encontrados en las carreteras de Andalucía para cada grupo de vertebrados. Nótese que, durante el primer período del estudio, no se realizaron muestreos en septiembre ni en octubre.

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Mammals Birds 57 58 59–175 176–216 217––329 Reptiles

52 53–86 87–105 106–144 145–168

Amphibians 5–9 10–11 12–19 20–24 25–29

4 5–6 7–9 10–23 24–29

Total mortality 135 136–177 178–367 368–408 409–448

Fig. 3. Vertebrate roadkill in Andalusia according to the ecological units (ecoregions). The grey scale reflects the absolute number of roadkilled animals (Natural Breaks Classification). Fig. 3. Número de vertebrados atropellados encontrados en Andalucía en función de las unidades ecológicas (ecorregiones). La escala gris refleja la cifra absoluta de animales atropellados (clasificación mediante umbrales naturales o de Jenks).

(x22 = 10.5, P < 0.005) and no spatial patterns were evident for reptiles (x22 = 1.9, P = 0.38). Roadkills were not randomly distributed across the surveyed roads. Rather they were spatially clustered, i.e., a few road sections accumulated a disproportionately large fraction of all casualties (fig. 1s in supplementary material). Moreover, we confirmed the absence of spatial randomness (CRS) with a 0.95 confidence level. Hotspots identified by Kernel and Malo et al.'s methods were spatially consistent (Spearman correlation between hotspots: rs = 0.58, p < 0.001; fig. 4). Within vertebrate classes, roadkills were also aggregated, but the identity of the road stretches underlying such a pattern varied by taxonomic group (fig. 4 and table 1s and fig. 2s in supplementary material). Discussion Our extensive road survey allowed us to unravel the relative magnitude, composition and spatiotemporal patterns of vertebrate roadkill across Andalusia, a region located in a globally recognized biodiversity hotspot. Temporally, roadkills showed seasonal peaks and, spatially, some ecoregions showed disproportionally more casualties than others. Further, certain sections had higher mortality rates than others and a higher incidence on some particular taxa.

Traffic–related mortality affects a wide range of taxonomic groups and is currently considered one of the main causes of non–natural mortality of wildlife (Forman et al., 2003; Bissonette et al., 2008; Morelle et al., 2013). We recorded casualties of 102 vertebrate species during our survey, including some species listed as threatened in Spain. Based on the total number of casualties recorded in this study, one might consider the incidence of traffic–related mortality on endangered species as anecdotal. However, due to differences in species detectability, the detected numbers of casualties during the surveys do not represent the real number of roadkill for some species (Teixeira et al., 2013). Furthermore, as endangered species are characterized by low effective population sizes, the roadkill of even a few individuals may seriously threaten the long–term viability of a population (Jackson and Fahrig, 2011; Borda–de–Água et al., 2011). In effect, traffic mortality is considered a major threat for the endangered Iberian lynx, Lynx pardinus (Ferreras et al., 1992; IUCN, 2015), the Florida scrub–jay, Aphelocoma coerulescens (Mumme et al., 2000), and several populations of turtles in the United States (Gibbs and Shriver, 2002; Andrews et al., 2008), among others. Among the threatened species recorded during our survey, we would like to draw attention to the mortality recorded for European rabbits and fire salamanders, both species listed as 'Vulnerable' in Spain. The former is a keystone species in Mediterranean ecosystems

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Mammals

Birds

Reptiles

291

Amphibians

All taxa

Mammals

Birds

0 > –1.2700 1.2701–3.4834 3.4835–8.084

0 > –1.47 1.48–3.66 3.67–9.45 Reptiles

Amphibians

0 > –0.58 0.59–1.76 1.77–4.27

0 > –0.54 0.55–1.49 1.50–4.38

All taxa

0 > –2.10 2.11–5.31 5.32–17.84 Fig. 4. Collision hotpots found on the Andalusian roads: A, Malo et al’s method (Malo et al., 2004). The hotspots highlighted were defined by segments of 500-m with more than two fatalities for amphibians and reptiles (0.0003 Poisson probability), more than three fatalities for mammals (0.002 Poisson probability) and birds (0.001 Poisson probability), and more than four fatalities for all pooled taxa (0.007 Poisson probability); B, Kernel density estimation (bandwith = 500 m; Okabe et al., 2013). Fig. 4. Puntos calientes de colisiones en las carreteras de Andalucía: A, método de Malo y colaboradores (Malo et al., 2004). Los puntos calientes destacados fueron definidos por tramos de 500 m con más de dos atropellos mortales de anfibios y reptiles (probabilidad de Poisson de 0,0003), más de tres atropellos mortales de mamíferos (probabilidad de Poisson de 0,002) y de aves (probabilidad de Poisson de 0,001) y más de cuatro atropellos mortales cuando todos los taxones fueron analizados conjuntamente (probabilidad de Poisson de 0,007); B, estimación de la densidad Kernel (anchura = 500 m; Okabe et al., 2013).

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(Delibes and Hiraldo, 1981) as it is the main prey of other key endangered species, such as the Iberian lynx and the Spanish imperial eagle Aquila adalberti. Despite being locally abundant in some parts of its natural range, the rabbit has suffered a sharp population decline over the last decades (70 % between the 1970s and 2000s; Virgós et al., 2007). Thus, although rabbits have high reproductive rates and probably only those inhabiting road verges are affected by roadkill, the high prevalence of rabbits recorded in our survey (298 individuals; 35 % of road killed mammals) might play a negative role in the recovery of certain rabbit populations. In the case of the fire salamander, all road–killed individuals were found within a single 2–km section (local road HU–9116/SE–6405) near a main river, suggesting all carcasses belonged to the same local population and that roadkill has a signficant impact on this species in the area. Traffic–related mortality in our survey showed seasonal peaks both across and within vertebrate groups. Several factors may have jointly contributed to shape the temporal distribution of traffic–related mortality across our study region. First, weather conditions seem to affect the incidence of roadkill, since mortality peaked in spring (March, April and May) and in autumn (October, November) and decreased in summer. The Mediterranean climate in Andalusia is characterized by a long summer drought. High temperatures and low water availability during this season often lead to a low level of activity in the studied vertebrate groups. In contrast with other regions across Spain, many areas of Andalusia, particularly those nearby to the coast, have mild autumn and winter temperatures. The moderate decrease in temperatures after summer combined with the increased water availability after the autumn rainfalls allow for high to moderate levels of activity in vertebrates in autumn and winter in Andalusia, even in cold–blooded taxa. Possibly, the annually bimodal (spring–autumn) roadkill pattern found in amphibians was largely influenced by rainfall and the typically mild temperatures in the region during these seasons. In the case of reptiles, the higher number of reptile casualties in late spring and early summer may be related to long photoperiods and high temperatures that may promote high activity levels (Colino–Rabanal and Lizana, 2012; D’Amico et al., 2015). It should be noted that the number of herpetofauna roadkills recorded was low (possibly due to different sources of carcass removal; Guinard et al., 2012; Teixeira et al., 2013; Santos et al., 2015) and, as a consequence, the mortality patterns of this group should be cautiously interpreted. Second, phenological factors such as migrations (Santos et al., 2007), breeding periods (Grilo et al., 2009; D’Amico et al., 2015), rutting seasons (Madsen et al., 2002; Zuberogoitia et al., 2014) and the hunting period (Sudharsan et al., 2006) may affect wildlife behavior and habitat use and, consequently, roadkill risk (Rytwinski and Fahrig, 2012). Phenological factors seem to be particularly important for birds, in which mortality peaks match the pre–breeding (i.e. migration) and breeding periods in spring on one hand, and autumn southward migration and the arrival of birds from northern latitudes on the other

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(Andalusia is an important overwintering place for a large proportion of the Western European migratory birds; SEO/BirdLife, 2012). Regarding mammals, the number of casualties increased in spring (April) and in late autumn–early winter (October, November and December). This increase could be due to the combined effect of breeding (e.g. European hedgehogs and red foxes feeding their young in spring; Grilo et al., 2009), the subsequent dispersal of inexperienced young (e.g. red foxes in autumn), and the impact of disturbance from hunting activities on animals' propensity to move (Sudharsan et al., 2006; Morelle et al., 2013). Overall, our results are in line with those found in previous studies in the Iberian Peninsula (e.g. Frias, 1999; Grilo et al., 2009; Garriga et al., 2012; Zuberogoitia et al., 2014; D’Amico et al., 2015) showing that roadkill patterns reflect species–specific differences in activity and mobility caused by weather, and phenological and hunting events (Santos et al., 2007; Grilo et al., 2009; D’Amico et al., 2015). Other factors, such as short– or long–term fluctuations in traffic volume (Seiler, 2005; Zuberogoitia et al., 2014; Gagné et al., 2015) or poor visibility conditions (El Faouzi et al., 2010; Mitra, 2014), may also have affected the roadkill patterns found in this study. However, given the correlative nature of our study, it is not possible to disentangle the relative contribution of the above factors to the roadkill patterns in each vertebrate group. Roadkill patterns differed among ecoregions, with the Guadalquivir lowlands and the arid zones in southeastern Andalusia showing the highest and lowest number of roadkill, respectively. Except for mammals, the spatial variation of vertebrate roadkill is in accordance with the general patterns of species richness across ecoregions (Martin and Ferrer, 2015). Possible explanations for this unexpected result in the case of mammals include a higher attraction to roads or a reduction in the effectiveness (and/or number) of mitigation measures in areas with less species richness, as compared with more species–rich areas. Road type (a proxy of traffic intensity and speed) was also an important predictor of roadkill, as shown in other studies (e.g. Morelle et al., 2013; Zuberogoitia et al., 2014). Overall, the highest vertebrate mortality occurred on highways, followed by national and local roads. This is in accordance with the pattern of traffic volume and speed expected for these types of roads, assumed to decrease gradually from highways to nationals and local roads. However, traffic volume and other road features may also vary at microgeographic scales, and thus confound the general patterns of roadkill mortality. For example, some particular stretches within local or national roads caused similar or even higher mortality than some sections within highways, whereas vertebrate mortality varied markedly between close sections of the same highway (e.g. 26 and 56 roadkills in two different sections of A–66; table 1). Furthermore, the overall mortality recorded in each road section varied substantially between vertebrate classes. For example, we recorded the highest mortality rates for amphibians (25.3 %) on the surveyed stretch of road HU–9116, whereas no birds or reptiles were found

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along this stretch. As a consequence, the areas with high risk of collision varied among vertebrate groups as shown by kernel analyses. These results imply that a large fraction of roadkill variance is not explained by traffic parameters but rather by road and landscape features and/or the spatial behavior of the affected species. For example, diverse road features, such as elevation changes, road sinuosity and the presence of road crosses, may also increase roadkill (e.g. Malo et al., 2004; Seiler, 2005; Grilo et al., 2009; Gunson et al., 2011; Zuberogoitia et al., 2014). Besides this, proximity to water bodies may be related to the rate of amphibian roadkill (Santos et al., 2007; Colino–Rabanal and Lizana, 2012), whereas proximity to forested areas often increases mortality risk in ungulates (Madsen et al., 2002; Seiler, 2005; Langbein and Putman, 2006). Overall, we have shown that roadkills in southern Spain were spatially (across regions) and temporally (throughout the year) aggregated. Our results suggest that species’ behavior, landscape and road features were the main factors determining the probability of roadkill. Remarkably, the spatial pattern of mortality found (both overall and within taxa) resembles that caused by other anthropogenic infrastructures, such as power lines or wind farms, wherein few pylons or wind turbines typically account for the majority of casualties (Janss and Ferrer, 2001). Thus, as successfully shown in these research areas (Janss and Ferrer, 2001), the identification and subsequent application of mitigation measures in collision hotspots would dramatically reduce overall road mortality. Acknowledgements We thank Miguel A. Sanza for his dedicated assistance during the road surveys. We are grateful to Mario Diaz and an anonymous reviewer for their constructive comments on a previous draft. This project was funded by Consejería Medio Ambiente (Junta de Andalucía). C. Camacho was supported by the Spanish Ministry of Economy and Competitiveness (SVP–2013–067686). References Andrews, K., Gibbons, J., Jochimsen, D., 2008. Ecological effects of roads on amphibians and reptiles: a literature review. In: Urban herpetology: 121–143 (J. C. Mitchelland, R.E. Jung, Eds.). Society for the Study of Amphibians and Reptiles, Salt Lake City. Bissonette, J. A., Kassar, C. A., Cook, L. J., 2008. Assessment of costs associated with deer–vehicle collisions: human death and injury, vehicle damage, and deer loss. Human–Wildlife Interactions, 2: 17–27. Blanco, J. C., González, J. L. (Eds.), 1992. Libro Rojo de los Vertebrados de España. ICONA, Madrid. Borda–de–Água, L., Navarro, L., Gavinhos, C., Pereira, H. M., 2011. Spatio–temporal impacts of roads on the persistence of populations: analytic and numerical approaches. Landscape Ecology,

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Supplementary material

Table 1s. Ecoregion (E), road type (RT), situation of the roads, and roadkill recorded (number of casualties), separated according to taxonomic groups (M, mammals; B, birds; R, reptiles; A, amphibians): KmO, km origin; KmE, km end; * 10–km long sections including the same road designated with two different names, so that the official assignation of km also varies. Tabla 1s. Ecorregión (E), tipo de carretera (RT), situación de la carretera, y número de atropellos mortales registrados, separados por grupo taxonómico (M, mamíferos; B, aves; R, reptiles; A, anfibios): KmO, km origen; KmE, km final; * tramos de 10 km de una misma carretera que se designa con dos nombres diferentes, con lo que la asignación oficial de los puntos kilométricos también varía. Ecoregion Taxonomic group Type

Province

KmO

KmE

Lowlands and greenfields of Guadalquivir River (LGG) Highways

A–92

Seville

A–4

Seville

497

507

A–49

Huelva

National roads A–380

Seville

A–407/A–456

Seville

38 *

A–364

Seville

Local roads

JA–6108

Jaen

SE–7200

Seville

SE–8105

Seville

Atlantic and Mediterranean coastline (AMC) Highways

A–49

Huelva

107

117

A–48

Cadiz

A–7

Almeria

483

493

Huelva

National roads A–494

A–405

Cadiz

A–377

Malaga

Local roads

A–2227

Cadiz

A–2101

Cadiz

AL–3106

Almeria

Atlantic and Continental biogeographic regions of Sierra Morena (ACSM) Highways

A–66

Huelva

755

765

A–66

Seville

766

776

A–66

Seville

782

792

National roads A–461

Huelva

N–433

Huelva

A–424

Córdoba

Local roads

HU–9116/SE–6405 Huelva/Seville

15 *

CO–6103

Córdoba

A–3200

Córdoba

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Table 1s. (Cont.)

Ecoregion Type

Province

KmO

KmE

Taxonomic group M

Median and high mountain areas of the Baetic system (MHMB) Highways

A–381

Cadiz

A–92M

Malaga

A–92

Malaga

161

171

Granada

Seville

National roads A–308

A–406

A–333/A–328

Malaga/Córdoba

69 *

Local roads

CA–5102

Cadiz

SE–8205

Seville

Subdesert and arid zones in the southeastern (AZS) Highways

A–92N

Granada

A–92N

Granada

A–92

Almeria

365

375

National roads A–334

Granada

A–330

Granada

A–349

Almeria

Local roads

GR–7100

Granada

GR–9109

Granada

AL–3102

Almeria

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Table 2s. Temporal distribution of roadkill on the Andalusian roads according to taxonomic group and study period: M, month; P1, period 1; P2, period 2; Und., undetermined; NA, roads not surveyed in the second period in September or October. Tabla 2s. Distribución temporal de los individuos atropellados encontrados en las carreteras de Andalucía, separados por grupo taxonómico y período de estudio: M, mes; P1, período 1; P2, período 2; Und., sin determinar; NA, carreteras que no se muestrearon en el segundo período en septiembre ni en octubre.

Mammals

Birds

Reptiles

Amphibians

Und.

127

150

179

131

130

141

212

155

Total

448

387

320

235

1,535

Total

Table 3s. Spatial distribution of roadkill on the Andalusian roads according to taxonomic group and study period: P1, period 1; P2, period 2; Und., undetermined; E, ecoregion. (For abbreviations of ecoregions, see figure 1). Tabla 3s. Distribución espacial de los individuos atropellados encontrados en las carreteras de Andalucía, separados por grupo taxonómico y período de estudio: P1, período 1; P2, período 2; Und., sin determinar; E, ecorregión. (Para las abreviaturas de las ecorregiones, véase la figura 1).

Mammals

Birds

Reptiles

LGG

141

188

AMC

100

MHMB

146

107

ACSM

Amphibians P1

Und.

Total

448

367

408

177

AZS

135

Total

448

387

320

235

1535

Road section

A–92N_S1 A–66_S1 A–92_S1 A–92N A–66_S2 A–7_S1 A–92M A–49_S1 A–66_S3 A–49_S2 A–4 A–381 A–92_S2 A–7_S2 A–92 A–494 A–461 A–424 A–330 A–377 A–334 A–308 A–349 A–433 A–406 A–405 A–333/A–328 A–364 A–380 A–407/A–456 GR–9109 A–3200 GR–7100 CO–6103 JA–6108 CA–5102 SE–8205 MA–5102 AL–3106 AL–3102 HU–9116/SE–6405 SE–8105 SE–7200 A–2101 A–2227

Number of roadkills

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100

Fig. 1s. Histograms of the accumulated mortality by road section.

Fig. 1s. Histogramas de la mortalidad acumulada por tramo de carretera.

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120

Highway National road Local road

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300

14 12

Mammals

10 8 6 4 2 0 14 12

Birds

10 6 4 2 0 25

Amphibians

15 10 5 0 14 12

Reptiles

10 8 6 4 2 0

A–2101 A–2227 A–308 A–3200 A–330 A–333/A–328 A–334 A–348 A–364 A–377 A–380 A–381 A–4 A–405 A–406 A–407/A–456 A–424 A–461 A–48 A–49 A–49 (23–33) A–484 A–66 (755–765) A–66 (766–776) A–66 (782–792) A–7 A–92 A–92 (161–171) A–92 (42–52) A–92M A–92N (10–20) A–92N (50–60) AL–3102 AL–3106 CA–5102 CO–6103 GR–7100 GR–9109 HU–9116/SE–6405 JA–6108 MA–5102 N–433 SE–7200 SE–8105 SE–8205

Number of roadkills

Fig. 2s. Number of vertebrate roadkills found in the Andalusian roads according to taxonomic group and road section. Y–axis was calculated as the number of roadkills detected on a section of road in relation to the total number of roadkills found for the taxonomic group. Note that the scale of y–axis varies between taxonomic groups. Fig. 2s. Número de vertebrados atropellados encontrados en las carreteras de Andalucía, separados por grupo taxonómico y tramo de carretera. El eje de las Y indica el número de atropellos observados en cada tramo de carretera respecto al total de atropellos encontrados para el grupo taxonómico. La escala del eje de las Y difiere entre grupos taxonómicos.

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Foraging habitat selection by gull–billed tern (Gelochelidon nilotica) in Central Spain (Castilla–La Mancha)

V. O. Britto, J. A. Gil–Delgado, R. U. Gosálvez, G. M. López–Iborra, A. Velasco

Britto, V. O., Gil–Delgado, J. A., Gosálvez, R. U., López–Iborra, G. M., Velasco, A., 2018. Foraging habitat selection by gull–billed tern (Gelochelidon nilotica) in Central Spain (Castilla–La Mancha). Animal Biodiversity and Conservation, 41.2: 301–310. Abstract Foraging habitat selection by gull–billed tern (Gelochelidon nilotica) in Central Spain (Castilla–La Mancha). The gull–billed tern breeds in temporary lakes in Castilla–La Mancha in Central Spain but depends on surrounding land habitats to feed its chicks. It is therefore vital to know the type of environments it selects to capture prey to feed nestlings. The aim of this study was to evaluate the use of habitats for hunting by adult gull–billed tern. Of 66 lakes monitored between 1996 and 2016, we found the gull–billed tern used 12 for breeding. Each lake was used during this period for 1–14 breeding seasons. We selected circular areas around the three wetlands where the species bred in 2013 and 2014. Within these circles, we sampled a total of 60 random points and recorded 125 gull–billed tern contacts (including between 1 and 39 birds). We estimated the same environmental variables at contact and random points, including land use and the distance to the nearest wetland, the nearest colony and to several types of anthropic uses (paved roads, houses, and cities). To evaluate habitat selection we calculated the Manly selection index for soil use variables, and fitted linear mixed models to evaluate differences in the distance variables. Land uses selected for foraging by the gull–billed tern were mainly cereal crops, whereas vineyards were avoided. The birds foraged on average up to 2 km from the colonies and tended to avoid proximity of towns and paved roads, suggesting that the species is sensitive to human disturbance. Vineyards are the main land use in this region and the intensity is increasing. Our results suggest vineyards should be limited in areas around these wetlands so that gull–billed terns may forage in their preferred sites. Key words: Breeding colony, Agricultural landscape, Intensive vineyards, Foraging, Habitat selection, Temporary lake Resumen Selección del hábitat alimentario de la pagaza piconegra (Gelochelidon nilotica) en el centro de España (Castilla–La Mancha). La pagaza piconegra se reproduce en lagos temporales de Castilla–La Mancha, pero depende de los hábitats terrestres de los alrededores para alimentar a los pollos. Por consiguiente, es fundamental conocer el tipo de ambientes que selecciona para capturar presas con las que alimentarlos. La finalidad de este estudio fue evaluar la utilización que los adultos de la pagaza piconegra hacen de los hábitats para la caza. De los 66 lagos estudiados entre 1996 y 2016, constatamos que la pagaza se reproducía en 12. Durante este período, cada lago se utilizó para 1–14 temporadas de cría. Seleccionamos las zonas circulares alrededor de los tres humedales en los que la especie crio en 2013 y 2014. Dentro de estos círculos, muestreamos 60 puntos aleatorios y registramos 125 contactos con pagazas piconegras (incluidas entre 1 y 39 aves). Estimamos las mismas variables ambientales en los puntos de contacto y los aleatorios, con inclusión del uso de la tierra y la distancia al humedal más cercano, a la colonia más cercana y a varios tipos de usos antrópicos (carreteras asfaltadas, viviendas y ciudades). Para evaluar la selección del hábitat, calculamos el índice de selección de Manly para las variables de uso del suelo y utilizamos modelos lineales mixtos para evaluar las diferencias entre las variables de distancia. El principal uso de la tierra que la pagaza piconegra seleccionó para la alimentación fue el cultivo de cereales, mientras que evitó los viñedos. De media, las aves se alejaban para alimentarse hasta 2 km de las colonias y tendían a evitar la proximidad de ciudades y carreteras asfaltadas, lo que sugiere que la especie es sensible a las perturbaciones antrópicas. Los viñedos constituyen el ISSN: 1578–665 X eISSN: 2014–928 X

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principal uso de la tierra en esta región y su extensión está aumentando. Nuestros resultados sugieren que los viñedos se deberían limitar en zonas cercanas a estos humedales, para que la pagaza piconegra pueda alimentarse en sus lugares preferidos. Palabras clave: Colonia de reproducción, Paisaje agrícola, Viñedos intensivos, Alimentación, Selección de hábitat, Lago temporal Received: 10 VII 17; Conditional acceptance: 20 IX 17; Final acceptance: 15 XII 17 Vanessa Oliveira Britto, José Antonio Gil–Delgado, ICIBYBE/Dept. of Microbiology and Ecology, Univ. of Valencia, c/ Catedrático José Beltrán 2, 46980 Paterna, Valencia, Spain.– Rafael Ubaldo Gosálvez, Dept. of Geography and Land Planning, Univ. Castilla–La Mancha, Avda. Camilo José Cela s/n., Ciudad Real.– Gérman M. López–Iborra, Dept. of Ecology/IMEM Ramon Margalef, Univ. of Alicante, Apartado 99, E–03080 Alicante, Spain.– Angel Velasco, Dept. of Environmental Science, Univ. Castilla–La Mancha, Av. Carlos III s/n., Campus Fábrica de Armas, edifc. Sabatini Lab. 0.3 CP 45071 Toledo, Spain. Corresponding author: V. O. Britto E–mail: vanessa.obritto@gmail.com

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Introduction Habitat selection is closely linked to the need to extract resources necessary to complete life cycles. It is therefore necessary that wildlife managers are aware of a species' preferred habitats (Caughley, 1994). Habitat selection is the innate or learned behavioral response that allows a bird to choose among the various environmental components, habitats or structures in a location that will influence survival or adaptation (Block and Brennan, 1993). The selection of breeding places seems to be a major factor in the breeding strategy of gulls and terns (Vargas et al., 1978; Goutner, 1991; Sánchez et al., 1991; Erwin et al., 1998), since many potential breeding sites are highly unstable and unpredictable habitats. Despite this, many individuals often use the same locality for years or decades, occupying the same places in consecutive breeding seasons, while in some years, colonies move to new breeding sites (Sánchez et al., 2004; Corbacho et al., 2009). The gull–billed tern, Gelochelidon nilotica, has a worldwide distribution range, but its breeding colonies show a patched spatial distribution (Del Hoyo et al., 1996). In Spain, the gull–billed tern breeds in the Delta del Ebro, in a few temporary lakes in Central Spain, in the south of Andalusia, and in a few sites in Extremadura and Castilla–León (Martí and Del Moral, 2003). Breeding colonies tend to be located on beaches, wetlands and sedimentary islands but they are also found in man–modified habitats (Moller, 1981; Cramps, 1985; Palacios and Mellink, 2007). The gull–billed tern breeds in monospecific or mixed colonies, with other waterbirds (Vargas et al., 1978; Sánchez et al., 2004; Molina et al., 2009; Barati et al., 2012). They forage at distances ranging from 2 km to 20 km (Fasola and Bogliani, 1990) of breeding places. The population of gull–billed tern in Central Spain appears to be increasing (Corbacho et al., 2009), but there is some also evidence of a reversing trend (Del Hoyo et al., 1996). The agricultural intensification that is occurring in Central Spain (Ruiz–Pulpón, 2013, 2015) could be related to a decrease in appropriate feeding habitats for the Gull–billed tern, since the lakes with breeding colonies are surrounded by agriculture fields, mainly vineyards and cereal crops (Ruiz–Pulpón, 2013, 2015). Changes in the landscape affect some bird species, especially those that breed in regions with intensive agriculture (Chamberlain et al., 2000), because these fields are prone to changes linked to crops considered more profitable. To provide information that may help to preserve breeding colonies and their breeding success (Litvaitis, 2000; Molina et al., 2009, 2010), in this study we aimed to determine which habitats surrounding wetlands were preferentially used for hunting by the gull–billed tern and which habitats were avoided Study area The study area is located in the Reserva de La Biosfera Mancha Húmeda (hereafter RBMH), a wetland–rich area in Central Spain that stretches

303

over almost 7.551 ha (GIA, 2015) with 117 humid zones. Most Castilla–La Mancha lakes are temporary and salty, and face drought periods which may dry them for years (Cirujano and Medina, 2002; Cirujano and Cobelas, 2011). We monitored 66 wetlands in this region. In some cases, this monitoring began in 1996, but it began ten years ago in most cases (fig. 1). These wetlands are located in a large agricultural region whose main crops are vineyards and cereals. Fields with tree species, such as olive trees (Olea europaea), are scarce. Only 12 of the 66 wetlands had breeding colonies of gull–billed tern in one or more years (fig. 2). We recorded the number of couples that bred from 2007 to 2016 in all the wetlands. Of the wetlands used in 2013–2014, we selected four to study habitat selection in the following two breeding seasons. One of these wetlands (Camino de Villafranca) was not occupied in the following years, reducing the number of wetlands studied to three: Manjavacas (39° 24' 54'' N, 2° 51' 59'' W), Mermejuela (39° 32' 22'' N, 3° 8' 18'' W) and Longar (39° 42' 10'' N, 3° 19' 31'' W). Data collection We used a 7–km radius circle, centered in each study lake, to record habitat use and availability. In other habitats, gull–billed terns may search for food farther than this distance (Molina and Marschalek, 2003), but in our study area, a larger radius would encompass parts of nearby towns. Data collection was carried out in spring and summer, from April to June, in 2015 in Longar and in 2015 and 2016 in the other two lakes. Overall, we covered 27 itineraries, totalling 2,248.46 km (mean = 83.30 km, SD = 48.1). We used a GIS (Geographic Information System) to delimit the itineraries in such a way that they were distributed across the entire surface of the circle. Itineraries were visited by car as many times as needed to reach a minimum of 30 foraging observations per circle of each lake. Contacts with gull–billed terns looking for food, identified through their hunting flight, were positioned by GPS (Garmin ETREX30). The hunting flight is characterized by the bird flying slowly, with head down and frequent dipping movements towards land, corresponding to attempts to catching prey (Cramp, 1985; Molina and Marschalek, 2003; Molina et al., 2009). Contacts with birds in direct displacement flight, when they did not seem to be inspecting territory, were not considered in the analyses. At every contact point (hereafter CP), the following variables were obtained: the number of birds searching for food and the type of environment (land use type) in which they fed. We used GIS to estimate distance from the individual bird or flock to: i) the nearest breeding colony; ii) to nearest paved road; iii) to nearest urban place (towns and villages); iv) to nearest wetland (used or not for breeding); and v) to the nearest isolated building (including abandoned buildings, inhabited houses or shelters for cattle) (table 1). To evaluate habitat availability, we selected 20 random points (hereafter RP) within the 7–km

Britto et al.

38º

35º 9ºW 6º 3º 0º 3º 6ºE

l üe ig 2 G

iana

Guad

550000

Ja b

4400000

35º

500000

Záncara

41º

450000

Tajo

4350000

41º

400000

4350000

44º

4300000

9ºW 6º 3º 0º 3º 6ºE

4300000

44º

4400000

304

400000

ón

450000

500000

UTM Projection and Grid zone 30 ETRS89 Datum

550000 20

40 km

Fig. 1. Left, map of Spain with black square marking the study area. Right, map of Castilla–La Mancha wetlands showing the three sites (1, Manjavacas; 2, Mermejuela; 3, Longar) where habitat selection was studied: l breeding colonies (studied lakes); l breeding colonies (previous studied years); ¡ lake without breeding colonies. Fig. 1. Izquierda: mapa de España con la zona de estudio indicada con un cuadro negro. Derecha: mapa de los humedales de Castilla–La Mancha en el que se muestran los tres lugares (1, Manjavacas; 2, Mermejuela; 3, Longar) en los que se estudió la selección de hábitat: l colonias de cría (lagos estudiados); l colonias de cría (años estudiados previamente); ¡ lago sin colonias de cría.

radius circle centered in each of the three wetlands studied. For this purpose, we randomly selected 20 random points, distributed along the length of the itinerary previously defined with GIS around each lake, using the random numbers function in Excel. At each of these points, this function was also used to randomly select the left or right side of the itinerary and a random perpendicular distance between 0 and 80 m from the point. Thus random points were situated within a 160 m wide band centered in the itineraries. This band was selected because gull–billed terns were detected from the itineraries within it, and in this way random points were selected within the area where terns were detectable. In RP, we estimated the same variables as in CP. Given that Manjavacas and Mermejuela were monitored in 2015 and 2016, the same RP were visited in the second year to check if type of land use had changed, but in all cases it remained the same. Land use types considered (table 1) comprised two types of vineyards (traditional goblet vineyards and vertical trellis vineyards, in which canes are secured to trellis wires running the length of the row of vines), grain fields in different stages, and fallows (uncultivated fields resting between harvesting seasons). The wetland edge category includes areas with low vegetation surrounding lakes that may be flooded after high rainfalls.

Data analysis Manly’s Index (Manly et al., 2004) was used to evaluate the selection of land use types. This index computes the relation between available environments and those used by the species, as in equation 1:

Wj =

uj aj

(eq. 1)

where uj is the proportion of use of habitat in category j and aj is the availability of habitat j. These indices were calculated using package adehabitat HS (Calenge, 2006) in R (R development Core Team, 2016). We used linear mixed models to evaluate whether distance variables differed between RP and CP. In these models, each distance variable was included as a dependent variable and the type of point (coded RP as 0 and CP as 1) as a fixed factor. Lake was included as a random factor. We used the lme function in the R nlme package (Pinheiro et al., 2017) to fit these models. Results The gull–billed tern bred in 12 of the 66 temporary wetlands monitored (fig. 1). Colonies used each lake from 1 to 8 times (fig. 2). Some colonies occupied the

Animal Biodiversity and Conservation 41.2 (2018)

305

1996 1997 1998 1999 2000 2001 2002 2003 2004 2005 2006 2007 2008 2009 2010 2011 2012 2013 2014 2015 2016

Wetlands

Pajares

108 240

Salicor Camino de Villafranca

176

195

120

70 3

Mermejuela 200 473

Longar 215

Manjavacas

239

170

145

106

221

258

183

820

200

322 120

238

137

200 140

220 167

Huevero 12

Inesperada

Cucharas 350 325 350

Peñahueca

Larga de Villacañas

Not breeding

35 10

190

Quero 26

Breeding

No registry

Attempt to breed

Fig. 2. Monitoring data of the 12 wetlands used as breeding colonies by gull–billed tern. Fig. 2. Datos de seguimiento de los 12 humedales empleados como colonias de reproducción por la pagaza piconegra.

Table 1. Description of habitat variables (land uses and minimum distances to potentially relevant territory features) used in this study. Tabla 1. Descripción de las variables del hábitat (usos de la tierra y distancias mínimas a características del territorio potencialmente relevantes) empleadas en este estudio. Substrate

Code

Description

Land use Cereal production

Fallows

Crops of oats or wheat Unploughed cereal fields harvested at least one year

ago and with dense herbaceous cover

Wetland

Naturally flooded areas

Wetland edge Ploughed fields Traditional vineyards

W WE P TV

Herbaceous plants and bushes around wetlands Ploughed fields, mostly without vegetation Non–irrigated vineyards grown in a traditional way

(goblet) Intensive vineyard

Irrigated vineyards grown on metal trellises

Trees

Stands of Pinus ssp, Prunus dulcis or Olea europaea

Distance (km) Distance to breeding colony Distcolony Distance to closest breeding colony Distance to wetland Distance to paved roads

Distwetland Distance to closest wetland (used or not for breeding) Distroad

Distance to closest roads

Distance to urban places

Disttowns Distance to closest town or village

Distance to houses

Disthouse Distance to closest abandoned building, inhabited

house or shelter for cattle

Britto et al.

306

Table 2. Percentage of random (RP) and contact (CP) points where each habitat type was present in the 7–km radius circle and distances to variables (mean ± SD). Number of contact points in parentheses. Number of random points is 20 in each lake. Selection index shown is the Manly index (Wj), its standard error (SE) and significance level (p). Land uses W and WE were not considered to calculate selection indexes as no random points were selected in therein. Table 1 shows the codes of variables. Variable Land use

Manjavacas CP (45)

Mermejuela

RP (20)

CP (45)

RP (20)

31.11

11.11

26.67

–

4.44

–

13.33

–

8.89

35.56

–

6.67

2.22

–

Distance Distcolony

2.73 ± 1.52

3.21 ± 1.19

1.36 ± 1.35

4.10 ± 1.26

Distwetland

1.77 ± 1.43

2.09 ± 1.01

1.36 ± 1.35

3.56 ± 1.12

Distroad

1.16 ± 0.89

1.60 ± 1.23

2.51 ± 1.00

1.35 ± 1.17

Disttowns

6.65 ± 1.49

6.57 ± 1.38

6.23 ± 2.76

4.60 ± 3.16

Disthouse

0.45 ± 0.35

0.56 ± 0.55

0.98 ± 0.93

0.47 ± 0.53

same wetland in consecutive years. In some cases, colonies appeared to have moved to other wetlands in the same breeding season owing to flooding of nesting sites. For instance, in 2007, the gull–billed terns abandoned the Camino de Villafranca colony because heavy rains in May flooded the sedimentary island where the colony was settled. In parallel, in the Salicor wetland, this rainfall turned some areas that were previously connected with surrounding fields into islands. A gull–billed tern colony settled on these islands soon after the abandonment of the breeding colony at Camino de Villafranca. This suggests a movement of the gull–billed tern colony of Camino de Villafranca toward Salicor Lake. In 2010, heavy May rains flooded the sedimentary islands at Manjavacas, causing the desertion of its colony. A colony then settled in Huevero Lake, likely including the individuals that left Manjavacas. Foraging habitat A total of 136 contacts with gull–billed terns were recorded; 11 of these were located outside the 7 km radius and are not considered here. The size of gull– billed tern flocks searching for food ranged from 1 to 39 individuals (mean = 3.4; SD = 4.5; N = 125). The largest flock observed was following a tractor in a grain field. Distances between the contacts and

the nearest colony ranged from 0.03 to 6.24 km (mean = 1.98; SD = 1.50; N = 125). The gull–billed tern showed a positive selection towards cereal fields (p = 0.032), clearly avoiding traditional vineyards (p = 0.000) (table 2). Ploughed fields presented the highest selection index, but it was not significant. Fallows were used according to their availability. No use of intensive vineyards or areas with trees was detected. These habitat preferences were similar in the three studied sites (table 2). The analysis of distance variables showed that gull–billed tern seeks food closer to the colony than expected according to random points (p < 0.0001), at 2 km on average. On the contrary, they were found farther than expected from paved roads (p < 0.0001) and towns (p = 0.0195), even if these were near colonies (fig. 3). The variable distance from solitary houses scattered in the fields had no effect (p = 0.8482) (table 2, fig. 3). Discussion Habitat selection We observed that cereal crops were the preferred habitat for hunting in the agricultural landscapes studied, and this pattern was similar between the

Animal Biodiversity and Conservation 41.2 (2018)

307

Tabla 2. Porcentaje de puntos aleatorios (RP) y de contacto (CP) en los que cada tipo de hábitat estaba presente en un radio de 7 km y las distancias a las variables (media ± DE). Número de puntos de contacto entre paréntesis. El número de puntos aleatorios es de 20 en cada lago. Se muestran el índice de selección de Manly (Wj), su error estándar (SE) y el grado de significación (p). Los usos de la tierra W y WE no se tuvieron en cuenta para calcular los índices de selección porque no contenían puntos aleatorios. En la tabla 1 se muestran los códigos de las variables. Longar CP (35)

Total RP (20)

Selection index RP

Wj SE p

42.86

44.80

3.24 1.04 0.032

0.83 0.24 0.492

19.20

2.86

–

2.40

– – –

2.86

–

5.60

– – –

31.43

24.80

–

3.20

–

6.67 30 10

4.04 2.04 0.138 0.11

0.06

0.000

0 – –

13.33 0 – –

F df P

1.81 ± 1.25

3.41 ± 1.06

1.98 ± 1.50

3.57 ± 1.22

53.38

1.181

< 0.0001

1.63 ± 1.28

2.86 ± 1.35

1.59 ± 1.36

2.84 ± 1.29

35.41

1.181

< 0.0001

2.50 ± 1.60

0.94 ± 0.69

2.02 ± 1.33

1.30 ± 1.08

15.13

1.181

< 0.0001

4.40 ± 2.82

3.52 ± 1.79

5.87 ± 2.56

4.90 ± 2.55

5.55

1.181

0.0195

0.92 ± 0.61

1.41 ± 1.47

0.77 ± 0.71

0.82 ± 1.04

0.03

1.181

0.8482

study colonies. Clearly, gull–billed terns avoided both vineyards types present in the study area, since we only found a slight percentage of contacts in traditional vineyard and we did not detect any use of intensive vineyards. Avoidance of vineyards could be explained by several factors. On the one hand, cereal fields are less intensive crops than vineyards, and it is likely that availability of prey is higher in the former. Differences in composition of the arthropod community may also be important, since gull–billed terns prey mainly on large ground–dwelling insects (beetles and grasshoppers, Britto et al., in prep.) while canopy–dwelling arthropods are more abundant in vineyards (Nash et al., 2010). In addition, vertical habitat structure may have an important effect on habitat selection by this species. Gull–billed terns capture prey in flight and do not pursue them on foot (Molina et al., 2009), so searching for food over open areas, such as cereal crops or plowed fields, may facilitate prey capture, while woody vegetation (such as trees and vineyards) would make it more difficult. The dense vegetation cover of vineyards in spring may limit prey visibility and, in intensive vineyards, the vertical metal posts on which branches are fixed to wires running the length of the row of vines may limit birds' flight. Our results show that gull–billed terns forage mainly in specific land use types (cereal crops, ploughed fields and fallows accounted for 89 % of contacts) in

areas surrounding the colonies, so chicks depend on preys captured by the parents n these zones (Vargas et al., 1978; Fasola et al., 1989; Díes et al., 2005; Aourir et al., 2013). On average, in our studied colonies, parents fly about 2 km to look for their prey and rarely more than 6 km from the colony. In an Italian coastal lagoon, Fasola and Bogliani (1990) found that the gull–billed tern was one of the species, together with the Little Tern, whose density decreased faster with distance from the nesting wetland, although in this area average distance (9.3 km) was higher than in La Mancha. In California, individuals foraged at least 8–9 km from the nesting colony (Molina and Marschalek, 2003). The longer distance covered for hunting in the studies of Fasola and Bogliani (1990) and Molina and Marschalek (2003) could be explained because these colonies lived in coastal habitats where they catch larger prey that are more profitable energetically (crustaceans, fishes and lizards). In our study area, the main preys are insects and these smaller preys could set a shorter limit to the hunting distance. The changes in land use within the belt around the breeding site may have a strong effect on colony success. In particular, according to our results, agricultural changes favoring intensive vineyards could have the worst effect. If intensive vineyards were extended to surround the breeding wetlands, the cost of displacement of adults to seek food could increase

Britto et al.

Frequency (%)

Distance to breeding colony 45 40 35 30 25 20 15 10 5 0 0 1 2 3 4 5 6 km Distance to paved roads

50 40 30 20 10 0

3 km

Distance to wetland Contact points (CP) Random points (RP)

3 4 km

Distance to houses 90 80 70 60 50 40 30 20 10 0

30 Frequency (%)

Frequency (%)

45 40 35 30 25 20 15 10 5 0

Frequency (%)

308

2 km

Distance to urban places

25 20 15 10 5 0

0 1 2 3 4 5 6 7 8 9 10 111 5 22 km

Fig. 3. Frequency distribution (%) of the distance variables (km) analyzed in this study. Fig. 3. Distribución de la frecuencia (%) de las variables de distancia (km) analizadas en este estudio.

to the point of not being energetically profitable and lakes at the study area could become unsuitable for breeding. Intensive vineyards are widespread in the RBMH, and they are extending in detriment of other crop types, including traditional vineyards (Ruiz–Pulpón, 2013, 2015). Other bird species are also reported to avoid vineyards (García et al., 2006; Benítez–López et al., 2017) and we agree with these authors that the establishment of intensive vineyards should be prohibited in areas that are most appropriate for endangered birds. The gull–billed tern does not seem to choose a feeding habitat independently of its surroundings. In the agricultural landscapes where human interference is constantly present, we found a significant effect of several landscape features. The species avoids feeding close to towns, villages and paved roads; however, it is not affected by isolated farm houses. The lack of effect of the proximity of farm houses can be explained by the fact that they are scattered and human presence is sporadic. In a meta–analysis of infrastructural effects, Benítez–López et al. (2010) detected a negative effect of roads and other infrastructures on bird abundance that extended up to 1 km. Our results agree with this result because we found fewer contacts than expected within 1 km from roads. The effects of roads on wildlife are multiple (Trombulak and Frissell, 2000), but given the mobility of gull–billed terns and the relative low traffic on most of these roads, we hypothesize that in our study, a

likely effect is the reduction in prey availability through changes in conditions near roads. The same could be true for the distance from towns, where the percentage of contacts is lower than the percentage of random points within the first four kilometers. Although in some cases fallows and cereal fields reach the edges of towns, they are usually surrounded by land use types that are not so attractive for this species. Distance variables were not analyzed in other studies of gull– billed terns, so we do not know if the effect of these man–made structures is similar in other landscapes. In our study area, where rainfall and water level at lakes varies from year to year, it is not uncommon that gull–billed tern colonies are forced to abandon a wetland and establish a new breeding colony in another lake within the same breeding season. These changes seem to be associated to adverse weather conditions. Spring abundant rainfalls may change the availability of suitable islands for nesting by flooding sedimentary islands harboring a gull–billed tern colony. In other lakes, the same rains may induce the formation of new sedimentary islands by isolating pieces of land from the shore. Besides our study, other authors have also reported nesting site changes within the same breeding season (Costa, 1986; Sánchez et al., 2004). Therefore, the possibility arises of an interaction between the environmental factors affecting nesting place selection inside the lake and habitat selection in their surroundings. Further research is needed to assess whether, in these dry

Animal Biodiversity and Conservation 41.2 (2018)

Mediterranean landscapes, the lack of appropriate habitats for feeding around the lakes would limit the chances to move colonies between lakes when rainfalls alter wetlands suitability for breeding. Conservation implications This study offers new information about gull–billed tern ecology that may help in the design of regional and local level conservation management strategies. Habitat selection analysis allowed us to identify factors that determine habitat preferences for hunting and the potential influence of human disturbance on this species. At a landscape scale, our results call for limitations in agricultural intensification around potential nesting places. Vineyards should be limited or avoided in areas surrounding these wetlands. Given the importance of rotation of cereal crops and fallow lands as hunting areas, these uses should be maintained and incentivized in lands surrounding wetlands in this region. In addition, the lakes where breeding colonies regularly settle must be protected. The awareness of farmers and people is extremely important for the implementation of conservation strategies of this species that should involve the implementation of specific agri–environment schemes. Acknowledgements The authors would like to thank the following colleagues for support with fieldwork: M. S. Gonçalves, P. Pons, A. Paredes and A. O. Britto. We also acknowledge the receipt of a PhD study grant from the CNPq–Conselho Nacional de Desenvolvimento Científico e Tecnológico (No. 200004/2014–0) for V. O. Britto. This study was jointly supported by the Spanish Ministry of Economy and Competitiveness and the European Regional Development Fund (FEDER) 'One way to make Europe', through the projects: CLIMAWET (Climate change mitigation and adaptation in the main types of Iberian Mediterranean wetlands: carbon budged and response models of species and habitats, CGL2015–69557–R) and ECOLAKE (Ecological patterns in endorheic lakes: keys to their conservation, CGL2012–38909). References Aourir, M., Radi, M., Znari, M., 2013. Foraging habitat and diet of Gull–billed Tern, Gelochelidon nilotica, during the nesting period in Sebkha Zima, West– central Morocco. Ecologia mediterrânea, 39: 31–38. Barati, A., Etezadifar, F., Esfandabad, B. S., 2012. Nest–site and hatching success at a mixed–species colony of Black–wingerd Stilts Himantopus himantopus and Gull–billed Tern Gelochelidon nilotica. Avian Biology Research, 5: 142–146. Benítez–López, A., Alkemade, R., Verweij, P. A., 2010. The impacts of roads and other infrastructure on mammal and birds populations: a meta–analysis. Biological Conservation, 143: 1307–1316. Benítez–López, A., Viñuela, J., Mougeot, F., García,

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Brief communication 311

A predation event by free–ranging dogs on the lowland tapir in the Brazilian Atlantic Forest A. Gatti, J. B. Seibert, D. O. Moreira

Gatti, A., Seibert, J. B., Moreira, D. O., 2018. A predation event by free–ranging dogs on the lowland tapir in the Brazilian Atlantic Forest. Animal Biodiversity and Conservation, 41.2: 311–314. Abstract A predation event by free–ranging dogs on the lowland tapir in the Brazilian Atlantic Forest. Dogs are associated with humans and human–dominated landscapes, and when they become feral and free–roaming, they can negatively impact wildlife through processes such as predation. In this study, we report a predation event of free–ranging dogs on lowland tapirs in a protected area (PA) of the Brazilian Atlantic Forest. As tapirs can be vulnerable to dog attacks, especially in a protected area surrounded by farming activities, research programmes and monitoring of these areas are crucial to understand the impact of free–ranging domestic species on wildlife. Additionally, education programs and dog control should be incorporated into conservation plans in such areas around PAs. Key words: Conservation, Tropical forest, Alien species, Canis familiaris, Predation, Protected area, Tapirus terrestris Resumen Un evento de depredación del tapir por perros criados en libertad en parches forestales de mata atlántica de Brasil. Los perros se relacionan con los humanos y los paisajes dominados por humanos, y cuando se asilvestran y crían en libertad, pueden incidir negativamente en la fauna silvestre mediante procesos como la depredación. En el presente estudio, describimos un evento de depredación del tapir por perros criados en libertad en una zona protegida (PA) de la mata atlántica del Brasil. El tapir puede ser vulnerable a los ataques de los perros, en especial en zonas protegidas rodeadas de actividades agrícolas, y es fundamental hacer un seguimiento de estas zonas y analizarlas para comprender los efectos de las especies domésticas criadas en libertad en la fauna silvestre. Además, deberían incorporarse programas de control y educación canina en las cercanías de las zonas protegidas a los planes de control de dichas zonas. Palabras clave: Conservación, Bosque tropical, Especies exóticas, Canis familiaris, Depredación, Área protegida, Tapirus terrestris Received: 10 VIII 17; Conditional acceptance: 17 X 17; Final acceptance: 18 XII 17 Andressa Gatti, Jardel B. Seibert, Danielle O. Moreira, Programa de Pós–Graduação em Ciências Biológicas (Biologia Animal), Depto. de Ciências Biológicas, Univ. Federal do Espírito Santo, Av. Fernando Ferrari, nº 514, Vitória, ES 29075–910, Brazil. Corresponding author: Andressa Gatti. E–mail: gatti.andressa@gmail.com

Introduction and/or invasion of exotic species is a major threat to biodiversity worldwide as a consequence of the increasing human population around, and even in, protected areas (Gompper, 2013; Roy et al., 2014). One species that has had a serious impact on biodiversity is the domestic dog (Canis familiaris Linnaeus, 1758), ISSN: 1578–665 X eISSN: 2014–928 X

the most abundant carnivore in the world (Young et al., 2011; Paschoal et al., 2016). Dogs are strongly linked to humans and human–dominated landscapes, in both rural and urban settings (Lacerda et al., 2009). Domestic dogs protect property and reduce human– wildlife conflicts by protecting livestock from people or © 2018 Museu de Ciències Naturals de Barcelona

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predators (Paschoal et al., 2016). However, when they are neglected or left without human care, dogs often become feral or free–roaming (Young et al., 2011), seriously affecting the wildlife (Gompper, 2013). Interactions between domestic dogs and wildlife can be negative, especially in areas where dogs can roam freely throughout the landscape (Paschoal et al., 2016). Interactions between dogs and native species, for example, can result in competition (Lessa et al., 2016), as well as disease transmission and predation (Young et al., 2011; Lessa et al., 2016). Predation on ungulates by domestic dogs has previously been reported elsewhere (e.g., Buuveibaatar et al., 2009; Lacerda et al., 2009; Young et al., 2011; Silva–Rodríguez and Sieving, 2012; Lessa et al., 2016). However, the extent of the problem is poorly reported in South America. In Brazil, for example, attacks on the lowland tapir Tapirus terrestris (Linnaeus, 1758) have been reported only for the Cerrado biome (Lacerda et al., 2009; Lessa et al., 2016). Here, we describe the first report of predation by domestic dogs on lowland tapir in a protected area of the Brazilian Atlantic Forest. The lowland tapir represents an important functional group in the Neotropical region, because it feeds on a large variety of fruits and plant parts (Tobler et al., 2010), making them the largest surviving frugivores in tropical forests. However, the species is considered 'Vulnerable', both globally and in Brazil (Naveda et al., 2008; ICMBio, 2016), and is 'Endangered' in the Atlantic Forest (Medici et al., 2012). In the Atlantic Forest, the lowland tapir populations are distributed among very fragmented areas and threatened by high hunting pressure, traffic collisions, and rapid urbanization (Medici et al., 2012). The Biological Reserve Córrego do Veado (BRCV) protects 2,342 ha of forest and is located in the municipality of Pinheiros, northwestern Espírito Santo (40° 8' W 18° 22' S), Brazil. The BRCV is a federal protected area (IUCN category Ia), without direct human interference, and dedicated to the preservation of biological diversity. The BRCV is the last remnant of forest in the region as natural vegetation has been largely replaced by agriculture and pasture. A variety of mammals remain in the protected area, however, such as anteaters (Tamandua tetradactyla), titi monkeys (Callicebus personatus), ocelots (Leopardus pardalis), deer (Mazama sp.), white–lipped peccary (Tayassu pecari), and lowland tapirs. On June 20th 2012 , during a fieldwork campaign in the BRCV, we found two lowland tapirs under attack by a pack of approximately five dogs. The attack occurred about 300 m east of the reserve headquarters at around 7:00 a.m. (local time; UTC/GMT–3 h). The volunteers from the fire brigade of the BRCV alerted our team to the attacks. One of the tapirs was an adult female and the other was a young female, approximately 18 months old (fig. 1). The adult female managed to escape, but when our team arrived, the young tapir (which was standing in a stream) was still being attacked by two dogs, only one of which was caught and removed by the BRCV employers. The young tapir had several superficial injuries on

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its body, including head, limbs, and tail (fig. 1), and it was treated in situ by veterinarians. Soon after its recovery, the young tapir was released some 30 m into the forest from where we found it. About two hours passed from the moment we intervened in the attack until the young tapir was released. Considering the presence of dogs in BRCV, we expect attacks on tapirs and other wildlife to become more common in this area for several reasons. First, the BRCV is a totally isolated forest fragment that is easily invaded. It is located within a landscape matrix of cattle pastures, crops (such as coffee, cassava and papaya), and eucalyptus and seringa (rubber) plantations (Flesher and Gatti, 2010). Second, it is surrounded by small and large farms from which dogs can likely roam freely. Third, the forest is frequented by poachers and their dogs (A. Gatti, pers. obs.). Fourth, the shape of the reserve and the type of the surrounding matrix increase its accessibility to people and domestic animals. Lowland tapirs can be vulnerable to dog attacks as both species occur in a variety of habitats within landscapes (Lacerda et al., 2009), increasing their encounter rate, especially just beyond reserve edges. In these areas, lowland tapirs can take advantage of the heterogeneous landscape provided by the nearby matrix and disturbed areas, where they often feed in plantations and gardens (Lacerda et al., 2009). This is true for the BRCV, where farmers frequently complain about tapirs feeding on rubber trees and cassava plantations (Flesher and Gatti 2010; A. Gatti, pers. obs.). This situation prompts them to keep dogs to protect their livelihood, but increases the threats to the protected area. Also in these protected area edges, our team confirmed the recurring presence of lowland tapirs through direct observation and footprints on the dirt roads along the reserve and along the streams. Assuming that lowland tapir population should have between 30 and 200 individuals to be demographically and genetically viable in areas of the Atlantic Forest, it has been suggested that, in the same way, viable populations would require habitat areas with at least 7,500 and 50,000 ha, respectively (Gatti et al., 2011). Because of the aforementioned problems, and considering the minimum population size, we expect that the lowland tapir population in the BRCV are neither demographically nor genetically viable in the long–term (the BRCV is an area of 2,382 ha), and that the presence of free–ranging dogs can act negatively in tandem with other types of impact (e.g., poaching) on the lowland tapir population, as well on the forest diversity (e.g., Gompper, 2013). Their local or functional extinction means loss of seed dispersal agents, especially for plant species that have large fruits and seeds (Bueno et al., 2013), and may also disrupt long–distance dispersal of many plant species (Giombini et al., 2016). Furthermore, regardless of the degree of threat in a given protected area, the presence of dogs prompts the need for discussions on methods to mitigate their negative impact, many of which are controversial. For example, Zapata–Ríos and Branch (2016) suggested

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Fig. 1. A young individual of Tapirus terrestris with injuries after being attacked by a pack of dogs in the Biological Reserve of Córrego do Veado, southeastern Brazil. Fig. 1. Individuo joven de Tapirus terrestris con lesiones después de ser atacado por una jauría de perros en la Reserva Biológica de Córrego do Veado, sureste de Brasil

that robust populations of the mountain tapir Tapirus pinchaque and other native species in Cayambe–Coca National Park in the northern Ecuadorian Andes, may persist in the long–term, but only through control and eradication of domestic dogs. However, such actions are in generally expensive, rarely effective and non– viable in the long–term (Gompper, 2013). Given the above, we recommend law enforcement to control the presence of dogs in protected and surrounding areas and its deep consideration by managers and stakeholders. For example, there is currently no mention of management or actions to control dogs in the BRCV management plan, other than the acknowledgement that ''[...] the susceptibility to invasions by domestic animals, such as cattle and dogs from neighboring farms, are [...] weaknesses with a high degree of intensity [...]'' (Ministério do Meio Ambiente, 2000; encarte 6, p. 5). We also suggest that studies aimed at estimating the density of dogs in the BRCV and other protected areas of the region be performed to identify the potential effects of free–ranging dogs on the abundance and behavior of native mammals, including the lowland tapir. Additionally, management plans for protected areas and wildlife key conservation actions should include specific measures to minimize dog incidence in protected areas. Furthermore, for managers, we recommend special efforts be made to inform and educate the population living in surrounding protected areas about this problem, particularly the dog owners. Likewise, engaging local people to participate in con-

servation actions should be encouraged. Paschoal et al. (2012) emphasized that, in this particular case, community involvement needs to be mandatory. Understanding how the existing conflict occurs is one way to reduce the pressure on the remaining populations of lowland tapir in different ecoregions, especially in the Atlantic Forest. Acknowledgements We wish to thank the Biological Reserve Córrego do Veado, the Instituto de Ensino, Pesquisa e Preservação Ambiental Marcos Daniel (IMD), the veterinarians Igor Acosta and Herbert Soares, and field assistant José Roberto de Oliveira. We are also grateful to Ryan Huang for revising the English and the two anonymous reviewers for their suggestions and comments. Fieldwork was supported by FIBRIA Celulose S. A. This study was conducted with the authorization of the Biodiversity Authorization and Information System (Sisbio–ICMBio/IBAMA nr. 25642). References Bueno, R. S., Guevara, R., Ribeiro, M. C., Culot, L., Bufalo, F. S., Galetti, M., 2013. Functional redundancy and complementarities of seed dispersal by the last Neotropical Megafrugivores. PLoS ONE, 8(2): e56252.

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Buuveibaatar, B., Young, J. K., Fine, A. E., 2009. Mongolian saiga in Sharga Nature Reserve: are domestic dogs a threat to saiga? Mongolian Journal of Biological Sciences, 7(1–2): 37–43. Flesher, K. M., Gatti, A., 2010. Tapirus terrestris in Espírito Santo, Brazil. Tapir Conservation, 19/1(26): 16–23. Gatti, A., Brito, D., Mendes, S. L., 2011. How many lowland tapirs (Tapirus terrestris) are needed in Atlantic Forest fragments to ensure long–term persistence? Studies on Neotropical Fauna and Environment, 46(2): 77–84. Giombini, M. I., Bravo, S. P., Tosto, D. S., 2016. The key role of the largest extant Neotropical frugivore (Tapirus terrestris) in promoting admixture of plant genotypes across the landscape. Biotropica, 48(4): 499–508. Gompper, M. E., 2013. Free–ranging dogs and wildlife conservation. Oxford University Press, Oxford, UK. ICMBio, 2016. Sumário executivo do livro vermelho da fauna brasileira ameaçada de extinção. Instituto Chico Mendes de Conservação da Biodiversidade (ICMBio), Brasília. Lacerda, A. C. R., Tomas, W. M., Marinho–Filho, J., 2009. Domestic dogs as an edge effect in the Brasília National Park; Brazil: interactions with native mammals. Animal Conservation, 12: 477–487. Lessa, I., Guimarães, T. C. S., Bergallo, H. de G., Cunha, A., Vieira, E. M., 2016. Domestic dogs in protected areas: a threat to Brazilian mammals? Natureza & Conservação, 14(2): 46–56. Medici, E. P., Flesher, K., Beisiegel, B. M., Keuroghlian, A., Desbiez, A. L. J., Gatti, A., Pontes, A. R. M., Campos, C. B., Tófoli, C. F., Moraes, E. A., Azevedo, F. C., Pinho, G. M., Cordeiro, J. L. P., Santos, T. S. J., Morais, A. A., Mangini, P. R., Rodrigues, L. F., Almeida, L. B., 2012. Avaliação do Risco de Extinção da Anta brasileira Tapirus terrestris Linnaeus, 1758, no Brasil. Biodiversidade Brasileira, Ano II, 3: 103–116. Ministério do Meio Ambiente, 2000. Plano de manejo da Reserva Biológica Córrego do Veado. MMA/

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IBAMA, Brasília. Naveda, A., Thoisy, B. de, Richard–Hansen, C., Torres, D. A., Salas, L., Wallace, R., Chalukian, S., Bustos, S. de., 2008. Tapirus terrestris. The IUCN Red List of Threatened Species 2008: e2456. Paschoal, A. M. O., Massara, R. L., Santos, J. L., Chiarello, A. G., 2012. Is the domestic dog becoming an abundant species in the Atlantic forest? A study case in southeastern Brazil. Mammalia, 76: 67–76. Paschoal, A. M. O., Massara, R. L., Bailey, L. L., Kendall, W. L., Doherty Jr., P. F., Hirsch, A., Chiarello, A. G., Paglia, A. P., 2016. Use of Atlantic Forest protected areas by free–ranging dogs: estimating abundance and persistence of use. Ecosphere, 7(10): e01480. Roy, H. E., Peyton, J., Aldridge, D. C., Bantock, T., Blackburn, T. M., Britton, R., Clark, P., Cook, E., Dehnen–Schmutz, K., Dines, T., Dobson, M., Edwards, F., Harrower, C., Harvey, M. C., Minchin, D., Noble, D. G., Parrott, D., Pocock, M. J. O., Preston, C. D., Roy, S., Salisbury, A., Schönrogge, K., Sewell, J., Shaw, R. H., Stebbing, P., Stewart, A. J. A., Walker, K. J., 2014. Horizon scanning for invasive alien species with the potential to threaten biodiversity in Great Britain. Global Change Biology, 20: 3859–3871. Silva–Rodríguez, E. A., Sieving, K. E., 2012. Domestic dogs shape the landscape–scale distribution of a threatened forest ungulate. Biological Conservation, 150(1): 103–110. Tobler, M. W., Janovec, J. P., Cornejo, F., 2010. Frugivory and Seed dispersal by the Lowland Tapir Tapirus terrestris in the Peruvian Amazon. Biotropica, 42(2): 215–222. Young, J. K., Olson, K. A., Reading, R. P., Amgalanbaatar, S., Berger, J., 2011. Is wildlife going to the dogs? Impacts of feral and free–roaming dogs on wildlife populations. Bioscience, 61: 125–132. Zapata–Ríos, G., Branch, L. C., 2016. Altered activity patterns and reduced abundance of native mammals in sites with feral dogs in the high Andes. Biological Conservation, 193: 9–16.

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315 Brief communication

Using non–invasive genetic techniques to assist in maned wolf conservation in a remnant fragment of the Brazilian Cerrado N. Mannise, R. G. Trovati, J. M. B. Duarte, J. E. Maldonado, S. González

Mannise, N., Trovati, R. G., Duarte, J. M. B., Maldonado, J. E., González, S., 2018. Using non–invasive genetic techniques to assist in maned wolf conservation in a remnant fragment of the Brazilian Cerrado. Animal Biodiversity and Conservation, 41.2: 315–319. Abstract Using non–invasive genetic techniques to assist in maned wolf conservation in a remnant fragment of the Brazilian Cerrado. The maned wolf is a South American canid considered a keystone species of the Cerrado. We performed a genetic assessment of maned wolves that inhabit a small remnant fragment of the Cerrado in Brazil. We collected 84 fecal samples over a year and also included two tissue samples from road–killed animals. We successfully identified the species, sex, and individuals using molecular markers. Using microsatellite loci analysis we identified 13 different individuals, eight females and five males. The genetic variability level found and the high number of individuals detected indicates the presence of an open population. Key words: Fecal DNA, Microsatellite loci, ZFX–ZFY, Real time PCR, Neotropical canid Resumen Utilización de técnicas genéticas no invasivas para contribuir a la conservación del aguará guazú en un fragmento residual del Cerrado de Brasil. El aguará guazú es una especie de cánido sudamericano que se considera clave en el Cerrado. Se realizó un estudio genético de varios individuos de aguará guazú que habitan en un pequeño fragmento del Cerrado de Brasil. Se colectaron 84 muestras fecales durante un año y se incluyeron también dos muestras de tejidos de animales atropellados. Se determinaron la especie, el sexo y los individuos mediante marcadores moleculares. Asimismo, se identificaron 13 individuos mediante la amplificación de loci de microsatélites, de los cuales ocho eran hembras y cinco, machos. El grado de variabilidad genética observado y el elevado número de individuos detectados indican la presencia de una población abierta. Palabras clave: ADN fecal, Loci de microsatélites, ZFX–ZFY, PCR en tiempo real, Cánidos neotropicales Received: 9 VII 17; Conditional acceptance: 17 X 17; Final acceptance: 20 XII 17 Natalia Mannise, Susana González, Depto. de Biodiversidad y Genética, Inst. de Investigaciones Biológicas Clemente Estable–MEC, Av. Italia 3318, Montevideo 11600, Uruguay.– Guilherme R Trovati, Centro de Energia Nuclear na Agricultura, Escola Superior de Agricultura 'Luiz de Queiroz'–Univ. de São Paulo, Av. Centenário 303, 13400–970 Piracicaba, SP–Brasil.– J. Mauricio B. Duarte, Núcleo de Pesquisa e Conservação de Cervídeos, Depto. de Zootecnia, Univ. Estadual Paulista Via de Acesso Paulo Donato Castellane s/n., 14884–900 Jaboticabal, SP–Brasil.– Jesús E. Maldonado, Center for Conservation Genomics, Smithsonian Conservation Biology Inst., Smithsonian Institution, 3001 Connecticut Ave, NW, Washington D.C. 20008, USA.– Susana González, Sección Genética Evolutiva, Fac. de Ciencias–UdelaR, Iguá 4225, Montevideo 11400, Uruguay. Corresponding author: Natalia Mannise. E–mail: natymanni@gmail.com

ISSN: 1578–665 X eISSN: 2014–928 X

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Introduction The maned wolf is the keystone canid species of the Brazilian Cerrado, the largest and richest tropical savanna in the world, and one of the most threatened biomes in South America (Klink and Machado, 2005). The IUCN listed it as Near Threatened because of the drastic reduction of suitable habitats for their survival. In Brazil it is categorized as Endangered (De Paula et al., 2008). The Cerrado is considered 'the central range of this species', and has lost approximately 50 % of its native vegetation (Klink and Machado, 2005). In this scenario, fragmented and isolated protected areas are too small for their long–term viability (Do Passo Ramalho et al., 2014; Lion et al., 2011). It is therefore critical to understand the amount of movement in and out of these areas. However, only two studies of maned wolves have been carried out to monitor population size and genetic diversity at a local scale in these protected areas (Do Passo Ramalho et al., 2014; Lion et al., 2011). Do Passo Ramalho et al. (2014) analyzed a maned wolf population in Jataí Ecological Station (São Paulo). The geographic distance between Jataí and our study area, the Estação Ecológica de Itirapina (EEI), is about 180 km. Habitat loss and fragmentation could lead to a rapid and dramatic decline if maned wolves are unable to disperse from one Cerrado patch to another. Our aim was to assess the population genetic diversity of maned wolves in the EEI, an area that contains a remnant fragment of the Cerrado biome in São Paulo (Brazil) near other protected areas where maned wolf population status have previously been assessed. Our results provide valuable information to evaluate the population dynamics of maned wolves at a fine, local scale. Material and methods The EEI (22º 11'–22º 15' S and 47º 51'–48º 00' W, São Paulo–Brazil) is inside an array of cultivated areas (fig. 1). We conducted monthly, six–day sampling surveys using the linear transect method. Each transect was 3 m wide and 200 m long. From March 2007 to February 2008, we collected 84 feces samples. Samples were stored in 50 ml sterile plastic tubes with 100 % ethanol and deposited in a freezer at –20 ºC until DNA extraction. Two tissue samples were collected from road–killed animals of opposite sexes. DNA extractions and PCR amplifications were conducted under sterile conditions with negative controls in Núcleo de Pesquisa e Conservação de Cervídeos (UNESP, Brazil). We used DNA Stool Mini Kit (Qiagen Inc. Valencia–California), following the manufacturer's instructions. DNA extractions from tissues were conducted following González et al. (2015a) procedures. For species identification, we amplified a species– specific D–loop fragment of mitochondrial DNA (mtDNA) using the protocols of González et al. (2015b). The PCR products were purified using Zymo Research DNA Clean and ConcentratorTM and sequenced in Ma-

crogen and Institute Pasteur–Montevideo. Sequences were compared with the GenBank database using BLAST. We used MEGA 5 software to construct the alignment by Clustal X (Tamura et al., 2011). Sex was determined through Real Time–PCR (RT–PCR) amplification of a fragment of the ZFX– ZFY genes, and High Resolution Melting Analysis (HRMA) (González et al., 2015a). Experiments were conducted in duplicate in a Rotor Gene 6000® (Corbett Research) (software version 1.7, Qiagen, UK). We used twelve microsatellite loci that were previously established to reliably amplify in maned wolves and had sufficient levels of polymorphism and adequate non–exclusion probabilities of identity to assign parent pairs (Mannise et al., 2017). Multiplex PCR reactions and thermal profile were conducted as suggested by Mannise et al. (2017). PCR products were run using the LIZ 500 size standard on an ABI PRISM 3100 sequencer. We analyzed fragment size and genotypes in GENMARKER V1.75 (SoftGenetics). PCRs were repeated three and two times for homozygotes and heterozygotes, respectively. We identified capture histories and the number of maned wolves resampled from multiple scats using GENECAP software (Wilberg and Dreher, 2004). We detected matching genotypes with a probability of 0.01, assuming that individuals could be siblings (Wilberg and Dreher, 2004). The probability of identity (P(id)) was calculated for Hardy–Weinberg equilibrium (HWE) and for sibling presence. Genotyping errors were estimated with MICROCHECKER (Van Oosterhout et al., 2004); genotypes were corrected following Brooksfield´s (1996) estimation if genotyping errors were detected. We used GENEPOP (Raymond and Rousset 1995) to compute HWE, linkage disequilibrium (LD) and inbreeding coefficients (Fis). In case of disequilibrium, Bonferroni corrections were applied. We calculated number of alleles, expected and observed heterozygosity and allele frequencies using CERVUS (Marshall, 1998–2007). We applied the ML–RELATE software to analyze the relatedness coefficients and pedigree relationships among individuals (Kalinowski et al., 2006). Results We identified 58 fecal samples as being from maned wolves and we determined the presence of three D– loop haplotypes previously described in other maned wolf populations from Brazil (table 1) (González et al., 2015b). Of the twelve microsatellite loci tested, one (AHTK253) showed low amplification success. Samples with at least six reliable genotypes after replicates were included in further analyses. We determined the presence of 13 different individuals sampled with P(id) = 3.57 x 10–11 for HWE and 1.18 x 10–4 for siblings (table 1, fig. 1). P(id) values were acceptable for accurate individual identification (< 0.05) (Woods et al., 1999). We assigned eight females and five males based on the melting curve pattern (table 1). The mean melting curve was 82.95 ºC (SD = 0.22) for ZFX, and

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LG1 LG10 LG12 LG13 LG17 LG20 LG21 LG22 LG23 LG27 LG6 LG8 LG9

1 mi

Fig. 1. Samples collected in the EEI. Each symbol represents a different individual; same shape represents re–sampled individuals. Road–killed animals and its re–sampled feces are shown by balloons. Fig. 1. Muestras colectadas en la estación ecológica de Itidapina (EEI). Cada símbolo representa distintos individuos; la misma forma representa individuos que se han muestreado más de una vez. Los animales atropellados y sus muestras fecales muestreadas más de una vez se indican con globos.

82.08 ºC (SD = 0.17) for ZFY. The observed sex ratio was not significantly different from the expected 1:1 (x2 = 0.69, p = 0.41). The presence of null alleles in FH2561 has been previously detected in other maned wolf populations, for this reason we recommend that this locus should not be used in future studies. We identified a minimum of three family groups (table 1). One comprised a mother with her son, two aunts and an uncle (table 1). Another included two full–siblings (brother and sister), and the third had two half–sisters (table 1). Discussion Our non–invasive genetic protocols allowed us to reliably identify species, gender, and individuals. We successfully amplified fecal DNA using microsatellite multiplex protocols to effectively genotype 67 % of the samples. Maned wolves in the EEI had high levels of polymorphism and similar levels of observed and expected heterozygosity (table 2). We found that maned wolves in the EEI have moderate levels of mitochondrial genetic diversity and high levels of nuclear genetic diversity. Mean levels of heterozygosity (0.646) were slightly lower than

other population heterozygosity estimates of maned wolf populations in Brazil (0.71 and 0.72) (Do Passo Ramalho et al., 2014; Lion et al., 2011). The presence of three family groups in the EEI area during our surveys suggests that if dispersal in and out of this area were restricted, it would result in inbreeding and loss of genetic variability. In strong support of a recent genetic analysis conducted in maned wolves throughout their range (Mannise et al., 2017), our results here suggest that wolves maintain high levels of genetic diversity and may avoid inbreeding by dispersing through cultivated lands. Furthermore, these high levels of genetic variability coupled with the large number of individuals detected indicate the presence of an open population in the EEI. Effective conservation management strategies can benefit from approximate estimates of population size and non–invasive genetic studies such as this one (Kohn et al., 1999). Our study provides valuable information regarding the genetic diversity of an imperiled canid species in a human–dominated landscape. It suggests that this small fragmented remnant of Cerrado can maintain several maned wolf family groups with adequate levels of genetic variability, and it highlights the essential need to protect the Cerrado habitat for their conservation. Future studies that include non–invasive

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Table 1. For each individual we show: ID, identification number; ST, sample type; D–lH, D–loop haplotype; Genbank, Genbank accession number; sex; N, number of recaptures; and family group (PO, parent–offspring; FS, full–sibling; HS, half–sibling). Tabla 1. Para cada individuo se muestra: ID, el número de identificación; ST, el tipo de muestra; D–IH, el haplotipo para D–loop; GenBank, el número de accesión del Genbank; el sexo; N, el número de recapturas; y el parentesco (PO, padre–hijos; FS, hermanos; HS, medio hermanos).

ID ST D–lH

Genbank

Sex

Family group

LG1

Feces

KM406503

Female

PO: LG10, HS: LG8 – LG21 – LG22

LG6

Feces

KM406508

Female

HS: LG12

LG8

Feces

KM406503

Female

HS: LG1 – LG21 – LG22

LG9

Feces

KM406503

Female

–

LG10

Feces

KM406503

Male

PO: LG1, HS: LG8 – LG21 – LG22

LG12

Skin

KM406508

Female

HS: LG6

LG13 Skin G KM406508 Male 2 – LG17

Feces

KM406503

Female

–

Lg20 LG21

Feces

KM406503

Male

–

Feces

KM406505

Male

FS: LG22, HS: LG1 – LG8

LG22

Feces

KM406505

Female

FS: LG21, HS: LG1 – LG8

LG23

Feces

KM406503

Female

FS: LG27

LG27

Feces

KM406503

Male

FS: LG23

Table 2. For the overall dataset and for each locus: k, number of alleles; Ta, annealing temperature; Ho, observed heterozygosity; He, expected heterozygosity; Pic, polymorphic information content; Fis, inbreeding coefficient; Ge, presence/absence of genotyping errors; * loci that after Bonferroni corrections slightly deviate from HWE. Tabla 2. Para el conjunto de datos y para cada locus se muestra: k, número de alelos; Ta, temperatura de unión al cebador; Ho, heterocigosidad observada; He, heterocigosidad esperada; Pic, contenido de información polimórfica; Fis, coeficiente de endogamia; Ge, presencia/ausencia de errores de genotipado; * loci que después de las correcciones de Bonferroni se desvían ligeramente del equilibrio Hardy–Weinberg (HWE).

Locus

Ta (ºC)

Pic

Fis

FH2140*

0.883

0.831

–0.139

FH2328*

0.556

0.928

0.864

0.416

Null alleles

FH2137*

0.636

0.788

0.713

0.2

FH2535

58 0.846 0.748 0.687 –0.138

FH2848

58 0.833 0.808 0.741 –0.033

REN105

2 58 0.25 0.489 0.359 0.5

PEZ19

58 0.444 0.386 0.327 –0.164

FH2561*

FH2226

9 58 0.846 0.886 0.836 0.047

FH2054*

REN169

3 60 0.167 0.163 0.15 –0.023

Overall

0.583 0.25

0.899 0.757

0.836 0.683

0.361 0.679

Null alleles No

Null alleles

6.27 – 0.589 0.646 0.583 0.168

No –

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surveys outside these protected areas will be crucial to evaluate movement and corridor connectivity within this fragmented landscape. Acknowledgements Permits to collect samples were granted by the Instituto Brasileiro do Meio Ambiente e dos Recursos Naturais Renováveis (IBAMA 036/2007–CGFAU) and the Instituto Florestal de São Paulo with the research license for the project entitled 'Mammals burrowing (Dasypodidae and Echimyidae) Cerrado of Itirapina region and its role in terrestrial vertebrate communities'. This research was funded by the Agencia Nacional de Investigación e Innovación (ANII Fondo Clemente Estable_3_2011_1_6619) from Uruguay, and the Programa de Desarrollo de las Ciencias Básicas (PEDECIBA–Uruguay). We thank Mariana Cosse and Aline Meira Bonfim Mantellato for assistance with lab techniques and data analysis. References De Paula, R. C., Medici, P., Morato, R. G., 2008. Maned Wolf Action Plan. Icmbio, Brasilia. Do Passo Ramalho, F., Miotto, R. A., Martins, N., Galetti, P. M., 2014. Maned wolf (Chrysocyon brachyurus) minimum population size and genetic diversity in a Cerrado protected area of southeastern Brazil revealed by fecal DNA analysis. Mammalia, 78(4): 465–472. González, S., Cosse, M., Franco, M. R., Emmons, L., Vynne, C., Duarte, J. M. B, Beccacesi, M. D, Maldonado, J. E., 2015b. Population Structure of mtDNA Variation due to Pleistocene Fluctuations in the South American Maned Wolf (Chrysocyon brachyurus, Illiger, 1815): Management Units for Conservation. Journal of Heredity, 106(S1): 459–468. González, S., Mannise, N., Repetto, L., Maldonado, J. E., 2015a. Sex determination of three Neotropical canids by high resolution melting analysis. Conservation Genetics Resources, 7(3): 643– 645. Kalinowski, S. T., Wagner, A. P., Taper, M. L., 2006. ML–Relate: a computer program for maximum like-

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lihood estimation of relatedness and relationship. Molecular Ecology Notes, 6(2): 576–579. Klink, C. A., Machado, R. B., 2005. Conservation of the Brazilian Cerrado. Conservation Biology, 19(3): 707–713. Kohn, M. H., York, E. C, Kamradt, D. A, Haught, G., Sauvajot, R. M, Wayne, R. K., 1999. Estimating population size by genotyping faeces. Proceedings of the Royal Society of London B: Biological Sciences, 266(1420): 657–663. Lion, M. B., Eizirik, E., Garda, A. A., Fontoura– Rodrigues, M. L., Guimaraes Rodriguez, F. H., Marinho– Filho, J. S., 2011. Conservation genetics of maned wolves in a highly impacted area of the Brazilian Cerrado biome. Genetica, 139(3): 369–381. Mannise, N., Cosse, M., González, S., Emmons, L. H., Duarte, J. M. B., Beccaceci, M. D., Maldonado, J. E., 2017. Maned wolves retain moderate levels of genetic diversity and gene flow despite drastic habitat fragmentation. Endangered Species Research, doi: 10.3354/esr00859 Marshall, T., 1998–2007. Cervus 3.0 Help. Field Genetics Ltd., Copyright Tristan Marshall. Raymond, M., Rousset, F., 1995. GENEPOP (Version 1.2): Population Genetics Software for Exact Tests and Ecumenicism. Journal of Heredity, 86(3): 248–249. Tamura, K., Peterson, D., Peterson, N., Stecher, G, Nei, M., Kumar, S., 2011. MEGA5: Molecular Evolutionary Genetics Analysis Using Maximum Likelihood, Evolutionary Distance, and Maximum Parsimony Methods. Molecular Biology and Evolution, 28(10): 2731–2739. Van Oosterhout, C., Hutchinson, W., Wills, D. P. M., Shipley, P., 2004. Micro–checker: software for identifying and correcting genotyping errors in microsatellite data. Molecular Ecology Notes, 4(3): 535–538. Wilberg, M. J., Dreher, B. P., 2004. Genecap: a program for analysis of multilocus genotype data for non–invasive sampling and capture–recapture population estimation. Molecular Ecology Notes, 4(4): 783–785. Woods, J. G., Paetkau, D., Lewis, D., McLellan, B. N., Proctor, M., Strobeck, C., 1999. Genetic tagging of free–ranging lack and brown bears. Wildlife Society Bulletin, 27: 616–627.

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The role of semi–natural grasslands and livestock in sustaining dung beetle communities (Coleoptera, Scarabaeoidea) in sub–Mediterranean areas of Slovenia J. Jugovic, N. Koprivnikar, T. Koren

Jugovic, J., Koprivnikar, N., Koren, T., 2018. The role of semi–natural grasslands and livestock in sustaining dung beetle communities (Coleoptera, Scarabaeoidea) in sub–Mediterranean areas of Slovenia. Animal Biodiversity and Conservation, 41.2: 321–332. Abstract The role of semi–natural grasslands and livestock in sustaining dung beetle communities (Coleoptera, Scarabaeoidea) in sub–Mediterranean areas of Slovenia. We studied the richness and structure of the coprophagous Scarabaeoidea community in two pastures (Hrastovlje and Zazid) in sub–Mediterranean Slovenia. In each pasture, we examined three habitat patches characterised by different levels of grazing (S1, the active part of the pasture; S2, the overgrown part of the pasture, mainly spiny shrubs; S3, a meadow with some overgrown patches of shrubs outside the fenced pasture). The main results were as follows: (1) 29 species were sampled, corresponding to about three quarters of the species presumably present at the two study sites; (2) species richness and abundance in Zazid are were similar in all three patches; (3) the species richness and abundance in Hrastovlje (in total, and separately for dwellers and tunnelers) were highest in S2. In Hrastovlje, dwellers were most abundant in S1. As the two different habitat patches were shown to positively influence the dung beetle community, we recommend maintaining a traditionally–managed mosaic landscape. Key words: Karst pasture/meadow, Species biodiversity, Microhabitat, Bait, Pitfall traps Resumen La importancia de los pastizales seminaturales y la ganadería en el mantenimiento de las comunidades de coleópteros coprófagos (Coleoptera, Scarabaeidae) en las zonas submediterráneas de Eslovenia. Estudiamos la riqueza y la estructura de la comunidad de escarabeoideos coprófagos en dos pastizales (Hrastovlje y Zazid) en la zona submediterránea de Eslovenia. En cada pastizal, analizamos tres fragmentos caracterizados por diferentes grados de pastoreo (S1, la zona activa de pastoreo; S2, la zona de crecimiento del pasto, principalmente arbustos espinosos; y S3, una pradera con algunos fragmentos arbustivos con crecimiento fuera del pastizal vallado). Los resultados principales fueron los siguientes: (1) se muestrearon 29 especies que correspondían aproximadamente a tres cuartas partes de las especies previsiblemente presentes en los dos sitios de estudio; (2) la riqueza y la abundancia de especies en Zazid fueron parecidas en los tres fragmentos; y (3) la riqueza y la abundancia de especies en Hrastovlje (en total y los residentes y los cavadores por separado) fueron más elevadas en S2. En Hrastovlje, los residentes fueron más abundantes en S1. Como se constató que ambos fragmentos de hábitat influían positivamente en la comunidad de coleópteros coprófagos, recomendamos mantener un territorio en mosaico gestionado de forma tradicional. Palabras clave: Pasto y pradera cársticos, Biodiversidad de especies, Microhábitat, Cebo, Trampas de caída Received: 24 II 17; Conditional acceptance: 24 V 17; Final acceptance: 20 XII 17 Jure Jugovic, Nataša Koprivnikar, Dept. of Biodiversity, Fac. of Mathematics, Natural Sciences and Information Technologies, Univ. of Primorska, Glagoljaška 8, 6000 Koper, Slovenia.– Toni Koren, Association Hyla, Lipovac I no. 7, 10000 Zagreb, Croatia. Corresponding author: Jure Jogovic. E–mail: jure.jugovic@upr.si, jure.jugovic@gmail.com ISSN: 1578–665 X eISSN: 2014–928 X

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Introduction Within the topographically diverse Mediterranean Basin (S Europe) lying at the intersection of Europe, Asia and Africa (cf. Blondel, 2006; Blondel et al., 2010; Christodolou et al., 2016), grasslands (e.g. meadows, pastures) and shrublands support exceptionally high biodiversity (Lumaret and Kirk, 1991; Verdú et al., 2000; Allen, 2003). Karst meadows belonging to the class Festuco–Brometea, for example, are regarded as species–rich habitats of national and European importance, and are among the most species–rich environments within the semi–natural habitat types (Jugovic et al., 2013a). Two main associations can be found there: the association Carici–Centaureetum in pastures and the association Dianthonio–Scorzeneretum in meadows (Kaligarič, 2005). These grasslands are the result of past human activities (Kaligarič, 2005; Stergaršek, 2009) that strongly influenced the layout and biodiversity of the landscape. Many turned into overgrown areas (shrublands and later pioneer forests: Jugovic et al., 2013a) after traditional extensive agricultural practices were abandoned (Zeiler, 2000; Stefanescu et al., 2004). In the Mediterranean, the maintenance of open (meadows, pastures) and semi–open (shrubland) habitats ceased with the abandonment of traditional agricultural practices (extensive grazing and occasional extensive mowing: see Jogan et al., 2004; Kaligarič, 2005). The consequent fragmentation of grasslands decreased the size of habitat patches and their connectivity (Polus et al., 2007). However, a dense network of suitable habitats is crucial to maintain metapopulations and enable dispersal of animals inhabiting the remaining open spaces (cf. Anthes et al., 2003; Bergman and Landin, 2001; Mousson et al., 1999; Polus et al., 2007; Thomas et al., 1992). Grazing activity increases the extent of open areas (Lumaret, 1994), but it also strongly modifies the vegetation structure. Hence, grazing is often seen to have a negative impact on biodiversity (e.g. Jugovic et al., 2013a, 2014a, 2017), but some invertebrates, such as dung beetles (Coleoptera: Scarabaeoidea), rely on the presence of livestock, depending on their feces as a main source of nutrients (Verdú and Galante, 2004; Zamora et al., 2007). Such ephemeral habitats with prominent temporal instability (Finn, 2001) are subjected to succession that is short– termed (days or weeks, rarely months), and the organisms specialised in their exploitation compete with one another (e.g. Metazoa, Fungi; Sladecek et al., 2017). Different animal taxa can occupy different spatial compartments: while the community inhabiting the surface or outer rim of the excrement consist mainly of adult dipterans (Diptera), the internal community consists mainly of beetle (Coleoptera) adults and dipteran larvae (Sladecek et al., 2017; Mohr, 1943). Temporal segregation and use of the excrement has already been extensively studied, showing that copro– and necrophilous species are segregated along the successional gradient by their oviposition preferences (see Sladecek et al., 2017 for a review). Although not as highly specialised as co-

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prophagous Scarabaeoidea, some other invertebrate taxa can also be consistently attracted to vertebrate excrement, either because it is a food source (coprophagous invertebrates such as dipterous larvae and earthworms) or because it contains their (e.g. Chilopoda, Scorpiones) prey species (Curry, 1994). Coprophagous dung beetles (Scarabaeoidea) are an invertebrate group that is attracted to (predominantly) fresh herbivore and omnivore excrement; and many species have developed complex nesting behaviour that includes the use of dung as a food supply for their offspring (Cambefort and Hanski, 1991). Coprophagous dung beetles are a part of the diverse superfamily Scarabaeoidea, and members of three families, Aphodiidae, Geotrupidae and Scarabaeidae (Halffter and Matthews, 1966), share a coprophagous life style. Dung beetles fall into three basic nesting categories, rollers (telecoprids), tunnelers (paracoprids) and dwellers (endocoprids; cf. Errouissi et al., 2009). Rollers make a ball from the excrement and use it as a food source or as a brooding chamber. Tunnelers bury the dung wherever they find it, while dwellers neither roll nor burrow, but stay and live in the dung (Scholtz et al., 2009). In a wide array of habitats in ecosystems, Coprophagous Scarabaeoidea play an important role in many processes, such as recycling of animal excreta, nutrient cycling, bioturbation, pollination, seed dispersal and primary production, and parasite suppression (Losey and Vaughan, 2006; Nichols et al., 2008; Hanski and Camberfort, 1991; Lobo et al., 2004). Through these roles, they help establish and sustain other ground living invertebrate communities. Coprophagous Scarabaeoidea may be used as bioindicators (cf. Favila and Halffter, 1997; Lobo et al., 2002; Verdú et al., 2000) as they are susceptible to slight changes in their environment and compositionally respond to local changes (Nichols et al., 2008). They are therefore useful to detect small differences in habitat changes on a local scale. Moreover, significant differences in species presence, abundance and diversity indices can help conservation practitioners assess the relative importance of natural, semi–natural and highly transformed habitats (Davis et al., 2004). The transformation in habitat considered to be mainly responsible for the decrease in the dung beetles’ species richness and abundance is the abandonment of pasturelands. Nevertheless, a variety of impacts on the dung beetle community during the succession that follows grazing abandonment have been proposed (e.g. Macagno and Palestrini, 2009; Negro et al., 2011, Verdú et al, 2000, Tonelli et al., 2017a). In an attempt to highlight the differences in diversity of a dung beetle community at a local scale, we (i) present data on the species community of dung beetles from two study sites located in sub–Mediterranean dry grasslands of SW Slovenia, and (ii) test for possible differences in dung beetle species richness and abundance between three habitat patches within each site differing in their levels of grazing (S1, S2, S3: see section on study sites). We aimed to determine what level of grazing impact can support the highest dung beetle diversity.

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Hungary

Austria Italy

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Hrastovlje (H)

Slovenia Croatia

S2 Gulf of Trieste

Trieste

Koper

S2 H

5 km

Zazid (Z)

500 m

Fig. 1. Geographic position (inset) and habitat patches (S1, grazed part of the pasture; S2, overgrowth part of the pasture; S3, dry karst meadow) at two study sites (Hrastovlje and Zazid) in SW Slovenia. Fig. 1. Posición geográfica (recuadro) y fragmentos de hábitat (S1, zona activa de pastoreo; S2, zona de crecimiento del pasto; S3, pradera cárstica seca) en los dos sitios del estudio (Hrastovlje y Zazid), en el sudeste de Eslovenia.

Material and methods Study sites The study sites selected were two pastures and their surroundings with a total size of 22.5 ha, located in south–western Slovenia (45º 30' 40.55'' N, 13º 54' 47.55'' E). Study sites were between 90 and 450 m above sea level. Both sites lie on the predominately carbonate bedrock near the villages of Hrastovlje (study site H) and Zazid (study site Z) and are 4.3 km apart (measured centre to centre). Each study site consisted of three habitat patches that represent different levels of grazing impact (fig. 1). Within the fenced pasture, two habitat patches were present: (1) a grazed part of the pasture (stage S1: goats in H, and cattle in Z), and (2) an overgrown part of the pasture, mainly with some spiny shrubs, as a result of selective grazing (stage S2: in Z, Prunus spinosa Linnaeus, Prunus mahaleb Linnaeus, Crataegus monogyna Jacquin, Cotinus coggygria Scopoli, and in H: Paliurus spina–christi Miller; cf. Jugovic et al., 2013a, 2017; field observations). The shrubs were relatively evenly distributed across the patch and close to each other. Outside the fenced pasture, (3) abandoned and partly overgrown dry karst meadow ('islands' of predominantly C. coggygria or C. monogyna) grazed in the past made up the third patch (stage S3). Sampling design Sampling took place between 12 March 2012 and 7 November 2012. We used pitfall traps to capture

dung beetles at each habitat patch in the two study sites. Traps consisted of 500 ml plastic jars with a depth of 13 cm and a diameter of 10 cm. Traps were evenly distributed across the habitat patches, at least 50–100 m away from each other (Larsen and Forsyth, 2005; Silva and Hernández, 2015). The only exception was the smallest patches (S3 at both sites), where any given pair of two nearest traps could be closer, but still at least 20 m apart. Even then, such a pair consisted of one baited and one unbaited (control) trap. Four sampling traps were used per habitat type × study site, i.e.; 24 traps in total. On every occasion, two traps were used as a control (without bait). In the other two, we added a ball of fresh cow excrement (cca. 4–5 cm3) as bait to additionally attract the animals. The use of cow excrement for bait has shown to be highly effective for sampling Scarabaeoidea in Mediterranean regions (e.g. Martin–Píera and Lobo, 1996; Barbero et al., 1999; Dormont et al., 2007, 2010; Errouissi et al., 2004). The bait was wrapped in gauze and attached to a wire over the trap opening. Although some authors recommend larger amounts of bait (e.g. Errouissi et al., 2004), we found the baited traps to be highly effective in comparison with the control traps. Animals were collected 14 days after placement of traps. Traps were emptied a total of 15 times. Propylene glycole was used as a fixative. Laboratory work Dung beetles were separated from other trapped ground invertebrates. Each specimen was identified to species level (following Ballerio et al., 2010:

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45 40

Total (95 % CI)

Taxa

S1 S2

S1 S3 S2

15 10 0 0

Hrastovlje Zazid

0 5 10 15 20 25 30 35 40 45 0 10 20 30 40 50 60 70 80 90 100110 120 130 140 150 Sampling units Sampling units

Fig. 2. A, species accumulation curves (dotted) for Scarabaeoidea from Hrastovlje (grey) and Zazid (black) with 95 % confidence interval (outer lines). The accumulation curve is extrapolated by the factor 3. B, species accumulation curves for Scarabaeoidea from three habitat patches (S1, S2, S3) from Hrastovlje (grey) and Zazid (black). The accumulation curve is extrapolated by the factor 10. Each sampling unit represents: A, twelve traps per locality; B, four traps per habitat patch. Traps were emptied 15 times, at 14–day intervals. Fig. 2. A, curvas de acumulación de especies (punteadas) de la familia Scarabaeidae de Hrastovlje (gris) y Zazid (negro) con un intervalo de confianza del 95 % (líneas externas). La curva de acumulación se ha extrapolado por el factor 3. B, curvas de acumulación de especies de la familia Scarabaeidae de los tres fragmentos de hábitat (S1, S2 y S3) de Hrastovlje (gris) y Zazid (negro). La curva de acumulación se ha extrapolado por el factor 10. Cada unidad de muestreo representa: A, doce trampas por localidad; B, cuatru trampas por fragmento de hábitat, respectivamente. Las trampas se vaciaron 15 veces, a intervalos de 14 días.

when not identifiable by external morphology, genital structures were also inspected). Each animal was labelled and stored in the collection of the Department of Biodiversity, Faculty of Mathematics, Natural Sciences and Information Technologies, University of Primorska, for further studies. Numbers of specimens and taxa were counted per trap. For dung beetles, numbers of specimens and species were counted in total, and further subdivided into four groups from three guilds (cf. Jay–Robert et al., 2008a; Errouissi et al., 2009: Aphodiidae–dwellers, Scarabaeidae–tunnelers, Geotrupidae–tunnelers and Scarabaeidae–rollers. Data analysis We constructed species accumulation curves with 95 % confidence intervals using the program EstimateS 9.1.0 (purl.oclc.org/estimates), separately (i) for each of the two study sites and (ii) for each of the three habitats (separately for the two study sites) to estimate inventory completeness. A single sampling unit (SU) consisted of the specimens collected per (i) locality and (ii) habitat patch over each 14–day interval (i.e. number of sampling occasions (= 15) correspond to number of SUs). At the locality level, each SU consisted of 12 pitfall traps, and at the habitat patch level, each SU consisted of 4 pitfall traps. Table 1 shows the complete list of collected species and individuals.

Diversity of taxa is represented by the number of species. The Shannon index (H’, here with ln) was calculated for each study site and study site × habitat patch. Mean Shannon values were then calculated using the data of the 24 traps. For these calculations, we used program PAST (Palaeontological Statistics: Hammer, 1999–2012). The mean Shannon index values between the two study sites (t–test, significance accepted at p < 0.05) and between the three habitat patches within each study site (ANOVA, at p < 0.05) were compared using the SPSS Statistical package ver. 20.0 (SPSS inc., 1989, 2011). Due to the deviations of data from the normal distribution (Kolmogorov–Smirnov test, p < 0.05), non–parametric statistical tests executed within SPSS Statistical package ver. 20.0 were used to test the following hypotheses. To test for possible differences in the number of species and specimens between the two study sites (data for all pitfall traps combined) we implemented a Mann–Whitney U–test of ranks, and accepted the difference to be significant at p < 0.05. We also tested for possible differences in abundance of species and specimens between the three habitat patches using a Kruskal–Wallis test of ranks (p < 0.05) and post–hoc pairwise comparisons with the Dunn–Bonferroni post–hoc method. Finally, to test for differences in species richness and abundance between the three patches, we treated the data at each study site separately in order to

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Table 1. A list of Scarabaeoidea species collected in pitfall traps in three habitat patches (S1, S2, S3) at two study sites (Hrastovlje and Zazid) in SW Slovenia in 2012. Diversity indices (No. of species, H’) are shown at the bottom of the table for the two study sites, and three habitat patches within each of them. Tabla 1. Lista de especies de la familia Scarabaeidae recogidas en trampas de caída en tres fragmentos de hábitat (S1, S2 y S3) en los dos sitios del estudio (Zazid y Hrastovlje), situados en el sudeste de Eslovenia en 2012. Al final de la tabla se muestran los índices de diversidad (nº de especies y H') de los dos sitios del estudio y los tres fragmentos de hábitat en cada uno de ellos. Family Hrastovlje Zazid Species S1 S2 S3 S1 S2 S3 Aphodiidae–dwellers Oxyomus sylvestris (Scopoli, 1763) 2 0 3 0 12 0 Sigorus porcus (Fabricius, 1792) 2 4 8 0 0 0 Loraphodius suarius (Faldermann, 1836) 0 0 0 0 0 1 Subrinus sturmi (Harold, 1870) 1 0 0 0 0 0 Aphodius fimetarius (Linnaeus, 1758) 0 0 12 0 0 0 Eurodalus paracoenosus (Balthasar & Hrubant, 1960) 0 0 1 0 0 1 Teuchestes fossor (Linnaeus, 1758) 0 0 0 0 0 1 Phalacronothus biguttatus (Germar, 1824) 0 0 0 0 0 1 Esymus pusillus (Herbst, 1789) 0 0 0 1 0 0 Colobopterus erraticus (Linnaeus, 1758) 0 0 0 2 0 0 Acrossus luridus (Fabricius, 1775) 0 0 0 2 0 0 Melinopterus prodromus (Brahm, 1790) 3 0 0 0 0 0 Geotrupidae–tunnelers Geotrupes spiniger Marsham, 1802 0 0 0 3 0 2 Trypocopris vernalis (Linnaeus, 1758) 68 201 46 54 134 147 Scarabaeidae–tunnelers Onthophagus grossepunctatus Reitter, 1905 51 82 96 54 59 165 Onthophagus verticicornis (Laicharting, 1781) 0 7 2 26 21 52 Onthophagus medius (Kugelann, 1792) 3 0 0 19 2 20 Onthophagus ruficapillus Brullé, 1832 4 2 1 0 1 0 Onthophagus coenobita (Herbst, 1783) 0 2 0 0 0 12 Onthophagus lemur (Fabricius, 1781) 0 0 0 11 0 0 Onthophagus ovatus (Linnaeus, 1767) 3 4 3 7 6 7 Onthophagus joannae Goljan, 1953 4 5 4 5 6 18 Onthophagus taurus (Schreber, 1759) 10 3 1 1 1 3 Onthophagus illyricus (Scopoli, 1763) 2 0 1 0 0 5 Onthophagus fracticornis (Preyssler, 1790) 0 2 2 7 3 15 Caccobius schreberi (Linnaeus, 1767) 20 0 2 0 0 0 Euoniticellus fulvus (Goeze, 1777) 3 1 0 0 2 0 Copris lunaris (Linnaeus, 1758) 2 0 0 0 0 0 Scarabaeidae–rollers Sisyphus schaefferi (Linnaeus, 1758) 35 364 148 403 459 968 No. of species (study site × habitat patch) 16 No. of species (study site) No. of individuals (study site × habitat patch) 232

12 15 14 21 677 330 595

12 16 23 706 1,418

No. of individuals (study site) 1,220 2,719 H' (study site × habitat patch) 1.91 1.18 1.51 1.25 1.13 1.16 H' (study site) 1.54 1.21

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Table 2. Comparisons of median values (Mann–Whitney U–test) in numbers of Scarabaeoidea species and specimens between the two study sites (H, Hrastovlje; Z, Zazid) for different datasets: All, whole dataset, and separately for dwellers, tunnelers and rollers; n/a, not applicable; 1 rollers were omitted from the analysis as only one species was present in the samples (Sisyphus schaefferi); 2 only two species of Geotrupidae–tunnelers were present in the samples, so the analyses were conducted on data pooled together for Scarabaeidae–tunnelers + Geotrupidae–tunnelers. Tabla 2. Comparación de los valores medianos (prueba U de Mann–Whitney) de la cantidad de especies e individuos de la familia Scarabaeidae en los dos sitios del estudio (H, Hrastovlje; Z, Zazid) para diferentes conjuntos de datos: All, todo el conjunto de datos; y por separado para dwellers (residentes), tunnelers (cavadores) y rollers (rodadores); n/a: no aplicable. 1 se omitió a los rodadores del análisis porque en las muestras solo se encontró una especie (Sisyphus schaefferi); 2 solo se encontraron dos especies de la familia Geotrupidae–cavadores en las muestras, de tal forma que los análisis se realizaron con los datos agrupados de Scarabaeidae–cavadores + Geotrupidae–cavadores.

All

Dwellers

Z p Compared groups

Tunellers2

Rollers

Z p Z p Z p

No. of species

H:Z

–0.839 0.401

–1.099 0.272 –0.899 0.369 n/a n/a

No. of individuals

H:Z

–0.58 0.562

–1.108 0.268

avoid the influence of geographic distance, altitudinal difference, and different livestock characteristics between the two sites. Results Species richness In total, 3,939 Scarabaeoidea specimens (Hrastovlje: 1,220; Zazid: 2,719) were collected. Of these, we identified 29 species of Scarabaeoidea belonging to all four groups from the three guilds (Aphodiidae–dwellers: 12 species, Geotrupidae–tunnelers: 2 species, Scarabaeidae–tunnelers: 14 species, Scarabaeidae–rollers: 1 species). This corresponds to ca. 25 % of known Slovenian species belonging to these three families (cf. Brelih et al., 2010). In Hrastovlje, we collected 21 species, i.e. 87 % of species we would collect at doubled sample effort, or 84 % of species at tripled effort. In Zazid, we collected 23 species, and the respective percentages were 79 % and 74 %. At higher sampling effort, the extrapolation line for Hrastovlje started to approach the maximum at 5–times greater effort (i.e. calculated maximum of 25 species: collected species represent 84 %), whereas for Zazid, the extrapolation line reached its maximum at 9–times larger effort (i.e. calculated maximum of 32 species: collected species represent 72 %). Both accumulation curves with their 95 % confidence intervals overlapped (fig. 2A). The results for the three habitat patches from Hrastovlje and Zazid (fig. 2B) were: 16 and 14 collected species in S1 (corresponding to 79 % and 55 % of expected maximum number of species [H: 20 species, Z: 26 species] at 7 and 19–times larger sampling effort), 12 species in S2 from each of the

–0.974 0.330 –1.58 0.114

two localities (corresponding to 59 % and 76 % of expected maximum number of species [H: 20, Z: 16] at 11– and 8–times larger sampling effort), and 15 and 16 collected species in S3 (corresponding to 54 % and 80 % of expected maximum number of species [H: 28, Z: 20] at 11– and 8–times larger sampling effort). There was an evident overlap among the 95 % confidence intervals for all six accumulation curves. In total, six species (Aphodiidae–dwellers: Aphodius fimetarius (Linnaeus, 1758), Aphodius prodromus (Brahn, 1790), Sigorus porcus (Fabricius, 1792), Subrinus sturmi (Harold, 1870); Scarabaeidae–tunellers: Copris lunaris (Linnaeus, 1758), Caccobius schreberi (Linnaeus, 1767)) were present exclusively at the study site in Hrastovlje, and eight species (Aphodiidae–dwellers: Acrossus luridus (Fabricius, 1775), Colobopterus erraticus (Linnaeus, 1758), Esymus pusillus (Herbst, 1789), Loraphodius suarius (Faldermann, 1836), Phalacronothus biguttatus (Germar, 1824), Teuchestes fossor (Linnaeus, 1758); Geotrupidae–tunellers: Geotrupes spiniger Marsham, 1802; Scarabaeidae–tunellers: Onthophagus lemur (Fabricius, 1781)) were found only in Zazid. The rest of the species were collected at both study sites (table 1). Diversity and factors influencing species richness and structure Diversity index values showed no statistically significant differences between Hrastovlje (H' = 1.54) and Zazid (H' = 1.21) (t–test, p > 0.05; table 1). The most abundant species, Sisyphus schaefferi (Linnaeus, 1758), represented 44.8 % and 67.3 % of the Scarabaeoidea community in Hrastovlje and Zazid, respectively. The Shannon index was highest

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Table 3. Comparison of median values of the number of species and specimens of Scarabaeoidea among the three habitat patches (S1, S2, S3) (KW, Kruskal–Wallis test) and post–hoc pairwise comparisons (PC, Dunn–Bonferronni) from two study sites (SS: H, Hrastovlje; Z, Zazid) in SW Slovenia: STS, standardised test statistic; n/a, not applicable; 1 rollers were omitted from the analysis as only one species was present in the samples (Sisyphus schaefferi); 2 only two species of Geotrupidae–tunnelers were present in the samples, so the analyses were conducted on data pooled together for Scarabaeidae–tunnelers + Geotrupidae–tunnelers. Tabla 3. Comparación de los valores medianos de la cantidad de especies e individuos de la familia Scarabaeidae en los tres fragmentos de hábitat (S1, S2, S3) (KW, prueba de Kruskal–Wallis) y comparación por pares a posteriori (PC, Dunn–Bonferronni) de los dos sitios del estudio (SS: H, Hrastovlje; Z, Zazid), situados en el sudeste de Eslovenia: STS, prueba estadística estandarizada; n/a: no aplicable; 1 se omitió a los rodadores del análisis porque en las muestras solo se encontró una especie (Sisyphus schaefferi); 2 solo se encontraron dos especies de la familia Geotrupidae–cavadores en las muestras, de tal forma que los análisis se realizaron con los datos agrupados de Scarabaeidae–cavadores + Geotrupidae–cavadores. SS

All

Compared groups

Test

x /STS 2

Dwellers

x /STS p

Tunellers2

x /STS 2

Rollers

x /STS p 2

Number of species 1

H S1, S2, S3 (df = 2) KW

9.743 0.008

H S1 : S2

–2.309 0.063

n/a

–1.972 0.146

n/a

H S1 : S3

0.664 1.000

n/a

0.355 1.000

n/a

H S2 : S3

2.973 0.009

n/a n/a

S1, S2, S3 (df = 2) KW

2.069 0.355

6.283 0.043 n/a n/a

2.326 0.060 n/a

n/a n/a

2.306 0.316

1.003 0.606

5.156 0.076

n/a

H S1, S2, S3 (df = 2) KW

11.154 0.004

9.502 0.009

8.098 0.017 13.120 0.001

H S1 : S2

–2.641 0.025

2.889 0.012

–2.332 0.059 –3.334 0.003

H S1 : S3

0.450 1.000

0.514 1.000

–0.246 1.000 –0.442 1.000

H S2 : S3

3.091 0.006 –2.375 0.053

2.578 0.030 2.892 0.011

2.031 0.362

5.910 0.052 1.366 0.505

Number of specimens

S1, S2, S3 (df = 2) KW

in habitat patch S1 and lowest in S2 (both study sites; table 1). However, no significant differences were detected between the three patches, either within or between the two study sites (ANOVA, p > 0.05). The number of collected species or specimens (total or subdivided by guilds) did not differ significantly between the two study sites (p > 0.10; table 2). Except for two datasets from Hrastovlje (all traps combined: p = 0.008, tunnelers: p = 0.043), there were no significant differences in the number of species between the three habitat patches (Kruskall–Wallis test: p > 0.05; table 3) in either of the two study sites (see also figure 3 for Hrastovlje–all traps combined). Furthermore, we found significant differences in the numbers of specimens only in Hrastovlje (Kruskall–Wallis test: p < 0.02; table 3). However, differences were found for the whole dataset as well as for each of the three guilds. Further post–hoc pairwise comparisons of habitat patches in Hrastovlje revealed that more species at S2 than at S3 (the whole dataset; p = 0.009); however, no differences were found between S1 and S3, and S1

0.884 0.643

and S2 (p > 0.05). There was a higher abundance of specimens at S2 than at S1 or S3 in Hrastovlje in most cases (p < 0.05), and no significant differences between S1 and S3 (see figure 3 for pooled dataset), whereas in Zazid, no statistically significant differences were found between the pairs of habitat patches (Kruskal–Wallis test: p > 0.36). Discussion Species richness We report the community richness and structure of Scarabaeoidea dung beetles from dry karst grasslands with different habitat types (pastures under different impact of livestock, and dry karst meadows) in southwestern Slovenia. Although the use of pitfall traps alone rarely reflects the majority of species present in a given study area (Barbero et al., 1999), the sampling throughout 2012 with many pitfall traps and additional baits provided a high proportion of

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NS 6

Number of species

Number of specimens

15 10

5 0

p = 0.06 S1

p < 0.01 S2

p = 0.05

0 S3

p < 0.01 S2

Fig. 3. Comparisons of numbers of specimens (A) and species (B) of Scarabaeoidea among the three habitat patches in Hrastovlje: NS, non–significant (significance p–values added). Fig. 3. Comparación de las cifras de individuos (A) y especies (B) de la familia Scarabaeidae entre los tres fragmentos de hábitat en Hrastovlje: NS, no significativo (valores p de significación añadidos).

species that are presumably present in the area, and exceeded at least 70 % of species presumably present at any of the two sampling sites. Some species that were not trapped, however, were additionally collected manually (Onthophagus vacca (Linnaeus, 1767), Coprimorphus scrutator (Herbst, 1783)) during our field trips (cf. Koprivnikar, 2012). Lobo et al. (1998) determinedthe number of traps needed to collect a certain proportion of species at local and regional scales; they suggest the use of five, seven and ten baited traps to successfully trap over 70 %, 78 % and 89 % of species at local level, respectively. Although we did not follow all their recommendations (i.e. we used six baited traps plus six control traps/ site) their estimated proportions (1998) are in line with the outcome of our study. At the habitat patch level, however, our success was expectedly lower as we used only four traps (two baited plus two control traps/habitat patch). Our trapping success ranged between 54 % (Hrastovlje, S3) and 80 % (Zazid, S3) and corresponded to the use of two to seven traps, respectively (cf. Lobo et al., 1998). Diversity and factors influencing species richness and structure Species that were exclusively recorded at a single study site may be related to the specific type of livestock present at those sites (cattle in Zazid and goats in Hrastovlje). The dung source has shown to be an important factor influencing species richness and abundance of Scarabaeiodea (e.g. Carpaneto et al., 2005; Lobo et al., 2006), and this factor may be the most important of all (e.g. Carpaneto et al., 2005). Lumaret et al. (1992) showed that changes in resources from sheep to cattle grazing resulted

in a two to three–fold increase in beetle numbers and biomass over a five–year period. However, we could not relate the species' exclusiveness on one site with the preference for a specific type of excrement or habitat type (Lobo et al., 2006) alone. Other unknown factors could influence their presence, as could the altitudinal difference (Lobo et al., 2006; Negro et al., 2011; Tocco et al., 2013) of over 300 m between the two sites. The Shannon index showed an insignificantly higher value in Hrastovlje (goats) than in Zazid (cattle). However, the abundance of dung beetles at the former site was lower. Barbero et al. (1999) reported that higher diversity is more common for pastures with cattle than for those with sheep and goats because the former habitats can support not only generalists but also more specialist species. It is true that our results deviate somewhat from this conclusion as we recorded two more species on pasture with goats, but it should also be noted that cattle excrement was used as bait. Nevertheless, the number of present species alone is only a rough estimate for species diversity. The lowest Shannon index was shown in habitat patch S2 (the overgrown part of the pasture); however, this patch supported the highest numbers of species and specimens of Scarabaeoidea. The numbers at S2 may be higher because the shadowed areas of the pasture provide a preferred microclimate for many ground living invertebrates (i.e. higher relative humidity, shade, and cooler temperatures during hot summers), and the difference then favoured S2 in comparison with any other patch (all datasets except for dwellers exclusively in Hrastovlje) or there was no difference (all datasets from Zazid). In Hrastovlje, fewer specimens of tunnelers (abundance mostly on account of O. grossepunctatus)

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and rollers were present in S3 than in S2, indicating their preference towards the shrubby areas and simultaneous requirement for availability of herbivore excrements nearby. The existence of such shrub patches within pastures is always a result of selective grazing (Stergaršek, 2009), when usually inedible and/or spiny plants (see Material and methods) are left intact. These plants can therefore serve as a temporary shelter for ground living invertebrates, as the microclimate (cool, shady and humid during the spring and early summer) is more appropriate for invertebrates. A lower abundance of animals and species of dung beetles in shrubby areas (as in our case S3) has previously been noted for the Mediterranean (e.g. Numa et al., 2009). Thus, the complete abandonment of grazing activity can consequentially lead to lower species richness and abundance of the dung beetles. A decrease in alpha diversity and biomass density in dung beetles has been shown after the pasture abandonment (Tonelli et al., 2017a, 2017b), while long–term grazing continuity and size of the pastures both have a positive effect on species richness (Buse et al., 2015). Nevertheless, Numa et al. (2009) pointed out that in homogeneous conditions of trophic resources and climate, the landscape structure is more important in determining the dung beetle assemblage than the characteristics of a given habitat patch. Synthesis and implications for conservation Recurring disturbance to agro–ecosystems is human influenced (management) and usually has a negative effect on biodiversity. Scalercio et al. (2007) reported that in the case of butterflies and moths, semi–natural habitat patches can support a higher diversity than highly transformed patches. Furthermore, species' communities in agricultural areas have high proportions of migrant species, and although some specialist species are found in these highly changed environments, it is assumed these patches primarily act only as stepping stones for species looking for permanent habitats. Although grazing and overgrowth both have a negative impact (in relation to open karst meadows) on some taxa (cf. Jugovic et al., 2013a, 2013b, 2014a, 2014b, 2017), the coprophagous dung beetle community can to some extent benefit from the presence of livestock in pastures (e.g. Lobo et al., 2006; Jay–Robert et al., 2008b). Despite some possible negative effects, extensive pastures with different types of livestock are important, especially when wild, large herbivores are absent (Zamora et al., 2007). Coprophagous animals at such places are thus commonly present because of the abundance of food (excrement; see Koprivnikar, 2012). Since the inter– and intra–guild competition of dung beetles (and some other coprophagous animals) is well expressed and can be avoided by temporal and spatial segregation (cf. Errouissi et al., 2009; Jugovic et al., 2015; Sladecek et al., 2017), we suggest that mosaic landscapes consisting of various open– to semi–open habitats for different species should be maintained for them to be able to find suitable (micro)

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habitats. In such systems, distances between patches should not overlook the movement abilities of different taxa (e.g. Jugovic et al., 2017). Mosaic landscape is consistently mentioned among the richest and most temporally heterogeneous habitats, and should be further maintained through traditional human activities (e.g. extensive grazing, only occasional and late extensive mowing; cf. Jugovic et al., 2017; Barbero et al., 1999). This landscape is human–shaped and has been managed in a way that reflects the high spatio–temporal changes (Allen, 2003), where changes in agricultural, stockbreeding activities, and abandonment or intensification can lead to the loss of biodiversity (review in Zamora et al., 2007). Acknowledgements The authors would like to thank (in alphabetic order): Miran Gjerkeš (for valuable identification keys and help with identification), Katja Kalan, Mladen Kučinić, Martina Lužnik (for constructive comments during the study), and Elena Bužan, Mitja Črne, Daniel Jehart, Domen Trkov and Sara Zupan (for help in the field). Further thanks go to the owners of the study sites (Trček family) and to the Šiler family (for providing the bait). The study was partially financed through the framework of the CBC programme Italy–Slovenia 2007–2013 (project BioDiNet–Network for the protection of biodiversity and landscape). We also thank Tilen Genov for linguistic support, and also Dr J. M. Lobo and three anonymous reviewers who significantly improved the initial version of the manuscript with their constructive comments. References Allen, H. D., 2003. Response of past and present Mediterranean ecosystems to environmental change. Progress in Physical Geography, 27: 359–377. Anthes, N., Fartmann, T., Hermann, G., Kaule, G., 2003. Combining larval habitat quality and metapopulation structure – the key for successful management of pre–alpine Euphydryas aurinia colonies. Journal of Insect Conservation 7(3): 175–185. Ballerio, A., Rey, A., Uliana, M., Rastelli, M., Rastelli, S., Romano, M., Colacurcio, L., 2010. Coleotteri Scarabeoidei d´Italia. Museo Civico di Storia Naturale Carmagnola (TO), Progetto Biodiversita Comitato Parchi, Centro Studi Roma, Muse odi Storia Naturale Venezia. (DVD). Barbero, E., Palestrini, C., Rolando, A., 1999. Dung beetle conservation: effects of habitat and resource selection (Coleoptera: Scarabaeoidea). Journal of Insect Conservation, 3: 75–84. Bergman, K. O., Landin, J., 2001. Distribution of occupied and vacant sites and migration of Lopinga achine (Nymphalidae: Satyrinae) in a fragmented landscape. Biological Conservation, 102(2): 183–190. Blondel, J., 2006. The ‘design’ of Mediterranean landscapes: a millennial story of humans and ecological

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Environmental representativeness and the role of emitter and recipient areas in the future trajectory of a protected area under climate change M. Mingarro, J. M. Lobo

Mingarro, M, Lobo, J. M., 2018. Environmental representativeness and the role of emitter and recipient areas in the future trajectory of a protected area under climate change. Animal Biodiversity and Conservation, 41.2: 333–344. Abstract Environmental representativeness and the role of emitter and recipient areas in the future trajectory of a protected area under climate change. We propose a protocol to estimate the effects of climate change on species inhabiting a reserve by assessing the location of areas with similar environmental conditions to a focal protected area, both now and in the future. Following this protocol it is possible to estimate: (i) the level of change that will occur in the current climatic conditions of a reserve; (ii) the present location of the areas with similar conditions to those this reserve will have in the future (emitter areas); and (iii) the location of the areas that in the future will have similar environmental conditions to those existing in the studied protected area (recipient areas). This knowledge can be used to anticipate and adapt the protected area against future changes. In this study, we used an Iberian reserve representative of the Mediterranean conditions, the Cabañeros National Park, as an example to calculate the extension, fragmentation and location of the areas with climatic conditions similar to those of the reserve. We also determined the connectivity between these areas and their degree of anthropic alteration. Key words: Cabañeros, Reserves, Climatic representativeness, Conservation, Climate change Resumen La representatividad ambiental y la importancia de las áreas emisoras y receptoras en la evolución ante el cambio climático de un área protegida. Proponemos un protocolo para estimar los efectos del cambio climático en las especies que habitan una reserva, evaluando la ubicación de las áreas con condiciones ambientales similares a un área focal protegida, tanto ahora como en el futuro. Siguiendo este protocolo es posible estimar: (i) el cambio que se producirá en las condiciones climáticas actuales de una reserva, (ii) la ubicación actual de las áreas con condiciones similares a las que tendrá esta reserva en el futuro (áreas emisoras) y (iii) la localización de las áreas que en un futuro tendrán condiciones ambientales similares a las existentes en el área protegida estudiada (áreas receptoras). Este conocimiento puede utilizarse para anticiparse y adaptar el área protegida a los futuros cambios. En este estudio se ha utilizado como ejemplo la reserva del Parque Nacional de Cabañeros, representativa de las condiciones mediterráneas, para calcular la extensión, la fragmentación y la localización de las áreas con condiciones climáticas similares a las de la reserva; asimismo, se ha determinado la conectividad de estas áreas y su grado de alteración antrópica. Palabras clave: Cabañeros, Reservas, Representatividad climática, Conservación, Cambio climático Received: 05 VII 17; Conditional acceptance: 18 X 17; Final acceptance: 12 I 18 Mario Mingarro y Jorge M. Lobo, Depto. de Biogeografía y Cambio Global, Museo Nacional de Ciencias Naturales–CSIC, c/ José Gutiérrez Abascal 2, 28006 Madrid, España (Spain). Corresponding author: Jorge M. Lobo. E–mail: mcnj117@mncn.csic.es

ISSN: 1578–665 X eISSN: 2014–928 X

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Introduction Protected areas (PAs) are essential for the conservation of biodiversity and most institutions attempt to integrate them into their national and international conservation strategies (Pressey et al., 2007; Palomo et al., 2014; Visconti et al., 2015). They are a crucial tool to mitigate threats related to human activity (Rodrigues et al., 2004), not only by limiting biodiversity loss (Dudley and Parish, 2006), but also by keeping natural ecosystems functional and by providing shelter for species therein. Historically, the creation of PAs has been driven by socio–economic, aesthetic and political criteria rather than by scientific or conservationist reasoning (Pressey, 1994; Fraschetti et al., 2002; Joppa and Pfaff, 2009), overlooking the fact that they should be ecologically and environmentally representative (Visconti et al., 2015). Determining the environmental representativeness of protected areas is thus a fundamental issue in systematic conservation planning and the maintenance of biodiversity (Margules and Pressey, 2000; Pressey et al., 2007; Laurance et al., 2012). Climate plays a key role when estimating environmental diversity (Faith and Walker, 1996a, 1996b; Parmesan, 2006; Chen et al., 2011; Triviño et al., 2013; IPCC, 2014) as it is a major factor conditioning biological assemblages and ecosystem characteristics (Woodward et al., 2004). However, climate is changing rapidly as a consequence of human actions. Reports from the Intergovernmental Panel on Climate Change indicate that substantial variations in climate have occurred due to the emission of greenhouse gases, and that these changes will continue to occur in the near future (IPCC, 2007, 2014). Keeping in mind that PAs have spatially fixed boundaries and are often surrounded by a matrix of transformed land uses, one might wonder what the environmental representativeness of protected areas is when the climate is changing. PAs could be considered islands representing particular environmental and biotic conditions and they may also serve to avoid the negative influence of anthropic actions. However, the effects of climate change could make these areas ineffective for their intended purpose (Lobo, 2011). On one hand, if the species that inhabit a protected area are influenced in their distribution and abundance by climatic conditions, each PA would become a recipient of outside fauna and flora. On the other hand, protected areas would also emit or export individuals to other settlements which, in the future, would represent the environmental conditions currently existing in this area. These processes of change could lead to (i) the disappearance of individuals, populations or species (Bestion et al., 2015); (ii) an increase in the evolutionary forces that promote the in situ adaptation to new conditions (Hoffmann and Sgrò, 2011); and (iii) the decline of populations and/or emigration of individuals into new territories (Mason et al., 2015; 'spatial adaptation' according to Hengeveld, 1997). Available evidence shows that the populations of some species have declined within PAs as consequence of climatic changes, while other species have undergone a population growth or colonized a reserve for the first time (Thomas and Gillingham, 2015).

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Although PAs may act as natural shelters against the effects of climate change (Thomas and Gillingham, 2015; Gaüzere et al., 2016), creating corridors between them can facilitate their inter–connection (Haddad et al., 2015). PAs representing different climate conditions should be connected in order to minimize the threat of local extinction and maximize the adaptive and dispersive possibilities of organisms. Most studies that select the location of possible reserves keeping climate change scenarios in mind have used distribution models able to anticipate the geographic response for each species to changes in climate (Jones et al., 2016). Such predictions have several drawbacks (Lobo, 2015). For example, they may produce inconsistent and unreliable results because they do not include estimations about the real and direct effect of climate variables in delimiting the occurrence and abundance of species (Araújo et al., 2011; Felicísimo et al., 2011). Using individual predictive species distribution models to estimate the possible future location of areas that should be protected is a hazardous strategy. This is because the many uncertainties of each individual model may lead to the misappropriation of conservation resources in some regions. Moreover, identifying climatically favourable territories for species without taking future and possible changes in land use into account can also lead to an inefficient selection of areas (Faleiro et al., 2013; Jones et al., 2016). Instead of trying to estimate the effects of climate change on the species inhabiting a reserve, we here propose an approach based on estimating the location of areas with environmental conditions similar to those of a focal PA, both now and in the future. Assuming that the environmental conditions of a PA are the main determinants of its conservation value (Albuquerque and Beier, 2015), we propose estimating (i) the present location of the areas with similar conditions to those this PA will have in the future (emitter areas), and (ii) the future location of the areas that will have similar environmental conditions to those currently existing in the focal PA (recipient areas). If we cannot reliably predict the future distribution of each species because we do not know the true and contingent effects of climate on each one, the proposed approach aims to estimate the environmental representativeness of a protected area to derive conservation strategies. This knowledge can be used to anticipate and adapt PAs against future changes. In this work, we used an Iberian reserve that is representative of Mediterranean conditions, Cabañeros National Park, as an example of focal PA: (i) to estimate the current and future climate representativeness of this reserve; (ii) to evaluate the level of change that will occur in its current climatic conditions, calculating the extension, fragmentation, connectivity and location of the climatic conditions that Cabañeros currently represents and will represent in the future; and (iii) to identify recipients and emitter areas under a future climate change scenario, as well as the connectivity of these areas to the focal PA, taking into account the degree of anthropic alteration of the entire Iberian territory.

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Fig. 1. Regions with similar edaphic conditions to those existing in the Cabañeros National Park, which in addition have natural land uses. The color range represents pH variation. Fig. 1. Regiones con condiciones edáficas similares a las del Parque Nacional de Cabañeros y que además poseen usos del suelo naturales. El gradiente de colores representa la variación en pH.

Material and methods The focal protected area Cabañeros was declared a National Park in 1995. It is located in the region of Montes de Toledo in central Spain (39.414 N, –4.509 W) between the provinces of Ciudad Real and Toledo. It covers an extension of 40,856 hectares. Its elevation oscillates between 520 and 1,448 m a.s.l., with an average altitude of 788 m. Cabañeros bioclimatically represents the Mediterranean region. Most of the territory belongs to the mesomediterranean bioclimatic type (520–1,000 m), while the supramediterranean bioclimatic type (1,000–1,450 m) only appears in the NE part of the region (Rivas Martinez, 1987). Origin of climatic data Data on current climate come from the University of Extremadura (see methodology in Felicísimo et al., 2011) and include data about mean maximum monthly daily temperature, minimum monthly daily temperature, average monthly temperature and total monthly precipitation from 1950 to 2007 for the whole Iberian Peninsula. Using this primary source of climatic information at a resolution of 1 km2 UTM grid cells and the formulas of Valencia–Barrera et

al. (2002) and López Fernández and López (2008) we built a total of 23 bioclimatic variables (table 1). As Felicísimo et al. (2011) do not provide future monthly climatic data, we used WorldClim data (http://www.worldclim.org/) at a resolution of 30 arc seconds (~ a cell of 0.82 km2). The model we selected was the IPSL–CM5A–LR from the Pierre Simon Laplace Institute (Dufresne et al., 2013), specifically that from the fifth assessment report (AR5) that predicts a mean increase in temperature of 1.3 ºC around 2050 (RCP6.0) (Van Vuuren et al., 2011). We selected this climatic projection for its intermediate character concerning greenhouse gas emissions and socioeconomic assumptions. We used the predicted values of the four primary climatic variables mentioned formerly for 2070 (mean maximum monthly daily temperature, minimum daily monthly temperature, average monthly temperature and total monthly precipitation) to derive the same 23 bioclimatic variables for the future as those obtained for present times following the explained procedure. Other environmental data Suitable climatic conditions do not guarantee that a given species can inhabit a locality. To restrict both recipients and emitter areas, we also considered soil characteristics and land uses. Unlike climate,

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Table 1. Bioclimatic variables obtained using the formulas provided by Valencia et al. (2002) and López Fernández and López (2008), and also correlations between the values of these variables and the three factors that emerged from a principal components analysis. Values > 0.7 are shown in bold. Tabla 1. Variables bioclimáticas obtenidas a partir de las fórmulas proporcionadas por Valencia et al. (2002) y López Fernández y López (2008) y correlación entre los valores de estas variables y los tres factores que surgieron de un análisis de componentes principales. Los valores > 0,70 se indican en negrita. Variable

Factor 1

Factor 2

Precipitation seasonality

0.9170 0.2740 0.1457

Temperature seasonality

–0.4825

–0.1465

Factor 3 –0.8254

Isothermality

0.5271 0.0435 0.5471

Aridity index (Martonne)

–0.8480 0.4296 –0.2883

Continentality index

–0.2414

Precipitation contrast

0.9212 0.2569 0.1743

Thermal contrast

–0.4221

Ombrothermic index I0

0.7930 –0.5183 0.2807

Ombrothermic index I5

0.8497 –0.3215 0.3637

Annual precipitation

0.9183 –0.2122 0.3173

–0.1544 –0.1351

–0.8325 –0.8939

Precipitation in wettest month

0.9464 –0.0040 0.2985

Precipitation in driest month

0.3365

Positive precipitation 0

0.9131 –0.1695 0.3315

Positive precipitation 5

0.8672 0.1156 0.3660

–0.7076 0.4210

Emberger’s pluviometric ratio

0.7314 –0.5931 0.1074

Maximum temperature in warmest month

–0.3436

0.6228

–0.6970

Average monthly maximum temperature

–0.3410

0.6318

–0.6898

Annual mean temperature

–0.1142

0.9800 –0.0915

Average monthly minimum temperature

0.1167

0.9479 0.2869

Minimum temperature in coldest month

0.1167

0.9479 0.2869

Absolute minimum temperature

0.0818

0.9673 0.2129

Positive temperature 0

–0.1120

0.9784 –0.0964

Positive temperature 5

–0.0763

0.9772 –0.0763

soil features are not subject to short–term modifications and are relatively independent of climatic alterations, at least in short time spans. Therefore, if the occurrence of a species is conditioned by both edaphic and climatic characteristics, it will be necessary to consider both requirements to delimit its probable distribution. In this study, we used pH as a general surrogate of the edaphic characteristics. We obtained pH data from the European Soil Data Centre (http://esdac.jrc.ec.europa.eu/; see Reuter et al., 2008) showing continuous pH values for each of the 1 km2 UTM grid cells of the Iberian Peninsula. Additionally, we used information on land use from the CORINE Land Cover project (www.eea.europa. eu) to limit the edaphic–climatic areas to those with natural conditions. To do this, we reclassified the

different land uses recognized in CORINE (level 2; resolution: 100 m2) for 2011 into three categories: anthropic, semi–anthropic, and natural (table 2), eliminating the localities categorized as anthropic or semi–anthropic from the climatic–edaphic suitable areas. Thus, suitable edaphic areas with natural land uses (fig. 1) constitute the most restricted geographical scenario to represent recipient and emitter areas. Finally, we downloaded a digital cartography representing the Iberian protected areas included in the Natura 2000 network from Protected Planet (www.protectedplanet.net/) and used this to describe which are, and will be the PAs that have and will have similar environmental conditions to those in Cabañeros.

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Selection of climatic variables After standardizing the values of all the considered climatic variables to mean zero and one standard deviation to eliminate the effect of different measurement scales, we conducted a principal components analysis (PCA) to reduce the number of climatic variables that would be used. PCA provided three non–correlated factors with eigenvalues higher than 1, representing 93.5 % of all the climatic variability in the Iberian Peninsula (factor 1 = 53.6 %, factor 2 = 31.7 %, factor 3 = 8.2 %). For each one of these three factors we chose the original variable with the highest factor loading; i.e. the primary variable best correlated with the values of each factor. The values of the first factor were positively correlated with different precipitation variables and negatively correlated with soil acidity (table 1), selecting the precipitation of the wettest month as representative (factor loading = 0.9464). The second factor was positively correlated with different temperature variables and negatively correlated with the precipitation of the driest month (table 1). On this occasion, the annual average temperature was chosen as the most representative variable (factor loading = 0.9800). Finally, the third factor was negatively correlated with temperature seasonality, continentality and thermal contrast (table 1), selecting thermal contrast as the representative variable (factor loading = – 0.8939).

Like isothermality, average monthly maximum temperature and maximum temperature of the warmest month were relatively poorly represented by the selected PCA factors (table 1); the first two were also selected to describe the climatic conditions of Cabañeros (only one of the two temperature variables was selected because both were highly and positively correlated; r = 0.997, p < 0.0001). Data analysis The five previously selected climatic variables were used to calculate the Mahalanobis distance (MD) from the conditions in the 1 km2 cells of the National Park of Cabañeros to all remaining Iberian cells. We thus obtained a continuous measure able to represent not only the places with the same conditions to those of Cabañeros, but also the places with relatively similar conditions. The process was repeated both for present and for future climatic data. MD was chosen to measure climate similarity because this multidimensional measure takes into account the correlations of the variables and it is scale–invariant regardless of the units used for each variable (Farber and Kadmon, 2003; Xiang et al., 2008). We used the value corresponding to the 90th percentile of the MD values appearing in Cabañeros as the decision threshold to delimit the areas with a climate highly similar to that in the national park. Subsequently, a similar estimate

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Table 2. The Iberian area currently represented by the climatic (C) or the climatic and edaphic (CE) conditions of Cabañeros National Park, and CE areas with natural land cover (CEN). CEN areas currently protected by any type of reserve, CEN areas within the Nature 2000 network (N2000), CEN area covered by large and continuous patches, total number of patches in CEN, and the value of the area–weighted mean shape index (AWMSI). The same values are provided for recipient areas (sites that in the future will have similar environmental conditions to those currently existing in Cabañeros) and emitter areas (that at present have similar conditions to those that Cabañeros will have in the future). C, CE and CEN percentages are computed considering the total area of the Iberian Peninsula, while remaining percentages are calculated on the basis of the CEN area. Tabla 2. Superficie de la península ibérica que actualmente está representada por las condiciones climáticas (C) o climático–edáficas (CE) del Parque Nacional Cabañeros, y áreas CE con cobertura natural (CEN). Zonas CEN actualmente protegidas por cualquier tipo de reserva, zonas CEN dentro de la Red Natura 2000 (N2000), área CEN cubierta por parches grandes y continuos, número total de parches en CEN y valor del índice medio de forma ponderado por el área (AWMSI). Se proporcionan los mismos datos para el área receptora (sitios que tendrán en el futuro condiciones ambientales similares a las existentes hoy en Cabañeros) y el área emisora (sitios que actualmente tienen condiciones similares a las que tendrá en el futuro Cabañeros). Los porcentajes de C, CE y CEN están calculados sobre la superficie total de la península ibérica, mientras que los restantes están calculados sobre el área CEN.

Current representativeness

km2

Future emitter areas

km2 %

C 157,327 27.0

C 48,990 8.4

CE 92,280 15.9

CE 35,244 6.1

CEN 42,030 7.2

CEN 16,100 2.8

Protected 19,355 46.05

Protected 6,806 42.27

N2000 17,419 41.45

N2000 5,937 36.87

Patches > 10.000 km2 25,509 60.7

Patches > 10.000 km2

0.0

Patches 1.000–10.000 km

Patches 1.000–10.000 km 11,508

71.5

5,442 12.9

Number of patches 6,985

Number of patches 2,881

AWMSI

59.96

27.84

Future recipient areas

km2 %

C 37,630 6.5 CE 9,048 1.6 CEN 5,023 0.9 Protected 2,218 44.17 N2000 2,002 39.86 Patches > 10.000 km2 Patches 1.000–10.000 km 2

0.0

1,382 27.5

Number of patches 1,505 AWMSI 10.02

was made taking into account both climatic and edaphic variables. For this purpose, we estimated the range of pH values appearing within Cabañeros (from 5 to 6), removing all the areas outside these pH values from the climatically favourable territory. However, considering that species can be relatively tolerant to pH variations (Prentice et al., 1992), pH ranges were modified in ± 0.5 (i.e. from 4.5 to 6.5)

in order to include those with relatively similar pH conditions as edaphically favourable regions. Once identified and mapped, the areas with favourable climatic and edaphic conditions (i.e. those with MD values lower than the 90th percentile value calculated for Cabañeros) were overlapped with the current natural areas according to CORINE land cover as well as with the polygons representing Natura

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Fig. 3. Regions with climatic (A), climatic and edaphic (B), and climatic–edaphic areas currently harbouring natural land cover conditions that, in the future, will have similar conditions to those currently existing at Cabañeros National Park (C). Climatic data come from the IPSL–CM5A–LR scenario of the fifth assessment report (AR5) (2060–2080). Fig. 3. Regiones con condiciones climáticas (A), climático–edáficas (B) y climático–edáficas con una cobertura natural que, en el futuro, serán similares a las existentes actualmente en el Parque Nacional Cabañeros (C). Los datos climáticos provienen del escenario IPSL–CM5A–LR del quinto informe de evaluación (AR5) (2060–2080).

2000 PAs. Finally, considering that fragmentation is one of the biggest threats to biodiversity conservation (Fahrig, 2003), we calculated the area, number and location of the groups of localities connected or adjacent (touching each other), assuming that a high fragmentation diminishes the conservation value of recipient and emitter areas. To do this, we used only those areas that are suitable from the climatic and edaphic point of view and, also have natural land uses. We also measured fragmentation using the area–weighted mean shape index (AWMSI). This index measures the average perimeter–to–area ratio, weighted by the size of the patches so that larger patches weigh more than smaller ones (McGarigal et al., 2012). This index is equal to 1 when all patches are circular, increasing in value without limit as patch shapes become more irregular. Results Current representativeness The climatic conditions of Cabañeros are found in a large area of the Iberian Peninsula (fig. 2), accoun-

ting for 27 % (157,327 km2) of its total area (table 2). Part of the northern sub–plateau just above Serra da Estrela, almost all of the southern sub–plateau to the Guadalquivir valley, and the Subbaetic mountains are climatically similar areas to those of Cabañeros. However, the region with similar climatic and edaphic conditions covers a considerably smaller area as the result of the elimination of eastern calcareous areas, totalling around 16 % (92,280 km2) of the Iberian Peninsula (fig. 2). That is a decrease of 41 % in the representative area (65,047 km2 less). Within the National Park, only 12 % of the territory is dedicated to anthropic land uses. In contrast, the representative Iberian climatic and edaphic area is highly anthropized (34.5 %) and only 45.5 % of it harbours natural landscapes (around 42,030 km2; see fig. 2 and table 2). Taking into account the climatic and edaphic conditions with natural land cover, around 19,355 km2 (46 % of this area) is included under some type of protection category, with 90 % corresponding to the Natura 2000 Network (table 2). Connectivity between the climatic and edaphic favourable area and natural land cover is high and its fragmentation low. Only two patches have more than 10,000 km2, representing 60.7 % of this total area, and another four patches

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Fig. 4. Regions with climatic (A), climatic and edaphic (B), and climatic–edaphic areas harbouring natural land cover conditions that currently have similar conditions to those that Cabañeros National Park will have in the future (C). Climatic data come from the IPSL–CM5A–LR scenario of the fifth assessment report (AR5) (2060–2080). Fig. 4. Regiones con condiciones climáticas (A), climático–edáficas (B) y climático–edáficas con una cobertura natural que, actualmente, son similares a las que el Parque Nacional de Cabañeros tendrá en el futuro (C). Los datos climáticos provienen del escenario IPSL–CM5A–LR del quinto informe de evaluación (AR5) (2060–2080).

embody 13 %. In total, there are 6,985 patches and the AWMSI index is 59.96 (table 2). Future recipient areas The places with the climatic conditions currently represented by Cabañeros are greatly reduced in the future scenario, and their geographical location also shifts (fig. 3). The climatically favourable area would be divided into two fragments, a smaller area located in the South of the 'Montes de León', in the Portuguese region of Tras–Os–Montes and Spanish territories bordering with Portugal, and a larger area located in a strip from the eastern part of the southern plateau below the Iberian System to Sierra Nevada. As a consequence, between 2060 and 2080 around 120,000 km2 of climatically representative area will have disappeared (table 2). This change could establish a new climatically favourable region equivalent to approximately 6.5 % (37,630 km2) of the total Iberian Peninsula area. When edaphic conditions are also considered, representative areas would cover a much smaller area (9,048 km2; 1.6 % of total Iberian area). About half of this future climatically and eda-

phically favourable territory currently has natural land cover (5,023 km2), being 44 % currently protected (2,218 km2) (fig. 3, table 2). In this case, no patch has more than 10,000 km2 and only one patch has more than 1,000 km2 representing 27.5 % of the total. Taken together, there are 1,505 patches and the value of the AWMSI index decreases to 10.2 (table 2). Future emitter areas The areas that currently have similar climatic conditions to those that Cabañeros will have in the future occupy 48,990 km2, which is equivalent to 8.4 % of the total area of the Iberian Peninsula (fig. 4, table 2). These areas are located in two main parts of the southern plateau: one between the Guadiana and Tajo valleys (Montes de Toledo, Villuercas, etc.), and another in the southeast around the Subbaetic mountain chain. Furthermore, there are also 643 km2 very close to the Tajo International Natural Park, located in the boundary between Spain and Portugal. When both the climatic and edaphic conditions are considered this area is reduced to 35,244 km2. Around 34 % of these favourable climatic and edaphic territories cu-

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Fig. 5. Regions that will act as emitter (A) or recipient (B) areas for the Cabañeros National Park in the future. Fig. 5. Regiones que en el futuro se comportarían como zonas emisoras (A) o receptoras (B) para el Parque Nacional de Cabañeros.

rrently have an anthropic land use, while 46 % have natural land cover (table 2). These favourable and natural areas are divided in two main patches (fig. 4), and they include part of the current territory of the Cabañeros National Park. About 43 % of this climatic and edaphic area is currently protected (6,806 km2), mainly by the Natura 2000 network (87 %). This area would be composed of 2,881 patches with an AWMSI value of 27.84 (table 2). Discussion The Iberian reserve selected in this exercise has a key role in terms of climate representativeness as it represents Mediterranean forest conditions better than other National Parks (Sánchez–Fernández et al., 2013), such as those in mountain areas (Lobo, 2011), which barely represent a few hectares beyond their protected area boundaries. Even when both climatic and edaphic conditions are considered together, this Iberian reserve remained representative of a large part of the Iberian territory (around 16 %). If the Iberian territory represented by Cabañeros in regard to climatic and edaphic conditions is large, rather than non–fragmented, with little human impact and many

protected areas, we can assume that the species inhabiting this reserve and environmentally similar areas have great potential to maintain their connected and conserved populations. Indeed, almost half of this territory currently possesses a high degree of wilderness, three–quarters is protected, and the general degree of fragmentation is very low; large and continuous patches cover 73 % of the suitable natural conditions. Thus, the potential environmental niche of many of the species sheltered in Cabañeros would also, a priori, appear in these other protected and natural areas, and vice versa. Remarkably, it appears that the size of the areas representing the current environmental conditions of this reserve will be drastically reduced and fragmented in the future (future recipient areas). Climatic and edaphic future suitable areas that currently have natural land cover may be ten times smaller and only a quarter of them would be located in continuous and large patches. In addition, we should stress that only a small part of the territory that could act as a recipient area is located close to the examined reserve (fig. 5); the conservation of these localities should be given priority because they may ensure the maintenance of some of the organisms currently protected by this reserve. In this specific case, the most important area

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with optimal conditions to act as a recipient area is the Special Protection Area of Montes de Toledo because the species currently inhabiting lowlands will find suitable climatic conditions in highland areas even when these are located within the park. This whole set of results suggests a strong reduction in the environmental conditions currently represented by Cabañeros, thus probably diminishing the climatic–edaphic niche of many of the species that currently inhabit Cabañeros. This could result in the export of faunistic and floristic elements to areas in which these conditions will appear in the future and, in general, to a drastic reduction of the Mediterranean conditions that motivated the creation of this national park. According to our results, between 2060 and 2080, Cabañeros National Park will undergo changes in climatic conditions similar to those currently appearing in other areas. These probable emitter areas seem to be larger, currently protected, and not very fragmented (fig. 5). Around 16,000 km2 of natural land cover currently have climatic and edaphic conditions similar to those that Cabañeros will have in the future. A large part of this area is currently protected and located under continuous and larger patches that encompass the park itself. The International Tajo Natural Park, located on the border between Badajoz province and Portugal, will constitute the main protected emitter area, together with Sierra de las Villuercas, Tajo River, and Monfragüe National Park. These reserves can be important areas from which populations and species will eventually reach Cabañeros, but even the lower elevation parts of the National Park itself can act as emitter areas. It is necessary to promote the connectivity of these areas to facilitate the long–term stability of biodiversity (Haddad et al., 2015). Taken together, these results suggest that the import of new populations and species in Cabañeros is more probable than the export of species. If climatic and edaphic conditions determine the fauna and flora of this National Park, climate change will generate a deep alteration of its biotic elements (Bestion et al., 2015), basically due to the entry of new elements. These changes could increase the populations of some colonizing species and largely decrease those of other native species. As such changes could lead to various conservation and management problems (Thomas and Gillingham, 2015), it is necessary to anticipate possible alterations and solutions, such as avoiding the isolation of the park and facilitating flow to and from the areas indicated in this study. References Albuquerque, F., Beier, P., 2015. Global patterns and environmental correlates of high–priority conservation areas for vertebrates. Journal of Biogeography, 42: 1397–1405. Araújo, M. B., Guilhaumon, F., Neto, D. R., Pozo–Ortego, I., Gómez–Calmaestra, R., 2011. Impactos, vulnerabilidad y adaptación al cambio climático de la biodiversidad española. 2. Fauna de Vertebrados. Dirección general de medio Natural y Política

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Population dynamics of the endangered seahorse Hippocampus reidi Ginsburg, 1933 in a tropical rocky reef habitat N. Freret–Meurer, T. Fernández, N. Okada, A. Vaccani

Freret–Meurer, N., Fernández, T., Okada, N., Vaccani, A., 2018. Population dynamics of the endangered seahorse Hippocampus reidi Ginsburg, 1933 in a tropical rocky reef habitat. Animal Biodiversity and Conservation, 41.2: 345–356. Abstract Population dynamics of the endangered seahorse Hippocampus reidi Ginsburg, 1933 in a tropical rocky reef habitat. This study was conducted in Armação de Búzios, Brazil, a municipality where ecosystem degradation has been observed following large increases in tourism and population growth. The goal of this study was to determine seasonal variations in three Búzios populations of the long snout seahorse Hippocampus reidi. Monthly dives were conducted from November 2011 to October 2013. All three subpopulations had low densities of seahorses and no seasonality. The sex ratio differed at each site. The most commonly used microhabitats were the sponge Aplysina fulva and the seaweed Sargassum sp. There was no significant difference in temperature and salinity. The environmental trends could not explain the variation in seahorse density at the three beaches. The population showed no seasonality and no further decline. Key words: Monitoring, Seasonality, Fish, Syngnathidae, Brazil Resumen Dinámica de la población del hipocampo en peligro de extinción Hippocampus reidi Ginsburg, 1933 en un arrecife rocoso tropical. Este estudio se realizó en Armação de Búzios, en Brasil, un municipio en el que se ha observado la degradación de algunos ecosistemas tras el gran aumento del turismo y el crecimiento demográfico. El objetivo fue determinar las variaciones estacionales de tres poblaciones de Búzios del caballito de mar, Hippocampus reidi. Las inmersiones mensuales se realizaron entre noviembre de 2011 y octubre de 2013. Las tres subpoblaciones tenían baja densidad de caballitos de mar y carecían de estacionalidad. La razón de sexos fue diferente para cada sitio. Los microhábitats más utilizados fueron la esponja Aplysina fulva y el alga Sargassum sp. No hubo diferencias significativas ni en temperatura ni en salinidad. Las tendencias ambientales no pudieron explicar la variación de la densidad de caballitos de mar en las tres playas. La población no mostró estacionalidad y no disminuyó más. Palabras clave: Seguimiento, Estacionalidad, Peces, Syngnathidae, Brasil Received: 5 X 17; Conditional acceptance: 23 XI 17; Final acceptance: 16 I 18 Natalie Freret–Meurer, Tatiane Fernández, Nayara Okada, and Amanda Vaccani, Lab. de Comportamento Animal e Conservação, Inst. de Ciências Biológicas e Ambientais, Univ. Santa Úrsula, Rua Fernando Ferrari, 75, Prédio 4/ Sala 405, Botafogo, CEP 22231–040 Rio de Janeiro, RJ Brasil. Corresponding author: Natalie Freret–Meurer. E–mail: nataliefreret@yahoo.com.br

ISSN: 1578–665 X eISSN: 2014–928 X

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Introduction Seahorses are reef fishes belonging to the Syngnathidae family, with 42 species currently described (Lourie et al., 2016; Han et al., 2017). They are found all over the world in both tropical and sub–tropical waters (Lourie et al., 1999). Three seahorse species occur in Brazil, namely Hippocampus reidi Ginsburg, 1933, Hippocampus erectus Perry, 1810 and Hippocampus patagonicus Piacentino & Luzzatto 2004, and all can be found in several threatened species lists (Mazzoni et al., 2000; IUCN, 2010; MMA, 2010). H. reidi is commonly called the Long Snout Seahorse and is the most abundant species of the Rio de Janeiro state coast (Freret–Meurer, 2010), although it is classified as vulnerable by the List of Threatened Species of the Rio de Janeiro State (Mazzoni et al., 2000). Its population has been declining for the last 30 years (Costa–Neto, 2000) but little is known about its ecology, population structure, distribution, or behavior. The Armação de Búzios area comprises many costal ecosystems such as beaches, rocky reefs, and bays usually inhabited by seahorses (Rosa et al., 2007; Freret–Meurer and Andreata, 2008). The area has increasingly attracted large numbers of tourists because of its natural beauty, resulting in high city development in recent years (Xavier, 2006). In addition, local fishermen traditionally capture marine animals, including seahorses, to sell as souvenirs or handcrafts. The pressure associated with their capture for the aquarium market has a detrimental impact in the local marine species (Freret–Meurer, pers. observ.). Information regarding the seahorse population in Búzios over the last two decades is lacking but there are anecdotal reports that seahorses used to be frequently seen on beaches and were constantly captured in fishing nets. Currently, it is rare to find seahorses in the area. These animals have great potential as a flagship species for conservation due to their peculiar morphology and life history (Shokri et al., 2009) The present study aimed to check for seasonal variations in population parameters of H. reidi in three populations, at Armação de Búzios, Rio de Janeiro, Southeastern Brazil, by recording population size, reproductive state, sex ratio, ratio between juveniles and adults, behavior, substrate and occurrence depth. Environmental trends, such as salinity, temperature and visibility were also characterized at each study area. Material and methods Study sites The state of Rio de Janeiro, located in the southeastern region of Brazil, has a coastal area of 800 km and exhibits great ecosystem diversity, including a continuous rocky reef (Freret–Meurer, 2010). Surveys were conducted at three beaches in the southeast of the region, namely João Fernandes, Canto and Ossos. These beaches were selected due to the

previous report by Freret–Meurer (2010) and the fact that they might be good indicators of the seahorse population conservation status in the state of Rio de Janeiro (fig. 1). Field sampling Dives were conducted monthly from November 2011 to October 2013 by snorkeling, between 8:00 a.m. and 5:30 p.m. Scuba diving was not necessary, since the rocky reef depth is less than 4 m. Water visibility ranged from 80–100 % from surface to the bottom of the study site and did not interfere with the research. H. reidi were searched for in eight fixed transects (20 x 5 m, comprising 800 m2) by two divers at each site of the transect (2.5 m), parallel to the coast. The species was identified according to Lourie et al., (2004) and Figueiredo and Menezes (1980). Sex was recognized by the presence (male) or absence (female) of the brood pouch (Lourie et al., 1999) in individuals larger than 100 mm only (Silveira, 2005). Individual height was measured from the top of the coronet to the stretched tail (Lourie, 2003). Reproductive state was determined according to Lourie (2003), as follows: for males: 0, just given birth, pouch flabby; 1, ouch empty, pouch flat; 2, pregnant, pouch rounded; 3, about to give birth, pouch extremely rounded and shiny, and for females: 0, eggs recently entrusted to males, belly sunken; 1, no mature eggs, belly flat; 2, bearing mature eggs, belly slightly raised; 3, hydrated eggs, belly distended. Juveniles were identified by size, according to Rosa et al. (2007); specimens smaller than 100 mm were considered juveniles. Data on behavior were obtained by the scan method (Altmann, 1974) which recorded the behavior of the individual at the time it was found. Depth and substrate where the animals were anchored were also recorded. Water temperature and salinity were measured at each site. Statistical analyses Descriptive statistics were reported as percentages and means ± standard deviation. Comparative analyses of population densities between sites were performed by the Kruskal–Wallis test, followed by Dunn's test because these parameters did not pass the normality assumption (Kolmogorov–Smirnov test; α < 0.05) and homoscedasticity test (Bartlett's test; α < 0.05). Comparisons between individual height, depth, temperature and salinity were conducted using the ANOVA one–way test, followed by Tukey's test. The correlation between height and depth was conducted using Spearman correlation. Multiple regression was used to assess the relationship between temperature/salinity and seahorse density. Results Canto beach showed a mean density of 0.029 ± 0.572 ind/m2 throughout the study period. Higher densities were recorded from December

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Brazil

3 Armação dos Búzios

Rio de Janeiro

2 km

564 km

Fig. 1. Seahorse monitoring at three beaches in Armação dos Búzios, state of Rio de Janeiro, south–eastern Brazil: 1, João Fernandes; 2, Ossos; 3, Canto. Fig. 1. Seguimiento de los caballitos de mar en tres playas ubicadas en Armação dos Búzios, en el estado de Río de Janeiro, en Brasil: 1, João Fernandes; 2, Ossos; 3, Canto.

2012 to April 2013. No seasonality between wet (November–April) and dry season (May–October) was detected (p = 0.30; t = 1.093; df = 11). João Fernandes showed considerable variation in population density due to changes in the locations of seahorse population patches. A patch was initially detected, but due to the absence of any animals during population monitoring (November 2011 to June 2012), a methodological readjustment was required, with an increase in the number of transects (n > 8) and, in parallel, an intensive search for the seahorses, to find the patch again. In July 2012 the patch was found, with high numbers of seahorses for this month (average density = 0.05 ± 0.07 ind/m2), remaining constant throughout the rest of the study months, and peaking in August 2012, with an average density of 0.1 ± 0.25 ind/m2. Once methodology was adjusted, we used only one year data, from November 2012 to October 2013, to compare wet (November 2012–April 2013) and dry (May 2013–October 2013) seasons. We found a statistical difference between the dry and wet seasons for the population in João Fernandes (p = 0.003; t = 5.198; df = 5), representing a higher population in the dry season (0.039 ± 0.002 ind/m2). Density at Ossos beach remained relatively constant, with an average total of 0.04 ± 0.02 ind/m2. Density peaks occurred in March 2012, of 0.07 ± 0.12 ind/ m2, followed by peaks in April 2012, August 2012

and October 2012, all of 0.06 ind/m2 (standard deviations of 0.09; 0.09 and 0.07, respectively). The lowest densities were observed in November and September 2012, of 0.01 ± 0.04 ind/m2. No significant difference was found between dry and wet seasons (p = 0.191; t = 1.393; df = 11). Densities between the three beaches were significantly different (p = 0.01; KW = 8.149), being higher at Ossos Beach, followed by Canto Beach and, finally, João Fernandes beach, with the latter two showing no significant difference between them (table 1). It is important to note that the established density of João Fernandes after adjustment was higher than that at the other beaches (fig. 2). Sex and juvenile x adult ratios The operational sex ratio at Canto Beach was of 1:2, skewed towards females. Contrary to what was found during most of the study months, only males were recorded in June (3:0). Despite the unequal sex ratio and the fact that females were absent from the study area for many months, all observed adults (100 %, n = 14) were reproductively active. Only adults were observed in the area, except in August, when a juvenile specimen was found. The overall sex ratio for the population of seahorses at João Fernandes was 1:1, but this varied over the study months, with

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0.12

João Fernandes Ossos

Density (ind/m2)

0.1

Canto

0.08 0.06 0.04 0.02 0

XI XII I II III IV V VI VII VIII IX X XI XII I II III IV V VI VII VIII IX X 2011

2012

2013

Fig. 2. Density (ind/m2) of H. reidi seahorses during the study period at João Fernandes, Ossos and Canto beaches. Fig. 2. Densidad (ind./m2) de los caballitos de mar H. reidi durante todo el período de estudio en las playas de João Fernandes, Ossos y Canto.

a trend towards the presence of females in July and August 2012, and towards the presence of males in September and October 2012 (fig. 3). The number of juvenile individuals in the area remained constant over the study months (1 per month), but the number of adults was variable (fig. 3). All the recorded adults (100 %) were reproductively active. The total operational sex ratio found for this population was skewed towards males (2:1), dominant in March and April 2012 and from June to October 2012. The sole presence of males at Ossos Beach from June to September 2012 is noteworthy. In addition, only adults in this

Table 1. Matrix results of Dunn's post–test analysis, indicating the differences between seahorse population densities at the studied beaches (α = 0.05). * Statistically significant. Tabla 1. Resultados de la matriz del análisis a posteriori de Dunn, que indican las diferencias entre las densidades de población de caballitos de mar de las playas estudiadas (α = 0,05). * Estadísticamente significativo. Study sites

p–value

João Fernandes x Ossos

< 0.05*

João Fernandes x Canto

> 0.05

Ossos x Canto

< 0.05*

study area were observed, reproductively active at all times (100 %), including during the months when no females were found in the area. The sex ratio was different between the beaches, indicating distinctive characteristics. Canto Beach showed a total operational sex ratio skewed towards females (1:2), while a proportionate sex ratio (1:1) was observed at João Fernandes Beach. At Ossos Beach, the same parameter was skewed towards males (2:1) (fig. 3). Regarding the juvenile x adult ratio, João Fernandes and Canto Beaches were observed as areas that receive and house juveniles, while Ossos Beach showed no juvenile individuals throughout the study period. Size and depth Seahorses at Canto Beach had a mean size of 155 ± 34 mm. The mean size of adults observed at João Fernandes Beach was 144 ± 15 mm, while juveniles had a mean size of 83.75 ± 8.9 mm. The mean size of adult seahorses at Ossos Beach was 162.2 ± 21 mm, remaining constant throughout the study months. Adult sizes differed significantly between beaches (p = 0.023, F = 3.990). The multiple comparison Tukey–Kramer test indicated that animals from João Fernandes Beach were significantly smaller (p < 0.05; q = 3.89) than those from Ossos Beach, but similar to Canto Beach (p > 0.05, q = 2.87). No significant difference was observed between seahorse size from Ossos and Canto Beach (p > 0.05; q = 0.26). Regarding height, a higher number of

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Percentage

100 90 80 70 60 50 40 30 20 10 0

XI XII I II III IV V VI VII VIII IX X XI XII I II III IV V VI VII VIII IX X

2011

100 90 80 70 60 50 40 30 20 10 0

2012 2012

XI XII I II III IV V VI VII VIII IX X XI XII I II III IV V VI VII VIII IX X

2011

2012

2013

100 90 80 70 60 50 40 30 20 10 0

2012

2013

XI XII I II III IV V VI VII VIII IX X XI XII I II III IV V VI VII VIII IX X 2011 2011

100 90 80 70 60 50 40 30 20 10 0

XI XII I II III IV V VI VII VIII IX X XI XII I II III IV V VI VII VIII IX X

2011

2013 2013

Percentage

XI XII I II III IV V VI VII VIII IX X XI XII I II III IV V VI VII VIII IX X 2011 2011

100 90 80 70 60 50 40 30 20 10 0

2013

Percentage

100 90 80 70 60 50 40 30 20 10 0

2012

349

2012 2012

2013 2013

XI XII I II III IV V VI VII VIII IX X XI XII I II III IV V VI VII VIII IX X

2011

2012

2013

Fig. 3. Total operational sex ratio (%) (F, females; M, males) and juvenile (J) x adult (A) ratio at the three study areas João Fernandes (A), Ossos (B) and Canto beaches (C), Armação de Búzios (RJ). Fig. 3. Razón de sexos total (%) (F, hembras; M, machos) y proporción entre juveniles (J) y adultos (A) en las tres zonas del estudio de playa João Fernandes, playa Ossos y playa Canto, en Armação de Búzios (RJ).

individuals from Ossos Beach were categorized in higher height classes than individuals from the two other beaches (fig. 4). The mean occurrence depth of seahorses was of 141 ± 36 cm, constant throughout the study year. No correlation was found between the individual size and occurrence depth (p = 0.375, r = 0.256). The occurrence depth of the seahorses was significantly different between beaches (p < 0.001; F = 55.605), with João Fernandes showing the highest occurrence depths (fig. 5). The multiple comparison Tukey–Kramer analysis showed that seahorses found in significantly deeper locations at João Fernandes beach than at Ossos (p < 0.05; q = 14.185) and Canto (p < 0.05, q = 10.916) beaches. No significant difference was observed between occurrence depths between Canto and Ossos beaches (p > 0.05; q = 0.470). Spearman correlation showed no significant relation between depth and size (p = 0.168, r2 = 0.008).

Substrate and behavior Although we did not quantify this, we observed dominance in substrate availability for each site. Canto beach was dominated by algae and the gorgonian Phyllogorgia dilatata, while Ossos and João Fernandes beach were mainly composed of the sponge Aplysina fulva. Seahorses were recorded in Canto beach in association with seven different types of substrate, namely the algae Sargassum sp. C. Agardh, Amphiroa sp. (J. V. Lamouroux, 1812) and turf algae; the gorgonian Phyllogorgia dilatata (Esper 1806), the sponge Aplysina fulva (Pallas 1766), rocks and fishing nets. Of all the recorded substrates, the highest occurrence frequency was the seaweed Sargassum sp., representing 47 % of the records. Seahorses from João Fernandes Beach were observed associated with only three types of substrate, namely turf algae, the sponge Aplysina fulva, and the gorgonian Lepto-

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João Fernandes Ossos Canto

Number of seahorses

30 25 20 15 10 5

210–219

200–209

190–199

180–189

170–179

160–169

150–159

140–149

130–139

120–129

110–119

100–109

90–99

80–89

70–79

60–69

Height classes (mm) Fig. 4. Height classes (mm) of H. reidi seahorses at the three study areas João Fernandes, Ossos and Canto beaches, Armação de Búzios (RJ). Fig. 4. Clases de altura (mm) de los caballitos de mar H. reidi en las tres zonas del estudio, en las playas de João Fernandes, Ossos y Canto, en Armação de Búzios (RJ).

gorgia sp. Milne–Edwards, 1857. The substrate most used by H. reidi in this study area was the sponge A. fulva, with an occurrence frequency on this substrate of 92 % (n = 21). The other substrates accounted for only 4 % (n = 1) of occurrence each. Seahorses from Ossos Beach were found associated with five different types of substrate, with the Aplysina fulva sponge being that most frequently used, in 85 % (n = 33) of the total cases. Other substrates used were the bryozoan Schizoporella unicornis (Johnston in Wood 1844) (2 %; n = 1), the ascidian Phallusia nigra (Savigny 1816) (2 %; n = 1), seaweed Sargassum sp. (9 %, n = 3) and a submerged tree branch with epiphytes (2 %, n = 1) (fig. 6). The seahorses were found in association with a greater number of substrates at Canto Beach, followed by Ossos Beach and João Fernandes Beach. Of note was the relevance of the Aplysina fulva sponge as a substrate, recorded at all beaches. Behavioral diversity was greatest at Canto Beach, followed by João Fernandes and Ossos Beach (fig. 7). Despite greater behavioral diversity at Canto Beach, the most frequent behavior was resting at all beaches. The individuals showed the broadest behavior repertoire attached to the sponge Aplysina fulva and the seaweed Sargassum sp. (table 2). The other substrates were used basically to rest. Environmental characteristics The environmental variables at Canto Beach remained relatively stable over the study months,

except for transparency. The maximum recorded temperature was in March 2012, being 26 ºC, with a minimum of 20.2 ºC in May 2012 (fig. 8). Salinity showed subtle variations during the study period, with a recorded maximum of 40 in May and September 2012, and a minimum of 35 in November and December 2011, February to April 2012 and October 2012 (fig. 9). João Fernandes beach showed a normal curve, with peak temperatures in March and April 2012, and a gradual decrease in the other study months. The temperature varied by up to 7 ºC between months, with a maximum of 27 ºC in March/12 and a minimum of 20 ºC in October/12 (fig. 8). Salinity showed increasing values over the study months, with peaks at 40 in June and August 2012. Lower salinity values were observed from November/11 to April/12 (fig. 9). Ossos beach presented the warmest waters, from 24 to 25 ºC, from February to June 2012, and the coldest periods, 19 and 21 ºC, from November/11 to January/12 and from September to October 2012 (fig. 8). Salinity also presented marked periods, with lowest values, of 35 from November/11 to April/12, and highest values, ranging from 37 to 40, from May to October 2012 (fig. 9). Environmental variables were similar between areas, with no significant differences in temperature (p = 0.882; F = 0.125, df = 35) or salinity (p = 0.882; F = 0.125, df = 35). The environmental trends could not explain the variation in seahorse density variation at the three beaches (p = 0.743; R2 = 0.85; df = 71).

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351

João Fernandes

Ossos

Canto

300

Depth (m)

250 200 150 100 50 0

XI XII I II III IV V VI VII VIII IX X XI XII I II III IV V VI VII VIII IX X 2011

2012

2013

Fig. 5. Occurrence depth (m) of H. reidi seahorses at the three study areas João Fernandes, Ossos and Canto beaches, Armação de Búzios (RJ). Fig. 5. Profundidad de la presencia (m) de los caballitos de mar H. reidi en las tres zonas del estudio, en las playas de João Fernandes, Ossos y Canto, en Armação de Búzios (RJ).

100

João Fernandes

Ossos

Canto

70 60 50 40 30 20 10 0

5 6 Substrates

Fig. 6. Number of substrates used by H. reidi seahorses at the three study areas João Fernandes, Ossos and Canto beaches, Armação de Búzios (RJ): 1, Sargassum sp.; 2, Amphiroa sp.; 3, Turf algae; 4, Phillogorgia dilatata; 5, Aplysina fulva; 6, Schizoporella unicornis; 7, Phallusia nigra; 8, branch; 9, net gear; 10, rock. Fig. 6. Número de sustratos utilizados por los caballitos de mar H. reidi en las tres zonas del estudio, en las playas de João Fernandes, Ossos y Canto, en Armação de Búzios (RJ). (Para las abreviaturas de los sustratos, véanse arriba).

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100

Inactive behavior Foraging Swimming Courtship

Behavior pattern (%)

90 80 70 60 50 40 30 20 10 0

João Fernandes

Ossos

Canto

Fig. 7. Percentage of behaviors exhibited by H. reidi seahorses at the João Fernandes, Canto, and Ossos beaches, Armação de Búzios (RJ). Fig. 7. Porcentaje de comportamientos exhibidos por los caballitos de mar H. reidi en las playas de João Fernandes, Canto y Ossos, en Armação de Búzios (RJ).

Discussion Population densities at the three Armação de Búzios beaches were low when compared to H. reidi populations found in mangrove ecosystems in Northeastern Brazil (Pernambuco, Ceará, Piauí) (Silveira, 2005; Osório, 2008; Mai and Rosa, 2009; Aylesworth et al., 2015). Rosa et al. (2007) conducted a survey on the occurrence of H. reidi along the Brazilian coast, ob-

serving a total density of 0.026 ind/m2, similar to what was found herein. Freret–Meurer and Andreata (2008) found similar densities for this species on a rocky reef at Ilha Grande, Angra dos Reis, also Southeastern Rio de Janeiro, while Freret–Meurer (2010) reported densities of 0.003 ind/m2 for João Fernandes Beach and 0.001 ind/m2 for Geribá Beach in 2006. No seasonality has been identified for the three populations, as reported by Freret–Meurer and Andreata (2008).

Table 2. Frequency of occurrence behavior x holdfast: R, resting; S, swimming; C, courtship; F, foraging. Tabla 2. Frecuencia de comportamiento de presencia x substrato: R, descansando; S, nadando; C, cortejo; F, buscando comida.

R S C F

Turf algae

100 % 0 %

0 %

Protopalythoa sp.

100 % 0 %

0 %

Alga filamentosa

50 % 50 % 0 %

0 %

Phillogorgia sp.

83 %

17 %

0 %

Sargassum sp.

63 % 25 %

0 %

12 %

Schizoporella unicornis

Caulerpa racemosa 100 % 0 %

0 %

100 % 0 %

0 %

100 % 0 %

0 %

Dictyota sp.

100 % 0 %

0 %

Phallusia nigra

Amphiroa sp

50 %

0 %

50 %

Iron box

Padina sp.

100 % 0 %

0 %

(Anthopogenic waste) 100 % 0 %

0 %

Fishing net

100 % 0 %

0 %

100 % 0 %

0 %

100 % 0 %

0 %

Aplysina fulva

97 % 0.6 % 1.2 % 0.6 %

Millepora alcicornis 100 % 0 %

0 %

Branch

Palythoa sp.

0 %

Rock

100 % 0 %

Temperature (ºC)

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30 29 28 27 26 25 24 23 22 21 20 19 18 17 16 15

353

João Fernandes Ossos Canto

XI XII I II III IV V VI VII VIII IX X XI XII I II III IV V VI VII VIII IX X 2011

2012

2013

Fig. 8. Temperature (oC) recorded during the study at the João Fernandes, Ossos and Canto beaches, Armação de Búzios (RJ).

Salinity (ups)

Fig. 8. Temperatura (oC) registrada durante el estudio en las playas de João Fernandes, Ossos y Canto, en Armação de Búzios (RJ).

45 44 43 42 41 40 39 38 37 36 35 34 33 32 31 30 29 28 27 26

João Fernandes Ossos Canto

XI XII I II III IV V VI VII VIII IX X XI XII I II III IV V VI VII VIII IX X 2011

2012

2013

Fig. 9. Salinity (ups) recorded during the study at the João Fernandes, Ossos and Canto beaches, Armação de Búzios (RJ). Fig. 9. Salinidad (psu) registrada durante el estudio en las playas de João Fernandes, Ossos y Canto, en Armação de Búzios (RJ).

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Low densities are often found for several other seahorse species in other regions of the world, such as Hippocampus capensis in South Africa (Bell et al., 2003), Hippocampus comes in the Philippines (Perante et al., 2002; Morgan and Vincent, 2007) and Hippocampus whitei in Australia (Vincent et al., 2005). These densities are considered low and reflect the pressure suffered by these species over the years (Costa–Neto, 2000). Correia et al. (2015) also reported a variation of densities over time for H. guttulatus in Rio Formosa. Despite low densities, the populations studied herein remained stable throughout the study, with no major changes, except for João Fernandes Beach. Although seahorses were not marked, it has previously been reported that H. reidi has a home range exceeding 100 m2 (Freret–Meurer and Andreata, 2008), which suggests that seahorses herds move along the rocky reef. These movements could occur when males increase their home ranges, when they are not pregnant and become more active (Freret–Meurer et al, 2012). This could explain the density fluctuation, what may lead to misinterpretation of the population dynamics. These displacements are commonly observed (Freret–Meurer, pers. observ.) and can lead to erroneous inferences regarding the abundance of individuals in a given area. The distribution and abundance of some species of seahorse appeared to be related to the availability of habitat (Curtis and Vincent, 2005; Rosa et al., 2007; Aylesworth et al., 2015; Gristina et al., 2015; Harasti, 2016). The habitat structure and complexity is clearly important to maintain healthy populations of H. reidi and could influence their behavior. The holdfasts most commonly used by the seahorses in the present study were A. fulva and Sargassum sp. Both substrates apparently provide a good habitat for the seahorses to feed, protect and reproduce, particularly A. fulva. Other studies found seahorses resting on the substrate, a behavior also described for H. reidi (Rosa et al., 2007), H. abdominalis (Martin–Smith and Vincent, 2005) and H. capensis (Bell et al., 2003). This was also the most frequent behavior observed in our research. Hippocampus reidi is known to occur between depths of about 10 cm (Rosa et al., 2002) and 55 m Vari (1982), but we did not find any relation between size and depth. Although these data are not consistent with those of Dauwe (1992), who recorded larger individuals of this species in deeper water, they support findings of Oliveira and Freret–Meurer (2012), who neither found any relation between size and depth in a rocky reef in Arraial do Cabo, Rio de Janeiro. Population characteristics, such as sex ratios for juveniles and adults, were specific for each beach, indicating that each area may play an important role in maintaining the Armação de Búzios subpopulations. Canto Beach is directly influenced by currents that bring large amounts of solid waste to the area, with soda/beer cans and plastic bags constantly being found. João Fernandes and Ossos Beaches attract many swimmers and amateur divers who ignorantly trample the rocky shore to capture organisms out of

curiosity or to take pictures. The low occurrence of seahorses at the studied beaches increases the vulnerability of these animals to captures. Despite their cryptic nature, these are animals arouse interest when seen, making them frequent targets for curious people and the aquarium market. Besides tourists, fishermen are often seen stretching their nets along the rocky reefs, without any concern regarding specific animal populations in these areas. Seahorses from Armação de Búzios have a small population, although they reproduce all year long. This is of note in terms of conservation, because although they reproduce all year long, they are unable to increase their population density. This is probably due to a reef fish dispersal strategy. After birth, seahorses become planktonic for about 15 days (Foster and Vincent, 2004), being carried by currents and tides for kilometers before they settle. Studies about seahorse settlement are lacking, although it has been reported that the clownfish Amphiprion clarkia, which has a 7–11 day planktonic larval duration, is able to disperse 36 km (Green et al., 2014). The population font of A. clarkia may thus establsh other populations kilometers away. This dispersion is also likely among seahorses, so the stocks at our study site are possibly replenished from a nearby, non–studied population of seahorses. In spite of the studied populations of H. reidi are not driven by seasonality, stocks appear to remain stable. Temperature and salinity do not seem to have a major influence on density or the reproductive period. Each population has a particular structure and particularities should be considered for management plans. Although we did not record a decrease in seahorse abundance, the density was low. To conserve the small population from Búzios, we recommend controlled ecotourism on rocky reefs and projects of environmental education. Such actions could help the natural restoration of H. reidi populations in the area. Acknowledgements This research was supported by Projeto Coral Vivo and funded by Petrobrás Ambiental. The authors would like to thank the Universidade Santa Úrsula for the logistic support. References Altmann, J., 1974. Observational study of behavior: sampling methods. Behaviour, 49(3): 227–265. Aylesworth, L. A., Xavier, J. H., Oliveira, T. P. R., Tenorio, G. D., Diniz, A. F., Rosa IL., 2015. Regional–scale patterns of habitat preference for the seahorse Hippocampus reidi in the tropical estuarine environment. Aquatic Ecology, 49(4): 499–512. Bell, E. M., Lockyear, J. F., Mcpherson, A. D. M., Vincent, A. C. J., 2003. First field studies of an endangered south african seahorse Hippocampus capensis. Environmental Biology of Fishes, 67: 35–46.

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Correia, M., Caldwell, I., Koldewey, H., Andrade, J. P., Palma, J., 2015. Seahorse (Hippocampinae) population fluctuations in the Ria Formosa Lagoon, south Portugal. Journal of Fish Biology, 87: 679–690. Costa – Neto, E. M., 2000. Zootherapy based medicinal traditions in Brazil. Honey Bee, 11(2): 2 – 4. Curtis, J. M. R., Vincent, A. C. J., 2005. Distribution of sympatric seahorse species along a gradient of habitat complexity in a seagrass–dominated community. Marine Ecology Progress Series, 291: 81–91 Dauwe, B., 1992. Ecology of the seahorse Hippocampus reidi on the Bonaire coral reef (N.A.): habitat, reproduction and community interactions. MSc Dissertation, Rijksuniversiteit Groningen. Figueiredo, J. L., Menezes, N. A., 1980. Manual de peixes marinhos do sudeste do Brasil. III. Teleostei (2). Universidade de São Paulo, São Paulo. Foster, S. J., Vicent, A. C. J., 2004. Life history and ecology of seahorses: implications for conservation and management. Journal of Fish Biology, 65: 1–61. Freret–Meurer, N. V., 2010. Ecologia comportamental do cavalo–marinho brasileiro Hippocampus reidi Ginsburg, 1933 em recifes rochosos do estado do Rio de Janeiro. MSc Dissertation, Universidade do Estado do Rio de Janeiro. Freret–Meurer, N. V., Andreata, J. V., 2008. Field studies of a Brazilian seahorse population, Hippocampus reidi Ginsburg, 1933. Brazilian Archieves of Biology and Technology, 51(4): 743–751. Freret–Meurer, N. V., Andreata, J. V., Alves, M. A. S., 2012. Activity rate of the seahorse Hippocampus reidi Ginsburg, 1933 (Syngnathidae). Acta Ethologica, 15(2): 221–227. Green, A.. Maypa, A., Almany, G., Rhodes, K., Weeks, R., Abesamis, R., Gleason, M., Mumby, P., White, A., 2014. Larval dispersal and movement patterns of coral reef fishes, and implications for marine reserve network design. Biological Reviews, 90(4): 1215–1247, doi: 10.1111/brv.12155. Gristina, M., Cardone, F., Carlucci, R., Castellano, L., Passarelli, S., Corriero, G., 2015. Abundance, distribution and habitat preference of Hippocampus guttulatus and Hippocampus hippocampus in a semi–enclosed central Mediterranean marine area. Marine Ecology, 36: 57–66. Han, S.–Y., Kim, J.–K., Kai, Y., Senou, H., 2017. Seahorses of the Hippocampus coronatus complex: taxonomic revision, and description of Hippocampus haema, a new species from Korea and Japan (Teleostei, Syngnathidae). ZooKeys, 712: 113–139. Harasti, D., 2016. Declining seahorse populations linked to loss of essential marine habitats. Marine Ecology Progress Series, 546: 173–181. IUCN, 2010. Red List of Threatened Species [Internet]. Available from: http://www.iucnredlist.org [Accessed 16 Feb 2016]. Lourie, S. A., 2003. Measuring seahorses. Technical Report Series, 4: 15. Lourie, S. A., Foster, S. J., Cooper, E. W. T., Vincent, A. C. J., 2004. A Guide to the Identification of

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Microsatellite variability of the wood stork Mycteria americana (Aves, Ciconiidae) in Cuba: implications for its conservation A. Llanes–Quevedo, M. Alfonso González, R. Cárdenas Mena, C. Frankel, G. Espinosa Lopez

Llanes–Quevedo, A., Alfonso González, M., Cárdenas Mena, R., Frankel, C., Espinosa Lopez, G., 2018. Microsatellite variability of the wood stork Mycteria americana (Aves, Ciconidae) in Cuba: implications for its conservation. Animal Biodiversity and Conservation, 41.2: 357–364. Abstract Microsatellite variability of the wood stork Mycteria americana (Aves, Ciconiidae) in Cuba: implications for its conservation. Mycteria americana (Aves, Ciconiidae) is the only species of stork found in the Caribbean. It is a permanent yet rare resident in Cuba, with only two reproductively active colonies. In this work, we used five microsatellite loci to characterize 37 individuals from these colonies, located in two of the most important wetlands of Cuba, the Zapata Swamp and the Sabana–Camagüey Archipelago. We found low genetic variability, with similar values to those reported for North and South American populations of the species, and little but significant genetic differentiation between colonies. Our results highlight the need to improve the management and conservation planning of the species in Cuba because the combination of low genetic variation, small colonies, anthropogenic influence and climatic factors could threaten its persistence. Key words: Genetic diversity, Genetic structure, Waterbirds, Wetlands Resumen Variabilidad de los microsatélites de la cigüeña americana Mycteria americana (Aves, Ciconiidae) en Cuba: implicaciones para su conservación. Mycteria americana (Aves, Ciconiidae) es la única especie de cigüeña distribuida en el Caribe. En Cuba se considera residente permanente, pero rara, y solo se conocen dos colonias reproductivamente activas. En este trabajo se emplearon cinco loci de microsatélites para caracterizar genéticamente a 37 individuos de esas colonias, localizadas en dos de los humedales más importantes de Cuba: la ciénaga de Zapata y el archipiélago Sabana–Camagüey. Se observó una baja variabilidad en los índices de variabilidad genética, cuyos valores fueron similares a los referidos para las poblaciones de la especie en Norteamérica y Suramérica, y poca diferenciación genética entre las colonias que, sin embargo, era significativa. Nuestros resultados destacan la necesidad de mejorar la planificación del manejo y la conservación de la especie en Cuba debido a que la combinación de la baja variación genética, el pequeño tamaño de las colonias, la influencia humana y los factores climáticos podrían amenazar su persistencia. Palabras claves: Diversidad genética, Estructura genética, Aves acúaticas, Humedales Received: 03 XI 17; Conditional acceptance: 15 XII 17; Final acceptance: 19 I 18 Alexander Llanes–Quevedo, Massiel Alfonso González, Reinier Cárdenas Mena and Georgina Espinosa Lopez, Fac. de Biología, Univ. de La Habana, Cuba.– A. Llanes–Quevedo, Museo de Zoología "Alfonso Herrera", Fac. de Ciencias, Univ. Nacional Autónoma de México, México.– Carlos Frankel, Empresa Nacional para la Protección de la Flora y Fauna, Cuba.

ISSN: 1578–665 X eISSN: 2014–928 X

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Introduction Wetland birds face increasing threats due to factors such as the destruction and fragmentation of their habitats. Wetlands are among the most threatened ecosystems in the world due to accelerated climate change, the anthropogenic impact caused by expanding coastal development, agricultural land conversion, and urbanization (Junk et al., 2013). Waterbirds occupying islands could be especially susceptible due to their relative isolation, comparatively small population size, and the effects of climate change such as sea level rise. Mycteria americana Linnaeus, 1758 (Aves, Ciconiidae), the wood stork, is one of the most conspicuous and rare aquatic species found in the Cuban archipelago, being the only stork species present in the Caribbean (Raffaele at al., 2010). It is widely distributed from the southeastern of the United States to the north of Argentina (Coulter et al., 1999). In the Caribbean islands, it resides permanently only in Cuba and the Dominican Republic (Latta et al., 2010; Garrido and Kirckonnell, 2010), while in Jamaica, Bahamas and Dominica it is classified as a casual visitor (Latta et al., 2010). In Cuba, its distribution includes the southern coast of Pinar del Río, the Zapata Swamp, the Sabana–Camagüey Archipelago and the Birama Swamp according to Garrido and Kirkconnell (2010). However, only two nesting sites have been identified and described to date. Historically, these sites have low numbers of breeding pairs, always fewer than 50 pairs, and despite their legal protection, the number of colonies and nests has progressively decreased. In 2015, for example, the largest colony in the Sabana–Camagüey Archipelago contained only 26 nests, while there were only 8 nests at Las Salinas in the Zapata Swamp (Llanes–Quevedo et al., 2015). The species does not exhibit migratory habits and has a tendency to use the same reproductive sites over the years (Frederick and Ogden, 1997). In general, movements of individuals or colonies seem to be a response to the fluctuation of environmental conditions and successive failed attempts at reproduction at the sites (Coulter et al., 1999). Several studies have reported the dispersal movements of the species, but these are restricted to specific areas of occurrence, such as the United States (Hylton, 2004; Borkhataria, 2009), Argentina and Brazil (Antas, 1994; Del Lama et al., 2015). The entire range of the species has not been studied to date. However, the movements of wood storks are supposedly limited in the Carribean by the absence of thermals that birds require to fly across long distances (Coulter et al., 1999). The wood stork is considered as Least Concern according to IUCN (BirdLife International, 2016). Nevertheless, throughout its distribution, it has been affected by the loss and alteration of foraging and reproductive sites. Habitat loss and degradation due to anthropogenic activities such as occupation of areas, alteration of drainage for agriculture, and tourism have already led to declines in population numbers of this species in various regions. For example, in the United States, populations of over 100,000 individuals were reduced to about 3,000 between the

1960s and 1970s, (Coulter et al., 1999), and as a result, the species was classified as Endangered. In the Caribbean, wood stork populations have been classified as locally endangered and even as extirpated in the Dominican Republic (Raffaele at al., 2010; Latta et al., 2010), while in Cuba, although it is considered a rare species that occupies vulnerable habitats, it does not present any category of threat. Nevertheless, M. americana colonies in Cuba face considerable and increasing threats, such as local hunting and tourism development. These elements, together with the reduced census size of the reproductive colonies (Llanes–Quevedo et al., 2015), the regular occurrence of severe hydro–meteorological phenomena, and the lack of studies enabling development of adequate management measures for the species, may jeopardize the future of M. americana in the country. Hence, the goal of the present study was to examine levels of genetic diversity, inbreeding and population differentiation in breeding colonies of M. americana in Cuba, using microsatellite markers. Knowing the genetic characteristics of the species in Cuba would provide valuable information to design and improve management plans for its conservation. Material and methods Sample collection Sampling was conducted in M. americana breeding colonies of in Cayo Romano, Camagüey province, and Las Salinas in the Zapata Swamp, Matanzas province in the reproductive season 2015–2016 (fig. 1). These breeding sites, included in areas classified as Wildlife Refuges, are located in two of the most important wetlands in Cuba: the Zapata Swamp and Sabana– Camagüey Archipelago (Denis, 2006). In general, the nesting areas are similar in terms of vegetation, soil types, temperature and average rainfall; they are separated by approximately 400 km (Llanes–Quevedo et al., 2015). Samples consisted of feathers taken from chicks, one or two per nest: from 25 chicks from 17 nests in Cayo Romano colony, and from 12 chicks from 8 nests in Las Salinas colony. Chicks were captured and handled for a short time. Capture and sampling was performed following the Ornithological Council Guidelines for Ethical Animal Experimentation (Fair et al., 2010) and approved by the Cuban authorities from the Center for Environment Inspection and Control (CICA permits of access and collection in protected natural areas n° 2015/21 and 2015/87). The feathers were plucked from the dorsum and the belly. The injuries produced were treated with alcohol to prevent infection once the sample was obtained and the chicks were released at the capture site. Feathers were deposited in a 2–mL Eppendorf tube containing 90 % ethanol. DNA extraction and amplification of molecular markers Total genomic DNA was extracted from 10 mg of plucked feathers with a Nucleo Spin Tissue Kit (Ma-

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N 1 20 km

20 km

100

200 km

Fig. 1. Map of the Cuban Archipelago showing collection sites of Mycteria americana in the reproductive season 2015–2016: 1, Cayo Romano breeding colony, Cayo Romano, Camagüey (n = 25); and 2, Zapata Swamp breeding colony, Zapata Swamp, Matanzas (n = 12). Fig. 1. Mapa del archipiélago cubano en el que se muestran los sitios de recolección de Mycteria americana en la temporada reproductiva de 2015–2016: 1, colonia de reproducción Cayo Romano, en cayo Romano, Camagüey (n = 25); y 2, colonia de reproducción Zapata Swamp, en la ciénaga de Zapata, Matanzas (n = 12).

cherey–Nagel, Dün, Germany). DNA was diluted in 50 µl of sterile water, and the quality and amount of DNA were estimated by electrophoresis (agarose 1 %). The five most polymorphic microsatellite loci described by Tomasulo–Seccomandi et al. (2003) were used to genotype samples: WSµ03, WSµ08, WSµ09, WSµ14 and WSµ20. PCRs were performed in 15 µl of mix containing 1x buffer (2 mM MgCl2), 10 pM of each primer, 0.5 mM of each dNTP, one unit of Taq polymerase, and 50–150 ng of template DNA. PCR consisted of the following steps: 1) initial denaturing, 5 min at 94 °C; 2) 30 cycles of 35 sec of denaturing at 94 °C, 60 sec of annealing at the appropriate temperature (50–60 °C) and 1 min of extension at 72 °C; and 3) 10 min of extension at 72 °C. DNA fragments were separated by vertical polyacrylamide gel electrophoresis for 2 hr at 65 W using a Sequi–Gen GT sequencing cell. Genotypes were determined visually and to minimize error in alelle calling, at least two people performed the readings. In addition, to control for genotyping error, we re–amplified all microsatellite loci for 20 % of the samples. Allele sizes were estimated by comparison with the pGem molecular weight markers and GenePrint CTT Multiplex and GenePrint FFV Multiplex Systems (Promega, 2006). Genetic diversity We analysed genotypes for null allele with Micro– Checker 2.2.3 (Van Oosterhout et al., 2004). We

considered polymorphic all loci with at least two alleles, when the frequency of the most common allele did not exceed 95 % (Graur and Li, 2002). We calculated the standard measures of genetic variation for microsatellite loci: number of alleles per locus, number of private alleles, effective number of alleles, observed heterozygosity, expected heterozygosity, and inbreeding coefficients with Genalex v6.5 (Peakall and Smouse, 2012) and Arlequin v3.5 (Excoffier and Lischer, 2010) for each colony. Allelic richness was valued in FSTAT v2.9.3.2 (Goudet, 2001). The effective population size was estimated based on the sibship assignment method described by Wang (2009) and implemented in Colony v2.0 (Jones and Wang, 2010). We performed the linkage disequilibrium and Hardy–Weinberg equilibrium analyses with GenePop v4.1.0 (Rousset, 2008), according to the Markov chain method with 1,000 dememorizations, 1,000 batches and 10,000 iterations. Markers were tested for neutrality using the 'detsel' R package v1.0.2 (Vitalis, 2012). Genetic structure To assess the level of genetic differentiation between colonies, we computed the FST index of Weir and Cockerham (1984). The variance components in allele frequencies between colonies and individuals were calculated using a hierarchical AMOVA implemented in Arlequin v3.5 (Excoffier and Lischer, 2010).

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Genetic structure was was further tested using Bayesian clustering of the microsatellites data in Structure v.2.3.4 (Pritchard et al., 2000). The running parameters were no a priori information on sampling location, an admixture model, and correlated allele frequencies, λ of one. Run length was set to 1,000,000 MCMC (Markov chain Monte Carlo) replicates after a burn–in period of 500,000 and the number of clusters (K) from 1 to 4, with 20 replicates for each value of K. To select the most probable K number, we examined the obtained likelihood values for each run and the variance between these values for each K value. We selected the value with the best posterior probability and the smallest variance between repetitions (Pritchard et al., 2000) as that which best represented the number of population clusters present in the sample.

(AMOVA) we determined that 93.1 % of the variation was explained by differences between individuals within the localities, while the remaining 6.98 % was explained by differences between localities. On the other hand, the Bayesian clustering analysis showed that, for K = 2, all individuals had equal probability (50 %) of belonging to any group (fig. 2A). A single population cluster seems to best explain the variation in allele frequencies observed in the wood storks in Cuba, as K = 1 had the highest logarithm of the likelihood (close to zero) and also the lower variance (fig. 2B).

Results

In the present study, we used five autosomal microsatellite loci for genetic characterization of the two breeding colonies of M. americana in Cuba. Both colonies showed all polymorphic loci, but the number of alleles was low, coinciding with that reported for two populations, one from Fazenda Ipiranga, Mato Grosso, Brazil and one from Tamiami West, Everglades, Florida, USA (Tomasulo–Seccomandi et al., 2003), and also with later studies comprising several colonies from Pantanal (Lopes et al., 2007) and Amapa region (Miño et al., 2011) in Brazil. As already proposed by other authors, the low polymorphism found in the present study seems to be intrinsic to the species (Lopes et al., 2006; Miño et al., 2017). Low genetic diversity of M. americana has also been detected through mtDNA sequences (Lopes et al., 2006, 2007), other microsatellite loci (Van Den Bussche et al., 1999), and alloenzyme analyses (Strangel et al., 1990). According to Eo (2011), effective population sizes and ecological and environmental features of aquatic birds produce (compared with other types of birds) low rates of molecular evolution. Although it is not a universal pattern, genetic diversity is commonly reduced in peripheral populations (Volis et al., 2016). These populations differ from central populations mainly because of genetic drift. As a consequence of the random sampling in reproduction, there is usually a loss, by chance, of less frequent alleles and a fixation of the most frequent alleles, thus reducing the genetic diversity of the population. The two Cuban colonies could be considered peripheral because they are located at one of the extremes of the species distribution and their number of breeding pairs is low. However, they present similar and even higher genetic diversity indexes than those of the North American and Brazilian populations; that is, the discrete genetic variation found for M. americana in Cuba is not significantly lower than that reported in continental populations. In this sense, our results suggest that Cuban colonies may remain connected, to some extent, with those of North or South America through the exchange of individuals and genes, although this theory should be complemented by future studies using banding and mitochondrial markers.

Genetic diversity The genetic characterization of breeding colonies of M. americana in Cuba showed that the five autosomal microsatellite loci were polymorphic but with a low number of alleles (between 2 and 3). The total number of alleles was 13 with an average of 2.60 alleles per locus. The genotypic proportions of all loci evaluated were in agreement with expectations under Hardy Weinberg equilibrium, except for WS08 in the Cayo Romano colony. No evidence of non–random association was found between segregating alleles at any loci–pair (linkage disequilibrium); there was no evidence of null alleles or deviation from neutrality in any of the studied samples. The mean number of effective alleles (Ne) was 2.11, with locus WS03 contributing the most (at 2.75) and WS20 the least (1.65). The percentage of Ne per locus was greater than 50 % at all loci. The average value of allelic richness (Rs) was 2.49; WS03 was the most diverse locus with 3 alleles, while WS09 and WS20 presented the lowest value with 2 alleles. Heterozygosity (Ho) was less than 0.5 in all loci, with a minimum value of 0.25 for locus WS20 and a maximum of 0.48 for WS03.The WS03 locus showed the highest expected heterozygosity (He = 0.63) and WS20 showed the lowest (He = 0.38). No rare alleles were found (frequency ≤ 5 %) and only one private allele was detected for the WS14 locus in the Cayo Romano colony. Cayo Romano also showed slightly higher genetic variability indexes, but this difference was not statistically significant (table 1). The average inbreeding coefficient was 0.28, with individual inbreeding coefficients ranging from 0.26 (Cayo Romano colony) to 0.31 (Zapata Swamp colony). Effective population size was 16 (95 % confidence interval, CI: 8 to 34) for Cayo Romano and 9 (95 % CI: 4 to 26) for Zapata Swamp. Genetic structure The FST statistic between the two colonies was 0.06 (p = 0.03). From the analysis of molecular variance

Discussion Genetic diversity

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100 %

80 %

40 % 20 %

K=4

–200 –300 –400

Ma01 Ma03 Ma05 Ma07 Ma10 Ma12 Ma15 Ma18 Ma20 Ma22 Ma24 Ma26 Ma29 Ma40 Ma42 Ma44 Ma46 Ma48 Ma51

0 %

K=3

–100 Ln (P)

60 %

K=1 K=2

Fig. 2. A, Probability of assignment of Mycteria americana individuals to each colony estimated by Structure v.2.3, based on allele frequencies. Vertical bars represent the individuals and are divided into segments corresponding to the probability of assignment to Cayo Romano (gray) or Zapata Swamp (black). B, log of the likelihood value versus most probable number of populations (K) obtained with Structure Fig. 2. A, Probabilidad de asignación de individuos de Mycteria americana a cada colonia, estimada por Structure v.2.3 a partir de frecuencias alélicas. Las barras representan a los individuos y se dividen en dos segmentos que corresponden a la probabilidad de asignación a Cayo Romano (gris) y a Zapata Swamp (negro). B, logaritmo del valor de la probabilidad con respecto de el número más probable de poblaciones (K) obtenido con Structure.

The colonies analyzed in the present work showed low estimates of effective population size and positive and high values of the inbreeding coefficient. These elements constitute risks for their survival in the future because they indicate that the number of reproductive individuals contributing demographically and genetically to the next generation is small (Hedrick, 2000). The reduced effective polpulation size of M. americana colonies was similar to that obtained by Miño et al. (2017) for the Pantanal and Amapá regions in Brazil; these authors found lower and, in some cases, negative Fis values. In the present work, we found that the differences between colonies in the genetic diversity indexes were minimal. Nevertheless, all indices were slightly higher in the Cayo Romano colony (table 1). Regarding the higher number of alleles, this could be due to the larger sample size taken in this colony (N = 25). For the other indexes, the higher indices could be associated with the genetic health status of this larger nesting colony in Cuba (Cayo Romano) being better than that of the colony in the Zapata Swamp, which, according to nest counts from 1987 to the present, is declining (Llanes et al., 2015). Our findings could support the urgent management and protection of the species in Cuba where, the Zapata Swamp colony is likely to be impacted in the near future by habitat loss, saline inclusion and coastal erosion (Moya et al., 2005). In contrast, the nesting area of M. americana in Cayo Romano is currently classified as medium–low risk (Menéndez Carrera et al., 2015).

Genetic structure The value of FST obtained in the present study (FST = 0.06; p = 0.03) suggests genetic differentiation between the two colonies studied is low, a finding that does not agree with the results obtained by Structure. The significant genetic structuring between colonies of M. americana was only recently documented, by Miño et al. (2017), in their comparison of colonies of Pantanal and Amapá. In the remaining studies where this statistic has been used, no evidence of genetic structuring has been found between wood stork colonies in the Pantanal region (Del Lama et al., 2002; Lopes et al., 2004, 2006), in the Everglades (Stangel et al., 1990; Van Den Bussche et al., 1999), or in North and South America. However, the FST values reported in these studies were calculated for allozymes and microsatellites (described by Van Den Bussche, 1999) that are less variable than those used in the present study. The Cuban breeding colonies we studied are separated by approximately 400 km, but this does not seem to be an important limitation for individual interchange due to the species' great flight capacity. Studies in Florida have reported that wood storks can travel more than 100 km per day in their dispersal movements between winter and summer residence sites (Hylton, 2004). This incipient genetic differentiation thus suggests some degree of site fidelity, a behavior reported in the literature based on ecological (Frederick and Odgen, 1997) and genetic

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Table 1. Estimates of genetic diversity of Mycteria americana breeding colonies in Cuba based on five microsatellites loci: Na, number of alleles; Ne, number of effective alleles; Ho, observed heterozygosities; He, expected heterozygosities; and Rs, allelic richness. Tabla 1. Estimaciones de la diversidad genética de las colonias de reproducción de Mycteria americana en Cuba a partir de cinco loci de microsatélites: Na, número de alelos; Ne, número de alelos efectivos, Ho, heterocigosis observada; He, heterocigosis esperada; Rs, riqueza alélica.

Pop

WSµ03 WSµ08 WSµ09 WSµ14 WSµ20 Mean

Na Cayo Romano 3.00 3.00 2.00 3.00 2.00 2.60

Zapata Swamp 3.00 3.00 2.00 2.00 2.00 2.40

Ne Cayo Romano 2.98 2.21 1.70 2.85 1.75 2.30

Zapata Swamp 2.53 1.72 2.00 1.86 1.55 1.93

Ho Cayo Romano 0.46 0.40 0.34 0.49 0.30 0.40

Zapata Swamp 0.50 0.36 0.20 0.36 0.20 0.33

He Cayo Romano 0.66 0.55 0.41 0.64 0.42 0.54

Zapata Swamp 0.61 0.42 0.50 0.46 0.33 0.46

RS Cayo Romano 3.00 2.96 2.00 2.99 2.00 2.59

Zapata Swamp 3.00 2.90 2.00 2.00 2.00 2.38

evidence (Miño et al., 2017), although this was not recognized in previous works with molecular markers (Del Lama et al., 2002, Lopes et al., 2004, 2006). Implications for conservation The markers used in the present work are neutral and do not therefore allow direct evaluation of the diversity of genes that confer adaptation to current environmental conditions. Nevertheless, they provide data to make inferences regarding important variations that should be maintained as a source for natural selection to act upon when responding to future environmental changes. Although there is wide debate on the subject (Reed and Frankham, 2001; Volis et al., 2016), numerous studies have reported a correlation between the diversity of neutral markers and that of the loci subjected to selection (e.g. Vandewoestijne et al., 2008; Da Silva, 2006). Low genetic diversity associated with a loss of adaptive potential has been identified as one of the most important risks for species persistence and its importance goes beyond its effects on population dynamics, to affect the structure of communities and ecosystem processes (Hughes et al., 2008). The low genetic variability found for M. americana in Cuba coincides with that reported in the literature for the North and South American populations of the species. However, Cuban colonies are smaller, creating an extinction risk (Gilpin and Soule, 1986) because of the increased potential for inbreeding (Blomqvist et al., 2010). This situation is further aggravated by the negative influence of anthropogenic activities on reproductive colonies and

the use of chicks as food by local people. All these factors could cause the Cuban colonies to enter a vortex of extinction, that is, create negative dynamics that would increase the probability of extinction due to stochastic or catastrophic demographic events (Gilpin and Soulé, 1986). At this point, we emphazise the need to study fitness variables, given that the variation in quantitative traits is a better indicator of their evolutionary potential than neutral genetic variation (Navarro et al., 2005, Volis et al., 2016). These studies would be useful to assess the real adaptive potential of these colonies and allow more accurate decisions regarding conservation priorities. Besides the conservation importance of the Cuban colonies of the wood stork in view of their current risky demographic conditions, they could theoretically be of great importance from an evolutionary standpoint. Peripheral and central ones populations tend to differ because of the combined effect of genetic drift and natural selection, and the less predictable effects of the latter on genetic variation. Marginal populations, i.e., those distributed at the extremes of the range of occurrence of species, often occupy lower quality habitats and undergo different selection pressures, which may lead to the emergence of important new allele variants or may constitute a source potential of speciation episodes (Lesica and Allendorf, 1995). Our findings highlight the importance of planning and implementing measures to ensure effective and lasting protection of the species in Cuba. An essential part of these plans should be the conservation of the habitats that the species occupies, particularly the

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wetlands of northern Camagüey, which harbor the most current genetic diversity of the species in Cuba. To effectively conserve this area it is necessary to implement measures that allow the restoration of its ecosystems, mainly affected by the increased building development due to the expansion of human settlement and tourism. Mangroves and seagrasses in this area are already affected by increased turbidity, sediment suspension and hypersalination caused by the obstruction of channels with structures built in or between the cays (Iturralde–Vinent and Serrano, 2015). Special emphasis should also be given to the need for environmental education in order to prevent wood stork consumption by local people. Excessive hunting has already decimated the colonies and presumably led to the extinction of the species in Hispaniola (Latta et al., 2010). Although potentially more complex, ex–situ management of wood storks could also be considered in an attempt to alleviate the effect of human impact, at least while chicks are in the nest. Acknowledgements The authors thank all those who collaborated with this work. In particular, we thank Juan Escalona, Etiam Pérez, José M. Barrios, Lázaro Sosa, Idelfonso Bonachea and Ricardo Rodríguez for their support during the fieldwork, and Fermin Amaro and Dayana Ontivero for their valuable help in laboratory work. We also want to thank Carolina Miño and the editor of the Journal for the manuscript revision and the valuable suggestions. Finally, we thank IdeaWild organization for providing the field equipment used to carry out this work. We would also like to thank George Amato, Director of Conservation Genomics of the Sackler Institute for Comparative Genomics, American Museum of Natural History Central for the help provided with materials and reagents for the laboratory. References Antas, P. T. Z., 1994. Migration and other movements among the lower Parana River valley wetlands, Argentina, and the south Brazil/Pantanal wetlands. Bird Conservation International, 4: 181–190. BirdLife International, 2016. Mycteria americana. The IUCN Red List of Threatened Species 2016: e.T22697648A93627312, doi: 10.2305/IUCN. UK.2016–3.RLTS.T22697648A93627312.en. Blomqvist, D., Pauliny, A., Larsson, M., Flodin, L. Å, 2010. Trapped in the extinction vortex? Strong genetic effects in a declining vertebrate population. BMC Evolutionary Biology, 10(1): 33. Borkhataria, R., 2009. Modeling population viability and habitat suitability for the endangered Wood Stork (Mycteria americana) in the southeastern United States. Tesis doctoral, University of Florida. Coulter, M. C., Rodgers, J. A., Ogden, J. C., Depkin, F. C., 1999. Wood Stork (Mycteria americana). In: The Birds of North America: 1–28 (A. Poole and F. Gill, Eds.). The Birds of North America, Inc.,

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The mysterious bird outbreak of 1779 in southeastern Iberian peninsula: a massive irruption of the Spanish sparrow Passer hispaniolensis from Africa? J. J. Ferrero–García, L. M. Torres–Vila, P. P. Bueno

Ferrero–García, J. J., Torres–Vila, L. M., Bueno, P. P., 2018. The mysterious bird outbreak of 1779 in southeastern Iberian peninsula: a massive irruption of the Spanish sparrow Passer hispaniolensis from Africa? Animal Biodiversity and Conservation, 41.2: 365–377. Abstract The mysterious bird outbreak of 1779 in southeastern Iberian peninsula: a massive irruption of the Spanish sparrow Passer hispaniolensis from Africa? Several current and past bibliographical references mention the sudden pest outbreak of a mysterious sparrow–like bird in the southeastern Iberian peninsula in 1779. Based on these references, we investigated unpublished documentary sources from various historical archives that reflected the actions carried out by public authorities against the bird pest. Some narratives come from direct witnesses who sometimes provided relevant data on the origin and biology of the birds involved. From the analysis and interpretation of these data, it was clear that the bird outbreak was caused by an unusual passerine in southeastern Iberia. In May 1779, birds irrupted in large numbers into several localities in the current provinces of Alicante, Murcia and Almería, probably coming from North Africa. Damage caused to cereal crops was meaningful and the extraordinary alarm generated in the people motivated the intervention of both local authorities and government institutions. The birds formed large arboreal colonies, building multiple nests per tree. We discuss different hypotheses related to the taxonomic position of these birds within the Ploceidae and Passeridae families. The bird species whose distribution, morphology, life characteristics and behaviour agrees best with the testimonies analysed is the Spanish sparrow Passer hispaniolensis. We propose that this sparrow could be the protagonist of this historic bird pest outbreak. Key words: Bird pests, Passerines, 18th century, Historical archives, North Africa, Southeastern Spain Resumen La misteriosa plaga de aves de 1779 en el sureste de la península ibérica: ¿una irrupción masiva del gorrión moruno Passer hispaniolensis desde África? Algunas referencias bibliográficas, actuales y pasadas, señalan la abrupta llegada de una misteriosa plaga de aves, parecidas a los gorriones, al sureste de la península ibérica en 1779. Partiendo de dichas referencias, en este estudio hemos analizado en fuentes documentales inéditas de distintos archivos históricos, donde se recogen las medidas que adoptaron diferentes autoridades contra esta plaga. Algunos de estos relatos provienen de testigos directos, que a veces aportaron datos de interés sobre la biología de las aves y de su procedencia. Del análisis de esos datos se desprende que esta plaga fue provocada por un paseriforme no habitual en el sureste de la península ibérica. Las aves irrumpieron en gran número en mayo de 1779 en algunas localidades de las actuales provincias de Alicante, Murcia y Almería, probablemente procedentes del norte de África. Los daños causados en los cultivos de cereal debieron ser significativos y la extraordinaria alarma que suscitó en la población motivó la intervención tanto de autoridades locales como de instituciones estatales. Las aves formaron grandes colonias en árboles y construyeron muchos nidos por árbol. Planteamos diferentes hipótesis en relación con la identificación taxonómica de estas aves, dentro de las familias Ploceidae y Passeridae. Finalmente sugerimos que la especie cuya distribución, morfología, características y comportamiento concuerda mejor con los testimonios analizados es el gorrión moruno Passer hispaniolensis, por lo que creemos que pudo ser el protagonista de esta plaga histórica. Palabras clave: Plagas de aves, Paseriformes, Siglo XVIII, Archivos históricos, Norte de África, Sureste de España ISSN: 1578–665 X eISSN: 2014–928 X

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Received: 6 IX 17; Conditional acceptance: 19 XII 17; Final acceptance: 22 I 18 Juan J. Ferrero–García, Servicio de Calidad Agropecuaria y Alimentaria, Consejería de Medio Ambiente y Rural PAyT, Junta de Extremadura. Avda. Luis Ramallo s/n, 06800 Mérida, Spain.– Luis M. Torres–Vila, Pedro P. Bueno, Servicio de Sanidad Vegetal, Consejería de Medio Ambiente y Rural PAyT, Junta de Extremadura. Avda. Luis Ramallo s/n, 06800 Mérida, Spain. Corresponding author: Juan J. Ferrero–García. E–mail: juanjose.ferrerog@juntaex.es

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Introduction It is well–known that some bird species can cause extensive damage to agricultural crops, as reported from many parts of the world (Jones, 1972; Pinowski and Kendeigh, 1977; De Gracio, 1978; Wright et al., 1980; Bruggers and Ruelle, 1981; Bellatreche, 1986; Dhindsa and Saini, 1994; Huber et al., 2002; Contreras et al., 2003; Tracey et al., 2007; Behidj–Benyounes et al., 2011; De Melo and Cheschini, 2012; De Mey et al., 2012; Canavelli et al., 2014; Codesido et al., 2015; De Rijk, 2015; Loza–Del Carpio et al., 2016). Based on records of historical archives, we recently investigated the incidence and spatio–temporal variation of bird pests in the region of Extremadura (southwestern Spain) over four centuries (1501–1900) (Torres–Vila et al., 2009; Ferrero–García et al., 2014, 2016; Torres–Vila et al., 2015). Our research showed that passerines, particularly sparrows (house sparrow Passer domesticus and/or Spanish sparrow Passer hispaniolensis), were the birds that caused most harm to Spanish agriculture in the past. Our results provided a historical perspective of both the impact of the birds and the changing human perception towards them. Currently, the house sparrow is widely distributed in the Iberian peninsula, while the Spanish sparrow occurs mainly in southwestern Iberia (Bernis, 1989; Tellería et al., 1999; SEO/BirdLife, 2012; De Juana and Garcia, 2015). The past distribution of these two species has been known with relative precision since the 19th century and was broadly similar to the current distribution (De Juana and Garcia, 2015). In Spain, the house sparrow is basically sedentary and nests in rocks, trees and buildings, while the Spanish sparrow is partially migratory and mostly breeds in trees and shrubs, sometimes forming large colonies (Bernis, 1989; Tellería et al., 1999; De Juana and Garcia, 2015). In one of the above–cited works (Ferrero–García et al., 2014) we found that several municipalities in Extremadura echoed an Order of the Council of Castile, dated 19 October 1779, concerning a plague of 'overseas' birds that had arrived in Spain that same year. The Order mentioned that these birds had irrupted into southeastern Spain (provinces of Almería and Murcia) in May 1779. We hypothesized that it was likely the same bird species that, also in May 1779, struck Totana (Murcia), a crucial matter about which the Count of Floridablanca (probably the most relevant statesman in those years) wrote to Pedro Franco Dávila, Director of the Royal Cabinet of Natural History. Ferrero–García et al. (2014) stressed that the bird outbreak had such repercussion that its occurrence was announced in the Gaceta de Madrid (1779), the official Spanish bulletin of the epoch, where the birds were described as similar to sparrows ('páxaros como especie de gorriones'). However, the study focused on the region of Extremadura and did not investigate further into the reports or into the species of birds involved. Previous references discussed this mysterious bird outbreak, but none included strictly scientific background. García–Hourcade (1996) mentioned that in 1779 there was a plague of 'African sparrows' in

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Cartagena (Murcia), referring to an older reference mentioning the devastation produced that year in Cartagena by the same 'African sparrows' (Díaz, 1895). García–Abellán (1975) quoted the plague of sparrows of 1779 as one of the most remarkable in the region of Murcia in the 18th century. In a specific paper published in a local magazine (in which the Order of 19 October 1779 is reproduced) Grima (1990) remarked: 1) that the denomination of these birds as 'overseas' itself suggests that they came from the other side of the sea; 2) that birds settled initially in Cartagena (Murcia), from where they moved to Lorca (Murcia) and then to Vera (Almería); and 3) that the birds were the size of a sparrow. Unfortunately, Grima (1990) did not conduct an in–depth study and did not reach any conclusion about the birds that starred in the pest outbreak of 1779. In this paper we aimed to perform a more rigorous analysis of the available documentary sources (especially of the unpublished documents from different historical archives) in relation to the bird outbreak of 1779. Our objectives were: 1) to systematically collect and analyse the available information; 2) to find out if the described bird pest was caused by a rare or occasional species (in the sense of 'rarities'; see e.g., De Juana, 2006) of unusual presence in southeastern Spain; 3) to specify the territorial scope that birds came to encompass as well as to infer the lasting time; 4) to assess whether birds had a major impact on agricultural crops; and 5) to infer the taxonomic position of these birds. In short, we aimed to investigate a bird outbreak that had a notable social impact, not only in the affected localities, but also in other places in Spain. Such an historical event has been mentioned by several authors over the years, and in varied contexts, although always in a collateral, marginal or anecdotal manner, and often within studies with objectives other than wildlife research. Material and methods We consulted the historical and municipal archives in the cities and towns mentioned in the bibliographical references we found on the bird plague of 1779, as well as those archives in the localities suspected of having been affected. Documents were in paper format but digital reproductions were also used when available (e.g., AHRM, 2017). Target documents were read in full searching for and collecting all information relevant to the goals of this study. To assess the importance of the testimonies contained in the studied documents, it is important to note that several Spanish authorities from the period (during the reign of Carlos III) of different rank/status are named. The most relevant authorities (Gómez, 2006) were: 1) the Corregidor, the representative of the king in a municipality at that time, responsible for judicial and police matters; 2) the Major Mayor, a lawyer appointed by the monarch and the highest authority in municipalities that did not have a Corregidor or a Corregidor´s legal assistant; 3) the Governor of a city, a Corregidor to whom the military

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government was entrusted; and 4) the Regidor, the Trustee ('Procurador Síndico') and the Deputy of the Common, other local authorities. The historical archives consulted and their corresponding documents are as follows: The Books of Agreements (Municipal Archives) A large part of the documentary sources studied included the Books of Agreements (BAs thereafter) from several municipal archives. The BAs contain the agreements (organized per session) adopted by municipalities regarding matters within their legislative, executive and judicial jurisdiction (Torres–Vila et al., 2015). BAs constitute the most relevant documentary series that the Spanish municipalities have produced, and they date back back to the Late Middle Ages (García–Ruipérez, 2008). As emphasized in previous researches (Ferrero–García et al., 2014, 2016; Torres– Vila et al., 2015), aspects reflected in the BAs include matters related to agricultural pests (including birds), and the human (and divine) actions taken to combat these. We consulted the BAs of Vera (Almería) and Lorca (Murcia) where the bird pest occurred according to the Order of 19 October 1779. In this Order it is also stated that birds colonised Mojácar and Turre (Almería), but BAs of these municipalities before the 19th century are not preserved. We also consulted the BAs of Totana (Murcia) where the bird pest was referred to both in its time (Gaceta de Madrid, 1779) and today (Calatayud, 1987; Velasco–Pérez, 2012). Lastly, we checked the BAs of Cartagena (Murcia), where the occurrence of the bird pest in 1779 is referred by several authors (Díaz, 1895; García–Hourcade, 1996). This research sometimes motivated the consultation of further BAs from other municipalities, which were mentioned in the sources investigated. This was the case of Orihuela (Alicante) and city of Murcia. In short, we searched the archives in six municipalities now belonging to three provinces. Note that the current provincial distribution in Spain dates from 1833. In the Results section we show the findings: Document 1 (BA of Orihuela), Document 2 (BA of Murcia), Document 3 (BA of Cartagena) and Document 4 (BA of Lorca). File 558 (Archive of the Museo Nacional de Ciencias Naturales [MNCN] of Madrid) The existence of File 558 of the Document Catalogue from the Royal Cabinet of Natural History (1752–1786) has been cited previously (Calatayud, 1987; Velasco–Pérez, 2012) but these papers did not deepen in their content. File 558 consists of several documents, beginning with an official dispatch dated 15 May 1779 from the Royal site of Aranjuez (Madrid) by the Count of Floridablanca, and addressed to Pedro Franco Dávila. Both men were important figures: Floridablanca had the full confidence of King Carlos III from 1766 to 1788, and he held the position of Secretary of State from 1776 to 1792 (Hernández, 2009). Dávila in turn was the founder and first director of the Royal Cabinet of Natural History of Madrid, currently the National Museum of Natural Sciences (Museo

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Nacional de Ciencias Naturales, MNCN) (Calatayud, 1987; Velasco–Pérez, 2012). On 17 May 1779, Dávila acknowledged the receipt of the official dispatch to Floridablanca, and returned a copy of the documents that accompanied it. These documents, which deal with the studied bird pest, are the following: Document 5: letter dated 11 May 1779 written in Murcia by Marcos Mayoral and addressed to his superior, Miguel Múzquiz Goyeneche. The former was the General Intendant of Murcia, who at that time was the head of the Public Finance in the province, adjunct to the Ministry of Finance (Lillo and Lisón, 2002). The latter was a Minister of Carlos III, specifically the Secretary of the Treasury (Escobedo, 2007). Document 6: letter dated 10 May 1779 written in Totana by Bartolomé Fontana and addressed to Marcos Mayoral. According to the letter, Fontana was an employee of Spain´s 'Renta del Tabaco' (Tobacco Income), an important institution managed by Spanish Public Finance in the 18th century, and whose employees had extensive privileges (Escobedo, 2007). Document 7: letter–testimony also dated 10 May 1779 written in Pozo Estrecho (Cartagena) by Lucas Josef de Valera (including the testimonies of Francisco Meseguer, Fulgencio Saura and Miguel Sancho) and addressed to Marcos Mayoral. According to this letter, Valera, Meseguer and Sancho were employees of the Tobacco Income, whereas Saura was Deputy of the Common of Pozo Estrecho. File 472–019 (Municipal Archive of Vera, Almería) This File 472–019 'on the birds discovered in these surroundings' was formed, as specified in its beginning, by direct Order of the Governor of the Council of Castile. 'Governor of the Council' was a position of the highest status within the institutional framework of the Old Regime in Spain, as it was the maximum representative of the king (Granda, 2011). The main documents the file contains relevant to this work are: Document 8: it consists of four letters written in Madrid in early June 1779 and addressed to Manuel Serrano Cilleros, Major Mayor of Vera, as stated in the same document. At the beginning of the 19th century, the Mayor’s Office of Vera still consisted of four mayoralties including those of Vera and Mojácar (Almería), the latter in turn including the village of Turre (Morales, 1998). The first letter, dated 1 June 1779, is signed by Manuel Ventura Figueroa, Governor of the Council of Castile from 1773 to 1783 (Martínez, 2010). The second letter, dated 4 June 1779 is written by Antonio Martínez Salazar, notary of the Council of Castile and King’s Secretary (De Sancha, 1780). The other two letters are dated 6 and 9 June 1779 and are signed again by Manuel Ventura Figueroa. Document 9: Judicial Order ('Auto') dated 11 June 1779 written in Vera by Manuel Serrano Cilleros, and road letters (i.e., 'veredas', a type of mail notification to nearby areas) by the same author dated on the same day and addressed to several villages. Document 10: letter dated 29 June 1779 written in Vera by Manuel Serrano Cilleros and addressed to Manuel Ventura Figueroa.

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Document 11: letter dated 19 October 1779 written in Madrid by Antonio Martínez Salazar and addressed to the authorities of Vera. Document 12: Judicial Order dated 10 November 1779 written in Vera by Ginés Antonio García, and road letters by the same author dated on the same day addressed to several villages. Ginés Antonio García, as indicated in the same document, was a lawyer of the Royal Councils and a regent of the Major Mayor of Vera. Results We summarise the contents of the main documentary sources (all written in the Spanish of that time) that refer to events regarding the bird irruption (table 1). The eight (current) municipalities unequivocally affected by the outbreak (according to these sources) were: Vera, Mojácar and Turre (Almería); Cartagena, Lorca, Murcia and Totana (Murcia); and Orihuela (Alicante) (table 1). We developed a graph of the approximate spatial distribution of the bird irruption in Spain, which extended at least over the three (current) provinces indicated (fig. 1). We have superimposed municipalities for which there were indications of the birds but no solid documentary evidence (fig. 1). The most relevant information extracted from the documentary sources studied is assembled below: Document 1: on 8 May 1779, it was commented that for two days, large flocks of harmful birds called sparrows had arrived in Orihuela and quickly built nests and started laying. An order was issued requiring all inhabitants of La Huerta to 'eliminate' native sparrows and 'harmful aliens' (table 1), each neighbour having to contribute twelve 'heads of birds'. Several people were also selected to bring down the nests and kill the birds. Document 2: on 8 May 1779 it was commented in the city of Murcia that 'over a few days' birds similar to sparrows had been detected in great numbers and in the breeding stage at a place named El Campo (table 1). It was noticed that birds were destroying cereal crops, and that there were reports that there was a plague of the same bird in Cartagena and Orihuela. It was agreed to appoint an inspector to inform the Corregidor of Murcia. On 18 May 1779, the Deputy of the Common of Cañada Hermosa (Murcia) reported the occurrence of large numbers of strange birds resembling sparrows (table 1) eating the crops and causing great damage. Some men were sent to shoot the birds. In addition, a report dated 17 May 1779 detailed the results of the bird inspections carried out by Ginés Buitrago in certain parts of the municipality of Murcia (La Zarza, Barqueros, El Campo and Cañada Hermosa). It was certified that on the date of the report, birds remained only in some places, such as La Zarza, that birds bred and built their nests in olive trees, that there were eggs and sometimes nestlings in their nests, that birds were sparrows from other lands (table 1), and that in their place of origin either there were no available crops or crops were delayed with respect to their usual harvesting date. It

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was noticed that the greatest damage had occurred in early barley crops, that a plague 'of the same species' had occurred even more intensely 40 years previously, and that control actions had consisted of 'spells, gunpowder and nest removal'. Document 3: on 4 May 1779, Francisco Subiela, Deputy of the Common of Cartagena, noticed a plague of Barbary sparrows ('gorriones de Berbería') in 'Levante' and San Ginés (table 1). Sparrows appeared in great numbers, building numerous nests and causing great damage. It was said that such a plague had never been seen and it was agreed to inform the Governor of Cartagena. On 7 May 1779 the Governor replied that he had ordered the Regidor and the Deputy of the Common of Cartagena to perform inspections, and that subsequently he had been informed that the birds had vanished, as only ten or twelve individuals were seen. The inspectors verified the presence of many nests in trees, however, and orders were given for these to be knocked down. Document 4: on 17 April 1779, a mandatory imposition ('repartimiento') was agreed in Lorca to kill the birds, but no exceptional circumstances were reported. In contrast, on 4 May 1779, the Trustee of Lorca requested that measures be taken given the abundance of another bird species (table 1). It was agreed to continue with the measures adopted on 17 April. Document 5: Marcos Mayoral stated that a plague of birds had appeared between Murcia and Totana, causing great damage. He sent two inspectors from the Tobacco Income (Bartolomé Fontana and Francisco Meseguer) to various places for information. Mayoral attached the two writings by the inspectors to his report dated 11 May 1779, stating he had sent to Miguel Múzquiz two birds (collected by Fontana) and one egg; he added he was not including a nest as he considered they were not special in any way. Mayoral explains that the birds arrived dead, that he was forwarding them conserved in 'wine spirit' (i.e. alcohol), that one bird seemed to be an old adult because it had a white feather as a necktie, and that the other bird looked like a female or nestling, having no white feather. Document 6: on 10 May 1779, Bartolomé Fontana wrote that he had gone with some assistants to the field on 4 May, as he was informed that numerous small birds resembling sparrows had been observed in the previous days (table 1), nesting in the trees and causing great damage to crops in 'El Paretón' (Totana) and Alcanara (Lorca), and also in groves along the Sangonera River. He stated that: 1) he saw birds, especially in 'El Paretón' ; 2) they nested in olive trees and sometimes in black poplars; 3) clutch size ranged from 3 to 7 eggs; 4) some nests had two parts (each one with its eggs inside) but a single conduit to enter and exit; 5) in many olive trees there were between 100 and 250 nests; 6) they chased the birds away with daily repetition of gun shots; this did not get rid of them but it avoided their breeding; 7) each day they killed many birds; and to conclude, 8) the birds were a serious pest. In addition, Fontana informed Mayoral that he sent him four live birds captured with birdlime, and two nests. Document 7: Lucas Josef de Valera began his writing on 10 May 1779 stating that he left the city of

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Murcia on the same day in the morning, along with Francisco Meseguer, to whom he was helping. At first they did not find the birds, so they continued to Pozo Estrecho (Cartagena), where they were joined by Fulgencio Saura and Miguel Sancho. These latter informed them that, in the orchards of the Convent of San Ginés de la Jara (Cartagena), large numbers of birds with the size, feathers, beak and general look of sparrows (table 1) were damaging trees and crops. Saura and Sancho informed Valera that they had notified the authorities of Cartagena, who had sent large numbers of people to eliminate the birds. In fact, on 8 May, when Saura and Sancho were in command of 150 men, they found no birds, although villagers told them that there had been a lot of birds, that when the birds arrived they remained quite still but that in less than twelve hours they had built innumerable nests in trees, and that when men fired shots, the birds flew away towards Orihuela, and perhaps some of them also flew towards Lorca and Totana. These sources report that the bird damage was not too high because of the rapidity with which control actions were undertaken. It was also reported that one bird was captured alive and was sent to the Governor of Cartagena. He ordered the bird to be recognized by a Maghrebian from the Military Arsenal, who said that plagues of this bird had occurred many years in his land (table 1), that the birds were very harmful, and that Maghrebians cooked wheat with poison and spread the mixture in the fields to kill the birds. Document 8: on 1 June 1779 Manuel Ventura commented on a letter from Manuel Serrano dated 25 May informing about the appearance in Vera (Almería) of a new species of bird similar to sparrows (table 1) which was damaging wheat fields. Ventura stated that he already had heard about this bird pest from the Corregidor of Murcia and that he agreed with the measures taken by Serrano: prizes to collaborator villagers, mandatory impositions ('repartimientos') and the use of birdlime as hunting method applied on rods, canes and esparto plants ('en varetas, cañas y espartos'). In the other three letters written in Madrid, first Antonio Martínez —at the suggestion of the Council of Castile— and then Ventura, informed to Serrano in early June 1779 that they continue with the measures adopted to extinguish the bird pest. All four letters issued from Madrid are answers to successive Serrano’s writings (which are not kept in the file). In one of them, apparently dated 1 June 1779, Serrano specified that bird damages were centered on Vera, Mojácar and Turre (and such data are reproduced in the letter from Ventura dated 9 June 1779). Document 9: on 11 June 1779, Manuel Serrano transmitted a Judicial Order through road letters addressed to those towns and villages infested by the pest or at risk of being affected, but at no time he did distinguish between both situations. The complete list of localities to which road letters were addressed has not been able to be identified completely, although there is evidence that they were received in about thirty villages (almost all within the present province of Almería, but also seven villages in the province of Granada). Road letters communicated, among other

things, that the best way to exterminate the birds was by knocking down their nests. Document 10: Manuel Serrano writes on 29 June 1779 to Manuel Ventura informing about the diligences practiced in order to eradicate the bird pest. Document 11: Antonio Martínez writes on 19 October 1779, recalling how in the month of May of that year a great number of unknown birds arrived into Vera, Mojácar and Turre (table 1), reason why the Council of Castile gives the instructions summarised in the Martínez’s letter dated 4 June 1779. Appropriate measures are now dictated to defray the costs of gunpowder used trying to eliminate the pest of 'overseas' birds (table 1), and the affected towns are ordered to pay those costs to the Corregidor [Major Mayor] of Vera. In addition, he informs that the Council of Castile has agreed an Order warning local authorities in case the same bird pest returns next year. Document 12: in addition to recall the bird pest that took place in the spring of 1779, Ginés A. García comments on 10 November 1779 the damage that were causing to the olive harvest the 'innumerable flocks and clouds of foreign birds that now have come upon', in the Almanzora River and Cabrera Mountain Range, in the province of Almería. Therefore, he recommends to the authorities of the affected villages to shoot the birds at night in the places where they go to congregate and rest (letters were issued to half a hundred localities). Discussion Unconventional expressions were used in at least one dozen texts written in 1779 to refer to the birds involved in the pest outbreak occurring that year in Spain (table 1). The texts used the Spanish vernacular names 'gorrión' (sparrow) and, more often, 'pájaro' (bird). The latter word is used to name any passerine, any small bird, and also any bird in general, a peculiar trait of the Spanish language (Bernis, 1995). In most old texts, these names do not differ from the customary use and, in fact, they are the most abundant denominations for birds we have found in previous studies (Ferrero–García et al., 2014, 2016). However, the texts from 1779 all reflect the underlying doubts people had about the taxonomic ascription of the birds, or the conviction that the birds involved in the pest outbreak belonged to a species that today we would qualify as rare or unusual in the southeastern Iberian peninsula. Birds were described with epithets such as 'strange birds', 'unknown birds', 'foreign birds', 'sparrow–like birds', 'sparrow–shaped birds', 'sparrows from other lands', 'overseas birds' and 'Barbary sparrows'. In Document 3, the alleged provenance of birds is even indicated with relative precision ('Barbary sparrows'), providing valuable zoogeographical information regarding the target bird species. 'Barbary' ('Berbería') is an old European word that remained in use until the second half of the 19th century, roughly defining the current region of the Maghreb and, by extension, the whole of North Africa (De Bunes, 1989). This report, provided by an authority of Cartagena, increases its

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credibility because it coincides with the testimonies from other report by an authority of Pozo Estrecho, and also with those of three employees from the Tobacco Income (Document 7) who narrated how a Moor working in the Military Arsenal of Cartagena had examined the birds and had recognised them 'as characteristic of his homeland'. Thousands of people worked in this military centre in the 18th century, including many Maghrebians as prisoners (Henares, 2010). It is likely that Díaz (1895) collected his testimonies when talking about the plague of 'African sparrows' in Cartagena, a city that is less than 200 km from the North African coast at its nearest point (Oran, Algeria). A key aspect of the documents in table 1 is that the protagonists were not usually in contact with one another, so they would not be mutually influenced. In any case, these two texts (Documents 3 and 7) are those that most accurately point to the geographical origin of the bird pest. The documentary sources locate the bird pest with certainty in an array of eight municipalities (fig. 1). This does not mean that the whole of each municipality was occupied by birds, or that some lands within neighbouring municipalities were not partially populated. Indeed it is likely that the birds arrived in another thirty municipalities in the provinces of Almería and Granada (fig. 1, Document 9). The historical documents sometimes allow the birds to be placed in relatively definite places such as small villages that still exist today (Pozo Estrecho, La Zarza, Barqueros, Cañada Hermosa, El Paretón and Alcanara). Some natural areas are also mentioned as populated by birds. Examples include the 'Campo de Murcia', which currently corresponds to the southern half of the municipality of Murcia (Aliaga, 2008), the section of the Sangonera River between Lorca, Totana and Murcia, better known as the Guadalentín Valley (Ramírez and Baños, 1997), or the 'Huerta de Orihuela', which constitutes a well–defined territory around the Segura River (Ferrández and Diz, 2015). In one case it is even possible to locate the bird pest with complete accuracy (Document 7): it is the Convent of San Ginés de la Jara, located a few kilometers east of the city of Cartagena, with origins dating back to the 13th century, and whose building remains standing although in a ruinous state (Sánchez, 2012). Regarding the dates the birds appeared we agree with Grima’s (1990) version. Birds irrupted into the coast of Cartagena on the first days of May 1779 (Document 3), continuing to Lorca and Totana (Documents 4 and 6), and then to Murcia (Document 2) and Orihuela (Documents 1 and 2). No bibliographic source, as far as we know, has previously cited the bird pest in the latter municipality belonging to Alicante province. Testimonies indicate that around May 7, a few birds remained in Cartagena (Documents 3 and 7) but most of them disappeared during the second ten–day period of the month. An exception is found in the municipalities of Almería, where birds arrived from the Guadalentín Valley in the second half of May (Document 8) and probably persisted at least until some time in June (Documents 9 and 10). In short, birds appeared suddenly and then completely vanished in just two months. One possible explanation

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for the massive bird arrival could be poor harvests in their homelands, which would have deprived them of trophic resources (Document 2) and set off a migratory process. Lack of food is a main ecological factor triggering migratory behaviour in birds, possibly leading to a large–scale massive irruption in the destination areas (Nelson, 1985; Berthold, 2001; Newton, 2006). In this regard, it should be noted that the period 1760–1800 was characterised in the western Mediterranean basin by a notorious climatic oscillation known as the Maldá anomaly, with extreme hydrometeorological episodes and major losses in agriculture (Barriendos and Llasat, 2003). In many areas of the Iberian peninsula, this anomaly provoked climatic events such as cold and heat waves, and floods and droughts (Oliva et al., 2018). In relation to the municipalities of Almería, the judicial order and road letters issued on 10 November 1779 by the regent of Major Mayor of Vera (Document 12) were surprising. This figure, who had not intervened until then, documented a supposed return flight of the birds in November. However, what he seems to describe is the typical attack by wintering flocks of starlings (Sturnus spp.) and/or thrushes (Turdus spp.) that come to the Iberian peninsula to feed on mature olives, an occurrence documented both in the past (Ferrero–García et al., 2014) and today (Tellería et al., 1999; SEO/BirdLife, 2012; De Juana and Garcia, 2015). Such a surprising judicial action could have something to do with the fact that the Council of Castile had decided a few weeks earlier that a part of the expenses caused in the control of 'overseas' birds had to be paid to the authorities of Vera by the affected villages (Document 11). It is of note that some documents written as of October 1779 include statements that are difficult to believe. As these documents were generated after a three–month parenthesis without any news regarding the bird plague, it is assumed they were probably written by people who were not direct witnesses of the events that occurred the previous spring. Thus, in the Order of 19 October 1779, bird nests were said to contain 28–30 eggs (Grima, 1990) in contrast with the 3–7 eggs mentioned in May by Bartolomé Fontana (Document 6), or the 5–6 eggs published in the Spanish official bulletin in the same month (Gaceta de Madrid, 1779). The latter clutch sizes appear to be much more reliable. Klom (1970) noted pheasants (8–15 eggs) and partridges (14–20 eggs) as examples of bird groups with a very large clutch size, neither of which are passerines. More recently, Jetz et al. (2008) analysed data from 5,290 bird species and calculated a mean clutch size of 2.8 eggs (mode = 2) with only six bird species having mean values higher than 14 eggs: Rhea americana (19.7), Rhea pennata (17.3), Aepypodius arfakianus (20.0), Perdix perdix (16.0), Perdix dauurica (19.0) and Callipepla californica (14.9). None of these six birds are passerines. In fact, in this latter study, only three passerines (out of almost three thousand analyzed) were found to laid more than 9 eggs, and never above 10: Aegithalos caudatus (9.8), Regulus regulus (9.9) and Parus ater (9.4).

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Table 1. Main textual references found in the unpublished documentation studied (written in the Spanish language of the 18th century) referring to birds causing the mysterious plague of 1779 in the Iberian peninsula. A free translation into the English language is given. The source document and the affected municipalities (as well as the corresponding current provinces) according to these documentary sources are also indicated. Tabla 1. Principales referencias textuales encontradas en la documentación inédita estudiada (escritas en la lengua española del siglo XVIII) en las que se mencionan las aves que causaron la misteriosa plaga de 1779 en la península ibérica. Se da una traducción libre al idioma inglés. También se indican la fuente documental y los municipios afectados (y las provincias actuales correspondientes) según estas fuentes. Municipality Main textual references

Document

Orihuela (Alicante) "a matar assi los gorriones del pais, como los extranjeros dañeros" [to kill both native sparrows and "harmful aliens"]

Murcia (Murcia) "espezie de paxaros gorriones" [birds similar to sparrows]

"multitud de paxaros extraños a forma de gorriones" [a multitude of strange birds resembling sparrows]

"siendo cierto que son gorriones de otras tierras" [being true that they are sparrows from other lands]

Totana and Lorca (Murcia) "abundancia de pájaros pequeños con similitud á los gorriones" [abundance of small birds resembling sparrows]

Lorca (Murcia) "abundancia de otra especie" [abundance of another bird species]

Cartagena (Murcia) "plaga de gorriones de Berbería" [plague of Barbary sparrows]

Cartagena (Murcia) "muchedumbre de paxaros, su grandaria, pluma, pico, y todo él de gorrión" [swarm of birds, their body size, feathers, beak and a general sparrow appearance]

"un moro manifestó que en su tierra la morisma muchos años se experimenta esta plaga" [a Moor stated that in his homeland, the Maghreb, this bird plague occurred over many years]

Vera (Almería) "nueba especie de paxaros semejantes a gorriones" [new bird species similar to sparrows]

Vera, Mojácar and Turre (Almería) "pájaros desconocidos" [unknown birds]

"plaga de pájaros ultramarinos" [plague of overseas birds]

Another relevant aspect to consider is whether the bird pest actually caused severe damage to agricultural crops. Most evidence indicates that they did, particularly to cereals, with wheat (Document 8) and barley (Document 2) being expressly mentioned. It is unquestionable that people of that time was alarmed in an extraordinary way. Therefore, actions against birds were deployed, not only with the available

means in each municipality (as it was usual in the case of 'normal' bird pests; Ferrero–García et al., 2014, 2016, Torres–Vila et al., 2015) but also with the intervention of major State institutions: the Tobacco Income in the province of Murcia, and the Council of Castile in the province of Almería. Indeed, even the highest authorities in the country, including two ministers of Carlos III (the Count of Floridablanca

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Alicante Murcia

Granada Almería 100 km

Fig. 1. Municipalities in the current provinces of Alicante, Murcia and Almería (southeastern Iberian peninsula) affected by the mysterious bird plague of 1779 (dark grey). Other possibly affected municipalities within the provinces of Almería and Granada are also shown (light grey, see text). Fig. 1. Municipios de las actuales provincias de Alicante, Murcia y Almería (sureste de la península ibérica) afectados por la misteriosa plaga de aves de 1779 (gris oscuro). También se muestran otros municipios que pudieron verse afectados en las provincias de Almería y Granada (gris claro, véase el texto).

and Miguel Múzquiz Goyeneche) and the Governor of the Council of Castile (Manuel Ventura Figueroa), knew about the bird pest. The last question to be dealt with is the most complex and transcendent: what bird, presumably alien to southeastern Iberia avifauna, could provoke such an unusual pest outbreak? We are looking for a bird with the size and appearance of a sparrow, that populates trees in huge colonies, builds multitude of nests per tree, has a clutch size of at least 6–7 eggs, eats mostly cereals or other small grains, and potentially causes severe damage to crops. These features agree with those summarized by the Spanish official bulletin of the epoch (Gaceta de Madrid, 1779). In Africa (the continent of origin that some testimonies pointed out) there are several candidate birds that fulfill the above features, at least in part. Within the Ploceidae family, we find some species accused of producing important plagues in sub–saharan Africa: the Red–billed Quelea Quelea quelea and the Village Weaver Ploceus cucullatus (Bernis, 1989; Craig, 2010). However, at least today, no Ploceidae species arrives as far north in its distribution as to the coasts of the Mediterranean Sea, and what is more, they do not venture into the Sahara desert (Craig, 2010; Del Hoyo and Collar, 2016). In addition, clutch size in these species does not exceed four eggs and males have easily recognisable and conspicuous colours in some body parts (Craig, 2010). It should be noted that testimonies contained in our documents never mention any details regarding the colour of

birds, which make us suspect that surely there was nothing to emphasize. Regarding the Passeridae, several species are able to form local concentrations and damage crops (Summers–Smith, 1988; Bernis, 1989). In Africa, damage caused by the Sudan Golden sparrow Passer luteus can be important (Bernis, 1989; Summers–Smith, 2009). This species, however, inhabits only a narrow strip from Mauritania to Sudan (Summers–Smith, 2009; Del Hoyo and Collar, 2016), has a clutch size of less than 5 eggs, and the male has a conspicuous golden colour on the head and belly (Summers–Smith, 2009). It follows that all the above mentioned passerine species are unlikely to be the cause of the 1779 pest outbreak. In contrast, there is another Passeridae that we must carefully consider: the Spanish sparrow, curiously called the 'Moorish sparrow' in Spanish 'gorrión moruno'. The Spanish sparrow (sometimes considered conspecific with the house sparrow and the Italian sparrow P. italiae; Fulgione and Rippa, 2012; Del Hoyo and Collar, 2016) is currently distributed in some parts of Asia, eastern Europe and most of northern Africa, from Morocco to Libya (Summers–Smith, 1988; Bernis, 1989; Summers–Smith, 2009; Del Hoyo and Collar, 2016). Already in the 19th century a similar distribution was mentioned, with great abundance in Algeria (Degland and Gerbe, 1867). The species also inhabits some areas in the Iberian peninsula, mainly in the southwestern sector, and its presence in the south east is marginal or occasional only (Alonso, 1997; Bernis, 1989; Tellería et al., 1999;

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Roviralta, 2003; SEO/BirdLife, 2012; De Juana and Garcia, 2015; Calvo et al., 2017). Despite its English common name, for unknown reasons, the Spanish sparrow in Spain has undergone marked fluctuations in both population abundance and distribution since the second half of the 19th century, being prone to abandon colonies and establish new ones (De Juana and Garcia, 2015). Neither Reyes' (1886) or Arévalo's (1887) decimononic recompilations mentioned the occurrence of the Spanish sparrow in any locality of the provinces of Almería, Murcia and Alicante (nor in those of the neighbouring provinces). In the 1960s–1980s, the species restricted its already scarce populations to certain areas of the Tajo and Guadiana basins, especially within the Extremadura region (Bernis, 1989; De Juana and Garcia, 2015). However, by the late 20th century the species had greatly expanded its range and in Extremadura it even became a pest, with colonies of almost 30,000 nests and yield losses of up to 90 % in certain crops (Prieta, 2003), probably motivating its non–protected status in Spain after 1999 (Ferrero–García et al., 2014). As above mentioned, the Spanish sparrow is not only characterised by its high population fluctuations, but also by its irregular and unpredictable movements, particularly those in North Africa. Both characteristics are closely related to interannual variations in the ecological conditions (Summers–Smith, 1988). In any case, in the Iberian peninsula the Spanish sparrow is a wandering species outside the breeding season and part of the population may overwinter in northern Africa (De Juana and Garcia, 2015). Therefore, the explanation of the massive arrival of the birds in May 1779 (within the breeding season) would have to be sought in an exceptional situation, maybe related to the aforementioned Maldá anomaly. The Spanish sparrow is known to provoke serious damages to cereal crops in North African countries such as Algeria (Metzmacher and Dubois, 1981; Bellatreche, 1986; Summers–Smith, 1988; Behidj– Benyounes et al., 2011; Belkacem et al., 2012). The species normally breeds in trees forming large colonies, with a number of nests reaching well over a hundred per tree (Bernis, 1989; Tellería et al., 1999; Roviralta, 2003; Belkacem et al., 2012), and a clutch size of up to seven eggs (Summers–Smith, 1988; Metzmacher, 1990; Tellería et al., 1999; Marques, 2002), so that it is widely considered the most gregarious sparrow in the Palaearctic (Summers–Smith, 1988). Bernis (1989) stated that the Spanish vernacular name of this species ('gorrión moruno' = Moorish sparrow) was influenced by the knowledge of the severe damage that these birds caused in Morocco, and the huge magnitude of some of their colonies there. At least in the 19th century, this vernacular denomination was used in Spain together with other common names (Arévalo, 1887). Summers–Smith (1988) mentions that the main English vernacular name (Spanish sparrow) is rather unfortunate as it simply derives from the fact that the species was described based on some specimens from Spain that were sent from Gibraltar but captured in Algeciras, Cadiz (Temminck, 1820).

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Spanish sparrow females have a coloration almost the same as that of house sparrow females. Males of the two species are somewhat more distinct, and more showy in spring in both cases (see De Juana and Varela, 2000; Summers–Smith, 1988, 2009; Del Hoyo and Collar, 2016). The house sparrow is widely distributed throughout Iberia, including the southeastern sector (Bernis, 1989; Tellería et al., 1999; Molina, 2003; SEO/BirdLife, 2012; De Juana and Garcia, 2015; Calvo et al., 2017). This scenario was the same in the second half of the 19th century (Reyes, 1886; Arévalo, 1887) and we may infer that it was similar in the 18th century (see Ferrero–García et al., 2014). Therefore, if people of that century had to describe the Spanish sparrow (assuming that they had never seen it before, or at least they had not noticed it because its presence was sparse or occasional) in a geographical area where the house sparrow was well–known, surely the simplest and easy description would be to use expressions invoking the similarity of the first sparrow species with the second one. In the documents analysed, only Marcos Mayoral (Document 5) comments on a perceptible physical feature that apparently one of the birds exhibited (a male adult): a white feather. The Spanish sparrow male has notably very white cheeks (De Juana and Varela, 2000; Summers–Smith, 1988, 2009; Del Hoyo and Collar, 2016). However, it was a dead bird (Mayoral had never seen it alive) and perhaps somewhat deteriorated, as he defined the specimen as 'old'. The bird had in fact been caught by using birdlime, a hunting procedure that damages feathers (Murgui, 2014). It is relevant to note that nest removal is mentioned as one of the most effective bird control methods in some of the documents studied. At present, this method is used mutatis mutandis for the control of the Spanish sparrow in some African countries such as Algeria (Belkacem et al., 2012). In Tunisia there was even a law of 1892 that required people to destroy the Spanish sparrow nests (Summers–Smith, 1988). Note also that during an anti–pest campaign carried out in Kyrgyzstan (Central Asia), almost two million Spanish sparrows were eliminated with poisoned wheat (Summers–Smith, 1988; Bernis, 1989). Thanks to Document 7, we know that this dangerous control method was also probably used in the Maghreb to control this species during the second half of the 18th century. It is also interesting to verify the use of birdlime in the provinces of Murcia (Document 6) and Almería (Document 8) in the same period, a non–selective hunting method whose use in other regions of Spain has been a source of conflict since the first half of the 20th century to the present (Ferrero–García, 2017a, 2017b). In conclusion, we have the certainty that an extraordinary bird outbreak occurred in Spain in May 1779, damaging cereal crops in several municipalities in Alicante, Murcia and Almería, and lasting in the latter province until June. The pest outbreak was probably caused by the irruption of a sparrow–sized passerine bird that formed large arboreal colonies with many nests per tree and whose presence was totally unusual in southeastern Iberian peninsula.

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Descriptions narrated by direct witnesses coincide regarding the biology and behavior of the (badly named) Spanish sparrow and are consistent with the distribution of the species in southern Iberia and the Maghreb, as well as with its morphology and other species–specific features. We thus hypothesize that the Spanish sparrow would be the protagonist of this historic bird pest. Acknowledgements We thank the staff at the municipal archives of Vera, Orihuela and Lorca for their kind collaboration and two anonymous referees for their comments and suggestions. This research was supported in part by the Servicio de Sanidad Vegetal (SSV), Junta de Extremadura. References AHRM [Achivos Históricos de la Región de Murcia], 2017. Proyecto Carmesí http://www.regmurcia.com/ servlet/s.Sl?METHOD=CUADROCLASIFICACION &sit=c,373,m,139,serv,Carmesi Aliaga, I., 2008. Nuevos desarrollos urbanísticos en el Campo de Murcia. Implicaciones territoriales y planeamiento territorial. Papeles de Geografía, 47–48: 5–24. Alonso, J. C., 1997. Gorrión moruno Passer hispaniolensis. In: Atlas de las aves de España (1975–1995): 498–499 (F. J. Purroy, Ed.). Lynx Edicions, Barcelona. Arévalo, J., 1887. Aves de España (Memorias de la Real Academia de Ciencias Exactas, Físicas y Naturales, tomo XI). Imprenta de viuda e hijo de Aguado, Madrid. Barriendos, M., Llasat, M. C., 2003. The Case of the 'Maldá' Anomaly in the Western Mediterranean Basin (AD 1760–1800): An Example of a Strong Climatic Variability. Climatic Change, 61: 191–216. Behidj–Benyounes, N., Bissaad, F. Z., Behidj, K. K., Chebouti, N., Doumandji, S., 2011. Variations inter parcellaires des dommages dus au moineau hybride Passer domesticus x P. hispaniolensis sur céréales dans un milieu agricole de l’extrême partie orientale de la Mitidja (Algerie). Sciences & Technologie, C–34: 61–71. Belkacem, A. A., Sekour, M., Doumandji, S., 2012. Effectiveness of mesh netting and nest’ destruction in protection of crops against attack by Spanish sparrow Passer hispaniolensis. African Journal of Agricultural Research, 7: 4575–4580. Bellatreche, M., 1986. Approche économique des dégâts aviaires en Algérie. Ann. Inst. Nati. Agro., El Harrach, 10: 181–195. Bernis, F., 1989. Los gorriones. Con especial referencia a su distribución y eto–ecología en las mesetas españolas (Comunicaciones INIA. Serie Recursos Naturales). Ministerio de Agricultura, Pesca y Alimentación, Madrid. – 1995. Diccionario de nombres vernáculos de aves.

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Bird roadkill occurences in Aragon, Spain D. Vidal–Vallés, A. Rodríguez, E. Pérez–Collazos

Vidal–Vallés, D., Rodríguez, A., Pérez–Collazos E., 2018. Bird roadkill occurences in Aragon, Spain. Animal Biodiversity and Conservation, 41.2: 379–388. Abstract Bird roadkill occurences in Aragon, Spain. The increase in road networks and vehicular traffic has posed a major threat to vertebrates over the last century. Although it is difficult to determine the annual number of avian–vehicle collisions, 2 to 9 million roadkills have been estimated for Europe, with numbers varying from country to country. Few studies have been conducted at a national or regional level in Spain. In this study we used data from La Alfranca Wildlife Rehabilitation Centre database to determine location, season and incidence of avian–vehicle collisions in the autonomous county of Aragon (Spain). A total of 643 wild birds representing 71 species were killed on roads between 2012 and 2014. Nine of these species have a high incidence of avian–vehicle collisions, four a moderate incidence, and 57 a low incidence. The species with the highest incidence was the griffon vulture (120 individuals). Spatial distribution of avian–vehicle collisions was heterogeneous, and the incidence was highest in July, August and September. We identified 41 reas of high roadkill occurrence, using a number of roadkills per km index (RI): 28 in the province of Zaragoza, 9 in Huesca and 4 in Teruel. Management strategies are proposed to reduce this threat on wild birds. Key words: Areas of high roadkill occurrence, Mitigation, Necrophagous, Number of roadkills per km index, Raptor, Road ecology Resumen Atropellos mortales de aves en Aragón, España. La expansión de la red viaria y el aumento del tráfico de vehículos se han convertido en una amenaza importante para los vertebrados en el último siglo. Algunas estimaciones sitúan entre 2 y 9 millones los individuos afectados anualmente en Europa, con variaciones entre países. En España existen pocos estudios a escala nacional y autonómica. En este estudio se ha utilizado la base de datos del Centro de Recuperación de Fauna Silvestre de La Alfranca para determinar la localización, temporalidad e incidencia de los atropellos de avifauna que se producen en la comunidad autónoma de Aragón. Se han detectado 643 atropellos mortales de aves silvestres (71 especies) entre los años 2012 y 2014. Nueve de estas especies presentan una incidencia de atropellos alta; cuatro, una incidencia moderada, y 57, una incidencia baja. La especie con la mayor incidencia fue el buitre leonado (120 individuos). La distribución espacial de los atropellos fue heterogénea y julio, agosto y septiembre fueron los meses con mayor incidencia. Se identificaron 41 áreas de incidencia alta de atropellos, usando el índice de atropellos mortales por kilómetro (RI): 28 en la provincia de Zaragoza, 9 en la de Huesca y 4 en la de Teruel. Se proponen una serie de estrategias de gestión con el fin de reducir esta amenaza para las aves silvestres. Palabras clave: Áreas de incidencia alta de atropellos, Mitigación, Necrófago, Índice de atropellos mortales por kilómetro, Rapaz, Ecología de carreteras Received: 29 IX 17; Conditional acceptance: 21 XII 17; Final acceptance: 15 II 17 David Vidal–Vallés, Ernesto Pérez–Collazos, Escuela Politécnica Superior de Huesca, Univ. de Zaragoza, carretera de Cuarte s/n., 22071 Huesca, España.– Airam Rodríguez, Depto. de Ecología Evolutiva, Estación Biológica de Doñana–CSIC, Av. Américo Vespucio 26, 41092 Sevilla, España. Corresponding author: David Vidal–Vallés. E–mail: davidvidalvalles@gmail.com ISSN: 1578–665 X eISSN: 2014–928 X

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Introduction Road networks can generate major impacts on wildlife, the two most important being habitat fragmentation and direct mortality. Habitat fragmentation creates barriers that prevent wildlife movements, thereby increasing the rate of local extinction (Fahrig, 2003; Martínez–Freiría and Brito, 2012; Loss et al., 2014), while direct mortality involves death of individuals due to vehicle collisions (Loss et al., 2014; Sáenz–de–Santa–María and Tellería, 2015). Although precise global estimates of the number of collisions are lacking, the number of vertebrate roadkills per year could exceed 1,000 million individuals (PMVC, 2003). Among vertebrates, small mammals are those most likely to be killed on roads because their size allows them to get through highway containment barriers (Haigh, 2012; Saranholi et al., 2016). Scavengers or species with opportunistic feeding habits –such as badgers and martens or vultures and kites– may also have increased possibility of becoming roadkill when they are attracted to corpses on the road. Although it is complex to estimate the real number of individuals hit on a road network, studies from various countries have shown that the number of bird roadkills is high and a serious problem for conservation. Recent studies indicate an annual average of 199 million birds in the United States (Loss et al., 2014), 9.4 million in Germany (Fuellhaas et al., 1989), 7 million in Bulgaria (Nankinov and Todorov, 1983), 4.6 million in Canada (Bishop and Brogan, 2013), and 2 million in the Netherlands (Schrijver, 1993). In Spain, few estimates are available at a national level. However, 74,600 wildlife collisions were estimated between 2006 and 2012 (Sáenz–de–Santa–María and Tellería, 2015). For birds, the most relevant data may be the 16,036 roadkilled individuals, corresponding to 193 species, during 1990–1992 (PMVC, 2003). Recent studies have been conducted in entire Autonomous Communities (Catalonia: Garriga et al., 2012; Aragón: Vidal–Vallés and Pérez–Collazos, 2016) and large natural areas (Rico–Guzmán et al., 2011; Espinosa et al., 2012; Martínez–Freiría and Brito, 2012). This study aimed to identify the non–hunted wild bird species with the highest incidence of vehicle collisions in Aragón, Spain, and to determine the locations and the temporality with the greatest incidence. Material and methods Roadkill maps Our study area included the entire road network of the Autonomous Community of Aragon, in northeastern Spain. Aragon comprises three provinces (from North to South: Huesca, Zaragoza and Teruel) and covers an area of 47,710 km2. It has a population of around 1,300,000 people. We obtained information on wild bird collisions from the database of the Wildlife Recovery Center (WRC) of La Alfranca (Zaragoza, Spain), which is the only WRC in Aragon. Injured birds were kept at the WRC until death or release. In the case of animals that

arrived dead, the cause of the death was determined by the WRC veterinarians through necropsies. WRC database holds information on terrestrial vertebrates collected in the field, injured or dead, from any part of Aragon and admitted for any cause. Given that it is a database from a WRC, it does not include game species, especially large ungulates, which are managed by the SEPRONA, Servicio de Protección de la Naturaleza of the Guardia Civil (Nature Protection Service of the National Police). We selected only records referring to vehicle collisions of birds in the WRC database and we assumed that all admitted birds (deceased or recovered and released into the wild) would have died without admission to the WRC. When available, UTM coordinates and place–names of the locations helped us to georeference collisions. By using geoprocesses of ArcMap 10.1, we created a layer of points in shapefile format that represents the bird collisions recorded in the road network from 2012–2014. For each record, we manually added the following information: bird species, month and year of collision detection, and category of threat according to regional Aragón law (EN, Endangered; VU, Vulnerable; SAH, Sensitive to the alteration of its habitat; Decree 181/2005). We intersected the road network of Aragon, downloaded from the Instituto Geográfico de Aragón (Spatial Data Infrastructure of Aragon: available online at http://idearagon.aragon.es/portal/), to the point layer representing bird collisions. To determine the species with the highest incidence of roadkills, we adapted the incidence categories proposed by Tenés et al. (2007): low (0–3 roadkills per year), moderate (4–6 roadkills per year), high (7–23 roadkills per year) and very high (more than 24 roadkills per year, i.e. two or more roadkills every month). Data analyses To determine roadkill temporality, we classified each roadkill record within seasons (spring: April–June; summer: July–September; autumn: October–December; winter: January–March). We then conducted a Kruskal–Wallis test to assess potential roadkill mean differences between seasons and Nemenyi post–hoc tests to test which levels (seasons) differed from each other. To study the spatial distribution of roadkills in Aragon, we created a grid of 10 x 10 km cells. We inserted the point layer of collisions with the grid to identify the grids with the highest incidences of roadkills (black squares). We calculated the number of roadkills per kilometre Index (RI), taking into account the kilometres of road in each 10 x 10 km cell. We conducted a Pearson correlation test to assess the relationship between number of roadkills and RI (number of roadkills per kilometre index). The areas of high roadkill occurrence were identified by the RI in each of them. We assigned four incidence ranges (very low, low, moderate and high) according to the number of roadkills: very low incidence (RI = 0–0.09, results not shown), low incidence (RI = 0.1–0.19), moderate incidence (RI = 0.2–0.29) and high incidence (RI > 0.3). This information was turned onto the GIS, generating an Aragon map of areas of high roadkill occurrence

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ordered by wild bird roadkills. In addition, a heat map was created with the Heatmap tool in QGIS (version 2.18.12) with a radius of 10,000 m. Results Involved species A total of 643 records of bird collisions (64 % of the total database) were detected, involving 71 species (table 1). The griffon vulture (Gyps fulvus) was the species with the highest incidence (120 cases), with a mean of 3.3 vehicle collisions per month during 2012–2014, and it was classified in the Very High category (mean of 40 roadkills per year). In addition, nine species showed a high incidence: common buzzard, black kite, white stork, common kestrel, eagle owl, short–eared owl, little owl, marsh harrier and barn owl (table 1). Four species reached a moderate incidence: tawny owl, red kite, sparrow hawk and scops owl. The remaining species showed a low incidence (table 1). Eleven species affected by road casualties are catalogued in the maximum category of protection according to the Aragón law (Decree 181/2005; table 1). Spatial distribution and temporality The spatial distribution of bird collisions was heterogeneous but associated with the main roads of the Aragon road network, 88.3 % of the bird collisions were reported in the provinces of Zaragoza and Huesca (302 and 266 roadkills, respectively) (fig. 1). The highest incidence of roadkill occurred in September (83 cases), followed by July and August with 80 and 69 roadkills, respectively. Lowest values were recorded in October, November and December (34, 30 and 31, respectively) (fig. 2). The number of roadkills varied among seasons (K–W = 16.127; d.f. = 3; P < 0.001), although pairwise comparisons only detected significant differences between summer and autumn (P < 0.001; fig. 2). The remaining pairwise comparisons reached P–values higher than 0.097. In case of the griffon vulture, the species with the highest incidence, the highest mortalities were reached in September and March (19 and 17 roadkills, respectively) and the minimum in May (5). Identification of areas of high roadkill occurrence We identified 41 areas of high roadkill occurrence, with a total of 238 roadkills. Two cells had a high incidence, while the rest were classified as moderate (5) or low (34) incidence (fig. 3). We found a positive correlation between the number of roadkills and RI (r = 0.734, P < 0.001). The number of roadkills also correlated positively with the length of roads in each cell in kilometres (r = 0.533; P < 0.001). The number of areas of high roadkill occurrence (28) was highest in the province of Zaragoza, followed by Huesca (9) and Teruel provinces (4) (fig. 3). The species with the highest incidence at areas of high roadkill occurrence were the white stork (5 roadkills) and the black kite (3 roadkills).

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Discussion Methodological aspects Conducting a sampling of wildlife collisions in large areas such as Aragon is a difficult task due to multiple factors, such as detectability (detection of roadkills and injured birds) and carcass persistence times (removal of carcasses by scavenger species or disappearance by multiple vehicle run over) (Bishop and Brogan, 2013; Beckmann and Shine, 2015; Santos et al., 2015; Vidal–Vallés and Pérez–Collazos, 2016). La Alfranca WRC database has relevant data missing due to: (i) lack of game species records, (ii) bias related to areas with a lower/greater presence of Nature Protection Agents, (iii) data deletion of non–informative records (i.e. records without date or location), and iv) lack of citizen awareness to inform the WRC about run over species. Despite these limitations, this database allowed us to determine a list of bird species affected by vehicle collisions in Aragon (table 1), the spatial and temporal distribution of the roadkills (figs. 1, 2), and to identify the areas of high roadkill occurrence where incidence is higher in the region (fig. 3). Furthermore, we suggest management recommendations (see below), linked to more detailed studies confirming the magnitude of bird mortality in the detected black squares. Involved species The griffon vulture was the species with the highest incidence (120 vehicle collisions, table 1), and given its body mass (around 6–11 kg), it poses a serious threat to road safety. Its large population in Aragon (the autonomous community with the second largest population in Spain, the largest being Castilla–León (Hernández, 2009), its high capacity of movement to explore large areas, its heavy weight, making it difficult to take off from the ground, and its strict scavenger feeding habits could explain the high number of collisions. The number of individuals run over annually (34 in 2012, 42 in 2013 and 44 in 2014) may affect the population dynamics of the species (Oschadleus and Harebottle, 2002; Fahrig and Rytwinski, 2009). The fact that a species with strict necrophagous feeding is the most vulnerable to vehicle collision suggests the accidents occur while the birds are feeding on corpses of other run over vertebrates. This is similar to findings for other species of opportunistic feeding habits, such as black kite, red kite, booted eagle, and common buzzard, which, in some cases, feed on dead animals (Planillo et al., 2015). These predatory opportunistic species can prey 73 % of the animals run over in a few hours (Beckmann and Shine, 2015). Black kite, red kite and common buzzard can become accustomed to travelling and exploiting roads as a source of food (Blanco and Viñuela, 2004; Planillo et al., 2015). This effect might be enforced by high densities of keystone prey species of the Mediterranean ecosystem near roads, according to a study on small mammals (Apodemus sylvaticus, Crocidura russula, Mus spretus; Ruiz–Capillas et

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Table 1. List of avian–vehicle collisions in Aragon (2012–2014) with number (N) and incidence (I). Species included in the Aragon Red List: * Endangered; ** Vulnerable; and *** Sensitive to habitat alteration. Tabla 1. Listado de las especies de aves silvestres atropelladas en Aragón (2012–2014). Se indica el número de atropellos (N) y su incidencia (I). Especies incluidas en el Catálogo de Especies Amenazadas de Aragón: * en peligro de extinción; ** vulnerable; *** sensible a la alteración de su hábitat. Family

Common name

Scientific name

Accipitridae

Griffon vulture

Gyps fulvus

120

Very High

Common buzzard

Buteo buteo

High

Black kite

Milvus migrans

High

Western marsh harrier

Circus aeruginosus

High

Red kite ***

Milvus milvus

Moderate

Eurasian sparrowhawk

Accipiter nisus

Moderate

Short–toed snake eagle

Circaetus gallicus

Low

Booted eagle

Aquila pennata

Low

European honey buzzard

Pernis apivorus

Low

Northern goshawk

Accipiter gentilis

Low

Bearded vulture *

Gypaetus barbatus

Low

Egyptian vulture **

Neophron percnopterus

Low

Hen harrier ***

Circus cyaneus

Low

Montagu's harrier **

Circus pygargus

Low

Golden eagle

Aquila chrysaetos

Low

Cinereous vulture

Aegypius monachus

Low

Alaudidae

Crested lark

Galerida cristata

Low

Alcedinidae

Common kingfisher

Alcedo atthis

Low

Apodidae

Common swift

Apus apus

Low

Ardeidae

Grey heron

Ardea cinerea

Low

Western cattle egret

Bubulcus ibis

Low

Eurasian bittern *

Botaurus stellaris

Low

Purple heron **

Ardea purpurea

Low

Burhinidae

Eurasian stone–curlew

Burhinus oedicnemus

Low

Caprimulgidae

European nightjar

Caprimulgus europaeus

Low

Red–necked nightjar

Caprimulgus ruficollis

Low

White stork

Ciconia ciconia

High

Ciconiidae Corvidae

Common raven

Corvux corax

Low

Red–billed chough **

Pyrrhocorax pyrrhocorax

Low

Cuculidae

Great spotted cuckoo

Clamator glandarius

Low

Common cuckoo

Cuculus canorus

Low

Falconidae

Common kestrel

Falco tinnunculus

High

Lesser kestrel ***

Falco naumanni

Low

Peregrine falcon

Falco peregrinus

Low

Merlin

Falco columbarius

Low

Red–footed falcon

Falco vespertinus

Low

Eurasian hobby

Falco subbuteo

Low

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Table 1. (Cont.) Family

Common name

Scientific name

Fringillidae

European greenfinch

Chloris chloris

Low

Gruidae

Common crane ***

Grus grus

Low

Hirundinidae

Barn swallow

Hirundo rustica

Low

Common house martin

Delichom urbicum

Low

Laniidae

Woodchat shrike

Lanius senator

Low

Meropidae

European bee–eater

Merops apiaster

Low

Motacillidae

White wagtail

Motacilla alba

Low

Muscicapidae

European pied flycatcher

Ficedula hypoleuca

Low

Oriolidae

Eurasian golden oriole

Oriolus oriolus

Low

Otidae

Great bustard *

Otis tarda

Low

Paridae

Great tit

Parus major

Low

Passeridae

Rock sparrow

Petronia petronia

Low

Phalacriciracidae Great cormorant

Phalacrocorax carbo

Low

Picidae

European green woodpecker

Picus viridis

Low

Great spotted woodpecker

Dendrocopos major

Low

Podicipedidae

Great crested grebe

Podiceps cristatus

Low

Rallidae

Common moorhen

Gallinula chloropus

Low

Strigidae

Eurasian eagle–owl

Bubo bubo

High

Long–eared owl

Asio otus

High

Little owl

Athene noctua

High

Tawny owl

Strix aluco

Moderate

Eurasian scops owl

Otus scops

Moderate

Short–eared owl

Asio flammeus

Low

Sylviidae

Sardinian warbler

Sylvia melanocephala

Low

Common whitethroat

Sylvia communis

Low

Turdidae

Common blackbird

Turdus merula

Low

Black redstart

Phoenicurus ochruros

Low

Common redstart

Phoenicurus phoenicurus

Low

Whinchat

Saxicola rubetra

Low

European stonechat

Saxicola rubicola

Low

Northern wheatear

Oenanthe oenanthe

Low

European robin

Erithacus rubecula

Low

Tytonidae

Western barn owl

Tyto alba

High

Upupidae

Hoopoe

Upupa epops

Low

Total roadkill number

643

Total number of species

al., 2013) and another on European rabbits (Planillo and Malo, 2013; and references therein). Roadside verges have shown to be an important refuge for abundant populations of small mammals, attracting avian and mammal predators (Ruiz–Capillas et al.,

2013). Abundant records on actively hunter species, such as the marsh harrier, the short–toed eagle and the lesser kestrel, suggest that they can also be killed while overflying roads and ditches in search of rodents and run–over vertebrates (PMVC, 2003).

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384

Aragon Spain Huesca

Zaragoza

Roads Roadkill density 1.32 2.65 3.09 4.42 5.74 7.06

Teruel

Threat categories EN VU SAH LC

100 km

Fig. 1. Spatial distribution of avian–vehicle collisions (n = 643) detected in the Aragon autonomous county road network (2012–2014). Threat categories: EN, Endangered; VU, Vulnerable; SAH, Sensitive to the habitat alteration; LC, Least concern. Fig. 1. Distribución espacial de los atropellos de aves (n = 643) detectados en la red viaria de la comunidad autónoma de Aragón (2012–2014). Categorías de amenazas: EN, en peligro de extinción; VU, vulnerable; SAH, sensible a la alteración de su hábitat; LC, preocupación menor.

Although some species, such as the Spanish imperial eagle and vultures, avoid roads with high intensity traffic, other species such as the common kestrel, the Eurasian buzzard, the booted eagle, red and black kites, and goshawks exploit roads independently of traffic intensity (Bautista et al., 2004). Therefore, the probability of these opportunistic and necrophagous species being hit seems to be higher than for other bird species. Another factor that may increase the incidence is migration (Pérez–Tris and Santos, 2004; Santos et al., 2015). For example, although some individuals are resident, a large part of the black kite population migrates to the Iberian peninsula in spring to breed and

returns to Africa in September–November (Scholerl et al., 2016), favouring the formation of roosts often located in groves near landfills and urban centres (Viñuela, 1997), and increasing roadkills during spring and summer (38 cases) rather than autumn and winter (8 cases). This species shows a high incidence of roadkills, especially if we take into account the number of months that the species occurs in Aragon. Thus, road mortality could constitute an even more relevant threat for the species during the breeding period. The red kite (7 roadkills in winter, 5 in spring, 1 in summer and 4 in autumn), classified as EN and SAH by the Spanish and Aragón laws, respectively, migrates in winter from its Central European breeding areas to

Animal Biodiversity and Conservation 41.2 (2018)

Winter

Spring

a,b,c

385

Summer

Autumn

a,b

a,c

Number of roadkills

J J A Months

Fig. 2. Temporality of avian–vehicle collisions in Aragon over three years (2012–2014). Black dots indicate the number of avian roadkills per month. The large grey dots indicate the mean and standard deviation (vertical grey lines) of the number of avian roadkills per season. Letters indicate significant differences between seasonal roadkill means (Nemenyi post–hoc test). Differences were detected between summer and autumn, but no differences were reached between winter–spring and summer–autumn. Fig. 2. Estacionalidad de los atropellos de aves en la comunidad autónoma de Aragón durante tres años (2012–2014). Los puntos negros indican el número de atropellos mortales registrados por mes. Los círculos grandes grises y las líneas verticales grises indican la media y la desviación estándar de los atropellos mortales por estación, respectivamente. Las letras indican las diferencias significativas observadas entre las medias de los atropellos mortales de las cuatro estaciones (prueba a posteriori de Nemenyi). Se observaron diferencias entre las estaciones de verano y otoño, pero no entre los períodos invierno–primavera y verano–otoño.

the Iberian Peninsula (Viñuela, 1997), although many couples are resident in Aragon (400–600 pairs; Viñuela, 2004). The incidence of road mortality in this species is not as serious as that of the black kite, but 50 % of the records occurred within in a small area of highway A–23 between Sabiñánigo and Jaca, in the province of Huesca. Temporality The months with the highest incidence of road mortality were July, August and September, the latter being the most critical month (fig. 2). This period is linked to the first flights of inexperienced young (note that most species breed in spring, and in summer juveniles are dispersing), which might be affected to a greater extent by collisions, as well as the postnuptial migration of species, such as black kite, white stork, European bee–eater or Egyptian vulture, which occurs in August and September (Pérez–Tris and Santos, 2004).

Spatial distribution and areas of high roadkill occurrence As previously mentioned, determining the magnitude of the road mortality and identifying the areas of high roadkill occurrence is complex and depends on the methodology and the scale of the study (Santos et al., 2015). Our study is a first approach to map the areas with a high incidence in Aragon. Subsequent studies should evaluate each of the areas of high roadkill occurrence in greater depth, combining different models, such as traffic flow, road crossing models depending on the species' ethology and geometric models, which allows to calculate a probability of collision for each grid (Lin, 2016). Our results indicate a distribution of heterogeneous bird–vehicle collisions in space, although it is associated with the main axes of the Aragon road network, especially to the highways AP–2, A–2, AP–68, A–68 and A–23 (figs. 1, 3). The areas of high roadkill occurrence with a high incidence are located in the Tarazona area and Pastriz–La Alfranca (7 and 19 roadkills, respectively; fig. 3).

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Aragon 1

Spain Huesca

5 7

4 5 4

10 9 1

Zaragoza

12 5

7 20 19 8

11 11

6 2 3 5 3

3 3 3

3 4

6 1

Incidence (RI) Low Moderate High Roads

Teruel

100 km

Fig. 3. Location of avian–vehicle collisions in Aragon in areas of high roadkill occurrence; the numbers therein are the number of roadkills per km index (RI) (2012–2014). Low, moderate and high incidence spots are indicated. The number of roadkills detected in each cell is also shown ('very low' incidence cells are not shown). Fig. 3. Localización de los atropellos de aves en las áreas de incidencia alta de atropellos; los números indican el índice de atropellos por kilómetro (RI) (2012–2014). Se señalan los puntos con incidencia de atropellos baja, moderada y alta. También se muestra el número de atropellos detectado en cada celda (la categoría de atropellos "muy baja" no se muestra en el mapa).

No particular high traffic density is present in either areas of high roadkill occurrence. The high number of roadkills in the area of high roadkill occurrence Pastriz–La Alfranca is probably influenced by the special attention given by wildlife police due to the proximity of two protected areas and the WRC of La Alfranca (fig. 3). Two areas of high roadkill occurrence with a moderate incidence of roadkill are located in Zaragoza province, one next to Tarazona and the other on highway A–2, close to the town of La Almunia de Doña Godina town, this latter having a high number of roadkills (11; fig. 3). Another two areas of high roadkill

occurrencewith a moderate incidence are located in the province of Huesca, one in the Monegros area (6 cases) and the second on the Pyrenean road of Benasque, with only one case in very few kilometres in that area of high roadkill occurrence, but many in that road (figs. 1, 3). Only one moderate area of high roadkill occurrence was detected in Teruel province, close to the Natural Reserve of Gallocanta (6 cases), which is an important resting and reproductive bird area (fig. 3). Elevated numbers of roads and high traffic intensity explain the presence of areas of high roadkill occurrence (Fahrig and Rytwinski, 2009; Bec-

Animal Biodiversity and Conservation 41.2 (2018)

kman and Shine, 2015). Nevertheless, our areas of high roadkill occurrence do not match the roads with a significant increase in traffic in these specific areas, suggesting that a high density of birds might be the cause of the elevated number of vehicle collisions detected. Detailed field prospection in high and moderate areas of high roadkill occurrence should be carried out to confirm this hypothesis or to test other possible causes, such as closeness of areas of high roadkill occurrence to lagoons, bird resting or reproduction areas, or presence of bird corridors. Low incidence cells appear to be associated with the Ebro Valley, demonstrating a significant presence of birds in the valley and its functionality as a biological corridor for this taxonomic group. However, these road networks are also the roads with most traffic in Aragon. Determining the points of the road network where road mortality is highest is essential to implement management measures that reduce the main effects of roads on wildlife (mortality due to collisions and habitat fragmentation). We therefore propose following management measures for the 41 areas of high roadkill occurrence detected, especially in the moderate (5) and high incidence (2) areas of high roadkill occurrence (fig. 3). Management recommendations Given that our results are extracted from a database and not from on–site sampling, we recommend monitoring roadkills in the areas of high roadkill occurrence to confirm the magnitude of vehicle collision and develop management actions. Given that scavengers are killed when they feed or look for food on the roads, we suggest roadkill carcasses be removed from roads, verges and gutters as soon as possible to minimise the likelihood of collisions. The finding of two species (great bittern and great bustard) catalogued as in danger of extinction by Spanish law (the maximum threat category) emphasises the need to monitor the locations where these roadkills were found (N–232, PK.138–141, and N–II, Bujaraloz–Pina de Ebro) to determine whether these road sections constitute a serious threat to the populations. Road signs for drivers should be reinforced in the areas of high roadkill occurrence. Acknowledgements We thank Manuel Alcántara, María Cortés, Chabier González, José Manuel Sánchez and Juan Fernández for their help in multiple ways. We also thank the Dirección General de Conservación del Medio Natural (Departamento de Agricultura, Ganadería y Medio Ambiente, Gobierno de Aragón) for allowing access to information concerning birds admitted to the La Alfranca Wildlife Rehabilitation Center. References Beckmann, C., Shine, R., 2015. Do the Numbers and Locations of Road–Killed Anuran Carcasses

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Accurately Reflect Impacts of Vehicular Traffic? Journal of Wildlife Management, 79: 92–101. Bishop, C. A., Brogan, J. M., 2013. Estimates of avian mortality attributed to vehicle collisions in Canada. Avian Conservation and Ecology, 8(2): 2. Blanco, G., Viñuela, J., 2004. Milano Negro, Milvus migrans. In: Libro Rojo de las Aves de España: 116–119 (A. Madroño, C. González, J. C. Atienza, Eds.). Dirección General para la Biodiversidad, SEO/BirdLife, Madrid, España. Bautista, L. M., García, J. T., Calmaestra, R. G., Palacín, C., Martín, C. A., Morales, M. B., Bonal, R., Viñuela, J., 2004. Effect of weekend road traffic on the use of space by raptors. Conservation Biology 18: 726–732. Decreto 181/2005, de 6 de septiembre, del Gobierno de Aragón, por el que se modifica parcialmente el Decreto 49/1995, de 28 de marzo, de la Diputación General de Aragón, por el que se regula el Catálogo de Especies Amenazadas de Aragón. Boletín Oficial de Aragón nº 114, de 23 de septiembre de 2005: 11527–11532. Espinosa, A., Serrano, J. A., Montori, A., 2012. Incidencia de los atropellos sobre la fauna vertebrada en el Valle de El Paular. LIC “Cuenca del río Lozoya y Sierra Norte”. Munibe–Ciencias Naturales, 60: 209–236. Fahrig, L., 2003. Effects of habitat fragmentation on biodiversity. Annual Review of Ecology, Evolution and Systematics, 34: 487–515. Fahrig, L., Rytwinski, T., 2009. Effects of roads on animal abundance: an empirical review and synthesis. Ecology and Society, 14: art. 21. Fuellhaas, U., Klemp, C., Kordes, A., Ottersber, H., Pirmann, M., Thiessen, A., Tshoetschel, C., Zucchi, H., 1989. Untersuchungen zum Strassentod von Vögeln, Säugetieren, Amphibien und Reptilien. Beiträge Naturkunde Niedersachsens, 42: 129–147. Garriga, N., Santos, X., Montori, A., Richter–Boix, A., Franch, M., Llorente, G. A., 2012. Are protected areas truly protected? The impact of road traffic on vertebrate fauna. Biodiversity and Conservation, 21: 2761–2774. Haigh, A., 2012. Annual patterns of mammalian mortality on Irish roads. Hystrix, the Italian Journal of Mammalogy, 23: 58–66. Hernández, F., 2009. El buitre leonado en Aragón. In: El buitre leonado en España. Población reproductora en 2008 y método de censo: 51–60 (J. C. Del Moral, Ed.). SEO/BirdLife, Madrid, España. Lin, S. C., 2016. Landscape and traffic factors affecting animal road mortality. Journal of environmental engineering and landscape management, 24: 10–20. Loss, S. R., Will, T., Marra, P. P., 2014. Estimation of bird–vehicle collision mortality on US roads. Journal of wildlife management, 78: 763–771. Martínez–Freiría, F., Brito, J., 2012. Quantification of road mortality for amphibians and reptiles in Hoces del Alto Ebro y Rudrón Natural Park in 2005. Basic and Applied Herpetology, 26: 7–16. Nankinov, D. N., Todorov, N. M., 1983. Bird casualties on highways. Soviet Journal of Ecology, 14:

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288–293. Oschadleus, D. H., Harebottle, D. A., 2002. Survey of road–kills, with special emphasis of bird. Bird Numbers, 11: 42–44. Pérez–Tris, J., Santos, T., 2004. El estudio de la migración de aves en España: trayectoria histórica y perspectivas de futuro. Ardeola, 51: 71–89. Planillo, A., Malo, J. E., 2013. Motorway verges: Paradise for prey species? A case study with the European rabbit. Mammalina Biology, 78: 187–192. Planillo, A., Kramer–Schadt, S., Malo, J. E., 2015. Transport Infrastructure Shapes Foraging Habitat in a Raptor Community. PLOS One, 10(3): e0118604, doi: 10.1371/journal.pone.0118604 PMVC., 2003. Proyecto de Mortalidad de Vertebrados en Carreteras. Documento técnico de conservación nº 4. Sociedad para la Conservación de los Vertebrados (SCV), Madrid, España. Rico–Guzmán, E., Cantó, J. L., Terrones, B., Bonet, A., 2011. Impacto del tráfico rodado en el Parque Natural del Carrascal de la Font Roja. ¿Cómo influyen las características de la carretera en los atropellos de vertebrados? Galemys, Spanish Journal of Mammalogy, 23: 113–123. Ruiz–Capillas, P., Mata, C., Malo, J. E., 2013. Road verges are refuges for small mammal populations in extensively managed Mediterranean landscapes. Biological Conservation, 158: 223–229. Sáenz–de–Santa–María, A., Tellería, J. L., 2015. Wildlife–vehicle collisions in Spain. European Journal of Wildlife Research, 61: 399–406. Santos, S. M., Marques, J. T., Lourenco, A., Medinas, D., Barbosa, A. M., Beja, P., Mira, A., 2015. Sampling effects on the identification of roadkill

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hotspots: Implications for survey design. Journal of environmental management, 162: 87–95. Saranholi, B. H., Bergel, M. M., Ruffino, P. H. P., Rodriguez, K. G., Ramazzotto, L. A., de Freitas, P. D., Galetti, P. M., 2016. Roadkill hotspots in a protected area of Cerrado in Brazil: planning actions to conservation. Revista MVZ Cordoba, 21: 5441–5448. Scholerl, M. N., Martin, B., Ferrer, M., Onrubia, A., Bechard, M. J., Kaltenecker, G. S., Carlisle, J. D., 2016. Variable shifts in the autumn migration phenology of soaring birds in southern Spain. Ardea, 104: 83–93. Schrijver, B., 1993. Nederlands verkeer rijdt jaarlijks. Vogels, 1: 16–17. Tenés, A., Cahill, S., Llimona, F., Molina, G., 2007. Atropellos de mamíferos y tráfico en la red viaria de un espacio natural en el área metropolitana de Barcelona: quince años de seguimiento en el parque de Collserola. Galemys, Spanish Journal of Mammalogy, 19: 169–188. Vidal–Vallés, D., Pérez–Collazos, E., 2016. Incidencia de atropellos de mamíferos silvestres no cinegéticos en la red viaria de la Comunidad Autónoma de Aragón (2012–2014). Lucas Mallada, 18: 47–66. Viñuela, J., 1997. Road transects as a large–scale census method for raptors: The case of the Red Kite Milvus milvus in Spain. Bird study, 44: 155–165. – 2004. Milano real, Milvus milvus. In: Libro Rojo de las Aves de España: 120–125 (A. Madroño, C. González, J. C. Atienza, Eds.). Dirección General para la Biodiversidad, SEO/BirdLife, Madrid, España.

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MHC class II DRB variability in wild black howler monkeys (Alouatta pigra), an endangered New World primate L. E. Arguello–Sánchez, J. R. Arguello, L. M. García–Feria, C. A. García–Sepúlveda, D. Santiago–Alarcon, A. Espinosa de los Monteros Arguello–Sánchez, L. E., Arguello, J. R., García–Feria, L. M., García–Sepúlveda, C. A., Santiago–Alarcon, D., Espinosa de los Monteros, A., 2018. MHC class II DRB variability in wild black howler monkeys (Alouatta pigra), an endangered New World primate. Animal Biodiversity and Conservation, 41.2: 389–404. Abstract MHC class II DRB variability in wild black howler monkey (Alouatta pigra), an endangered New World primate. The genes of the major histocompatibility complex (MHC) are the most important genetic component of the immune system in vertebrates. Their variability is known to influence a species' ability to recognize and respond to pathogens. Here, we present the first data of the MHC class II DRB exon 2 for the endangered black howler monkey (Alouatta pigra), one of the most northerly distributed platyrrhines. Twenty–one DRB sequences corresponding to four new lineages were identified in 44 individuals through a combination of cloning and reference strand conformational analysis. The detection of up to eight sequences per individual suggests the existence of at least four loci in the species. A relatively low DRB sequence diversity, but similar lineage and loci numbers. were found in A. pigra when compared to other platyrrhines. The reduced DRB allelic diversity in the species appears to be a consequence of drift, reflecting the colonization by its ancestors from South to Central America. Finally, the allelic diversity in the species might be enabling an adequate immune response in wild populations to cope with current pathogens, but it might entail a risk for these populations in case of the emergence of new pathogens. Key words: Major histocompatibility complex, Polymorphism, Howler monkey, Genetic bottleneck Resumen Variabilidad del gen DRB del CMH de clase II en el mono aullador negro (Alouatta pigra), un primate del Nuevo Mundo en peligro. Los genes del complejo mayor de histocompatibilidad (CMH) son el componente genético más importante del sistema inmunitario en vertebrados. Su variabilidad puede influir en la habilidad de una especie para detectar patógenos y reaccionar ante ellos. En este trabajo presentamos por primera vez datos relativos al exón 2 del gen DRB del MHC de clase II del mono aullador negro (Alouatta pigra), una de las especies de platirrinos con una distribución más septentrional. Se identificaron 21 secuencias de DRB, pertenecientes a cuatro nuevos linajes en 44 individuos de A. pigra. Se detectaron hasta ocho secuencias por individuo, lo que sugiere la existencia de al menos cuatro loci en la especie. En comparación con otros platirrinos, en A. pigra se observó una diversidad relativamente baja de secuencias de DRB, pero un número parecido de linajes y loci. La reducida diversidad alélica del gen DRB en la especie parece ser consecuencia de deriva génica, lo que refleja la colonización por parte de sus ancestros desde América del Sur hacia Centroamérica. Por último, la diversidad alélica de la especie está permitiendo que las poblaciones silvestres muestren una respuesta inmunitaria adecuada frente a los patógenos actuales, pero podría implicar un riesgo para esas poblaciones en caso de que surjan otros nuevos. Palabras clave: Complejo mayor de histocompatibilidad, Polimorfismo, Mono aullador, Cuello de botella genético Received: 23 X 17; Conditional acceptance: 02 II 18; Final acceptance 20 II 18 Laura Elisa Arguello Sánchez, Luis Manuel García Feria, Diego Santiago Alarcon, Alejandro Espinosa de los Monteros, Inst. de Ecología, A. C. Carretera Antigua a Coatepec #351, El Haya, 91070 Xalapa, Veracruz, México.– J. Rafael Arguello, Fac. de Medicina Unidad Torreón, Univ. Autónoma de Coahuila, Av. Morelos #900 Ote, C.P. 27000 Torreón, Coahuila, México.– Christian Alberto García Sepúlveda, Lab. de Genómica Viral y Humana, Fac. de Medicina, Univ. Autónoma de San Luis Potosí, Avenida Carranza #2405, Colonia Filtros Las Lomas, 78210 San Luis Potosí, SLP, México. Corresponding author: L. E. Arguello–Sánchez. E–mail: elisillos@hotmail.com ISSN: 1578–665 X eISSN: 2014–928 X

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Introduction The major histocompatibility complex (MHC) is the most functionally important genetic region involved in infectious disease resistance among jawed vertebrates (Hedrick, 1994). Classical class I and II genes encode for cell surface glycoproteins that bind antigens and present them to effector cells, triggering an adaptive immune response (Hedrick, 1994; Knapp, 2005b). Class I molecules are expressed on the surface of all nucleated cells, and present peptides of intracellular origin (such as those naturally produced by a cell as well as those originating from a virus) to cytotoxic T cell lymphocytes (CD8+ T cells), targeting infected cells for destruction. Class II molecules are only expressed on specialized antigen–presenting cells, such as macrophages, dendritic cells and B cells, and present antigens with an extracellular origin (such as those derived from cellular pathogens) to helper T lymphocytes (CD4+ T cells), initiating the antigen–specific immune response (Janeway et al., 2001). The MHC is characterized by a particularly high polymorphism, which is mainly located at residues involved in antigen binding (antigen binding sites, ABS) (Hedrick, 1994). Such diversity appears to arise from pathogen–driven balancing selection (Ujvari & Belov, 2011; van Oosterhout, 2009). This selective force is thought to operate through the heterozygote advantage (i.e. overdominance; Doherty and Zinkernagel, 1975) or through the emergence of rare alleles (i.e. negative frequency–dependent selection; Takahata & Nei, 1990). Nevertheless, other mechanisms like mating preferences, maternal–fetal interactions and olfactory–based markers for kin recognition and inbreeding avoidance, have also been proposed to play a significative role in MHC polymorphism (Ekblom et al., 2010; Ujvari and Belov, 2011; Penn and Potts, 1999). Although most natural vertebrate populations that have been studied exhibit high levels of MHC diversity (Klein, 1986), some species show a limited variation (Sommer et al., 2002; Seddon and Baverstock, 1999). This indicates that balancing selection may not always play the most important role in shaping MHC diversity in natural populations, and that neutral forces, such as genetic drift and gene flow, might also influence variation in these functionally important genes (Miller and Lambert, 2004). It has been proposed that low MHC polymorphism in some cases can be explained by reduced selection pressure due to a solitary lifestyle, or low exposure to pathogens in high latitudes or in marine environments (Mainguy et al., 2007; Caballero et al., 2010; Ellegren et al., 1996). Also, low MHC polymorphism has been suggested as a consequence of a restrictive mating system, a small population size, or past population bottleneck (Sommer et al., 2002; Bollmer et al., 2011; Hedrick et al., 2000). In such cases, the strength of genetic drift may exceed that of balancing selection (Miller and Lambert, 2004). There is growing evidence showing a correlation between MHC gene variability and disease resistance (Bernatchez and Landry, 2003; Froeschke, 2005; Westerdahl et al., 2005; Worley et al., 2010). However, the consequences of a low MHC variation for most wildlife are not clear (Radwan et al., 2010; Bollmer

Arguello–Sánchez et al.

et al., 2011), and it has been hypothesized that low levels of variation might increase susceptibility to infectious diseases and limit the capacity of populations to respond to emerging pathogens. This possibility is of particular concern for endangered species (O’Brien et al., 1985; Lafferty and Gerber, 2002; Siddle et al., 2007; Zhang et al., 2015). Transmittable virus–associated cancers such as Tasmanian Devil Facial Tumour Disease and Canine Transmissible Venereal Tumour are examples of the relevance of MHC variation in conservation biology, since both diseases are assumed to have emerged and disseminated due to low MHC diversity in Tasmanian devil, wolf and dog populations (Siddle et al., 2007; Belov, 2011; McCallum, 2008); some populations of Tasmanian devils have been driven to a rapid decline (Belov, 2011). Assessing the degree of polymorphism of MHC can thus provide critical baseline information on a species’ immune response and adaptive potential. In the case of non–human primates, many of which are classified as endangered species, the development of standardized techniques for MHC studies using non–invasive samples (e.g., faeces, hair) has increased our possibility to target vulnerable wild populations (Lukas et al., 2004; Knapp, 2005a; Müller et al., 2014; Hans et al., 2015; Lukas and Vigilant, 2005). Nonetheless, current knowledge of MHC diversity from wild New World monkey (NWM) populations is restricted to isolated studies of the family Cebidae (Nino–Vasquez et al., 2000; Gyllensten et al., 1994; Suárez et al., 2006). In this study, non–invasive samples were used to characterize the polymorphism at MHC class II loci from wild A. pigra, a representative species of the Atelidae family and one of the most northerly distributed platyrrhines. A. pigra is distributed in a relatively small area in the Yucatan Peninsula of Mexico, Belize and central Guatemala (Rylands et al., 2005). This species is classified as 'Endangered' (Marsh et al., 2008), mainly due to habitat destruction, illegal hunting and the pet trade. Furthermore, the fast and continuous loss of their habitat and human encroachment is increasing their exposure to both human and domestic animal pathogens (Vitazkova and Wade, 2006; Nunn and Gillespie, 2016; Martínez–Mota et al., 2015), raising the risk of emerging infectious diseases. This highlights the importance of the assessment of immune gene variability in the species for its inclusion in conservation strategies. Analyses of genetic variation on neutral markers (i.e., microsatellites and mtDNA) have detected low levels of variation in A. pigra compared to other Alouatta species distributed further south (Amendola, 2009; Van Belle et al., 2012; James et al., 1997). These low levels of variation have been mainly attributed to a founder effect due to the colonization of A. pigra ancestors from South to Central America (Ellsworth and Hoelzer, 2006; Cortés–Ortiz et al., 2003). Contrary to neutral markers, MHC genes are thought to be under balancing selection. It is therefore possible that variation in these genes is retained in species that have undergone a bottleneck like A. pigra (Aguilar et al., 2004). Nevertheless, in numerous cases, low variation at neutral loci has been

Animal Biodiversity and Conservation 41.2 (2018)

correlated with reduced MHC diversity, indicating that balancing selection had been overpowered by genetic drift (Bollmer et al., 2011; Miller and Lambert, 2004). In this study, we assessed the level of variation on exon 2 of MHC–DRB loci of wild A. pigra, in terms of its allelic, lineage, and loci polymorphism. Our aim was to to generate baseline information on immune gene diversity for this endangered NWM species. Nonetheless, we also intended to determine whether the evolutionary forces shaping the neutral genetic variation in this species had also influenced the variation in genes linked to adaptive immunity like the MHC, or whether balancing selection had managed to counteract the effects of genetic drift. We chose to study the MHC–DRB since it is the most widely studied MHC loci in NWM and other non–human primates (De Groot et al., 2012), thus allowing a better comparison of MHC diversity. Furthermore, we focused on exon 2 of DRB loci, which encodes most of the ABS and has a primary immune function, making it relevant for conservation purposes (Ujvari and Belov, 2011; Hedrick, 1994; Zhang et al., 2015).

391

Table 1. Number of faecal samples of Alouatta pigra individuals (Ni); number of social groups (Nsg) collected from June to November 2014 at each locality; and number of successfully analyzed samples (Nsas): PNP, Palenque Nacional Park; CBR, Calakmul Biosphere Reserve; and CQR, Coba Quintana Roo. Tabla 1. Número de muestras fecales de individuos (N); número de grupos sociales (Ngs) de Alouatta pigra recogidas entre junio y noviembre de 2014 en cada localidad; y número de muestras analizadas (Nsas): PNP, Parque Nacional de Palenque; CBR, Reserva de la Biosfera de Calakmul; CQR, Cobá Quintana Roo. Population

CBR

7 23 23

Nsg

Nsas

PNP

6 20 17

CQR 2 6 4 Material and methods

Total

15 49 44

Sample collection Considering the conservation status of the species and the possible health risks of a chemical immobilization procedure to obtain blood samples, we chose to use a non–invasive faecal DNA sampling approach for this study. A total of 49 faecal samples from adult animals belonging to 15 troops and two solitary individuals were collected from June to November 2014 (table 1). Sampling took place at three different sites: (1) Calakmul Biosphere Reserve (CBR) in the state of Campeche (18º 36' 43'' N, 89º 32' 53'' W), the largest Mexican tropical forest having a surface area in excess of 720,000 ha; (2) Palenque National Park (PNP) in the state of Chiapas (17º 29' 00'' N, 92º 03' 00'' W), a protected natural area of 1,772 ha; and (3) a seven hectare patch close to Coba, Quintana Roo (CQR) of low and medium subdeciduous forest (20º 28' 51.58'' N, 87º 44' 18.40'' W), see figure 1. Monkey troops were visually tracked and closely monitored to allow expedite collection of fresh faecal samples from well–identified and sexed adult individuals. Faecal samples were collected using single–use gloves that were changed between samples to avoid cross–contamination. Approximately 2 g of stool sample were placed in sterile tubes with 3 ml of RNAlater solution (Ambion, Austin, TX, USA) and stored at ambient temperature for a maximum of 18 days. Once in the laboratory, they were stored at –20 ºC until DNA extraction. Genotyping of MHC–DRB by cloning–RSCA and sequencing DNA was extracted from approximately 15 mg of stool sample using the Fecal DNA Extraction Kit (Zymo Research, Irving, California, USA) following the manu-

facturer’s instructions. To amplify a 269 bp of exon 2 we used the primers DRBP1 (5' CCGGATCCTTCGTGTCCCCACAGCACG 3') and DRBP2 (5' TCGCCGCTGCACTGTGAAG 3') (Tiercy et al., 1990). Primer selection was based on the alignment of DRB exon 2 sequences from all known NWM lineages, including species most closely related to Alouatta (i.e., families Atelidae and Pitheciidae). PCR amplification was carried out in a total volume of 25 µl with 20 pmol of each primer, 200 µM of dNTPs, 2 mM MgCl2, 1 U of GoTaq (Promega, Madison, WI, USA) and 4 µl of DNA extract. In order to minimize the differences in the effectiveness of allele amplifications, a 'Touchdown' protocol for the PCR was implemented (Hans et al., 2015). Thermocycling conditions included an initial denaturalization step of 94 ºC for 5 min, followed by two cycles of denaturing 94 ºC for 30s, annealing at 68 ºC for 30s, extension at 72 ºC for 30s, two cycles of 94 ºC for 30s, 63 ºC for 30s, and 72 ºC for 30s, and 30 cycles of 94 ºC for 30s, 58 ºC for 30s, and 72 ºC for 30s, with a final extension step of 72 ºC for 10 min. Products were visualized and documented in 1.5 % agarose gels. For DRB genotyping, a combination of two molecular methods was used: the classic cloning–sequencing approach, and a Reference Strand Conformational Analysis (RSCA) strategy (Argüello et al., 1998). Both techniques achieve high resolution for the genotyping of human and non–human MHC (Drake et al., 2004; Babik, 2010; Strand and Höglund, 2011; Oppelt and Behrmann–Godel, 2012). However, given the source of DNA samples, we found amplicon cloning yielded better results for detecting all DRB sequence variants within a sample (data not shown). In order to avoid sequencing a great number of identical clones,

Arguello–Sánchez et al.

392

we used RSCA as a screening tool to differentiate sequence variants among the obtained clones, thus reducing the set of clones subjected to nucleotide sequencing. To ensure reliability of genotyping result, two or three independent amplification products from each individual sample were cloned into a pGEM–T easy vector (Promega, Madison, WI, USA) and the ligation mix was transformed into Top10 cells (Invitrogen) using heat shock. Verification of the cloned fragment size relied on subsequent PCR amplification using primers DRBP1 and DRBP2; those having the expected size were then subjected to RSCA. RSCA allows the differentiation of sequence variants present in amplicons that are hybridized to a fluorescently labelled pre–sequenced reference strand (FLR). The electrophoretic mobility of the heteroduplex is affected by mismatches, allowing the identification of different sequence variants. For this study, we used two FLRs of human origin (HLA–DRB1*10:01:01 and HLA–DRB3*01:16) so as to better ensure that no homoduplex would be formed when hybridizing with the A. pigra samples (Argüello et al., 1998). Lastly, we selected three to five clones (from different individuals) for each putative allelic variant identified and subjected them to bidirectional Sanger sequencing at LANBAMA, Mexico and Macrogen, South Korea using M13 universal primers. Sequence analysis and artefact detection The alignment and edition of sequences were accomplished in the sequencer 5.0.1 software (Gen Codes Corporation). Visual inspection of sequences was performed to detect possible base calling errors based on the consensus sequence. All sequences were translated into protein sequences to verify whether there was evidence of pseudogenes, such as the presence of premature stop codons or indels. Samples exhibiting information suggestive of chimeric sequence artefacts (resulting from template switching by the DNA polymerase) were detected using six simultaneous analyses in RDP 4.56 (Martin et al., 2015): RDP, MaxChi2, BootScan, Chimaera, Siscan and 3Seq (Martin and Rybicki, 2000; Smith, 1992; Posada and Crandall, 2001; Martin et al., 2005; Gibbs et al., 2000; Boni et al., 2007). Finally, we considered true alleles as those present in at least two different individuals or independent PCR products from the same individual (Cutrera and Lacey, 2006; Real–Monroy et al., 2014). Statistical and phylogenetic analyses To assess variation at the DRB gene, we considered the number of alleles, number of lineages, nucleotide diversity (π), and number of segregating sites (S). These calculations were made using DnaSP 5.1 software (Rozas, 2009). Since the primers used may amplify multiple copies of DRB in the samples, a phylogenetic analysis was carried out to identify groups of sequences that potentially represent different loci, based on the assumption that alleles of a particular MHC locus have a tendency

to form a phylogenetic cluster (Oppelt et al., 2010; Lukas et al., 2004). A Bayesian inference analysis was implemented using Mr. Bayes 3.2.6 (Ronquist et al., 2012). The best nucleotide substitution model (HKY+I) was selected using the Akaike information criterion in the jModelTest software (Posada, 2009). Markov chain Monte Carlo (MCMC) were run for 15 million generations, sampling trees and parameters every 1,000 generations. Analysis consisted of two independent runs with one cold and three hot chains. The majority–rule consensus tree was obtained, with their respective posterior probabilities, after discarding the initial 25 % of the accumulated trees. Sequences representing the 35 DRB lineages from platyrrhine species were included in the analysis. In addition, sequences with the highest degree of homology available in the IPD–MHC NHP database (De Groot et al., 2012) from catarrhine species were also included. A human DRB sequence (HLA–DRB1*01:01:01) was used as the outgroup. Lastly, to detect molecular level evidence of selection on DRB, the relative rates of synonymous substitutions (dS) and non–synonymous substitutions (dN) were calculated using the Nei and Gojobori with the Jukes and Cantor correction in DnaSP 5 (Rozas, 2009). Calculations were performed for the antigen binding sites (ABS) [extrapolated from those described in humans (Brown et al., 1993)] and non–ABS sequence regions for each cluster of A. pigra DRB sequences (putative loci) and for all sequences. Finally, a Z–test was carried out to evaluate departure from neutrality for the dN/dS ratios. Tests were executed in Mega 7.0.2 (Kumar et al., 2016). Results From a total of 49 samples of A. pigra, amplification and cloning of DRB exon 2 PCR products of 44 individuals was achieved (table 1). RSCA analysis of clones (20 to 42 clones per sample) identified 33 DRB sequence variants. Twelve of these variants were discarded from subsequent analyses as ten of them were only found in one or two clones from a single amplification of one individual. These discarded sequences showed a substitution of one or two nucleotides relative to true DRB sequences present in the same sample (eight variants), or they had a deletion of a single nucleotide at position 61 or 117 that shifted the reading frame (two variants). The other two sequence variants were found to be in vitro recombinants, having as parental sequences Alpi–DRB*W107:04 and Alpi–DRB*W105:05 (recombinant 1), Alpi–DRB*W105:04 and Alpi–DRB*W106:04 (recombinant 2). Finally, all sequence variants considered true alleles (N = 21) were identified in clones from two or more independent amplifications of one individual (Alpi–DRB3*06:09), or in clones of more than two individuals (table 2). DRB variation Based on the nucleotide sequences, 21 DRB alleles were detected in the analysed samples of A. pigra.

Animal Biodiversity and Conservation 41.2 (2018)

88º 00' W

Yucatan

Campeche

22º 00' N

93º 00' W

393

CQR

Quintana Roo

CBR

Gulf of Mexico

Mexico

PNP

17º 00' N

Tabasco

Belize

Chiapas

Guatemala

Pacific Ocean

Honduras 0

93º 00' W

100

200 km

88º 00' W

Fig. 1. Location of the three sites where fecal samples of Alouatta pigra were collected from June to November 2014. Sites are marked with stars: PNP, Palenque National Park; CBR, Calakmul Biosphere Reserve; and CQR, Coba Quintana Roo. (The light–shaded area indicates the putative distribution range of the species.) Fig. 1. Ubicación de las tres localidades donde se recogieron las muestras fecales de Alouatta pigra entre junio y noviembre de 2014. Las localidades están marcadas con estrellas: PNP, Parque Nacional de Palenque; CBR, Reserva de la Biosfera de Calakmul; CQR, Cobá Quintana Roo. (El área con sombreado claro indica el rango de distribución supuesto de la especie).

These alleles had a 93 % to 95.9 % nucleotide similarity to DRB sequences from other primates registered in GenBank, mainly from platyrrhine species (e.g. Aotus nancymaae, Callithrix jacchus, and Cebus apella). A total of 45 segregating sites and a general nucleotide diversity of 0.057 were found (table 3). The pairwise difference among A. pigra alleles ranged from 1 to 33 bp, with a mean of 15.26 differences along the 269 bp fragment. The 21 alleles translated to 20 different amino acid sequences as the difference between two alleles is a single synonymous substitution (fig. 2). Within the 89 amino acid positions, 27 (30.34 %) were found to be variable. All alleles were submitted to GenBank/IPD–MHC databases (accession numbers: MF136723–MF136743) and were named by the IPD–MHC NHP nomenclature committee (De Groot et al., 2012). Since it was not clear whether alleles belonged to different lineages from a single or multiple loci, A. pigra alleles were designated as W (workshop): Alpi–DRB*W105,

DRB*W106, DRB*W107, and DRB*W108 (table 4). No insertions or deletions were observed within the true allele set, and when translated, no premature stop codons were revealed. Therefore, no evidence suggesting the presence of pseudogenes was found. In the Bayesian analysis, all A. pigra alleles were recovered into three clusters with moderate and high support values (fig. 3). The four alleles from cluster I were closely related to lineages DRB5*03 and DRB*W37 from Macaca spp., as well as several other lineages from NWM species (e.g. DRB11*01, DRB*W38, DRB*W12 and DRB*W19). Cluster II contained alleles from A. pigra from lineages DRB*W105 and DRB*W108, and was the sister group to Aotus nancymaae (Aona–DRB*W89:01). The third cluster corresponded to alleles from lineage DRB*W106. For this cluster, the relationship between the sequences was unclear. The number of alleles detected per individual ranged from two to eight (table 2), with a mean of 4.75 alleles (table 3). Most of the individuals ana-

Arguello–Sánchez et al.

394

Table 2. MHC–DRB exon 2 sequences detected per Alouatta pigra individual analyzed. The last letter in an individual's code indicates gender (M, male; F, female). The names of the alleles correspond to those assigned by the IPD–MHC NHP nomenclature committee (De Groot et al., 2012): Indiv, individuals; N, number of alleles. Tabla 2. Secuencias del exón 2 del gen DRB del CMH detectadas por individuo analizado de Alouatta pigra. La ultima letra del código del individuo indica el sexo (M, macho; F, hembra). Los nombres de los alelos corresponden a los asignados por el Comité de Nomenclatura del IPD–MHC NHP (de Groot et al., 2012): Indiv, individuos; N, número de alelos.

Alpi–DRB*W107:04

Alpi–DRB*W107:03

Alpi–DRB*W107:02

Alpi–DRB*W107:01

Alpi–DRB*W106:09

Alpi–DRB*W106:08

Cluster I Alpi–DRB*W106:07:02

Alpi–DRB*W106:07:01

Alpi–DRB*W106:06

Alpi–DRB*W106:05

Alpi–DRB*W106:04

Alpi–DRB*W106:03

Alpi–DRB*W106:02

Alpi–DRB*W106:01

Alpi–DRB*W108:01

Cluster III Alpi–DRB*W105:06

Alpi–DRB*W105:05

Alpi–DRB*W105:04

Alpi–DRB*W105:03

Alpi–DRB*W105:02

Alpi–DRB*W105:01

Cluster II

Indiv N IS01M + + + + + + + 7 IS02H + + + + + + 6 IS03H * + + + + + + 7 IR01H + + + 3 IR02M

+ + + + + + 6

IR03H

+ + + + 4

IR04H + + + + + 5 IR05H + + + + + 5 IR06M + + + + + + 6 K261M + + + + + + + 7 K262M + + + 3 K263H + + 2 K264H

+ + + + + + + 7

K265H + + + 3 DR01H

+ + + + + + + + 8

DR02M + + + + 4 DR03H + + + 3 ZA21M

+ + + + + + + 7

ZA22H

+ + + + + 5

ZA11H

+ + + + + 5

PU01M

+ + + + + 5

PU02H + + + + 4 PU03H + + + + + 5 FM01M + + + + + 5 FM02H

+ + + + + + + 7

FM03H + + + 3 FM04H + + 2 CP01H + + + + + + 6 CP02H + + 2

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395

Table 2. (Cont.)

Alpi–DRB*W107:04

Alpi–DRB*W107:03

Alpi–DRB*W107:02

Alpi–DRB*W107:01

Alpi–DRB*W106:09

Alpi–DRB*W106:08

Alpi–DRB*W106:07:02

Alpi–DRB*W106:07:01

Cluster I

Alpi–DRB*W106:06

Alpi–DRB*W106:05

Alpi–DRB*W106:04

Alpi–DRB*W106:03

Alpi–DRB*W106:02

Alpi–DRB*W106:01

Alpi–DRB*W108:01

Cluster III Alpi–DRB*W105:06

Alpi–DRB*W105:05

Alpi–DRB*W105:04

Alpi–DRB*W105:03

Alpi–DRB*W105:02

Alpi–DRB*W105:01

Cluster II

Indiv N AP01M + + + 3 AP02H

+ + + + 4

AP03M + + + + 4 CN01M

+ + + + + 5

CN02M + + + + 4 CN03H + + + + 4 CN04H

+ + + + + + 6

CN05H + + + + + + 6 PC01M + + + + + + + 7 P01H + + + + + 5 EM01H + + 2 CU01M + + + + + 5 CU02H + + + + + 5 CU03H + + + + + 5 CU04H + + 2

Table 3. MHC–DRB polymorphism detected in faecal samples of Alouatta pigra from: PNP, Palenque National Park; CBR, Calakmul Biosphere Reserv; CQR, Coba Quintana Roo; N, number of individuals; π, nucleotide diversity; S, segregating sites; MAI, mean number of alleles per individual. Tabla 3. Polimorfismo del gen DRB del MHC detectado en muestras fecales de Alouatta pigra de: PNP, Parque Nacional de Palenque; CBR, Reserva de la Biosfera de Calakmul; CQR, Cobá Quintana Roo; N, número de individuos; π, diversidad de nucleótidos; S, sitios segregados; MAI, número promedio de alelos por individuo. Locality

No. alleles

No. lineages

MAI 5.09

CBR 23

0.057

PNP 17

0.057

44 4.41

CQR 4

0.055

39 4.25

Total 44

0.057

4.75

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396

Fig. 2. Amino acid sequence alignment of the 21 DRB alleles of Alouatta pigra. The numbers at the top indicate the amino acid position based on HLA–DRB1*01:01:01 allele. The shaded columns correspond to antigen binding sites deduced from human positions. Fig. 2. Alineamiento de secuencias de aminoácidos de los 21 alelos DRB de Alouatta pigra. Los números en el margen superior indican la posición de los aminoácidos basada en el alelo HLA–DRB1*01:01:01. Las columnas sombreadas corresponden a los sitios de unión al antígeno deducidos a partir de las posiciones descritas en humanos.

Table 4. Alouatta pigra MHC–DRB alleles and lineages detected per locality: PNP, Palenque Nacional Park; CBR, Calakmul Biosphere Reserve; CQR, Coba Quintana Roo. Tabla 4. Alelos y linajes del gen DRB del MHC de Alouatta pigra detectados por localidad: PNP, Parque Nacional de Palenque; CBR, Reserva de la Biosfera de Calakmul; CQR, Cobá Quintana Roo.

Locality

Lineage Allele

CBR PNP CQR N = 23 N = 17 N = 4

DRB*W105

Locality

Lineage Allele

CBR PNP CQR N = 23 N = 17 N = 4

Alpi–DRB*W106:06 +

Alpi–DRB*W105:01 +

Alpi–DRB*W105:02 +

Alpi–DRB*W105:03 +

Alpi–DRB*W105:04 +

Alpi–DRB*W105:05 +

Alpi–DRB*W105:06 +

Alpi–DRB*W106:07:01 + + +

+ +

Alpi–DRB*W106:07:02 + Alpi–DRB*W106:08 +

Alpi–DRB*W106:09 +

DRB*W107 Alpi–DRB*W107:01 +

DRB*W106

Alpi–DRB*W107:02 +

Alpi–DRB*W106:01 +

Alpi–DRB*W106:02 +

Alpi–DRB*W106:03 +

Alpi–DRB*W106:04 +

Alpi–DRB*W106:05 +

Alpi–DRB*W107:03 +

Alpi–DRB*W107:04 +

DRB*W108 Alpi–DRB*W108:01 + Total

21 18 8

Animal Biodiversity and Conservation 41.2 (2018)

Ceap–DRB3*06:01 Aona–DRB*W91:01 Aona–DRB*W13:01 0.96 Aona–DRB*W41:01 1 Aona–DRB*W30:01 1 Aovo–DRB*W92:01 1 Aona–DRB*W44:01 Aoni–DRB*W43:01 Aona–DRB*W45:01 0.99 Aona–DRB*W46:01 Aona–DRB*W47:01 0.85 Aona–DRB*W29:01 0.64 Aona–DRB*W42:01 0.99 Aovo–DRB*W90:01 Aovo–DRB*W88:01 0.99 Alpi–DRB*W107:01 Alpi–DRB*W107:02 0.99 Cluster 1 Alpi–DRB*W107:03 Alpi–DRB*W107:04 0.79 Caja–DRB*W16:01 0.87 Camo–DRB*W17:01 Ceap–DRB*W15:01 0.75 Saoe–DRB*W22:01 Saoe–DRB11*01:01 Caja–DRB1*03:01 0.99 Saoe–DRB1*03:01 0.90 0.95 Ceap–DRB*W32:01 Camo–DRB*W14:01 Camo–DRB11*01: 0.91 0.62

397

Sasc–DRB*W19:01 0.87 Atbe–DRB*W19:01 1 Mafa–DRB*W37:01 Mamu–DRB*W37:01 0.94 Masp–DRB5*03:03 1 Saoe–DRB5*07:01 0.84 Aoaz–DRB*W38:01 Atbe–DRB*W38:01 0.89 Alpi–DRB*W105:01 Alpi–DRB*W105:02 Alpi–DRB*W105:05 0.57 Alpi–DRB*W105:03 Cluster 2 0.7 Alpi–DRB*W105:04 0.68 Alpi–DRB*W105:06 0.53 Alpi–DRB*W108:01 Aona–DRB*W89:01 Alpi–DRB*W106:01 Alpi–DRB*W106:02 0.96 Alpi–DRB*W106:03 Alpi–DRB*W106:04 0.71 Alpi–DRB*W106:05 Cluster 3 Alpi–DRB*W106:06 Alpi–DRB*W106:08 Alpi–DRB*W107:07:01 Alpi–DRB*W107:07:02 Alpi–DRB*W106:09 1 Aona–DRB3*06:01 0.57 Aona–DRB3*06:01 0.58 Aona–DRB3*06:01 1 0.9 Aoni–DRB3*06:01 Aoaz–DRB3*06:01 Ceap–DRB*W13:01 0.99 Pipi–DRB3*05:01 1 Saoe–DRB*05:05 Camo–DRB3*05:01 0.7 Camo–DRB3*07:01 0.75 Camo–DRB1*03:01 Aona–DRB*W18:01 Aovo–DRB*W93:01 0.51 Atbe–DRB*W39:01 HLA–DRB1*01:01:01

Capy–DRB*W12:01

Fig. 3. Consensus Bayesian tree of DRB exon 2 of Alouatta pigra and selected sequences representing all known lineages of NWM. A human sequence was used as outgroup (HLA–DRB1*01:01:01): A. pigra (Alpi–DRB) sequences are shown in bold. Other sequences are: Ceap–DRB, Cebus apella; Saoe, Saguinus oedipus; Camo–DRB, Callithrix moloch; Pipi–DRB, Pithecia pithecia; Aoaz–DRB, Aotus azarai; Aoni–DRB, Aotus nigriceps; Aona–DRB, Aotus nancymaae; Atbe–DRB, Ateles belzebuth; Aovo–DRB, Aotus vociferans; Capy–DRB, Callithrix pygmaea; Caja–DRB, Callithrix jacchus; Maca–DRB, Macaca mulatta; Mafa–DRB, Macaca fascicularis; Masp–DRB, Mandrillus sphinx. The numbers above the lines correspond to Bayesian posterior probabilities. Fig. 3. Árbol bayesiano consenso del exón 2 del gen DRB de Alouatta pigra y secuencias que representan todos los linajes conocidos de platirrinos. Se usó una secuencia de humano como grupo externo (HLA–DRB1*01:01:01). Las secuencias de A. pigra (Alpi–DRB) están marcadas en negritas. Los números encima de las líneas corresponden a las probabilidades posteriores bayesianas. (Para las abreviaturas de las distintas secuencias, véase arriba.)

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Table 5. Z–test for positive selection at putative antigen binding sites (ABS) as described for humans by Brown et al. (1993) and the non–ABS positions of DRB exon 2 alleles of Alouatta pigra: dS, synonymous substitutions; dN, non–synonymous substitutions; dN/dS, ratio of nonsynonymous to synonymous substitutions; P, probability value. Sequences were grouped according to their position in the phylogenetic analysis. Tabla 5. Prueba de Z para la selección positiva en sitios de unión al antígeno (ABS) descritos para humanos por Brown et al. (1993) y en las posiciones de aminoácidos no pertenecientes a ABS (non–ABS) de los alelos del exón 2 del gen DRB detectados en Alouatta pigra: dS, sustituciones sinónimas; dN, sustituciones no sinónimas; dN/dS, tasa entre sustituciones no sinónimas a sinónimas; P, valor de probabilidad. Las secuencias se agruparon según la posición que ocupan en el análisis filogenético.

d N

dN/dS

Z–value

ABS N = 24 Clade I

0.114 ± 0.047

0.009 2.387

Clade II

0.097 ± 0.057

0.004 2.696

Clade III

0.014 ± 0.029

0.072 ± 0.042

5.14

0.081

All sequences

0.078 ± 0.071

0.181 ± 0.094

2.305

0.045 1.535

Clade I

0.021 ± 0.014

0.022 ± 0.012

1.04

0.468

Clade II

0.020 ± 0.022

0.023 ± 0.019

1.13

0.401

0.252

Clade III

0.033 ± 0.037

0.024 ± 0.027

0.73

–0.327

All sequences

0.028 ± 0.022

0.039 ± 0.022

1.39

0.282

0.578

1.407

Non–ABS N = 65

lyzed (26 out of 44 individuals) had more than five DRB sequences. Some individuals presented more than two DRB alleles (up to four) that belonged to the same phylogenetic cluster. Seventeen individuals harboured three or four alleles from cluster II, and nine individuals had three or four alleles from cluster III (table 2). Alouatta pigra from CBR showed a slightly higher allelic variation (N = 21) than PNP (N = 18). In CQR, where only four samples were successfully analysed, eight sequence variants were detected. Allele Alpi–DRB*W106:07:01 was not detected in PNP and alleles Alpi–DRB*W106:07:02 and Alpi– DRB*W107:02 were unique for CBR (table 4). As for the nucleotide variation and the number of segregating sites, values were similar for CBR (π = 0.057; S = 45) and PNP (π = 0.057; S = 44) (table 3). CQR values were slightly lower than those for the other two best–represented localities (π = 0.055, S = 39, and an average number of alleles per individual of 4.25). The greatest difference in the mean number of alleles per individual was between CBR (5.09) and CQR (4.25) (table 3). Positive selection Within the 24 amino acid positions putatively belonging to the ABS, 13 (54.2 %) were detected as variable. On the other hand, only 14 (20 %) out of 65 positions of the non–ABS part were variable, and half of them

0.081

were located next to an ABS. The values of the rates of non–synonymous (dN) vs. synonymous (dS) substitutions were higher for the ABS (dN/dS = 2.305) than for the non–ABS sequence region (dN/dS = 1.39), for all sequences and for each cluster. Even though both values of dN/dS were higher than 1, the Z–test for positive selection was significant (P = 0.045) only for the ABS (table 5). For the ABS of Cluster III, although the dN/dS rate showed a high value, Z test was not significant. Discussion Current knowledge on MHC–DRB polymorphism in platyrrhines is concentrated in a few species of the Cebidae family, which are commonly used in biomedical research (Nino–Vasquez et al., 2000; Suárez et al., 2006; Mee et al., 2011). This study is therefore, the first to investigate DRB gene nucleotide sequence polymorphism in a representative species of the Atelidae family. In A. pigra, we detected a relatively low level of allelic diversity (N = 21) and four different DRB lineages from those previously reported in other primate species. DRB allelic diversity in A. pigra MHC allelic diversity can vary considerably among primate species. Many catarrhine species have been

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found to possess a high DRB allelic diversity (e.g. 2311 in humans, 211 crab–eating macaques), while in general, a more modest allelic diversity has been detected in NWM (e.g. 44 in common marmoset, 110 in owl monkeys, and 47 in cotton–top tamarins) (De Groot et al., 2012; Robinson et al., 2013). Thus, comparing the DRB allelic diversity found in A. pigra (i.e. 21 sequences in 44 individuals) to that reported in other studies from NWM (e.g. Aotus nancymaae, 34 sequences in 15 individual and 67 sequences in 71 individuals; Saguinus oedipus, 28 sequences in 13 individuals; Callithrix jaccus, 21 sequences in 30 individuals [Antunes et al., 1998; Suárez et al., 2006; Nino–Vasquez et al., 2000; Gyllensten et al., 1994]), it can be considered as low. This difference in the level of DRB sequence diversity of A. pigra compared to other NWM could be related to several factors, such as differences in selection pressure (Mainguy et al., 2007; Hambuch and Lacey, 2002). Although NWM have mainly arboreal lifestyles, parasite exposure could vary between species given their differences in: (a) geographical distribution, as evidence suggests a decrease in parasite richness from the equator to poles (Guernier et al., 2004; Mainguy et al., 2007); (b) the size of the social groups (i.e. 2–8 individuals in A. pigra; 15–30 in C. jaccus; 2–5 in A. nancymaae; and 2–13 in S. oedipus), the population density and their contact with related species, which influence lateral transfer of parasites (Nunn et al., 2003); and (c) the species diet, as those that include insects in their diet (A. nancymaae, S. oedipus, and C. jacchus) are expected to have higher parasite diversity than herbivores (A. pigra), as insects can serve as intermediate hosts for parasites (Lafferty, 1999; Vitone et al., 2004). Alternatively, the reduced allelic variation of A. pigra could be explained by a process of past purifying selection against a parasite (Bollmer et al., 2011), favouring the fixation of only a limited number of DRB sequences. Historical events such as founding events and bottlenecks, however, can be translated into low MHC sequence variation. In such events, the strength of balancing selection to maintain polymorphisms is exceeded by genetic drift. therefore, randomly losing genetic diversity (Sommer, 2005). This seems the most likely explanation for the sequence diversity found in A. pigra. First, due to its biogeographic history, as Central American Alouatta species (i.e., A. pigra and A. palliata) have their origin in populations of northern South America that invaded Central America and southern Mexico after the formation of the Panama isthmus (Fleagle, 1999). Microsatellite loci studies of both Central American Alouatta species show that this founder effect derived from the process of colonization, creating a pattern of gradual loss of genetic variation northward of their distribution (Ellsworth and Hoelzer, 2006). Contrary to neutral markers such as microsatellites, MHC loci are subjected to positive selection due to their role in the host immune response. Nevertheless, the strong genetic drift derived from this process of colonization could have affected MHC allelic diversity in this northern species. This hypothesis is strongly supported by the low neutral variability (i.e.

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microsatellites) previously detected in the species (Amendola, 2009; Van Belle et al., 2012). It is also in agreement with the almost homogenous geographical distribution of DRB sequences detected among the sampled localities (table 4). Thus, although neither of the two former explanations can be ruled out, they cannot necessarily account for the low variation in neutral markers reported for A. pigra in other studies. Therefore, the relatively low level of DRB allelic diversity in A. pigra is most likely a consequence of drift due to the biogeographical history of this species. DRB loci and lineages Some of the known DRB platyrrhine alleles show high similarity to catarrhine alleles from DRB1*03, DRB3*01, and DRB5 loci. However, evidence indicates that the similarity between DRB alleles of platyrrhines and catarrhines has arisen through convergent evolution, given the differences of the codons determining the shared amino acid motifs (Kriener et al., 2000). None of the sequences detected in A. pigra belong to this group of alleles, and the four lineages to which they were assigned (DRB*W105, DRB*W106, DRB*W107, and DRB*W108) are different from those previously reported in other primate species, although they show a high nucleotide similarity (93–95.9 %) to alleles of other NWM lineages. The number of alleles detected per animal ranged from two to eight, indicating the existence of at least four DRB loci in A. pigra. This variation in the number of MHC–DRB loci between individuals of a species has been reported for several mammals, such as humans, other non–human primates, bats, tree shrews, and voles (Bontrop, 2006; Oppelt et al., 2010; Kloch et al., 2010; Salmier et al., 2016). These gene duplications contribute to MHC diversity and are known to play an important role in the adaptive evolution of organisms (Hughes and Yeager, 1998). Phylogenetic analysis revealed three A. pigra DRB allele clusters, which could be interpreted as hypothetical loci, on the premise that alleles from a particular DRB locus tend to cluster together. However, if a relatively recent gene duplication has occurred, the alleles of both genes will be highly related, difficulting the distinction of the two loci by phylogenetic clustering of sequences (Oppelt et al., 2010). This is probably the case in A. pigra since data indicate the existence of at least four loci, but only three clusters were recovered. Furthermore, the allelic composition of the screened animals suggests that two of the phylogenetic clusters likely represent more than a single loci. This is revealed by a high number of individuals harbouring more than two alleles from cluster II or III (table 2). In contrast, no more than two alleles from cluster I were found in a single individual. Nonetheless, the evolution and mechanisms generating polymorphisms in DRB exon 2 are complex and do not follow specific patterns (Takahata and Satta, 1998). Consequently, the exclusive use of exon 2 sequences to evaluate allele phylogenetic relationships may lead in some cases to erroneous conclusions. Moreover, it has been shown that the relationship between alleles is more accurately inferred by using information from

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introns or exons that are not under selective pressure (Kriener et al., 2000). The inclusion of intron sequence data could thus help clarify the relationships between the alleles of A. pigra, providing a more precise identification of loci. In addition, the observed variation in the composition of lineages present in each animal could indicate the existence of haplotype (i.e. combinations of sequences that are inherited as blocks) polymorphism in the species. Althoughwe cannot rule out the possibility that mutations in the hybridization sites of the primers and allelic drop out may have led to a failure to detect some of the allelic variants , the methodological approach used here minimizes this possibility, and a real haplotype variation is likely occurring. However, due to the lack of data on mother–infant pairs or kinship information of the analysed animals, it was not possible to define DRB haplotypes. Since genomic DNA was used, it cannot be ascertained whether the DRB alleles amplified for A. pigra are expressed. However, no evidence indicating the presence of pseudogenes was found, as none of the studied sequences have indels or premature stop codons. Also, an excess of non–synonymous substitution at the ABS reveals that selection has acted on these genes, but such a process is not necessarily on–going (Garrigan and Hedrick, 2003). Therefore, transcription studies are needed to confirm their functionality, particularly because genomic and expression analyses in other NWM species have uncovered the presence of non–functional alleles from different lineages (Doxiadis et al., 2006; Trtková et al., 1993). Finally, although A. pigra has a reduced set of alleles, sequence level divergence between some of the alleles from the different lineages is marked (up to 33 p). Thus, they may likely recognize antigens from very different pathogens (Hedrick et al., 2000). Furthermore, polymorphism is observed in the number of lineages and loci, which is comparable to that found in other NWM species, except Aotus nancymaae (Trtková et al., 1993; De Groot et al., 2012; Gyllensten et al., 1994). It is also observed in allelic configurations of individuals, which suggests haplotype variation. This observed polymorphism, taken as a whole, appears to have been sufficient for wild A. pigra to cope with parasites under current ecological conditions, given that there is no evidence that the species suffers from increased susceptibility to parasites when compared with other Alouatta species (Trejo–Macías and Estrada, 2012; Vitazkova and Wade, 2006; Stoner and Gonzalez Di Pierro, 2006). However, this relatively low variation in adaptive immune response genes may not be enough to give populations the necessary protection in case of the emergence of novel pathogens. This could be a cause of concern for the conservation of the species if a greater loss of this diversity occurs derived from the ongoing population declines. Nevertheless, the relatively low level of polymorphism detected in these functionally important loci (MHC–DRB), together with previous low variation reported in neutral markers (microsatellites and mtDNA), suggests that genetic drift due to the biogeographical history of the species had an impact on the general

genetic diversity of the species. As a result, conservation programs should consider the maintenance of the maximal genetic diversity (including MHC and other immune genes). Future research should also consider spatio–temporal variations of MHC–DRB diversity in A. pigra populations so as to monitor and detect possible increases in the risk of infectious diseases. This applies particularly to those at higher risk of pathogen introduction,and those suffering from steep population declines or isolation derived from human encroachment to their habitat (Ujvari and Belov, 2011; Grogan et al., 2017). Another point of relevance is the need to incorporate in future studies other functionally important immune genes, such as MHC class I genes, particularly MHC–B and MHC–G like genes that have shown to be highly polymorphic and that are expressed in several NWM, including some from the Atelidae family (van der Wiel et al., 2013; Lugo and Cadavid, 2015; Cao et al., 2015). Non–MHC immune genes, especially those related to viral resistance like the OAS1 (oligoadenylate synthetase) gene (Acevedo–Whitehouse and Cunningham, 2006; Rios et al., 2007), should also be considered since there have been reports of yellow fever outbreaks in populations of other Alouatta species (De Almeida et al., 2012; Crockett, 1998; Holzmann et al., 2010), that could represent a possible threat to the conservation of A. pigra populations. All this would allow a better understanding of the immunogenetic status of the species and gauge its response to novel pathogen threats. In summary, a relatively low MHC–DRB exon 2 variation was detected at the allelic level in A. pigra compared to that reported for other NWM species. The consistency of these results with those from previous studies on neutral marker variation, and the nearly homogeneous allele and lineage distribution observed across localities of this study, suggest that this reduced allelic diversity in free–ranging A. pigra is most likely associated with a historical founder effect during the northward expansion of their ancestors from South America. However, the excess of nonsynonymous substitutions at the antigen binding sites indicates that sequence evolution was driven at some point by positive selection. Finally, despite the relatively low DRB allelic diversity in the species, polymorphism in lineage and loci, as well as the level of differentiation between some of the alleles, might be enabling wild populations to initiate an adequate immune response to cope with current pathogens, but it might entail a risk for these populations in case of the emergence of new ones. Acknowledgements We thank the Secretaría del Medio Ambiente y Recursos Naturales, and Comisión Nacional de Areas Naturales Protegidas of Calakmul Biosphere Reserve and Palenque National Park, for permits and logistic support in the field for sample collection. We are grateful to Jonas Morales Linares, Ricardo Gómez Enríquez and Sebastián Montejo Narvaez for field assistance and Janet Nolasco Soto for her assistance during lab

Animal Biodiversity and Conservation 41.2 (2018)

analysis. We appreciate the support given by the staff of the molecular analysis lab of the Instituto de Ciéncia y Medicina Genómica in the RSCA sample analysis, particularly to Ana Laura Astorga Sifuentes, Claudia Alejandra Gómez Martínez and Fernando Hernández Terán. We also thank Ruben López Revilla and Gloria Carrión for allowing the use of their respective laboratories for the processing of samples, and Ivo Borges Bombarda for the elaboration of the sampled localities map. Finally, we thank Natasja de Groot, Ronald Bontrop and Nel Otting for their help in naming the alleles. LEAS received a Graduate Student grant from Consejo Nacional de Ciencia y Tecnología Mexico (250327). This manuscript is submitted in partial fulfillment of the requirements for the degree of Doctor of Science in the graduate studies program of INECOL. References Acevedo–Whitehouse, K., Cunningham, A., 2006. Is MHC enough for understanding wildlife immunogenetics? Trends in Ecology & Evolution, 21(8): 433–438. Aguilar, A,. Roemer, G., Debenham, S., Binns, M., Garcelon, D., Wayne, R. K., 2004. High MHC diversity maintained by balancing selection in an otherwise genetically monomorphic mammal. Proceedings of the National Academy of Sciences, 101(10): 3490–3494, doi: 10.1073/pnas.0306582101 Amendola, M., 2009. Estudio de la variabilidad genetica en poblaciones de Alouatta pigra del estado de Campeche: implicaciones para la conservación. Ph D Thesis, Instituto de Ecologia, A.C. Antunes, S. G., de Groot, N. G., Brok, H., Doxiadis, G., Menezes, A. A. L., Otting, N., Bontrop, R. E., 1998. The common marmoset: a new world primate species with limited Mhc class II variability. Proceedings of the National Academy of Sciences of the United States of America, 95(20): 11745–11750. Argüello, J. R., Little, A. M., Bohan, E., Goldman, J. M., Marsh, S. G. E., Madrigal, J. A., 1998. High resolution HLA class I typing by reference strand mediated conformation analysis (RSCA). Tissue antigens, 52(1): 57–66, doi: 10.1111/j.1399-0039.1998. tb03024.x Babik, W., 2010. Methods for MHC genotyping in non–model vertebrates. Molecular Ecology Resources, 10(2): 237–251. Belov, K., 2011. The role of the Major Histocompatibility Complex in the spread of contagious cancers. Mammalian Genome, 22(1–2): 83–90. Bernatchez, L., Landry, C., 2003. MHC studies in nonmodel vertebrates: What have we learned about natural selection in 15 years? Journal of Evolutionary Biology, 16(3): 363–377. Bollmer, J. L., Hull, J. M., Ernest, H. B., Sarasola, J. H., Parker, P. G., 2011. Reduced MHC and neutral variation in the Galápagos hawk, an island endemic. BMC Evolutionary Biology, 11(1): 143, doi: 10.1186/1471-2148-11-143. Boni, M. F., Posada, D., Feldman, M. W., 2007. An exact nonparametric method for inferring mosaic

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Breeding activity of the agile frog Rana dalmatina in a rural area M. Biaggini, I. Campetti, C. Corti

Biaggini, M., Campetti, I., Corti, C., 2018. Breeding activity of the agile frog Rana dalmatina in a rural area. Animal Biodiversity and Conservation, 41.2: 405–413. Abstract Breeding activity of the agile frog Rana dalmatina in a rural area. Rural landscapes can host many protected species that are constantly threatened by agriculture intensification and abandonment of traditional managements. Amphibians are severely affected by both processes due to loss and alteration of aquatic and terrestrial habitats. We monitored the breeding activity of Rana dalmatina in a lowland rural area focusing on spawning sites in open habitats, namely ditches amid traditional arable lands and pastures with varying vegetation features, size, and distances from woodlots. Egg clump density and clump size differed between sites, probably depending on environmental and ecological factors (i.e., larval competition, food availability, and predation). The sites next to woodlots showed the highest clump density (up to 0.718 n/m2). Our observations indicate that the maintenance and correct management of water bodies connected to traditional rural activities can be key to amphibian conservation in agricultural areas. Key words: Agriculture, Amphibians, Breeding sites, Conservation, Rana dalmatina Resumen Actividad reproductiva de la rana ágil Rana dalmatina en una zona rural. Los territorios agrícolas pueden ser refugio de varias especies protegidas que, sin embargo, están constantemente amenazadas por la intensificación de la agricultura y el abandono de las prácticas tradicionales. Los anfibios están gravemente amenazados por ambos procesos, a causa de la pérdida y la alteración de los hábitats acuáticos y terrestres. Estudiamos la actividad reproductiva de Rana dalmatina en una zona rural de tierras bajas, centrándonos en los sitios de desove en hábitats abiertos, es decir, zanjas entre tierras arables y pastos con diferente vegetación, tamaño y distancia a las parcelas forestales. La densidad y el tamaño de las masas de huevos difirieron según el sitio, probablemente como resultado de factores ambientales y ecológicos (p. ej., la competición entre las larvas, la disponibilidad de alimentos y la depredación). Los sitios cercanos a los parcelas forestales mostraron la densidad más alta (hasta 0,718 n/m2). Nuestras observaciónes indican que la conservación y el mantenimiento adecuado de las masas de agua destinadas a las actividades rurales tradicionales pueden ser esenciales para la conservación de los anfibios en zonas agrícolas. Palabras clave: Agricultura, Anfibios, Sitios reproductivos, Conservación, Rana dalmatina Received: 30 XI 17; Conditional acceptance: 23 I 18; Final acceptance: 26 II 18 Marta Biaggini, Claudia Corti, Museo di Storia Naturale dell'Università degli Studi di Firenze, Sez. di Zoologia "La Specola", Via Romana, 17, 50125 Firenze, Italy.– Irene Campetti, Dipto. di Biologia dell'Università di Pisa, Via Luca Ghini 13, Pisa, Italy. Corresponding author: M. Biaggini. E–mail: marta.biaggini@virgilio.it

ISSN: 1578–665 X eISSN: 2014–928 X

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Introduction Most European countries are widely characterized by cultural landscapes that have been deeply modified by human activities over the centuries. Agriculture, in particular, has played a major role in shaping large extensions of territories (Emanuelsson, 2009).Agricultural lands in Italy occupy over 40 % of the total surface area and about 30 % of the Natura 2000 network, and they include many endangered habitats and species (http://www.crea.gov.it/lagricoltura–italiana– conta–2016; http://www.minambiente.it/pagina/rete– natura–2000). However, due to the intensification of cultivation practices, agriculture is now a major threat for biodiversity worldwide (Wake, 1991; Foley et al., 2005). Therefore, monitoring biodiversity and planning conservation actions in rural landscapes is essential to develop effective environmental protection policies. Amphibian populations are declining worldwide (Stuart et al., 2004; Gallant et al., 2007) due to several threats such as habitat loss and fragmentation (Blaustein and Wake, 1990; Cushman, 2006), emerging infectious diseases (Daszak et al., 2003), and climate change (Corn, 2005; Pounds et al., 2006). Moreover, amphibians are particularly vulnerable to land use changes and habitat alterations due to their complex lifecycle that requires both aquatic and terrestrial habitats at different life stages and, after metamorphosis, throughout the seasons (i.e., for breeding, hibernating) (Beebee, 1996; Alford and Richards, 1999). For these reasons, they can be considered good indicators in human–altered environments. The agile frog, Rana dalmatina Fitzinger, in Bonaparte, 1838 is a brown frog, broadly distributed in central and southern Europe, and listed in Habitat Directive 92/43/EEC (Annex IV). It is an explosive breeder; in early spring, females lay a single egg clump in some type of water body or, less frequently, in slow–flowing watercourses. Such breeding features make it particularly easy to monitor the population density and the breeding trend of this species through standardized egg counts (Grossenbacher et al., 2002; Rossi et al., 2016). R. dalmatina has terrestrial habits and occurs in various environments, mainly deciduous forests, but also open habitats like grasslands within forested areas, from sea level up to 1,500 m a.s.l. (Speybroeck et al., 2016). In some areas, the lifecycle can involve both forested (terrestrial post–metamorphic phase) and open habitats (breeding period, larval phase). As such, R. dalmatina may be considered a useful target species for studies on environmental conservation in lowland landscapes where alternation of both kinds of habitats often occurs. In these areas, most females spawn in open habitats (frequently within 200 m from forests) where egg and larval development can be favored by higher temperatures (Ponsero and Joly, 1998; Wederkinch, 1988). In particular, R. dalmatina can also be found in agricultural areas near forests or residual woodlots (Pavignano et al., 1990; Hartel et al., 2009). Consequently, waterbodies, such as permanent or ephemeral ponds, ditches and watercourses included in the matrix of agricultural land uses can acquire key importance for the species’ persistence at a local

scale. The risk of alteration and loss of both breeding sites and habitats suitable for dispersion and survival of the post–metamorphic phase is particularly high in rural areas where the intensification of agricultural practices can strongly threaten R. dalmatina. In this study, we monitored R. dalmatina female breeding activity and tadpole development in a rural area in central Italy, partially included in a Natura 2000 site (special area of conservation, SAC IT5120101). In a previous study (Biaggini and Corti, 2011), we observed that, in the same area, the species also spawned in ditches and watercourses outside the forested patches. We focused on these breeding sites in order to: i) verify if and how rural habitats are exploited by R. dalmatina for the breeding activity; ii) identify the main threats affecting the species in a rural lowland area; and iii) suggest possible conservation actions. In particular, in the study area, one of the two woodlots frequented by R. dalmatina is fenced and included in a WWF oasis and is thus at least partially protected. Fields and grasslands, on the contrary, are involved in economic activities (sheep pasture, arable lands) and the SAC still lacks a management plan. Among amphibians, Lissotriton vulgaris, Triturus carnifex, Bufo bufo, Hyla intermedia, Pelophylax lessonae–esculentus complex, also occur in the same open habitats (Biaggini and Corti, 2011). Therefore, the need for conservation proposals for the rural habitats is pressing. Material and methods Study area and sampling The study area (43° 46' N, 10° 37' E, 1.71 km2) is partially included in the SAC IT5120101 'Ex alveo del Lago di Bientina', a large reclaimed area dominated by agricultural lands that are periodically flooded, and smaller patches of wetlands and wet woods. From mid–February to the end of May 2017, we visited the area once a week in search of egg clumps in all potential breeding sites, focusing on open habitats. We detected five breeding sites (table 1) where we counted egg clumps during each visit, by walking along the ditch banks and back again. For each egg clump found, we recorded the following variables: position (using a GPS device), distance of the clump from the water surface (spawning depth), water depth, the three main diameters of clumps (two horizontal, one vertical) and egg diameter (with and without envelope, using a caliper. For each clump, we considered the mean diameter of three eggs). We estimated the number of eggs per clump from the ratio between the volume of the clump (assimilated to an ellipsoid) and the volume of the egg (assimilated to a sphere). Considering that eggs and egg clumps tend to expand with maturation, we took measures on newly laid clumps (six days at most). We assessed the distance of each clump from the nearest woodlot (distance from the wood) using Google Earth aerial images and tools (www. google.com/intl/it/earth/).

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Statistical analyses We used the whole dataset for descriptive statistic (egg and clump size) and to compare egg clump density between the five breeding sites found in the study area (table 1). For the other analyses, we considered only sites 1, 2, 3, for which larger samples were available (table 1). We performed a Friedman test to compare clump density between the five breeding sites, using the ratio between the number of clumps recorded during each sampling, and the studied surface for each watercourse. For the following pairwise comparisons, we used the Wilcoxon matched pairs test. Focusing on sites 1, 2 and 3 to explore the relationships between the main reproductive variables, water depth, and the distance from the nearest woodlot, we performed a Spearman rank correlation between egg size (considered with envelope hereafter), number of eggs per clump, spawning depth, water depth, and distance from the wood. To verify whether spawning activity varied between ditches, we compared the egg size, the number of eggs per clump, and the spawning depth in the three breeding sites, using log–linear models. We adjusted egg size for the number of eggs per clump and vice versa, based on Spearman rank correlation results. When analyzing the spawning depth we assumed a Poisson distribution of the data due to the high number of egg clumps recorded on the water surface (spawning depth equal to zero); spawning depth was adjusted for water depth and distance from the nearest woodlot (covariates), based on Spearman correlation results. Results The spawning activity occurred from mid–February to mid–March, peaking during the first two weeks (fig. 1). We measured 141 clumps. Eggs had a mean diameter (± SD) of 2.04 ± 0.18 mm, and 10.14 ± 1.99 mm considering the envelope (table 1). We estimated that clumps contained from about 90 to 1,480 eggs, with a mean value of 578.19 ± 295.55 (table 1). The density of egg clumps (n/m2) varied significantly between the five breeding sites (Friedman test: n = 8, x2 = 23.771, p < 0.001). Site 2 had the highest density, differing significantly from all sites but 1 (Wilcoxon tests: 2 vs 1, Z = 1.400, p = 0.161; 2 > 3, Z = 2.028, p = 0.042; 2 > 4, 5, Z = 2.201, p = 0.028), followed by site 1 (1 > 3, 4, 5, Z = 2.521, p = 0.012), and 3 (3 > 4, 5, Z = 2.201, p = 0.028); site 4 and 5 had comparable low densities (4 vs 5, Z = 1.604, p = 0.109) (fig. 1, table 1). Spearman correlation revealed that in the biggest clumps eggs tended to be smaller (correlation between egg size and number of eggs per clump: n = 137, r = –0.642, p < 0.001, fig. 2); both egg and clump sizes were not correlated with any other variable (p > 0.05 in all cases). The spawning depth was positively correlated with water depth (n = 137, r = 3.149, p = 0.002) and with the distance from the woodlots (n = 137, r = 2.660, p = 0.009). These two last variables

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were highly correlated (n = 137, r = 0.448, p < 0.001) because the deepest ditch was also the furthest from woodlots (table 1). No other correlations were significant (p > 0.05). Log–linear models revealed that egg size (adjusted for the number of eggs per clump) did not vary between breeding sites, while clumps contained significantly fewer eggs at site 2 than at sites 1 and 3, when controlling for egg size (even if significance was low, table 2, fig. 3). The spawning depth was greater at site 3, with respect to sites 1 and 2, and the distance from the wood had no significant influence on the model (table 2). Discussion In the present study, we further verified that Rana dalmatina exploits rural habitats for spawning in lowland environments (Pavignano et al., 1990; Hartel et al., 2009). Eggs were laid in slow running waters, both in drainage ditches with no aquatic or semi– aquatic vegetation and completely in the open, and in ditches rich in semi–aquatic plants and with trees and bushes on their banks (table 1). The two breeding sites with the highest clump density were next to a woodlot: in site 2, in particular, we recorded a very high clump density, 0.718 n/m2 (table 1) compared with values available in literature. Ponsero and Joly (1998) recorded a maximum clump density of 0.30 n/m2 in a marsh, Hartel (2005) a maximum of 0.026 n/m2 in a rural area, and Romano et al. (2016), a maximum of 0.47 n/m2 in a woodland. In the study area, the breeding sites exploited by R. dalmatina in open habitats were the same as in 2010 (Biaggini and Corti, 2011), with the exception of site 2, where no egg clumps were found during the first observations. Clumps varied in size, from few eggs to almost 3,000, with the mean number of eggs per clump being nearly 600. In previous studies, Ponsero and Joly (1998) found a mean of t 316 to 406, Weddeling et al. (2005) 950, Bernini et al. (2007) 800, and Solský et al. (2014) 1,300. In the largest clumps, eggs tended to be smaller and vice versa, suggesting a possible trade–off between egg number and size, as found in other R. dalmatina (Weddeling et al., 2005) and Rana spp. populations (e.g., Cummins, 1986). When comparing clump and egg size between the three main breeding sites, we found that clumps in site 2 contained, on average, fewer eggs (when correcting for egg size), while we found no differences in egg size. According to Weddeling et al. (2005), clump size does not depend on age, or size, or somatic condition of females. Alternatively, the lower number of eggs per clump in site 2 could be a response to some environmental constraints. Site 2 was the ditch with the highest clump density and the only one lacking aquatic vegetation, contradicting the more commonly observed association between R. dalmatina breeding sites and the presence of abundant aquatic vegetation (Pavignano et al., 1990; Kecskés and Puky, 1992; Hartel, 2008). Moreover, no bushes or trees occurred on the banks of the ditch, just grasses. Consequently, site 2 was also the only site lacking

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Table 1. Main features of Rana dalmatina spawning sites and breeding characteristics in the study area: BS, breeding site; Land uses, nearby land uses; Vegetation, semi–aquatic vegetation and vegetation along the ditch banks; Width, mean width (± SD) (m); Depth, mean water depth (± SD) (cm); NClump, maximum number of clumps (number of measured clumps); DClump, maximum egg clump density (N/m2); NEggs, Estimated number of eggs per clump (mean ± SD, range); EggSz, mean egg size (± SD) (cm, with envelope); SpDepth, mean spawning depth (± SD, and range) (cm); Dist, mean distance of clumps from wood edge (± SD) (m).

BS 1

Land uses

Deciduous woodlot,

arable lands.

Vegetation

Width

Depth

Dominated by

1.482 ± 0.205

15.800 ± 4.165

Iris pseudacorus

On the SCI boundaries

and Arundo donax

from the end of March 2

Deciduous woodlot,

Absent

0.533 ± 0.052

Outside the SCI

Sheep pasture and abandoned

pastures managed with

periodic cutting.

Arundo donax,

On the SCI boundaries

Phragmites australis

Dominated by

with periodic cutting.

Outside the SCI

Sheep pasture.

Inside the SCI

1.552 ± 0.383

Mainly

2.267 ± 0.441

19.333 ± 8.042

1.228 ± 0.225

18 ± 3.606

Carex elata Shrubs, mainly Rubus spp., Crataegus monogyna

0.8

0.8 Site Site Site Site Site

0.7 0.6 0.5

1 2 3 4 5

0.6 0.5

4 IV

0.1 28 III

0.1 21 III

0.2

14 III

0.2

7 III

0.3

28 II

0.3

21 II

0.4

14 II

Median 25 %–75 % Min–max

0.7

0.4

7 II

31.727 ± 4.968

Iris pseudacorus,

Abandoned pastures managed

Egg clump density (n/m2)

19.230 ± 4.938

arable lands.

3 Sites

Fig. 1. Egg clump density at the various breeding sites: trend throughout the spawning period (left); comparison between sites (right). Fig. 1. Densidad de las masas de huevos en los diferentes sitios reproductivos: tendencia durante el período de puesta (izquierda); comparación entre sitios (derecha).

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Tabla 1. Principales características reproductivas de Rana dalmatina y principales características de sus sitios de desove en la zona del estudio: BS, sitio de desove; Land uses, usos de las tierras cercanas; Vegetation, vegetación semiacuática y vegetación en el lecho de la zanja; Width, anchura media (± DE) (m); Depth, profundidad media del agua (± DE) (cm); NClump, número máximo de masas de huevos (número de masas medidas); DClump, densidad máxima de las masas de huevos (N/m2); NEggs, número estimado de huevos por masa (media ± DE, rango); EggSz, tamaño medio de los huevos (± DE) (cm, con envoltura); SpDepth, profundidad media de la puesta (± DE, rango) (cm); Dist, distancia media de las masas de huevo a las zonas forestales (± DE) (m).

NClump

DClump

NEggs

EggSz

SpDepth

Dist

129 (84)

0.295

614.767 ± 276.366

0.976 ± 0.158

0.875 ± 1.893

(130–1391.602) (0–8)

51 (41)

0.718

471.811 ± 299.564

1.085 ± 0.271

1.158 ± 2.428

(91.594–1367.187)

40.634 ± 12.369

(0–9)

89 (12)

0.073

685.634 ± 335.169

0.954 ± 0.094

6.417 ± 5.760

(330.625–1481.481)

224.750 ± 18.733

(0–16)

4 (4)

0.011

249.278 ± 78.342

1.075 ± 0.050

325.232 ± 13.722

< 0.001

–

149.957 ± 16.106

2,0 1,8

Egg size (cm)

1,6 1,4 1,2 1,0

Site 1 Site 2 Site 3

0,8 0,6

200

400

600 800 1,000 1,200 1,400 1,600 Eggs per clump

Fig. 2. Correlation between egg size and number of eggs per clump. Fig. 2. Correlación entre el tamaño de los huevos y el número de huevos por masa.

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Table 2. Comparison of egg size, clump size and spawning depth between the three main breeding sites: * comparisons significant at p < 0.05. Tabla 2. Comparación del tamaño de los huevos, el tamaño de las masas y la profundidad de puesta entre los tres sitios reproductivos principales: * comparaciones significativas con p < 0,05.

Wald stat.

Intercept

92.888

< 0.001

Number of eggs per clump

208.915

< 0.001

0.017

0.991

Dependent variable: egg size

Water course

Goodness of fit. Deviance: stat/DF = 0.003; Pearson x2: stat/DF = 0.003; AIC = –381.169 Dependent variable: number of eggs per clump Intercept

3218.385

< 0.001

Egg size

160.633

< 0.001

5.909

0.048

Water course Comparisons between water courses: 2 > 1, 3*

Goodness of fit. Deviance: stat/DF = 0.138; Pearson

x2: stat/DF = 0.138; AIC = 123.863

Dependent variable: spawning depth Intercept

0.028

0.868

Distance from wood

3.641

0.056

Water depth

25.886

0.000

Water course

6.231

0.044

Comparisons between water courses: 3 > 1, 2* Goodness of fit. Deviance: stat/DF = 3.157; Pearson

evident debris –the tadpoles' food source– on the bottom. These observations suggest intraspecific larval competition was probably high at this site (in addition, at the end of March, Hyla intermedia also spawned at this site). Larval density can have detrimental effects on metamorphosis, influencing the development rate, size and survivorship of larvae (e.g., Wilbur, 1976; Semlitsch and Caldwell, 1982; Petranka and Sih, 1986; Murray, 1990). At high densities and/or low food levels, larvae generally grow more slowly and are smaller at metamorphosis, both because of reduced resources, and because of the production of growth inhibitors (Licht, 1967; Wilbur, 1977; Steinwascher, 1978). As such, high larval competition can induce females to choose alternative breeding sites (Hartel, 2008) or, to lay fewer eggs (Resetarits and Wilbur, 1989). A similar mechanism could be hypothesized at site 2 but an extended sampling is clearly necessary to verify and understand the observed pattern. Nonetheless, some other features at site 2, such as the presence of fewer predators and favorable thermal conditions (e.g., Resetarits and Wilbur, 1989; Petranka et al., 1994; Laurila and Aho, 1997; Binckley and Resetarits, 2003; Ficetola and De Bernardi, 2004) may have made it an attractive breeding site. If, on

x2: stat/DF = 4.243; AIC = 547.843

one hand, the absence of vegetation surely reduced the availability of food and shelter for tadpoles, it could also have reduced the presence of predators, which, among invertebrates, are usually associated with more abundant vegetation (Gunzburger and Travis, 2004). Indeed, during sampling, we did not observe potential predators (such as, Odonata or Dytiscidae larvae), at site 2 on any occasion, with the exception of the allochthonous species Procambarus clarkii (adult individuals), which was present in the whole area but was not observed spawning in this ditch. In the other R. dalmatina breeding sites, on the contrary, we recorded P. clarkii larvae (at all sites), Gambusia affinis and, sporadically, Dytiscidae. Previous studies have shown that the choice of the breeding site of anuran females can be driven by the attempt to avoiding larvae predators (Resetarits and Wilbur, 1989; Petranka et al., 1994; Laurila and Aho, 1997; Binkley and Resetaris, 2003). Moreover, site 2, like site 1, was characterized by shallow water, which is usually an advantageous feature for the development of both embryos and larvae, because temperature can quickly rise, hastening the growth processes and easing survival rates (Semlitsch, 2002).

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1,000

Eggs per clump

900 800 700 600 500 400 300

2 Site

Fig. 3. Comparison of the estimated egg number per clump between sites 1–3 (mean values ± SD are shown). Fig. 3. Comparación del número estimado de huevos por masa entre los sitios 1–3 (se muestran las medias ± DE).

Most importantly, sites 1 and 2 were located next to wet woodland (dominated by Quercus rubur, Alnus glutinosa, Robinia pseudacacia), the presence of which guarantees an optimal habitat for metamorphosed juveniles and adults (Bernini et al., 2007). The positive relationship between the occurrence and density of R. dalmatina and the presence of deciduous woodlands has been widely observed (e.g., Pavignano et al., 1990; Hartel et al., 2009), and it is particularly strong at a small spatial scale (Ficetola et al., 2009). In a study in a floodplain, Ponsero and Joly (1998), found that most egg clumps were laid within 50–100 m from the wood. Most breeding sites in the study area dried up before the end of April, and the only site that guaranteed tadpoles’ complete metamorphosis was site 1 (where we observed the first metamorphosed juvenile in less than three months, at the end of May). This caused the loss of 54 % of the breeding effort (in terms of egg clumps) in the surveyed area. An analogous situation was observed in 2010 (Biaggini and Corti, 2011), when most of the drainage ditches and the flooded portions of fields dried out before larval metamorphosis (within April). On that occasion, a very large amount of Procambarus clarkii (in early life stages) was also recorded, especially at site 3, where although an egg clump density of 0.067/m2 was recorded at the beginning of March, we did not observe R. dalmatina larvae at the end of April (Biaggini and Corti, 2011). In 2010, we hypothesized a possible influence of P. clarkii in R. dalmatina loss of fitness, as observed in other areas (Ficetola et al., 2011). In spring 2017, all egg clumps developed without evidence of predation and the main cause of mortality was the early drying up of the breeding sites. It is possible that the red swamp crayfish may have

different impacts over time, due to natural population fluctuations (Martelloni et al., 2012). The variability in the breeding activity of the agile frog over the years (Cummins, 1986) and the species relatively high longevity (Sarasola–Puente et al., 2011) underscore the need for long–term monitoring in order to understand the population trends of the species, and to identify the involved biotic and abiotic factors in a certain area. Our observations, however, can provide some prompts for the conservation of the species in rural environments. The high clump densities recorded and the persistent spawning activity in artificial ditches over the years, suggest that the minor water bodies strictly connected to agricultural activities can be attractive breeding sites for R. dalmatina. In view of this, their maintenance and correct management may be of key importance for the persistence of amphibian populations, at least in traditional agricultural areas. The management of such water bodies must take several requirements into account. In the study area, the hydroperiod of R. dalmatina breeding sites should be extended to guarantee the metamorphosis of larvae. However, maintaining the ephemeral nature of ditches is important to lower the presence of potential predators and invasive species (e.g., Ficetola et al., 2012). The observed preference of the agile frog for a site without aquatic vegetation could indicate an opportunistic behavior of this species, probably shifting its more common habits to avoid predation. Should this be the case, the increase of sites similar to site 2, by simply managing already existing drainage ditches in the fields next to woodlots, would enhance the availability of breeding sites suitable for R. dalmatina and less fit for P. clarkii. Finally, promoting the control of red swamp crayfish abundance and dispersion could have positive effects for all the amphibian species in the area.

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Phylogeographic patterns of Capreolus capreolus in the centre of the Iberian peninsula F. Horcajada, L. Alcaraz, I. Barja, A. Machordom

Horcajada, F., Alcaraz, L., Barja, I., Machordom, A., 2018. Phylogeographic patterns of Capreolus capreolus in the centre of the Iberian peninsula. Animal Biodiversity and Conservation, 41.2: 415–425. Abstract Phylogeographic patterns of Capreolus capreolus in the centre of the Iberian peninsula. One hundred and one samples of muscle tissue were obtained from roe deer in the centre of the Iberian peninsula. We compared the sequences of the control region (D–loop) of the mitochondrial DNA of these samples with those obtained in previous studies. Adding the information from microsatellite markers and derived genetic parameters to study the population structure, we found a philopatric structure, with females maintaining mitochondrial haplotype diversity, while males showed a pattern of genome homogenization. The population can thus be considered panmictic. Different times of palaeohistory of the species may explain these results: glacial–interglacial stages of the Pleistocene and the reduction and recovery of populations throughout the 20th century. Key words: Capreolus capreolus, Central Spain, Microsatellites, Mitochondrial, Phylogenetic Resumen Patrones filogenéticos de Capreolus capreolus en el centro de la península ibérica. Se obtuvieron 101 muestras de tejido muscular de ejemplares de corzo en el centro de la península ibérica. Se compararon las secuencias de la región control (D–loop) del ADN mitocondrial con las obtenidas en otros estudios anteriores. Al añadir la información relativa a los marcadores microsatélites y los parámetros genéticos derivados para estudiar la estructura de la población, se constató la existencia de una estructura filopátrida, en la que las hembras mantenían la diversidad de haplotipos mitocondriales, mientras que los machos seguían un patrón de homogeneización del genoma. Por lo tanto, la población puede considerarse panmíctica. Diferentes épocas de la paleohistoria de la especie pueden explicar estos resultados: las etapas glaciales e interglaciales del Pleistoceno, por un lado, y una marcada reducción y recuperación de las poblaciones a lo largo del siglo XX, por otro. Palabras clave: Capreolus capreolus, España central, Microsatélites, Mitocondrial, Filogenética Received: 20 IV 17; Conditional acceptance: 01 XII 17; Final acceptance: 28 II 18 Fernando Horcajada Centro de Investigación, Seguimiento y Evaluación, Parque Nacional de la Sierra de Guadarrama, ctra. M–604 km 28, 28740 Rascafría, Madrid, Spain.– Lourdes Alcaraz, Annie Machordom, Depto. de Biodiversidad y Biología Evolutiva, Museo Nacional de Ciencias Naturales (MNCN–CSIC), c/ José Gutiérrez Abascal, 2, 28006 Madrid, Spain.– Fernando Horcajada, Isabel Barja, Depto. de Biología, Unidad de Zoología, Fac. de Ciencias, Univ. Autónoma de Madrid, c/ Darwin 2, Campus Universitario de Cantoblanco, 28049 Madrid, Spain. Corresponding author: F. Horcajada. E–mail: horcajadas@gmail.com

ISSN: 1578–665 X eISSN: 2014–928 X

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Introduction The roe deer (Capreolus capreolus) was part of the fauna of the Iberian peninsula during the last glaciations (Gliozzi et al., 1997; Hufthammer and Aaris–Sørensen, 1998). Some authors claim that the populations followed different routes of dispersal and recolonization toward northern Europe after the Ice Age (Hewitt, 1999), while others (Lorenzini and Lovari, 2006) propose that the Iberian peninsula remained an isolated geographical area, not acting as a main source for postglacial recolonization (Sommer and Zachos, 2009). Capreolus capreolus was present in the Middle Pleistocene of Central Spain, where paleontological remains have been found (Buitrago, 1992). In addition to functioning as a shelter in Pleistocene times, the Sierra de Guadarrama area might have played a fundamental role in safeguarding populations of roe deer (Tellería, 1999), although there are references indicating that populations of Central Spain underwent severe reductions in the 19th and early 20th century (Tellería and Virgós, 1997; Gortázar et al., 2000) that may have left their mark on their genetic parameters. Among the few genetic analyses that exist on the Iberian populations, only the studies of Lorenzini and Lovari (2006), Lorenzini et al. (2014), Randi et al. (2004) and Royo et al. (2007) included some locations near the Sierra de Guadarrama. With the aim of contributing to a comprehensive molecular analysis of the populations in the Sierra de Guadarrama we examined two types of markers, mitochondrial and nuclear (microsatellites), with the largest number of samples analysed to date in the Iberian peninsula, in order to increase our knowledge of the phylogeographic pattern of this species in Spain and dispel doubts about its heterogeneity, and to analyse the impact of isolation from recent centuries. Material and methods Study area Samples were collected between 2002 and 2007 at 22 locations in the Sierra de Guadarrama, in the mountains north of Madrid, in the centre of the Iberian peninsula (fig. 1, table 1). Currently, four of these localities are within the National Park of the Sierra de Guadarrama (declared a National Park on June 25th, 2013) (fig. 1). Sampled locations are within public forests of Pinus sylvestris and Quercus pyrenaica. Sample collection and DNA extraction We collected muscle tissue from 101 individuals distributed evenly throughout the study area. Samples were collected post–mortem from hunted deer or deer killed in road accidents. A small muscle biopsy was kept in absolute ethanol at –20 ºC until analysis was undertaken in the laboratory. DNA was extracted using the standard phenol/ chloroform protocol (Sambrook et al., 1989). Agarose

gels were used to test presence, concentration and possible degradation of the DNA extracted from the samples. Analysis of the mitochondrial DNA We amplified the complete mitochondrial D–loop using the primers L–Pro and H–Phe (Jäger et al., 1992). Internal primers used from the monitoring region (D–loop) were L–362 and H–493 (Randi et al., 1998). The PCRs were conducted with 1–3 ml DNA, 5 ml of the corresponding buffer (with 10 x 2 mM MgCl2), 1 ml of dNTPs mix (10 mM), 0.8 ml of forward and reverse primers (10 mM), 0.3 ml Taq DNA polymerase (5 U/ml) (Biotools) and double– distilled water (ddH2O) for a total volume of 50 ml. The PCR cycles were: 94 ºC (4 min), 40 cycles of 94 ºC (45 s), 56 ºC (1 min), 72 ºC (1 min) and a final extension at 72 ºC (10 min). Amplifications were purified for sequencing both forward and reverse using BigDye Terminator and an ABI 3730 genetic analyser (SECUGEN, Madrid). Chromatograms resulting from each specimen reaction were combined and primers were cleaned with Sequencher (Gene Codes Corporation). To compare the sequences obtained with those from other European populations, we downloaded GenBank available sequences representing the different haplotypes cited mainly by Wiehler and Tiedemann (1998), Versini et al. (2002), Randi et al. (2004) and Royo et al. (2007). We used the complete control region sequence of Capreolus pygargus and part of one of its subspecies, Capreolus pygargus ochracea, as an outgroup to root the phylogenetic tree. Two treatments were used to analyse these mitochondrial DNA data, phylogenetic inferences and networks (using Haploviewer, http://www.cibiv. at/~greg/haploviewer) based on the haplotypes, their frequencies, and the localities where each sample was collected. For the phylogenetic inferences, we used maximum parsimony (using PAUP; Swofford, 2003). Both the complete matrix of data, and a matrix reduced to 336 characters were prepared, the latter to avoid the problem of missing data, due to the reduced length of some GenBank sequences. A Mantel test was used to calculate the coefficient of association between genetic distance and altitude. Microsatellite analyses We amplified 12 unlinked microsatellite loci (BM1706, BM757, BM848, CSSM39, HUJ1177, NVHRT48, BMC1009, CSSM41, CSSM43, IDVGA29, IDVGA8 and OarFCB304) which had been previously used in other roe deer studies, with the concentrations there referred (Galán et al., 2003; Coulon et al., 2004). Forward primers were labelled with 3 different fluorescent dyes (6–FAM, HEX or NED). Fragment sizes were analysed on an ABI 3730 sequencer, and individuals were genotyped using the program GenMapper 3.1 (Do and Rahm, 2004). We tested for deviation from Hardy–Weinberg equilibrium at each locus and over all using the

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5000

5000 m

National Park Location code

Fig. 1. Location of the study area (Sierra de Guadarrama, Madrid), in the centre of the Iberian peninsula, where the roe deer muscle samples were obtained. Circles contain the location code (table 1). Fig. 1. Mapa de localización de la zona de estudio (Sierra de Guadarrama, Madrid) en la península ibérica donde se obtuvieron las muestras de músculo de corzo. Cada círculo contiene el código de la localidad (tabla 1).

Fisher's exact test and Markov chain algorithms (1,000 batches, 10,000 iterations), and FIS (Weir and Cockerham, 1984) was calculated using the program Genepop 3.4 (Raymond and Rousset, 1995). To as-

sess the level of genetic diversity, we calculated the mean expected and observed heterozygosities using Arlequin 3.0 (Schneider et al., 2000). In addition, we calculated the molecular coancestry coefficient and

Table 1. Data on tissue samples collected from roe deer analyzed during the study: L, location code; N, number of samples; and H, haplotypes found. Tabla 1. Datos sobre las muestras de tejido de corzo recogidas y analizadas durante el estudio: L, código de cada localidad; N, número de muestras; y H, haplotipos encontrados. L N H

L N H

13 4

A, E

2 3 C

A, C

3 3 C

A, C

4 1 A

A, B, C

5 1 C

17 1

A, C

A, B

19 5

A, B, C, E C

8 2 C

20 3

9 1 D

21 1

A, B

22 1

A, C

23 1

A, C

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Guadarrama France Northern Europe Italy Eastern Europe Iberian peninsula

5 10 20 40

Fig. 2. Network based on 336 base pairs of the mitochondrial D–loop of roe deer analysed (Sierra de Guadarrama) and the available sequences of several European populations. Circles represent the different haplotypes and are proportional to their frequency among the analysed data, as indicated. Fig. 2. Red formada por los 336 pares de bases de la región D–loop del ADN mitocondrial del corzo analizados (Sierra de Guadarrama) y las secuencias disponibles de varias poblaciones europeas. Los círculos representan los distintos haplotipos y, como se indica, son proporcionales a su frecuencia relativa entre los datos analizados.

PIC (Polymorphism Information Content) for each locus using the program Molkin 2.0 (Gutiérrez et al., 2005). To define the genetic structure of the populations of roe deer in the study area we used the program Structure 2.1 (Pritchard and Wen, 2003). This program uses a Bayesian algorithm to group samples into genetically distinct clusters, K, based on the similarities between the genotypes of individuals. We tested K = 1–15, with 10 replicates for each K–level. The most likely number of clusters was identified using the posterior probability (Pritchard and Wen, 2003). Results Mitochondrial DNA analysis The complete D–loop region (928–930 base pairs) was sequenced for the 101 studied samples (GenBank accession numbers MG760247 to MG760343). A maximum of only five gaps was necessary to align the complete data set, rendering a matrix length of 933 positions. We found 13 parsimony–informative characters in this matrix that exclusively contained the newly sequenced samples. Subsequently, sequences from GenBank were incorporated, including the different haplotypes previously found, most of them from partially sequenced D–loop, with lengths from 340–342, 427, 678 to the complete sequence, rea-

ching a matrix of 201 samples and 936 characters. In this last matrix, the number of parsimony informative characters increased to 62. The 101 individuals analysed comprised five different haplotypes, A, B, C, D and E (table 1), grouped in four clusters. Differences between haplotypes A and E were due to 9 polymorphic positions in haplotype E, probably because of some degree of heteroplasmy. To avoid spurious results and reticulations, haplotype E was eliminated from the matrix to the following data treatments. The network constructed showed the complex relationships between the samples from different locations (fig. 2). Usually in this kind of figure, all the haplotypes are included to show the relative weight of each haplotype and deduce the potential centre of dispersion. Here we have only represented the different haplotypes or those representatives of groups of haplotypes cited in the principal roe deer D–loop analyses (Wiehler and Tiedemann, 1998; Versini et al., 2002; Randi et al., 2004; Royo et al., 2007) plus the samples analysed here. This rendered a matrix of 68 haplotypes of 201 specimens. The large number of haplotypes in the Italian and French populations stands out, distributed throughout the network. The Iberian samples were situated in the centre of this representation, and their different haplotypes showed more similarity to other groups, than among them. By analysing the phylogenetic relationship of the total matrix (201 specimens, 940 pb), the unique-

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66 62 85 93

51 56

C. capreolus from France, Germany, Italy, Serbia, Sweden and Norway

C C. capreolus from Germany, Poland, Slovakia, Netherlands, Serbia, Italy and France 54

63 78

A 64 81 100

B 85

–0,0005 substitutions/site

C. capreolus from France, Portugal, Italy and Spain

67 100

C. pygargus

Fig. 3. Maximum parsimony phylogenetic tree for European roe deer. The numbers on the branches indicate bootstrap values above 50 %. A, B, C and D are the haplotypes found in this study. Fig. 3. Árbol filogenético basado en el principio de la máxima parsimonia para el corzo europeo. Los números sobre las ramas indican los valores de bootstrap superiores al 50 %. A, B, C y D son los haplotipos encontrados en este estudio.

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C. capreolus from Los Alcornocales (Cádiz)

C. capreolus from Sueve (Asturias) B C. capreolus from Muniellos (Asturias) Los Alcornocales (Cádiz)

C. capreolus from Sierra de Guadarrama (Madrid)

D 74

C. capreolus from Muniellos (Asturias) Pobla de Segur (Lérida) Picos de Europa (Asturias) C. capreolus from Muniellos (Asturias) Pobla de Segur (Lérida) Picos de Urbión (Soria), Ancares (Galicia) Sueve (Asturias) C. pygargus

–0,0005 substitutions/site Fig. 4. Maximum parsimony phylogenetic tree with individuals analysed in this study together with the haplotypes found by Royo et al. (2007) for populations in the Iberian peninsula. The numbers in the circles indicate the four main nodes, and those located on the branches indicate bootstrap values of above 50 %. A, B, C and D are the haplotypes found in this study. Fig. 4. Árbol filogenético basado en el principio de la máxima parsimonia con los individuos analizados en este estudio junto con los haplotipos encontrados por Royo et al. (2007) para la península ibérica. Los números dentro de los círculos señalan los cuatro nodos principales y los situados sobre las ramas indican los valores de bootstrap superiores al 50 %. A, B, C y D son los haplotipos encontrados en este estudio.

ness of some of the specimens from the Sierra de Guadarrama was confirmed and the relationships with some European haplotypes (fig. 3) was again demonstrated. Furthermore, this study showed the non–monophyly (common source) of Iberian haplotypes, since some appeared more related to other European haplotypes than with others found in the Sierra de Guadarrama. Fundamentally, lineages A and B were related to others analysed from Spain and Portugal, but also to some from France

and Italy. The C group was related to haplotypes from Germany, Poland, Slovakia, the Netherlands, Serbia, Italy and France, while the groups related to the unique roe deer with haplotype D were mostly from central and northern Europe (France, Germany, Italy, Serbia, Sweden and Norway). However, the most part of the groups signalled were hardly supported (moderate to low bootstrap values), since the number of differences between sequences was limited.

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2,200

Altitude

1,800 1,400 1,000 600 200 A

C C1 Haplotypes

Fig. 5. Relationship between genetic distance and altitude. Fig. 5. Relación entre distancia genética y altitud.

In figure 4, centring attention on the Iberian samples, we divided the clades into four groups. The first of these groups included specimens with A and B haplotypes, where haplotype 'A' coincided exactly with one haplotype found in the study by Royo et al. (2007). This coincident haplotype of Royo et al. (2007) comes from Alcornocales (Cádiz, south of Spain). In this first cluster, there were others Iberian specimens coming from Asturias (Sueve and Picos de Europa), Asturias (Muniellos), and France (Bordeaux), although the latter location is not decisive, as reintroductions from different sources were conducted. In our haplotype clustering C, in the group numbered 2 (fig. 4), there was also an individual sequenced by Royo et al. (2007), also coming also from the Sierra de Guadarrama. Group 3 included animals presenting the 'D' haplotype, which coincided with one haplotype sequenced by Royo er al. (2007) from northern locations such as the Picos de Europa or Muniellos, related to a lineage composed of specimens from the Picos de Europa, plus others from Pobla de Segur (Lleida) and Bordeaux. The group marked 4 was composed entirely of specimens from the northern Iberian peninsula. An analysis based on the Mantel test (fig. 5) showed that genetic distance was not significantly correlated with altitude (p > 0.05), suggesting that altitude was not the principal factor influencing genetic differentiation in the 101 individuals analysed. Microsatellite analysis Of the 12 loci analysed, 11 were polymorphic (table 2), rendering a polymorphism equal to 0.91. The average number of alleles for the population was 4.75 ± 0.96, varying between 2 (CSSM41, CSSM43 and IDVGA29) and 13 (OarFCB304). The

PIC reached its lowest value for one of the less polymorphic loci (PIC IDVGA29 = 0.1035) and the highest value for one of the most polymorphic loci (PIC BM757 = 0.7465). The observed heterozygosity ranged from 0.11 (IDVGA29) to 0.73 (BM1706), with a mean of 0.43. The mean molecular coancestry, which characterizes the degree of genetic similarity among individuals in a population, was 0.5067 regardless of the informational value of different markers; the average value considering the PIC was 0.3701. The population deviated significantly from Hardy– Weinberg equilibrium (p < 0.05) due to 5 markers (BM757, HUJ1177, BMC1009, IDVGA8 and OarFCB304). These 5 markers had positive FIS (reflecting a deficit of heterozygotes). This was a general trend since most FIS values were positive. Only the microsatellites IDVGA29 and BM1706 with negative FIS, of –0.031 and –0.056, respectively, showed a weak excess of heterozygotes. The population FIS calculated by Weir and Cockerham (1984) was positive and equal to 0.1115. The various simulations with the computer program Structure 2.1 determined that the estimated log–likelihood of the data (lnPr (X/K) was maximum when K = 1, i.e, when considering a single population. When a factorial correspondence analysis was performed, where roe deer were in the space with a number of dimensions equal to the markers analysed, the specimens differed slightly, appearing overlapped because of the genotype similarity (fig. 6). The lower right area of figure 6 distinguished a unique individual, which could force the greater overlap of the other specimens that shared alleles other than this animal. This distinction is produced because this individual had a private allele at one of the loci analysed.

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Table 2. Summary of basic analysis of genetic variability for the population of roe deer: HO, observed heterozygosity; HE, expected heterozygosity under the Hardy–Weinberg equilibrium; FIS, was calculated according to Weir and Cockerham (1984). Values of the average number of alleles, average heterozygosity and FIS calculated for all loci are shown in the last row. Tabla 2. Resumen del análisis básico de variabilidad genética para la población de corzo: HO, heterocigosis observada; HE, heterocigosis esperada bajo equilibrio de Hardy–Weinberg; FIS, se calculó según Weir y Cockerham (1984). En la última fila aparecen los valores del número medio de alelos, la heterocigosis media y la FIS calculada para todos los loci. Locus name

Allele range

Nº alleles

HO HE

FIS

BM1706

205–251

0.7303 0.7085 –0.031

BM757

193–295

0.3529 0.7829 0.551

BM848

355–361

0.6235 0.6235

CSSM39

177–181

0.6046 0.6242 0.032

HUJ1177

198–226

0.5662 0.6990 0.191

NVHRT48 86 1 0 0 – BMC1009

277–281

0.1785 0.1851 0.036

CSSM41

121–123

0.3222 0.3880 0.170

CSSM43

240–246

0.4588 0.4960 0.075

IDVGA29

140–148

0.1162 0.1102 –0.056

IDVGA8

207–227

0.6363 0.6851 0.072

OarFCB304 151–193

0.6140 0.7543 0.187

4.75

0.4337 0.5047 0.112

Total

–

Discussion The combined analysis of two types of markers increased knowledge of the roe deer population in the Sierra de Guadarrama. Data from mitochondrial genomes (maternally inherited) and nuclear (from both parents) were complementary. Analysis of the variation in the mtDNA control region revealed a multiple relationship between roe deer from the Sierra de Guadarrama and those from the Iberian peninsula populations (Randi et al., 2004; Royo et al., 2007) and European populations (Douzery and Randi, 1997; Vernesi et al., 2002; Randi et al., 2004). This relationship could be interpreted as a multiple origin of the population, with migration coming from different lineages of roe deer to the mountains of Madrid, either naturally by range expansion or through reintroductions. If migration had been natural by expansion of the distribution range, the observed variation should be lower than in the centre of origin, as it would represent the limit distribution of these populations; however, it is inconsistent with the results obtained. The variation stands in the range that it was shown in other studies, even in the upper part of this range (Wiehler and Tiedemann, 1998; Vernesi et al., 2002; Randi et al., 2004). The possibility of re–introductions in this area is unlikely because there are no records of such strategies in the Community of Madrid (FIDA, 2008).

The most likely explanation is therefore the acceptance of the theories that argue for the existence of shelters in the Pleistocene in the Iberian peninsula (Gliozzi et al., 1997; Hufthammer and Aaris–Sørensen, 1998; Taberlet et al., 1998), followed by later recolonization routes to northern Europe (Hewitt, 1999), or the area remaining as an isolated geographical region that would not have been influenced as a source for postglacial recolonization (Lorenzini and Lovari, 2006). It should be noted that various proposals of phylogenetic relationships between haplotypes can be obtained when the characters are taken in whole or in part. The sequences provided by Royo et al. (2007) have different lengths. In the present study, the data were handled by taking all available information (930 pb) for the analyses together with the European available data or by reducing our matrix to the number of characters in common with other information studies (up to 446 pb), thereby obtaining various proposed relationships between the detected lineages. Lineages A to D showed different clustering, with C being more closely related to D in the global analysis, while in the treatment where only the specimens from the Iberian peninsula were considered, C was apparently closer to A and B. In any case, it should be kept in mind that most of the nodes were poorly supported, and thus, if the nodes not fully supported were collapsed, no incongruence occurred. Besides, in these proposals,

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Axe 2 (5.54 %)

–1

0 Axe 1 (5.57 %)

Fig. 6. Factorial correspondence analysis of the roe deer individuals in the Sierra de Guadarrama (Madrid, Spain), determined by the alleles detected in 12 microsatellite loci. The colours of the individuals indicate the different localities. Fig. 6. Análisis factorial de correspondencias de los individuos de corzo en la Sierra de Guadarrama (Madrid, España), determinado por los alelos detectados en 12 loci de microsatélites. Los colores de los individuos indican las diferentes localidades.

not only the relationship between Sierra de Guadarrama samples and other specimens from central or southern Iberian peninsula remained constant in different clusters, but also their relationship with samples from the north, contrary to what is expected according to the findings of Royo et al. (2007). This confirms the point made in the comparative analysis of the study population with other European studies (Douzery and Randi, 1997; Vernesi et al., 2002; Randi et al., 2004). As the history of very recent populations, from an evolutionary point of view, shows this lack of support, since the number of diagnostic characters is limited, in this study we considered other markers such as microsatellites, which could elucidate the genetic structure of the population as they are more variable than mitochondrial markers. In addition, results from the population structure and Mantel test suggest that the relationship between genetic diversity and altitude is not significant, and hence it is possible to hypothesize that the species has not had sufficient time for evolutionary differentiation to occur along an altitude gradient. Analysis of the genetic variability using microsatellite markers indicated that there is probably a single genetically homogeneous population. The existence of a single population can be expected since the roe deer population has recovered recently in the mountains of Madrid (Tellería and Virgós, 1997). Only if movements of recolonization arose from different populations or if different nuclei had remained isolated for a long time would a clear genetic structure be expected. This also implies that gene flow between groups of roe deer is adequate

and that, in principle, there are no major barriers that prevent adequate dispersion of individuals in the Madrid mountains. While this is a good sign, in order to preserve the population, it should be kept in mind that although the roe deer is a species with a high ecological plasticity (Tellería and Virgós 1997; Acevedo et al., 2005), it moves in small areas and is therefore highly susceptible to fragmentation and the presence of barriers (Coulon et al., 2004). In–depth analysis of the genetic variability reveals that the population deviates from the Hardy– Weinberg equilibrium, so it is violating some of the assumptions of this principle (infinite population panmictic reproduction, absence of migration, mutation or selection). This was shown in five markers with positive values of FIS. This parameter has values ranging from –1 to 1, with negative values indicating an excess of heterozygotes and positive values indicating a deficit. Thus, the results point to some inbreeding in the population. Studies in other populations of roe deer (Wang and Schreiber, 2001; Coulon et al., 2004) also found deviations from the Hardy–Weinberg equilibrium due to a deficit of heterozygotes. In a study on the population in south–western France (Coulon et al., 2004), the same set of markers used in the present study (except NVHRT48) was analysed in 1,148 individuals in an area of 40 x 55 km. Two of the four microsatellites showing deficit of heterozygotes in the French population also showed this deficit in the Sierra de Guadarrama population. The average number of alleles and the heterozygosity values were higher than those observed here and the FIS

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for this population was lower, but also positive. The results suggest that, in comparison with other studies with similar methodologies in European populations (Coulon et al., 2004), the genetic variability of the Guadarrama population is somewhat impoverished, although respect to the Spanish populations, it shows a genetic variability similar to or even superior to other more southern populations (Lorenzini et al., 2003; Royo et al., 2007). The study by Royo et al. (2007) shared two microsatellites used in this study, NVHRT48 and BM757. These authors found four and 20 alleles for each marker respectively, while in our case, BM757, being one of the most polymorphic, showed only eight alleles and NVHRT48 was monomorphic. Although nine different populations were analysed in the study by Royo et al. (2007) without detailing the number of alleles in each of them, it should be noted that if in the rest of the Iberian peninsula four alleles were present for NVHRT48, in the population of Sierra de Guadarrama the variation was reduced to only one. The average coefficient of coancestry helps us to analyse these results in more detail, giving an idea of how individuals are alike. Analyses in other populations of roe deer found mean coancestry values ranging between 0.277 and 0.476 (Royo et al., 2007). If we compare these results with the average molecular coancestry in our study, the value in the Madrid population was higher, indicating lower polymorphism or increased homozygosity, related to some inbreeding, probably due to the bottleneck the population underwent in the recent past (Tellería and Virgós, 1997; Gortázar et al., 2000). However, it should be kept in mind that the set of microsatellites used in the two studies and the population sizes were different, so it is more appropriate to compare the coancestry coefficient corrected by PIC. In this case, our value falls within the range found by Royo et al. (2007) and the value that these authors found for the 'central' population, including individuals from the Sierra de Guadarrama and Toledo (average coancestry coefficient of 0.373). In conclusion, the combination of both results provided part of the complex story of the roe deer in the Iberian peninsula and especially in the Sierra de Guadarrama. On one hand, there is a philopatric structure with females maintaining the variation in mitochondrial haplotypes, clustered in four different groups without a common origin. These haplotypes are related to those from different populations of the Iberian peninsula and Europe. This complex origin leads to the consideration that no C. capreolus subspecies can be identified in the Iberian peninsula, agreeing in this case with Royo et al. (2007). On the other hand, as males contribute to the homogenisation of the genome, the population cannot thus be subdivided into different breeding units, but has to be considered as a single panmictic population. Additionally, heterozygosity was below the expected level in most markers and allelic richness was generally lower than in other populations. This provides evidence of a relative genetic impoverishment of the population. In view of the above, the different times of the palaeohistory of the species may explain these re-

sults: glacial–interglacial stages of the Pleistocene (Gliozzi et al., 1997; Hufthammer and Aaris–Sørensen, 1998) and the reduction and recovery of populations throughout the 20th century (Tellería and Virgós, 1997; Gortázar et al., 2000). The two markers used, mitochondrial DNA of maternal inheritance and nuclear DNA of both parents, have a different heritage and unequal mutation rate, reflected in each of these two main stages in the recent evolution of the roe deer. Our results have a strong implication for the management and conservation of roe deer in the Sierra de Guadarrama. Fortunately, in recent years, the roe deer population in the Iberian peninsula has benefited from a decline in ranching, abandonment of agriculture in foothill areas, and a significant increase in forest areas (Acevedo et al. 2005), all positive events to avoid population isolation. Even so, to preserve the populations, management must keep in mind that the species is highly susceptible to fragmentation and to barriers (Coulon et al., 2004) such as highways and fencing. Finally, taken together, our results indicate that restocking within the Sierra de Guadarrama should be prevented if we want to preserve the diversity among populations and the patterns of natural gene flow. Genetic assessment of structure and connectivity of roe deer populations that recently recolonized a fragmented landscape could be a promising approach for future studies. Acknowledgements We are grateful to the Consejería de Medio Ambiente y Ordenación del Territorio de la Comunidad de Madrid, for making this study in the Sierra de Guadarrama possible. We also thank M. J. Ruiz for her full involvement in genetic testing. Thanks too to R. Fernández–Mellado, A. Navarro–Castilla and the Editor and referees for their comments and improvement of the manuscript. References Acevedo, P., Delibes–Mateos, M., Escudero, M. A., Vicente, J., Marco, J., Gortázar, C., 2005. Environmental constraints in the colonization sequence of roe deer across the Iberian Mountains, Spain. Journal of Biogeography, 32: 1671–1680. Buitrago, A. M., 1992. Estudio de los Artiodáctilos del yacimiento del Pleistoceno medio de Pinilla del Valle (Madrid). PhD thesis, Universidad Complutense de Madrid, España. Coulon, A., Cosson, J. F., Angibault, J. M., Cargnelutti, B., Galan, M., Morellet, N., Petit, E., Aulagnier, S., Hewison, A. J., 2004. Landscape connectivity influences gene flow in a roe deer population inhabiting a fragmented landscape: an individual–based approach. Molecular Ecology, 13: 2841–2850. Do, H. H., Rahm, E., 2004. Flexible integration of molecular–biological annotation data: The genmapper approach. In: Advances in database technology–EDBT 2004: 811–822 (E. Berlino, S. Christodoulakis, D. Plexousakis, V. Christophi-

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des, M. Koubarakis, K. Böhm, E. Ferrari, Eds.). Springer, Berlin Heidelberg. Douzery, E., Randi, E., 1997. The mitochondrial control region of Cervidae: evolutionary patterns and phylogenetic content. Molecular Biology and Evolution, 14: 1154–1166. FIDA (Fundación para la Investigación y el Desarrollo Ambiental), 2008. Filogenia, evolución y variabilidad genética. In: Ecología y caracterización genética de las poblaciones de corzo de la Comunidad de Madrid: 79–95 (Fundación para la Investigación y el Desarrollo Ambiental, Eds.). Consejería de Medio Ambiente, Vivienda y Ordenación del Territorio, Madrid. Galán, M., Cosson, J. F., Aulagnier, S., Maillard, J. C., Thévenon, S., Hewison, A. J. M., 2003. Cross– amplification tests of ungulate primers in roe deer (Capreolus capreolus) to develop a multiplex panel of 12 microsatellite loci. Molecular Ecology Notes, 3: 142–146. Gliozzi, E., Abbazzi, L., Argenti, P., Azzaroli, A., Caloi, L., Capasso Barbato, L., Torre, D., 1997. Biochronology of selected mammals, molluscs and ostracods from the middle pliocene to the late pleistocene in Italy. The state of the art. Rivistaitaliana di paleontologia e stratigrafia, 103: 369–388. Gortázar, C., Herrero, J., Villafuerte, R., Marco, J., 2000. Historical examination of the status of large mammals in Aragón, Spain. Mammalia, 64: 411–422. Gutiérrez, J. P., Royo, L. J., Álvarez, I., Goyache, F., 2005. MolKin v2.0: A computer program for genetic analysis of populations using molecular coancestry information. Journal of Heredity, 96: 718–721. Hewitt, G. M., 1999. Post–glacial re–colonization of European biota. Biological Journal of the Linnean Society, 68: 87–112. Hufthammer, A. K., Aaris–Sørensen, K., 1998. Late–and postglacial European roe deer. In: The European roe deer: the biology of success: 47–69 (R. Andersen, P. Duncan, J. D. C. Linnell, Eds.). Scandinavian university press, Oslo. Jäger, F., Hecht, W., Herzog, A., 1992. Untersuchungenanmitochondrialer DNS (mtDNS) von hessischemRehwild (C. capreolus). Zeitschriftfür Jagdwissenschaft, 38: 26–33. Lorenzini, R., Garofalo, L., Qin, X., Voloshina, I., Lovari, S., 2014. (Artiodactyla: Cervidae), a Palaearctic meso–mammal. Zoological Journal of the Linnean Society, 170: 209–221. Lorenzini, R., Lovari, S., 2006. Genetic diversity and phylogeography of the European roe deer: the refuge area theory revisited. Biological Journal of the Linnean Society, 88: 85–100. Lorenzini, R., San José, C., Braza, C., Aragón, S., 2003. Genetic differentiation and phylogeography of roe deer in Spain, as suggested by mitochondrial DNA and microsatellite analysis. Italian Journal of Zoology, 70: 89–99. Pritchard, J. K., Wen, W., 2003. Documentation for STRUCTURE software: version 2. Available from: http://pritch.bsd.uchicago.edu. Randi, E., Alves, P., Carranza, J., Milosevi–Zlata-

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novi, S., Sfougaris, A., 2004. Phylogeography of roe deer (Capreolus capreolus) populations: the effects of historical genetic subdivisions and recent nonequilibrium dynamics. Molecular Ecology, 13: 3071–3083. Randi, E., Pierpaoli, M., Danilkin, A., 1998. Mitochondrial DNA polymorphism in populations of Siberian and European roe deer (Capreolus pygargus and C. capreolus). Heredity, 80: 429–437. Raymond, M., Rousset, F., 1995. Genepop (version 1.2): population genetics software for exact tests and ecuminicism. Journal of Heredity, 86: 248–249. Royo, L. J., Pajares, G., Álvarez, I., Fernández, I., Goyache, F., 2007. Genetic variability and differentiation in Spanish roe deer (Capreolus capreolus): A phylogeographic reassessment within the European framework. Molecular Phylogenetics and Evolution, 42: 47–61. Sambrook, J., Fritsch, E. F., Maniatis, T., 1989. Molecular Cloning: a Laboratory Manual, 2nd Edition. Cold Spring Harbor Laboratory Press, New York. Schneider, S., Roessli, D., Excoffier, L., 2000. Arlequin v.2.0: a software for population genetics data analysis. University of Geneva, Geneva, Switzerland. Sommer, R. S., Zachos, F. E., 2009. Fossil evidence and phylogeography of temperate species: 'glacial refugia' and post–glacial recolonization. Journal of Biogeography, 36: 2013–2020. Swofford, D. L., 2003. PAUP*: Phylogenetic analysis using parsimony (*and other methods), version 4.0b 10. Sinauer associates, Sunderland, Massachusetts. Taberlet, P., Fumagalli, L., Wust–Saucy, A. G., Cosson, J. F., 1998. Comparative phylogeography and postglacial colonization routes in Europe. Molecular Ecology, 7: 453–464. Tellería, J. L., 1999. La diversidad de vertebrados del Valle del Paular (Madrid). In: Primeros Encuentros Científicos del Parque Natural de Peñalara: 155–162 (Consejería de Medio Ambiente, Eds.). Comunidad de Madrid, Madrid. Tellería, J. L., Virgós, E., 1997. Distribution of an increasing roe deer population in fragmented Mediterranean landscape. Ecography, 20: 247–252. Vernesi, C, Pecchioli, E., Caramelli, D., Tiedermann, R., Randi, E., Bertorelle, G., 2002. The genetic structure of natural and reintroduced roe deer (Capreolus capreolus) populations in the Alps and central Italy, with reference to the mitochondrial DNA phylogeography of Europe. Molecular Ecology, 11: 1285–1297. Wang, M., Schreiber, A., 2001. Impact of the habitat fragmentation and social structure on the population gentics o roe deer (Capreolus capreolus L.) in Central Europe. Heredity, 86: 703–715. Weir, B. S., Cockerham, C. C., 1984. Estimating F– statistics for the analysis of population structure. Evolution, 38: 1358–1370. Wiehler, J., Tiedemann, R., 1998. Phylogeography of the European roe deer (Capreolus capreolus) as revealed by sequence analysis of the mitochondrial control region. Acta Theriologica, 5: 187–197.

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Isotopic niche overlap between the invasive leiothrix and potential native competitors E. Pagani–Núñez, M. Renom, C. Furquet, J. Rodríguez, F. Llimona, J. C. Senar

Pagani–Núñez, E., Renom, M., Furquet, C., Rodríguez, J., Llimona, F., Senar, J. C., 2018. Isotopic niche overlap between the invasive leiothrix and potential native competitors. Animal Biodiversity and Conservation, 41.2: 427–434. Abstract Isotopic niche overlap between the invasive leiothrix and potential native competitors. We analysed niche overlap between the red–billed leiothrix Leiothrix lutea, a spreading exotic invasive bird, and the European robin Erithacus rubecula and the blackcap Sylvia atricapilla, similar native species, which are commonly considered as potential competitors in Mediterranean forests. We analyzed stable isotope ratios (δ13C and δ15N) from feathers of birds sampled in several locations within the Collserola mountain range (Barcelona, NE Spain) where leiothrix have strongly increased their numbers in the last decade and quantified niche overlap with nicheROVER. Blackcap individuals showed the lowest probability to be found in the isotopic niche of the other two species (around 60 %), while leiothrix and robins showed a high probability to share the same isotopic niche (82 %). Our results stress that competition for resources is potentially high and the species shared marked niche asymmetries, which may have implications for community dynamics in the long term. Key words: Interspecific competition, Niche overlap, Niche asymmetries, Red–billed leiothrix, Stable isotopes Resumen Solapamiento de nichos isotópicos entre el ruiseñor del Japón, invasivo, y los posibles competidores nativos. Analizamos el solapamiento de nichos entre el ruiseñor del Japón, Leiothrix lutea, que es un ave exótica invasiva en expansión, y el petirrojo europeo, Erithacus rubecula, y la curruca capirotada, Sylvia atricapilla, que son especies nativas parecidas que se suelen considerar competidores potenciales en los bosques mediterráneos. Analizamos la relación de isótopos (δ13C y δ15N) en las plumas de las aves capturadas en varios lugares dentro de la cordillera de Collserola (Barcelona, NE España) en los que la población de ruiseñor del Japón ha aumentado notablemente en el último decenio, y cuantificamos el solapamiento de nichos con nicheROVER. Los individuos de curruca presentaron la menor probabilidad de encontrarse en el nicho isotópico de las otras dos especies (alrededor del 60 %), mientras que el ruiseñor y el petirrojo mostraron una elevada probabilidad de compartir el mismo nicho isotópico (el 82 %). Nuestros resultados destacan que la competencia por los recursos puede ser elevada y que las especies comparten marcadas asimetrías de nichos, lo que puede tener consecuencias para la dinámica de comunidades a largo plazo. Palabras clave: Competencia interespecífica, Solapamiento de nichos, Asimetrías de nichos, Ruiseñor del Japón, Isótopos estables Received 29 VI 18; Conditional acceptance: 30 VII 18; Final acceptance: 05 IX 18 Emilio Pagani–Núñez, State Key Lab. of Biocontrol, Dept. of Ecology/School of Life Sciences, Sun Yat–sen Univ., No.135 West Xingang Road, 510275 Guangzhou, People’s Republic of China.– María Renom, Juan Carlos Senar, Museu de Ciències Naturals de Barcelona, Psg. Picasso s/n., 08003 Barcelona, Spain.– Carles Furquet, Institut Català d’Ornitologia (ICO), Museu de Ciències Naturals de Barcelona, Psg. Picasso s/n., 08003 Barcelona, Spain.– Jordi Rodríguez, Grup d’Anellament Parus, Institut Català d’Ornitologia (ICO), c/ Salze 36, 08186 Lliçà d’Amunt, Barcelona, Spain.– Francesc Llimona, Estació Biològica Can Balasc, Consorci del Parc Natural de Collserola, ctra. de l’Església 92, 08017 Barcelona, Spain. Corresponding author: Emilio Pagani–Núñez. E–mail: pemilio@mail.sysu.edu.cn ISSN: 1578–665 X eISSN: 2014–928 X

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Introduction Invasive species are a major threats to biodiversity (Lockwood et al., 2007; Kolar and Lodge, 2001) but they also offer an opportunity to study ecological and evolutionary processes (Mooney and Cleland, 2001). A topic of current interest is the process by which animals become successfully established in a newly invaded area and how they accommodate within the host community (Marsico et al., 2010; Henriksson et al., 2016; Sol and Maspons, 2016; Duncan et al., 2003). Two main hypotheses have been suggested: the competition hypothesis and the opportunistic hypothesiss (Duncan et al., 2003). The competition hypothesis suggests that the main factor facilitating successful colonization of novel environments is the superior competitive ability of the invader over native species (Duncan et al., 2003; Petren and Case, 1996; Duncan, 1997). This assumes that invasive and native species show high niche overlap, which may drive intense competition. The opportunistic hypothesis, however, proposes that the invading species does not compete with natives, but simply takes advantage of niche opportunities not used by native species (Duncan et al., 2003; Shea and Chesson, 2002; Batalha et al., 2013). However, it has been difficult to find clear empirical support for either of these hypotheses (Duncan et al., 2003). The red–billed leiothrix Leiothrix lutea (hereinafter leiothrix) exemplifies this controversy. A worldwide– scale meta–analysis of avian exotic introductions identified the leiothrix as one of the bird species with the highest negative local impact on bird communities (Martin–Albarracin et al., 2015). This species has successfully invaded several regions, from Japan to Hawaii and SW Europe (Herrando et al., 2010; Male et al., 1998; Tojo and Nakamura, 2004; Pereira et al., 2017), while it is native to South Asia (from West Himalaya to South and Central China) (Collar et al., 2018). A recent work in the Iberian Peninsula, based on behavioural observations during foraging and on habitat use, concluded that leiothrix exhibited little niche overlap with most native species, and that this invasion had relatively few consequences for the populations of the other species (Vall–llosera et al., 2016). As a consequence, they suggested that the success of leiothrix in this area could be explained by the opportunistic hypothesis. However, another two works, also conducted in Iberia, used morphology, diet and exploratory behaviour to show a competitive advantage of leiothrix over native species rather than an opportunistic occupation of an empty ecological niche (Pereira et al., 2017). Behavioural observations also supported interference competition by an active displacement of native species such as European robins Erithacus rubecula and blackcaps Sylvia atricapilla through aggressive attacks of leiothrix (Pereira et al., 2018). Discrepancy between the two hypotheses about the establishment of the leiothrix in Iberia probably lies in the different approaches used by different authors. However, resolving this discrepancy may have relevant consequences in the design of con-

servation policies in the areas invaded by leiothrix. The aim of our work was therefore to shed new light on this controversy by using stable isotopes as an alternative approach. Trophic interactions are fundamental to study the impact of exotic on native species (David et al., 2017). Comparing the niches of invasive and native species in the invasion distribution range has been suggested as a good approach to assess the occurrence of interspecific competition (Cornell and Lawton, 1992). Analyses using stable isotopes ratios of δ13C and δ15N, which are mostly indicative of habitat use and diet respectively (Pagani–Núñez et al., 2017), can help understand niche overlap over longer time spans than behavioural observations (Pagani–Núñez et al., 2017; Inger and Bearhop, 2008; Layman et al., 2012). Moreover, stable isotopes are commonly used to assess interspecific competition (see e.g. Karlson et al., 2015; Smith et al., 2017). This is why it has been suggested as a powerful tool to study niche overlap between native and invasive species (Jackson et al., 2012; Kamenova et al., 2017). In this paper we compared the isotopic niche of the leiothrix with that of the European robin (hereinafter robin) and the blackcap, which have been identified as the two main native competitors of leiothrix in the Mediterranean area (Pereira et al., 2017). We assessed niche overlap among these species using nicheROVER, a new statistical tool developed by Swanson et al. (2015). This allowed us to contrast the competition and niche opportunity hypotheses in wild conditions and using the most advanced research techniques. Material and methods This study was carried out in the Natural Park of Collserola (45° 27' N, 2° 8' E; Catalonia, northeast Spain). Our study area may be characterized as Mediterranean mixed forest (Pagani–Núñez et al., 2014). The three main tree species were Aleppo pine Pinus halepensis, holm oak Quercus ilex and the oak Q. cerrioides. The first leiothrix was captured in this area on 14th July 1990 (La Espinagosa, Vallvidrera; C. Jordà, J. L. Copete and J. C. Senar, pers. obs.). Since then, the species has spread exponentially along the whole park, across forest and riverside zones, now reaching the close mountain range of Serra de Marina (Herrando et al., 2010). We used data from field surveys to illustrate and quantify this pattern (see below). The study is based on 67 leiothrix, 43 robins and 20 blackcaps captured in the area using mist nets and funnel traps (Senar et al., 1997) between January and August 2013. We took the 5th most external tail feather of the right side of all individuals for laboratory analyses. Given the long sampling period, some of the sampled individuals had grown their feathers in 2012 and some in 2013, so we assessed year–related effects on our sample. There was a significant effect of year on δ13C, with δ13C being more negative in 2013 than in 2012 feathers, but not on δ15N (data

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now shown). We thus extracted residuals of a model including δ13C as the dependent variable and year as the categorical factor. We added average δ13C to these residuals and obtained a standardized δ13C, which we used in our analysis. The three species, leiothrix, robins and blackcaps, make their nests in shrubs, have similar body sizes, and show a comparable low degree of sexual dimorphism (Tellería and Carbonell, 1999; Ellrich et al., 2010; Pagani–Núñez et al., 2013). They coexist throughout the year in this area and have generalist diets, feeding on fruit and small invertebrates (Herrera, 1998; pers. obs.). They also show similar moult strategies, with the three species carrying out a partial moult in their first year of life and a complete moult subsequently (see Jenni and Winkler (1994) for information on robins and blackcaps, further research is needed to confirm this pattern in leiothrix). They may differ, however, in their migratory strategies. Robins are partial migrants displaying short– to long–distance movements (Collar, 2018), although in our study area they are sedentary (pers. obs.). Blackcaps have both sedentary and short–distance migratory populations (Aymí et al., 2018), although in order to avoid trapping birds that had moulted away from our study area, most of our birds (18/20) were sampled during the breeding season. Leiothrix performs altitudinal migrations in its native distribution range (Collar et al., 2018; pers. obs.), but our invasive population seems fully resident (pers. obs.). We performed stable isotopes analyses of tail feathers in 2014. Feathers were stored in cold conditions (–20 ºC) within the shortest time possible after collection in the field. Tail feathers were cleaned in a solution of NaOH (0.25 M) and oven–dried at 40 ºC for 12 hours. We analysed feather tips (Vitz and Rodewald, 2012), which were carefully extracted using sterilized metal scissors. These subsamples of 0.35 mg were loaded into tin recipients and crimped for combustion for both δ13C and δ15N analyses. We used an elemental analysis–isotope ratio mass spectrometry (EA–IRMS) with a Flash 1112 (for C and N) elemental analyzer coupled to a Delta C isotope ratio mass spectrometer via a CONFLOIII interface (Thermo Fisher Scientific, Bremen, Germany). We carried out the laboratory work at the Scientific Technical Services Department at the University of Barcelona. We expressed stable isotope ratios as parts per thousand (%), according to the following equation: δX = [(Rsample/Rstandard)–1] * 1,000 where X is 13C or 15N and R is the corresponding ratio 13C/12C or 15N/14N. We referenced our samples against international standards: Pee Dee Belemnite (VPDB) for 13C, atmospheric nitrogen (AIR) for 15N. The measurement precisions were 0.15 % for δ13C and 0.25 % for δ15N. We characterized population trends of this exotic population of leiothrix. We pooled data from transects in habitats where oaks and pines predominate, carried out in forest and riverine zones during and

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outside the breeding season. Each transect was carried out two times during the breeding season and two times during the wintering season in the same area of the Natural Park of Collserola where we collected the feathers. Transects had a length of 1.5 km and all the leiothrix seen or heard were recorded within 50 m from the observer. We included data from 1998, when the species was detected for first time in our study area, to 2014. We analysed differences in δ13C and δ15N values between the leiothrix, robins and blackcaps using an analysis of variance. We computed niche overlap and asymmetries between leiothrix and potential competitors using the R package nicheROVER v1.0 (Swanson et al., 2015). We estimated isotopic niches (a 95% probability region based on δ13C and δ15N isotopic values of feathers) as the probability to find an individual of one species in the isotopic niche of the other species. We carried out 1,000 runs with a probability level of alpha = 0.95. These analyses were carried out in R (R Development Core Team, 2014). Results Our data illustrate a marked pattern of population growth for leiothrix in our study area (fig. 1). One–way analysis of variance showed significant differences in δ13C values between the three species (SS = 6.9, F2,127 = 4.37, P = 0.01) (fig. 2). Robins showed the least negative mean values of δ13C (mean ± SD = –23.27 ± 0.70), followed by blackcaps (–23.35 ± 1.14), while leiothrix showed the most negative values (–23.75 ± 0.91). In the case of δ15N, inter–specific differences were close to significance (SS = 21.50, F2,127 = 2.96, P = 0.05) (fig. 2). Blackcaps showed the highest mean values of δ15N (3.90 ± 3.03), followed by leiothrix (2.89 ± 1.43), while robins displayed the lowest mean values (2.66 ± 1.91). Blackcap individuals showed the lowest probability to be found in the niche regions of robins (66.15 %) and leiothrix (58.52 %), while these species showed an extremely high probability to be recorded in the blackcap niche region (98.73 % and 95.58 % respectively) (fig. 3). Leiothrix and robins showed a similarly high probability to be found in each other´s niche regions (leiothrix in robin’s niche region 82.48 %, robin in leiothrix's niche region 82.89 %) (fig. 3). Discussion The comparison of isotopic niches between leiothrix, an exotic invasive species, and robins and blackcaps showed a considerable (> 80 %) niche overlap between leiothrix and robins. Blackcaps, on the other hand, showed the broadest niche and a low probability to be found in the niches of these two other species. The major difference between species was a δ15N 1 ‰ higher in the blackcap, which suggests that this species consumed relatively more insects or insects of higher trophic levels (in opposition to fruits or in-

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Nº individuals/km

25 20 15 10

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Fig. 1. Graph showing the progression in number of red–billed leiothrix Leiothrix lutea in our study area (Collserola Natural Park, Barcelona) over the last sixteen years based on lineal transects (values and means and standard error; n = 4 transects per year in each of the seven habitat types within Collserola Natural Park). Fig. 1. Gráfico en el que se muestra la progresión del número de individuos del ruiseñor del Japón, Leiothrix lutea, en la zona de estudio (Parque Natural de Collserola, Barcelona) durante los últimos 16 años, basada en transectos lineales (valores, medias y errores estándar; n = 4 transectos por año en cada uno de los siete tipos de hábitat del Parque Natural de Collserola).

8 7 6

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Fig. 2. Bi–plot showing δ13C and δ15N mean values (± standard deviation) in the 5th rectrix of the three species considered in this study. The red square corresponds to red–billed leiothrix Leiothrix lutea (n = 67), the blue diamond corresponds to European robins Erithacus rubecula (n = 43), and the black triangle to blackcaps Sylvia atricapilla (n = 20). Fig. 2. Gráfico biplot en el que se muestran los valores medios de δ13C y δ15N (± desviación estándar) en la quinta rectriz de las tres especies analizadas en este estudio. El cuadro rojo corresponde al ruiseñor del Japón, Leiothrix lutea (n = 67); el rombo azul, al petirrojo europeo, Erithacus rubecula (n = 43), y el triángulo negro, a la curruca capirotada, Sylvia atricapilla (n = 20).

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δ13C

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100 30

Overlap probability (%) – niche region size 95 %

Fig. 3. Niche plots (random elliptical projections of trophic niche region for each species; up–right), density distributions (probability of the random variable to fall within the range of observed values; up–left: δ15N, down–right: δ13C), and raw data (down–left: scatterplot of δ13C vs δ15N) for the pairwise combination of δ13C and δ15N from feathers of red–billed leiothrix Leiothrix lutea, European robin Erithacus rubecula and blackcap Sylvia atricapilla (see Swanson et al. (2015) for a more detailed description). The numbers below the plots indicate the probability that an individual from the species indicated by row will be found within the niche of the species indicated by the column header. Fig. 3. Gráficos de nicho (proyecciones elípticas aleatorias de la región del nicho trófico para cada especie; arriba a la derecha), distribuciones de densidad (probabilidad de que la variable aleatoria se encuentre en el intervalo de valores observados; arriba a la izquierda: δ15N, abajo a la derecha: δ13C), y datos no elaborados (abajo a la izquierda: diagrama de dispersión de δ13C y δ15N) para la combinación por pares de δ13C y δ15N de las plumas del ruiseñor del Japón, Leiothrix lutea, el petirrojo europeo, Erithacus rubecula, y la curruca capirotada, Sylvia atricapilla (véase Swanson et al. (2015) para obtener una descripción más detallada). Los números que figuran debajo de los gráficos se refieren a la probabilidad de que un individuo de la especie indicada en el encabezamiento de la fila se encuentre en el nicho de la especie indicada en el encabezamiento de la columna.

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sects of lower trophic levels) than robins and leiothrix (see e.g., Pagani–Núñez et al., 2017). Given that during the winter there are fewerinsects available, this may indicate a potential constraint for blackcap populations. This supports the view of Pereira et al. (2017), based on morphology, that the robin is the species that can potentially suffer from competition with leiothrix to a greater extent than other species. This high trophic overlap among these species suggests that leiothrix use similar niches to native species, rather than exploiting empty niches (Vall–llosera et al., 2016). Thus, our results also suggest that leiothrix’s invasion success could be linked to its behavioural dominance over native species (Pereira et al., 2018). However, although an overlap in diet is a potential source of competition, it does not always mean that competition occurs (Cody, 1974; Dhondt, 2012), but it may in fact be attenuated by different factors. First, the different species could exploit the shared resources at different times of year, decreasing or avoiding competence (León et al., 2014; Gidoin et al., 2015). Nevertheless, since the three species moult during the same period over a short time window, approximately from July to August (Jenni and Winkler, 1994; pers. obs.), the diet overlap inferred among the three species corresponds to the simultaneous use of resources by the three species. However, this does not rule out the possibility the diet of our study species differ in different periods of the year out of our study period, in which case the competence could be relaxed. This possibility could be addressed by analysing other tissues, such as blood or nails, which are renewed by the birds at a shorter and more constant turnover rates, reflecting the diet over periods of a month or less, prior to extraction (Inger and Bearhop, 2008; McKechnie, 2004: Layman et al., 2012). This could also solve a drawback of our study, that in capturing some of the blackcaps and robins in winter, some of the individuals sampled could be migratory and have moulted in their breeding ranges; this could had broadened the isotopic niche of the species artificially. Nevertheless, this would not invalidate the results of our paper in that the robin highly overlapped in its diet with the leiothrix, and any sampling of non–resident birds would make our results more conservative. Moreover, most blackcap individuals were sampled during the breeding season (18 of 20), meaning that they were local birds, and most robins in the area are sedentary (pers. obs.). A third factor that may attenuate competence between leiothrix and robin could be that although they have a similar diet, the two species could use different foraging techniques or substrate to find their prey. For instance, they may use different sections of the same trees (such as inner and outer branches) and leiothrix seem to have a more diverse array of foraging techniques than robins. This is clearly the case of leiothrix in Japan and to some degree in NE Spain (Vall–llosera et al., 2016; Tojo and Nakamura, 2004; Amano and Eguchi, 2002). Similarly, the competence between species could be reduced by the use of different subareas within the study area. In line

with this, we found that leiothrix displayed the lowest mean δ13C values at the species level, which in our study area is indicative of foraging at valley bottoms (Pagani–Núñez et al., 2017). In fact, this is where most leiothrix were captured. However, the use of different areas or foraging techniques did not prevent the high levels of competence and negative impact of leiothrix on native species in Hawaii (Male et al., 1998; Mountainspring and Scott, 1985). A fourth factor relates to the niche overlap hypothesis, which establishes that when resources are very abundant, potential competitors can tolerate a relatively high degree of overlap in resource use without experiencing critical levels of competition (Rusterholz, 1981). This could be the case in the Mediterranean area, where food resources for these species (e.g. insects and fruits) are very abundant (Blondel et al., 2010). This could additionally be enhanced by the generalist nature of this community (Vall–llosera et al., 2016). To conclude, it is clear that the exotic invasive red–billed leiothrix shows high isotopic niche overlap with the European robin. However, further research is clearly needed to ascertain whether leiothrix success in these exotic locations is due to their high competitive ability, or to their capacity to expand their niches and to use particular features or resources of the invaded areas. Understanding this process may help us to prevent potential impacts of leiothrix on European forest birds. Acknowledgements We thank Ignasi Toranzo, Sergio Hernández–Gómez, Ángel Fernández, Antonio España, Seán Cahill, Daniel Díaz–Diethelm, Andrea Garmendia and David Meca for their invaluable help in the field, and Pili Rubio for her help in the lab. We also thank Jacob González Solís for his comments on the MS. This study was funded by research project CGL–2016–79568–C3– 3–P (to JCS) from the Ministry of Economy and Competitivity, Spanish Research Council and by the FPI grant BES2010–040359 (to EPN) from the Ministry of Science and Innovation, Spanish Research Council. Birds were handled with the permission of the Departament de Medi Ambient, Generalitat de Catalunya. Rings were provided by the Catalan Ringing Office (ICO). References Aymí, R., Gargallo, G., Christie, D. A., 2018. Eurasian Blackcap (Sylvia atricapilla). In: Handbook of the Birds of the World Alive (J. del Hoyo, A. Elliott, J. Sargatal, D. A. Christie, E. de Juana, E., Eds.). Lynx Edicions, Barcelona, https://www.hbw.com/ node/58952 [Accessed on September 5, 2018]. Amano, H. E., Eguchi, K., 2002. Foraging niches of introduced Red–billed Leiothrix and native species in Japan. Ornithological Science, 1: 123–131. Batalha, H. R., Ramos, J. A., Cardoso, G. C., 2013.

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Animal Biodiversity and Conservation 41.2 (2018)

Animal Biodiversity and Conservation Animal Biodiversity and Conservation és una revista interdisciplinària publicada, des de 1958, pel Museu de Ciències Naturals de Barcelona. Inclou articles d'inves tigació empírica i teòrica en totes les àrees de la zoologia (sistemàtica, taxo nomia, morfologia, biogeografia, ecologia, etologia, fisiologia i genètica) procedents de totes les regions del món amb especial énfasis als estudis que permetin compendre, desde un punt de vista pluridisciplinar i integrat, els patrons d'evolució de la biodiversitat en el seu sentit més ampli. La revista no publica compilacions bibliogràfiques, catàlegs, llistes d'espècies o cites puntuals. Els estudis realitzats amb espècies rares o protegides poden no ser acceptats tret que els autors disposin dels permisos corresponents. Cada volum anual consta de dos fascicles. Animal Biodiversity and Conservation es troba registrada en la majoria de les bases de dades més importants i està disponible gratuitament a internet a www.abc.museucienciesjournals.cat, de manera que permet una difusió mundial dels seus articles. Tots els manuscrits són revisats per l'editor executiu, un editor i dos revisors independents, triats d'una llista internacional, a fi de garantir–ne la qualitat. El procés de revisió és ràpid i constructiu. La publicació dels treballs acceptats es fa normalment dintre dels 12 mesos posteriors a la recepció. Una vegada hagin estat acceptats passaran a ser propietat de la revista. Aquesta es reserva els drets d’autor, i cap part dels treballs no podrà ser reproduïda sense citar–ne la procedència. Els drets d’autor queden reservats als autors, els qui autoritzen la revista a publicar l’article. Els articles es publiques amb una Llicència de Reconeixement 4.0 Internacional de Creative Commons: no es podrà reproduir ni reutilitzar cap part dels treballs publicats sense citar-ne la procedència.

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Robinson i Redford (1991)...", "...les prospeccions realitzades (Begon et al., 1999)...". Taules. Es numeraran 1, 2, 3, etc. i han de ser sempre ressenyades en el text. Les taules grans seran més estretes i llargues que amples i curtes ja que s'han d'encaixar en l'amplada de la caixa de la revista. Figures. Tota classe d’il·lustracions (gràfics, figures o fotografies) entraran amb el nom de figura i es numeraran 1, 2, 3, etc. i han de ser sempre ressenyades en el text. Es podran incloure fotografies si són imprescindibles. Si les fotografies són en color, el cost de la seva publicació anirà a càrrec dels autors. La mida màxima de les figures és de 15,5 cm d'amplada per 24 cm d'alçada. S'evitaran les figures tridimensionals. Tant els mapes com els dibuixos han d'incloure l'escala. Els ombreigs preferibles són blanc, negre o trama. S'evitaran els punteigs ja que no es reprodueixen bé. Peus de figura i capçaleres de taula. Seran clars, concisos i bilingües en la llengua de l’article i en anglès. Els títols dels apartats generals de l’article (Introducción, Material y métodos, Resultados, Discusión, Conclusiones, Agradecimientos y Referencias) no aniran numerats. No es poden utilitzar més de tres nivells de títols. Els autors procuraran que els seus treballs originals no passin de 20 pàgines (incloent–hi figures i taules). Si a l'article es descriuen nous tàxons, caldrà que els tipus estiguin dipositats en una institució pública. Es recomana als autors la consulta de fascicles recents de la revista per tenir en compte les seves normes. Comunicacions breus Les comunicacions breus seguiran el mateix procediment que els articles y tindran el mateix procés de revisió. No excediran de 2.300 paraules incloent–hi títol, resum, capçaleres de taula, peus de figura, agraïments i referències. El resum no ha de passar de 100 paraules i el nombre de referències ha de ser de 15 com a màxim. Que el text tingui apartats és opcional i el nombre de taules i/o figures admeses serà de dos de cada com a màxim. En qualsevol cas, el treball maquetat no podrà excedir de quatre pàgines.

Animal Biodiversity and Conservation 41.2 (2018)

Animal Biodiversity and Conservation Animal Biodiversity and Conservation es una revista inter disciplinar, publicada desde 1958 por el Museo Ciencias Naturales de Barcelona. Incluye artículos de investigación empírica y teórica en todas las áreas de la zoología (sistemática, taxo nomía, morfología, biogeografía, ecología, etología, fisiología y genética) procedentes de todas las regiones del mundo, con especial énfasis en los estudios que permitan comprender, desde un punto de vista pluridisciplinar e integrado, los patrones de evolución de la biodiversidad en su sentido más amplio. La revista no publica compilaciones bibliográficas, catálogos, listas de especies sin más o citas puntuales. Los estudios realizados con especies raras o protegidas pueden no ser aceptados a no ser que los autores dispongan de los permisos correspondientes. Cada volumen anual consta de dos fascículos. Animal Biodiversity and Conservation está registrada en todas las bases de datos importantes y además está disponible gratuitamente en internet en www.abc.museucienciesjournals.cat, lo que permite una difusión mundial de sus artículos. Todos los manuscritos son revisados por el editor ejecutivo, un editor y dos revisores independientes, elegidos de una lista internacional, a fin de garantizar su calidad. El proceso de revisión es rápido y constructivo, y se realiza vía correo electrónico siempre que es posible. La publicación de los trabajos aceptados se realiza con la mayor rapidez posible, normalmente dentro de los 12 meses siguientes a la recepción del trabajo. Una vez aceptado, el trabajo pasará a ser propiedad de la revista. Ésta se reserva los derechos de autor, y ninguna parte del trabajo podrá ser reproducida sin citar su procedencia. Los derechos de autor quedan reservados a los autores, quienes autorizan a la revista a publicar el artículo. Los artículos se publican con una Licencia Creative Commons Atribución 4.0 Internacional: no se podrá reproducir ni reutilizar ninguna de sus partes sin citar la procedencia.

Normas de publicación Los trabajos se enviarán preferentemente de forma electrónica (abc@bcn.cat). El formato preferido es un documento Rich Text Format (RTF) o DOC, que incluya las figuras y las tablas. Las figuras deberán enviarse también en archivos separados en formato TIFF, EPS o JPEG. Debe incluirse, con el artículo, una carta donde conste que el trabajo versa sobre investigaciones originales no publicadas anterior mente y que se somete en exclusiva a Animal Biodiversity and Conservation. En dicha carta también debe constar, para trabajos donde sea necesaria la manipulación de animales, que los autores disponen de los permisos necesarios y que han cumplido la normativa de protección animal vigente. Los autores pueden enviar también sugerencias para asesores. ISSN: 1578–665X eISSN: 2014–928X

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Las pruebas de imprenta enviadas a los autores deberán remitirse corregidas al Consejo Editor en el plazo máximo de 10 días. Los gastos debidos a modificaciones sustanciales en las pruebas de im prenta, introducidas por los autores, irán a cargo de los mismos. El primer autor recibirá una copia electrónica del trabajo en formato PDF. Manuscritos Los trabajos se presentarán en formato DIN A–4 (30 líneas de 70 espacios cada una) a doble espacio y con las páginas numeradas. Los manuscritos deben estar completos, con tablas y figuras. No enviar las figuras originales hasta que el artículo haya sido aceptado. El texto podrá redactarse en inglés, castellano o catalán. Se sugiere a los autores que envíen sus trabajos en inglés. La revista ofrece, sin cargo ninguno, un servicio de corrección por parte de una persona especializada en revistas científicas. En cualquier caso debe presentarse siempre de forma correcta y con un lenguaje claro y conciso. Los caracteres en cursiva se utilizarán para los nombres científicos de géneros y especies y para los neologismos que no tengan traducción; las citas textuales, independientemente de la lengua en que estén, irán en letra redonda y entre comillas; el nombre del autor que sigue a un taxón se escribirá también en redonda. Se evitará el uso de términos extranjeros (p. ej.: latín, aleman,...). Al citar por primera vez una especie en el trabajo, deberá especificarse siempre que sea posible su nombre común. Los topónimos se escribirán bien en su forma original o bien en la lengua en que esté redactado el trabajo, siguiendo el mismo criterio a lo largo de todo el artículo. Los números del uno al nueve se escribirán con letras, a excepción de cuando precedan una unidad de medida. Los números mayores de nueve se escribirán con cifras excepto al empezar una frase. Las fechas se indicarán de la siguiente forma: 28 VI 99 (un único día); 28, 30 VI 99 (días 28 y 30); 28–30 VI 99 (días 28 al 30). Se evitarán siempre las notas a pie de página. Formato de los artículos Título. Será conciso pero suficientemente explicativo del contenido del trabajo. Los títulos con designaciones de series numéricas (I, II, III, etc.) serán aceptados excepcionalmente previo consentimiento del editor. Nombre del autor o autores Abstract en inglés de 12 líneas mecanografiadas (860 espacios como máximo) y que exprese la esencia del manuscrito (introducción, material, métodos, resultados y discusión). Se evitarán las especulaciones y las citas bibliográficas. Irá encabezado por el título del trabajo en cursiva. Key words en inglés (un máximo de seis) que especifiquen el contenido del trabajo por orden de importancia. © 2018 Museu de Ciències Naturals de Barcelona

Resumen en castellano, traducción del abstract. Su traducción puede ser solicitada a la revista en el caso de autores que no sean castellano hablantes. Palabras clave en castellano. Dirección postal del autor o autores. (Título, Nombre, Abstract, Key words, Resumen, Palabras clave y Dirección postal conformarán la primera página.) Introducción. En ella se dará una idea de los antecedentes del tema tratado, así como de los objetivos del trabajo. Material y métodos. Incluirá la información referente a las especies estudiadas, aparatos utilizados, metodología de estudio y análisis de los datos y zona de estudio. Resultados. En esta sección se presentarán únicamente los datos obtenidos que no hayan sido publicados previamente. Discusión. Se discutirán los resultados y se compararán con otros trabajos relacionados. Las sugerencias sobre investigaciones futuras se podrán incluir al final de este apartado. Agradecimientos (optativo). Referencias. Cada trabajo irá acompañado de una bibliografía que incluirá únicamente las publicaciones citadas en el texto. Las referencias deben presentarse según los modelos siguientes (método Harvard): * Artículos de revista: Conroy, M. J., Noon, B. R., 1996. Mapping of species richness for conservation of biological diversity: conceptual and methodological issues. Ecological Applications, 6: 763–773. * Libros y otras publicaciones no periódicas: Seber, G. A. F., 1982. The estimation of animal abundance. C. Griffin & Company, London. * Trabajos de contribución en libros: Macdonald, D. W., Johnson, D. P., 2001. Dispersal in theory and practice: consequences for conservation biology. In: Dispersal: 358–372 (T. J. Clober, E. Danchin, A. A. Dhondt, J. D. Nichols, Eds.). Oxford University Press, Oxford. * Tesis doctorales: Merilä, J., 1996. Genetic and quantitative trait variation in natural bird populations. Tesis doctoral, Uppsala University. * Los trabajos en prensa sólo se citarán si han sido aceptados para su publicación: Ripoll, M. (in press). The relevance of population studies to conservation biology: a review. Animal Biodiversity and Conservation.

Las referencias se ordenarán alfabéticamente por autores, cronológicamente para un mismo autor y con las letras a, b, c,... para los trabajos de un mismo autor y año. En el texto las referencias bibliográficas se indicarán en la forma usual: "...según Wemmer (1998)...", "...ha sido definido por Robinson y Redford (1991)...", "...las prospecciones realizadas (Begon et al., 1999)...". Tablas. Se numerarán 1, 2, 3, etc. y se reseñarán todas en el texto. Las tablas grandes deben ser más estrechas y largas que anchas y cortas ya que deben ajustarse a la caja de la revista. Figuras. Toda clase de ilustraciones (gráficas, figuras o fotografías) se considerarán figuras, se numerarán 1, 2, 3, etc. y se citarán todas en el texto. Pueden incluirse fotografías si son imprescindibles. Si las fotografías son en color, el coste de su publicación irá a cargo de los autores. El tamaño máximo de las figuras es de 15,5 cm de ancho y 24 cm de alto. Deben evitarse las figuras tridimensionales. Tanto los mapas como los dibujos deben incluir la escala. Los sombreados preferibles son blanco, negro o trama. Deben evitarse los punteados ya que no se reproducen bien. Pies de figura y cabeceras de tabla. Serán claros, concisos y bilingües en castellano e inglés. Los títulos de los apartados generales del artículo (Introducción, Material y métodos, Resultados, Discusión, Agradecimientos y Referencias) no se numerarán. No utilizar más de tres niveles de títulos. Los autores procurarán que sus trabajos originales no excedan las 20 páginas incluidas figuras y tablas. Si en el artículo se describen nuevos taxones, es imprescindible que los tipos estén depositados en alguna institución pública. Se recomienda a los autores la consulta de fascículos recientes de la revista para seguir sus directrices. Comunicaciones breves Las comunicaciones breves seguirán el mismo procedimiento que los artículos y serán sometidas al mismo proceso de revisión. No excederán las 2.300 palabras, incluidos título, resumen, cabeceras de tabla, pies de figura, agradecimientos y referencias. El resumen no debe sobrepasar las 100 palabras y el número de referencias será de 15 como máximo. Que el texto tenga apartados es opcional y el número de tablas y/o figuras admitidas será de dos de cada como máximo. En cualquier caso, el trabajo maquetado no podrá exceder las cuatro páginas.

Animal Biodiversity and Conservation 41.2 (2018)

Animal Biodiversity and Conservation

Manuscripts

Animal Biodiversity and Conservation is an interdisciplinary journal published by the Natural Science Museum of Barcelona since 1958. It includes empirical and theoretical research from around the world that examines any aspect of Zoology (Systematics, Taxonomy, Morphology, Biogeography, Ecology, Ethology, Physiology and Genetics). Special emphasis is given to integrative and multidisciplinary studies that help to understand the evolutionary patterns in biodiversity in the widest sense. The journal does not publish bibliographic compilations, listings, catalogues or collections of species, or isolated descriptions of a single specimen. Studies concerning rare or protected species will not be accepted unless the authors have been granted the relevant permits or authorisation. Each annual volume consists of two issues. Animal Biodiversity and Conservation is registered in all principal data bases and is freely available online at www.abc.museucienciesjournals.cat assuring world–wide access to articles published therein. All manuscripts are screened by the Executive Editor, an Editor and two independent reviewers so as to guarantee the quality of the papers. The review process aims to be rapid and constructive. Once accepted, papers are published as soon as is practicable. This is usually within 12 months of initial submission. Upon acceptance, manuscripts become the property of the journal, which reserves copyright, and no published material may be reproduced or cited without acknowledging the source of information. All rights are reserved by the authors, who authorise the journal to publish the article. Papers are published under a Creative Commons Attribution 4.0 International License: no part of the published paper may be reproduced or reused unless the source is cited.

Manuscripts must be presented in DIN A–4 format, 30 lines, 70 keystrokes per page. Maintain double spacing throughout. Number all pages. Manuscripts should be complete with figures and tables. Do not send original figures until the paper has been accepted. The text may be written in English, Spanish or Catalan, though English is preferred. The journal provides linguistic revision by an author’s editor. Care must be taken to use correct wording and the text should be written concisely and clearly. Scientific names of genera and species as well as untranslatable neologisms must be in italics. Quotations in whatever language used must be typed in ordinary print between quotation marks. The name of the author following a taxon should also be written in lower case letters. Foreing terms (e.g. Latin, German,...) should not be used. When referring to a species for the first time in the text, both common and scientific names should be given when possible. Do not capitalize common names of species unless they are proper nouns (e.g. Iberian rock lizard). Place names may appear either in their original form or in the language of the manuscript, but care should be taken to use the same criteria throughout the text. Numbers one to nine should be written in full within the text except when preceding a measure. Higher numbers should be written in numerals except at the beginning of a sentence. Specify dates as follows: 28 VI 99 (for a single day); 28, 30 VI 99 (referring to two days, e.g. 28th and 30th), 28–30 VI 99 (for more than two consecutive days, e.g. 28th to 30th). Footnotes should not be used.

Information for authors

Formatting of articles

Electronic submission of papers is encouraged (abc@bcn.cat). The preferred format is DOC or RTF. All figures must be readable by Word, embedded at the end of the manuscript and submitted together in a separate attachment in a TIFF, EPS or JPEG file. Tables should be placed at the end of the document. A cover letter stating that the article reports original research that has not been published elsewhere and has been submitted exclusively for consideration in Animal Biodiversity and Conservation is also necessary. When animal manipulation has been necessary, the cover letter should also specify that the authors follow current norms on the protection of animal species and that they have obtained all relevant permits and authorisations. Authors may suggest referees for their papers. Proofs sent to the authors for correction should be returned to the Editorial Board within 10 days. Expenses due to any substantial alterations of the proofs will be charged to the authors. The first author will receive electronic version of the article in PDF format.

Title. Must be concise but as informative as possible. Numbering of parts (I, II, III, etc.) should be avoided and will be subject to the Editor’s consent. Name of author or authors Abstract in English, no longer than 12 typewritten lines (840 spaces), covering the contents of the article (introduction, material, methods, results and discussion). Speculation and literature citation should be avoided. The abstract should begin with the title in italics. Key words in English (no more than six) should express the precise contents of the manuscript in order of relevance. Resumen in Spanish, translation of the Abstract. Summaries of articles by non–Spanish speaking authors will be translated by the journal on request. Palabras clave in Spanish. Address of the author or authors.

ISSN: 1578–665X eISSN: 2014–928X

Introduction. Should include the historical background of the subject as well as the aims of the paper. Material and methods. This section should provide relevant information on the species studied, materials, methods for collecting and analysing data, and the study area. Results. Report only previously unpublished results from the present study. Discussion. The results and their comparison with related studies should be discussed. Suggestions for future research may be given at the end of this section. Acknowledgements (optional). References. All manuscripts must include a bibliography of the publications cited in the text. References should be presented as in the following examples (Harvard method): * Journal articles: Conroy, M. J., Noon, B. R., 1996. Mapping of species richness for conservation of biological diversity: conceptual and methodological issues. Ecological Applications, 6: 763–773. * Books or other non–periodical publications: Seber, G. A. F., 1982. The estimation of animal abundance. C. Griffin & Company, London. * Contributions or chapters of books: Macdonald, D. W., Johnson, D. P., 2001. Dispersal in theory and practice: consequences for conservation biology. In: Dispersal: 358–372 (T. J. Clober, E. Danchin, A. A. Dhondt, J. D. Nichols, Eds.). Oxford University Press, Oxford. * PhD thesis: Merilä, J., 1996. Genetic and quantitative trait variation in natural bird populations. PhD thesis, Uppsala University. * Works in press should only be cited if they have been accepted for publication: Ripoll, M. (in press). The relevance of population studies to conservation biology: a review. Animal Biodiversity and Conservation. References must be set out in alphabetical and chronological order for each author, adding the letters a, b, c,... to papers of the same year. Bibliographic citations in the text must appear in the usual way: "...according to Wemmer (1998)...", "...has been defined by Robinson and Redford (1991)...", "...the prospections that have

been carried out (Begon et al., 1999)..." Tables. Must be numbered in Arabic numerals with reference in the text. Large tables should be narrow (across the page) and long (down the page) rather than wide and short, so that they can be fitted into the column width of the journal. Figures. All illustrations (graphs, drawings, photographs) should be termed as figures, and numbered consecutively in Arabic numerals (1, 2, 3, etc.) with reference in the text. Glossy print photographs, if essential, may be included. The Journal will publish colour photographs but the author will be charged for the cost. Figures have a maximum size of 15.5 cm wide by 24 cm long. Figures should not be tridimensional. Any maps or drawings should include a scale. Shadings should be kept to a minimum and preferably with black, white or bold hatching. Stippling should be avoided as it may be lost in reproduction. Legends of tables and figures. Legends of tables and figures should be clear, concise, and written both in English and Spanish. Main headings (Introduction, Material and methods, Results, Discussion, Acknowledgements and References) should not be numbered. Do not use more than three levels of headings. Manuscripts should not exceed 20 pages including figures and tables. If the article describes new taxa, type material must be deposited in a public institution. Authors are advised to consult recent issues of the journal and follow its conventions. Brief communications Brief communications should follow the same procedure as other articles and they will undergo the same review process. They should not exceed 2,300 words including title, abstract, figure and table legends, acknowledgements and references. The abstract should not exceed 100 words, and the number of references should be limited to 15. Section headings within the text are optional. Brief communications may have up to two figures and/or two tables but the whole paper should not exceed four published pages.

Animal Biodiversity and Conservation 41.2 (2018)

VII

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Animal Biodiversity and Conservation 41.1 (2018)

Animal Biodiversity and Conservation 41.2 (2018)

Arxius de Miscel·lània Zoològica vol. 15 (2017) Museu de Ciències Naturals de Barcelona ISSN: 1698–0476 www.amz.museucienciesjournals.cat

Índex / Índice / Contents López–Soriano, J., Quiñonero–Salgado, S., Cadevall, J., 2017. Presència del bivalve invasor Sinanodonta woodiana (Lea, 1834) al delta del Llobregat (Baix Llobregat). Arxius de Miscel·lània Zoològica, 15: 1–7. Abstract Presence of the invasive bivalve Sinanodonta woodiana (Lea, 1834) in the Llobregat delta (Baix Llobregat).— The Llobregat delta is among the most important wetlands in Catalonia. The diverse malacofauna of the area is threatened, however, by the large infrastructures surrounding the delta and the area´s close proximity to the urban centre of Barcelona. Here we present the first report on invasive bivalves in the delta following findings of a well–established population of the Chinese pond mussel, Sinanodonta woodiana. The distribution of the species in this natural area is also discussed. Key words: Invasions, Bivalves, Llobregat delta Fontana–Bria, L., Frago, E., Prieto–Lillo, E., Selfa, J., 2017. Biogeographic evaluation of the dragonflies and damselflies in the Eastern Iberian Peninsula. Arxius de Miscel·lània Zoològica, 15: 8–29. Abstract Biogeographic evaluation of the dragonflies and damselflies in the Eastern Iberian Peninsula.— Insects are one of the most diverse groups of animals in terrestrial ecosystems, and are thus a good model system to study macrogeographic patterns in species’ distributions. Here we perform a biogeographical analysis of the dragonflies and damselflies in the Valencian Country (Eastern Iberian Peninsula). We also compare the species present in this territory with those in the adjacent territories of Catalonia and Aragon, and with those present in the whole Iberian Peninsula. Furthermore, we update the list of species of dragonflies and damselflies in the Valencian territory (65 species), and discuss the current status of two of them: Macromia splendens and Lindenia tetraphylla. Our results highlight that the Valencian Country has a higher proportion of Ethiopian elements but a lower proportion of Eurosiberian elements than Catalonia and Aragon. We also emphasize the importance of volunteer work in providing new knowledge on this group of iconic insects, and the relevance of museum collections in preserving them. The role of climate change in the distribution of Odonata is also discussed. Key words: Odonata, Valencian Country, Spain, Iberian Peninsula, Biogeography, Climate change Viñolas, A., Caballero–López, B., Masó, G., 2017. The collection of type specimens belonging to the subfamily Pimeliinae (Coleoptera, Tenebrionidae) in the Natural Sciences Museum of Barcelona, Spain. Arxius de Misceŀlània Zoològica, 15: 30–92. Abstract The collection of type specimens belonging to the subfamily Pimeliinae (Coleoptera, Tenebrionidae) in the Natural Sciences Museum of Barcelona, Spain.— The type collection of the subfamily Pimeliinae (Coleoptera, Tenebrionidae) deposited in the Natural Sciences Museum of Barcelona, Spain, was organised, revised and documented. The collection contains 438 type specimens representing 140 different taxa. Of note is a considerable number of species belonging to a subfamily described by Francesc Español, Maurice Antoine, Zoltán Kaszab and Carlo Koch. In this paper we provide all the available information relating to these type specimens and for all taxa (species or subspecies) we give the following information: original and current taxonomic status, original citation of type material, the exact transcription of the original label, and the preservation condition of the specimen. We also discuss the differences between the original descriptions and labels. If a taxonomic change has occurred, the references describing those changes are included at the end of the description. Key words: Collection type, Coleoptera, Tenebrionidae, Pimeliinae subfamily, Taxonomic revision ISSN: 1578–665 X eISSN: 2014–928 X

Animal Biodiversity and Conservation 41.2 (2018)

Macià, R., Caballero–López, B., Masó, G., 2017. The collection of Lepidoptera type specimens deposited in the Natural Sciences Museum of Barcelona (MCNB), Spain. Arxius de Miscel·lània Zoològica, 15: 93–206. Abstract The collection of Lepidoptera type specimens deposited in the of Natural Sciences Museum of Barcelona (MCNB), Spain.— The Lepidoptera type specimens in the collection of the Natural Sciences Museum of Barcelonawere reorganized and catalogued. The revised collection –the vast majority of which belonged to Ignasi de Sagarra– now consists of 1,128 type specimens, corresponding to 168 taxa. This study provides all the available information relating to each of the revised specimens, including the current taxonomic status, the original name if different, the original reference for the type material, the exact transcription of the original label, and the conservation status of the specimen. Differences between the original description of species and their labels, as well as any taxonomical changes that have occurred, are also discussed. The corresponding bibliography is included in the references. If a taxonomic change has occurred since the description of taxa, the references discussing such changes are given. Key words: Collection type, Lepidoptera, Catalogued and reorganized, Sagarra collection Gómez–Hoyos, D. A., Ríos–Franco, C. A., Vanegas–Guerrero, J., Velasco, J. A., González–Maya, J. F., 2017. Estado y prioridades de conservación de los anfibios del departamento del Quindío, Colombia. Arxius de Miscel·lània Zoològica, 15: 207–223. Abstract Conservation status and priorities of amphibians from the Quindio Department, Colombia.— We reviewed the conservation status and priorities for amphibians from the Quindio region of Colombia, with the purpose of proposing conservation actions. We modeled the potential distribution of threatened species using the maximum entropy algorithm in MaxEnt and evaluated representability in the Departmental System of Protected Areas for Quindio (Spanish acronym: SIDAPQ). We prioritized areas for amphibian conservation using the algorithm ILV4 adjacency in ConsNet. We recorded 45 species, 24.4% of which are included in threatened categories on the IUCN Red List. Over 50% of the distribution and records of the threatened amphibians occurred inside the SIDAPQ. Prioritized areas to achieve representation goals of 10%, 20% and 30% of amphibian distribution are highly fragmented and have only approximately 30% of prioritized distribution in the SIDAPQ. Considering this scenario we propose a conservation strategy on the landscape level that includes agroecosystems, maintaining their heterogeneity and eliminating or mitigating threat factors. Key words: Amphibians, Central Andes, Colombia, Protected Areas Vergilov, V., N. Natchev, N., 2017. First record of tail bifurcations in the snake–eyed skink Ablepharus kitaibelii Bibron & Bory de Saint–Vincent, 1833 from Pastrina hill (northwestern Bulgaria). Arxius de Miscel·lània Zoològica, 15: 224–228. Abstract First record of tail bifurcations in the snake–eyed skink Ablepharus kitaibelii Bibron & Bory de Saint–Vincent, 1833 from Pastrina hill (northwestern Bulgaria).— We report for the first time on the occurrence of tail bifurcations in the snake–eyed skink (A. kitaibelii). This morphological anomaly was identified during a four–year monitoring program conducted in a herpetological hot–spot at Pastrina hill (northwestern Bulgaria). From a total of 415 captured specimens, four animals (0.96%) showed symmetrical or asymmetrical lateral duplication of the tail. Only bifurcations of the distal–most caudal section were detected in contrast to some other lizards (e.g. Gekkonidae, Lacertidae, Teiidae) that are reported to survive with bifurcations at more proximal tail sections. Key words: Lizard, Autotomy, Regeneration GBIF Agudelo Martínez, J. C., Pérez–Buitrago, N., 2017. New records of hunting ants (Poneroids and Ectatomminoids) in the northern part of the Colombian Orinoquia region. Arxius de Miscel·lània Zoològica, 15: 229–248. Abstract New records of hunting ants (Poneroids and Ectatomminoids) in the northern part of the Colombian Orinoquia region.— We reviewed 466 specimens of hunting ant species collected in flooded savanna environments and their adjacent forest fragments in the rural area of Arauca municipality (Arauca, Colombia). Samples were taken from eight forest fragments with sizes between 0.25 and 220 ha. In each location we set linear transects with seven sampling points separated by 20 m. In each sampling point we used three capture methods: a pitfall trap (left 24 hours), a sample of 1 m2 of soil and litter to be processed with a mini Winkler extractor, and direct capture. Fifteen species were recorded; the most diverse genus was Neoponera with six species, followed by Odontomachus with two species and Ectatomma, Anochetus, Gnamptogenys, Prionopelta Pseudoponera, ISSN: 1578–665 X eISSN: 2014–928 X

Animal Biodiversity and Conservation 41.2 (2018)

Pachycondyla and Thaumatomyrmex each with one species. This study extends the geographical distribution of 15 hunting ant species to the northern part of the Colombian Orinoquia region and documents for the first time the presence of Thaumatomyrmex cf. mutilatus in Colombia. Data published through GBIF and SiB Colombia (doi: 10.15472/vp8v42). Key words: Biodiversity, Ponerinae, Orinoco, Biogeography GBIG: Datapaper Martínez–Ortí, A., 2017. Las colecciones del Museu Valencià d’Història Natural. Arxius de Miscel·lània Zoològica, 15: 249–252. Abstract The collections of the Museu Valencià d’Història Natural.— The records of database collections of the Museu Valencia d’Història Natural are published, with a clear dominance of mollusks, arthropods and chordates. Data published through GBIF (doi:10.15470/8oedep). Key words: Biodiversity information, Mediterranean region GBIF: Datapaper Belda, A., Belenguer, R., Zaragozí, B., Ferri, V., 2017. Presència del gat domèstic, Felis silvestris catus (Schreber, 1775), i del gat serval, Felis silvestris, en un espai natural protegit: el cas del Parc Natural Serra de Mariola (sud–est espanyol). Arxius de Miscel·lània Zoològica, 15: 253–263. Abstract Presence of domestic cat, Felis silvestris catus (Schreber, 1775), and wild cat, Felis silvestris, in a natural protected area: the case of Serra de Mariola Natural Park (SE Spain).— The main objective of this research was to determine the ecological aspects, distribution, and impact of domestic cat in the Serra de Mariola. The study area is a natural park of 17,500 hectares located in the south of the Valencia Community. Better knowledge of the distribution of this cat will help define management measures in the park. Using camera traps, we collected 29,941 images of animal contact. A total of 0.62 % of these photographs were of domestic cat, the presence of which was detected in 29 of 63 grids (2 x 2 km) in Serra de Mariola Natural Park (46.03 %). Sampling was performed from January to September 2010. The study allowed us to integrate the information collected in the field into databases and to evaluate the presence of domestic cat in the Serra de Mariola. Data published through GBIF (doi:10.15470/p7evig). Key words: Distribution, Camera trap, Domestic cat, Landscape, Serra de Mariola Natural Park Martínez–Ortí, A., Arco, M. C., López–Alabau, A., 2017. Hallazgo de una población de Quickella arenaria (Potiez & Michaud, 1835) (Mollusca, Gastropoda, Succineidae) en Castilla–La Mancha (España). Arxius de Miscel·lània Zoològica, 15: 264–268. Abstract On the finding of a population of Quickella arenaria (Potiez & Michaud, 1835) (Mollusca, Gastropoda, Succineidae) in Castilla–La Mancha (Spain).— After studying many springs in the province of Cuenca (Spain) we report the finding of Quickella arenaria, a little known succinid gastropod in Spain, for the first time in Castilla–La Mancha. We discuss the threats that the population faces in this habitat and propose conservation management strategies. Key words: Molluscs, Succineidae, Quickella arenaria, Conservation, Cuenca, Spain

ISSN: 1578–665 X eISSN: 2014–928 X

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321–332 Jugovic, J., Koprivnikar, N., Koren, T. The role of semi–natural grasslands and livestock in sustaining dung beetle communities (Coleoptera, Scarabaeoidea) in sub–Mediterranean areas of Slovenia 333–344 Mingarro, M., Lobo, J. M. Environmental representativeness and the role of emitter and recipient areas in the future trajectory of a protected area under climate change 345–356 Freret–Meurer, N., Fernández, T., Okada, N. Vaccani, A. Population dynamics of the endangered seahorse Hippocampus reidi Ginsburg, 1933 in a tropical rocky reef habitat 357–364 Llanes–Quevedo, A., Alfonso González, M., Cárdenas Mena, R., Frankel, C., Espinosa Lopez, G. Microsatellite variability of the wood stork Mycteria americana (Aves, Ciconidae) in Cuba: implications for its conservation 365–377 Ferrero–García, J. J. Torres–Vila, L. M. Bueno, P. P. The mysterious bird outbreak of 1779 in southeastern Iberian peninsula: a massive irruption of the Spanish sparrow Passer hispaniolensis from Africa?

379–388 Vidal–Vallés, D., Rodríguez, A., Pérez–Collazos, E. Bird roadkill occurences in Aragon, Spain 389–404 Arguello–Sánchez, L. E., Arguello, J. R., García–Feria, L. M., García–Sepúlveda, C. A., Santiago–Alarcon, D., Espinosa de los Monteros, A. MHC class II DRB variability in wild black howler monkeys (Alouatta pigra), an endangered New World primate 405–413 Biaggini, M., Campetti, I., Corti, C. Breeding activity of the agile frog Rana dalmatina in a rural area 415–425 Horcajada, F., Alcaraz, L., Barja, I., Machordom, A. Phylogeographic patterns of Capreolus capreolus in the centre of the Iberian peninsula 427–434 Pagani–Núñez, E., Renom, M., Furquet, C., Rodríguez, J., Llimona, F., Senar, J. C. Isotopic niche overlap between the invasive leiothrix and potential native competitors

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Índex / Índice / Contents Animal Biodiversity and Conservation 41.2 (2018) ISSN 1578–665 X eISSN 2014–928 X 195–208 Huang, Q., Liu, X., Li, Y., Kraus, J., Songer, M. Understanding nutrient landscapes for giant pandas in the Qinling Mountains, China: the relationships between bamboo mineral content and giant panda habitat selection during migration 209–215 Martín, B., Pérez, H., Ferrer, M. Effects of natural and artificial light on the nocturnal behaviour of the wall gecko 217–225 Garrote, G., Fernández–López, J., López, G., Ruiz, G., Simón, M. A. Prediction of Iberian lynx road–mortality in southern Spain: a new approach using the MaxEnt algorithm 227–243 Carlisle, J. D., Sanders, L. E., Chalfoun, A. D., Gerow, K. G. Raptor nest–site use in relation to the proximity of coalbed–methane development 245–256 Trujillo–Caballero, S., González–Oreja, J. A. Efficient vs. structured biodiversity inventories: reptiles in a Mexican dry scrubland as a case study 257–266 Ávila–Nájera, D. M., Palomares, F., Chávez, C., Tigar, B., Mendoza, G. D. Jaguar (Panthera onca) and puma (Puma concolor) diets in Quintana Roo, Mexico

267–274 Taybi, A. F., Mabrouki, Y., Berrahou, A., El Abd, A. A. Bio–ecology of Potamon algeriense (Herbst, 1785) (Crustacea, Decapoda) in eastern Morocco 275–280 Zieliński, D., Schwarz, C. J., Ehrmann, R. Evaluation of the expansion of Mantis religiosa (L.) in Poland based on a questionnaire survey 281–300 Canal, D., Camacho, C., Martín, B., de Lucas, M., Ferrer, M. Magnitude, composition and spatiotemporal patterns of vertebrate roadkill at regional scales: a study in southern Spain 301–310 Britto, V. O., Gil–Delgado, J. A., Gosálvez, R. U., López–Iborra, G. M., Velasco, A. Foraging habitat selection by gull–billed tern (Gelochelidon nilotica) in Central Spain (Castilla–La Mancha) 311–314 Gatti, A., Seibert, J. B., Moreira, D. O. A predation event by free–ranging dogs on the lowland tapir in the Brazilian Atlantic Forest 315–319 Mannise, N., Trovati, R. G., Duarte, J. M. B., Maldonado, J. E., González, S. Using non–invasive genetic techniques to assist in maned wolf conservation in a remnant fragment of the Brazilian Cerrado

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