Animal Biodiversity and Conservation issue 37.1 (2014) by Museu Ciències Naturals de Barcelona MCNB

Dibuix de la coberta: Anax imperator, libèl·lula emperador, libélula emperador, Emperor dragonfly (Jordi Domènech). Editor en cap / Editor responsable / Editor in Chief Joan Carles Senar 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 – SEE) Biologia Evolutiva / Biología Evolutiva / Evolutionary Biology: Santiago Merino (Sociedad Española de Biología Evolutiva – SESBE) Editors / Editores / Editors Pere Abelló Inst. de Ciències del Mar CMIMA–CSIC, Barcelona, Spain Javier Alba–Tercedor Univ. de Granada, Granada, Spain Russell Alpizar–Jara Univ. of Évora, Évora, Portugal Xavier Bellés Inst. de Biología Evolutiva UPF–CSIC, Barcelona, Spain Luís Mª Carrascal Museo Nacional de Ciencias Naturales–CSIC, Madrid, Spain Michael J. Conroy Univ. of Georgia, Athens, USA Adolfo Cordero Univ. de Vigo, Vigo, Spain Mario Díaz Museo Nacional de Ciencias Naturales–CSIC, Madrid, Spain José A. Donazar Estación Biológica de Doñana–CSIC, Sevilla, Spain Jordi Figuerola Estación Biológica de Doñana–CSIC, Sevilla, Spain Gary D. Grossman Univ. of Georgia, Athens, USA Damià Jaume IMEDEA–CSIC, Univ. de les Illes Balears, Spain Jordi Lleonart Inst. de Ciències del Mar CMIMA–CSIC, Barcelona, Spain Jorge M. Lobo Museo Nacional de Ciencias Naturales–CSIC, Madrid, Spain Pablo J. López–González Univ. de Sevilla, Sevilla, Spain Jose Martin Museo Nacional de Ciencias Naturales–CSIC, Madrid, Spain Santiago Merino Museo Nacional de Ciencias Naturales–CSIC, Madrid Juan J. Negro Estación Biológica de Doñana–CSIC, Sevilla, Spain Vicente M. Ortuño Univ. de Alcalá de Henares, Alcalá de Henares, Spain Miquel Palmer IMEDEA–CSIC, Univ. de les Illes Balears, Spain Javier Perez–Barberia The James Hutton Institute, Scotland, United Kingdom Oscar Ramírez Inst. de Biologia Evolutiva UPF–CSIC, Barcelona, Spain Montserrat Ramón Inst. de Ciències del Mar CMIMA–CSIC, Barcelona, Spain Ignacio Ribera Inst. de Biología Evolutiva UPF–CSIC, Barcelona, Spain Alfredo Salvador Museo Nacional de Ciencias Naturales–CSIC, Madrid, Spain José L. Tellería Univ. Complutense de Madrid, Madrid, Spain Francesc Uribe Museu de Ciències Naturals de Barcelona, Barcelona, Spain Carles Vilà Estación Biológica de Doñana–CSIC, Sevilla, Spain Rafael Villafuerte Inst. de Estudios Sociales Avanzados (IESA–CSIC), Cordoba, Spain Rafael Zardoya Museo Nacional de Ciencias Naturales–CSIC, Madrid, Spain Secretària de Redacció / Secretaria de Redacción / Managing Editor Montserrat Ferrer Assistència Tècnica / Asistencia Técnica / Technical Assistance Eulàlia Garcia Anna Omedes Francesc Uribe

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Animal Biodiversity and Conservation 37.1, 2014 © 2014 Museu de Ciències Naturals de Barcelona, Consorci format per l'Ajuntament de Barcelona i la Generalitat de Catalunya Autoedició: Montserrat Ferrer Fotomecànica i impressió: XXXXXXXXX ISSN: 1578–665 X eISSN: 2014–928 X Dipòsit legal: B. 5357–2013 Animal Biodiversity and Conservation es publica amb el suport de: l'Asociación Española de Ecología Terrestre, la Sociedad Española de Etología i la Sociedad Española de Biología Evolutiva The journal is freely available online at: www.abc.museucienciesjournals.cat

Animal Biodiversity and Conservation 37.1 (2014)

Diversity of large and medium mammals in Juchitan, Isthmus of Tehuantepec, Oaxaca, Mexico M. Cortés–Marcial, Y. M. Martínez Ayón & M. Briones–Salas

Cortés–Marcial, M., Martínez Ayón, Y. M. & Briones–Salas, M., 2014. Diversity of large and medium mammals in Juchitan, Isthmus of Tehuantepec, Oaxaca, Mexico. Animal Biodiversity and Conservation, 37.1: 1–12. Abstract Diversity of large and medium mammals in Juchitan, Isthmus of Tehuantepec, Oaxaca, Mexico.— The Isthmus of Tehuantepec in Oaxaca, Mexico, is one of the country’s most important regions from a zoogeographical perspective due to the large number of endemic Neotropical species found there. Between September 2007 and August 2008, we sampled medium–sized and large mammals in the Juchitan municipality and compared their diversity in two areas with distinct levels of anthropogenic impact, defined according to estimates of human activities, livestock density and habitat degradation, We obtained 167 records of 18 species, with a 79% representation according to species accumulation models in both areas. The highest species richness and alpha diversity were recorded in the preserved area, whereas the disturbed area exhibited half the diversity found in the preserved area. A high interchange of species was also observed between zones. The two species with the largest number of records were Urocyon cinereoargenteus (n = 52) and Didelphis virginiana (n = 42). In both areas, the highest relative abundance occurred during the rainy season. Habitat degradation and human activities seem to affect the diversity of mammal species in the region. Key words: Biodiversity, Conservation, Disturbance, Isthmus of Tehuantepec, Tropical deciduous forest. Resumen La diversidad de los mamíferos de talla grande y mediana en Juchitán, istmo de Tehuantepec, Oaxaca, México.— El istmo de Tehuantepec en Oaxaca, México, es una de las regiones más importantes del país desde el punto de vista zoogeográfico, ya que alberga una gran cantidad de especies endémicas neotropicales. Entre septiembre de 2007 y agosto de 2008, se realizó un muestreo de mamíferos de talla mediana y grande en el municipio de Juchitán, y comparamos su diversidad en dos zonas con distintos niveles de impacto antropogénico definido de acuerdo con las estimaciones de las actividades humanas, la densidad de ganado y la degradación del hábitat. Se obtuvieron 167 registros de 18 especies, con una representatividad del 79% según el modelo de acumulación de especies en ambas zonas. La mayor riqueza de especies y de diversidad alfa se registraron en la zona conservada, mientras que la zona perturbada presenta la mitad de la diversidad encontrada en la zona conservada. Se observó un fuerte intercambio de especies entre ambas zonas. Dos especies, Urocyon cinereoargenteus (n = 52) y Didelphis virginiana (n = 42), tuvieron el mayor número de registros. En ambas zonas, la mayor abundancia relativa se observó durante la época de lluvias. La degradación del hábitat y las actividades humanas al parecer afectan a la diversidad de especies de mamíferos en la región. Palabras clave: Biodiversidad, Conservación, Perturbación, Istmo de Tehuantepec, Bosque deciduo tropical. Received: 14 VI 13; Conditional acceptance: 14 X 13; Final acceptance: 8 I 14 Malinalli Cortés–Marcial, Yazmín del Mar Martínez Ayón & Miguel Briones–Salas, Lab. de Vertebrados Terrestres (Mastozoología), Centro Interdisciplinario de Investigación para el Desarrollo Integral Regional (CIIDIR–Unidad Oaxaca), Inst. Politécnico Nacional, Oaxaca, México. Corresponding author: M. Cortés–Marcial. E–mail: mali_cor@yahoo.com.mx ISSN: 1578–665 X eISSN: 2014–928 X

Introduction One of the issues of greatest interest in ecology is the relationship between habitat structure and the structure of animal communities. Habitat disturbance and habitat fragmentation influence both the original plant communities and the heterogeneity and complexity of the entire ecosystem. This, in turn, influences the availability of resources, and affects the birth and death rates of several species, thus affecting vertebrate diversity (August, 1983; Soule et al., 1992; Collins et al., 1995; Murcia, 1995; Zarza, 2001). Large and medium–sized mammals are particularly sensitive to habitat changes, and they are common victims of poaching and illegal trading (Michalski & Peres, 2005; Laurance et al., 2006). The functional significance of these species lies in their ecological roles, such as seed dispersal and predation on numerous plant species. These functional roles may change the structure and composition of the ecosystem. Moreover, these species influence the community structure and complexity on the trophic levels in which they are involved, due to their regulatory role as preys and predators (Roemer et al., 2009). The loss of these organisms could have devastating effects because they contribute in many ways to the functioning of the natural ecosystem (Alonso et al., 2001; Bolaños & Naranjo, 2001). Given the importance of these species, studies identifying and predicting the environmental changes that may affect their diversity are essential, and in such studies, relative abundance and species diversity are usually used as indicators (Carrillo et al., 2000). The Isthmus of Tehuantepec (Mexico) is one of most diverse regions within this country (Briones–Salas & Sánchez–Cordero, 2004; González et al., 2004). Furthermore, this area has a particular importance from a zoogeographical perspective because it lies in the zone where the Nearctic and Neotropical regions overlap. This important corridor between the Atlantic and costal Pacific plains represents a significant barrier for highland mammal species, and also favors a high degree of endemicity (Peterson et al., 1999; García–Trejo & Navarro, 2004; Barragan et al., 2010). However, this diversity may be declining dramatically, due to hunting and habitat modification derived from crops and livestock. Therefore, the aim of this study was to identify the differences in diversity, in terms of abundance and heterogeneity, of medium–sized and large mammals in two areas with differing degrees of anthropogenic disturbance. If anthropogenic environmental changes affect mammal communities, we hypothesized that the area with greater human disturbance would exhibit a lower diversity of medium and large mammals. Methods The study area is located in the coastal plain of Tehuantepec, northeast of the city of Juchitan, Oaxaca, Mexico, at 200 m a.s.l., within the coordinates 94° 55' to 94° 50' W, and 16° 38' to 16° 30' N (fig. 1). The climate is sub–humid and warm. There is a marked dry season from December to May, and a rainy season from June to

Cortés–Marcial et al.

November, with an average annual rainfall of 932.2 mm. The annual average temperature is 27.6°C (Garcia, 1988). The first sampling area was located on the hill of Tolistoque, northeast of Juchitan (16° 35' 5.91'' N, 94° 52' 20.63'' W) within an area —protected by the regional indigenous communities— known as Ojo de Agua Tolistoque Protected Communal Area (Ortega et al., 2010). The vegetation is tropical deciduous forest. The second sampling area was south of the Protected Communal Area northeast of Juchitan (16° 32' 12.95'' N, 94° 50' 53.95'' W), in an area of secondary vegetation. This area is dedicated to farming activities, with gallery forest areas around irrigation canals, and tropical deciduous forest remnants (fig. 1). We applied an indirect sampling method. Such methods are sometimes the only option available to study the distribution and abundance of inaccessible vertebrates such as medium–sized and large mammals (Sutherland, 1996). These methods also have some advantages over direct methods as they are easier to implement and independent of the time of day, which is important when target species are nocturnal, cryptic and difficult to capture or recapture because their traces remain for long periods of time (Bilenca et al., 1999; Simoneti & Huareco, 1999; Aranda, 2000; Carrillo et al., 2000; Ojasti, 2000). In both areas the level of disturbance was evaluated according to the index proposed by Peters & Martorell (2000) and Martorell & Peters (2005). In order to measure the contribution of different agents, we recorded 14 metrics at each site by means of two 50 m long transects at each site (table 1). Disturbance was measured on a scale of 0–100, where zero is the least disturbance. The values were calculated as follows: Disturbance = 3.41 Goat – 1.37 Catt + + 27.62 Brow + 49.20 Ltra – 1.03 Comp + + 41.01 Fuel + 0.12 Tran + 24.17 Prox + + 8.98 Core + 8.98 Luse – 0.49 Fire + + 26.94 Eros + 17.97 Isla + 26.97 Toms + 0.2 The medium and large mammals were classified using the system of Robinson & Redford (1986), who divided mammals into four categories based on a logarithmic scale of average weight: small < 100 g; medium > 100~ < 1,000 g; large > 1,000 g < 10,000 g; very large > l0,000 g. To search for traces of medium and large mammals, monthly samples were taken from September 2007 to August 2008. During each period, four transects (two in each zone) of 4.5 km each were sampled, resulting in a total sampling of 108 km walked in each zone. We used a Mexican mammal field guide (Aranda, 2000) to identify tracks and feces, and compared these with the reference material on traces of mammals of Oaxaca, of the Collection of Mammalogy (OAX.MA.026.0497) at the Centro Interdisciplinario de Investigación para el Desarrollo Integral Regional (CIIDIR–Oaxaca), National Polytechnic Institute (IPN). Ten camera traps (Cuddeback Expert ®) were also used for the last six sampling periods to confirm the presence of the species (five in each zone). These were placed at approximately 1.5 km from each other.

Animal Biodiversity and Conservation 37.1 (2014)

Mexico

Oaxaca Study site

94º 50' 0'' W

94º 45' 0'' W

Santo Domingo Ingenio

16º 35' 0'' N

Preserved area

La Venta

Disturbed area 16º 30' 0'' N

8 km

94º 55' 0'' W

Union Hidalgo

Tropical deciduous forest

Tropical deciduous forest with secundary forest

Tropical deciduous thorn

Halophytic vegetation and gypsophila

Agricultural management

Induced grassland

Community protected area

Highway

Municipality

Urban area

94º 50' 0'' W

16º 30' 0'' N

Vegetation type

La Ventosa

16º 35' 0'' N

94º 55' 0'' W

94º 45' 0'' W

Fig. 1. Geographic location and vegetation types around the study area in the Isthmus of Tehuantepec, Oaxaca, Mexico. Fig. 1. Ubicación geográfica y tipos de vegetación en el entorno de la zona de estudio en el istmo de Tehuantepec, Oaxaca, México.

Each camera trap was installed approximately 40–50 cm above ground level, depending on the topography and slope of the sampling area. The camera circuit was programmed to remain active for 24 hours, and the camera locations were geo–referenced with a GPS (Garmin Etrex®). Cameras were checked monthly. Photographic records were prepared according to Botello et al. (2007) and deposited in the Collection of Mammalogy (OAX. MA.026.0497) of CIIDIR–OAX. Data analysis Species inventories were evaluated using Clench’s asymptotic models of species accumulation with the program Species Accumulation, for which the data were previously randomized 100 times with the EstimateS program, version 8.0 (Colwell, 2000). We also calculated the sampling effort required to include 95% of the species in the inventories.

The relative species abundance index for each area and season (dry and rainy) was calculated as the total number of signs found per species, divided by the distance sampled (Carrillo et al., 2000). A Mann–Whitney U test was applied to determine whether there were significant differences in relative abundance between areas and seasons (Zar, 1999). The species diversity of each area and season was determined according to the Shannon–Wiener entropy index (H’). Dominance (D) was estimated with the Berger–Parker index (Whittaker, 1972), which is an indirect method to measure species diversity: The lower the dominance, the higher the species diversity, and vice versa. Pielou’s eveness index (J’) was determined as the proportion of diversity observed in relation to the maximum diversity expected (Magurran, 1988). To compare the Shannon index between areas, we applied the Student’s t test modified by Hutchenson (Magurran, 1988).

Cortés–Marcial et al.

Table 1. Metrics of the disturbance index of livestock density variables, human activities variables and land degradation variables. Tabla 1. Valores del índice de perturbación de las variables relativas a la densidad de ganado, las actividades humanas y la degradación del suelo. Variable

Acronym

Description

Livestock density Goat droppings frequency Goat Cattle droppings frequency

Catt

Computed from presence of goat dung in ten randomly chosen 1 m squares along the transect; frequency was deﬁned as the fraction of squares with positive records. Bovine and equine dung, computed as for Goat.

Browsing Brow

All shrubs and trees that were rooted within the transect were thoroughly examined for signs of browsing. The ratio of browsed to total plants was calculated as an index of browsing intensity.

Livestock trail density Ltra

Livestock uses well–deﬁned trails to move while browsing. The number of these per meter along the transect was recorded.

Soil compaction Comp

The constant trampling of livestock along tracks causes soil compaction, which affects water inﬁltration. A cylinder of 10.4 cm of diameter was driven 4 cm into the ground in a randomly chosen trail. 250 ml of water were then poured into the cylinder, and the time needed for complete inﬁltration was recorded. This procedure was repeated on a spot with no evidence of trampling. The degree of soil compaction was calculated as the ratio of the time recorded on the trail and in the untrampled terrain.

Human activities Fuelwood extraction Fuel

Peasants cut branches for fuel. This metric was measured as Brow, but taking machete cuts into account.

Human trails density Tran

It was measured as Ltra, but recording trails used by people to travel.

Settlement proximity Prox

Proximity was deﬁned as the multiplicative inverse of the distance to the closest towns in km.

Contiguity to activity cores Core

A core was deﬁned as a place where human activities normally take place, such as houses, cornﬁelds, mines and chapels. Contiguity was recorded at each transect if a core was less than 200 m away. The fraction of transects contiguous to a core was used as a metric.

Land use Luse

In several studies the percent of land cover devoted to agriculture, cultivated or induced pastures, or urban areas is used as a measure of disturbance. Here, the fraction of the study area used for these purposes was visually estimated.

Evidence of ﬁres Fire

Most of these are initiated by people, either to clear an area, promote pasture growth for livestock, or accidentally. The presence or absence of evidence at a study site was recorded as one or zero.

Animal Biodiversity and Conservation 37.1 (2014)

Table 1. (Cont.) Variable

Acronym

Description

Land degradation Erosion Eros

Overgrazing and human activities increase erosion. We only considered spots where the soil showed tracks of strong and frequent removal of material by water (such as ravines) as unequivocal evidence of erosion. Twenty points were selected randomly along the transect for its estimation, and the fraction of eroded spots was recorded.

Presence of soil islands Isla

When severe erosion takes place, soil is only held where large shrubs are rooted. As a result, a landscape of small mounds can be observed. The presence or absence of these ‘‘islands’’ was recorded either as one or zero.

Totally modiﬁed surfaces Toms

Land may be so severely modiﬁed that measuring most of the previous metrics makes no sense, as it can happen on a paved road, a house, or on artiﬁcial water– ways. When the transect crossed such surfaces, their cover was measured by means of the line intercept method.

Furthermore, to analyze diversity more effectively, we calculated the effective number of species (true diversity) to know how much diversity was lost or gained between areas and between seasons. We used the exponential Shannon–Wiener index, in which all the species in the community are weighted in exact proportion to their abundance (Jost, 2006; Moreno et al., 2011). Beta diversity (change in species composition) between areas was evaluated using the Whittaker index (Wilson & Schmida, 1984; Magurran, 1988), which in this case can have values between 1 and 2, and the degree of similarity between habitats was evaluated according to the Jaccard similarity index (Magurran, 1988). Results The least disturbed area was located on the Tolistoque hill, hereafter called the 'preserved area'. The area located southeast of La Venta was named the 'disturbed area', and it showed greater disturbance due to its proximity to centers of activity, changes in land use, and islands (table 2). Clench’s species accumulation model was the best choice for the data, although asymptote was not reached in the study area. The model predicted 23 species (a = 6,806 and b = 0.297), meaning that our mammal inventory was 79% complete. According to this model, a total of 63 months would be required to record 95% of the medium and large mammal species living in the study site.

We obtained 167 records, of which 61% were traces and 28% were feces. Of all the records, 79 (47.30%) were found in the preserved area and 88 (52.70%) in the disturbed area (table 3). The records belonged to 18 species, 18 genera, 12 families and six orders of medium and large mammals (table 4). Through the use of camera traps, 82 photographs of mammals were obtained, confirming the presence of ten of the species recorded by indirect methods. In terms of relative abundance, Urocyon cinereoargenteus was the species with the highest abundance in the preserved area (0.23/km), while in the disturbed area the most abundant species were Didelphis virginiana (0.29/km) and U. cinereoargenteus (0.25/km) (table 4). According to the Mann–Whitney test, significant differences were found between the relative abundance in the two study areas (N1 = 79, N2 = 88, U = 63.5, p = 0.032). U. cinereoargenteus and D. virginiana were the most abundant species during the two seasons. In both areas, the highest relative abundance of species was observed in the rainy season. However, the seasonal variation in relative abundance was not statistically significant (U = 72, p = 0.76, and U = 23.5, p = 0.72, in preserved and disturbed areas, respectively). The preserved area exhibited the highest diversity (H’ = 2.33) and evenness (J’ = 0.82), and the lowest dominance (D = 30.86) (table 4). Significant differences were observed in the Shannon–Wiener index between the diversity of the preserved and disturbed areas (t = 4.9, d.f. = 160). The highest diversity was recorded during the rainy season in the preserved area

Cortés–Marcial et al.

Discussion Table 2. Values obtained with the disturbance index in two areas with different levels of perturbation near the Isthmus of Tehuantepec, Oaxaca. (For the abbreviations of variables see table 1.) Tabla 2. Valores obtenidos con el índice de perturbación en dos zonas con diferentes grados de perturbación en el istmo de Tehuantepec, Oaxaca. (Para las abreviaturas de las variables ver tabla 1.)

Area Preserved

Disturbed

Livestock density Goat

0.000

Catt

0.400

Fire

0.000

Brow

0.027

0.050

Ltra

0.020

Comp

0.185

0.260

Human activities Fuel

0.126

0.046

Fire

0.000

Tran

0.051

0.040

Prox

0.191

0.301

Core

0.000

1.000

Luse

0.000

1.000

Eros

0.125

0.700

Isla

0.000

1.000

Toms

0.000

0.050

14.339

67.074

Land degradation

Total disturbance

(H’ = 2.30). This area also showed lower dominance (D = 18.18) and higher evenness (J‘ = 0.92). No significant differences were found in the Shannon index between seasons for the preserved and disturbed areas (t = 1.40, g.l. = 80.56 and t = 1.68, g.l. = 73.60, respectively). According to the measure of true diversity, the diversity of medium and large mammals in the preserved area was double that of the disturbed area. During the dry season, the diversity of mammal species was lower than during the rainy season in both the preserved (24%) and disturbed areas (28%). Our data revealed a high turnover of species between zones (βw = 1.48). Of the 17 species recorded in this study, eight were found in both areas, while nine species were exclusively found in the preserved area. Spilogale gracilis was recorded only in the disturbed area. Finally, the two areas showed a similarity of 47% in species composition according to the Jaccard similarity index.

The indirect method was an efficient way to study mammal diversity in this study. Using this method we recorded 18 species of medium–sized and large mammals, whereas the camera traps only recorded the presence of ten species. However, this sampling was not standardized, as camera trapping was only used during the last six months of sampling. Consequently, we recommend the use of complementary methods to record a greater number of species. Indirect methods could however underestimate species richness and abundance as they focused mainly on recording terrestrial species and can overlook tree–dwellers (Aranda, 2000). Combining various techniques also reduces the influence of environmental and methodological factors, providing a more reliable estimate of diversity and abundance in a particular study site (Botello et al., 2008). Zarco (2007) recorded the same number of species as in this study using camera traps in the same vegetation type. This technique facilitates the determination of species' activity patterns, but it is expensive to implement compared to indirect methods. The species richness found in the area is equivalent to 34.62, 45.00, 57.89 and 66.67% of the total species, genera, families and orders of medium and large mammals present in Oaxaca. These values are higher than the 17 species reported by Santos–Moreno & Ruiz–Velásquez (2011) in the region of Isthmus of Tehuantepec in similar vegetation type, while Monroy– Vilchis et al. (2011) recorded 19 species with camera traps in an area where the main vegetation type, was tropical deciduous forest. These results show that the study area maintains a diverse community of medium and large mammals, despite the effects of disturbance (habitat deterioration and a high presence of human activities) in the south of the Protected Communal Area. The species richness found at the site, however seems low compared to the study by Cervantes & Yepez (1995) around Salina Cruz, in the coastal plain of Tehuantepec, Oaxaca. In their study, the authors recorded 30 species of medium–sized and large mammals. This difference may be due to the fact that Cervantes & Yepez (1995) conducted their study in tropical deciduous forest, mangrove forest, thorn scrub and dune vegetation, so a greater number of species occupying different ecological niches and ecosystems was recorded. This was seen in the case of Lontra longicaudis, for example, which is located only in aquatic environments. The number of species found in our study was similar to that reported by Lavariega et al. (2012) in the municipality of Santiago Camotlán. However, their study was conducted in cloud forest, oak forest, evergreen forest, crop fields and coffee plantations. Species typical of highly conserved sites, such as Panthera onca and Tamandua mexicana, are reported in some of these habitats. They are also recorded in association with cattle in disturbed areas, but in a lower proportion (Treves & Karanth, 2003). According to the Clench model, the species inventory is not fully represented, and it is likely that more species are still to be found in the area. We

Animal Biodiversity and Conservation 37.1 (2014)

Table 3. Number of records of medium and large mammals recorded in La Venta, Juchitan. Record types: F. Footprint; f. Feces; Sr. Skeletal remains; S. Sighting. Tabla 3. Número de registros de especies de mamíferos de talla mediana y grande registrados en La Venta, Juchitán. Tipos de registros: F. Huella; f. Excrementos; Sr. Restos óseos; S. Avistamiento.

Preserved area

Disturbed area

Total

S Total

Canis latrans

–

Coendou mexicanus

–

Conepatus leuconotus

–

Dasypus novemcinctus

–

Didelphis virginiana

–

Herpailurus yagouaroundi

–

Leopardus pardalis

–

Mustela frenata

–

Nasua narica

–

Odocoileus virginianus

–

Pecari tajacu

–

Philander opossum

–

Procyon lotor

–

Puma concolor

–

Spilogale putorius

–

Sciurus aureogaster

–

Sylvilagus floridanus

–

Urocyon cinereoargenteus

–

recorded the presence of Ateles geoffroyi at the north of the Tolistoque hill on April 2007 (16° 35' 52.97'' N / 94° 52' 35.56'' W), although its presence had not been reported by Ortiz–Martinez et al. (2008) in a study on the distribution of Alouatta palliata and A. geoffroyi. We did not include this latest species in our analysis given that we saw it only once, several months before the present study, in the north of the preserved area. By including A. geoffroyi, our inventory would reach 83% of completeness, and we would be missing only three species. One factor that could affect estimates of the relative abundance of species is the difference in the detectability of their traces, which is related to the size of the species (Litvaitis et al., 1994), their habits, their inclination while walking, and the type of substrate. It is therefore more likely to find tracks of D. virginiana because their weight facilitates track impressions and makes them easier to detect. On the contrary, the genus Sciurus may be more abundant than deer Odocoileus sp., but their habits are primarily arboreal, making track observations more difficult. It is noteworthy that the rainy season facilitated the record of tracks, mainly in areas of flooding, and at this season we recorded the greatest abundance of species.

The relative abundance of Dasypus novemcinctus was lower than that reported by Navarro (2005) in secondary forest and oak forest, as this author reported densities of 0.2 individuals/km at each vegetation type. Likewise, Perez–Irineo & Santos– Moreno (2012) reported an even higher relative abundance for the same species (0.07 individuals/ km) in a deciduous forest in northeastern Oaxaca. In our study, particularly the disturbed area is affected by strong human intervention, which may explain the low observed abundance of this species. Hunting may also contribute to decrease the abundance and increase the secretive and evasive behavior of some species. It is well known that medium and large sized mammal species are the most affected by hunting. In our study area local inhabitants and people from the surroundings were observed hunting. The most hunted species for meat consumption are armadillos D. novemcinctus, squirrels Sciurus aureogaster and rabbits Silvilagus floridanus. The high abundance of U. cinereoargenteus and D. virginiana corresponds with the findings of Orjuela & Jimenez (2004) and Luna (2005), who report that the fox has the highest relative abundance values. These high values of abundance may be related to

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Table 4. List of species of medium and large mammals recorded in La Venta, Juchitan, following the taxonomic arrangement proposed by Ramirez et al. (2005) and including the number of records (n) and relative abundance (Rel ab) in each of the areas and seasons. Index of diversity α and β. Status conservation NOM 059 (* Threatened, ** Endangered). (For the abbreviations of record types, Rec, see table 3.) Tabla 4. Lista de las especies de mamíferos de talla mediana y grande registradas en La Venta, Juchitán, siguiendo la taxonomía propuesta por Ramírez et al. (2005) e incluyendo el número de registros (n) y la abundancia relativa (Rel ab) en cada zona y temporada. Índices de diversidad α y β. Estado de conservación NOM 059 (* Amenazada, ** En peligro). (Para las abreviaturas de los tipos de registro, Rec, véase la tabla 3.)

Preserved area Rainy

Taxonomic list

Rec

Dry

Disturbed area

Total

n Rel ab n Rel ab n Rel ab

Rainy

Dry

Total

n Rel ab n Rel ab n Rel ab

O. Didelphimorphia / F. Didelphidae Didelphis virginiana

F, P

3 0.0556 7 0.1296 10 0.0926

21 0.3889 11 0.2037 32 0.2963

Philander oposum

1 0.0185 0 0.0000 1 0.0093

O. Cingulata / F. Dasypodidae Dasypus novemcinctus F, Sr, P

3 0.0556 0 0.0000 3 0.0278

11 0.2037 2 0.0370 13 0.1204

O. Canivora / F. Canidae Canis latrans

F, f, P

3 0.0556 2 0.0370 5 0.0463

–

0 0.0000

Urocyon cinereoargenteus F, f, S, P

6 0.1111 19 0.3519 25 0.2315

11 0.2037 16 0.2963 27 0.2500

O. Canivora / F. Felidae Herpailurus yagouaroundi* F Puma concolor

Leopardus pardalis**

F, P

0 0.0000 1 0.0185 1 0.0093

1 0.0185 0 0.0000 1 0.0093

1 0.0185 2 0.0741 3 0.0463

–

0 0.0000 2 0.03704 2 0.0185

0 0.0000 0

O. Canivora / F. Mustelidae Mustela frenata

0 0.0000 1 0.0185 1 0.0093

–

0 0.0000

O. Carnivora / F. Mephitidae Conepatus leuconotus

F, P

Spilogale putorius

F, P

1 0.0185 2 0.0370 3 0.0278 0 0.0000

–

0 0.0000

1 0.0185 2 0.0370 3 0.0278

O. Carnivora / F. Procyonidae Nasua narica

0 0.0000 1 0.0185 1 0.0093

1 0.0185 0 0.0000 1 0.0093

Procyon lotor

F, P

0 0.0000 1 0.0185 1 0.0093

5 0.0926 2 0.0370 7 0.0648

O. Artiodactyla / F. Tayassuidae Pecari tajacu

3 0.0556 0 0.0000 3 0.0278

–

0 0.0000

O. Artiodactyla / F. Cervidae Odocoileus virginianus F, f, Sr, P

6 0.1111 5 0.0926 11 0.1019

–

0 0.0000

O. Rodentia / F. Sciuridae Sciurus aureogaster

F, S

1 0.0185 2 0.0370 3 0.0278

–

0 0.0000

O. Rodentia / F. Erethizontidae Coendou mexicanus*

3 0.0556 0 0.0000 3 0.0278

–

0 0.0000

O. Lagomorpha / F. Leporidae Sylvilagus floridanus F, S, P Total records

2 0.0370 1 0.0185 3 0.0278

Total species

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Table 4. (Cont.)

Preserved area

Diversity α

Rainy

Shannon–Wiener Evenness (J’) Dominance Effectivenes diversity

2.304 0.927 0.181 10.010

Dry 2.023 0.789 0.395 7.563

Disturbed area

Total

Rainy

2.331 0.823 0.308 10.289

1.653 0.752 0.388 5.225

Dry 1.324 0.739 0.470 3.757

Total 1.597 0.727 0.363 4.939

% Diversity loss / areas 52.00 % Loss / seasons

24.449

28.099

Diversity β Whittaker 1.48 Jaccard 47%

the characteristics of the species; as omnivores, they are more likely to find food. Consequently, its presence is favored on disturbed areas, or in crops such as sorghum, one of the crops found in the region. We found evidence of sorghum consumption by foxes. The diversity values recorded for both the preserved and the disturbed areas are lower than those reported by Cueva et al. (2010) (H' = 2.4). However, their study area represents a very well preserved area with a greater extension, since it belongs to a biological reserve of about 730 ha in the reserve community Santa Lucía (Ecuador). Contrary, the mammal diversity in our study area is higher than that reported by Perez–Irineo & Santos–Moreno (2012) in a deciduous forest in Oaxaca (H' = 0.89). Therefore, our results are significant because this index is usually between 1.5 and 3.5 (Magurran, 1988). It also has been observed that H’ decreases as disturbance increases, varying from 0.98 to 2.16 according to the degree of the environmental disturbance. The results obtained in this study show that the preserved area is the most diverse, since in this area we found the lowest dominance and the highest evenness. The total values of diversity indexes in both areas of study show that the populations of medium–sized and large mammals respond to anthropogenic factors, which is reflected in a decrease in their diversity. The preserved area offers the best conditions under which species can develop their activities: find shelter, search for food, and reproduce. The greatest diversity in the preserved area may be due to the greater vegetation richness and greater canopy height, which increases the potential niches and provides more food resources, shelter, protection and escape opportunities to mammals (Gallina et al., 2007). In this area we also found species such as Puma concolor, Leopardus pardalis and Pecari tajacu, which can be considered indicators of well–preserved environments (Cruz–Lara et al., 2004).

The disturbed area may present lower diversity due to several processes found in the area. Human activities such as deforestation, the opening of roads, and noise pollution, affect the habitat directly and indirectly, and modify wildlife activity (Herrera–Flores et al., 2002). Nevertheless, this area still maintains moderate diversity because of the fast–growing vegetation used as habitat and a food source for mammals (Soto & Herrera–Flores, 2003). The presence of water bodies near the site also attracts some mammal species that can find food, water and shelter in the surrounding vegetation (Guzmán–Lenis & Camargo–Sanabria, 2004). Species diversity can change or remain stable in response to disturbances in the forest. Certain groups of animals, such as foxes, can increase their abundance. Thus, some species may increase their dominance, while the community species richness remains constant in the area. A change like this may decrease the diversity in the area. According to Rocha & Dalponte (2006), the absence of deer and puma in the disturbed area may be because the site does not meet the needs of a predator at on the top of the food chain, such as P. concolor, and does not provide a suitable habitat for the occurrence of O. virginianus. The values of beta diversity and similarity suggest a high species turnover. The medium and large mammals found in both areas are considered different communities according to the proposal of Sanchez & Lopez (1988), who propose that for two communities to be similar they should have a similarity of above 66.6%. The high species turnover may be mainly due to the fragmentation of local populations throughout the environment, derived from disturbances such as the presence of the Panamerican Highway 185, which separates the two areas and thus creates a barrier that limits the movement of organisms between areas. Also, the isolation of populations may cause local extinctions due to lack of genetic exchange with other

individuals from different populations (Arroyave et al., 2006). In this way the presence of human activity can have also an adverse effect on the dispersion pattern of animals. Acknowledgements We thank the authorities of La Venta, Juchitan for their support and the facilities provided. M. Cortes thanks CONACyT for the scholarship granted during the period August 2007 and June 2009. The study was supported by the Secretaría de Investigación y Posgrado of the IPN (SIP 20100263). The authors would also thank C. Moreno for her valuable suggestions and comments. M. Briones was funded by Estímulos al Desempeño de la Investigación, Comisión de Operación y Fomento a las Actividades Académicas of IPN and Sistema Nacional de Investigadores. We thank M. Lavariega for the map design. References Alonso, A., Dallameier, F. & Campbell, P., 2001. Urubamba: The biodiversity of a Peruvian rainforest. Smithsonian Institution, Washington, D.C. Aranda, M., 2000. Huellas y otros rastros de los mamíferos grandes y medianos de México. Primera edición. Ed. Instituto de Ecología, AC, Veracruz– México. Arroyave, M. P., Gómez, C., Gutiérrez, M. E., Múnera, D. P., Zapata, P. A., Vergara, I. C., Andrade, I. M. & Ramos, K. C., 2006. Impactos de las carreteras sobre la fauna silvestre y sus principales medidas de manejo. Revista Escuela de Ingeniería Antioquia (EIA), 5: 45–57. August, P., 1983. The role of habitat complexity and heterogeneity in structuring tropical mammal communities. Ecology, 64: 1495–1507. Barragán, F., Lorenzo, C., Morón, A., Briones–Salas, M. & López, S., 2010. Bat and rodent diversity in a fragmented landscape on the Isthmus of Tehuantepec, Oaxaca, Mexico. Tropical Conservation Science, 3(1): 1–16. Bilenca, D., Balla, P., Álvarez, M. & Zalueta, G., 1999. Evaluación de dos técnicas para determinar la actividad y abundancia de mamíferos en el bosque chaqueño, Argentina. Revista Ecológica Latino Americana, 6(1): 13–18. Bolaños, C. & Naranjo, J. E., 2001. Abundancia, densidad y distribución de las poblaciones de ungulados en la cuenca del río Lacatún, Chiapas, México. Revista Mexicana de Mastozoología, 5: 45–57. Botello, F., Monroy, G., Illoldi–Rangel, P., Trujillo–Bolio, I. & Sánchez–Cordero, V., 2007. Sistematización de imágenes obtenidas por fototrampeo: una propuesta de ficha. Revista Mexicana de Biodiversidad, 78: 207–210. Botello, F., Sánchez–Cordero, V. & González, G., 2008. Diversidad de carnívoros en Santa Catarina Ixtepeji, Sierra Madre de Oaxaca, México. In: Avances en el estudio de los mamíferos de México. Publicaciones

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Animal Biodiversity and Conservation 37.1 (2014)

Plankton composition and environmental parameters in the habitat of the Iranian cave barb (Iranocypris typhlops) in Iran A. Farashi, M. Kaboli, H. Reza Rezaei, M. Reza Naghavi & H. Rahimian

Farashi, A., Kaboli, M., Reza Rezaei, H., Reza Naghavi, M. & Rahimian, H., 2014. Plankton composition and environmental parameters in the habitat of the Iranian cave barb (Iranocypris typhlops) in Iran. Animal Biodiversity and Conservation, 37.1: 13–21. Abstract Plankton composition and environmental parameters in the habitat of the Iranian cave barb (Iranocypris typhlops) in Iran.— The Iranian cave barb (Iranocypris typhlops Bruun & Kaiser, 1944) is 'Vulnerable' in the IUCN Red List. It is an endemic species of ray–finned fish of the family Cyprinidae from a single locality in the Zagros Mountains, western Iran. This species is an omnivore that depends on plankton for food. We studied the spatial and seasonal distribution of plankton in the native habitat of the Iranian cave barb between May 2012 and February 2013. We measured various environmental parameters and related these to plankton distribution. The plankton assemblage included 13 genera and five species. Rotifera had the highest number of genera (4) and species (4), followed by Arthropoda (3), Ochrophyta (3), Myzozoa (2), Charophyta (2), Chlorophyta (2), Ciliophora (1) and Cryptophyta (1). In terms of numbers, the dominant species of phytoplankton and zooplankton were Achnanthidium sp. and Lecane sp. Pearson correlation coefficients showed a low but significant relationship between plankton communities and environmental parameters. Among the environmental parameters, total suspended solids and turbidity seemed to have the most important influence on the temporal distribution of plankton species. We also observed that dissolved oxygen played an important role for most plankton species, as did temperature for most zooplankton species. The diversity and abundance of phytoplankton and zooplankton were low throughout the year in the cave with an annual mean of 96.4 ind./l and they did not show any peaks during the year. Key words: Iranian cave barb, Endemic, Habitat, Phytoplankton, Zooplankton. Resumen La composición planctónica y los parámetros ambientales en el hábitat del barbo cavernícola iraní Iranocypris typhlops.— El barbo cavernícola iraní (Iranocypris typhlops Bruun & Kaiser, 1944) es una especie catalogada como "Vulnerable" en la Lista Roja de la IUCN. Endémica de una única localidad situada en las montañas Zagros, en Irán occidental. Se trata de una especie omnívora que depende del plancton para alimentarse. Se estudió la distribución espacial y estacional del plancton en el hábitat original del barbo cavernícola iraní entre mayo de 2012 y febrero de 2013. Se midieron varios parámetros ambientales y se relacionaron con la distribución del plancton. La comunidad planctónica comprendía 13 géneros y cinco especies. El filo Rotifera tenía el mayor número de géneros (4) y de especies (4), seguido por Arthropoda (3), Ochrophyta (3), Myzozoa (2), Charophyta (2), Chlorophyta (2), Ciliophora (1) y Cryptophyta (1). Por lo que respecta a la cantidad, las especies dominantes de fitoplancton y zooplancton fueron Achnanthidium sp. y Lecane sp. Los coeficientes de correlación de Pearson pusieron de manifiesto que la relación entre las comunidades de plancton y los parámetros ambientales era baja pero significativa. Entre los parámetros ambientales, el total de sólidos en suspensión y la turbidez parecieron ser los más influyentes en la distribución temporal de las especies de plancton. Asimismo, se observó que el oxígeno disuelto desempeñaba una función importante para la mayoría de las especies de plancton, al igual que la temperatura para la mayoría de las especies de zooplancton. La diversidad y la abundancia de fitoplancton y zooplancton eran bajas durante todo el año en la cueva con una media anual de 96,4 ind./l y no mostraron ningún máximo durante el año.

ISSN: 1578–665 X eISSN: 2014–928 X

Farashi et al.

Palabras clave: Barbo cavernícola iraní, Endémico, Hábitat, Fitoplancton, Zooplancton. Received: 4 XI 13; Conditional acceptance: 11 XII 13; Final acceptance: 21 II 14 Azita Farashi, Dept. of Environmental Sciences, Fac. of Natural Resources and environment, Ferdowsi Univ. of Mashhad, Iran.– Mohammad Kaboli, Dept. of Environmental Sciences, Fac. of Natural Resources, Univ. of Tehran, Iran.– Hamid Reza Rezaei, Dept. of Environmental Sciences, Fac. of Natural Resources, Gorgan Univ., Iran.– Mohammad Reza Naghavi, Dept. of Biotechnology, Fac. of Agricultural Sciences, Univ. of Tehran, Iran.– Hassan Rahimian, Dept. of Animal Biology, Fac. of Biology, Univ. of Tehran, Iran. Corresponding author: Mohammad Kaboli, Daneshkade Street, Karaj, 31587–77871 Iran.

Animal Biodiversity and Conservation 37.1 (2014)

Introduction The Iranian cave barb (Iranocypris typhlops Bruun & Kaiser, 1944) is a rare species of the family Cyprinidae endemic to the Zagros Mountains, western Iran (Mahjoorazad & Coad, 2009). The distribution of the species seems to be restricted to a single cave. I. typhlops is sympatric with Paracobitis smithi (Greenwood, 1976) and both are listed as 'Vulnerable' in the IUCN Red List (IUCN, 2013). As such, Coad (2000), using 18 criteria that focused on distribution and habitat, found this species to be one of the top four threatened species of freshwater fishes in Iran. Zalaghi (2011) estimated the population size of the species at between 353 and 625 individuals. Conservation of this species has received little attention so far. The major conservation objective, perhaps reinforced by legislation, must be habitat restoration and management. Knowledge on this species habitat is poor. Assemblages of species in ecological communities reflect interactions between organisms and the abiotic environment as well as among organisms (Hughes, 2000). Plankton species are valuable indicators of environmental conditions (Beaugrand, 2004; Bonnet & Frid, 2004) since they are ecological indicators of many physical, chemical and biological factors. On the other hand, the diet of this cave species is extremely dependent on plankton (> 60%), as found in field observations. Food density is a main environmental variable for appearance and abundance of fishes (McNamara & Houston, 1987; Hüppop, 2005). Therefore, information on the species’ feeding could prove useful

for urgently needed conservation measures, such as breeding programs, stock maintenance or translocation, as well as habitat rehabilitation measures (Kalogianni et al., 2010), similar to those successfully implemented for the conservation of its related species, the native Iberian toothcarp Valencia hispanica (Planelles & Reyna, 1996; Risueño & Hernández, 2000; Caiola et al., 2001). A habitat and diet overlap study could improve our knowledge of their effects of such measures on the species, and help develop appropriate management strategies. In this study we outline a one–year study of the plankton community in the habitat, in order to analyse species composition and seasonal dynamic of the plankton community, and to assess their dependence on environmental parameters. Methods Study area The Iranian cave barb’s original locality is a water cave, the natural outlet of a subterranean limestone system in the Zagros Mountains. The stream below the cave locality is the 'Ab–e Sirum', a tributary of the Dez River, in Lorestan province. The Dez flows into the Karun River which drains to the head of the Persian Gulf. The cave is located at 33° 04' 39'' N and 48° 35' 33'' E (fig. 1). Recently, this fish has been reported in another locality. The new locality is at 131 km in a direct line from the type locality. The

Caspian Sea

Seymareh River

Persian Gulf

< Omman Sea

33° 16' 56' N 47° 12' 16'' E

Ab–e Sirum 33° 04' 39'' N 48° 35' 33'' E

Bagh–e Levan

Dez Dam Lake Karkheh Dam Lake Karkheh River

Dez River

50 km

Fig. 1. Location of Iranocypris typhlops habitat in Bagh–e Levan and the new locality reported by Mahjoorazad & Coad (2009) in Iran. Fig. 1. Ubicación del hábitat de Iranocypris typhlops en Bagh–e Levan y la nueva localidad registrada por Mahjoorazad & Coad (2009) en Irán.

Farashi et al.

Table 1. Environmental parameters with their mean, range of variation and statistical difference observed during the study period. Statistical results are for two–way ANOVA (main effects: S. Season, D. Depth; interaction: S x D): *p < 0.05; **p < 0.01. Tabla 1. Parámetros ambientales con sus medias, rango de variación y diferencia estadística observados durante el período de estudio. Los resultados estadísticos son para ANOVA de dos factores (efectos principales: S. Estación, D. Profundidad; interacción: S x D): * p < 0,05; ** p < 0,01.

ANOVA Variables Mean ± SD Min Max S D S × D Physical variables pH 7.61 ± 0.17 7.10 7.90 ** Electrical conductivity (EC, µs/cm) 458.59 ± 20.38 430.00 506.00 ** Turbidity (NTU) 0.56 ± 0.07 0.42 0.71 ** ** ** Water temperature (T, °C) 18.14 ± 1.83 15.00 24.00 ** ** ** Dissolved oxygen (DO, mg/l) 5.83 ± 0.85 4.50 7.60 ** ** ** Total suspended solids (TSS, mg/l) 0.53 ± 0.05 0.26 0.80 ** ** Total dissolved solids (TDS, mg/l) 244.46 ± 7.39 226.00 258.00 ** Metals Magnesium (mg/l) 19.48 ± 0.79 Potassium (mg/l) 3.13 ± 0.20 Sodium (mg/l) 19.20 ± 0.94 Calcium (mg/l) 54.68 ± 4.29 Iron (II) (mg/l) 00.00 ± 0.00 Total iron (TFe, mg/l) 00.00 ± 0.00 Inorganic (non–metallic) matter Chlorine (mg/l) 29.16 ± 1.64 Bicarbonate (mg/l) 149.54 ± 3.85 Carbonate (mg/l) 0.00 ± 0.00 Total nitrogen (TN, mg/l) 1.30 ± 0.07 Nitrite (mg/l) 0.00 ± 0.00 Nitrate (mg/l) 0.51 ± 0.03 Total phosphorus (TP, mg/l) 0.59 ± 0.04 Phosphate (mg/) 0.34 ± 0.05 Total sulfur (TS, mg/l) 88.79 ± 4.63 Sulfate (mg/l) 58.57 ± 2.84 Organic matter Biological oxygen demand (BOD, mg/l) 0.01 ± 0.00 Chemical oxygen demand (COD, mg/l)

0.22 ± 0.05

17.3 2.70 17.15 46.50 0.00 0.00

21.18 3.70 22.80 61.50 ** 0.00 0.00

24.80 136.00 0.00 1.20 0.00 0.46 0.50 0.20 79.80 50.14

33.00 ** 158.70 ** 0.00 1.45 ** 0.00 0.60 ** 0.68 ** 0.45 ** 99.30 ** 66.00 **

0.00

0.15

0.01

0.31

construction of a dam on the Seymareh River on the Lorestan–Ilam provincial border, 30 km northwest of Darrehshahr at 33° 16' 56'' N and 47° 12' 16'' E, involved excavation of an intake tunnel for a power house, 11 m in diameter and 1,500 m in length, at 597 m altitude. The tunnel intersected many faults, joints and small karstic features. Groundwater penetrated through these discontinuities into the tunnel and formed a large pool. The tunnel is now encased in concrete and the karst environment is no longer accessible (Mahjoorazad & Coad, 2009).

Sample collection and analysis The Iranian cave barb’s original locality was found to have a depth of 28 m by local divers but for sampling only 5 m from the surface was attainable by a Ruttner sampler. Water samples were seasonally taken from the surface (20 cm) to a depth of 5 m below the cave surface at ten sites of each depth of the cave using a Ruttner sampler (volume of 10 l, 1 l for environmental variables and 9 l for plankton samples) at successive depth intervals of 1 m (Talling, 2003). This sampling

Animal Biodiversity and Conservation 37.1 (2014)

PC2

Ca2+

Na+

Spring

NO3

Winter

Mg2+

Cl– K+ –

Summer

SO42–

HCO3–

Turbidity TSS

EC TDS

TS BOD5

PC1

Autumn PO43+

Fig. 2. PCA ordination diagram for the cave seasons and environmental factors. It shows the distribution of environmental parameters over the four seasons during the study period. (The abbreviations used for environmental variables are given in table 1.) Fig. 2. Diagrama de ordenación mediante análisis de componentes principales de las estaciones y los factores ambientales en la cueva. Se muestra la distribución de los parámetros ambientales en las cuatro estaciones durante el período de estudio. (Las abreviaturas utilizadas para las variables ambientales se indican en la tabla 1.)

took place from May 2012 to February 2013 at 8:00–10:00 a.m. The environmental physicochemical variables are listed in table 1. Inorganic (non–metallic) matter, organic matter, metals and TDS were analyzed in the laboratory according to APHA (2012) methods and DO, T, pH, EC, TSS and turbidity were detected in situ. Plankton samples filtered through a net of mesh size 30 μm. All the concentrated plankton samples (total volume of 100 l) were divided into two parts; 50 ml preserved with Lugol’s solution for the enumeration and identification of phytoplankton, and 50 ml preserved with 4% neutral formalin for the enumeration and identification of zooplankton. For each sample (total volume of 50 ml), 20 counts of 1–ml subsamples were counted using an inverted microscope under at 40–600X magnifications. Plankton samples were identified to the lowest taxonomic level possible.

variables that could affect plankton distribution. Due to low Pearson correlations, we were unable to use an analysis technique to elucidate the relationships between biological assemblages of species and their environment. These analyses were performed with SAS software (SAS Institute Inc., Cary, NC, USA). Principal component analysis (PCA), an indirect gradient analysis technique, was used to detect the main environmental variables in the cave in CANOCO version 4.5 (Braak & Šmilauer, 2002). Data were logarithmically transformed to normalize the distribution prior to statistical analysis. We used the Shannon–Wiener diversity index H′ to ascertain the structural features of the plankton community.

Statistical analysis

Environmental variables

We used a two–way ANOVA followed by Duncan’s tests to examine the effects of depth and season on environmental variables and plankton densities. Pearson correlations were run between environmental variables and plankton density to distinguish key biotic

The main environmental variables of the water are reported in table 1. Most environmental variables exhibited significant difference between seasons. However, water temperature underwent a typical seasonal trend, with a minimum of 15°C in winter and a maximum of

Results

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Table 2. Planktonic species with their mean, range of variation and statistical difference observed during the study period. Statistical results are for two–way ANOVA (main effects: S. Season, D. Depth; interaction: S x D): * p < 0.05; ** p < 0.01. Tabla 2. Especies de plancton con sus medias, rango de variación y diferencia estadística observados durante el período de estudio. Los resultados estadísticos son para ANOVA de dos factores (efectos principales: S. Estación, D. Profundidad; interacción: S x D): * p < 0,05; ** p < 0,01. ANOVA Mean ± SD Plankton (ind./l) S D S x D Rotifera 9.90 ± 6.69 ** ** Lecane sp. 3.72 ± 3.90 Brachionus sp. 3.50 ± 3.91 * Trichocerca sp. 0.98 ± 1.41 * * Philodina sp. 1.70 ± 2.03 ** Arthropoda 0.21 ± 0.24 ** Tropocyclops sp. 0.07 ± 0.10 ** Nauplius sp. 0.09 ± 0.20 Mesocyclops sp. 0.05 ± 0.08 Ciliophora 0.06 ± 0.09 Vorticella similis 0.06 ± 0.09 Total zooplankton 10.17 ± 6.72 ** ** ** Ochrophyta 62.79 ± 31.47 ** ** ** Melosira varians 19.53 ± 13.52 ** ** Synedra sp. 20.62 ± 13.60 ** ** * Achnanthidium sp. 22.64 ± 18.35 Cryptophyta 0.49 ±0.77 Cryptomonas sp. 0.49 ± 0.77 Myzozoa 3.72 ± 3.91 * Gymnodinium sp. 0.13 ± 0.26 Peridinium sp. 3.72 ± 3.90 ** Chlorophyta 10.14 ± 8.20 ** Pediastrum boryanum 1.39 ± 1.70 ** Botrycoccus braunii 8.75 ± 7.94 ** Charophyta 12.09 ± 12.82 * ** ** Staurastrum ophiurum 0.11 ± 0.71 Spirogyra sp. 11.98 ± 12.83 * ** ** Total phytoplankton 86.16 ± 39.16 ** ** Total plankton 96.39 ± 42.47 ** ** ** 24°C in summer. pH was mostly neutral and measured between 7.1 and 7.9. The maximum concentration of DO was recorded in spring (7.60 mg/l) within the surface layer. The predominant anions and cations can

Table 3. Seasonal variation in zooplankton abundance (%): Sm. Summer; Sp. Spring; At. Autumn; Wn. Winter. Tabla 3. Variación estacional en la abundancia de zooplancton (%): Sm. Verano; Sp. Primavera; At. Otoño; Wn. Invierno.

Zooplankton Rotifera

Sm 97.09

Sp At Wn 97.21 97.47 97.82

Arthropoda

2.21

2.00

1.64

Ciliophora

0.69

0.58

0.53

0.55

Table 4. Seasonal variation in phytoplankton abundance (%): Sm. Summer; Sp. Spring; At. Autumn; Wn. Winter. Tabla 4. Variación estacional en la abundancia de fitoplancton (%): Sm. Verano; Sp. Primavera; At. Otoño; Wn. Invierno.

Phytoplankton Ochrophyta Cryptophyta Myzozoa Chlorophyta Charophyta

Sm Sp At Wn 77.41 73.58 68.05 69.31 0.54 0.58 0.63 0.49 0.42 0.35 0.32 5.06 11.27 11.10 13.48 11.11 10.36 14.39 17.51 14.03

be arranged in the following sequence in decreasing order of their average concentration: HCO– > TS > SO42– > Cl– > T N > TP > NO3– > PO43– and Ca2+ > Mg2+ > Na+ > K+. Also CO32– and NO2– concentration in anions and Fe+2 and TFe concentration in cations were zero during the study period. The PCA of environmental variables showed that most variability (86%) can be explained by two main principal components (fig. 2). Variables most responsible for differentiating samples in the PCA included T, TS, NO3–, Ca2+, PO43–, BOD5, and COD (fig. 2). A PCA biplot clearly indicates the correlation between variables as well as the relative importance of each variable in explaining the overall variability in the environmental data (fig. 2). In general, similar variables clustered together: (i) Ca2+, Mg2+, K+ and Na+; (ii) BOD5 and COD; (iii) TDS and turbidity. The distribution of each parameter over the seasons can be analyzed by the positions of the seasons with respect to the environmental factors. For example, the winter season (lower left quadrant) showed high values for turbidity, DO, TN and PO43–, whereas the autumn season (lower right quadrant) tended to be

Animal Biodiversity and Conservation 37.1 (2014)

Table 5. Pearson correlation coefficients between environmental variables and plankton density: * p < 0.05; ** p < 0.01. (For abbreviations see table 1.) Tabla 5. Coeficientes de correlación de Pearson entre las variables ambientales y la densidad de plancton: * p < 0,05; ** p < 0,01. (Para las abreviaturas ver tabla 1.) Plankton

Turbidity

Rotifera

0.287** 0.189**

Arthropoda

0.145*

Ciliophora

TSS

TDS

0.243** 0.247**

0.146*

0.204**

NO3

BOD5

–0.150* 0.166* –0.157* 0.144*

0.196**

0.140*

Total zooplankton

0.292** 0.195**

0.245** 0.255** 0.128* -0.155* 0.172**

Ochrophyta

0.148*

0.166** 0.160

Cryptophyta

0.142*

0.159*

COD

Myzozoa

0.217** 0.181**

Chlorophyta

0.182**

0.248** 0.274**

Charophyta

0.392**

0.471** 0.388**

Total phytoplankton

0.156*

0.198**

0.227** 0.214**

Total plankton

0.152*

0.231**

0.239** 0.226**

associated with higher concentrations of variables such as TDS, TSS, BOD5, TS and EC (fig. 2). Variables such as Ca2+, Mg2+, K+ and Na+ were significantly correlated with axis 2. In contrast, factors BOD5, COD, TS and T were significantly associated with axis 1. These findings indicate that the second axis is likely related to the cations while the first axis is likely related to the degradation of organic matter, a process that influences the BOD5 and COD values. Plankton species composition and density The plankton species identified in the cave are shown in table 2. The plankton assemblage included 13 genera and cinco species. Rotifera had the highest number of genera (4) and species (4) followed by Arthropoda (3), Ochrophyta (3), Myzozoa (2), Charophyta (2), Chlorophyta (2), Ciliophora (1) and Cryptophyta (1). In terms of numbers, the dominant species of phytoplankton and zooplankton were Achnanthidium sp. and Lecane sp. with an annual mean of total numbers of 22.64 and 3.72 ind./l, respectively. The seasonal total phytoplankton and zooplankton abundance ranged from 6.40 to 288.20 and from 0.00 to 37.10 ind./l respectively. The seasonal composition of the plankton community is presented in tables 3 and 4. Univariate analysis of variance revealed significant differences between seasons for a few species and between depth layers for most species (table 2). During all seasons, the zooplankton and phytoplankton communities were mainly composed of Rotifer and Ochrophyta, respectively, contributing to the total abundance with a percentage ranging from 97.09% (in summer) to 97.82% (in winter) for Rotifer

0.172**

0.157**

0.128

0.142

and 68.05% (in autumn) to 77.41% (in summer) for Ochrophyta (tables 3, 4). The Shannon–Wiener diversity index (H´) ranged from 0.84 to 0.98 decits, generally showing lower values in autumn, winter and summer, and higher values in spring. Influence of environmental variables on plankton Pearson correlation coefficients between environmental variables and plankton density (table 5) varied between –0.150 and 0.471 for all species, showing a low significant relationship. Among the environmental parameters, TSS and turbidity seemed to have the highest influence on the temporal distribution of plankton species (table 5). Furthermore, DO play an important role for most plankton species and temperature for most of zooplankton species. All the environmental variables correlated positively with plankton except NO3–, which correlated negatively with several zooplankton species. Among the other parameters, BOD5, COD and NO3– were important factors for several zooplankton species, but they played a less important role for phytoplankton (table 5). Discussion and conclusions Temperate caves are, in general, stable and characterized by a permanent absence of light and temperatures similar to those in the external environment (Ferreira & Martins, 2001; Prous et al., 2004), but tropical and subtropical caves show a great degree of variability in their environmental parameters. A popular misconcep-

tion about cave environments is that they are always poor in biodiversity and biomass. This misconception stems from the fact that most cave research has been conducted in temperate caves where biodiversity and biomass are rather poor (Romero, 2009). Population size in fish is limited by food density (McNamara & Houston, 1987). In the present study, the diversity and abundance of phytoplankton and zooplankton in the cave was low. This low food density might account for the small population of the species, estimated to be only 330 and 526 individuals (Zalaghi, 1997). There are no previous reports providing data on the plankton community of the cave. The cave differed significantly from other nearby aquatic ecosystems as results of its unusual nature: low transparency, low plankton abundance, and low concentrations of nutrients. Our results provide a spatial and temporal account of the plankton communities in the cave. The highest species diversity was found in the spring, when Cl– was lowest and pH was highest. We did not observe seasonal patterns in plankton density in the cave. The most abundant zooplankton and phytoplankton were Rotifers and Ochrophyta, respectively. Compared with the terrestrial environment, the aquatic ecosystem has few physical barriers obstructing the mixing of planktonic species (Prous et al., 2004). In this study, we aimed to develop parameters to predict the relationships between the plankton community and environmental parameters. Although the cave is not a homogametic aquatic ecosystem there is a significant but low relationship between the plankton community and environmental parameters. Our study of the plankton community in relation to environmental parameters in the cave showed that the environmental variables could not have been responsible for the present species composition in the cave. Numerous studies state the importance of environmental variables on community structure and the plankton community in aquatic ecosystems. In the present study, TSS and turbidity were the most important factors for all plankton species. Nogueira et al. (1999) and Bonecker & Aoyagui (2005) showed that increased turbidity and the consequent decrease in light penetration to the deeper water layers influence plankton density. DO is considered one of the most important abiotic parameters affecting the plankton occurrence and distribution (Zurek, 2006; Vanderploeg et al., 2009; Chen et al., 2011). In the present study, DO played an important role for most plankton species, and the highest significant correlation was between DO and charophyta abundance. Temperature is considered to be a crucial factor that influences many aspects of the biology and ecology of the zooplanktonic organisms (Wetzel, 2001). It has been reported that temperature affects zooplankton occurrence and distribution (Akbulut et al., 2008; Huber et al., 2010), and in the present study temperature played an important role for most zooplankton species. BOD5, COD and NO3– were among the major environmental variables influencing zooplankton in the cave, similar to results reported by others (Arora & Mehra, 2009; Chalkia et al., 2012) NO3–, however,

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was negatively correlated with zooplankton density, a finding also supported by other studies (Tolotti et al., 2006; Chalkia et al., 2012). Our study presented spatial and temporal variation in environmental variables and plankton species and confirmed that the habitat has low plankton density. The results from this study could be useful for conservation efforts such as habitat rehabilitation and animal translocation programs as well as a basis for future research and monitoring efforts References Akbulut, N., Akbulut, A. & Park, Y. S., 2008. Relationship between zooplankton (Rotifera) distribution and physico–chemical variables in Uluabat Lake (Turkey). Fresenius Environmental Bulletin, 17(8): 947–955. APHA, AWWA, WEF, 2012. Standard Methods for the Examination of Water and Wastewater, 22nd ed. American Public Health Association, Washington DC. Arora, J. & Mehra, N. K., 2009. Seasonal dynamics of zooplankton in a shallow eutrophic, man–made hyposaline lake in Delhi (India): role of environmental factors. Hydrobiologia, 626: 27–40. Beaugrand, G., 2004. The North Sea regime shift: evidence, causes, mechanisms and consequences. Progress in Oceanography, 60: 245–262. Bonecker, C. C. & Aoyagui, A. S. M., 2005. Relationships between rotifers, phytoplankton and bacterioplankton in the Corumba reservoir, Goias State, Brazil. Hydrobiologia, 546: 415–421. Bonnet, D. & Frid, C. L. J., 2004. Seven copepod species considered as indicators of water–mass influence and changes: results from a Northumberland coastal station. ICES Journal of Marine Science, 61: 485–491. Braak, C. J. F. ter & Šmilauer, P., 2002. CANOCO reference manual and CanoDraw for Windows User’s guide: Software for Canonical Community Ordination (version 4.5). Microcomputer Power, Ithaca. Bruun, A. F. & Kaiser, E. W., 1944. Iranocypris typhlops n. g., n. sp., the first true cave fish from Asia. Danish Scientific Investigations in Iran, Copenhagen, 4: 1–8. Caiola, N., Vargas, M. J. & De Sostoa, A., 2001. Feeding ecology of the endangered Valencia toothcarp, Valencia hispanica (Actinopterygii: Valenciidae). Hydrobiologia, 448: 97–105. Chalkia, E., Zacharias, I., Thomatou, A. A. & Kehayias, G., 2012. Zooplankton dynamics in a gypsum karst lake and interrelation with the abiotic environment. Biologia, 67: 151–163. Chen, P. Y., Lee, P. F., Ko, C. J., Ko, C. H., Chou, T. C. & Teng, C. J., 2011. Associations Between Water Quality Parameters and Planktonic Communities in Three Constructed Wetlands, Taipei. Wetlands, 31: 1241–1248. Coad, B. W., 2000. Criteria for assessing the conservation status of taxa (as applied to Iranian freshwater fishes). Biologia, 55(5): 539–557.

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Animal Biodiversity and Conservation 37.1 (2014)

A spatially explicit approach for estimating space use and density of common genets P. Sarmento, J. Cruz, C. Eira & C. Fonseca

Sarmento, P., Cruz, J., Eira, C. & Fonseca, C., 2014. A spatially explicit approach for estimating space use and density of common genets. Animal Biodiversity and Conservation, 37.1: 23–33. Abstract A spatially explicit approach for estimating space use and density of common genets.— Many species that occur at low densities are not accurately estimated using capture–recapture methods as such techniques assume that populations are well–defined in space. To solve this bias, spatially explicit capture–recapture (SECR) models have recently been developed. These models incorporate movement and can identify areas where it is more likely for individuals to concentrate their activity. In this study, we used data from camera–trap surveys of common genets (Genetta genetta) in Serra da Malcata (Portugal), designed to compare abundance estimates produced by SECR models with traditional closed–capture models. Using the SECR models, we observed spatial heterogeneity in genet distribution and density estimates were approximately two times lower than those obtained from the closed popula� tion models. The non–spatial model estimates were constrained to sampling grid size and likely underestimated movements, thereby overestimating density. Future research should consider the incorporation of cost–weighed models that can include explicit hypothesis on how environmental variables influence the distance metric. Key words: Camera–trapping, Capture–recapture, Genet, Program MARK, SECR package, Spatial models. Resumen Un método espacialmente explícito para estimar el uso del espacio y la densidad de la jineta común.— Muchas especies con baja densidad de población no se estiman con precisión utilizando métodos de captura y recaptura, puesto que tales técnicas suponen que las poblaciones están bien definidas en el espacio. Para resolver este sesgo, recientemente se han elaborado modelos de captura y recaptura espacialmente explícitos (SECR). Estos modelos incorporan el movimiento y pueden determinar las zonas en la que es más probable que los individuos concentren su actividad. En el presente estudio, utilizamos datos obtenidos con cámaras de trampeo en estudios sobre la jineta común (Genetta genetta) en Serra da Malcata (Portugal) concebidos para comparar las estimaciones de abundancia producidas por los modelos de SECR y los modelos tradicionales de captura para una población cerrada. Utilizando los modelos de SECR, observamos la existencia de heterogeneidad espacial en la distribución de la jineta y las estimaciones de la densidad fueron aproximadamente dos veces inferiores a las obtenidas en los modelos para poblaciones cerradas. Las estimaciones del modelo no espacial se limitaron al tamaño de la cuadrícula de muestreo y probablemente infravaloraron los movimientos, lo que conllevaría que se sobreestimara la densidad. Los estudios futuros deberían sopesar la incorporación de modelos de carga–distancia que puedan incluir hipótesis explícitas sobre la forma en que las variables medioambientales influyen en la métrica de la distancia. Palabras clave: Trampeo con cámara, Captura y recaptura, Jineta, Programa MARK, Paquete de SECR, Modelos espaciales. Received: 27 VIII 13; Conditional acceptance: 20 XI 13; Final acceptance: 12 III 14 Pedro Sarmento, Catarina Eira & Carlos Fonseca, Center for Environmental and Marine Studies and Biology Dept., Aveiro Univ., Campus Universitário de Santiago, 3810–193 Aveiro, Portugal.– Joana Cruz, Environment Dept., Univ. of York, YO10 5DD, United Kingdom. Corresponding author: P. Sarmento. E–mail: sarmentop@gmail.com ISSN: 1578–665 X eISSN: 2014–928 X

Introduction Researchers seeking to understand population proc� esses need methods and models that provide accurate inferences on population status. Camera trap methods have proven to be very useful to study animals that live in areas that are too large to study as a whole, and ani� mals that are elusive and difficult to observe (Gardner et al., 2010a). Many of the wide ranging species studied with camera traps methods occur at low densities. In such species, the conventional capture–recapture (CR) theory to this species is likely unsatisfactory as it is more suited to high density species occurring over relatively small ranges (Foster & Harmsen, 2012). The main problems when dealing with wide–ranging and/or elusive species are the relatively small sample sizes, low capture probabilities, and the effects of individual heterogeneity (Royle et al., 2011a). The usual method to analyse density is to apply closed population mod� els (White & Burnham, 1999), and convert these to densities using fundamentally ad hoc methods (Trolle et al., 2007). This approach presents two important difficulties: (1) the assumption of geographic closure of the population, i.e., no movement in and out of the sampling grid (White et al., 1982) which is frequently violated (Karanth & Nichols, 1998), and (2) the difficulty of estimating the effective sampling area (Balme et al., 2009a). The typical methodology consists of generating a buffer around the grid with half the mean maximum linear distance moved by animals captured in more than one trap (1/2MMDM) (Karanth & Nichols, 1998). Other approaches to estimate the buffer width include the full MMDM (Trolle et al., 2007), and the radius of an average home range (Sarmento et al., 2009).These approaches also present major problems: (1) They lack theoretical justification, being mostly ad hoc methods (Bochers & Efford, 2008); and (2) comparisons of es� timates from different methodologies become difficult (Sollmann et al., 2011). Furthermore, conventional CR methods assume that populations are well–defined in the sense that one can randomly sample individuals associated with a defined study location or area (Royle et al., 2009). However, individuals within populations are spatially organized and during the trapping periods they can display move� ment patterns that increase or decrease their capture probability (Gardner et al., 2010b; Sollmann et al., 2011). Relating their movement patterns to the trapping grid therefore has important implications for sampling design, modelling, and interpretation of data (Foster & Harmsen, 2012). The problem with classical closed population models for estimating density from trapping arrays is that 'space' has no explicit manifestation. This dilemma is solved with the recently developed spatially explicit capture–recapture models (SECR�� ), which in� corporate movement by assuming that each individual i has an activity centre si, which remains constant over the survey, and that the capture probability in a trap j is a monotonically decreasing function of the distance between the activity centre and trap j (Borchers & Ef� ford, 2008). Using these models we can also obtain maps of density of activity centres, which correspond to areas where it is more likely that individuals concentrate

Sarmento et al.

their activity. For camera trapping studies, the Poisson model is usually used (Borchers & Efford, 2008). This model assumes that an individual can be captured an arbitrary number of times in an arbitrary number of traps (Borchers & Efford, 2008). Model parameters can include: (1) the baseline trap encounter rate (λ0), which is defined as the expected number of detections of the individual i in a hypothetical trap located at the activity centre (Royle et al., 2009); (2) the movement parameter (σ), which controls the shape of the distance function, being expressed in the same unit used for the trapping grid (km); this parameter can be converted to a 95% home range radius by assuming a circular bivariate normal model for movement; and (3) density (D), is calculated by dividing N by the area of the state space (S), which is user–defined, includes the trapping grid (Borchers & Efford, 2008), and needs to be large enough to contain all individuals potentially exposed to the grid. SECR models can �� be implemented in several soft� ware packages: DENSITY (Efford, 2008), SECR pack� age (Efford, 2011) in program R, version 2.10.1 (R, 2006), and SPACECAP (Gopalaswamy et al., 2012). These platforms differ in their statistical paradigms: (1) likelihood based classical inference (DENSITY and SECR) and (2) Bayesian inference (SPACECAP). Integrated likelihood is an adequate process for analyzing these models, since they are created in terms of a set of latent variables or random effects that corre� spond to individual locations (Borchers & Efford, 2008). The observation model is conceptualized conditionally on the random effects, and inference is formally based on the likelihood created from the marginal probability distribution of the observations (Bochers & Efford, 2008). The random effects are eliminated from the conditional likelihood by integration, which is achieved numerically in SECR models. In this paper we used data collected during three camera–trapping surveys designed for common genets (Genetta genetta) in Serra da Malcata (Portugal) (Sarmento et al., 2010) to compare density estimates produced by SECR models with traditional closed capture models generated by program MARK (White & Burnham, 1999). We tested several models that incorporate habitat and heterogeneity effects and we discuss the major differences and potential advantages of each statistical framework. Common genets are an important forest species, and their density can be used as an indicator of ecosystem fitness, particularly in hu� man altered landscapes and protected areas (Sarmento et al., 2010). The development of suitable methods to evaluate the abundance of this species can therefore be of great importance. Materials and methods Study area The Serra da Malcata (fig. 1) is a Mediterranean moun� tainous area of 200–km2 located in Portugal close to the Spanish border (40º 08' 50'' N – 40º 19' 40'' N and 6º 54' 10'' W – 7º 09' 14'' W). Vegetation is dominated

Animal Biodiversity and Conservation 37.1 (2014)

SM02 SM01

N W

SM03

Portugal

Spain

Camera traps Serra de Malcata Nature Reserve

0 750 1,500 3,000 4,500 6,000 m

Fig. 1. Geographic location of the three trapping campaigns for common genets in Serra da Malcata Nature Reserve, Portugal, 2005–2007. The buffer around the camera–traps corresponds to the 1/2MMDM distance. Fig. 1. Localización geográfica de las tres campañas de trampeo de la jineta común en la reserva natural de Serra da Malcata (Portugal), entre 2005 y 2007. El área de influencia alrededor de las cámaras de trampeo equivale a un 1/2MMDM de la distancia.

by dense scrublands of Cytisus spp., Halimium spp., Cistus spp., Erica spp., Chamaespartium tridentatum and Arbutus unedo covering 43% of the area. Scattered woodlands of Quercus rotundifolia and Q. pyrenaica trees compose 15% of Serra da Malcata. Thirty percent (30%) of the area is covered by plantations of Pinus spp., Eucalyptus globulus and Pseudotsuga menziesii and the remaining 12% is cropland. Around 60% of Serra da Malcata is a protected area included in Serra da Malcata Nature Reserve. Camera–trapping We studied the distribution and abundance of common genets from October 2005 to November 2007. We used four different camera devices: (1) CamTracker® analogical system; (2) DeerCam® analogical system (DeerCam – Scouting Camera, Non Typical Inc., USA); (3) Bushnell trophy cam® digital camera (Bushnell Corp., Overland Park, KS, USA); and (4) GameSpy® digital camera (EBSCO Industries, Inc., East Birming� ham, AL, USA). The cameras were positioned 20 cm (average) above ground and distanced 2 to 4 m from the lure, which consisted of domestic cat urine sprayed on a piece of cork–tree bark connected to a wooden stake at a 40–50 cm height (Sarmento et al., 2009).

Trap–stations were placed in a trapping grid arrange� ment where the distance between cameras should be, at most, the diameter of a circle encompassing the smallest home–range described for the target species in the study area (Sarmento et al., 2010). Therefore, cameras were placed at a distance of 300 to 500 m, according to Cruz (Cruz, 2002). We divided the study into three trapping surveys (table 1), which were divided in an average of seven capture occasions of seven days (Karanth & Nichols, 1998). Genets were individually identified based on their distinct pelage patterns (Sarmento et al., 2010). Spatially explicit models Spatially explicit models combine a state model and an observation model (Borchers & Efford, 2008). The state model expresses the geographic distribution of individual home ranges, while the observation or spatial detection model estimates the probability of detecting an individual at a given detector (e.g. camera trap) to the distance of this detector from a central point in every animal’s home range (Borchers & Efford, 2008). The distribution of range centres in the population can be treated as a homogeneous Poisson point process. We estimated genet density by using the likelihood based classical inference in SECR package.

The SECR package uses SECR models based on the maximum likelihood approach (ML SECR) (Efford, 2011). One important step when using this package is to define the detector type. Since camera traps do not capture animals but simply record their passage we use a detector type called 'proximity'. These 'proximity' detectors can be considered to act independently of each other and may catch more than one animal at a time (Efford et al., 2009). Input data is expressed in two files: (1) one that contains the name and geographic coordinates of the detec� tors (cameras); (2) another containing the capture histories, which include season, animal identification, the occasion, and the detector. Considering that we want to model the detection probability of each individual i on occasion s at detector k, and considering we have observed n individuals on S occasions at K detectors, we will have n*S*K detection probabilities. In this framework a null model assumes that all n*S*K detection probabilities are the same. The usual sources of variation in capture probability can emerge in the n dimension as individual heterogeneity (corresponds to the Mh model), in the S dimension (corresponds to the Mt model of time variation) or as a particular interaction in these two dimensions (as a behavioural response to capture, corresponding to a Mb model). In these models, therefore, the detection probability can have two parameters (e.g. λ0, б for a half–normal function), or three parameters (e.g. λ0, б, z). These parameters may vary with respect to individual (i), occasion (s), or detector (k). Considering these specifications, we defined six models (Efford, 2011): (1) a null model with the notation 'λ0 ~ 1', where λ0 is constant across animals, occasions and detectors; (2) a behaviour model with the notation 'λ0 ~ b', where λ0 is affected by a reaction of the individuals to traps; (3) a model with learned response that affects both λ0 and б with the notation 'list (λ0 ~ b, б ~ b)'; (4) a heterogeneity model with a 2–class finite mixture for λ0 noted as 'λ0 ~ h2'; (5) a time variation model with the notation 'λ0 ~ T'; and (6) a learned response model in λ0 combined with trend over occasions noted as 'λ0 ~ b + T'. Candidate models were ranked using the Akaike Information Criterion corrected for small samples sizes (AICc) by calculating their Akaike weights (Burnham & Anderson, 2002). Models with ΔAICc values ≤ 2 from the most parsimonious model were considered strongly supported. Akaike weights (ω) were used to further interpret the relative importance of each model´s independent variable. ΔAICc values were used to compute ωi, which is the weight of evidence in favour of a model being the best approximating model given the model set (Burnham & Anderson, 2002). Unless a single model had a ωi > 0.9, other models were con� sidered when drawing inferences about the data, by calculating the averaged parameters using ω values (Burnham & Anderson, 2002). Population closure was estimated using the sta� tistical test of Stanley & Richards (2005). Finally, we computed the probability density of home range centres of detected animals (Borchers & Efford, 2008) for the best ranked model by using the function 'fxi'.

Sarmento et al.

Capture–recapture non–spatial model Genet abundance under a non–spatial scenario was estimated using the model 'full closed captures with heterogeneity' available in program MARK (White & Burnham, 1999). These models include a finite mixture as an estimate of individual heterogeneity in capture probability. The finite mixture is character� ized by a parameter π, which is the probability of an individual belonging to mixture a, for one or more mixtures. These models allow capture probability (p) and recapture probability (c) to vary with time or as a behaviour response to traps. Model selection was performed using Akaike’s information criterion (AICc) corrected for small sample size (Burnham & Anderson, 2002) as described above. The effective sampled area was determined using the 1/2MMDM (fig. 1) to generate a buffer around the trapping polygon (Balme et al., 2009a). Density was estimated by dividing N (obtained from the best ranking model) by this area. We evaluated the likelihood of population closure using the Close Test Program (Stanley & Richards, 2005). Each trapping campaign corresponded to 7–day sampling occasions to generate a sufficient number of captures, thereby maximizing the number of sampling occasions without violating population closure assumptions. Results Genet captures From 2005 to 2007, we obtained a capture success of 2.44 captures/100 trap–nights (table 1), which equals one genet capture for every 40.98 nights of trapping. Genets were photographed in 41% of trapping stations (n = 34). On average, we obtained 0.68 (SE = 0.14) captures per trap (min = 0; max = 5). Falsely triggered images constituted 36.19% of all images and were mostly caused by rain, wind and extreme heat. Spatially explicit models For trapping survey SM01 closed test results indicated that the population was closed to gains and losses during the trapping period (z = –0.08; p = 0.46). Only the null model had ΔAICc ≤ 2 with ωi = 0.95 (table 2), indicating no effects of time, behaviour or heterogeneity in the parameters of the models. We estimated a sampling area of 64.99 km2 (table 3). The baseline encounter rate (λ0), estimated at 0.22 (table 3), reached the asymptotic zero at a distance of approximately 2,000 m, indicating that an animal whose activity centre was located at this distance from a given trap had a theoretical capture probability of zero (the probability of being detected in that trap was zero) (fig. 2). Considering the contours of the net probability of detection, we observed that animals with an activity centre within a buffer of 250 m around the trapping polygon had a 0.99 probability of being

Animal Biodiversity and Conservation 37.1 (2014)

Table 1. Camera trapping periods and sampling effort during three trapping campaigns in Serra da Malcata Nature Reserve, Portugal, 2005–2007. Tabla 1. Períodos de trampeo con cámara y esfuerzo de muestro durante tres campañas de muestreo llevadas a cabo en la reserva natural de Serra da Malcata, Portugal, entre 2005 y 2007. Area

Sampling period

Trap stations

Camera–days

Photos

Captures

Individuals

SM01

10 X–6 XII 2005

1,653

SM02

6 X–20 XI 2006

1,350

SM03

14 X–19 XI 2007

828

2,331

Total

caught in any trap (fig. 3). For this model, we esti� mated a genet density of 0.30 (95% CI, 0.61–0.42), corresponding to an average estimated population of 20 individuals (95% CI, 11–27) (table 3). The premise of population closure was observed for trapping campaign SM02 (z = –0.103; p = 0.151). For this campaign we obtained 2 models with a ∆AICc < 2 (table 2). Therefore, no single model emerged as the top ranking model, i.e. ωi > 0.90. The model with the greatest support was SECR–T, followed by SECR–0 (table 2), suggesting that capture probability could be time dependent. The application of a LR test revealed no significant differences between the two models (x12 = 0.94; p = 0.75), so the averaged model was used to calculate the final parameters. An averaged value of 0.13 (CI 95% = 0.06–0.25) was obtained for λ0 (table 3).

The estimated detection probability was 0 at an aver� aged distance of 1,800 m from the traps (fig. 2). Using this model we estimated a genet density of 0.39/km2 (95% CI, 0.19–0.79). corresponding to an estimated population of 23 individuals (95% CI, 11–47) (table 4). We also observed population closure during the SM03 trapping campaign (z = –0.275; p = 0.392). The null model (SECR–0) emerged as the most robust, being the only model with ∆AICc < 2, closely followed by the behaviour model (SECR–b) (table 2). Considering this proximity and the results of an LR test, which revealed no significant differences between the two models (x12 = 0.98; p = 0.75), we used the averaged model to calculate the final parameters. The detection probability decreased consistently with increasing distance from the trapping polygon, presenting an estimated value of 0 at

Table 2. Best two ranked models of a model selection analysis (ΔAIC < 2) of spatially explicit capture– recapture models for genets obtained during three camera–trapping campaigns in Serra da Malcata Nature Reserve, Portugal, 2005–2007. Tabla 2. Los dos modelos mejor clasificados en un análisis de selección (ΔAIC < 2) de modelos de captura y recaptura espacialmente explícitos para la jineta obtenidos durante tres campañas de muestreo con cámaras llevadas a cabo en la reserva natural de Serra da Malcata, Portugal), entre 2005 y 2007. Model

–2Log–likelihood

AICc

∆AICc

AIC wgt

Trapping campaign SM01 SECR–0

287.98

239.90

0.00

0.95

SECR–b

225.42

233.42

6.85

0.03

Trapping campaign SM02 SECR–T

251.38

259.49

0.00

0.49

SECR–0

256.74

262.74

0.11

0.46

Trapping campaign SM03 SECR–0

168.10

178.90

0.00

0.75

SECR–b

163.08

181.09

2.19

0.25

Sarmento et al.

Table 3. Parameter estimates of spatially explicit capture–recapture models for genets obtained during three camera–trapping campaigns in Serra da Malcata Nature Reserve, Portugal, 2005–2007: S area. Effective sampled area (km2); λ0. Baseline encounter rate / occasion; б. Movement parameter (m); D. Genet density (genets / km2); N. Number of genets in the sampled area. Tabla 3. Estimaciones de los parámetros de los modelos de captura y recaptura espacialmente explícitos para la jineta obtenidas durante tres campañas de muestreo con cámaras llevadas a cabo en la reserva natural de Serra da Malcata, Portugal, entre 2005 y 2007: S area. Superficie efectiva muestreada (km2); λ0. Índice de referencia de encuentros / ocasión; б. Parámetro de movimiento (m); D. Densidad de la jineta (jinetas / km2); N. Número de jinetas en la zona muestreada.

Campaign SM01

Model

S area

SECR–0

64.99

SM02

SECR–T

58.76

SECR–0

Model

averaged

SM03

SECR–0

81.12

SECR–b

Model

averaged

λ0

0.22

759

0.30

(0.12–0.36)

(540–1066)

(0.16–0.42)

(11–27)

0.17

573

0.38

(0.07–0.34)

(392–837)

(0.18–0.76)

(11–45)

0.08

567

0.41

(0.04–0.16)

(420–1667)

(0.20–0.83)

(12–49)

0.13

570

0.39

(0.06–0.25)

(405–1239)

(0.19–0.79)

(11–47)

0.16

737

0.16

(0.07–0.34)

(501–1084)

(0.07–0.32)

(11–45)

0.49

742

0.19

(0.12–0.87)

(503–1093)

(0.09–0.42)

(8–34)

0.24

738

0.18

(0.14–0.46)

(502–1086)

(0.08–0.35)

(11–42)

an averaged distance of 2,000 m (fig. 2). An averaged density of genets was estimated at 0.18 individuals/km2 (95% CI, 0.19–0.79), corresponding to an estimated population of 21 individuals (95% CI, 11–42) (table 4). Closed capture–recapture models For all the trapping campaigns in the best–ranked models, p and c did not vary with time and were different (i.e. they showed a behaviour response to traps) (table 4). The recapture probability was always lower than p, which reflects a negative effect of the traps after the first capture. Average capture proba� bility was estimated at 0.61 (95% CI = 0.37–0.78), and the average recapture probability was 0.40 (95% CI = 0.26–0.52) (table 5). Density values varied bet� ween 0.45 and 0.73 individual/km2, with an average of 0.61 (95% CI = 0.58–0.67) (table 5). Discussion Over the last 10 years, camera trapping has become one of the most useful tools to estimate animal abun� dance, particularly in species that can be individually identified (Balme et al., 2009b; Negrões et al., 2010;

Sarmento et al., 2010). The most common approach is to combine this technique with standard closed population models. The theoretical constraint of closed population estimators is that, although abundance estimates may be suitable to calculate the numbers of a population that are exposed to traps, individual movements cannot be precisely associated to an accurate area (Royle et al., 2011b). At the same time, closed population models cannot incorporate moving traps, open systems, or multiple captures in a single occasion. The recent development of SECR models was crucial to deal with the baseline problem of abundance interpretation resulting from the ad–hoc approaches to estimate the sampling area in non–spatial capture– recapture models (Borchers & Efford, 2008). Using the classic estimation of abundance, it is difficult to compare different areas. However, the flexibility of SECR�� models allows this comparison and also in� cludes other aspects such as capture heterogeneity and covariate effects on capture probabilities (Borch� ers & Efford, 2008). The inclusion of trap–specific encounter histories in SECR models overcomes the problem of non–spatial models ignoring trap identity. In these models, if an animal is captured multiple times during a trapping period, these will count as

Animal Biodiversity and Conservation 37.1 (2014)

0.20 0.15 0.10 0.05 0.00 0

SM02

0.08 0.06 0.04 0.02

0.00 500 1,000 1,500 2,000 2,500 3,000 0 1,000 Distance (m)

0.20 Detection probability

0.10

SM01 Detection probability

Detection probability

0.25

2,000 3,000 Distance (m)

4,000 5,000

SM03

0.15

0.10

0.05

0.00 0

1,000

2,000 3,000 4,000 Distance (m)

5,000

Fig. 2. Variation of the detection probability as a function of distance from an individual trap for three trapping surveys for common genets in Serra da Malcata Nature Reserve, Portugal, 2005–2007. Fig. 2. Variación de la probabilidad de detección en función de la distancia desde una cámara determinada para tres estudios de trampeo de la jineta común en la reserva natural de Serra da Malcata, Portugal, entre 2005 y 2007.

a single capture only, possibly leading to oss of information. Furthermore, in non–spatial models it is difficult to include different periods of individual trap activity (i.e. traps that were not always active during the entire trapping period) (Efford et al., 2013), which is not necessary in SECR models because they are based on trap–level encounters of individuals (Royle et al., 2011b). SECR models assume the demographic closure of the population (i.e. no births, deaths, emigration or immigration during the study period). This population closure is assumed in the fixed nature of the estimated activity centres, which are considered to be constant over the trapping period. In this case, the presence of transient animals can be a factor of

population closure violation. According to Royle et al. (2011b)�� this non–closure can be overcome by model� ling an individual–specific encounter probability scale parameter, σ, that incorporates individual variability in home–range size. More extensions to these models are currently being developed to include moving activ� ity centres. Such centres will be crucial in multi–year studies, since home–ranges can vary with changes in resource availability and in other biological aspects (Royle et al., 2009). The density estimates obtained using the non–spa� tial model were, on average, 2.15 times higher than those obtained using the spatial model. This findings was also observed by other authors who performed the same type of comparison with Andean cats (Leopardus

Sarmento et al.

Probability curves of activity centers

Probability of detection 365000

0.10 capture probability curve

368000 364000 366000

363000

364000

362000

361000

360000

362000

SM01

360000

288000 289000 290000 291000 292000 293000

302000

288000

290000

292000

294000

296000

298000

370000

300000

368000

370000

366000

358000

364000

366000

362000

SM02 294000

SM02

296000

298000

300000

302000 364000

364000 362000

362000

360000

358000

356000

354000

356000

SM03

SM03 285000

290000

295000

284000

288000

292000

296000

Probability of detection 0.90

0.80

0.70

0.60

0.50

0.40

0.30

0.20

0.10

Fig. 3. Geographic distribution of the probability of detection around the trapping polygon for each potential home–range center (left column) and curves of probability of distribution of home range centers of animals that were detected for the three trapping campaigns (right column) for common genets in Serra da Malcata Nature Reserve, Portugal, 2005–2007. Fig. 3. Distribución geográfica de la probabilidad de detección alrededor del polígono de trampeo para cada posible centro de área de distribución (columna izquierda) y curvas de probabilidad de la distribución de dichos centros para los animales que se detectaron en las tres campañas de trampeo (columna derecha) de la jineta común en la reserva natural de Serra da Malcata, Portugal, entre 2005 y 2007.

Animal Biodiversity and Conservation 37.1 (2014)

Table 4. Model selection statistics for the two best ranking models of full closed captures analysis on common genet capture–recapture data from Malcata, Portugal, 2005–2007: Np. Number of parameters; Dv. Deviance; pi. Probability of mixture; p. Capture probability; c. Recapture probability; N. Population size. The parameters p and c were modeled such as could be constant over sampling occasions (.) or varying with time (t). Tabla 4. Valores de los criterios estadísticos para seleccionar modelos relativos a los dos mejores modelos de análisis de capturas en poblaciones totalmente cerradas aplicados a los datos obtenidos con la captura y recaptura de la jineta común en Malcata, Portugal, entre 2005 y 2007: Np. Número de parámetros; Dv. Desviación; pi. Probabilidad de mezcla; p. Probabilidad de captura; c. Probabilidad de recaptura; N. Tamaño de la población. Los parámetros p y c se utilizaron en el modelo como si pudieran ser constantes durante todas las ocasiones de muestreo (.) o variables con el tiempo (t). Model AICc ∆AICc

AICc Weights

Model Likelihood

SM01 {pi(.) p(.) c(.) N(.)}

76.35

0.00

0.87

1.00

58.23

{pi(.) p(t) c(.) N(.)}

80.95

4.59

0.09

0.10

56.09

SM02 {pi(.) p(.) c(.) N(.)}

38.02

0.00

0.85

1.00

20.70

{pi.p(.) = c(.).N}

42.04

4.01

0.11

0.13

24.71

SM03 {pi(.) p(.) c(.) N(.)}

37.17

0.00

0.87

1.00

24.48

{pi.p(t) = c(t).N}

41.75

4.59

0.09

0.10

16.32

jacobita) (Reppucci et al., 2011) and jaguars (Panthera onca) (Sollmann et al., 2011). In conclusion, the use of SECR models overcomes several problems that can arise when estimating

density of genets or other cryptic, low–density spe� cies (Sollmann et al., 2011). One of the main advan� tages of these models is that they can be applied to any capture–recapture technique that is based on

Table 5. Results of population closure, probability of mixture (π), capture (p) and recapture (c) probabilities and estimated abundance (N) and density (individuals / km2 [D]) of genet samples in Serra da Malcata Nature Reserve, Portugal, 2005–2007, using the best ranked model for each trapping campaign of table 4. Tabla 5. Resultados del grado en qué una población pueda ser considerada cerrada, la probabilidad de mezcla (π), las probabilidades de captura (p) y de recaptura (c) y la abundancia (N) y densidad (individuos / km2 [D]) estimadas a partir de las poblaciones de jineta estudiadas en la reserva natural de Serra da Malcata, Portugal, entre 2005 y 2007, utilizando el modelo mejor clasificado para cada campaña de trampeo de la tabla 4.

π p c N D SM01 SM02 SM03

0.47

0.60

0.42

0.59

(0.44–0.50)

(0.34–0.80)

(0.30–0.55)

(9–11)

(0.53–0.61)

0.47

0.63

0.32

0.74

(0.30–0.63)

(0.37–0.82)

(0.18–0.49)

(10–12)

(0.67–0.80)

0.48

0.75

0.42

0.46

(0.31–65)

(0.45–0.91)

(0.27–0.58)

(9–10)

(0.45–0.50)

1/2MMDM area (km2) 16.88 14.96 19.74

individual identification and trap–specific encounter histories (Royle et al., 2011b). In addition, because SECR�� models can integrate covariates that can influ� ence capture probabilities, camera–trapping sampling designs can be modified to improve capture success. One of the main constraints of SECR models is the use of encounter probabilities based on the Eu� clidean distance between traps and animal activity centres, thus presuming that home ranges are fixed and symmetric. Home ranges are therefore nchanged by landscape or habitat composition. If we apply these models in areas with significant geographic barriers, the results could be potentially biased. The probability of capturing an animal in a trap placed on the opposite side of the barrier would basically be a function of distance, while in reality the probability of capture should consider both distance and barrier permeability. These scenarios are very common in CR studies considering that most landscapes are hetero� geneous and animals tend to use linear features such as trails, corridors, or rivers. Future research should therefore consider the use of the models developed by Royle et al. (2013) that include explicit hypotheses on the effects of environmental variables on distance metrics. These hypotheses should then be directly incorporated into SECR models, so that they may be evaluated statistically. Their accuracy could be further increased by integrating other components, in the model, such as the complex relationships between movement, habitat heterogeneity, and drivers such as energetic costs, conspecific competition, social status and prey availability. SECR models could be greatly improved by incorporating data on habitat use, and social and population dynamics. By integrating movement into existing SECR methods, it will be pos� sible to study the effects of environmentally coupled movement on estimates of abundance and density (Rowcliffe et al., 2012). References Balme, G. A., Hunter, L. T. B. & Slotow, R. O. B., 2009a. Evaluating Methods for Counting Cryptic Carnivores. The Journal of Wildlife Management, 73: 433–441. Balme, G. A., Slotow, R. & Hunter, L. T. B., 2009b. Im� pact of conservation interventions on the dynamics and persistence of a persecuted leopard (Panthera pardus) population. Biological Conservation, 142: 2681–2690. Borchers, D. L. & Efford, M. G., 2008. Spatially Explicit Maximum Likelihood Methods for Capture–Recap� ture Studies. Biometrics, 64: 377–385. Burnham, K. P. & Anderson, D. R., 2002. Model selection and multimodel inference: a practical information–theoretic approach. Springer–Verlag, New York, USA. Cruz, J., 2002. Resource use and spatial organiza� tion of the genet (Genetta genetta). Ph. D. Thesis, Biology Department, Coimbra University. Efford, M. G., 2008. Program DENSITY. Software for spatially explicit capture–recapture. <http://www.

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otago.ac.nz/density/> – 2011. SECR: spatially explicit capture–recapture models. R package version 2.1.0. <hhttp://cran.r– project.org/i> Efford, M. G., Borchers, D. L. & Mowat, G., 2013. Varying effort in capture–recapture studies. Methods in Ecology and Evolution, 4: 629–636. Efford, M. G., Dawson, D. K. & Borchers, D. L., 2009. Population density estimated from locations of individuals on a passive detector array. Ecology, 90: 2676–2682. Foster, R. & Harmsen, B., 2012. A critique of density estimation from camera–trap data. Journal of Wildlife Management, 76: 224–236. Gardner, B., Reppucci, J., Lucherini, M. & Royle, J. A., 2010a. Spatially explicit inference for open populations: estimating demographic parameters from camera–trap studies. Ecology, 91: 3376–3383. Gardner, B., Royle, J. A., Wegan, M. T., Rainbolt, R. E. & Curtis, P. D., 2010b. Estimating Black Bear Density Using DNA Data From Hair Snares. The Journal of Wildlife Management, 74: 318–325. Gopalaswamy, A. M., Royle, J. A., Hines, J. E., Singh, P., Jathanna, D., Kumar, N. S. & Karanth, U., 2012. Program SPACECAP: software to estimate animal density using spatially explicit capture–recapture models. Methods in Ecology and Evolution, 3(6): 1067–1072. Karanth, K. U. & Nichols, J. D., 1998. Estimating tiger (Panthera tigris) populations form camera– trap data using capture–recaptures. Ecology, 79: 2852–2862. Negrões, N., Sarmento, P., Cruz, J., Eira, C., Revilla, E., Fonseca, C., Sollmann, R., Torres, N. M., Fur� tado, M. M., Jácomo, A. T. A. & Silveira, L., 2010. Use of Camera–Trapping to Estimate Puma Den� sity and Influencing Factors in Central Brazil. The Journal of Wildlife Management, 74(6): 1195–1203. R Development Core Team (RDCT), 2006. A language and environment for statistical computing. Vienna, Austria. ISBN 3–900051–07–0. <http:// www.R–project.org> Reppucci, J., Gardner, B. & Lucherini, M., 2011. Estimating detection and density of the Andean cat in the high Andes. Journal of Mammalogy, 92: 140–147. Rowcliffe, J. M., Carbone, C., Kays, R., Kranstauber, B. & Jansen, P. A., 2012. Bias in estimating animal travel distance: the effect of sampling frequency. Methods in Ecology and Evolution, 3: 653–662. Royle, J. A., Chandler, R. B., Sun, C. C. & Fuller, A. K., 2013. Integrating resource selection information with spatial capture–recapture. Methods in Ecology and Evolution, 4: 520–530. Royle, J. A., Kéry, M. & Guélat, J., 2011a. Spatial cap� ture–recapture models for search–encounter data. Methods in Ecology and Evolution, 2: 602–611. Royle, J. A., Magoun, A. J., Gardner, B., Valkenburg, P. & Lowell, R. E., 2011b. Density estimation in a wolverine population using spatial capture–recap� ture models. The Journal of Wildlife Management, 75: 604–611. Royle, J. A., Nichols, J. D., Karanth, K. U. & Gopa�

Animal Biodiversity and Conservation 37.1 (2014)

laswamy, A. M., 2009. A hierarchical model for estimating density in camera–trap studies. Journal of Applied Ecology, 46: 118–127. Sarmento, P., Cruz, J., Eira, C. & Fonseca, C., 2009. Evaluation of Camera Trapping for Estimating Red Fox Abundance. Journal of Wildlife Management, 73: 1207–1212. – 2010. Habitat selection and abundance of common genets Genetta genetta using camera capture– mark–recapture data. European Journal of Wildlife Research, 56: 59–66. Sollmann, R., Furtado, M. M., Gardner, B., Hofer, H., Jácomo, A. T. A., Tôrres, N. M. & Silveira, L., 2011. Improving density estimates for elusive carnivores: Accounting for sex–specific detection and move� ments using spatial capture–recapture models for

jaguars in central Brazil. Biological Conservation, 144: 1017–1024. Stanley, T. R. & Richards, J. D., 2005. Software Re� view: A program for testing capture–recapture data for closure. Wildlife Society Bulletin, 33: 782–785. Trolle, M., Noss, A., Lima, E. & Dalponte, J., 2007. Camera–trap studies of maned wolf density in the Cerrado and the Pantanal of Brazil. Biodiversity and Conservation, 16: 1197–1204. White, G. C., Anderson, D. R., Burnham, K. P. & Otis, D. L., 1982. Capture–recapture and Removal Methods for Sampling Closed Populations. National Laboratory Publications, LA–8778–NERP. White, G. C. & Burnham, K. P., 1999. Program MARK: survival estimation from populations of marked animals. Bird Study, 46: 120–138.

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Animal Biodiversity and Conservation 37.1 (2014)

Assessing the response of exploited marine populations in a context of rapid climate change: the case of blackspot seabream from the Strait of Gibraltar J. C. Báez, D. Macías, M. de Castro, M. Gómez–Gesteira, L. Gimeno & R. Real Báez, J. C., Macías, D., De Castro, M., Gómez–Gesteira, M., Gimeno, L. & Real. R., 2014. Assessing the response of exploited marine populations in a context of rapid climate change: the case of blackspot seabream from the Strait of Gibraltar. Animal Biodiversity and Conservation, 37.1: 35–47. Abstract Assessing the response of exploited marine populations in a context of rapid climate change: the case of blackspot seabream from the Strait of Gibraltar.— There is a growing concern over the decline of fisheries and the possibility of the decline becoming worse due to climate change. Studies on small–scale fisheries could help to improve our understanding of the effect of climate on the ecology of exploited stocks. The Strait of Gibraltar is an important fishery ground for artisanal fleets. In this area, blackspot seabream (Pagellus bogaraveo) is the main species targeted by artisanal fisheries in view of its relevance in landed weight. The aims of this study were to explore the possible effects of two atmospheric oscillations, the North Atlantic Oscillation (NAO) and the Arctic Oscillation (AO), on the capture of blackspot seabream in the Strait of Gibraltar, to determine their association with oceanographic conditions, and to improve our knowledge about the possible effects of climate change on fisheries ecology so that fishery management can be improved. We used two types of data from different sources: (i) landings per unit of effort reported from a second working group between Morocco and Spain on Pagellus bogaraveo in the Gibraltar Strait area, for �� the period 1983–2011�� , and (ii) the recorded blackspot seabream land� ings obtained from the annual fisheries statistics published by the Junta de Andalucía (Andalusian Regional Government). Our results indicate that the long–term landing of blackspot seabream in the Strait of Gibraltar is closely associated with atmospheric oscillations. Thus, prolonged periods of positive trends in the NAO and AO could favour high fishery yields. In contrast, negative trends in NAO and AO could drastically reduce yield. Key words: Arctic Oscillation, Blackspot seabream, Climate, Fisheries collapse, North Atlantic Oscillation, Oceanography. Resumen Evaluación de la respuesta de las poblaciones marinas explotadas en un contexto de cambio climático rápido: el caso del besugo de la pinta en el estrecho de Gibraltar.— Existe una creciente preocupación por la disminu� ción de la pesca y la posibilidad de que esta disminución se acelere debido al cambio climático. Los estudios sobre la pesca a pequeña escala podrían ayudar a mejorar nuestra comprensión de los efectos del clima en la ecología de las poblaciones explotadas. El estrecho de Gibraltar es una importante zona de pesca para la flota artesanal. En esta zona, el besugo de la pinta (Pagellus bogaraveo) es la especie más importante para la pesca artesanal en vista de su volumen de descarga. Los objetivos de este estudio consisten en estudiar los posibles efectos de dos oscilaciones atmosféricas: la oscilación del Atlántico Norte (NAO) y la oscilación del Ártico (AO), en la captura del besugo de la pinta en el estrecho de Gibraltar con objeto de determinar su relación con las condiciones oceanográficas, y mejorar nuestro conocimiento sobre los posibles efectos del cambio climático en la ecología de la pesca, para poder mejorar la gestión de la actividad pesquera. Utiliza� mos dos tipos de datos de diferentes fuentes: (i) los desembarques por unidad de esfuerzo registrados por un segundo grupo de trabajo entre Marruecos y España sobre el besugo de la pinta en la zona del estrecho de Gibraltar, para el período 1983–2011, y (ii) los desembarques registrados de besugo de la pinta obtenidos de las estadísticas anuales de pesca publicadas por la Junta de Andalucía. Nuestros resultados indican que el desembarque a largo plazo del besugo de la pinta en el estrecho de Gibraltar está íntimamente relacionado con las oscilaciones atmosféricas. Por lo tanto, los períodos prolongados de tendencias positivas en la NAO y ISSN: 1578–665 X eISSN: 2014–928 X

Báez et al.

la AO podrían favorecer altos rendimientos pesqueros. En contraste, las tendencias negativas de la NAO y la AO reducen drásticamente el rendimiento pesquero. Palabras clave: Oscilación del Ártico, Besugo de la pinta, Clima, Colapso pesquero, Oscilación del Atlántico Norte, Oceanografía. Received: 14 I 14; Conditional acceptance: 11 III 14; Final acceptance: 1 IV 14 José Carlos Báez, Inst. Español de Oceanografia (IEO), Centro Oceanográfico de Málaga, Puerto pesquero de Fuengirola s/n., 29640 Fuengirola, Málaga, España (Spain); investigador asociado de la Fac. de Ciencias de la Salud, Univ. Autónoma de Chile, Chile.– David Macías, Inst. Español de Oceanografia (IEO), Centro Oceanográfico de Málaga, Puerto pesquero de Fuengirola s/n., 29640 Fuengirola, Málaga, España (Spain).– Maite de Castro, Moncho Gómez–Gesteira & Luis Gimeno, Ephyslab, Fac. de Ciencias de Ourense, Univ. de Vigo, Ourense, España (Spain).– Raimundo Real, Depto. de Biología Animal, Fac. de Ciencias, Univ. de Málaga, 29071 Málaga, España (Spain). *Corresponding author: J. C. Báez. E–mail: granbaez_29@hotmail.com

Animal Biodiversity and Conservation 37.1 (2014)

Introduction Fisheries are an important source of food and income for many local communities, and their value as a source of animal protein was recently emphasized in a Food and Agriculture Organization report (FAO, 2010). Sev� eral studies (e.g. Thurstan et al., 2010) have suggested that over the last decade, 88% of monitored marine fish stocks in EU waters have been overfished, and some authors have predicted a global collapse of fish� eries within the next few decades (Worm et al., 2006, 2009). The observed decline in fisheries is mainly due to overfishing at an industrial scale (Worm & Myers, 2004; Pitcher, 2005). However, this situation could be aggravated by the response of fish populations to cli� mate change (e.g. see Brandt & Kronbak, 2010). Thus, with the aim of integrating fisheries within sustainable ecosystems, Pitcher (2005) proposed studying the ef� fect of climate parameters and their temporal variability on global fisheries. Some fisheries have been shown to respond to multi–decadal oscillations, such as the oscillation of El Viejo (The Old Man), or La Vieja (The Old Woman), in the Pacific (Chavez et al., 2003), and decadal oscillations, such as the North Atlantic Oscilla� tion (Báez et al., 2011; Báez & Real, 2011). The North Atlantic Oscillation (NAO) is a dominant pattern of coupled ocean–climate variability in the North Atlantic and Mediterranean basin (Hurrell, 1995). Many authors have observed a relationship between the NAO and changes in fishery abundance (Graham & Harrod, 2009; Báez et al., 2011; Báez & Real, 2011) and recruitment (Fromentin, 2001; Borja & Santiago, 2002; Mejuto, 2003). The NAO reflects fluctuations in atmospheric pressure at sea–level between the Icelandic Low and the Azores High. The NAO is associated with many meteorological variations in the North Atlantic region, affecting wind speed and direction and differences in temperature and rainfall (Hurrell, 1995). Recent studies (e.g. Overland et al., 2010) have discussed the effect of large–scale climate variability on several marine ecosystems and suggest that marine ecosystems could respond to climate change. Straile & Stenseth (2007) have suggested that the NAO can be used to explain inter–annual variability in ecological series, citing the following reasons: (1) a strong relationship between the NAO and weather conditions during the winter season; (2) qualitative changes in environmental conditions in response to winter weather conditions, especially temperature; and (3) the great importance of these environmental conditions in the distribution and population dynamics of species in temperate and boreal regions. Nevertheless, the dominant mode of variability in atmospheric circulation variability in the Northern Hemisphere is determined by the Arctic Oscillation (AO). The AO is characterized by a meridional di� pole in atmospheric sea level pressure between the northern Polar Regions and mid–latitudes (Thompson & Wallace, 1998). The NAO and AO are closely cor� related (Thompson et al., 2000). The AO has been attributed to stratosphere–troposphere coupling. Ac� cording to Thompson et al. (2000), this includes the NAO, which may be considered a different view of the

same phenomenon. Thus, the AO and the NAO both tend to be in a positive phase during winters when the stratospheric vortex is strong (Douville, 2009). Few studies have analyzed the possible effect of the AO on fisheries ecology, for example Gancedo (2005) and Yatsu et al. (2005). The possible effects of global climate change on the ecology of exploited stocks are difficult to study due to the multitude of other factors affecting these stocks, such as overfishing, coastal development, and pollution. Regional studies focused on small–scale fisheries could help to understand the effect of global climate change on the ecology of exploited stocks due to the reduction of the number of other variables (e.g. Meynecke et al., 2012; Pranovi et al., 2013). The Strait of Gibraltar connects the western Medi� terranean Sea with the Atlantic Ocean, providing an important fisheries ground for artisanal fleet (Silva et al., 2002). Because of the high frequency of maritime traffic in the Strait of Gibraltar, the largest Spanish fishing boats do not operate in this area; thus, the fishery is carried out by small numbers of artisanal boats working near the coast (Báez et al., 2009). Ac� cording to Báez et al. (2013b), the physical condition of bluefin tuna (Thunnus thynnus) caught in this area, is correlate with both NAO and AO. Blackspot seabream (Pagellus bogaraveo) is the most important species targeted by the artisanal fish� eries, according to their importance in landed weight (Silva et al., 2002). In this context, the fishery landings and distribution by class of boat are easy to control at small–scale fisheries. The blackspot seabream is a typical small de� mersal fish distributed from Eastern Atlantic Ocean to Western Mediterranean Sea, extensively fished from the early 80s by the artisanal fleet home–base in Gibraltar Strait. Fleets of Algeciras and Tarifa fished the blackspot seabream exclusively using a vertical deep water longline called 'voracera' baited with small sardines (Sardina pilchardus), while the artisanal fleet from Conil used a traditional bottom longline in the western part of the Strait of Gibraltar (for a detailed description of the fishery see Czerwinski et al., 2009; Gil–Herrera, 2010, 2012). The aim of this study was to assess the responses of exploited marine populations in a context of rapid climate and oceanographic change using the landing of blackspot seabream in the Strait of Gibraltar as study case. Material and methods Fisheries data The study area coincides with the fishing ground, an area within Spanish waters of the Strait of Gibraltar between the Rock of Gibraltar and Cape Trafalgar, and it included the landing harbours of Algeciras, Tarifa and Conil (fig. 1). Data were collected from two different sources. First, in the period 1983–2011, we used landings per sale, reported in CopeMed II (2012) and Gil–Herrera

(2012) on Pagellus bogaraveo in the Gibraltar Strait area, as landings per unit of effort (LPUE), because each sale is equivalent to the trip per boat (which is typically the fisheries effort). The artisanal Moroccan fleet also fished blackspot seabream in the Strait of Gibraltar. However, we excluded these data because the data available from Moroccan fleet is a short–time series (CopeMed II, 2012). Second, we used the recorded blackspot seabream landings obtained from the annual fisheries statistics published by the Junta de Andalucía (Andalusian Regional Government) (Galisteo et al., 2001a, 2001b, 2002, 2004, 2005; Alonso–Pozas et al., 2007; Gali� steo et al., 2007, 2008, 2009a, 2009b, 2011, 2012, 2013) for the period 1985–2012 from Tarifa, the most important landing harbour in the study area (table 1). Atmospheric data Monthly values of the NAO index and AO index were taken from the website of the National Oceanic and Atmospheric Administration: http://www.cpc.noaa. gov/products/precip/CWlink/pna/nao_index.html and fttp://www.esrl.noaa.gov/psd/data/correlation/ao.data, respectively. The atmospheric oscillations present strong inter– annual and intra–annual variability (Hurrell, 1995). However, several studies have shown that changes in NAO/AO trends have a delayed effect on aquatic ecosystems due to ecosystem inertia (Maynou, 2008; Báez et al., 2011). For this reason, we used NAO and NAO in the previous year (NAOpy); and AO and AO in the previous year (AOpy). Oceanographic data Ocean temperature and salinity data were obtained using the Simple Ocean Data Assimilation (SODA) package (http://www.atmos.umd.edu/~ocean). SODA uses an ocean model based on Geophysical Fluid Dynamics Laboratory MOM2 physics. Assimilated data include temperature and salinity profiles from the World Ocean Atlas–94 (Levitus & Boyer, 1994), as well as additional hydrography, sea surface tempe� ratures (Reynolds & Smith, 1994), and altimeter sea levels obtained from the Geosat, ERS–1, and TOPEX/ Poseidon satellites. Re–analyses of world ocean climate variability are available from 1958 to 2007 at a monthly scale, with a horizontal spatial resolution of 0.5º × 0.5º and a vertical resolution of 40 levels (Carton et al., 2000a, 2000b; Carton & Giese, 2008). According to previous research, the Mediterranean water mass is produced by the transformation of fresh and warm surface Atlantic water (AW) that enters in the Mediterranean Sea by the Strait of Gibraltar. The surface AW is gradually modified during its displace� ment eastward in the Mediterranean Sea due to air–sea interactions and mixing processes. A portion of these dense water masses flows back (after seven to 70 years) through the Strait of Gibraltar, mixing with Eastern North Atlantic Central Water (ENACW) to form the Mediterranean Outflow Water (MOW; Bozec et al., 2011). In the eastern Gulf of Cadiz, the MOW is

Báez et al.

very dense and sinks under water with an Atlantic origin until it reaches an equilibrium level (around 1,100 m). In the western Gulf of Cadiz (8º W), MOW reaches density values similar to those of mid–depth Atlantic layers and splits in two cores separating from the bottom. The upper core is characterized by a maxima temperature (~13ºC) and a potential density anomaly between 27.40 and 27.65 kg/m3, and the lower core is characterized by a maxima salinity (~37.5) and a potential density anomaly between 27.70 and 27.85 kg/m3. MOW spreads in the North Atlantic westward to the central Atlantic and northward along the coasts of Portugal and the Iberian peninsula. For this reason, the oceanographic analysis was carried out in a region large enough to contain the number of measurement points needed for a suitable oceanographic study of the MW tak� ing into account ocean currents, coastal areas, and water properties. In the present study, the selected area ranged from 8º to 13.75º W and from 35.25º to 40.25º N (the southern coast of the Iberian peninsula). We used the first 24 vertical levels (which correspond to a water depth of 1,378 m) since the study focuses on the detection of upper Mediterranean water (MWu hereafter), whose core is located at 800 m. The thick� ness of the vertical layers increases from 10 m near the surface to 100 m below 300 m. The period under study ranges from 1980 to 2007. We identified MWu using temperature, salinity and density values, which should lie within the intervals 10.5–13.5ºC, 35.8–36.8, and 27.4–27.65 kg/m3, respec� tively. First, the grid points where MWu was not detected in at least 50% of the samples were discarded from the analysis. Salinity and temperature data for each grid point were averaged to transform them into annual values. All salinity and temperature data corresponding to the intervals mentioned above for a specific year were averaged, regardless of layer, to obtain the mean MWu salinity and temperature values for that year. Long–term processes, such as warming–cooling or salinification–freshening, and their effect on the water column stratification were analyzed using annual trends, which were assumed to be linear. All trends were calcu� lated using raw data, without using any filter or running mean. The Spearman rank correlation coefficient was used to analyze the significance of trends due to its robustness to deviations from linearity and its resistance to the influence of outliers (Saunders & Lea, 2008). Data analysis In a first step, we analysed the time series for each variable. We searched for common time trends and cyclicity in the time series using spectral analysis, to identify periodicity. Spectral analysis was performed with the software PAST (available from web site: http:// folk.uio.no/ohammer/past/) (Hammer et al., 2001; Hammer & Harper, 2006). We tested the relationship between LPUE of blacks� pot seabream versus NAOpy and AOpy using linear multiple regressions. We selected the best fit among several significant regressions when different degrees of freedom were involved in accordance with the high�

Animal Biodiversity and Conservation 37.1 (2014)

–6º 20' –6º 00' –5º 40' –5º 20' –5º 00' –4º 40' 36º 40'

36º 20'

Gulf of Cadiz

36º 00'

Strait of Gibraltar Alboran Sea

35º 40'

35º 20' GMT

2011 Oct 18 04:18:39

www.seaturtle.org/maptool/

Projection: Mercator

Fig. 1. The study area was centred on the Strait of Gibraltar. The Strait of Gibraltar separates two regions: the Gulf of Cadiz (in the Atlantic Ocean) and the Alboran Sea (within the Mediterranean Sea). Fig. 1. La zona del estudio tiene en su centro el estrecho de Gibraltar. El estrecho de Gibraltar separa dos regiones: el golfo de Cádiz (en el océano Atlántico) y el mar de Alborán (en el mar Mediterráneo).

est R2 value. Normality of the data was tested using the Kolmogorov–Smirnov test (Sokal & Rohlf, 1995). A probabilistic analysis was performed by taking a year at random and calculating the probability that this particular annual landing from Tarifa was higher or lower than the average landing for all the years avail� able pooled together. Báez et al. (2011) used binary logistic regression to model the response of albacore fisheries to changes in the accumulated NAO index. Similarly, using binary logistic regression, we modelled the probability of the value for blackspot seabream landings being higher than the average landings for this species for each specific year. Thus, we assigned a value of 1 or 0, respectively, when the landing in a specific year was higher or lower than the average landing for the 26 years taken together; these were considered to be good and poor landings, respectively. We performed a forward stepwise logistic regression where the independent variables were NAO, NAOpy, AO and AOpy. The goodness–of–fit of the model was assessed using an omnibus test (for model coefficients) and a Hosmer and Lemeshow test, which also follows a Chi–square distribution (Zuur et al., 2007), with the low p–values indicating a lack of fit of the model. We evaluated the discrimination capacity of our model using the area under the receiving operating characteristic (ROC) curve (AUC) (Lobo et al., 2008). Despite a good fit of the logistic regression model, it is sensitive to the presence/absence ratio (Real et al., 2006). The presence/absence ratio was 0.625 for blackspot seabream. To resolve this difference,

we applied the favourability function (Ff) (Real et al., 2006) based on a logistic regression model, which adjusts the model regardless of the presence/ab� sence ratio. Favourability was easily calculated from the probability obtained from the logistic regression according to the expression: Ff = [P / (1 – P)] / [(n1 / n0) + (P / [1 – P])] where P is the probability of the value for blackspot seabream landings per a specific year was higher than the average landings for this species for all years, and n1 and n0 are number of years with good or poor blackspot seabream landings, respectively. The correlation between the different climatic indices and landings can be also analyzed in terms of the accumulated values. Annual values were transformed into anomalies by subtracting the mean value calculated over the whole period 1985–2010. The accumulated variables corresponding to a specific year were then calculated as the sum of the anomalies of the previous years (e.g. the accumulated values corresponding to 2000 were calculated as the sum of the anomalies for the period 1985–2000). Results The landing of blackspot seabream from Tarifa for 1985–2011 was the only variable with significant periodicity trend (table 2).

Báez et al.

Table 1. Blackspot seabream (Pagellus bogaraveo) landing per year and corresponding average for the North Atlantic Oscillation (NAOpy) and Arctic Oscillation (AOpy) index in the year before the landing. Tabla 1. Desembarque por año del besugo de la pinta (Pagellus bogaraveo) y promedio correspondiente para los índices de oscilación del Atlántico Norte (NAO) y de oscilación del Ártico (AO) en el año anterior al desembarque. Year

Landing

NAOpy

AOpy

1985

209866

0.25

–0.19

1986

249000

–0.18

–0.52

1987

292732

0.5

0.08

1988

318578

–0.12

–0.54

1989

413375

–0.01

0.04

1990

426400

0.7

0.95

1991

421070

0.59

1.02

1992

629668

0.27

0.2

1993

764522

0.58

0.44

1994

854436

0.18

0.08

1995

501569

0.58

0.53

1996

659485

–0.08

–0.27

1997

527186

–0.21

–0.46

1998

282522

–0.16

–0.04

1999

198794

–0.48

–0.27

2000

193408

0.39

0.11

2001

154832

0.21

–0.05

2002

147793.6

–0.18

–0.16

2003

179146.5

0.04

0.07

2004

131692.6

0.1

0.15

2005

165616.8

0.24

–0.19

2006

161772.5

–0.27

–0.38

2007

273035

–0.21

0.14

2008

285481

0.17

0.27

2009

424849.4

–0.38

0.18

2010

227391

–0.24

–0.33

A significant association was found between the LPUE of blackspot seabream from the Gulf of Cadiz and the NAOpy index, according to the following function (fig. 2): LPUE = 66.687 + 15.01 NAOpy (adjusted R2 = 0.106; F = 4.306; P = 0.048)

Table 2. Results of spectral analysis. We show the peaks in observed periodicity (in years), and significance for the time series variables: Pbg–LPUE. Blackspot seabream (Pagellus bogaraveo) landing per unit effort (LPUE) from harbours Algeciras, Tarifa and Conil for the period 1983–2011; Pbg. Blackspot seabream (Pagellus bogaraveo) landing from Tarifa for the period 1985–2011; NAOpy. Corresponding average for the North Atlantic Oscillation index in the year before at landing; AOpy. Corresponding average for the Arctic Oscillation index in the year before at landing. Tabla 2. Resultados del análisis espectral. Mostramos los máximos en la periodicidad observada (en años) y la significación de las variables de series temporales: Pbg–LPUE. Desembarque de besugo de la pinta (Pagellus bogaraveo) por unidad de esfuerzo (LPUE) en los puertos de Algeciras, Tarifa y Conil para el período 1983–2011; Pbg. Desembarque de besugo de la pinta (Pagellus bogaraveo) de Tarifa para el período 1985–2011; NAOpy. Promedio de la oscilación del Atlántico Norte en el año anterior al desembarque; AOpy. Promedio de la oscilación del Ártico en el año anterior al desembarque.. Variables

Periodicity

p (random)

Pbg–LPUE

13.997 years

0.683

Pbg

18.18 years

0.003645

NAOpy

2.59 years

0.7846

AOpy

14.28 years

0.7866

In addition, we obtained a significant model for the probability of obtaining good blackspot seabream landings, according to logit (y) function (fig. 3): y = –0.645 + 3.344 * AOpy The statistical tests for the goodness–of–fit of the model indicated a good fit. An omnibus test for model coefficients obtained x2 = 6.774, p = 0.009, and the Hosmer AND Lemeshow test obtained x2 = 8.740, p = 0.272. The AUC of the model was 0.756, which can be considered acceptable discrimination (Hosmer & Lemeshow, 2000). The Nagalkerke test obtained R2 = 0.312. The favourability function showed that the condi� tions that favour good blackspot seabream landings for a specific year coincided almost completely with the positive phase of the AOpy (fig. 3). Accumulated values for the NAO and AO were highly correlated (R2 = 0.91, p < 0.01) (fig. 4). The temperature and salinity trends corresponding to MWu over the period 1982–2007 are shown in figures 5A and 5B. Black dots represent the grid points where

Animal Biodiversity and Conservation 37.1 (2014)

110 100 LPUE

90 80 70 60 50 –1.5

–1

–0.5

0 NAOpy

0.5

Fig. 2. Linear relationship between the landing per unit effort (LPUE) of blackspot seabream from the Algeciras, Tarifa and Conil harbours (Gulf of Cadiz) and the North Atlantic Oscillation previous year (NAOpy) to the landing, for the period 1983–2011. Fig. 2. Relación lineal entre el desembarque por unidad de esfuerzo (LPUE) del besugo de la pinta en los puertos de Algeciras, Tarifa y Conil (golfo de Cádiz) y la oscilación del Atlántico Norte del año anterior (NAOpy) al desembarque para el período 1983–2011.

trends with a significance level greater than 90% were obtained. The blanks areas correspond to points with few measurements of MWu for the period under study following the protocol described above. The tempera�

ture trend (fig. 5A) was positive for the significant area with maximum values close to 0.2ºC per decade near to the Portuguese coast. A similar pattern was observed for salinity trends (fig. 5B) with maximum values close

1 0.9 0.8 Probability

0.7 0.6 0.5 0.4 0.3 0.2 0.1

–0.8

–0.6

–0.4

–0.2

0.2 0.4 AOpy

0.6

0.8

1.2

Fig. 3. Probability of obtaining good blackspot seabream landings from Tarifa harbour compared to the average Arctic Oscillation (AO) index for the year prior to landing (AOpy, gray circles), and the adjusted favorability for good blackspot seabream landings (black triangles). We plotted the years with good blackspot seabream landings (top squares) and years with poor blackspot seabream landings (bottom squares). Fig. 3. Probabilidad de obtener buenos desembarques de besugo de la pinta en el puerto de Tarifa en comparación con el índice medio de oscilación del Ártico (OA) para el año anterior al desembarque (AOpy, círculos grises) y favorabilidad ajustada de los buenos desembarques de besugo de la pinta (triángulos negros). Elaboramos un gráfico con los años de desembarques buenos de besugo de la pinta (cuadrados superiores) y los malos (cuadrados inferiores).

Báez et al.

Accumulated

0.5

–0.5 1986

1990

1994 1998 Time (year)

2002

2006

Fig. 4. Evolution of accumulated values for the North Atlantic Oscillation (gray line), Arctic Oscillation (dashed gray line), and Pagellus bogaraveo landings (black line). Signs were normalized so they could be represented in combination. Fig. 4. Evolución de los valores acumulados para la oscilación del Atlántico Norte (línea gris), la oscilación del Ártico (línea discontinua gris) y los desembarques de Pagellus bogaraveo (línea negra). Se normalizaron los signos para que pudieran representarse en combinación.

to 0.05 per decade near the Portuguese coast. Warming and salinification were almost negligible at locations far from the coast. Salinity and temperature time series were calculated by averaging the grid points in the area under study where trends with a significance level greater than 90% were obtained. Figure 6 shows the time evolution of Pagellus bogaraveo landings and bac� kward averaged MWu salinity and temperature, where the mean values (S and T) corresponding to a certain year were calculated by averaging the previous five years (e.g. backward averaged values for 1985 were calculated using values for the period 1980–1984). Both water properties were negatively correlated with landings (salinity: R = –0.71, p < 0.01; temperature: R = –0.68, p < 0.01). Discussion Few studies have shown that large–scale atmospheric phenomena could affect deep–sea population dyna� mics (e.g. Ruhl & Smith, 2004; Maynou, 2008; and references therein). Maynou (2008) found that the annual strength of red shrimp (Aristeus antennatus) landings is affected by variations in NAO (especially in winter) in the previous two or three years our re� sults indicate that the long–term landing of blackspot seabream from the Strait of Gibraltar is associated with the atmospheric oscillations. The positive NAO results in stronger–than–average westerly winds across northern mid–latitudes, affect�

ing both marine and terrestrial ecosystems, while a positive AO phase is characterized by a strong polar vortex (from the surface to the lower stratosphere). In this situation, storms increase in the North Atlantic and drought prevails in the Mediterranean basin. Strong winds agitate the water, favouring the mixing of deep water and surface water, and thus increasing the supply of nutrients at the surface. When the NAO and AO is in a negative phase, the continental cold air sinks into the Midwestern United States and Western Europe, while storms bring rain to the Mediterranean region (Ambaum et al., 2001). According to Maynou (2008), 'decreased rainfall during positive NAO years may increase water–mass mixing in the NW Mediterranean, enhancing meso– zooplankton production and food resources to Aristeus antennatus, especially in late winter when females are undergoing ovary maturation and require higher energy input. During years of enhanced food resources the reproductive potential of females would increase, and strengthen particular year classes that appear in the landings two to three year later'. Our results suggest the same explanation. We observed a significant nega� tive correlation between blackspot seabream landings and the temperature and salinity values obtained by calculating MWu. According to Báez et al. (2013a) a positive NAOpy and AOpy increases the amount of snow in the mountains surrounding the Alboran Sea, thus increasing the amount of continental freshwater entering the sea the following year, which in turn reduces surface salinity, and blocks water upwelling.

Animal Biodiversity and Conservation 37.1 (2014)

40º N

0.2 0.15

39º N

0.1

38º N

0.05

37º N

36º N

–0.05 14º W 13º W 12º W 11º W 10º W 9º W

0.05

40º N

0.04

39º N

0.03 0.02

38º N

0.01

37º N

0 –0.01

36º N

–0.02

14º W 13º W 12º W 11º W 10º W 9º W

Fig. 5. Temperature (A, in ºC/d) and salinity (B, in psu/d) trends corresponding to upper Mediterranean water (MWu). Dots mark locations where trends with a significance level greater than 90% were obtained. Blank areas correspond to points with few measurements of MWu. Fig. 5. Tendencias de la temperatura (A, en ºC/d) y la salinidad (B, en psu/d) correspondientes a la corriente superior de agua del Mediterráneo (MWu). Los puntos indican los lugares en los que se obtuvieron tendencias con un grado de significación superior al 90%. Las zonas en blanco corresponden a los puntos con pocas mediciones de la MWu.

We hypothesize that deep cold waters in the Al� boran Sea are prevented from upwelling in the years following positive NAO and AO phases, and appear in the Atlantic as colder MWu. This chain of events seems to benefit the eco–physiology of blackspot seabream by increasing their biomass. In this context, the dependence link could be due to an increase in survival of larvae related to higher amounts of

food. This hypothesis is reinforced by the strongest correlation found for the AO with a lag of two years (R2 = 0.95, p < 0.01). In recent years, a decreasing trend in blackspot seabream landings has been observed. However, this trend has coincided with the end of a long positive NAO and AO cycle between the 1980s and 1990s (Fyfe et al., 1999). Thus, prolonged periods of a

Báez et al.

0.5

–0.5

1986

1990

1994 1998 Time (year)

2002

2006

Fig. 6. Time evolution of Pagellus bogaraveo landings (black line) and backward averaged MWu temperatures (gray line) and salinity (dashed gray line). Signs were normalized so they could be represented in combination. Fig. 6. Evolución de los desembarques de Pagellus bogaraveo (línea negra) y los promedios de la temperatura (línea gris) y la salinidad (línea gris discontinua) de la MWu saliente. Se normalizaron los signos para que pudieran representarse en combinación.

positive AO trend could favour high fishery yields. In contrast, a negative NAO and AO phase drasti� cally reduces production. In the context of global change, this situation could have major implications for fisheries management. Thus, during positive NAO and AO phases, high exploitation levels could be allowed while maintaining the stocks within safety limits. During negative NAO and AO phases, more restrictive management measures should be adopt� ed, such as lower exploitation levels, or temporary fishery closure, to preserve fishery sustainability and population safety. Changes in the NAO and AO are correlated over long time periods (Feldstein & Franzke, 2006). Given the strong impact of the AO and NAO on the weather and climate of the wealthiest areas of the planet, and their large socioeconomic impact on energy, agri� culture, fisheries, industry, traffic and human health throughout the whole of Europe and eastern North America, there has been great interest in quantifying the extent to which the phenomena are predictable and the ability of climate numerical models to simu� late them. Bojariu & Gimeno (2003a) provide a good review of the topic. Predictive patterns have been iden� tified in the Atlantic SSTs preceding specific phases of the AO and NAO by up to six months (Rodwell & Folland, 2003), in Eurasian snow cover by up to one year (Bojariu & Gimeno, 2003b), and in the extent of sea–ice over the Arctic (Deser et al., 2000). Thus, the Atlantic SST, Eurasian snow cover, and Arctic sea–ice are good candidates to explore fisheries in Strait of Gibraltar up to one year in advance.

It is widely accepted that the planet is experiencing a period of rapid global warming (Oreskes, 2004), pri� marily driven by human activity (Keller, 2007). Although there is increasing concern over the impact of global warming on marine biodiversity and fisheries ecology (Yatsu et al., 2005), it is difficult to predict how the climate could alter marine biodiversity. In this context, climate change simulations with greenhouse gas and aerosol forcing for the period 1900–2100 indicate a positive trend in the AO (Fyfe et al., 1999). On the other hand, the AO responds to natural changes, such as the increase in stratospheric aerosols due to volcanic eruptions (Christiansen, 2007). Thus, via the NAO and AO, global warming could affect the fisheries ecology of blackspot seabream from the Strait of Gibraltar. This possibility could be extrapolated to other northern hemisphere stocks. Acknowledgements This work was partially supported by projects from the IEO based in Malaga, GPM–4 programs (IEO) and PNDB (EU–IEO), project CGL 2009–11316/ BOS from the Spanish Government and FEDER, and from the Xunta de Galicia under 'Programa de Consolidación e Estruturación de Unidades de Investigación' (Grupos de Referencia Competitiva) funded by FEDER. An �� anonymous reviewer pro� vided helpful comments on earlier versions of the manuscript. We would also like to thank to Andrew Paterson for style corrections.

Animal Biodiversity and Conservation 37.1 (2014)

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Sarpa salpa herbivory on shallow reaches of Posidonia oceanica beds L. Steele, K. M. Darnell, J. Cebrián & J. L. Sanchez–Lizaso

Steele, L., Darnell, K. M., Cebrián, J. & Sanchez–Lizaso, J. L., 2014. Sarpa salpa herbivory on shallow reaches of Posidonia oceanica beds. Animal Biodiversity and Conservation, 37.1: 49–57. Abstract Sarpa salpa herbivory on shallow reaches of Posidonia oceanica beds.— Here, we examined the temporal and small–scale spatial variability of grazing by the herbivorous fish Sarpa salpa on shallow beds of the temperate seagrass Posidonia oceanica. Herbivory intensity expressed as the percent of leaf area taken by fish bites was higher in September 2006 than in February 2007, and at 0.5 m than at 1.5 m during both sampling times. All S. salpa feeding at the shallow locations studied were juveniles, with bite sizes ranging from 0.03 to 0.62 cm2. Juveniles feeding at 1.5 m were larger in February 2007 than in September 2006, as evidenced by significant differences in mean bite size per shoot. However, the larger juveniles feeding at 1.5 m in February 2007 did not appear to feed as frequently as the comparatively smaller juveniles feeding at the same depth in September 2006, as suggested by significant differences in number of bites per shoot. The number of bites per shoot was also lower at 1.5 m than at 0.5 m in February 2007, although mean bite size did not differ significantly between the two depths at that sampling time. In general S. salpa juveniles did not select a particular range of leaf ages when feeding in the study locations, although the juveniles feeding at 1.5 m in September 2006 appeared to select mid–aged leaves. Fish did not show a preference for more epiphytized leaves. These results show that grazing activity by S. salpa juveniles in shallow reaches of P. oceanica meadows may vary temporally and across small changes in depth, which in turn may affect the overall intensity of herbivory on the seagrass. Key words: Grazing, Herbivory, Seagrass, Sarpa salpa, Posidonia oceanica. Resumen El herbivorismo de Sarpa salpa en los tramos someros de los lechos de Posidonia oceanica.— En el presente artículo analizamos la variabilidad temporal y espacial en pequeña escala de la actividad de alimentación del pez herbívoro Sarpa salpa en los lechos someros de la pradera submarina de clima templado Posidonia oceanica. La intensidad del herbivorismo expresada como el porcentaje de superficie foliar mordida por el pez fue superior en septiembre de 2006 que en febrero de 2007, y a una profundidad de 0,5 m que de 1,5 m durante los dos períodos de muestreo. Todos los individuos de S. salpa que se alimentaban en las zonas someras estudiadas eran juveniles y el tamaño de mordedura se situaba entre 0,03 y 0,62 cm2. Los juveniles que se alimentaban a 1,5 m de profundidad fueron mayores en febrero de 2007 que en septiembre de 2006, tal como ponen de relieve las diferencias significativas existentes en el tamaño medio de mordedura por brote. No obstante, los juveniles más grandes que se alimentaban a 1,5 m de profundidad en febrero de 2007 no parecían alimentarse con tanta frecuencia como los juveniles comparativamente más pequeños, que lo hacían a la misma profundidad en septiembre de 2006, tal como sugieren las diferencias significativas halladas en el número de mordeduras por brote. Asimismo, el número de mordeduras por brote fue inferior a 1,5 m de profundidad que a 0,5 m en febrero de 2007, si bien el tamaño medio de las mordeduras no difería en medida significativa entre las dos profundidades en aquel período de muestreo. En general, los juveniles de S. salpa no elegían un intervalo de edad concreto de las hojas a la hora de alimentarse en los lugares del estudio, a pesar de que los juveniles que se alimentaban a 1,5 m de profundidad en septiembre de 2006 parecían elegir hojas de edad mediana. Los peces no mostraron preferencia por las hojas con mayor cobertura de epífitos. Estos resultados muestran que la actividad de alimentación de los juveniles de S. salpa en los tramos someros de las praderas de P. oceanica puede variar con el tiempo y con pequeños cambios de profundidad, lo que a su vez puede afectar a la intensidad general del herbivorismo sobre las praderas submarinas. ISSN: 1578–665 X eISSN: 2014–928 X

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Palabras clave: Alimentación, Herbivorismo, Praderas submarinas, Sarpa salpa, Posidonia oceanica. Received: 11 III 13; Conditional acceptance: 18 VIII 13; Final acceptance: 17 IV 14 LaTina Steele & Just Cebrián, Dauphin Island Sea Lab, 101 Bienville Blvd., Dauphin Island, AL 36528, USA.– Kelly M. Darnell, Marine Science Inst., Univ. of Texas, 750 Channel View Dr., Port Aransas, TX 78373, USA.– Just Cebrián, Dept. of Marine Sciences, Univ. of South Alabama, 307 N. University Blvd., Mobile, AL 36688, USA.– Jose Luis Sanchez–Lizaso, Dept. of Marine Sciences and Applied Biology, Univ. of Alicante, Ap. 99, E–03080 Alicante, Spain. Corresponding author: LaTina Steele, Sacred Heart Univ., Dept. of Biology, 5151 Park Ave., Fairfield, CT 06825, USA. E–mail: steelel@sacredheart.edu

Animal Biodiversity and Conservation 37.1 (2014)

Introduction

Materials and methods

It is increasingly recognized that herbivory can influence the structure of seagrass populations (Mariani & Alcoverro, 1999; Alcoverro & Mariani, 2004; Valentine & Duffy, 2006; Heck & Valentine, 2006; Prado et al., 2007a), as well as that of seagrass–associated organisms (Prado et al., 2007a). Grazing of seagrasses has historically been perceived to be relatively unimportant in temperate species such as Posidonia oceanica (Vizzini et al., 2002; Cebrian et al., 1996a; Gobert et al., 2006; Mateo et al., 2006). However, studies have increasingly highlighted the important role of herbivory on beds of P. oceanica (Tomas et al., 2005; Verges et al., 2007a, 2007b; Prado et al., 2007b, Ruiz et al., 2009). In some cases, herbivory may even exacerbate damage to P. oceanica beds caused by other stressors, such as excessive input of organic matter, making herbivory particularly important to consider in this species (Ruiz et al., 2009). The disagreement over the importance of herbivory may be attributed to considerable seasonal and spatial variation in grazing rates and amounts of tissue lost to herbivory in P. oceanica beds (Cebrián & Duarte, 1998; Boudouresque & Verlaque, 2001; Peirano et al., 2001). For example, Tomas et al. (2005) assessed patterns of grazing by the fish Sarpa salpa on P. oceanica and found higher grazing on seagrass leaves at 5 m depth than at 10 m, as well as a seasonal grazing peak in summer. Similarly, Peirano et al. (2001) and Prado et al. (2007b) documented higher levels of S. salpa herbivory in summer months. Seagrass herbivory also varies at small spatial scales (among shoots, leaves, and other plant parts; Verges et al., 2007a). Some seagrass herbivores are known to feed selectively, based on leaf nutritional quality (McGlathery, 1995; Goecker et al., 2005) as well as structural and chemical deterrents (Verges et al., 2007a, 2007b; Prado & Heck, 2011; Steele & Valentine, unpublished data). Epiphyte cover can also be an important feeding cue for herbivorous fishes. For example, two fish species feeding on the Australian seagrass Posidonia australis displayed higher consumption of epiphytized leaves than of leaves with no epiphytes (Wressnig & Booth, 2007). Similarly, S. salpa and the sea urchin Paracentrotus lividus may selectively consume P. oceanica leaves with higher epiphyte loads (Cebrián et al., 1996a; Peirano et al., 2001). Assessing the spatial and temporal variability of herbivory on P. oceanica beds, as well as patterns of feeding selectivity by seagrass grazers, is important to understand the role that grazers have on the structure and functioning of these Mediterranean ecosystems. Here we examined the variability in the magnitude and selectivity of herbivory by S. salpa across a 1–m depth range in shallow P. oceanica beds between summer and winter sampling dates. We focused on a finer spatial scale than previous studies, and simultaneously addressed spatial variability in both the intensity and leaf–age selectivity of grazing by juvenile S. salpa in two seasons.

Posidonia oceanica shoots were collected from shallow beds in the southwestern Mediterranean near the town of Santa Pola (Spain), where the sea urchin Paracentrotus lividus only occurs in extremely low abundance (L. Steele, personal observation), in September 2006 and February 2007. These two collection dates captured most of the leaves produced throughout the year. Shoots collected in September 2006 bore leaves born since the previous spring (i.e. February–March 2006) through that summer, whereas the shoots collected in February 2007 bore leaves produced in between the previous fall and that winter (Ott, 1980; Cebrián et al., 1994). On each sampling date (September 2006, February 2007), 25 shoots were collected from each of three sites at least 10 m apart at two depths, 0.5 m and 1.5 m, with a total of 75 shoots per depth per date. Collection sites were not marked so different sites within each depth were sampled on each date. Upon collection, the leaves on the shoots were separated and digitally photographed for later analysis. S. salpa leaves distinct, half–moon shaped bite marks on the leaf blades of P. oceanica that are clearly discernible (Boudouresque et al., 1984). Apparent herbivory intensity (defined as the percentage of total leaf area per shoot taken by bite marks visible at the moment of sampling), number of bites per shoot, and mean bite size per shoot were derived using Sigma Scan software. Epiphyte cover on the leaf (0–25%, 26–50%, 51–75%, or 76–100%) was also visually estimated from the photographs of shoots collected in September 2006. Apparent herbivory intensity, number of bites per shoot, and mean bite size per shoot were compared between sampling times and depths using two–way ANOVA, with time and depth both as fixed factors. The focus of this study is to examine the overall variability between the two dates and depths sampled, and not within each depth, and thus sites were pooled for each depth and time. Post–hoc multiple comparisons were done with Tukey tests between times for each depth and between depths within each time (Quinn & Keough, 2002). For each depth in September 2006, a one–way ANOVA was used to determine if apparent herbivory intensity varied among the four groups of epiphyte cover, followed by post–hoc Tukey multiple comparison tests to determine which groups of epiphyte cover differed from each other. We also inspected the relationship between apparent herbivory intensity on the leaf (expressed as percentage of leaf area taken by the bite marks visible at the time of sampling) and leaf age to determine whether S. salpa fed selectively on a certain leaf age range. Leaf age was calculated using the plastochrone interval (PI) method described by Cebrián et al. (1994). Briefly, this method uses the length of the sheath (Lsheath) of the oldest leaf and the length of the youngest (leaf 1) and second to youngest (leaf 2) leaves (Lleaf 1, Lleaf 2) to calculate the age of the youngest leaf on the shoot expressed as a fraction of one PI. Namely:

Steele et al.

Age leaf 1 =

ln(Lleaf 1) – ln(Lsheath) ln(Lleaf 2) – ln(Lleaf 1)

(1)

The age of the youngest leaf can then be converted to days by multiplying its fractional age in PI units by the average chronological equivalence of the PI (i.e. mean number of days per PI) at the time of sampling. Similarly, the age of any other leaf on the shoot can be estimated by summing the average chronological equivalences of the PIs for that given leaf and all younger leaves. In turn, the average chronological equivalence of each PI can be obtained from the well–established seasonal cycle of the PI chronological duration (Ott, 1980; Romero, 1985; Cebrián et al., 1994). This technique has been used to characterize leaf age–dependent processes in seagrasses (Sand–Jensen et al., 1994; Cebrián et al., 1996a, 1996b; Cebrián et al., 1999). For more details we direct the reader to Cebrián et al. (1994). To reduce scatter in the relationship between apparent herbivory intensity and leaf age, and allow for better comparability among sampling depths and times, we first calculated the mean apparent herbivory intensity for leaves within 30–day age intervals (i.e. mean intensity for leaves with ages comprised between 0 and ≤ 30 days; mean intensity for ages comprised between 30 and ≤ 60 days, and so on). For each sampling depth and time we then identified the age interval at which consumption was first detected, and defined leaf age at the onset of herbivory as the lower limit of that interval. We subsequently plotted the mean apparent herbivory intensity for intervals with registered herbivory versus the upper limit of the interval after subtracting the age at the onset of herbivory (i.e. leaf age since the onset of herbivory, La). Herbivory intensity at the onset of herbivory was considered zero since all of the intervals beneath that age did not have any registered herbivory. This approach, which has been previously used in studies of seagrass herbivory (Cebrián et al., 1996a; Cebrián & Duarte, 1998), allows for the comparison of trends in herbivory intensity with increasing leaf age among seagrass populations regardless of the age at which herbivory first occurs. We used regression analysis to assess whether the plotted relationships followed a linear, parabolic, sigmoidal, or exponential fit. The linear fit: Herbivory intensity = a + b1(La)

(2)

indicates no feeding selectivity by S. salpa on a certain leaf age range since herbivory intensity increases linearly with the time of exposure to herbivory (i.e. leaf age). The parabolic fit: Herbivory intensity = a + b1(La) – b2(La)2

(3)

and the sigmoidal fit: Herbivory intensity = b1/(1 + e

b2 La

)

(4)

indicate preference for mid–aged leaves, with herbivory intensity peaking equation (3) or leveling off

equation (4) at the mid–age range. The exponential fit: Herbivory intensity = a + (La)b3

(5)

indicates preference for old leaves, with non–proportionally higher consumption values at old ages in comparison with younger ages. Results Apparent herbivory intensity varied between the two sampling times and depths (fig. 1, table 1). Namely, it was higher in September 2006 than in February 2007 at each depth, and higher at 0.5 m than at 1.5 m within each sampling time (table 1). On average, shoots at 1.5 m had larger bites in February 2007 than in September 2006 (fig. 1, table 1). However, shoots at 1.5 m had a lower number of bites in February 2007 than in September 2006 (fig. 1, table 1). In addition, in February 2007 shoots at 0.5 m had more bites than those at 1.5 m, although mean bite size per shoot did not differ significantly between the two depths on that sampling time (table 1). Regression analysis showed a linear relationship between apparent herbivory intensity and leaf age since the onset of herbivory at 0.5 m in September 2006 and at both 0.5 m and 1.5 m in February 2007 (fig. 2, table 2), indicating no feeding preference for a certain leaf age range. A parabolic curve was found to be the best fit at 1.5 m in September 2006 (fig. 2, table 2), indicating preference for mid–aged leaves. ANOVA showed significant differences in apparent herbivory intensity across different levels of epiphyte cover (0–25%, 26–50%, 51–75%, or 76–100% of leaf surface covered by epiphytes) for both depths in September 2006 (0.5 m, P < 0.01; 1.5 m, P < 0.01). Multiple comparison tests revealed significantly lower herbivory intensity on leaves with 0–25% epiphyte cover than on leaves with > 25% cover, but no differences among leaves with 26–50%, 51–75%, and 76–100% cover. The innermost and youngest leaves on the shoot had the least epiphyte cover (0–25%) and lowest herbivory intensity. Discussion The majority of Sarpa salpa feeding in our study locations were juveniles, as indicated by the size of their bite marks (Tomas et al., 2005; Criscoli et al., 2006), since the locations were probably too shallow for the large shoals that adult S. salpa typically form at greater depths. Our study focuses on grazing by S. salpa juveniles within a narrow depth gradient (i.e. comparison between 0.5 m and 1.5 m). We showed that grazing activity and behavior by S. salpa juveniles in shallow P. oceanica beds may vary significantly over time and across small changes in depth. Namely, larger juveniles fed at 1.5 m in February 2007 than in September 2006 (as indicated by bite size), although not as intensively (as indicated by number of bites on the shoot). For the sampling date of February

Animal Biodiversity and Conservation 37.1 (2014)

Apparent herbivory intensity

1.4 1.2 1 0.8 0.6 0.4 0.2 0

B Mean bite size/shoot cm2

0.35 0.3 0.25 0.2 0.15 0.1 0.05

Number of bites/shoot

0 4 3.5 3 2.5 2 1.5 1 0.5

0.5 m 1.5 m September 2006

0.5 m 1.5 m February 2007

Fig. 1. Apparent herbivory intensity (A) (% leaf area per shoot taken by bite marks present at the sampling time), mean bite size per shoot (B), and number of bites per shoot (C) at 0.5 m and 1.5 m in September 2006 and February 2007. Bars denote mean values, and lines ± SE, for the three sites at each combination of depth and time. Fig. 1. Intensidad del herbivorismo visible (A) (% de superficie foliar por brote con marcas de mordeduras en el período de muestreo), tamaño medio de mordedura por brote (B) y número de mordeduras por brote (C) a 0,5 y 1,5 m de profundidad en septiembre de 2006 y febrero de 2007. Las barras denotan los valores medios y las líneas ± ES, para los tres sitios en cada combinación de profundidad y tiempo.

2007, we also found differences between 0.5 m and 1.5 m in that juveniles fed more intensively at the shallower depth, although they seemed to be of similar size between the two depths (fig. 1). We recognize, however, that we would need direct observations of fish size at our study locations to fully demonstrate these inferences. Other studies have focused on greater depth ranges and included both juvenile and adult S. salpa. For instance, Tomas et al. (2005) compared herbivory

by S. salpa on P. oceanica between 5 and 10 m in summer and winter months. They found that herbivory by juvenile S. salpa was higher at 5 m than at 10 m and, during the summer months when present at those depths (adult S. salpa move to deeper waters in winter to spawn, Peirano et al., 2001; Criscoli et al., 2006), grazing by adult S. salpa was also higher at 5 m. Here, we showed that changes in the grazing activity of juvenile S. salpa may occur at a much narrower depth gradient.

Steele et al.

Table 1. Results of ANOVA and post–hoc Tukey tests. Tukey comparisons are between sampling times for each depth, and between depths for each sampling time. For instance, ‘(Sept '06 vs. Feb '07) 0.5 m’ denotes the comparison between September 2006 and February 2007 for 0.5 m, and ‘(0.5 m vs. 1.5 m) Sept '06’ the comparison between 0.5 and 1.5 m for September 2006. Tabla 1. Resultados de la ANOVA y la prueba a posteriori de Tukey. Las comparaciones de Tukey se realizaron entre los períodos de muestreo para cada profundidad y entre las profundidades para cada período de muestreo. Por ejemplo, ''(Sept '06 vs. Feb '07) 0.5 m'' hace referencia a la comparación entre septiembre de 2006 y febrero de 2007 a 0,5 m de profundidad y ''(0.5 m vs. 1.5 m) Sept '06'', a la comparación entre 0,5 y 1,5 m para septiembre de 2006. Herbivory intensity

Time

Depth

Interaction

< 0.01

0.359

Tukey tests (Sept '06 vs. Feb '07) 0.5 m: P < 0.01

(Sept '06 vs. Feb '07) 1.5 m: P < 0.01

(0.5 m vs. 1.5 m) Sept '06: P = 0.025

Mean bite size

< 0.01

0.786

0.048

(0.5 m vs. 1.5 m) Feb '07: P < 0.01 (Sept '06 vs. Feb '07) 0.5m: P = 0.46

(Sept '06 vs. Feb '07) 1.5 m: P < 0.01

(0.5 m vs. 1.5 m) Sept. ’06: P = 0.32

(0.5 m vs. 1.5 m) Feb. ’07: P = 0.66

Number of bites

< 0.01

0.264

(Sept '’06 vs. Feb '07) 0.5 m: P = 0.087

(Sept '06 vs. Feb '07) 1.5 m: P < 0.01

(0.5 m vs. 1.5 m) Sept '06: P = 0.17

(0.5 m vs. 1.5 m) Feb '07: P < 0.01

Past reports have documented higher levels of herbivory by S. salpa on P. oceanica in summer than winter (Tomas et al., 2005; Peirano et al., 2001; Prado et al., 2007b). Here, we also found higher herbivory intensity in our summer sampling time (September 2006) in comparison with the winter sampling time (February 2007). However, this difference seemed more related to the presence of longer leaves (and thus smaller percent of leaf area accounted for by bite marks), rather than to lower grazing activity, in the winter sampling time. At any rate, our temporal analysis was limited because it was only based on two dates. It certainly did not allow for any inferences regarding seasonal variability in herbivory since we did not have replicated times per season. Therefore, our analysis shows that significant temporal variability in grazing activity by S. salpa juveniles does occur in the shallow reaches of P. oceanica beds, but it falls short at describing the nature and extent of that variability among seasons. Notwithstanding the temporal limitations of our study, our results suggest that small–scale changes in feeding activity by S. salpa juveniles may have important implications on the overall intensity of herbivory on P. oceanica; in February 2007 the more intense feeding rates at 0.5 m than at 1.5 m may have contributed to higher herbivory at the former depth. Our results suggest preference for mid–aged leaves by S. salpa juveniles at 1.5 m on the shoots collected in September 2006, and no feeding preference in any

other case (0.5 m in September 2006, and 0.5 m and 1.5 m in February 2007; fig. 2). Cebrián et al. (1996b) examined a number of P. oceanica beds at depths ranging from 4 to 12 m along the Spanish Mediterranean coast during the summers of 1992 and 1993. They found that S. salpa showed preference for mid–aged leaves in some beds, but no preference for any particular leaf age range in others. However, the authors did not make any effort to determine the development stage of the fish based on the size of the bite marks. In concert, these results indicate that patterns of leaf age selectivity by S. salpa vary spatially and temporally. Differences in patterns of leaf age selectivity may have important implications for P. oceanica populations. Mid–age leaves are metabolically more active than old leaves (Alcoverro et al., 1998), and selective herbivory on mid–aged leaves may therefore have a greater impact on the productivity of P. oceanica. In contrast to the results of Peirano et al. (2001), S. salpa at our study locations did not feed preferentially on leaves with higher epiphyte cover. We found no differences in apparent herbivory intensity among leaves with 26–50%, 51–75%, or 76–100% epiphyte cover. There was significantly more grazing on leaves with 26–100% epiphyte cover than on leaves with 0–25% epiphyte cover; however, the leaves with few epiphytes were young leaves located in the inner portion of the seagrass shoot, and they were less accessible to grazers than older leaves.

Animal Biodiversity and Conservation 37.1 (2014)

September 2006

10 Apparent herbivory intensity (% of leaf area occupied by bite marks)

0.5 m

0.5 m 2

6 4

2 0

February 2007

120 150 180

1.6

0 30 60 90 120 150 180 210 240

0.012 1.5 m

1.2

1.5 m 0.008

0.8 0.004

0.4 0

0 90 120 150 0 30 60 90 120 150 180 210 240 Leaf age since the onset of herbivory (days)

Fig. 2. Apparent herbivory intensity (% of leaf area per shoot taken by bite marks present at the sampling time) vs. leaf age since the onset of herbivory for Posidonia oceanica leaves at two depths in September 2006 and February 2007. Fig. 2. Intensidad del herbivorismo visible (% de superficie foliar por brote con marcas de mordeduras en el período de muestreo) en función de la edad de las hojas desde el inicio del herbivorismo de Posidonia oceanica a las dos profundidades en septiembre de 2006 y febrero de 2007.

Our measurements of herbivory (i.e., apparent herbivory intensity) most likely underestimated the true consumption by S. salpa because heavily grazed leaves tend to break off more easily and, thus, they may have been grossly under sampled. Indeed, prior studies (Kirsch et al., 2002; Tomas et al., 2005; Prado et al.,

2007b) have shown that the extent of underestimation generated by herbivory calculations based on extant bite marks on the leaves may be considerable. Yet our results should be robust, at least qualitatively. For the September 2006 samples, mean bite size and number of bites per shoot were similar between 0.5 and 1.5 m.

Table 2. Parameters and statistics of the equations describing the relationship between herbivory intensity and leaf age since the onset of herbivory (see text for details): A. Age at the onset of herbivory (in days). Tabla 2. Parámetros y datos estadísticos de las ecuaciones que describen la relación entre la intensidad del herbivorismo y la edad de las hojas desde el inicio del herbivorismo (véase el texto para conocer más detalles): A. Edad en el inicio del herbivorismo (en días).

Time

Depth R 2

September 06

0.5 m

0.85

1.5 m

February 07

b1 ± SE (% removed leaf–1 day–1)

b2 ± SE (% removed leaf–1 day–2)

0.023 ± 0.004

–

0.91

0.014 ± 0.003

0.000073 ± 0.000026

0.5 m

0.84

0.0066 ± 0.0011

–

1.5 m

0.95

0.000030 ± 0.000003

–

Thus, herbivory should have been underestimated to a similar extent for the two depths on that sampling time since leaves should have had a similar likelihood of breakage based on the number and size of bites present. A similar underestimation for both depths should not negate the result that herbivory intensity was higher at 0.5 than at 1.5 m in September 2006. For the February 2007 samples, mean bite size was similar at the two depths, but there were more bites per shoot at 0.5 than at 1.5 m (fig. 1). On this basis, it could be expected that leaf breakage probability (and thus potential herbivory underestimation) would be higher at 0.5 m than at 1.5 m, but again this should not negate the result that herbivory intensity was higher at 0.5 m than at 1.5 m on this sampling date. Herbivory underestimation could perhaps affect the shape of leaf age–feeding selectivity relationships, as older leaves may sustain large cumulative herbivory and break off more easily. On these grounds, our three linear fits might have been misrepresentations of exponential curves, but this possibility seems rather unlikely judging by the levels of apparent grazing found on mid–aged leaves. Our parabolic fit, however, could be a misrepresentation of a linear process (i.e. no selectivity for any particular leaf age range). In conclusion, we have shown that grazing activity and behavior by S. salpa juveniles may vary significantly over time and within small depth ranges in shallow P. oceanica beds. Such variations may have important implications on the overall levels of herbivory intensity on the seagrass. In these shallow reaches, S. salpa juveniles may feed preferentially on mid–age leaves, although they often showed no preference for a particular leaf age range. Our results add to previous findings of depth–related gradients in S. salpa herbivory. However, more accurate assessments of the temporal and spatial variability of feeding behavior by S. salpa juveniles in shallow P. oceanica beds are needed to better understand their role in these environments. Acknowledgements The authors would like to thank CIMAR in Santa Pola, Spain for their hospitality during our stay. In particular, we thank Alfonso Ramos for his guidance. Andres Izquierdo provided invaluable assistance with field collections in February 2007. Ashley McDonald helped with data treatment and figure preparation. The University of South Alabama Department of Marine Sciences and Dauphin Island Sea Lab provided travel funds and monetary support. References Alcoverro, T., Manzanera, M. & Romero, J., 1998. Seasonal and age–dependent variability of Posidonia oceanica (L.) Delile photosynthetic parameters. Journal of Experimental Marine Biology and Ecology, 230: 1–13. Alcoverro, T. & Mariani, S., 2004. Patterns of fish

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and sea urchin grazing on tropical Indo–Pacific seagrass beds. Ecography, 27: 361–365. Boudouresque, C. F., Jeudy de Grissac, A. & Olivier, J. (Eds.), 1984. I. International workshop on Posidonia oceanica beds. GIS Posidonie, Marseille. Boudouresque, C. F. & Verlaque, M., 2001. Edible sea urchins: biology and ecology. In: Developments in Aquaculture and Fisheries Science: 177–216 (J. M. Lawrence, Ed.). Elsevier, Amsterdam. Cebrián, J. & Duarte, C. M., 1998. Patterns of leaf herbivory on seagrasses. Aquatic Botany, 60: 67–82. Cebrián, J., Duarte, C. M., Marba, N., Enriquez, S., Gallegos, M. & Olesen, B., 1996a. Herbivory on Posidonia oceanica: magnitude and variability in the Spanish Mediterranean. Marine Ecoogy Progress Series, 130: 147–155. Cebrián, J. Enríquez, S., Agawin, N., Duarte, C. M., Fortes, M. & Vermaat, J., 1999. Epiphyte accrual on Posidonia oceanica (L.) Delile leaves: implications on light absorption. Botanica Marina, 42: 123–128. Cebrián, J., Marba, N. & Duarte, C. M., 1994. Estimating leaf age of the seagrass Posidonia oceanica (L.) Delile using the plastochrone interval index. Aquatic Botany, 49: 59–65. – 1996b. Herbivory on Cymodocea nodosa (Ucria) Ascherson in contrasting Spanish Mediterranean habitats. Journal of Experimental Marine Biology and Ecology, 204: 103–111. Criscoli, A., Colloca, F., Carpentieri, P., Belluscio, A. & Ardizzone, G., 2006. Observations on the reproductive cycle, age and growth of the salema, Sarpa salpa (Osteichthyes: Sparidae) along the western central coast of Italy. Scientia Marina, 70: 131–138. Gobert, S., Cambridge, M. L., Velimirov, B., Pergent, G., Lepoint, G., Bouquegneau, J. M., Dauby, P., Pergent–Martini, C. & Walker, D. I., 2006. Biology of Posidonia. In: Seagrasses: Biology, Ecology, and Conservation: 387–408 (A. W. D. Larkum, R. J. Orth & C. M. Duarte, Eds.). Springer, Dordrecht. Goecker, M. E., Heck Jr., K. L. & Valentine, J. F., 2005. Effects of nitrogen concentrations in turtlegrass Thalassia testudinum on consumption by the bucktooth parrotfish Sparisoma radians. Marine Ecology Progress Series, 286: 239–248. Heck Jr., K. L. & Valentine, J. F., 2006. Plant–herbivore interactions in seagrass meadows. Journal of Experimental Marine Biology and Ecology, 330: 420–436. Kirsch, K. D., Valentine, J. F. & Heck, K. L., 2002. Parrotfish grazing on turtlegrass Thalassia testudinum: evidence for the importance of seagrass consumption in food web dynamics of the Florida Keys National Marine Sanctuary. Marine Ecology Progress Series, 227: 71–85. Mariani, S. & Alcoverro, T., 1999. A multiple–choice feeing–preference experiment utilising seagrasses with a natural population of herbivorous fishes. Marine Ecology Progress Series, 189: 295–299. Mateo, M. A., Cebrián, J., Dunton, K. & Mutchier, T., 2006. Carbon flux in seagrass ecosystems. In: Seagrasses: Biology, Ecology, and Conservation: 159–192 (A. W. D. Larkum, R. J. Orth & C. M. Duarte, Eds.). Springer, Dordrecht. McGlathery, K. J., 1995. Nutrient and grazing influen-

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ces on a subtropical seagrass community. Marine Ecology Progress Series, 122: 239–252. Ott, J. A., 1980. Growth and production in Posidonia oceanica (L.) Delile. Marine Ecology, 1: 47–64. Peirano, A., Niccolai, I., Mauro, R. & Bianci, C. N., 2001. Seasonal grazing and food preference of herbivores in a Posidonia oceanica meadow. Scientia Marina, 65: 367–374. Prado, P., Alcoverro, T., Martinez–Crego, B., Verges, A., Perez, M. & Romero, J., 2007b. Macrograzers strongly influence patterns of epiphytic assemblages in seagrass meadows. Journal of Experimental Marine Biology and Ecology, 350: 130–143. Prado, P. & Heck Jr., K. L., 2011. Seagrass selection by omnivorous and herbivorous consumers: determining factors. Marine Ecology Progress Series, 429: 45–55. Prado, P., Tomas, F., Alcoverro, T. & Romero, J., 2007a. Extensive direct measurements of Posidonia oceanica defoliation confirm the importance of herbivory in temperate seagrass meadows. Marine Ecology Progress Series, 340: 63–71. Quinn, G. P. & Keough, M. J., 2002. Experimental Design and Data Analysis for Biologists. Cambridge University Press, Cambridge. Romero, J., 1985. Estudio ecológico de las fanerógamas marinas de la costa catalana: producción primaria de Posidonia oceanica (L.) Delile en las Islas Medes. Ph. D. Thesis, Universidad de Barcelona. (In Spanish.) Ruiz, J. M., Perez, M., Romero, J. & Tomas, F., 2009. The importance of herbivory in the decline

of a seagrass (Posidonia oceanica) meadow near a fish farm: an experimental approach. Botanica Marina, 52: 449–458. Sand–Jensen, K., Jacobsen, D. & Duarte, C. M., 1994. Herbivory and resulting plant damage. Oikos, 69: 545–549. Tomas, F., Turon, X. & Romero, J., 2005. Seasonal and small–scale spatial variability of herbivory pressure on the temperate seagrass Posidonia oceanica. Marine Ecology Progress Series, 301: 95–107. Valentine, J. F. & Duffy, J. E., 2006. The central role of grazing in seagrass ecology. In: Biology, Ecology, and Conservation Seagrasses: 463–501 (A. W. D. Larkum, R. J. Orth & C. M. Duarte, Eds.). Springer, Dordrecht. Verges, A., Becerro, M. A., Alcoverro, T. & Romero, J., 2007a. Variation in multiple traits of vegetative and reproductive tissues influences plant–herbivore interactions. Oecologia, 151: 675–686. – 2007b. Experimental evidence of chemical deterrence against multiple herbivores in the seagrass Posidonia oceanica. Marine Ecology Progress Series, 343: 107–114. Vizzini, S., Sara, G., Michener, R. H. & Mazzola, A., 2002. The role and contribution of the seagrass Posidonia oceanica (L.) Delile organic matter for secondary consumers as revealed by carbon and nitrogen stable isotope analysis. Acta Oecologica, 23(4): 277–285. Wressnig, A. & Booth, D. J., 2007. Feeding preferences of two seagrass grazing monacanthid fishes. Journal of Fish Biology, 71: 272–278.

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Preferencia de hábitat del murciélago hortelano meridional Eptesicus isabellinus (Temminck, 1840) en ambientes mediterráneos semiáridos F. Lisón, Á. Haz & J. F. Calvo

Lisón, F., Haz, Á. & Calvo, J. F., 2014. Preferencia de hábitat del murciélago hortelano meridional Eptesicus isabellinus (Temminck, 1840) en ambientes mediterráneos semiáridos. Animal Biodiversity and Conservation, 37.1: 59–67. Abstract Habitat preference of the meridional serotine bat Eptesicus isabellinus (Temminck, 1840) in semiarid Mediterranean landscapes.— Several molecular studies have recently reported the presence of a second species of the genus Eptesicus in the Iberian peninsula, the meridional serotine bat, E. isabellinus. This species is present in the south of Iberia and it seems to have an allopatric distribution with its twin species, E. serotinus. Ecological studies are now needed to understand the biology of E. isabellinus in southeast Spain. In this study, we used presence–only data for E. isabellinus to perform an ecological niche factor analysis (ENFA) and to create a habitat suitability map (HSM). Our results show that the species has a low marginality index, suggesting it is well adapted to the semiarid conditions of the study area. The main habitats used by E. isabellinus are water courses, scrublands, and zones with high primary productivity. The species avoids non–irrigated cropland and shows no preference for human settlements or irrigated cropland. This study provides information about the ecology of E. isabellinus in southeast Spain and allows us to discuss relevant aspects for its conservation. Key words: Chiroptera, Conservation, Cryptic species, Distribution, ENFA. Resumen Preferencia de hábitat del murciélago hortelano meridional Eptesicus isabellinus (Temminck, 1840) en ambientes mediterráneos semiáridos.— Diversos estudios moleculares han revelado recientemente la existencia de una segunda especie del género Eptesicus en la península ibérica: el murciélago hortelano meridional E. isabellinus. Dicha especie se encuentra en el sur de la península y parece tener una distribución alopátrica con su especie gemela E. serotinus. Debido a la reciente separación entre ambas, es necesario realizar estudios que amplíen el conocimiento sobre la ecología de E. isabellinus. En este trabajo se han empleado datos sobre la presencia de la especie en la Región de Murcia (SE de España) para elaborar modelos de nicho ecológico (MNE) y el correspondiente mapa de idoneidad de hábitat. Los resultados revelan que la especie tiene un índice de marginalidad bajo, lo que sugiere una buena adaptación a las condiciones ambientales semiáridas de la zona de estudio. El principal hábitat utilizado por la especie son los cursos de agua, las zonas de matorral y, en general, los ecosistemas con una elevada productividad. Por el contrario, la especie parece evitar las zonas agrícolas de secano y no muestra especial preferencia por las zonas más humanizadas ni los cultivos de regadío. La información proporcionada por este estudio contribuye al conocimiento de la ecología de E. isabellinus y los resultados del modelo de distribución permiten discutir sobre los aspectos importantes para su conservación. Palabras clave: Chiroptera, Conservación, Especies crípticas, Distribución, MNE. Received: 18 XII 13; Conditional acceptance: 26 III 14; Final acceptance: 25 IV 14 Fulgencio Lisón & José F. Calvo, Depto. de Ecología e Hidrología, Univ. de Murcia, Campus de Espinardo, 30100 Murcia, España (Spain).– Ángeles Haz, Paseo Rosales 10, 4º D, Molina de Segura, 30500 Murcia, España (Spain). Corresponding author: Fulgencio Lisón. E–mail: lison@um.es

ISSN: 1578–665 X eISSN: 2014–928 X

Introducción El género Eptesicus Rafinesque, 1820 (Chiroptera, Vespertilionidae) se encuentra distribuido por zonas tropicales y templadas de todos los continentes (excepto la Antártida), y parece estar formado por una serie de grupos evolutivos que presentan una compleja consideración taxonómica (Juste et al., 2013). En la península ibérica se ha confirmado mediante análisis moleculares la existencia de dos especies del género Eptesicus (Ibáñez et al., 2006; Juste et al., 2009, 2013) que parecen tener una distribución alopátrica, aunque en el norte de Portugal se ha detectado la presencia conjunta de ambas especies (Barros, 2011) y es probable que exista una zona de contacto en Sierra Morena (Santos et al., en prensa). El murciélago hortelano común E. serotinus (Schreber, 1774) aparece distribuido por la zona septentrional de la península y el resto de Europa, mientras que el murciélago hortelano meridional E. isabellinus (Temminck, 1840) aparece en la mitad sur de la península y el norte de África (García–Mudarra et al., 2009; Ibáñez et al., 2013). A pesar de que ambas especies son abundantes y suelen aparecer ligadas a hábitats antrópicos (Cabrera, 1914; Balcells, 1963; Ibáñez, 2007; Flaquer et al., 2010), los cambios ambientales de los últimos decenios (urbanización, transformación de usos del suelo, proliferación de plaguicidas, etc.) podrían afectar negativamente a sus poblaciones (Ibáñez, 2007; Lisón et al., 2010). La separación entre ambas especies y la nueva condición de E. isabellinus como especie nueva de la península ibérica, hacen necesaria la ampliación de conocimientos sobre la biología y ecología de esta especie. Además, E. isabellinus, junto con E. serotinus, es el reservorio natural del virus de la rabia EBLV–1 (Vázquez–Morón et al., 2008) y, por tanto, la información sobre su posible distribución es importante desde el punto de vista epidemiológico. El objetivo de este estudio consiste en elaborar modelos de nicho ecológico de la especie utilizando los datos de presencia en una región mediterránea de carácter principalmente semiárido. De esta forma, se pretende: 1) determinar cuáles son las variables ambientales de importancia para la distribución de la especie; 2) elaborar el correspondiente mapa de idoneidad del hábitat y 3) discutir las necesidades de conservación de la especie. Material y métodos Especie estudiada El murciélago hortelano meridional E. isabellinus tiene una talla grande (longitud del antebrazo: 46–55 mm) y con un peso de entre 17 y 28 g, es uno de los murciélagos más grandes de la península ibérica. El pelaje tiene una característica tonalidad amarillenta pálida en la zona dorsal y mucho más blanquecina en la zona ventral. Las partes desprovistas de pelo tienen una tonalidad muy oscura y las orejas son redondeadas y con un trago arriñonado. La especie utiliza como refugios naturales las fisuras de las rocas

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y los huecos de los árboles, aunque también se ha adaptado a las construcciones humanas. Los hábitats de caza son diversos y caza a unos 5–15 m del suelo. Se alimenta de una amplia variedad de insectos con una gran proporción de coleópteros, lepidópteros y dípteros (Ibáñez, 2007). Aunque tiene una vasta distribución, se ha observado que las poblaciones tienen una elevada variación genética, lo que se traduce en diferentes grados de diferenciación a escala local y la formación de metapoblaciones, un fenómeno que está motivado principalmente por la acusada filopatría de las hembras (Juste et al., 2009). Zona de estudio y trabajo de campo La Región de Murcia se encuentra localizada en el sureste de la península ibérica y tiene una extensión de 11.314 km2, con un rango de altitudes entre 0 y 2.027 m s.n.m. El clima en gran parte de su superficie es de carácter mediterráneo semiárido, con escasas precipitaciones (promedio de 300–350 mm/año) y temperaturas cálidas (media anual de aproximadamente 18ºC; Alonso, 2007). El paisaje ha sufrido una substancial modificación antrópica: los bosques están compuestos principalmente por pino carrasco (Pinus halepensis) de repoblación y los valles y los cursos de agua se han usado para los cultivos de regadío en huertas tradicionales. Las zonas áridas están ocupadas principalmente por cultivos de secano (cereales, vid, almendros y olivos) y vegetación xerofítica. En los últimos decenios, con la expansión de la agricultura intensiva, se han reemplazado los cultivos de secano por cultivos de regadío (principalmente hortalizas y frutales) y se han urbanizado muchas zonas de cultivos tradicionales de huerta (Martínez–Fernández et al., 2000, 2012). Para satisfacer la creciente demanda de recursos hídricos, se han construido cientos de balsas de riego y se han alterado drásticamente los ríos y sus riberas (canalización, pérdida de vegetación de ribera o redistribución de flujos). La base de datos sobre la distribución de Eptesicus isabellinus se elaboró a partir de los muestreos realizados entre 2005 y 2009 en el contexto de un trabajo más amplio sobre la distribución de los quirópteros de la zona de estudio (Lisón et al., 2010). En el trabajo de campo se utilizaron diferentes metodologías: inspección de refugios, trampeo (redes japonesas y trampas en arpa) y empleo de detectores de ultrasonidos. El uso combinado de los datos obtenidos con diferentes procedimientos es una práctica común en estudios de modelos de nicho similares (Sattler et al., 2007; Rebelo & Jones, 2010; Rutishauser et al., 2012; Lisón & Calvo, 2013). El muestreo estaba diseñado para cubrir el mayor rango posible de condiciones ambientales y usos del suelo de la zona de estudio. En total, se analizaron 356 puntos de muestreo, incluidos 76 refugios (16 edificios y 60 cajas para murciélagos), 25 sitios trampeados y 255 puntos de grabación de ultrasonidos (2.550' de grabación). Los puntos de grabación se ubicaban en zonas no urbanizadas y en ausencia de iluminación artificial, para evitar el sesgo producido por su efecto de atracción sobre los insectos. Los individuos capturados fueron

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identificados como pertenecientes al género Eptesicus usando las claves de identificación de Dietz & Von Helversen (2004). Además, a los individuos capturados se les realizaba una pequeña biopsia en la membrana alar para obtener material genético, que posteriormente se enviaba a la Estación Biológica de Doñana para confirmar la especie (Javier Juste, com. pers.). Todas las muestras enviadas (n = 8) pertenecieron a E. isabellinus. Análisis de sonido Las grabaciones de las llamadas de ecolocación de los murciélagos se realizaron entre los años 2005 y 2008 con un detector Tranquility Transect (David J. Bale, UK) acoplado a una grabadora digital (Olympus VN–960PC, Olympus Imaging Corp., Tokyo, Japón; frecuencia de muestreo 22,5 kHz y 16 bits/muestra) y entre 2009 y 2010 con un detector de ultrasonidos Petterson D1000X (Petterson Elektronic, AB, Uppsala, Suecia, frecuencia de muestreo 300 kHz, 2 minutos de grabación y modo automático). Cada punto de escucha tenía una duración de 10 minutos. Las llamadas digitalizadas se analizaron con un programa específico de análisis de sonido (Batsound 4.03, Petterson Elektronic, AB, Uppsala, Suecia, frecuencia de muestreo 44,1 kHz y 16 bits/muestra). La identificación de las llamadas se basó en seis parámetros del pulso de ecolocación, a saber: el tipo de pulso, la frecuencia inicial, la frecuencia final, la frecuencia de máxima energía, la duración del pulso y el intervalo entre pulsos (Ahlén, 1990; Russo & Jones, 2002; Obrist et al., 2004; Papadatou et al., 2008). Las llamadas de ecolocación de Eptesicus pueden confundirse con las emitidas por Nyctalus leisleri (Kuhl, 1817), pero dicha especie no se encuentra en la zona de estudio (Agirre–Mendi, 2007; Lisón et al., 2011). Modelo de nicho ecológico y mapa de idoneidad de hábitat El mapa de hábitat idóneo se elaboró mediante un análisis factorial del nicho ecológico (Ecological Niche Factor Analysis, ENFA; Hirzel et al., 2002) utilizando el programa BIOMAPPER 4.0 (Hirzel et al., 2008). Este análisis relaciona datos únicamente de presencia de la especie con un conjunto de variables ecogeográficas. Esta técnica minimiza los errores relacionados con las “falsas ausencias”, que son inherentes al uso de información de presencia y ausencia (Hirzel & Le Lay, 2008). Los MNE se han concebido como una herramienta para la biología de la conservación y sus análisis se han utilizado en numerosos estudios (p. ej. Sattler et al., 2007; Rutishauser et al., 2012; Lisón & Calvo, 2013; Lisón et al., 2013; Sánchez–Fernández et al., 2013). En el presente trabajo se ha utilizado como sistema de referencia una rejilla UTM de 500 x 500 y se han considerado las tres categorías de descriptores ambientales siguientes: Topográficos Se han utilizado tres variables cuantitativas continuas (la altitud media, la pendiente media y la pendiente

máxima) que se obtenían a partir del modelo digital del terreno (20 m de resolución; Instituto Geográfico Nacional). Ecológicos Se ha empleado el índice normalizado diferencial de la vegetación (Normalized Difference Vegetation Index; NDVI) usando imágenes LANDSAT (sensor ETM+) de primavera (1 de marzo de 2000) y verano (26 de julio de 2001). El NDVI es un índice espectral relacionado con la productividad primaria (Tucker & Sellers, 1986) que adquiere valores entre –1 y +1. Una elevada productividad se corresponde con valores positivos elevados. Usos del suelo Se han empleado ocho variables de usos del suelo que se habían obtenido en el proyecto CORINE Land Cover 2000 de la Agencia Europea de Medio Ambiente. Las categorías incluidas eran: bosque, matorral, cultivos de regadío, cultivos de secano, cursos de agua, humedales, usos humanos y zonas mineras. Las variables de usos del suelo se estimaron con el módulo CircAn de la aplicación BIOMAPPER 4.0 (Hirzel et al., 2008) escogiendo una ventana circular de dos kilómetros de radio para calcular la frecuencia de presencia de cada clase de usos del suelo. Se empleaba este rango de distancia ya que en diferentes estudios de rastreo por radio se ha sugerido que los murciélagos hortelanos tienen un zona de campeo dentro de esta distancia, aunque con algunas variaciones (Dietz et al., 2009), y se califica a la especie como sedentaria (Ibáñez, 2007; Juste et al., 2009; Papadatou et al., 2011). Todas las variables fueron normalizadas con la transformación Box–Cox. El MNE proporciona un mapa ráster de idoneidad de hábitat, donde cada celda tiene un valor único que se encuentra entre 0 (menos idónea) y 100 (más idónea), en el cual se utilizaron los seis primeros factores del análisis (88,7% de varianza explicada) que se seleccionaron mediante el criterio broken– stick proporcionado por BIOMAPPER 4.0 (Hirzel et al., 2008). Posteriormente, el mapa se reclasificó en cuatro clases de idoneidad de hábitat (inadecuado, marginal, idóneo y óptimo), de acuerdo con un criterio de maximización del índice de Boyce B4 (Boyce et al., 2002; Hirzel et al., 2006; Sattler et al., 2007). El índice de Boyce varía entre –1 y +1. Los valores positivos indican que las predicciones del modelo se ajustan de manera coherente con la distribución de las presencias de la especie. Los valores próximos a 0 indican que el modelo escogido no difiere de uno elegido al azar, mientras que los valores negativos indican un mal ajuste a la distribución de presencias de la especie. El análisis del MNE proporciona además un índice global de marginalidad y otro de especialización. La marginalidad se define como la diferencia encontrada entre el nicho de la especie y la disponibilidad del hábitat en el zona estudiada, mientras que la especialización se describe como la varianza del hábitat ocupado por los individuos en relación con la varianza total (Hirzel et al., 2002).

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España

Presencia Ausencia Principales cursos de agua Embalses

0 4 8

32 km

Fig. 1. Mapa de la Región de Murcia donde se muestran los puntos de muestreo con presencia (circulo negro) y ausencia (cruz) de Eptesicus isabellinus y la distribución de los principales cursos de agua y de los embalses. Fig. 1. Map of the Region of Murcia showing the sampling points with presence (black circle) or absence (cross) of Eptesicus isabellinus, and the distribution of the main water courses and ponds.

Dado que algunos autores han sugerido que la autocorrelación espacial es una fuente importante de sesgo en los análisis espaciales (Segurado et al., 2006; Veloz, 2009), se ha calculado el índice de Moran para estimar dicha autocorrelación. Los valores de índice de Moran (I = 0,678, P = 0,1385) encontrados para nuestros datos indican que no existió autocorrelación espacial. Estos análisis se realizaron con el paquete estadístico R versión 2.15.1 (R Core Team, 2012). Resultados El muestreo de campo proporcionó un total de 86 datos de presencia de E. isabellinus en los 356 lugares analizados (fig. 1). El MNE indica que la marginalidad de la especie es baja (M = 0,569), aunque el grado de especialización es elevado (S = 1,430) y la tolerancia, baja (1/S = 0,699), lo que sugiere que los hábitats preferidos por la especie son abundantes en la zona de estudio. La especie muestra gran afinidad por estos hábitats preferidos y hace un uso exclusivo de los mismos a pesar de tener otros hábitats a su disposición. Los resultados del modelo muestran

que la especie tiene una gran afinidad por los cursos de agua (tabla 1), aunque influyen de manera significativa la presencia de zonas con matorral, una elevada productividad durante el verano, así como la pendiente máxima. La presencia de cultivos de secano y la altitud media afectan de manera negativa a la especie. Las actividades humanas, el bosque, los humedales y los cultivos de regadío tienen un impacto positivo moderado. El mapa de idoneidad de hábitat (fig. 2) presenta un índice de Boyce muy elevado (0,863 ± 0,151), lo que indica que la precisión del modelo es satisfactoria. Los porcentajes de hábitat inadecuado, marginal, adecuado y óptimo son 2,1%, 49,1%, 30,4% y 18,4%, respectivamente. Los valores de las variables ambientales para cada uno de los hábitat tipo (tabla 2) muestran que los hábitats adecuados y óptimos se encuentran en altitudes medias alrededor de los 500 m con una buena pendiente y que están compuestos por un mosaico de zonas de matorral y bosque. También aparece una importante proporción de cultivos de regadío. La distribución de estos hábitats favorables para la especie no se produce a lo largo de los cursos de agua, sino que aparecen concentrados en determinados tramos, especialmente

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Tabla 1. Correlación entre los factores del MNE y las variables ecogeográficas empleadas para caracterizar la distribución de Eptesicus isabellinus en Murcia. Se muestran los seis primeros factores utilizados para la elaboración del mapa de idoneidad de hábitat. Los porcentajes indican la cantidad de especialización representada por ese factor. Se indica el autovalor de cada factor (λ). Las variables ambientales están ordenadas en función de su valor de correlación con el primer factor: a Factor de marginalidad (explica el 100% de la marginalidad, los valores positivos indican que la especie tiene preferencia por dicha variable y los valores negativos indican lo contrario); b Factor de especialización (los valores positivos indican que la especie ocupa un estrecho rango de valores de los disponibles). Table 1. Correlation between the ENFA factors and ecogeographic variables used to characterize the distribution of Eptesicus isabellinus in Murcia. The table shows the first six factors used to create the habitat suitability map. The percentages indicate the quantity of specialization for each factor. The eigen value for each factor is shown (λ). Environmental variables are shown in function of their correlation with the first factor: a Marginality factor (explaining 100% of marginality, positive values indicate that the species has a preference for this variable and negative values indicate the opposite); b Specialisation factor (positive values show that the species occupies a narrow range of the studied variable).

Factor 1a (10.2%) λ1 = 2.922

Factor 2b (27.8%) λ2 = 7.955

Factor 3b (13.2%) λ3 = 3.782

Factor 4b (10.6%) λ4 = 3.044

Factor 5b (8.7%) λ5 = 2.487

Factor 6b (6.9%) λ6 = 1.964

Cursos de agua

0.571

0.189

–0.104

0.117

–0.143

–0.098

Matorral

0.347

0.335

–0.057

–0.015

–0.261

0.270

Pendiente máxima

0.308

–0.161

–0.018

–0.290

–0.335

–0.202

NDVI–Verano

0.307

–0.081

0.572

–0.332

0.293

0.079

NDVI– Primavera

0.246

0.106

–0.420

0.348

–0.024

–0.602

Pendiente media

0.161

–0.010

0.374

0.398

0.380

–0.136

Actividades humanas

0.152

–0.073

0.128

0.032

–0.113

–0.222

Bosque

0.151

0.137

–0.092

–0.200

0.137

0.282

Actividades mineras

0.134

–0.046

–0.036

–0.017

0.102

–0.011

Humedales

0.121

–0.691

–0.201

0.053

–0.036

–0.167

Cultivo de regadío

0.056

0.274

–0.331

–0.492

0.050

–0.237

Altitud media

–0.297

0.015

–0.207

–0.384

–0.674

–0.357

Cultivos de secano

–0.320

0.220

0.268

0.227

0.246

–0.334

en los que hay un embalse. Las zonas próximas a cursos de agua que están muy urbanizadas o tienen un uso agrícola intensivo constituyen un hábitat marginal para la especie. El hábitat inadecuado para la especie está compuesto principalmente por zonas dedicadas al cultivo de secano o de regadío. Discusión Desde hace tiempo se conocen las peculiaridades de los murciélagos hortelanos del sur de la península ibérica (Cabrera, 1914) con respecto a las poblaciones del norte e incluso se llegó a proponer la especie E. boscai Cabrera 1914, con holotipo en Muchamiel (Alicante), muy próxima a la zona de estudio (Juste et al., 2013). La principal característica diferencial era la coloración del pelaje (Ibáñez, 2007). Posteriormente,

se consideró que las poblaciones africanas eran una subespecie de E. serotinus (Harrison 1963), aunque más tarde volvieron a clasificarse como especie (Benda et al., 2004). Los estudios moleculares realizados en la península revelaron que se encontraban diferencias significativas entre las poblaciones de Eptesicus del norte y del sur, y estas últimas se asignaron a E. isabellinus (Ibáñez et al., 2006). La presencia de E. isabellinus en los territorios peninsulares y su carácter iberoafricano hacen que sea necesario realizar otros estudios que aporten conocimiento sobre su ecología y selección de hábitat. Los resultados del presente estudio sugieren que la presencia de cursos de agua tiene un fuerte impacto en la distribución de E. isabellinus, algo que siempre han destacado los investigadores del género Eptesicus en España (Cabrera, 1914; Balcells, 1963; Ibáñez, 2007). La elección de este tipo de hábitat

Lisón et al.

España

Inadecuado Marginal Adecuado Óptimo

0 4 8

32 km

Fig. 2. Mapa de idoneidad de hábitat para la especie Eptesicus isabellinus en la Región de Murcia obtenido a partir de los seis primeros factores del análisis factorial de nicho ecológico. Fig.2. Habitat suitability map for Eptesicus isabellinus in the Region of Murcia, obtained from the first six factors in the ecological niche factor analysis.

puede estar motivada por su uso como lugares de caza o bien como marcas del paisaje que conectan diferentes hábitats de caza (Verboom & Huitema, 1997; Ibáñez, 2007; Lisón & Calvo, 2011). Ciertos estudios realizados en la península ibérica han mostrado que la distancia a los cursos de agua es una variable predictora importante para E. isabellinus (Santos et al., en prensa). Asimismo, la preferencia por zonas con una elevada pendiente puede estar motivada por el hecho de que la especie utiliza como refugios naturales las grietas de los acantilados (Ibáñez, 2007; Juste et al., 2009). El murciélago hortelano meridional parece evitar las zonas llanas o bajas y se suele encontrar por encima de los 300 metros sobre el nivel del mar (Lisón & Calvo, 2011; Lisón et al., 2011). Los resultados de nuestro modelo muestran que la especie tiene especial preferencia por las zonas en las que pueda encontrar agua y que tengan abundante matorral, mientras que evita los cultivos de secano, algo que también se ha observado en Andalucía (Ibáñez, 2007). Posiblemente, estas zonas de matorral próximas a cursos de agua son zonas ideales donde encuentra una gran variedad de presas que componen su dieta en la zona de estudio (principalmente insectos voladores como

lepidópteros, coleópteros y dermápteros; F. Lisón, datos no publicados). También observamos una cierta preferencia por las zonas con una elevada producción primaria durante el verano, lo que puede estar relacionado con la época de cría de la especie en el sur peninsular (Ibáñez, 2007). En determinados estudios con modelos llevados a cabo en el conjunto de la península ibérica (Santos et al., en prensa) se encontró que la distribución de la especie está íntimamente ligada a la proximidad de plantaciones de eucaliptos. Curiosamente, en la Región de Murcia no existen este tipo de plantaciones y los bosquetes de eucaliptos son de escasas dimensiones o están formados por pies aislados. El estudio de Santos et al. (en prensa) incorporaba pocos datos relativos a nuestra zona de estudio. Por otro lado, a diferencia del elevado grado de asociación de la especie con las viviendas humanas que señalan diversos autores (Ibáñez, 2007; Dietz et al., 2009; Papadatou et al., 2009), los resultados de nuestro trabajo ponen de manifiesto que la presencia de actividades humanas o de cultivos de regadío no es especialmente importante para E. isabellinus. La preferencia de la especie por zonas próximas a cursos de agua, íntimamente ligados a la presencia

Animal Biodiversity and Conservation 37.1 (2014)

Tabla 2. Caracterización ambiental del mapa de idoneidad de hábitat de E. isabellinus en la Región de Murcia. Los valores de las variables topográficas y ecológicas representan el valor medio para cada clase. Los valores de las variables de usos del suelo representan el porcentaje de cada uso para cada clase de hábitat. Table 2. Environmental characterization of the habitat suitability map of E. isabellinus in the Region of Murcia. The values for the topographical and ecological variables represent the mean value for each type. The values of the land use variables represent the percentage of use for each type of habitat.

Variable

Óptimo

Adecuado

Marginal

Inadecuado

Altitud media (m)

542.7

494.8

482.3

629.2

Pendiente media (%)

10.1

7.5

4.6

4.5

Pendiente máxima (%)

12.6

9.9

7.1

6.9

NDVI–Primavera

0.10

0.07

0.05

0.06

NDVI–Verano

–0.19

–0.20

–0.24

–0.26

Actividades humanas

2.65

3.35

2.37

0.86

Cultivo de secano

2.78

5.56

23.02

48.43

Cultivo de regadío

20.17

37.88

51.78

38.33

Bosque

22.77

14.98

4.08

1.64

Matorral

51.09

37.57

18.34

9.32

Humedales

0.06

0.24

0.14

1.42

Cursos de agua

0.01

0.02

0.06

0.00

Actividades mineras

0.46

0.40

0.20

0.00

Topográficas

Ecológicas

Usos del suelo (%)

humana, puede haber motivado que de forma indirecta la especie comenzara a utilizar las casas para refugiarse. Sin embargo, se ha observado que E. isabellinus abandona las ciudades en bandadas al anochecer para dirigirse a las zonas de caza (Lisón et al., 2011), comportamiento que se ha observado en otras especies del género (Dietz et al., 2009). No obstante, las diferencias encontradas entre nuestros modelos y los datos publicados por otros autores pueden deberse a que se evitó incorporar al modelo datos procedentes de zonas urbanas. El mapa de idoneidad del hábitat muestra que los hábitats óptimo y adecuado para la especie se encuentran distribuidos a lo largo de toda la zona de estudio en zonas que se encuentran en torno a los principales cursos de agua, especialmente donde se localizan embalses. Los usos del suelo de estos hábitats parecen englobar un mosaico de zonas boscosas y de matorral en las que también aparecen cultivos de regadío de carácter tradicional. Además, muestra que E. isabellinus tiene marcadas preferencias por las zonas algo alejadas de los núcleos urbanos o con agricultura intensiva, a pesar de que con frecuencia se le califica como un murciélago antrópico (Ibáñez, 2007; Dietz et al., 2009).

La expansión de las zonas urbanas, la reducción de las zonas agrícolas tradicionales, la pérdida de matorral y el uso intensivo del suelo pueden perjudicar seriamente a sus poblaciones y a la conservación de la especie, además del posible efecto sinérgico que suponen las nuevas construcciones, que no permiten que la especie se refugie en ellas. En la zona de estudio se ha registrado la desaparición de numerosas colonias de cría por cerramientos inadecuados realizados por los propietarios (F. Lisón, datos no publicados). Posiblemente su grado de conservación haya podido pasar desapercibido hasta ahora, debido a su asociación con las construcciones humanas, a su fácil detectabilidad y a su consideración de murciélago urbano. Es necesario realizar estudios más específicos sobre este endemismo ibérico y como medidas de conservación deberían ejecutarse programas de educación ambiental para evitar el cerramiento de los refugios urbanos, potenciar la presencia de refugios artificiales y poner en marcha programas de traslado de las colonias urbanas para evitar problemas zoonóticos. Además, debería estudiarse su función como controlador de plagas y fomentar la reducción del uso de insecticidas inespecíficos.

El uso de los modelos de nicho ecológico se ha revelado como una herramienta útil para ampliar el conocimiento de muchas especies, en especial de los murciélagos (Sattler et al., 2007; Rebelo et al., 2009; Rebelo & Jones, 2010; Lisón & Calvo, 2013; Lisón et al., 2013; Santos et al., en prensa), y es especialmente interesante para analizar la distribución y los requerimientos ambientales de especies crípticas y especies endémicas. El mapa de idoneidad de hábitat elaborado para E. isabellinus en la Región de Murcia indica que una gran proporción del territorio representa hábitats favorables para la especie y aumenta el conocimiento sobre los requerimientos ecológicos de la misma. Agradecimientos Expresamos nuestro agradecimiento a Emilio Aledo, a los técnicos de la Dirección General de Patrimonio Natural y Biodiversidad de la Región de Murcia, al personal de la red de Espacios Naturales Protegidos y a los agentes medioambientales, por facilitar el trabajo de campo y la localización de algunos refugios. También agradecemos la colaboración del grupo de voluntariado de Espuña y Calblanque y en concreto de F. Almansa, N. D. Yelo y E. Martínez. Agradecemos la colaboración de J. A. López, F. J. López, S. Jiménez, M. L. Ortega, J. A. Palazón, A. Martínez y C. González. Los análisis de ADN fueron realizados por cortesía de J. Juste (Estación Biológica de Doñana). Gracias a SGS–TECNOS SA por el soporte técnico para la realización de este trabajo, con especial cariño a A. L. del Saz, A. Pérez, G. Romero, R. Cano y A. Borreguero. Agradecemos al Dr. Mario Díaz y al Dr. Javier Juste sus valiosos comentarios, que han mejorado sustancialmente el manuscrito original. Este estudio ha sido financiado en parte por el proyecto 102/08 de la Comunidad Autónoma de la Región de Murcia (Fondos FEDER). Referencias Agirre–Mendi, J., 2007. Nyctalus leisleri (Kuhl, 1817). In: Atlas y Libro Rojo de los Mamíferos Terrestres de España: 222–227 (L. J. Palomo, J. Gisbert & J. C. Blanco, Eds.). Dirección General para la Biodiversidad–SECEM–SECEMU, Madrid. Ahlén, I., 1990. Identification of bats in flight. Swedish Society Conservation of Nature & the Swedish Youth Association for Environmental Studies and Conservation. Katarina tryck Ab Press, Stockholm. Alonso, F., 2007. El clima. In: Atlas global de la Región de Murcia: 146–155 (A. Gómez, A. Romero & F. Alonso, Eds.). Murcia, La Verdad–CMM, S. A. Balcells, E., 1963. Datos españoles de Plecotus y Eptesicus (Chir. Vespertilionidae). Miscelania Zoológica, 1: 147–162. Barros, P., 2011. Contribución al conocimiento de la distribución de quirópteros en el norte y centro del Portugal. Barbastella, 5: 19–31. Benda, P., Ruedi, M. & Aulagnier, S., 2004., New data

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Animal Biodiversity and Conservation 37.1 (2014)

Habitat use pattern and conservation status of smooth–coated otters Lutrogale perspicillata in the Upper Ganges Basin, India M. S. Khan, N. K. Dimri, A. Nawab, O. Ilyas & P. Gautam

Khan, M. S., Dimri, N. K., Nawab, A., Ilyas, O. & Gautam, P., 2014. Habitat use pattern and conservation status of smooth–coated otters Lutrogale perspicillata in the Upper Ganges Basin, India. Animal Biodiversity and Conservation, 37.1: 69–76. Abstract Habitat use pattern and conservation status of smooth–coated otters Lutrogale perspicillata in the Upper Ganges Basin, India.— Smooth–coated otters inhabit several major river systems in southern Asia, and their environmental requirements link them to food and water security issues as the region is so densely populated by humans. The lack of baseline data on their distribution and ecology is another major constraint that the species is facing in India. The present study was stimulated by the rapid decline in the otter’s population in the country and focuses on estimating the conservation status, habitat use pattern, and associated threats in the upper Ganges River Basin (N India). Our findings contribute towards a better understanding of the complex ecological interactions and the design of effective conservation measures. Coupled with the habitat preferences, the study also provides new locations in the species distribution. This paper highlights the gap areas in the conservation of the species and suggests areas that should be prioritized for management. Key words: Otter, Ganges Basin, Conservation status, Habitat use. Resumen Modelo de uso del hábitat y estado de conservación de las nutrias lisas Lutrogale perspicillata en la zona alta de la cuenca del Ganges, India.— Las nutrias lisas habitan en varios sistemas fluviales importantes del Asia meridional y sus necesidades medioambientales las vinculan con problemas de seguridad alimentaria e hídrica, debido a la elevada densidad de humanos. La falta de datos de referencia sobre su distribución y ecología es otra limitación notable que la especie está afrontando en la India. El presente estudio se vio impulsado por el rápido descenso de la población de nutrias en el país y se centra en estimar el estado de conservación, el modelo de uso del hábitat y las amenazas asociadas en la zona alta de la cuenca del río Ganges (Asia septentrional). Nuestros resultados contribuyen a comprender mejor las complejas interacciones ecológicas y a elaborar medidas de conservación eficaces. Junto con las preferencias de hábitat, en el estudio también se informa sobre nuevas ubicaciones en la distribución de la especie. Asimismo se ponen de relieve las deficiencias existentes en la conservación de la especie y se sugieren las zonas cuya ordenación debería ser prioritaria. Palabras clave: Nutria, Cuenca del Ganges, Estado de conservación, Uso del hábitat. Received: 28 I 14; Conditional acceptance: 24 III 14; Final acceptance: 29 IV 14 Mohd. Shahnawaz Khan & Asghar Nawab, WWF India, 172 B Lodi Estate, New Delhi, 110 003 (India).– Nand Kishor Dimri, WII, 18, Chandrabani, Dehradun, Uttarakhand, (India).– Orus Ilyas, AMU, Aligarh, Uttar Pradesh 202 002 (India).– Parikshit Gautam, FES, NDDB House PB. 4906 Safdarjung Enclave, New Delhi, 110 029 (India).

ISSN: 1578–665 X eISSN: 2014–928 X

Introduction Natural floodplains are biologically the most productive and diversified ecosystems on earth (Mitsch & Gosselink, 2000) but due to their very slow recovery they are also the most threatened (Vitouesk et al., 1997; Ravenga et al., 2000). The Ganga River Basin is among the world’s largest productive floodplain ecosystems with enormous ecological, cultural and economical value (Ambastha et al., 2007). It has an extraordinary variety in altitude, climate, land use and biodiversity (O’Keeffe et al., 2012) The entire span of the Ganga River Basin in India can be divided into three stretches i.e. the upper reach from the origin to Narora, the middle reach from Narora to Ballia, and the lower reach from Ballia to its delta. The upper Ganga River Basin is a dynamic, bio– spatial complex eco–region. The natural landscape has been severely fragmented by anthropogenic factors and most of the wildlife endowments are restricted either to the Shivalik hills and their adjacent Bhabar–Terai tract or to protected areas (Rodgers & Panwar, 1988). These pockets in the upper Ganga River Basin provide refuge to some threatened populations of endangered aquatic and semi–aquatic mammalian species like the Ganges river dolphin Platanista gangetica and the smooth–coated otter Lutrogale perspicillata, respectively. The amphibious life styles of otters allow them to disperse over wide areas of riverine landscape, and as a result, they influence the ecological processes of the river floodplain in a direct and expansive manner. Smooth–coated otters play a vital role in balancing the freshwater ecosystems as a top carnivorous species (Sivasothi, 1995; Acharya & Lamsal, 2010), and they may therefore significantly influence the overall spatio–temporal dynamics of the eco–region over a long period of time (Naiman et al., 2000). There is little information available on the status of otter populations in India, although there seems to have been a rapid decline due to loss of habitat and intensive trapping (Hussain, 1999; Nawab, 2007, 2009; Nawab & Gautam, 2008). Presently, the population is severely fragmented throughout its distribution range and isolated populations are restricted mostly to protected areas (Hussain, 1999; Nawab, 2007, 2009). Although otter occurrence in the upper Ganga River Basin has been previously reported from the National Chambal Wildlife Sanctuary (Hussain, 1993), Corbett Tiger Reserve (Nawab, 2007), Dudhwa Tiger Reserve and Katerniaghat Wildlife Sanctuary (Hussain, 2002), the present study appends new geographical locations in the distribution range of smooth–coated otter, i.e. (i) Alaknanda–Ganga Basin in Uttarakhand and (ii) Hastinapur Wildlife Sanctuary in Uttar Pradesh. The present study was triggered by the rapid decline in the otter’s population in the country and it focuses mainly on assessing the otter’s conservation status, its habitat use pattern, and associated threats in the upper Ganga River Basin (N India). This will improve the understanding of the complex ecological interactions and will help to design effective conservation measures for this species (Stanford et al., 1996). The

Khan et al.

purpose of this paper is to highlight the gap areas in the conservation of the species and to suggest areas for management in the upper Ganga River Basin. Material and methods Study sites The Ganga River Basin is the largest river basin in India, constituting 26% of the country’s land mass and supporting about 43% of its population (448.3 million as per the 2001 census) (Ambastha et al., 2007). Rainfall and melt water from snow and glaciers are the main sources of water in the River Ganga (O’Keeffe et al., 2012). The present study was carried out at two selected sites, one in Uttar Pradesh and the other in Uttarakhand, states of India where the species has not been studied previously. Site I. Alaknanda–Ganga Basin (from Rudraprayag to Rishikesh) The River Alaknanda originates from the confluence of the Sathopanth and Bhagirathi Kharak Glacier and forms a unified stream of the upper Ganga River by merging with the River Bhagirathi at Devprayag. The Alaknanda–Ganga Basin (fig. 1) is characterized by rugged topography with major landforms comprising moderate to steep precipitous sloping mountainous terrain, narrow and broad valleys and highly dissected ridges with the formation of deep gorges (Anbalagan et al., 2008). Despite its unprotected status, the basin holds a good variety of wildlife, including endangered freshwater fauna like Golden Mahasheer Tor putitora. The general vegetation in the area is dominated by Pinus roxburghii, Anogeissus latifolius, Acacia catechu, Holoptelea integrifolia, Syzgium cumini and Aegle marmelos. The drainage system of the basin has been extensively regulated for hydroelectric production. Site II. Hastinapur Wildlife Sanctuary Hastinapur Wildlife Sanctuary spreads over an area of 2,073 km² along the banks of the River Ganges in western Uttar Pradesh (fig. 1). The Sanctuary was established in 1986 to conserve the fast vanishing, unique Ganga River grassland–wetland complex, locally known as Khadar. It is unique in the sense that it presents a variety of landforms and habitat types that include wetland, marshes, dry sandy beds and gently sloping ravines. River Ganga and its old bed, locally called Boodhi Ganga, forms the drainage system of the Sanctuary. River Ganga enters the Sanctuary area at Bijnor and leaves it at Garmukteshwer after flowing for 125 km. During summers, Boodhi Ganga becomes fragmented into a series of small swampy patches with nil or very insignificant water current. Because of this discontinuous belt of highly marshy land, there is profuse growth of vegetation like Phragmites species, Arundinella species and Typha species.

Animal Biodiversity and Conservation 37.1 (2014)

Site I. Alaknanda–Ganges Basin (from Rudraprayag to Rishikesh)

nda

kna

Ala

Uttarakhand Rishikesh

Haridwar

Site II. Hastinapur Wildlife Sanctuary

Haryana

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Bhagalpur Farakka Barrage

Jharkhand

oda

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Rih

Pad

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arm

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Madhya Pradesh

Bihar

Patna ne So

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Varanasioti

Mahananda

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Allahabad ns To

D ha san

l ba

Jamni

Parbati

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Uttar Pradesh Kampur

Ka li Ya Mathura mu na Agra Rin l d ba am Ch d Sin Betwa

B an as

Rajasthan

y Bak

Kauriyala

Delhi

Lal

Bhagirathi

and

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West Bengal Kas

Bay of Bengal Source: The original map is downloaded from www.sandrp.in on 20 January 2014

Fig. 1. Location of study sites in the Ganges River Basin. Fig. 1. Ubicación de las localidades de estudio en la cuenca del río Ganges.

Data collection

Data analysis

During the summer in 2010 we surveyed 35 kilometers of the River Alaknanda–Ganges (sampling sections, n = 7) and 145 km stretch of River Ganga (main stream) and its old bed Boodhi Ganga (sampling sections, n = 29). The selected river stretches were divided into 5 km sections using a Survey of India’s 1:50,000 topographic maps (Macdonald & Mason, 1983; Kruuk et al., 1994; Hussain & Choudhury, 1997; Nawab, 2007). Data on the habitat parameters and indirect evidences of otter occurrence such as tracks, spraints, den sites or scent marks were recorded from each section. Searches were made in 15 m wide strips along the edge of the river with the help of two trained researchers, by walking along both banks. In each study section, any location where spraints, tracks, den sites and other signs of otter presence were found was defined as a 'used plot' with dimensions 100 × 15 m; additionally, for each used plot, two available plots, one each at 500 m downstream as well as upstream, were considered. In case of spraint sites, a new site was registered only when spraints were separated by more than 5 m (Melquist & Hornocker, 1983; Newman & Grifﬁn, 1994; Medina, 1996; Nawab, 2007). At each section habitat parameters and human activities which are considered potentially threatening to otters were also recorded (Prenda & Granado–Lorencio, 1996; Prenda et al., 2001; Anoop & Hussain, 2004) (table 1). Species habitat selection was analyzed at plot scale.

The present study was based on the premise that otters live at low densities and are shy and often nocturnal or crepuscular, and hence difficult to track and to make direct estimates of population size and density. The distribution and frequency of occurrence of spraints and tracks were considered as the index of habitat use by the otters. The preference of habitat covariates was established following Bonferroni confidence intervals in combination with Chi–square goodness of fit test (Neu et al., 1974; Byers et al., 1984). Bonferroni confidence interval equation: Pi – Za/2k/Pi (1 – Pi) / n Pi Pi + Za/2k/Pi (1 – Pi) / n where Pi is the proportion of indirect evidences in the ith habitat category, n is the sample size, k is the number of categories of habitat studied, α is confidence interval while Z is the tabular value of standard curve. Chi–Square equation:

x = 2

3 (Oi – Ei)2 Ei

where Oi is the observed number of indirect evidence in the ith habitat category and Ei is expected number of indirect evidence in the ith habitat category. An independent sample t–test was performed to know the significance of difference between the used and available habitat covariates following Neu et al. (1974), Byers

Khan et al.

Table 1. Ecological parameters and human activities affecting the occurrence of smooth–coated otter, recorded during the study. Tabla 1. Parámetros ecológicos y actividades humanas registrados durante el estudio que afectan a la presencia de la nutria lisa.

Variable

Data type

Description and measurement details

Width of river (m)

Continuous

Distance between shorelines visually estimated

Average depth of river (m)

Continuous

The depth of the river was measured at both banks and

middle of the river and mean depth was calculated

Shoreline substrate type (%)

Approximate percentage of total area (100 m × 15 m)

Categorical

of the plot covered by rock/boulder, sand, mud, clay

or alluvial deposit was visually estimated

Water current (m/s)

The surface water velocity was calculated via floating

Continuous

ball method.

River bank slope (degree)

Measured via Clinometers

Continuous

Shoreline vegetation cover (%) Categorical

Approximate percentage of total area (100 m × 15 m)

of the plot covered by tree, shrub, herb or grass was

visually estimated

Escape distance (m)

Nearest distance from water’s edge to shoreline

Continuous

vegetation which provides cover for otter measured

by measuring tape

Disturbance (present/absent)

Presence of disturbing activities/evidences was

Binary

et al. (1984) and Zar (1984). Statistical package SPSS 7.3 (Norusis, 1994) was used for computing purposes. Results Site I. Alaknanda–Ganga Basin (from Rudraprayag to Rishikesh) The thirty–five kilometer stretch of the River Alaknanda–Ganga was divided into seven sampling sections of five kilometers. Otter occurrence was recorded only from two of these sections (i.e. 28.57% occupancy), at village Malysu and Papdasu (district Rudraprayag). Informal interviews with locals suggested occurrence of otters in the study area was common in the 1990s, but due to human disturbance, the habitat quality had declined and consequently the numbers of otters in the area had decreased. Sandy substrate was preferred over other available substrates by the species in the area (table 2). Of the 16 habitat parameters, the means of shoreline vegetation cover (P < 0.05), percentage of clay substrate (P < 0.001) and bank slope (P < 0.001) were used significantly different from their availability (table 3).

recorded at every plot

Site II. Hastinapur Wildlife Sanctuary The total 145 km stretch of River Ganga (main stream) and its old bed Boodhi Ganga was surveyed. The findings of the survey append the new locality record in the distribution range of smooth coated otter in north India. From a total of 29 sampling sections, only 6.89% (n = 2) were found occupied by otters. Interviews with locals revealed that the occurrence of otters in the sanctuary was common a decade before. However, excessive changes in land–use pattern and human disturbance led to a vast decline in habitat quality and hence the otter population also decreased. The result of Bonferroni confidence intervals indicates that smooth–coated otter prefer the most remote muddy parts of the river and avoid alluvial, sand and areas with clay as dominant substrate (table 2) as they are found adjacent to cultivated fields and easily accessible. Of the 15 parameters, the respective means of used and available plots of ten parameters were found significantly different at P < 0.001 level, while the differences between the mean of used and available plots for % sand was found significant at P < 0.05 level (table 3).

Animal Biodiversity and Conservation 37.1 (2014)

Table 2. Preference of shelter sites by the smooth–coated otter along site I and II: S. Substrate type; Pio. Proportion of total sampling plots; Oi. Number of used plots; Ei. Expected number of used plots; Pi. Proportion of indirect evidences at each sampling plot; x2. Chi–square distribution; Bonferroni. Bonferroni confidence interval proportions; C. Conclusion (+ Used more than available; – Used less than available) Tabla 2. Preferencia de la nutria lisa por los lugares de cobijo en las localidades I y II: S. Tipo de sustrato; Pio. Proporción en el total de parcelas de muestreo; Oi. Número de parcelas utilizadas; Ei. Número esperado de parcelas utilizadas; Pi. Proporción de pruebas indirectas en cada parcela de muestreo; x2. Distribución de la x2; Bonferroni. Intervalo de confianza de Bonferroni para las proporciones; C. Conclusión (+ Más utilizado de lo esperado; – Menos utilizado de lo esperado)

Pio

Site I Sand 0.29 (N = 14) Clay 0.02 (N = 1) Boulder 0.65 (N = 32) Alluvial 0.04 (N = 2) Site II Sand 0.16 (N = 71) Mud 0.49 (N = 218) Clay 0.29 (N = 126) Alluvial 0.06 (N = 27)

O i

E i

4 0 3 0

2.00 0.14 4.57 0.29

0.57 0.00 0.43 0.00

2.00 0.14 0.54 0.29

0.395 0.000 0.252 0.000

≤ ≤ ≤ ≤

Pi Pi Pi Pi

≤ ≤ ≤ ≤

0.748 0.000 0.605 0.000

+ – – –

0 18 7 0

4.02 12.33 7.13 1.53

0.00 0.72 0.28 0.00

4.02 2.16 0.00 1.53

0.000 0.667 0.227 0.000

≤ ≤ ≤ ≤

Pi Pi Pi Pi

≤ ≤ ≤ ≤

0.000 0.773 0.333 0.000

– + – –

Bonferroni

Mainly due to habitat loss and over–exploitation, the population of smooth–coated otters is declining throughout their range of distribution and the trend of population decline is expected to continue (Hussain et al., 2008). A deficiency of baseline data on the ecology of the species is another constraint for its conservation. Information on habitat selection by otters is further sketchier as compared to other aspects of their ecology (Hussain, 1996). In Europe and North America, many studies on Lutra lutra and Lutra canadensis have led to an increasing understanding of otter habitat preferences in temperate regions (Melisch et al., 1996), whereas in the case of the smooth–coated otter, availability of food, freshwater and shelter for resting, grooming and breeding are the important factors known to govern the process of habitat selection by otters (Mason & Macdonald, 1986; Kruuk, 1995; Anoop & Hussain, 2004; Nawab, 2009). In site I (Alaknanda–Ganga Basin), otters showed preference for sandy stretches in all the seasons, as these stretches provide sites for dens and grooming (Hussain, 1993); while in site II (Hastinapur Wildlife Sanctuary), the species preferred to use the muddy stretches of Boodhi Ganga which is almost inaccessible to humans and thus less disturbed. This ability of the species to adapt to diverse aquatic habitats accounts for its broad geographic distribution (Pocock, 1941). Otter occurrence was associated with shallow and calmer regions (with low water velocity) along the Gan-

ga River Basin in site I, as these conditions increase the rate of prey capture per efforts. Ease in capturing prey was interpreted to be the most important factor in selecting the habitat by the species, as also suggested by other studies (Kruuk, 1995; Anoop, 2001; Nawab, 2007; Acharya & Lamsal, 2010). Hastinapur Wildlife Sanctuary is one of the most populated and disturbed protected areas in Uttar Pradesh. As most of its land is cultivated, the area is highly accessible to humans, imposing an adverse effect on the inhabiting wildlife. Therefore, despite being a protected area, only 6.89% (n = 2) of otter occupancy was recorded in the area, far below the 28.57% (n = 2) recorded for otter occupancy at site I. Moreover, most of the animals like otters restricted themselves to the remaining inaccessible parts of the sanctuary, such as the swampy patches of the Boodhi Ganga River. Habitat features of Boodhi Ganga, such as deep waters forming pools, prey availability, presence of shoreline vegetation and gentle bank slopes, endorse the occurrence of otters. Other authors have also found a positive correlation between otter signs and the percentage of vegetation cover (Macdonald & Mason, 1983; Melisch et al., 1996; Anoop & Hussain, 2004; Nawab, 2007). Gentle bank slopes are favored by otters as they reduce energy expenditure while foraging or grooming (Kruuk, 1995). Otters are facing extreme threats by human–induced habitat destruction. The expansion of agriculture has led to the destruction of huge areas of natural habitats, including forests, grasslands and wetlands, in nearly all regions of the world (Ottino & Giller, 2004).

Discussion

Khan et al.

Table 3. Habitat variables influencing otter distribution along site I and II. (SE. Standard error) Talba 3. Variables del hﾃ｡bitat que influyen en la distribuciﾃｳn de la nutria en las localidades I y II.

Variables

Site I River bank characteristics % Alluvial % Boulder % Clay % Grass cover % Herb cover % Mud % Sand % Shrub cover % Total veg. cover % Tree cover Escape distance Slope River characteristics Average depth Average width Water current pH Site II River bank characteristics % Alluvial % Clay % Grass cover % Herb cover % Mud % Sand % Shrub cover % Total veg. cover % Tree cover Escape distance Slope River characteristics Average depth Average width Water current pH

Available plots

Used plots

Mean

10.12 55.86 5.29 19.88 17.98 1.55 27.19 30.95 28.33 9.76 7.07 50.76

1.69 4.22 1.53 2.52 2.15 0.65 4.20 4.03 3.02 1.69 0.82 3.91

7.14 55.71 0.00 22.14 22.14 1.43 35.71 46.43 47.14 9.29 5.29 14.29

4.91 28.58 1.28 7.81

0.44 2.96 0.08 0.02

9.70 32.81 86.47 6.16 41.49 16.01 0.94 31.74 0.35 48.68 14.17

Sig.

1.84 13.60 0.00 2.86 3.91 1.43 13.07 8.29 6.44 2.02 2.01 2.02

-0.70 -0.01 -3.45 0.36 0.75 -0.07 0.74 1.48 2.39 -0.18 -0.83 -8.29

0.486 0.990 0.001 0.723 0.454 0.944 0.463 0.146 0.021 0.859 0.412 < 0.001

3.06 26.14 0.93 7.83

0.74 6.24 0.23 0.02

-1.64 -0.32 -1.68 0.75

0.108 0.753 0.101 0.461

1.09 1.96 1.20 0.46 2.05 1.59 0.14 0.61 0.09 8.67 0.

0.00 25.80 94.00 4.40 74.20 0.00 1.40 65.60 0.60 2.78 9.00

0.00 7.34 1.35 0.97 7.34 0.00 0.68 2.13 0.33 0.49 0.82

8.89 0.86 -4.16 1.64 -4.30 2.46 -0.77 -13.28 -0.69 5.29 5.74

< 0.001 0.393 < 0.001 0.109 < 0.001 0.014 0.444 < 0.001 0.491 < 0.001 < 0.001

0.86 145.61 0.65

0.04 7.94 0.04

0.48 19.72 0.02

0.03 4.96 0.00

8.02 3.88 14.33

< 0.001 < 0.001 < 0.001

8.55

0.02

7.88

0.07

9.92

< 0.001

The expansion and development of urbanization and riverfront infrastructural developments, such as the construction of dams, has broken the continuum of natural habitats into small fragments (Nawab, 2007) and these patches of suitable habitat may be too small

to support a breeding pair or a functional social group. It is of note that area sensitive species (Lambeck, 1997) like otter, that have a low dispersal capacity, are unable to re窶田olonize such patches following extinction (Collinge, 1996).

Animal Biodiversity and Conservation 37.1 (2014)

Recommendations Site I. Alaknanda–Ganga Basin (from Rudraprayag to Rishikesh) Maximum evidence of otter occurrence was concentrated around the villages Malysu and Papdasu in the Rudraprayag district. These areas therefore merit special attention in terms of habitat management and protection. As evident from this study, otters are confined to small areas and the population seems to be vulnerable to anthropogenic and other stochastic disturbances. Detailed research on the population ecology of the species is necessary to implement better management practices to conserve the species in the region. Education and awareness programmes should be launched, focusing special emphasis on fishing and immigrant communities known to be involved in otter killings for meat and skin. Although otters are often in direct conflict with fishermen who view them as competitors for fish and kill them (Foster–Turley, 1992), in the Alaknanda– Ganga Basin, a tolerable association of otters and human presence was observed. From local sources we heard that otters damage nets and steal fish from the fishermen’s catch, but the conflict remains negligible; locals also appreciate the aesthetic and ecological importance of otters, accepting it within their environment and making co–existence possible. Site II. Hastinapur Wildlife Sanctuary Until the mid–twentieth century, extensive tracts of grassland–wetland complex (locally known as Khadar) harbored rich biodiversity all along the River Ganga. After India gained independence in 1947, Khadar received a large influx of Pakistani emigrants and in the following decades (i.e. 1980s) Punjabi emigrants also settled in the area, converting the Khadar into agricultural farms (Agarwal, 2009). Presently, the Hastinapur Wildlife Sanctuary is subjected to human disturbance, mainly due to large scale commercial exploitation of grasses (Phragmites), livestock grazing and illegal cultivation (Khan et al., 2003). Many swamps have been drained and converted into crop fields, or are in the process of such activity, like Boodhi Ganga. Modernised farming, i.e. unabated use of chemical fertilizers and pesticides in these agriculture fields, is deteriorating water quality (Agarwal, 2009). Indiscriminate fishing by use of gillnet, hooks and poison poses a major threat to aquatic fauna (Khan, 2010). There is a need for locals, especially fishermen and farmers, to become aware of the importance of aquatic ecosystems both for the conservation of wildlife and for their own sustenance. Local communities should be helped to obtain better educational opportunities. Otters are confined to small swampy patches of Boodhi Ganga and the population is vulnerable to anthropogenic and other stochastic disturbances in the sanctuary. The solution for their long–term survival in the sanctuary lies not only in taking stringent protection measures but also in developing and implementing long–term monitoring programs for otters along Boodhi Ganga in and around the Sanctuary. The illegal

encroachment and clearing of Boodhi Ganga that is currently in progress and encouraged by some migrant farmers severely affects the survival of the area´s wild inhabitants. The government needs to apply strict measures and stringently implement the law to prevent such illegal activities. Acknowledgements The data were collected during an M.Sc. internship of the first and second author, in the project Otter Conservation under the sponsorship of Living Ganga Programme of WWF India. We are thankful to Mr. Ravi Singh and Dr. Sejal Worah (WWF India) for constant encouragement and support. The help rendered by staff of WWF India’s Hastinapur field office is highly appreciated. We are also grateful to Dr. Anjana Pant (WWF India), Dr. Satish Kumar, Dr. Faiza Abbasi and Zarreen Syed (AMU, Aligarh) for their valuable suggestions on the manuscript. References Acharya, P. M. & Lamsal, P., 2010. A Survey for Smooth coated Otter Lutrogale perspicillata on the River Narayani, Chitwan National Park, Nepal. Hystrix Italian Journal of Mammology, 21(2): 203–207. Doi: 10.4404/Hystrix–21.2–4464 Agarwal, S., 2009. Angiosperm species diversity and ecological assessment of Hastinapur Wildlife Sanctuary, India. Ph. D. Thesis, Department of Botany, Aligarh Muslim University, Aligarh, India. Ambastha, K., Hussain, S. A. & Badola, R., 2007. Social and Economic Considerations in Conserving Wetlands of Indo–Gangetic Plains: A Case Study of Kabartal Wetland, India. Environmentalist, 27: 261–273. Anbalagan, R., Kohli, A. & Chakraborty, D., 2008. Stability Analysis of Harmony Landslide in Garhwal Himalaya, Uttarakhand State, India. Department of Earth Sciences, Indian Institute of Technology, Roorkee, India. Anoop, K. R., 2001. Factors affecting habitat selection and feeding habits of Smooth–coated Otter Lutra perspicillata in Periyar Tiger Reserve, Kerala. M. Sc. Dissertation, Saurashtra University, Rajkot, India. Anoop, K. R. & Hussain, S. A., 2004. Factors affecting habitat selection by Smooth–coated otter Lutra perspicillata in Kerala, India. Journal of Zoology, 263: 417–423. Byers, C. R., Randall, C., Stienhorst, R. K. & Krausman, P. R., 1984. Clarification of a technique for analysis of utilization availability data. Journal of Wildlife Management, 48: 1050–1053. Collinge, S. K., 1996. Ecological consequences of habitat fragmentation: implication of landscape architecture and planning. Landscape Urban Planning, 36: 59–77. Foster–Turley, P., 1992. Conservation ecology of sympatric Asian Otters Aonyx cinerea and Lutra perspicillata. Ph. D. Thesis, University of Florida. Hussain, S. A., 1993. Aspect of the ecology of Smooth–coated Otter in National Chambal Sanctuary.

Ph. D. Thesis, Centre for Wildlife and Ornithology, Aligarh Muslim University, India. – 1996. Group Size, Group Structure and Breeding in Smooth–Coated Otter Lutra perspicillata (Geoffroy) (Carnivora, Mustelidae) in National Chambal Sanctuary, India. Mammalia, 60: 289–297. – 1999. Otter conservation in India. Envis Bulletin – Wildlife and Protected Areas, 2(2): 92–97. – 2002. Conservation status of otters in the Tarai and Lower Himalayas of Uttar Pradesh, India. In: Otter Conservation – An Example for a Sustainable use of Wetlands (R. Dulfer, J. Conroy, J. Nel & A. Gutleb, Eds.). IUCN Otter Specialist Group Bulletin, 19: 131–142. Trebon, Czech Republic. Hussain, S. A. & Choudhury, B. C., 1997. Status and distribution of Smooth–coated Otter Lutra perspicillata in National Chambal Sanctuary. Biological Conservation, 80: 199– 206. Hussain, S. A., De Silva, P. K. & Mostafa Feeroz, M., 2008. Lutrogale perspicillata. In: IUCN 2013. IUCN Red List of Threatened Species. Version 2013.2. <www. iucnredlist.org> downloaded on 22 January 2014. Khan, M. S., 2010. Conservation status and habitat use pattern of Otters in Hastinapur Wildlife Sanctuary, Uttar Pradesh; India. M. Sc. Dissertation, Department of Wildlife Sciences, Aligarh Muslim University, India. Khan, J. A., Khan, A. & Khan, A. A., 2003. Structure and composition of barasingha habitat in Hastinapur Wildlife Sanctuary. Technical Report. Wildlife Society of India, Aligarh Muslim University, Aligarh: 5–7. Kruuk, H., 1995. Wild Otters – Predation and populations. Oxford University Press. Kruuk, H., Kanchanasaka, B. O’Sullivan, S. & Wanghongsa, S., 1994. Niche separation in three sympatric Otters Lutra perspicillta, L. Lutra and Aonyx cinerea. Biological Conservation, 69: 115–120. Lambeck, R. J., 1997. Focal species: a multi species umbrella for nature conservation. Conservation Biology, 11(4): 849–856. Macdonald, S. M. & Mason, C. F., 1983. Some factors influencing the distribution of Otters Lutra lutra. Mammlian Review, 13(1): 1–10. Mason, C. F. & Macdonald, S. M., 1986. Otters: ecology and conservation. Cambridge University Press, Cambridge, London. Medina, G., 1996. Conservation and status of Lutra provocax in Chile, Paciﬁc. Conservation Biology, 2: 414–419. Melisch, R., Asmoro, P. B., Kusumawardhani, L. & Lubis, I. R., 1996. The Otters of west java: A survey of their distribution and habitat use and a strategy toward a species conservation programme. PHPA/ Wetlands International–Indonesia Programme. Melquist, W. E. & Hornocker, M. G., 1983. Ecology of Otters in West Central Idaho. Wildlife Monograph, 83: 60. Mitsch, W. J. & Gosselink, J. G., 2000. Wetlands 3rd (edu.). John Wiley and Sons Inc., New York. Naiman, R. J., Bilby, R. E. & Bisson, P. A., 2000. Riparian ecology and management in the Pacific coastal rain forest. BioScience, 50: 996–1011. Nawab, A., 2007. Ecology of Otters in Corbett Tiger

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Reserve, Uttarakhand; India. Ph. D. Thesis, Forest Research Institute, Dehradun, India. – 2009. Aspects of the ecology of Smooth–coated Otter Lutrogale perspicillata Geoffroy St. Hilaire, 1826: A Review. Journal of Bombay Natural History Society, 106(1): 5–10. Nawab, A. & Gautam, P., 2008. Living on the edge: Otters in developing India. In Wetlands – The Heart of Asia. Proceedings of the Asian Wetland Symposium, Hanoi, Vietnam. Neu, C. W., Byers, C. R. & Peek, J. M., 1974. A technique for analysis of utilization – availability data. Journal of Wildlife Management, 38: 541–545. Newman, D. G. & Grifﬁn, R., 1994. Wetland use by river otters in Massachusetts. Journal of Wildlife Management, 58: 18–23. Norusis, M. J., 1994. SPSS/PC+ statistic 7.3. for 1BMPC/XT/AT and PS/2 SPSS. International Br, Netherlands. O’Keeffe, J., Kaushal, N., Smakhtin, V. & Bharati, L. 2012. Assessment of Environmental Flows for the Upper Ganga Basin. Summary Report. WWF India. Ottino, P. & Giller, P., 2004. Distribution, density, diet and habitat use of the Otter in relation to land use in the Araglin valley, Southern Ireland. Biology and Environment: Proceedings of the Royal Irish Academy, 104B(1): 1–17. Pocock, R. I., 1941. The Fauna of British India including Ceylon and Burma. Vol. II. Taylor and Francis, London. Prenda, J. & Granado–Lorencio, C., 1996. The relative influence of riparian habitat structure and ﬁsh availability on otter Lutra lutra sprainting activity in a small Mediterranean catchment. Biological Conservation, 76: 9–15. Prenda, J., López–Nieves, P. & Bravo, R., 2001. Conservation of otter (Lutra lutra) in a Mediterranean area: the importance of habitat quality and temporal variation in water availability. Aquatic Conservation: Marine and Freshwater Ecosystem, 11: 343–355. Ravenga, C., Brunner, J., Henninger, N., Kassem, K. & Payne, R., 2000. Pilot analysis of global ecosystems. Freshwater Systems. World Resources Institute, Washington D. C. Rodgers, W. A. & Panwar, H. S., 1988. Planning wildlife protected area network in India. Vol. 2 Project FO: IND/82/003. FAO, Dehra Dun. Sivasothi, N., 1995. The status of Otters in Singapore and Malaysia, and the diet of Smooth coated Otter Lutrogale perspicillata in Penang, West Malaysia. M. Sc. Thesis, National University of Singapore, Singapore. Stanford, J. A., Ward, J. V., Liss, W. J., Frissell, C. A., Williams, R. N., Lichatowich, J. A. & Coutant, C. C., 1996. A general protocol for restoration of regulated rivers. Regulated Rivers: Research and Management, 12: 391–413. Vitouesk P. M., Mooney, H. A., Lubchenco, J. & Melillo, J. M., 1997. Human domination of earth’s ecosystem. Science, 227: 494–499. Zar, J. H., 1984. Biostatistical analysis. IInd. Edn. Prentice–Hall Inc., New Jersey.

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Invertebrates outcompete vertebrate facultative scavengers in simulated lynx kills in the Bavarian Forest National Park, Germany R.–R. Ray, H. Seibold & M. Heurich

Ray, R.–R., Seibold, H. & Heurich, M., 2014. Invertebrates outcompete vertebrate facultative scavengers in simulated lynx kills in the Bavarian Forest National Park, Germany. Animal Biodiversity and Conservation, 37.1: 77–88. Abstract Invertebrates outcompete vertebrate facultative scavengers in simulated lynx kills in the Bavarian Forest National Park, Germany.— Understanding the role of scavengers in ecosystems is important for species conservation and wildlife management. We used road–killed animals, 15 in summer 2003 (June–August) and nine in winter 2003/2004 (from November to January), to test the following hypotheses: (1) vertebrate scavengers such as raven (Corvus corax), red fox (Vulpes vulpes) and wild boar (Sus scrofa) consume a higher proportion of the carcasses than in� vertebrates; (2) the consumption rate is higher in winter than in summer due to the scarcity of other food resources; and (3) vertebrate scavengers are effective competitors of Eurasian lynx. We monitored 65 animals belonging to eight different mammal and bird species with camera traps. Surprisingly, Eurasian lynx (Lynx lynx) was the most important vertebrate scavenger. However, in both seasons, the consumption of vertebrate scavengers was of minor impact. In summer, the carcasses were completely consumed within 10 days, mostly by invertebrates. In winter, only 5% of the carcasses were consumed within 10 days and 16% within 15 days. We conclude that vertebrates in the Bavarian Forest National Park are not strong competitors for lynx. Key words: Carrion, Lynx lynx, Scavenging, Kleptoparasitism, Decomposers, Food competition. Resumen Los invertebrados compiten con los vertebrados necrófagos facultativos por las presas simuladas de lince en el parque nacional del bosque de Baviera, Alemania.— Comprender la función de los necrófagos en los ecosistemas es importante para la conservación de especies y la ordenación de la fauna y la flora silvestres. Utilizamos animales que habían muerto en la carretera, 15 en verano de 2003 (de junio a agosto) y nueve en el invierno 2003/2004 (de noviembre a enero) para analizar las hipótesis siguientes: (1) los vertebrados necrófagos como el cuervo (Corvus corax), el zorro (Vulpes vulpes) y el jabalí (Sus scrofa) consumen una proporción mayor de los cadáveres que los invertebrados; (2) el índice de consumo es más elevado en invierno que en verano debido a la escasez de otras fuentes de alimentos, y (3) los vertebrados necrófagos son competidores reales del lince euroasiático. Seguimos a 65 animales que pertenecían a ocho especies diferentes de mamíferos y aves con cámaras de trampeo. Sor� prendentemente, el lince euroasiático (Lynx lynx) fue el vertebrado necrófago más importante. No obstante, en ambas estaciones, los efectos del consumo de los vertebrados necrófagos fueron de poca magnitud. En verano, los cadáveres fueron consumidos totalmente en 10 días, en su mayor parte por invertebrados. En invierno, solo el 5% de los cadáveres se consumieron en 10 días y el 16%, en 15 días. Concluimos que los invertebrados del parque nacional del bosque de Baviera no son fuertes competidores del lince. Palabras clave: Carroña, Lynx lynx, Necrofagia, Cleptoparasitismo, Descomponedores, Competencia por el alimento. Received: 7 IX 13; Condtional acceptance: 22 XI 13; Final acceptance: 6 V 14 Rena–Rebecca Ray, Zoological Research Museum A. Koenig, Adenauerallee 160, D–53113 Bonn, Germany.– Heidi Seibold, Inst. für Statistik, Ludwig–Maximilians–Universität München, Germany.– Marco Heurich, Dept of Conservation and Research, Bavarian Forest National Park and Chair of Wildlife Ecology and Management, Albert–Ludwigs–University of Freiburg, Germany. Corresponding author: Rena–Rebecca Ray. E–mail: r.ray@web ISSN: 1578–665 X eISSN: 2014–928 X

Introduction The availability of wild ungulate carcasses has been limited throughout the European ecosystems for many years. Ungulate populations are hunted intensively and are usually fed in winter. Therefore, their natural mortality rates are very low, and dead animals are quickly removed from the ecosystem. With the return of large predators —such as wolves (Canis lupus) and the Eurasian lynx (Lynx lynx)— to parts of Europe in recent years, the amount of carrion in the ecosystem may increase. Until recently, most studies on scavenging of predator kills have been carried out in areas that are home to specialized scavengers, such as vultures and hyenas, specifically in Africa, South America and southern Eu� rope (Kruuk, 1967, 1972; Houston, 1974, 1975, 1979, 1988; Schaller, 1972; Wallace & Temple, 1987; Hiraldo et al., 1991; Gomez et al., 1993; Travaini et al., 1998). In the last decade studies have also been conducted in the temperate zone (Heinrich, 1989, 1999; Stahler et al., 2002; Wilmers et al., 2003; Selva, 2004; Selva et al., 2005; Krofel et al., 2012). Results obtained indicate that the carcasses of animals, both those killed as prey and those dying from other causes, are mostly consumed by vertebrate scavengers. Selva (2004) observed that up to 36 vertebrate species consume animal carcasses. Such data are of relevance for the ecological community. In temperate regions, animal carcasses are an important source of nutrition in mid–winter when other resources are unavailable or depleted (Jedrzejewska & Jedrzejewski, 1998; Sidorovich et al., 2000), and they are crucial for survival during severe winters (Angerbjörn et al., 1991; Sidorovich et al., 2000; Selva et al., 2003). Carcasses have an important effect on the population dynamics of scavengers (DeVault et al., 2003; Roth, 2003) and can be highly significant for the stability and persistence of ecosystems (McCann et al., 1998). Turnover and distribution of nutrients from the carcasses is accelerated by scavengers, which thereby play an essential role in the nitrogen cycle (Putman, 1978; Braack, 1987; Towne, 2000). Especially in summer, the availability of animal car� casses increases the species richness, such as the abundance of beetles and the heterogeneity of plant communities. Certain developmental stages of some insect species may even depend on the presence of carcasses (Sikes, 1994; Towne, 2000; Melis et al., 2004). Vertebrate scavengers have the potential to con� sume large parts of prey. Ravens (Corvus corax) were shown to reduce the consumption rate of a single wolf by 70% and of a large wolf pack by 10% (Promberger, 1992; Hayes et al., 2000; Kaczenzky et al., 2005). In the Dinaric mountains, lynxes lost 15% of their kills to scavenging by other large predators such as brown bears (Ursus arctos) (Krofel et al., 2012). As a consequence, if parts of a predator’s kill are consumed by scavengers, the predator has to increase its kill rate to obtain enough food, as has been observed for wolves (Vucetich et al., 2004) and lynxes (Krofel et al., 2012). Like other cat species, lynx are solitary predators

Ray et al.

and stalk hunters. If a lynx is not disturbed and the carcass does not decay, the lynx will feed on the carcass for three to six days (Jobin et al., 2000). Scavengers therefore have a relatively broad window of time to discover and make use of the kill of lynx. To avoid scavengers, lynx usually cache their kills by dragging them into dense vegetation and cover� ing them with leaves or grass, or even snow or soil (Festetics, 1980; Hucht–Ciorga, 1988; Jedrzejewski et al., 1993). Scavenger species that use their olfac� tory sense for orientation, such as wild boar (Sus scrofa) and red fox (Vulpes vulpes), can find even well–hidden carcasses, though probably less often (Hucht–Ciorga, 1988; Jedrzejewska & Jedrzejewski, 1998; Jobin et al., 2000). Large predators compete with human hunters for the same prey species. An increase in the predator’s kill rate caused by scavengers removing parts of the killed prey (Vucetich et al., 2004; Krofel et al., 2012) could potentially exacerbate the conflicts between humans and large predators. As a result, illegal ki� llings of large predators might increase, with possible consequences for the conservation of small predator populations (Cerveny et al., 2002; Breitenmoser & Breitenmoser–Würsten, 2008). To understand this conflict and protect large predators, more information of the scavenger community and its possible influence on predator kills is needed. Our study in the Bavarian Forest National Park aimed to assess the scavenger community and its consumption of ungulate carcasses in summer and in winter. Our hypotheses were that (1) vertebrate scavengers such as ravens, red foxes and wild boars consume a higher proportion of the carcasses than invertebrates. (2) the consumption rate is higher in winter than in summer, because other food resources are scarce in this season, and (3) vertebrate scavengers are effective competitors of Eurasian lynx. For this purpose we set out road kills of ungulates that were simulated as lynx kills and monitored the scavengers using camera traps during the summer of 2003 and the winter of 2003/2004. Materials and methods Study area The study was performed in the Rachel–Lusen–Area of the Bavarian Forest National Park, which is situated in south–eastern Germany along the border with the Czech Republic (49° 3' 19'' N, 13° 12' 9'' E). The park covers an area of more than 240 km². Together with the Šumava National Park in the Czech Republic (690 km²), the Bavarian Forest Nature Park (3,070 km²) and the Šumava Protected Landscape Area (1,000 km²) it cons� titutes the Bohemian Forest Ecosystem. The area is mountainous, with altitudes varying between 600 and 1,453 m a.s.l. and mean annual temperature ranging between 6.5°C in the valleys and 3°C along the ridges and at higher elevations. The mean annual precipitation is between 830 and 2,230 mm, a considerable amount of which occurs as snowfall. Most of the area is forested (97%), consisting

Animal Biodiversity and Conservation 37.1 (2014)

mostly of Norway spruce (Picea abies) and European beech (Fagus sylvatica) (Heurich & Neufanger, 2005). Among the objectives of the National Park is the con� servation of natural processes, including the promotion of undisturbed dynamics within natural communities. This includes the avoidance of control measures for the wild ungulate population through human intervention. Therefore forest and wildlife management is not allowed in 75% of the area (Heurich et al., 2011). A small population of Eurasian lynx currently lives in the Bohemian Forest Ecosystem. The population derives from 17 lynx that were reintroduced to the Bohemian Forest in the 1980s. Initially, information on lynx presence indicated an increase in numbers and distribution. However, at the end of the 1990s, the growth of the lynx population stagnated and in recent years the number of individuals even decreased (Wölfl et al., 2001). Records from systematic camera trap� ping revealed a density of approximately 0.9 lynx per 100 km² in the national parks (Weingarth et al., 2012). Simulation of lynx kills We simulated the carcasses as lynx kills to obtain an idea as to whether lynx could be affected by kleptopa� rasitism in the Bavarian forest, meaning that scaven� gers of certain species feed on or take away parts of lynx kills. To determine the influence of scavengers, we monitored all animals that fed on the carcasses. Roe deer (Capreolus capreolus) is the most impor� tant lynx prey in our study area, followed by red deer (Cervus elaphus) (Podolski et al., 2013). This is similar to other areas occupied by lynx (Breitenmoser & Haller, 1987; Haller, 1992; Hucht–Ciorga, 1988; Okarma et al., 1997; Podolski et al., 2013) and we used road–killed ungulates for the simulations. When road–kills were reported, their bodies were collected within 1.5 h and immediately frozen to prevent decomposition. All of the 24 carcasses (19 roe deer, 4 wild boars, and 1 red deer) were frozen in a cold–storage house to avoid flies. Before the carcasses were placed in the wild they were thawed in a cooling chamber. We simulated a natural lynx kill by cutting each carcass at the neck or thigh and inflicted additional wounds on the carcasses to simulate natural kills, providing access points for scavenging birds. Human odour was minimised by wearing rubber boots and rubber gloves rubbed in lynx scat. We positioned the carcasses in the morning, simula� ting the natural behaviour of the lynx by dragging them along the ground for about 30 m to their final destina� tion to leave a scent trail and covering the carcasses with leaves. We also placed scat collected from lynx kept in an enclosure at the site since lynx often drop scat in the vicinity of their kills (Hucht–Ciorga, 1988). We usually positioned 1–3 carcasses at a time in summer (June, July, August; n = 15) and 1–3 carcas� ses at a time in winter (November, December, January) and monitored each of them until total depletion. This lasted 10 days in summer and 16 days (n = 6) or 32 days (n = 3) in winter, resulting in 150 and 192 observation and camera trap–days in summer and in winter, respectively.

We positioned the carcasses at 15 different loca� tions at least 1 km apart, distributed throughout the study area to avoid habituation of the scavengers living close to these sites (fig. 1). Carcass monitoring Each day of the observation period, we visually esti� mated the proportion of carcass remaining (in %) and assigned the level to one of eight classes (table 1). We decided against weighing the carcasses daily as done by Promberger (1992) to avoid disturbing the carcasses in the near–natural simulation experiment and because maggots and other decomposers in the flesh in summer would distort the results. We monitored the carcasses using camera traps (Camtrakker), consisting of compact analogue ca� meras equipped with motion sensors. The distance between the camera trap and the carcass varied from site to site, averaging 1–2 m, and motion sensors were set at the shortest interval to record as many move� ments as possible, with a time interval of 1' between the captures. We checked the camera traps daily and photographed the carcasses and recorded the decrease in the amount of tissue (table 1), changes in carcass position, and evidence of other animals in the immediate vicinity, such as tracks, scat and fur. When an animal moved the carcass, we moved the camera trap accordingly, and did not change the position of the carcass. During this procedure, which lasted no longer than 15', we minimized human odour by wearing rubber boots and gloves and chose the morning (two hours after sunrise) for our visits, be� cause mammalian scavengers did not appear before dusk. We assumed that in the intervening hours, our human scent was reduced to a level that did not deter the animals. We observed no indications of individual species avoiding or preferring the simulation sites, as indicated by tracks in the soil or snow. In general, one camera trap picture was regarded as an observation, but if an individual was captured in a consecutive series of pictures, these pictures were treated as a single observation. In addition, we used tracks in the soil around the carcasses to su� pport the camera trap observations. If an animal was not photographed by the camera but was definitely present based on the tracks, we counted the tracks as one observation. Determination of carcass consumption by scavengers We compared the amount of animal tissue consumed by the scavengers on each day with that normally con� sumed by the male lynx and by female lynx with kittens (family group) to assess the percentage of consumption by scavengers feeding on the carcasses, as reported by Jobin et al. (2000), who found an average of 3.2 kg per day consumed by a male lynx and 4.9 kg per day by a family group. The consumable fraction of an ungulate carcass is approximately 75%, including the digestive tract (Messier & Crete, 1985). Our measurements of live–trapped adult roe deer in the study area revealed an average weight of 21 kg. The mean killing rate of

Ray et al.

National Park Šumava Frauenau

Bavarian Forest National Park Speigelau Riedlhütte Mauth

St. Oswald Neuschönau Simulated lynx kills Forest 0 1 2 3 km

Grafenau Hohenau

Fig. 1. Map of the Bavarian Forest National Park, showing the placement of the simulated lynx kills. Fig. 1. Mapa del parque nacional del bosque de Baviera en el que se muestra la ubicación de las presas simuladas de lince.

a lynx in Poland was observed to be one roe deer in 5.4 days (Okarma et al., 1997); in Switzerland, the rate was one every 6.6 days for a male lynx and one every five days for a family group (Breitenmoser & Haller, 1987; Haller, 1992; Jobin et al., 2000). Depending on the size of the prey, we postulated that a lynx can feed 4–10 days on a single kill. Statistical analysis To account for the effects of the two different obser� vation periods (summer and winter), the absolute number of observations was converted to the num� ber of observations (individual animals) per camera trap–day (fig. 2, table 2). We computed a generalized linear mixed model (GLMM, Fahrmeir et al., 2013) using the number of vertebrates as outcome variable to compare the visitation of carcasses by vertebrates both in summer and winter. This mixed quasipoisson model was fitted using a Log–link function, with the observation site as a random intercept, an offset containing the trap days and season as a fixed factor. We used quasi–poisson regression instead of poisson regression in order to avoid overdispersion. We conducted a proportional hazards model, also known as the Cox–model (Cox & Oakes, 1984), to compare the time span until the carcasses were detec� ted by vertebrate species in summer and winter. The outcome is the time until the carcass was found, which is a right censored time–to–event variable. It is right

censored because some carcasses were not found within the observation time. To account for possible correlation on each observation site, we included a frailty term (Therneau & Grambsch, 2000). In the same way, we compared the time span until the carcasses were found between invertebrate and vertebrate spe� cies in summer. We used original incidence data and percentage–based data for the analyses. All computations were performed in R (R Core Developer Team, 2013), version 2.13.2, using the add–on package mgcv (Wood, 2010) for fitting GLMMs and the package survival for fitting the Cox–model (Therneau, 2014). Results Vertebrate species observed at the carcasses Over the course of the two study periods, June to August 2003 and November 2003 to January 2004, 280 successful observations of eight vertebrate spe� cies were recorded. In 45 photographs the species could not be identified (table 2). In summer, red foxes were the most common species observed at the sites (37%), whereas in winter, lynx was the most common species (31%) (fig. 2, table 2). During winter, we also observed three lynxes (female, male, and sex unknown) feeding on the same carcass. The average rate of observation was one vertebrate every five camera trap–days (fig. 2). GLMM analysis

Animal Biodiversity and Conservation 37.1 (2014)

Table 1. Criteria for classification of the degree of carcass consumption. Tabla 1. Criterios para clasificar el grado de consumo de los cadáveres. Remaining tissue (%)

Mean (%)

100

Carcass complete

90–99

Carcass with initial signs of consumption

70–89

Carcass missing one third of the animal tissue

50–69

More than half of the carcass remaining

30–49

Less than half of the carcass remaining

10–29

Carcass with less than one third of the animal tissue remaining

1–9

Carcass with scattered remains of animal tissue

Carcass with no remaining animal tissue

Characteristics

Elapsed time until scavengers found the carcasses The Cox–model with target variable days until detec� tion of the carcass and frailty term for the observation

0,20 0.18 0.16

Summer Winter

0.14 0.12 0.10 0.08 0.06 0.04

al To t

y C

om m

ar bo

ild

n ia as

ba E

R ed op ur E

0.00

0.02

fo x

Nº of observations / camera trap–day

using a random Intercept for the observation site showed that the number of vertebrates did not sig� nificantly vary between seasons (estimate = –0.752, p = 0.17878, SE = 0.541).

Class

Fig. 2. Observations of different species at the carcasses per camera trap–day. The absolute number of observations was divided by the total number of camera trap–days (in summer and in winter, 150 and 192 respectively). Fig. 2. Observaciones de diferentes especies en los cadáveres por día de trampeo con cámara. El número absoluto de observaciones se dividió por el total de días de trampeo con cámara (en verano y en invierno, 150 y 192 respectivamente).

Ray et al.

Table 2. Observations of various vertebrate species at the carcasses divided in summer and winter. Frequency of observations: percent of carcasses visited by each species, observations/camera trap day: percent of camera–trap days with a species present: Obsv. Frequency of observations; C. Carcasses visited; Obsv/ct Observation / camera trap days. Tabla 2. Observaciones de varias especies de vertebrados en los cadáveres en verano y en invierno. Frecuencia de las observaciones: porcentaje de cadáveres visitados por cada especie, observaciones por día de trampeo con cámara: porcentaje de días de trampeo con cámara en los que una especie estaba presente: Obsv. Frecuencia de observaciones; C. Cadáveres visitados; Obsv/ct. Observaciones por día de trampeo con cámara. Species

Summer

Winter

Obsv (%) C (%) Obsv/ct (%)

Red fox (Vulpes vulpes)

Obsv (%)

C (%) Obsv/ct (%)

Lynx (Lynx lynx)

122

Common buzzard (Buteo buteo)

Eurasian jay (Garrulus glandarius)

–

100

Wild boar (Sus scrofa)

Hedgehog (Erinaceus europaeus)

–

Badger (Meles meles)

–

Goshawk (Accipiter gentilis)

–

Total number

150

192

site showed that vertebrates found the carcasses more quickly in summer than in winter (estimate –1.16, p = 0.011, fig. 3). Vertebrate species found 67% of the carcasses within 10 days in summer, but only 22% were found within the same time period in winter, when they were frozen and covered with snow. A second Cox–model showed that invertebrates found the carcasses faster than vertebrates in summer (es� timate = –1.48, p = 0.058, fig. 3). Within four days in summer, decomposers such as maggots of blowflies and flesh–flies (Calliphoridae, Sarcophagidae) were observed on 95% of the deposited carcasses. Lynx, common buzzard, Eurasian jay, and hedgehog found the carcasses within two days. Red fox and wild boar did not arrive until the fourth day.

visited the carcasses when the maggots had reached peak numbers. Both species feed mostly on the ma� ggots and less on the carcass tissue and other remains.

Consumption of the carcasses

Discussion

Carcasses were consumed much more quickly in summer than in winter (fig. 4). In summer, all of the carcasses were totally consumed within 10 days. In that period, none of the carcasses was completely consumed in winter. Most of the scavenging in sum� mer was by invertebrates (85% of animal tissue, fig. 5). Vertebrates only played a minor role. Invertebra� tes reached peak numbers on days 4–6. Red foxes appeared before and after the maggots reached their peak numbers. Red foxes ate mostly internal organs or dragged away parts of the skeleton. Lynx mainly ate muscle tissue, and wild boars and hedgehogs mainly

Our study of use and consumption of carcasses pro� vides the first data on the scavenger community in the Bavarian Forest. We observed eight different vertebrate species at the simulated kills, with red foxes being the most frequent visitors. Contrary to other studies (Selva, 2004; Jedrzejewski et al., 1993; Jobin et al., 2000; Kellermann, 2001), we found that invertebrates consumed a much larger proportion of the carcasses than vertebrates and overall carcass consumption was much slower in winter than in summer. Similar to our study, Selva (2004) identified red fox, common buzzard, Eurasian jay and wild boar

Food competition with lynx Based on our results, if we assume that a male lynx killed a roe deer, competing scavengers would have consumed approximately 19% of the utilizable biomass of the carcass in summer and only 0.7% in winter. If we assume that a lynx family group was feeding on a killed roe deer, the scavengers would have consumed 0.7% of the utilizable biomass of the carcass in summer and 0% in winter.

Animal Biodiversity and Conservation 37.1 (2014)

Portion of detected carcasses (%)

100 90

Vertebrates in winter

Vertebrates in summer Invertebrates

70 60 50 40 30 20 10 0

1 2 3 4 5 6 7 8 9 10 11 12 13 14 15 16 17 18 19 20 21 22 23 24 Time until display of the carcasses (days)

Fig. 3. Observed time until detection of carcasses by invertebrates and vertebrates in summer and winter, respectively. Starting from 0% (unimpaired carcass, no sign of detection) to 100% (completely consumed carcass, detected). Fig. 3. Tiempo observado hasta la detección de los cadáveres por parte de los invertebrados y los vertebrados en verano y en invierno, respectivamente. Desde 0% (cadáver intacto, ninguna señal detectada) al 100% (cadáver completamente consumido).

100

Mean non–consumption of the carcasses (%)

90 80 Mean winter ± sd

70 60 50

Mean summer ± sd

40 30 20 10

3 4 5 6 7 8 9 10 11 12 13 14 15 Time until display of the carcasses (days)

Fig. 4. Observed relative progress of carcass consumption in summer (vertebrates and invertebrates) and winter (vertebrates only), starting from 100% (unimpaired carcass) to 0% (completely consumed). The solid lines indicate the average in each season. The dash–dotted lines on each side of each solid line indicate the standard deviation. Fig. 4. Progreso relativo observado del consumo de cadáveres en verano (vertebrados e invertebrados) y en invierno (solo vertebrados), desde 100% (cadáver intacto) al 0% (completamente consumido). Las líneas continuas indican el promedio de cada estación. Las líneas discontinuas a cada lado de las líneas continuas indican la desviación estándar.

as the main scavengers in the Białowieża Forest in Poland. However, they only observed lynx sporadically at the carcasses. In the same forest, Jedrzejewski et al. (1993) identified wild boar as the most frequent scavenger species, followed by raven and red fox. In the Swiss Alps and the Carpathian mountains of Romania, red fox was the predominant scavenger (Jobin et al., 2000; Kellermann, 2001). In our study during the summer, invertebrates such as blowflies and flesh–flies were the most efficient carcass find� ers and scavengers. Vertebrates such as red fox and wild boar usually found the carcass only after four days. Although red fox (Jobin et al., 2000) and wild boar (Jedrzejewska & Jedrzejewski, 1998) have little difficulty locating cached carcasses, days could pass until these scavengers discovered carrion by the scent. In contrast, the common buzzard and Eurasian jay sometimes found the carcass on the first day, but their consumption was very low. Other birds, namely vultures and ravens, are also very effective in locating carrion (Cortes–Avizanda et al., 2012). For example, turkey vultures (Cathartes aura) were able to find 96% of the carcasses within three days (Houston, 1986), and vultures can consume a 100 kg carcass within 30 min (Houston, 1974). In contrast to Poland, where ravens found 92% of the carcasses within three days and drew the attention of other species to their finds (Selva, 2004), this species was not observed in our study. In Poland, most of the carcasses, including European bison (Bison bonasus), red deer, and wild boar, were larger than in our study and were mostly found by avian scavengers in open rather than in closed woodland. The time it takes to locate a carcass is an important factor and depends strongly on the conditions under which a lynx leaves its kill. The risk of lynx kills being discovered by scavengers is probably low, because lynx often move the kill from the killing site and cam� ouflage it. They also keep the immediate area where they consumes their kill tidy, with the bones and hide thoroughly cleaned (Kruuk, 1986; Hucht–Ciorga, 1988; Jedrzejewska & Jedrzejewski, 1998; Breitenmoser & Breitenmoser–Würsten, 2008). In a large area covered by closed forest, it is unlikely that scavenging birds, such as ravens, can easily locate the hidden, relatively small lynx kills. This could explain why ravens, which are common in our study area, were not observed at the carcasses. While lynx are capable of relocat� ing and covering small roe deer kills, they are not able to do so with large red deer carcasses. For this reason, in the study of Jedrzejewska & Jedrzejewski (1998), scavengers were observed at only 28% of the roe deer kills but at 77% of the red deer kills. These authors also noted that scavengers visited 63% of the wolf kills, but only 38% of the lynx kills, which indicates that lynx are able to hide their kills from scavengers effectively. At the peak of maggot development, hedgehogs and wild boar visited the carcasses, eating mainly maggots and consuming little flesh. Besides a large proportion of plant matter, the omnivorous wild boars also consume a broad spectrum of invertebrates and carrion (Niethammer & Krapp, 1986; Briedermann, 1990).

Ray et al.

The differences in carcass consumption between summer and winter are probably primarily due to four factors. First, temperature seems to mediate competition between vertebrate scavengers and invertebrate decomposers (DeVault et al., 2004). At temperatures below 0°C, blowflies and flesh–flies become inactive, and as temperatures increase, the rate at which maggots consume a carcass increases (Rognes, 1991; Campobasso et al., 2001). Second, in winter, snow often covers the carcasses, making it difficult for scavengers to detect them, both visu� ally and olfactorily. Third, in contrast, defoliation and snow under a fresh kill increase the visibility of the carcasses from the air, possibly explaining the higher number of observations of European jay and goshawk (Accipter gentilis) in winter in our study. Poor visibility could also explain the decrease in scavenging by ravens in spring and summer (Kellermann, 2001). In our study, however, the common buzzard showed less scavenging activity in winter. This can be explained by the fact that buzzards leave their territories and disperse after the first snowfall. And lastly, in winter, food availability decreases and carrion becomes an important alternative trophic resource (Lockie, 1959; Goszczynski, 1974; Festetics, 1980; Pulliainen, 1981; Jedrzejewski & Jedrzejewska, 1992; Labhardt, 1996; Selva, 2004; Cortes–Avizanda et al., 2009). In summer, food is available in abundance, and most scavengers do not make use of carrion. Both our observations and those of Labhardt (1996) on the lower number of scavenging red foxes in spring and summer support these conclusions. Our observed scavenger consumption rate of 0.7 % in winter was much lower than that observed in winter studies in Poland (15%; Jedrzejewska & Jedrzejewski, 1998) and Romania (18%; Kellermann, 2001), probably because effective scavengers, such as ravens, did not visit the carcasses, and predators with conspicuous kills such as wolves, which occur in Poland and Romania (Selva, 2004), do not occur in our study area. Another explanation could be the lower scavenger densities in the mountainous, snow rich environment of the Bavarian Forest. This con� clusion is also supported by the relatively large time span the vertebrate scavengers needed to detect the carcasses. Although, in our study, various vertebrates were observed at the carcasses, their influence was so low that they cannot be considered to be competi� tors of lynx. Surprisingly, the second most frequently observed scavenger was the lynx. They consumed more flesh than any of the other observed vertebrates. Other studies have rarely recorded lynx consuming carrion other than their own kills (Hucht–Ciorga, 1988; Selva, 2004; Selva et al., 2005). Hucht–Ciorga (1988) documented for the Bavarian forest that lynxes sometimes only took foreign carrion when it was placed next to the lynx kill. One explanation for this behaviour might be the practice of hunters in the area to regularly leave the remains of shot deer, a predict� able food supply for lynx. Our observation of three lynx feeding on the same carcass might be explained by the fact that solitary individuals of different sex are more tolerant of each other at breeding time (Schmidt

Animal Biodiversity and Conservation 37.1 (2014)

Eurasian jay 2%

Common buzzard 3%

Common goshawk 1% European badger 2%

Wild boar 7%

Red fox 24% Eurasian lynx 61%

Wild boar 3% Red fox 6%

Common buzzard 1% European hedgehog 1%

Eurasian lynx 4%

Invertebrates 85%

Fig. 5. Carcass consumption by the different species in percent in summer (A) and winter (B), determined with the criteria shown in table 1. Fig. 5. Consumo de cadáveres por parte de las diferentes especies en verano (A) y en invierno (B), determinado con el criterio mostrado en la tabla 1.

et al., 1997; Breitenmoser & Breitenmoser–Würsten, 2008). Another explanation might be that sub–adults scavenge the kills of a territorial adult when it is not present or is alternately feeding on several simultane� ously available kills. Another interesting observation was two lynx feed� ing simultaneously on one carcass in winter, leaving their own scat alternating with other animals feeding on the same carcass, including red fox, wild boar and badger. In Switzerland, red foxes visited natural lynx kills only after lynx had abandoned the kill site (Jobin et al., 2000), whereas in Romania, red foxes frequented lynx kills while lynx were still utilising them (Kellermann, 2001). Despite the small number of simulated kills in this study, our results indicate that the impact of

vertebrates on the consumption was low all year. We conclude that other vertebrates in the national park are not strong competitors for lynx and that the im� pact of invertebrates on carrion decomposition should be given a higher priority in future research. In our study, blowflies and their kin were the most effective scavengers during the summer months. However, in natural lynx kills, lynx feed regularly on the carcass, thereby reducing the ideal nutritional conditions for maggots by not allowing their development to the same extent as in simulated kills (Smith, 1986). A high impact of scavenging maggots on lynx kills would be expected especially under natural conditions when the prey is large, such as a red deer, which takes a single lynx about seven days to completely consume. Nevertheless, invertebrates can out–compete other

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Chinchilla lanigera (Molina 1782) and C. chinchilla (Lichtenstein 1830): review of their distribution and new findings P. Valladares, C. Zuleta & Á. Spotorno

Valladares, P., Zuleta, C. & Spotorno, Á., 2014. Chinchilla lanigera (Molina 1782) and C. chinchilla (Lichtenstein 1830): review of their distribution and new findings. Animal Biodiversity and Conservation, 37.1: 89–93. Abstract Chinchilla lanigera (Molina 1782) and C. chinchilla (Lichtenstein 1830): review of their distribution and new findings.— Millions of Chinchilla chinchilla and C. lanigera were killed during the early twentieth century and they were nearly hunted to extinction. In order to establish the current range of distribution of these two wild species and to localize possible new colonies, we used the available scientific literature, technical reports, information from residents, and live trapping methods. Both species are 'critically endangered' since their current distribution is highly fragmented and all recognized colonies are small and isolated. We report a small new wild colony of C. lanigera in the Atacama region, Chile. Key words: Chinchilla, Critically endangered, Distribution, Endemism, New colonies, Chile. Resumen Chinchilla lanigera (Molina, 1782) y C. chinchilla (Lichtenstein, 1830): revisión de su distribución y nuevas observaciones.— Tanto Chinchilla chinchilla como C. lanigera estuvieron muy cerca de la extinción debido a la caza histórica y masiva de que fueron objeto, y que acabó con millones de ejemplares durante la primera parte del siglo veinte. Para determinar el rango de distribución de estas especies y localizar nuevas colonias, analizamos las publicaciones científicas, los informes técnicos, la información facilitada por personas residentes y los trampeos en vivo. Detectamos una nueva colonia silvestre de pequeño tamaño de C. lanigera en la región de Atacama, Chile. El estado de conservación de ambas especies sería de “en grave peligro de extinción”, ya que la distribución está muy fragmentada y la mayor parte de las colonias detectadas son pequeñas y están aisladas. Palabras clave: Chinchilla, En grave peligro de extinción, Distribución, Endemismo, Nuevas colonias, Chile. Received: 4 II 14; Conditional acceptance: 2 IV 14; Final acceptance: 23 V 14 Pablo Valladares Faúndez, Depto. de Biología, Fac. de Ciencias, Univ. de Tarapacá, Av. General Velásquez 1775, Arica, Chile.– Carlos Zuleta, Depto. de Biología, Fac. de Ciencias, Univ. de La Serena, Casilla 599, La Serena, Chile.– Ángel Spotorno Oyarzún, Lab. de Citogenética Evolutiva, Programa de Genética Humana, Inst. de Ciencias Biomédicas, Fac. de Medicina, Univ. de Chile, Santiago, Chile. Corresponding author: P. Valladares.

ISSN: 1578–665 X eISSN: 2014–928 X

Introduction Chinchilla chinchilla (Lichtenstein, 1830), commonly known as the short–tail or Andean chinchilla, is a rodent of the family Chinchillidae. The historical distribution of this chinchilla included the highlands of Chile, Argentina, Peru and Bolivia (Chacón, 1892; Walle, 1914; House, 1953; Grau, 1986; Jiménez, 1996; Anderson, 1997; Eisenberg & Redford, 2000; Parera, 2002; Woods & Kilpatrick, 2005). C. lanigera (Molina, 1782), on the other hand, is traditionally known as the long–tail or coastal chinchilla, and it is endemic to north–central Chile (Jiménez, 1996; Valladares, 2002; Spotorno et al., 2004a). In the past, however, it had a wider distribution (Grau, 1986; Jiménez, 1996), ranging from the Choapa River (32°S) to north Potrerillos (26°S). Over seven million chinchilla furs were exported from Chile during the first part of the twentieth century (Albert,1900). However, this represented only one third of the total number of captured chinchillas as many furs were damaged as a result of the hunting methods used and discarded (Albert, 1901). It is therefore estimated that more than twenty million specimens were killed in Chile during this period. Even though both species were considered extinct during the 1960s, C. chinchilla was rediscovered in the highlands of the Antofagasta region in Chile by Spotorno et al. (1998) and C. lanigera was found near Illapel, Coquimbo region, Chile (Mohlis, 1983). More recently, a colony of C. lanigera was documented near La Higuera, North of the Coquimbo region (Spotorno et al., 2004a). This study presents the new colony of C. lanigera. We also discuss the range of distribution of the two species of chinchillids in northern Chile based on the information available. Material and methods To assess the distribution, ecology and conservation status of the two species we analyzed all the scientific information available (e.g. Jiménez 1987, 1989, 1995, 1996; Spotorno et al., 1998; Cortés et al., 2002, Spotorno et al., 2004a, 2004b; Valladares, 2012; Valladares et al., 2012; Tirado et al., 2012; Lagos et al., 2012) in technical and public reports (e.g. Mohlis, 1983; Schlatter et al., 1987; Lagos et al., 2008; Martínez & Cortés, 2011; Povea et al., 2012). We revisited the sites where the chinchilla species have been observed to confirm their presence. Live trapping, feces and hair collection, and cave and pawprint identification were carried out to establish the presence of chinchillas. We describe the microhabitat, vegetation and the presence of other sympatric species of the newly discovered colonies. Results According to current scientific literature, C. chinchilla has been documented in restricted areas of Chile, most specifically around El Laco and Morro Negro towns, both near the Llullaillaco volcano, Antofagasta

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region (Spotorno et al., 1998; Spotorno et al., 2004b; Tirado et al., 2012) and also near the Nevado Tres Cruces National Park and its surroundings, in the Atacama region (Valladares et al., 2012) (fig. 1A). In Argentina it has been documented near the Antofalla, Catamarca (Walker et al., 2007), southwestern Jujuy (Olrog & Lucero, 1981), Salta (Ortiz et al., 2010), La Rioja (Parera, 2002) and northern San Juan (Cajal et al., 1981). In Bolivia, its distribution included the departments of La Paz, Oruro and Potosi (Anderson, 1997). The last wild specimen in this country was captured by residents of Huachacalla, Sabaya, and Caranga (Walle, 1914). The distribution of C. lanigera includes Las Chinchillas National Reserve (Jiménez, 1995, 1996) in Aucó (about 700 ind/km2, Cofré & Marquet, 1999) and Quebrada El Cobre (with 4.4 to 72.9 ind/km2; Lagos et al., 2010). Some colonies have been identified outside the limits of the reserve, in the areas of Quebrada Curico and Quebrada El Cuyano (between 17.5–82.6 ind/km2 and 12.3 to 58.3 ind/km2, respectively; Lagos et al., 2010), while a small and isolated colony of these chinchilla has been found in Corral de Piedras, La Higuera (Spotorno et al., 2004a). No estimate, however, is available on the size of the population. It appears that C. lanigera inhabited the Atacama region during the first part of the twentieth century, particularly around Vallenar (Wolffsohn, 1923), Quebrada El León (ca. 26° 57' 34.05'' S, 70° 41' 31.90'' O) (Gigoux, 1926), and Morro Copiapó (ca. 27° 7' 51.89'' S, 70° 55' 48.62'' W) (Gigoux, 1935). According to the literature, they had previously been abundant, but the massive captures were carried out in 1892, and the species was regarded as possibly extinct. Olave & Monroy (2006) published a photograph taken in 1923 of Pablo Trabucco Onetto, with their breeding of chinchillas captured near Chañaral. Based on this evidence, we thought that finding new colonies in this province was more likely, because local residents mentioned many places where chinchillas were captured at the beginning of the twentieth century. Other authors have mentioned wild colonies of chinchillas in the north of Chile, for instance, in Mejillones (Phillipi, 1860), the Licancabur volcano (Rudolph 1955), and La Ola and Potrerillos (Schlatter et al., 1987). Grau (1986) suggested that both species may have inhabited in sympatry around Potrerillos (north of Atacama region), corresponding to the traditional northern limit of the C. lanigera distribution, and the native southern limit of C. chinchilla (fig. 1A). A new colony was found by a group of miner– workers in the Atacama region, Chile (26° 55' 07'' S, 70° 21' 32'' W). They captured one chinchilla specimen, rescuing it from a group of domestic dogs. The specimen was taken to the Servicio Agrícola y Ganadero (SAG) of Copiapó and examined. It was a C. lanigera adult male, based on its long–tailed proportion (152 mm with hair and 75 mm without hair), the principal diagnostic character of C. chinchilla (with a tail < 110 mm long) (Spotorno et al., 2004b). Later, we visited the locality where this specimen was captured. We found another specimen between large cracked rocks. We identified another

Animal Biodiversity and Conservation 37.1 (2014)

N Rudolph (1955) Phillipi (1860)

Spotorno et al. (1998) Spotorno et al. (2004)

Gigoux 1926 Gigoux 1937

Schlatter et al. (1981) New colony Valladares et al. (2012) Valladares et al. (2012) Lagos et al. (2012)

Wolffsohn (1937)

Spotorno et al. (2004)

0 20 40 60 80 100 120 140 160 180 200 km

New colony

Región de Atacama Chile Caldera

0 10 20 30 40 50 60 km

Fig. 1. A. Distribution map of Chinchilla lanigera (squares) and C. chinchilla (circles); black squares and circles correspond to confirmed colonies, and white squares and circle to colonies mentioned in the literature but not confirmed by our recent field survey); B. Locality of the new colony of Chinchilla lanigera from Atacama region, Chile (26° 55' 07'' S, 70° 21' ´32'' W). Fig. 1. A. Mapa de la distribución de Chinchilla lanigera (cuadrados) y C. chinchilla (círculos); los cuadrados y los círculos negros corresponden a colonias confirmadas, mientras que los cuadrados y los círculos blancos corresponden a colonias mencionadas por otros autores, pero que no hemos confirmado en nuestro reciente estudio de campo; B. Localidad de la nueva colonia de Chinchilla lanigera de la región de Atacama, Chile (26° 55' 07'' S, 70° 21' 32'' O).

42 points with feces, footprints and/or wallows, nine of which showed recent activity. We roughly estimated a density of some 24.7 to 115.4 ind/km2. This new colony was located 44 km from the coast, inhabiting the middle of an extremely arid hill, approximately 1,150 m in height, and surrounded by extensive dunes of the Atacama Desert (fig. 1B). The vegetation was identified as Heliotropium sclerocarpum, Tetragonia microcarpa, Gymnophytum flexuosum, Nolana sp., and particularly Eriocyse aurata, probably the main source of water and food. Some 87% of cactus showed signs of being gnawed by rodents. No other sympatric species were reported, but Phyllotys darwini and Eligmodontia dunaris have been collected near this area (Valladares, 2012; Spotorno et al., 2013). An owl, Bubo magallanicus, was observed as possibly the only predator in the zone, although foxes have occasionally been observed by miners.

Another new colony was reported by the mining company 'Cerro Blanco' belonging to White Mountain Titanium Corporation, close to Vallenar, Atacama region. They mentioned in their line base a record of C. lanigera in winter, 2012 (http://seia.sea.gob.cl/expediente/expedientesEvaluacion.php?modo=ficha&id_ expediente=7895426). Regarding C. chinchilla, colonies were reported in the Atacama region by the mineral project Salares Lithium Company that was developing a survey of the 'Salares 7' (http://seia.sea.gob.cl/documentos/documento. php?idDocumento=6326647). They described the presence of vertebrates, showing a photo of the footprints of C. brevicaudata [sic]. The 'Salares Norte Mining' from Gold Fields Salares Norte Company showed a wild specimen (http://seia.sea.gob.cl/documentos/documento. php?idDocumento=8230878). However, the population density of these colonies was not assessed.

Discussion Based on the available evidence, it seems clear that the past distribution of both wild species was indeed extensive. C. chinchilla was distributed in southern Peru, Bolivia, and northern Argentina and Chile (Grau, 1986); nonetheless, it has not been documented in Bolivia, Peru and Argentina in the last 50 years. Furthermore, the colonies identified in Chile are small and restricted to the Antofagasta and Atacama regions. On the other hand, C. lanigera is an endemic species occupying an area from Antofagasta to the Coquimbo regions (Grau, 1986). However, after a massive extermination, their distribution became restricted to Las Chinchillas National Reserve (Mohlis, 1983), and to a small colony located in the north of the Coquimbo region (Spotorno et al., 2004a). Both species have been reduced to less than 95% of their original distribution and important biological variables regarding their conservation status have not been assessed in the observed colonies. The new colony reported here is a small and isolated population, inhabiting small hills surrounded by a vast desert. We were unable to locate any nearby colonies. One possible explanation is that the existing colonies of C. chinchilla are extremely small and isolated with respect to other group (Valladares et al., 2012; Lagos et al., 2012). The highly fragmented and small mammalian populations generally have a low genetic diversity and a high level of inbreeding. These factors have consequently reduced their fitness, thereby increasing their risk of extinction (Keller & Waller, 2002). It is imperative to analyze their genetic diversity to compare them with those of the other populations that have low reported values (Spotorno et al., 2004a). Such studies could provide greater insight into what its future conservation needs may be.

Acknowledgements We would like to thank José Andaur and Patricia Cáceres from the Servicio Agrícola y Ganadero (SAG) for allowing us to analyze wild chinchilla specimens and for the information they provided on locality of capture. We are grateful to Thomas Püschel and Hugo Benitez from the University of Manchester, UK, for their help in the revision of the manuscript. This work was supported by the University of Tarapacá [UTA Mayor de Investigación Científica y Tecnológica 4711–14], and Fondo de Protección Ambiental [FPA 4–G–042–2013] from Ministerio del Medio Ambiente, Chile. References Albert, F., 1900. La chinchilla. Anales de la Universidad de Chile, 107: 913–934. – 1901. Datos sobre la chinchilla. Revista Chilena de Historia Natural, 5(9): 201–209. Anderson, S., 1997. Mammals of Bolivia, taxonomy and distribution. Bulletin of American Museum of Natural History, 231: 1–652. Cajal, J. L., Reca, A. A. & Pujalte, J. C., 1981. La

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Reserva Provincial San Guillermo y sus asociaciones ambientales. SECYT. Ministerio de Cultura y Educación, Buenos Aires. Chacón, J. C., 1892. Descripción Zoológica.In: Monografía del Departamento de Potosí (Bolivia), 8: 197–240. Centro de Estudios, Potosí. Cofré, H. & Marquet, P. A., 1999. Conservation status, rarity, and geographic priorities for conservation of Chilean mammals: an assessment. Biological Conservation, 88: 53–68. Cortés, A., Miranda, E. & Jiménez, J., 2002. Seasonal food habits of the endangered long–tailed chinchilla (Chinchilla lanígera): the effect of precipitation. Mammalian Biology, 67: 167–175. Eisenberg, J. F. & Redford, K. H., 2000. Mammals of the Neotropics. The Central Neotropics. Ecuador, Perú, Bolivia y Brazil. The University of Chicago Press, Chicago. Gigoux, E. E., 1926. La Quebrada del León (Caldera). Revista Chilena de Historia Natural, 30(1): 288–297. – 1935. El Morro Copiapó. Revista Chilena de Historia Natural, 39(1): 253–265. Grau, J., 1986. La Chinchilla. Su crianza en todos los climas. 3ra edición. El Ateneo, Buenos Aires. House, R., 1953. Animales salvajes de Chile. Universidad de Chile, Santiago. Jiménez, J. E., 1987. Eficiencia relativa de seis modelos de trampa para la captura viva de micromamíferos silvestres, con énfasis en Chinchilla lanigera (Molina, 1782). Medio Ambiente (Chile), 8: 104–112. – 1989. Uso de la técnica de tarjetas ahumadas para evaluar la efectividad de cebos para micromamíferos silvestres, con énfasis en Chinchilla lanigera. Medio Ambiente, 10: 84–91. – 1995. Conservation of the last wild chinchilla (Chinchilla lanigera) archipielago: a metapopulation approach. Vida Silvestre Neotropical, 4: 89–97. – 1996. The extirpation and current status of wild chinchillas Chinchilla lanigera and C. brevicaudata. Biological Conservation, 77(1): 1–6. Keller, L. F. & Waller, D. M., 2002. Inbreeding effects in wild populations. Trends in Ecology and Evolution, 17(5): 230–241. Lagos, V., Rodríguez, J., Cortés, I., Fuenzalida, R., Silva, J., Segovia, R. & Saavedra, B., 2010. Catastro y georeferenciación de colonias de Chinchillas (Chinchilla laniger). Chile Forestal, 348: 32–38. Corporación Nacional Forestal, Santiago, Chile. Lagos, N., Villalobos, R. & Iriarte, A., 2012. Nuevos registros de poblaciones de chinchilla de cola corta, Chinchilla chinchilla (Rodentia, Chinchillidae) en la cordillera de la Región de Atacama. Boletín del Museo de Historia Natural (Chile), 61: 191–196. Lichtenstein, M. H. C., 1830. Eriomys chinchilla Licht. Ie Chinchilla–Wollmaus. In: Darstellungen neue oder wenigbekannte Saugethiere in Abbildungen und Bescreibungen von fünf und sechzig Arten auf funfzig colorirten Steindrucktaffe lnnach den original en des Zoologischen Museums der Universitätzu Berlin. Königlichen Akademie der Wissenschaften: Heft 5, palte 28, plus 2 unnumbered pages of text. Königlichen Akademie der Wissenschaften, Berlin,

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Germany. Martínez, P. & Cortés, S., 2011. Prospección de colonias de chinchillas (Chinchilla laniger) en Quebrada Curico, Comuna de Illapel. Corporación Nacional Forestal, Departamento de Áreas Silvestres Protegidas, Región de Coquimbo, Documento de Trabajo. Mohlis, C., 1983. Información preliminar sobre la conservación y manejo de la chinchilla silvestre en Chile. Boletín Técnico, Corporación Nacional Forestal, 3: 1–41. Molina, J. I., 1782. Saggia sulla storia naturale del Chili. Stamperia di S. Tommaso d’ Aquino, Boologna. Olave, N. & Monroy, O., 2006. Chañaral, imágenes del pasado. Catálogo. Primera edición, Talleres AMF, Ciudad de Chañaral, Región de Atacama, Chile. Olrog, C. C. & Lucero, M. M., 1981. Guía de los mamíferos argentinos. Ministerio de Cultura, Fundación Miguel Lillo, Tucumán, Argentina. Ortiz, P. E., González, R., Jayat, J. P., Pardiñas, U. F., Cirignoli, S. & Teta, P., 2010. Dieta del búho magallánico (Bubo magallanicus) en Los Andes del Noroeste Argentino. Ornitología Neotropical, 21: 591–598. Parera, A., 2002. Los mamíferos de la Argentina y la región austral de Sudamérica. Editorial El Ateneo, Buenos Aires. Philippi, R. A., 1860. Viaje al desierto de Atacama. Halle em Sajonia, Librería Autum. Povea, P., Lara, C., Cortés, R., Rodríguez, J. & Saavedra, B., 2012. Zonificación de colonias de Chinchilla en Quebrada El Cuyano. Corporación Nacional Forestal. Documento de Trabajo. Oficina Provincial Choapa, Región de Coquimbo, Coquimbo. Rudolph, W., 1955. Licancabur: mountain of the Atacameños. Geographic Review, 45: 151–171. Schlatter, R., Murúa, R. & Oltremari, J., 1987. Diagnóstico de la situación actual de la fauna silvestre (aves y mamíferos) más característica que habita entre la II y VII Región administrativa de Chile. Corporación Nacional Forestal (CONAF), DP/CHI/83/017, Documento de trabajo, Santiago de Chile. Spotorno, A. E., Valladares, P., Marín, J. C., Palma, R. E. & Zuleta, C., 2004a. Molecular divergence and phylogenetic relationships among chinchillids (Rodentia, Chinchillidae). Journal of Mammalogy, 85: 384–388. Spotorno, A. E., Zuleta, C., Gantz, A., Saiz, F., Rau, J., Rosenmann, M., Cortés, A., Ruiz, G., Yates, L., Couve, E. & Marín, J. C., 1998. Sistemática y

adaptación de mamíferos, aves e insectos fitófagos de la Región de Antofagasta. Revista Chilena de Historia Natural, 71: 501–526. Spotorno, A. E., Zuleta, C. A., Valladares, P., Deane, A. L. & Jiménez, J. E., 2004b. Chinchilla laniger. Mammalian Species, 758: 1–9. Spotorno, A. E., Zuleta C. A., Walker, L. I., Manriquez, G., Valladares, P. & Marin, J. C., 2013. A small, new gerbil–mouse Eligmodontia (Rodentia: Cricetidae) from dunes at the coasts and deserts of north–central Chile: molecular, chromosomic, and morphological analyses. Zootaxa, 3683(4): 377–394. Tirado, C., Cortés, A., Miranda–Urbina, E. & Carretero, M. A., 2012. Trophic preference in an assemblage of mammal herbivores from Andean Puna (Northern Chile). Journal of Arid Enviroment, 79: 8–12. Valladares, P., 2002. Divergencia Molecular de las Especies Silvestres y Cepas Domesticadas del Género Chinchilla (Rodentia: Chinchillidae) Basada en el gen para citocromo b. Mastozoología Neotropical, 9: 96–98. – 2012. Mamíferos terrestres de la Región de Atacama. Comentarios sobre su distribución y estado de conservación. Gayana, 76(1): 13–28. Valladares, P., Espinoza, M., Torres, M., Díaz, E., Zeller, N., De la Riva, J., Grimberg, M. & Spotorno, A., 2012. Nuevo registro de Chinchilla chinchilla (Rodentia, Chinchillidae) para la Región de Atacama. Extensión de su rango de distribución y estado de conservación. Mastozoología Neotropical, 19(1): 173–178. Walker, R. S., Novaro, A. J., Perovic, P., Palacios, R., Donadio, E., Lucherini, M., Pia, M. & López, M. S., 2007. Diets of three species of Andean carnivores in high–altitude deserts of Argentina. Journal of Mammalogy, 88(2): 519–525. Walle, P., 1914. Bolivia, its people and its resources, its railways, mines, and rubber–forest. T. Fisher Unwin, London. Wolffsohn, J. A., 1923. Medidas máximas y mínimas de algunos mamíferos chilenos colectados entre los años 1896 y 1917. Revista Chilena de Historia Natural, 1: 159–167. Woods, C. A. & Kilpatrick, C. W., 2005. Infraorder Hystricognathi. In: Mammal Species of the World. A Taxonomic and Geographic Reference, Third Edition, Vol. 2: 1538–1600 (D. E. Wilson & D. M., Reeder, Eds.). The Johns Hopkins University Press, Baltimore, U.S.A.

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Is the 'n = 30 rule of thumb' of ecological field studies reliable? A call for greater attention to the variability in our data A. Martínez–Abraín

Martínez–Abraín, A., 2014. Is the 'n = 30 rule of thumb' of ecological field studies reliable? A call for greater attention to the variability in our data. Animal Biodiversity and Conservation, 37.1: 95–100. Abstract Is the 'n = 30 rule of thumb' of ecological field studies reliable? A call for greater attention to the variability in our data.— A common practice of experimental design in field ecology, which relies on the Central Limit Theorem, is the use of the 'n = 30 rule of thumb'. I show here that papers published in Animal Biodiversity and Conservation during the period 2010–2013 adjust to this rule. Samples collected around this relatively small size have the advantage of coupling statistically–significant results with large effect sizes, which is positive because field researchers are commonly interested in large ecological effects. However, the power to detect a large effect size depends not only on sample size but, importantly, also on between–population variability. By means of a hypothetical example, I show here that the statistical power is little affected by small–medium variance changes between populations. However, power decreases abruptly beyond a certain threshold, which I identify roughly around a five–fold difference in variance between populations. Hence, researchers should explore variance profiles of their study populations to make sure beforehand that their study populations lies within the safe zone to use the 'n = 30 rule of thumb'. Otherwise, sample size should be increased beyond 30, even to detect large effect sizes. Key words: Sample size, Variance, Statistical power, Effect size, Field ecology, Reliability. Resumen ¿Es fiable la regla de oro de n = 30 de los estudios ecológicos de campo? Se debe prestar más atención a la variabilidad de nuestros datos.— La utilización de la regla de oro de n = 30 es una práctica común del diseño experimental en ecología de campo que se fundamenta en el teorema del límite central. A continuación se muestra que los artículos publicados en Animal Biodiversity and Conservation durante el período comprendido entre los años 2010 y 2013 se ajustan a esta regla. Las muestras recogidas cuyo tamaño se aproxima a esta cifra relativamente pequeña tienen la ventaja de relacionar resultados estadísticamente significativos con efectos de gran magnitud, lo cual es positivo porque por lo general los investigadores de campo están interesados en los efectos ecológicos de gran magnitud. No obstante, la posibilidad de detectar un efecto de gran magnitud no solo depende del tamaño de la muestra, sino también en gran medida de la variabilidad existente entre las poblaciones. Mediante un ejemplo hipotético, a continuación se muestra que la potencia estadística se ve poco afectada por los cambios pequeños o medios de varianza que pueda haber entre las poblaciones. Sin embargo, la potencia se reduce bruscamente a partir de un determinado límite, que nosotros establecemos aproximadamente en una diferencia de cinco veces en la varianza entre poblaciones. Por consiguiente, los investigadores deberían analizar los perfiles de varianza de sus poblaciones de estudio con el fin de asegurarse de antemano de que sus poblaciones en estudio se encuentran en la zona de seguridad en que puede emplearse la regla de oro de n = 30. De lo contrario, será necesario aumentar el tamaño de la muestra a más de 30, incluso para detectar efectos de gran magnitud. Palabras clave: Tamaño de la muestra, Varianza, Potencia estadística, Magnitud del efecto, Ecología de campo, Fiabilidad. Received: 1 VIII 13; Conditional acceptance: 11 XI 13; Final acceptance: 20 XII 13

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Alejandro Martínez–Abraín, Depto. de Bioloxía Animal, Bioloxía Vexetal e Ecoloxía, Univ. da Coruña, Fac. de Ciencias, Campus da Zapateira s/n., 15071 A Coruña, España (Spain); Population Ecology Group, IMEDEA (CSIC–UIB), Miquel Marquès 21, 07190 Esporles, Mallorca, España (Spain). E–mail: amartinez@imedea.uib–csic.es

It is common to read in introductory books on biostatistics that working with a sample size of at least 30 is safe for the design of field studies (e.g. pg. 43 in Cohen & Cohen, 1995). This recommendation relies on the Central Limit Theorem, according to which if random samples of size n are drawn from a normal population the means of these samples will conform to a normal distribution (Zar, 1999). Supposedly random samples of a minimum size = 30 allow recovery of a normal distribution of the mean even if samples are non–normal. A common consequence is that field researchers tend to use samples adjusted to this minimum. To verify this tendency, I analyzed the sample size used in the papers published in Animal Biodiversity and Conservation from 2010 to 2013 (n = 4 years), and the results suggest that this rule holds for ecological and conservation field studies, since the arithmetic mean of both the medians and averages of sample sizes used in each paper was very close to 30 (see table 1). This is in contrast with experimental design recommendations, where sample size is known to be directly dependent on the variance in the population (as estimated by the sample standard deviation), and indirectly dependent on the maximum allowable absolute difference between the estimated population parameter and the true population parameter (d) from equation 1, nr

z2 σ2 d2

where z is 1.96, the value for a 95% confidence interval from a normal distribution (Quinn & Keough, 2002). That is, the minimum required sample size to accurately estimate a parameter will be higher if a) population variance is high, in order to minimize the risk of our sample not being representative of the statistical population, and b) if the desired accuracy is high, for a given confidence level, since increasing the confidence level also increases the required sample size. Of course, deciding the value of d means that we have previous knowledge about the true magnitude of the study parameter, which is not usually the case in field ecological studies (e.g. Martínez–Abraín, 2008, 2013).

Sample size and null hypothesis testing This is not only the theoretical basis of experimental design for parameter estimation, but also for a priori or prospective power tests within the framework of null hypothesis statistical testing (Zar, 1999; Schneir & Gurevitch, 2001). The required sample size to couple biological and statistical significance (and hence to make sense of statistically non–significant results) is determined after providing alpha (i.e. the type–I error rate, typically fixed at 5%), power (i.e. 1–the type II error rate or probability of correctly failing to reject a null hypothesis, typically fixed at 80%), and effect size (the minimum magnitude of the difference between two populations that is considered to be biologically relevant if dealing with mean comparison, or the amount of variance of each variable that is explained by other variables considered, if dealing with regression problems). Again, this implies that we have some a priori knowledge of what represents a biologically–relevant effect in our study system, which unfortunately is seldom the case in field ecological studies. Empirically, sample size could be obtained by plotting the standard deviation against sample size until a plateau is reached, provided that our sample correctly represents the variability in the population. However, this is mostly viable for experimental studies (preferentially lab studies, although also some field studies), where sample size can be modified along a large range of possible values. A better–suited option for field studies could be to perform this plotting by applying resampling with repetition (bootstrap) to our data, if our sampling protocol includes samples of different size. Moreover, this exercise would be necessary for each variable under study, because each variable has its own profile. This means that the usual procedure of measuring many variables from the same sample of individuals —taking advantage of having captured them, for example— does not respect the prerequisite of accounting for variance to determine the right sample size, because variances for each trait of an individual do not necessarily co–vary in a strong way. Different sample sizes would thus be necessary to study different traits, something that seems impracticable in field studies where information

Animal Biodiversity and Conservation 37.1 (2014)

Table 1. Sample size extracted from n = 76 papers published in Animal Biodiversity and Conservation from 2010 to 2013: V. Journal volume; I. Journal isue; i. Initial page; f. Final page; n. Sample size; M1. Median sample size; M2. Average sample size; E. Cause of exclusion (1. Species description; 2. Ring recoveries; 3. Essay; 4. Survey, hunting–bag data, bioacoustics/RADAR data; 5. Species atlas or similar; 6. Review papers; 7. Genetics; 8. Museum collections; 9. Simulated data). Tabla 1. Tamaño de la muestra extraída de n = 76 artículos publicados en Animal Biodiversity and Conservation entre 2010 y 2013: V. Volumen de la revista; I. Número de la revista; i. Página inicial; f. Página final; n. Tamaño de la muestra; M1. Mediana del tamaño de la muestra; M2. Media del tamaño de la muestra; E. Motivo de exclusión (1. Descripción de especies; 2. Recuperación de anillas; 3. Ensayo; 4. Encuesta, datos de caza, datos de bioacústica y radar; 5. Atlas de especies o similar; 6. Artículos de revisión; 7. Genética; 8. Colecciones museísticas; 9. Datos simulados).

Year

n n

n n n n M1 M2 E

2010 33 1 1 13 22 19 11 13 13.2 2010

33 1 15 18

2010

33 1 19 29

4 17 39 23 26 18 27 24 21 11 27 22 21.6

2010

33 1 31 45

10 20 30 53 30 27.7

2010

33 1 47 51

2010

33 1 53 61

7 53 40.3

2010

33 1 63 87

13 16 11 20 11

2010

33 1 89 115

2010

33 1 117 117

2010

33 2 119 129 50 50 50 50 50 50 50.0

2010

33 2 131 142

2010

33 2 143 150 16 16 21 25 23 21 20.2

2010

33 2 151 185

2010

33 2 187 194 30 30 30 30.0

2010

33 2 195 203

2010

33 2 205 208

2011

34 1 1

2011

34 1 11 21

39 41 19 31 20 10 25.5 26.7

2011

34 1 23 29

60 60 60 60 60 60.0

2011

34 1 31 34

2011

34 1 35 45

20 14 36 24 16 11 16 36 18 21.6

2011

34 1 47 66

2011

34 2 229 247 17

2011

34 2 249 256 13 132 72.5 72.5

2011

34 2 257 264

2011

34 2 265 272

2011

34 2 273 285

2011

34 2 287 294

2011

34 2 295 308

2011

34 2 309 317

2011

34 2 319 330 10 10 10.0

2011

34 2 331 340

2011

34 2 341 353

2011

34 2 355 361 37 25 22 25 28.0

2012

35 1 1

9 10 10 12 14 10 19 12.5 21.8

1 6

8.0 4 1

3 10 10.0 5

17 10.5 10.5

8.0

3 5

Martínez–Abraín

Table 1. (Cont.) Year

n n

n n n n M1 M2 E

2012 35 1 13 21

2012 35 1 23 26

2012

35 1 27 50

2012

35 1 51 58

2012

35 1 59 69

2012

35 1 71 94

2012

35 1 107 117 73 51 62 62.0

2012

35 1 119 124 29 26 39 35 36 33 34 33.0

2012

35 1 125 139 30 30 30.0

2012

35 1 141 150 10

2012

35 2 151

2012

35 2 153 154

2012

35 2 155

2012

35 2 159 161

2012

35 2 163 170

2012

35 2 171 174

2012

35 2 175 188

2012

35 2 189 196 18 10 13 15 54 44 33 28 23 26.9

2012

35 2 197 207 27 64 60

2012

35 2 209 217

2012

35 2 219 220

2012

35 2 221 233 27 27 27.0

2012

35 2 235 246

2012

35 2 247 252

2012

35 2 253 265

2012

35 2 267 275

2012

35 2 277 283

2012

35 2 285 293 48 28 16 16 16 16 16 16 16 4 16 19.2

2012

35 2 295 306

2013

36 1 1

2013

36 1 13 31

2013

36 1 33 36

2013

36 1 37 46

2013

36 1 47 57

2013

36 1 59 67

2013

36 1 69 78

2013

36 1 79 88

2013

36 1 89 99

2013

36 1 101 111

2013

36 1 113 121

2013

36 1 123 139

5 5

29 42

7.5

48 12 5 27 31.7 3

21 15 18 20.2 5

20 49 158 42 42 54.8

12 12 29 31 32 25

3 20.5 19.8

9 17 17.0

50 51 41 50 50 50 45 49 59 22 51 30 30 21 30 6 47 39.7 3

30 14 91 30 21 45 11 13 20 30 28 24.5 28.0

Mean 27.3 28.0 SD 17.2 16.7

Animal Biodiversity and Conservation 37.1 (2014)

Table 2. Change in statistical power as between–population variance increases, using the n = 30 rule of thumb of field ecological statistics. Exercise using fictitious data in open software G*Power 3.1.3: n. Sample size; M1, M2. Means of both populations; Sd1, Sd2. Standard deviations of both populations; Ratio. Ratio Sd1/Sd2; Es. Effect size (Cohen’s d); Po. Statistical power. Tabla 2. Cambio en la potencia estadística a medida que aumenta la varianza entre poblaciones, utilizando la regla de oro de n = 30 de los estadísticos de la ecología de campo. Ejercicio que utiliza datos ficticios en el programa informático abierto G*Power 3.1.3: n. Tamaño de la muestra; M1, M2. Medias de ambas poblaciones; Sd1, Sd2. Desviaciones estándar de ambas poblaciones; Ratio. Coeficiente Sd1/Sd2; Es. Magnitud del efecto (d de Cohen); Po. Potencia estadística. ID

Sd1

Sd2

Ratio

1.5

1.0

0.2

1.0

2.50

1.00

1.5

1.0

0.3

0.2

1.5

1.96

1.00

1.5

1.0

0.4

0.2

2.0

1.58

1.00

1.5

1.0

0.5

0.2

2.5

1.31

1.00

1.5

1.0

0.6

0.2

3.0

1.12

0.99

1.5

1.0

0.7

0.2

3.5

0.97

0.96

1.5

1.0

0.8

0.2

4.0

0.86

0.90

1.5

1.0

0.9

0.2

4.5

0.77

0.83

1.5

1.0

0.2

5.0

0.69

0.75

1.5

1.0

1.1

0.2

5.5

0.63

0.67

1.5

1.0

1.2

0.2

6.0

0.58

0.60

1.5

1.0

1.3

0.2

6.5

0.54

0.53

1.5

1.0

1.4

0.2

7.0

0.50

0.48

1.5

1.0

1.5

0.2

7.5

0.47

0.43

for each individual is exploited as much as possible given the difficulty in obtaining it. The 'rule of thumb of n = 30' in field ecological studies comes from the fact that we are usually interested in 'large' effect sizes (of unknown absolute magnitude). The reasoning follows that if we are able to obtain a statistically–significant result (this only meaning that the properties of our sample can be applied to the whole statistical population, and hence that the desired inference from particular to general can be done) with a small sample size such as 30, the effect we are dealing with is probably large, and hence, most likely a biologically–relevant effect (Martínez–Abraín, 2007). However, this approach of reasoning around n = 30 in relation to the magnitude of the effects (in the denominator of equation 1) is influenced by variance (in the numerator of equation 1). Power decreases with increasing between–population variance, when sample size is kept constant at n = 30 (table 2, fig. 1). However, this decrease proceeds in a non–linear fashion, indicating a strong resilience of statistical power to small–to–medium changes in between–population variance. Only when the change in between–populations variance is large (around a

five–fold difference in the variance between groups, which corresponds to our study case #9) does power drop abruptly below the desired minimum value of 0.8 (fig. 1). In this hypothetical example, it is necessary to increase our sample size to n = 34 in case study #9, and to n = 73, in case study #14, to allow the recovery of a 0.80 power. The n = 30 rule of thumb also overlooks the possibility that small or medium effect sizes can be biologically relevant in some cases (Igual et al., 2005). Since we typically do not know when that is the case, we are forced or limited to work with large effect sizes. On the contrary, working with too large a sample size (as is commonplace among theoretical ecologists) could even be counter–productive at times because we could be focusing on small effects which could be biologically irrelevant. Hence, it seems reasonable to use the n = 30 rule in ecological field studies to make inferences on parameter values or to test null hypotheses, owing to our usual lack of knowledge on d or effect size of interest, but we should make an effort to explore beforehand the variance profile of our study populations in order to be able to detect large effects with that sample size and with a high power. Populations with

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Martínez–Abraín

1.2 1

Power

0.8 0.6 0.4 0.2

2 3 4 5 6 Relative change in standard deviation

Fig. 1. Non–linear decrease of statistical power with increasing variance between two populations (ratio of standard deviations between two groups) while keeping means and sample sizes constant (n = 30), for data in table 2. The reference line for a desired power of 0.80 is shown as well as the line for a five–fold difference in between–populations variance. Fig. 1. Para los datos que figuran en la tabla 2, la potencia estadística disminuye de forma no lineal a medida que aumenta la varianza entre dos poblaciones (coeficiente de las desviaciones estándar entre dos grupos) y con las medias y los tamaños de las muestras (n = 30) constantes. Se muestran también la línea de referencia de una potencia deseada de 0,80 y la línea que representa una diferencia de cinco veces en la varianza entre poblaciones.

high differences between them in variance will require larger sample sizes (compared to 30) to detect even large differences. Rules of thumb exist for a reason, but they should be used with great caution and as an approximation. The exercise of thinking, typical of the scientific enterprise, cannot and should not be set aside during the process of experimental design, despite the added complexities of ecological field studies compared to lab studies. Reduced power, such as increases in between–populations variance, can lead to increased prevalences of Type II errors (i.e. incorrectly failing to reject a null hypothesis of equality to zero), resulting in serious problems regarding decision–making in conservation, when evaluating the effect of human activities. We may conclude that nothing happens when indeed it does. Let’s hence give more attention to the variability of our field data for the benefit of proper knowledge acquisition and correct decision making. Acknowledgements I would like to thank the editor and two anonymous referees who helped improve the manuscript with their critical comments. Daniel Oro commented on an early version of the manuscript.

References Cohen, J. & Cohen, L., 1995. Statistics for ornithologists. BTO Guide 22. Igual, J. M., Forero, M. G., Tavecchia, G., González–Solís, J., Martínez–Abraín, A., Hobson, K. A., Ruiz, X., Oro, D., 2005. Short–term effects of data–loggers on Cory’s shearwater (Calonectris diomedea). Marine Biology, 146: 619–624. Martínez–Abraín, A., 2007. Are there any differences? A non–sensical question in ecology. Acta Oecologica, 32: 203–206. – 2008. Statistical significance and biological relevance: A call for a more cautious interpretation of results in ecology. Acta Oecologica, 34: 9–11. – 2013. Why do ecologists aim to get positive results? Once again, negative results are necessary for better knowledge accumulation. Animal Biodiversity and Conservation, 36.1: 33–36. Quinn, G. P. & Keough, M. J., 2002. Experimental design and data analysis for biologists. Cambridge University Press, Cambridge. Schneir, S. M. & Gurevitch, J. (Eds.), 2001. Design and analysis of ecological experiments. Oxford University Press, Oxford. Zar, J. H., 1999. Biostatistical analysis. Prentice Hall, Upper Saddle River, New Jersey.

Animal Biodiversity and Conservation 37.1 (2014)

Forum 101

Improving the reviewing process in ecology and evolutionary biology G. D. Grossman

Grossman, G. D., 2014. Improving the reviewing process in ecology and evolutionary biology. Animal Biodiversity and Conservation, 37.1: 101–105. Abstract Improving the reviewing process in ecology and evolutionary biology.— I discuss current issues in reviewing and editorial practices in ecology and evolutionary biology and suggest possible solutions for current problems. The reviewing crisis is unlikely to change unless steps are taken by journals to provide greater inclusiveness and incentives to reviewers. In addition, both journals and institutions should reduce their emphasis on publication numbers (least publishable units) and impact factors and focus instead on article synthesis and quality which will require longer publications. Academic and research institutions should consider reviewing manuscripts and editorial positions an important part of a researcher’s professional activities and reward them accordingly. Rewarding reviewers either monetarily or via other incentives such as free journal subscriptions may encourage participation in the reviewing process for both profit and non–profit journals. Reviewer performance will likely be improved by measures that increase inclusiveness, such as sending reviews and decision letters to reviewers. Journals may be able to evaluate the efficacy of their reviewing process by comparing citations of rejected but subsequently published papers with those published within the journal at similar times. Finally, constructive reviews: 1) identify important shortcomings and suggest solutions when possible, 2) distinguish trivial from non–trivial problems, and 3) include editor’s evaluations of the reviews including identification of trivial versus substantive comments (i.e., those that must be addressed). Key words: Publication process, Reviewing, Editorial, Editors. Resumen Mejora del proceso de revisión de artículos en ecología y biología evolutiva.— Se debaten los problemas actuales de la revisión y las prácticas editoriales en los campos de la ecología y la biología evolutiva, y se sugieren posibles soluciones para los mismos. La crisis por la que está pasando la revisión no cambiará a menos que las revistas tomen medidas para aumentar la inclusividad de los revisores y los incentivos a los mismos. Asimismo, tanto las revistas como las instituciones deberían prestar menos atención a las cifras relativas a la publicación (las unidades mínimas publicables) y los factores de impacto, y centrar el interés en la síntesis y la calidad de los artículos, lo que exigirá que las publicaciones sean más largas. Las instituciones académicas y de investigación deberían considerar la revisión de los manuscritos y las posturas de las editoriales como una parte importante de las actividades profesionales de un investigador, y compensarlas en consecuencia. Recompensar a los revisores, ya sea económicamente o con otros incentivos, como suscripciones gratuitas a revistas, puede alentar la participación en el proceso de revisión, para las revistas con y sin ánimo de lucro. Probablemente pueda mejorarse el rendimiento de los revisores con medidas que aumenten la inclusividad, como el envío a los revisores de las revisiones y las notificaciones de las decisiones adoptadas. Las revistas tal vez puedan evaluar la eficacia de sus procesos de revisión comparando las citas de los artículos rechazados que se hayan publicado posteriormente con las de los que se publicaron en la revista en el mismo momento. Por último, las revisiones constructivas deben: 1) determinar las deficiencias importantes y sugerir soluciones siempre que sea posible, 2) distinguir los problemas triviales de los que no lo sean y 3) contener las evaluaciones que el editor haga de las revisiones, incluida la determinación de las observaciones triviales y las sustantivas (las que deben abordarse). Palabras clave: Proceso de publicación, Revisión, Editorial, Editores. Received: 23 I 14; Conditional acceptance: 24 III 14; Final acceptance: 4 IV 14 Gary D. Grossman, Warnell School of Forestry & Natural Resources, Univ. of Georgia, Athens GA 30602, U.S.A.

ISSN: 1578–665 X eISSN: 2014–928 X

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Discussion of shortcomings in the peer review process and editorial practices within scientific journals likely started with publication of the first journal and employment of the first editor. Recently, multiple aspects of this topic have been described in publications dealing with ecology and evolutionary biology (EEB) (Hochberg et al., 2009; Mesnard, 2010; Statzner & Resh, 2010; Albuquerque, 2011; Rohr & Martin, 2012a) and the subject has received considerable attention in the biomedical research community (Smith, 2006; Tite & Schroter, 2007). Multiple critical issues of the peer review process have been raised including: 1) the difficulties of finding good reviewers*, 2) the lack of reward for reviewing, 3) the increased number of manuscripts submitted to journals exacerbating issue 1, and 4) negative institutional policies reduce incentives for participating in the editorial process. Although a number of potential solutions to the reviewer crisis have been suggested, there is little consensus regarding what should be done (DeVries et al., 2009; Montesinos, 2012; Rohr & Martin, 2012b; Duffy, 2013) and there appear to be few changes in editorial practices by journals (Grod et al., 2010). In this paper I will discuss additional issues contributing to the reviewer crisis and propose several additional solutions. Much of what I report is based on my own experiences as an author of 110+ papers and also as a reviewer/editorial board member/associate editor for five different journals run by both scientific societies and commercial publishers. Publication proliferation There is no doubt that the number of manuscripts submitted for publication in scientific journals has increased substantially in the last few decades, primarily due to an increase in the number of scientists. In addition, the pressures of promotion and high competition for jobs in the last four decades contribute to the pressure to 'slice' publications into what historically has been known as the Least Publishable Unit (LPU) or 'salami tactic'. The combination of increasing publication frequency and decreasing publication length was recognized decades ago (Broad, 1981; Lyman, 2013) and is one of the main factors contributing to the reviewer crises in EEB. There is no doubt that many will judge a scientist’s performance based on publication quantity rather than quality, and this is likely true for most scientific fields. The phenomenon itself is most easily observed in discussions of faculty search or tenure/promotion committees. It is clear that overall productivity (i.e., number of publications) should play a role in evaluations, but first assessments (and cuts) typically are made using simple criteria such as 'number of publications in refereed journals'. This criterion is easy, quick and may even be correlated with quality, but it also encourages vita padding. It is easily gamed by dividing larger potential research publications into LPUs, which contribute significantly to the editorial burden of the EEB community. Nonetheless, I doubt

* I will use the term reviewer and referee interchangeably.

Grossman

that publication frequency will ever disappear as an assessment criterion, but perhaps journal editors and referees should be more stringent in accepting papers that clearly are small slices of a complete pie. The LPU syndrome has been exacerbated by the proliferation of journals in EEB (Statzner & Resh, 2010); including the explosion of 'Letters' (i.e., short format) and open–access journals (Bohannon, 2013), all of which require enough papers to regularly fill issues. Some researchers appear to think that the publication process is slower than it was 25 years ago (Statzner & Resh, 2010), but recent studies provide surprising answers to that question. For example, there has been no demonstrable increase in average review time for journals in either behavioral sciences or natural history between 1980 and 2012 (Pautasso & Schaefer, 2010; Lyman, 2013). In addition, although there is a positive correlation between impact factor of a journal and the number of manuscripts submitted, there also is a negative trend between impact factor and time to acceptance (Pautasso & Schaefer, 2010). Hence, higher number of submissions does not necessarily result in more extensive editorial delays (Pautasso & Schaefer, 2010). It is possible; however, that the latter result is a consequence of many papers being rejected by journals without review (Pautasso & Schaefer, 2010) as has been the policy of a number of prominent EEB journals. This practice, although providing a quick turn–around for a manuscript, is quite susceptible to bias and cliquishness in publication, as noted in1974 (VanValen & Pitelka, 1974) and still in evidence today (Arnqvist, 2013). Nonetheless, in contrast to the results of Pautasso and Schaefer (2010) a recent survey of EEB editors showed a negative relationship between the number of papers handled and the proportion rejected without review (McPeek et al., 2009). The referee pool Given the increasing number of both journals and submissions, coupled with a pool of experienced referees that while increasing, still is insufficient to handle the current load (Hauser & Fehr, 2007; Statzner & Resh, 2010; Arnqvist, 2013; Duffy, 2013), it is obvious that the EEB community has yet to effectively deal with the ‘reviewer crises’. Several investigators have suggested ideas for dealing with the decreased willingness of referees to perform reviews, the high number of review requests received by ‘good referees’, and issues of review quality (Hauser & Fehr, 2007; Fox & Petchey, 2010; Rohr & Martin, 2012a; Duffy, 2013). These suggestions involve punishing slow reviewers and rewarding timely referees who provide thorough reviews, but as all authors admit, these solutions may do little to prevent some scientists employing ‘cheater’ strategies. Nonetheless, they all are right that changes are necessary to improve the current status of reviewing. Perhaps referees are no more nor less altruistic than they have been in the past, but what has changed

Animal Biodiversity and Conservation 37.1 (2014)

in the last 30 years are the external constraints on a researcher’s time. Most university researchers, at least in the United States, are now faced with a plethora of administrative responsibilities from both their own universities and governmental sources (e.g., faculty committees, training sessions for compliance with laws such as the Family Educational Rights and Privacy Act [FERPA], monthly documentation of graduate student performance, Institutional Animal Use and Care Committee (IACUC) requirements and training, federal data accessibility requirements, etc.). Concomitantly, both university and federal research budgets have been slashed in the United States and other countries; consequently researchers must devote much more time to seeking research funding than they have in the past. This is one of the major reasons referees are slow or reluctant to review papers; simply put, there is little time or energy left after performing one’s daily research responsibilities (Statzner & Resh, 2010). At the same time, the qualifications needed to obtain a research or faculty position are increasing (Statzner & Resh, 2010). Hence, even if someone is a 'good Samaritan' (McPeek et al., 2009), there are strong selective pressures acting against altruism, even if they are merely perceived rather than real. There is no solution to this problem until reviewing manuscripts, and editorial work in general, are viewed as normative responsibilities, with appropriate recognition and rewards from administrators. I suspect that in most institutions, editorial board membership or extensive reviewing rarely results in raises, increased release time or help from support staff. My supposition is that administrators resort to claims like ‘well everyone does that so we can just assume that it is a constant across faculty’ but the current crises suggest that reviewing and editorial work are not constant across faculty. In addition, an erroneous assumption by administrators that reviewing is equal across faculty promotes 'cheaters' who do no reviewing and devote all their time to writing grants or papers instead, especially when promotion decisions are made on a comparative basis. Faculty must become more proactive in demanding that incentives be provided for highly active and competent reviewers and associate editors, and managing editors should support them in this quest. Hopefully, this will result in administrators providing substantive rewards for participation in the editorial process as well as penalties for faculty who do not participate. How can journals and editors improve the situation? At present, there appear to be few journals that provide incentives for reviewers. A few journals provide free access to online versions of the journal although frequently this only extends over a month or two. Certainly one perquisite for reviewers that could increase referee responsiveness would be to give a free online subscription to the journal after a given number of reviews in a year. Even non–profit scientific societies could employ this incentive because it is not costly. Incentives could be provided on a graduated scale

103

where it might take four reviews in a year to obtain free access for a year, and a single review might earn only three months access. Of course this may penalize members of scientific societies who already receive a journal subscription, but they still might not have online access or they could be rewarded with free access in the next year or access to a journal they do not receive (many scientific societies publish multiple journals). No incentive scheme is perfect but it seems that some experimentation is called for at the present time, given the repeatedly voiced concerns by both editors and authors. It is possible that paying referees for reviews could improve both referee participation and performance, but its discussion mostly has occurred on online forums. I have found no published evaluation of this practice in EEB, although EEB outside examiners are paid by universities for dissertation reviews in both Australia and New Zealand and likely other countries. In addition, multiple European countries (Ireland, Poland and Spain) pay for proposal reviews, as does at least one commercial publisher for editorial board work. Nonetheless, a study of biomedical reviewers found that reviewers had mixed opinions regarding the positive impacts of financial rewards on the reviewing process (Tite & Schroter, 2007). The biggest objections to payment for services involve the end of volunteerism, and the assumption that financial rewards will bias the reviewing process, or pull referees away from journals that cannot provide incentives. There is a lack of evidence but I suspect this is unlikely. From a philosophical perspective, I deplore the loss of the volunteer ethic in science; however, the current crisis seems immune to philosophical regrets and perhaps represents the triumph of the market economy even in science. One of my goals is to suggest possible approaches leading to data on potential strategies to resolve the reviewer crisis. It would be useful for an EEB journal or society to conduct an experiment in which some reviewers are paid and others not and then compare the quality, timing, and responsiveness of the two reviewer treatment groups. There is no doubt that such an experiment would require a sophisticated design and still likely present logistical hurdles, however, it should aid in determining whether financial rewards would improve the reviewing process in EEB. Finally, it is true that payment for reviewers and editorial work may present logistical and financial difficulties for non–profit journals, however these obstacles are mostly irrelevant for the many journals published by highly profitable commercial publishers or open–access journals with high publication fees. One of the reasons for the poor performance of reviewers is that too many journals fail to cultivate a culture of inclusion in the editorial process. I suspect reviewer performance would be substantively improved if journals practiced a few simple steps that demonstrated the importance of individual reviews in the overall editorial process. For example, although some journals provide a reviewer with all reviews of a manuscript and editor’s decision letter, too many do not. Reading the comments of other reviewers and the editor always is an educational experience and is an excellent me-

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chanism for less experienced reviewers to learn from more experienced reviewers. In addition, it would be beneficial for everyone involved if editors explained their reasoning when they overrule a referee. Finally, I wonder how much effort is expended by journals in evaluating whether their editorial practices are efficient and unbiased, or whether the prevailing attitude is one of laissez faire (Grod et al., 2010). Certainly one way that journals could evaluate the accuracy of their reviewing practices would be to compare citation frequencies of a random sample of articles rejected by the journal but subsequently published in other journals with a sample of articles accepted in that same year. Although citation frequencies are not a perfect metric of quality, they are easily obtained and certainly indicative of quality if the citations are positive. Such an analysis should be conducted with historic data, for example volumes published 10, 7 and 4 years previously. If no difference exists between citation frequencies of the two sets of papers, and assuming that the rejected papers were appropriate subject matter for both journals, then it would be cause for examining historic editorial practices, or to determine if specific associate editors were the cause of these rejections. Of course the citations would have to be checked randomly to assure that the citations were comparable (i.e., to avoid the case where total citations are equal but one paper has all positive citations and the other has all negative citations). It also might help identify continuing trends in problematic editorial practices. In addition, recent work has shown that factors such as journal impact factors may affect reviews independent of manuscript quality, and that reviewer ratings of the same manuscript may not be highly correlated (Eyre–Walker & Stoletzki, 2013). It is likely that there is little formal or quantitative evaluation of associate editors for many journals, except where an editor’s behavior becomes intolerable, such as failing to act on multiple manuscripts. These issues all call for journals to evaluate the accuracy and precision of their reviewing policies. Improving reviews and reviewing There is no doubt that high quality referees and editors are both typically overworked. Nonetheless, if editors believe that reviewers should not use this excuse, then neither should they. My own experience suggests there has been a decline in the quality of review interpretation and decisions made by editors as well as a general decline in review quality. I have already mentioned fostering a sense of inclusion for referees in the editorial process and (Statzner & Resh, 2010) have covered many of the current negative trends in the editorial process. Having published my first paper in 1977, I have seen just about every constructive and inane comment possible, typically with no comments from the editor on inappropriate or obviously erroneous comments. I believe that it is an editor’s responsibility to ensure that an editorial decision letter does not come back to an author without commentary on the reviews. At the very least, editors should identify reviewer’s comments that must be addressed versus those that are optional.

Grossman

Nonetheless, the evaluation of reviewer’s comments by editors certainly is not general policy for scientific journals. Given the complaints by editors regarding the poor quality of many reviews, this is not a trivial issue, yet most editors provide an author with little guidance other than 'all comments must be addressed, especially revisions that you do not incorporate'. But how much detail must be provided by an author when a comment clearly is erroneous: a not infrequent situation? This can be particularly problematical for young scientists, especially given the many picayunish negative comments written by reviews of today. Frankly, if editors are actually reading reviews closely, as they should, then it does not take much more time to identify which comments need to be addressed and which do not. After all, how can an editor reach an informed decision without evaluating reviews, even when both ratings are reject? Every author deserves at least this much from an editor. An additional problem of today is that the category of 'accepted with revision' seems to have disappeared from many journals and instead the author is told that their manuscript has landed in the large gray category called ‘not acceptable in this form’. I have spoken with many researchers, especially young researchers, who have interpreted this as a rejection, when in fact it really is just code for 'significant revision'. Nonetheless, some editors have justified this change by saying that it was difficult to obtain substantive revisions from authors once the term 'accepted' had been used. What constitutes a good review? A thorough discussion of the reviewing process is provided by DeVries et al. (2009), an article that is particularly useful for young scientists. An interesting psychological question for both editors and reviewers is whether a paper should be viewed as acceptable until a sufficient number of problems render it unacceptable, or whether papers should be viewed as unacceptable until a sufficient number of positive points are identified so that it becomes acceptable. I favor the first view point, mainly because I believe it leads to more constructive reviewing and hopefully a more positive experience for the authors, even when a paper is rejected. Many journals do not have review templates that ask a reviewer to specifically identify both the strengths and weaknesses of the manuscript but this would lead to more objective reviewing and improved editorial decisions. For both referees and editors, clearly the criterion for any comment is whether or not it is truly constructive. Probably the most significant improvement would be to require referees to reference their criticisms. I have seen comments ranging from 'this simply is wrong' to the 'literature review was inadequate' without any subsequent explanation of why a given technique was wrong nor any subsequent listing of missing papers. Such comments are completely unhelpful to the author and certainly do not fall under the rubric of 'constructive criticisms'. It is not the reviewer’s responsibility to rewrite an author’s manuscript; nonetheless, unconstructive comments and reviews help no one and eventually result in a bad reputation for a journal. I know more than one scientist who simply has

Animal Biodiversity and Conservation 37.1 (2014)

stopped submitting manuscripts to journals that have persistently poor reviewing policies even when they have high impact factors. Nonetheless, clearly this is a luxury of the full professor, not the untenured assistant professor. A final comment on writing style is warranted, given that many current referees seem to have little tolerance for a style different from their own. I have received reviews stating that a manuscript is poorly written without any description of how this judgment was reached, let alone an 'example' paragraph that was rewritten to demonstrate good writing. In addition it is not uncommon to receive reviews in which one reviewer ranks the paper as well written while another says it is poorly written. Once again, this is the type of comment that should prompt an editor’s intervention but this is rare in my experience. Consequently, if you cannot identify specific problems in grammar, clarity or verbosity accompanied by examples of how this can be corrected, then it is likely that you and the author have different writing styles, and it should be left at that. An even more problematical stylistic issue is that of non–native English writers, and the level of grammatical 'stretch' that should be allowed in such manuscripts (Clavero, 2010) As with any large volunteer enterprise, problems exist with the current peer review system and whether or not they will be fixed depends on the EEB community itself. Nonetheless, I hope that the suggestions made in this paper are helpful, even if they only lead to small improvements in the overall EEB editorial system. Most importantly, journals should begin conducting experiments regarding changes in editorial practices that may improve the various aspects of the 'reviewing crisis', and ultimately communicate the results of these experiments to the EEB community. Acknowledgements I apologize in advance for any omissions contained in this article and to any journals, administrators, etc. who already employ the suggestions in this article. I am sure they are out there and should be congratulated. I would like to thank D. DeVries, E. Garcia–Berthou, A. Hildrew, D. Jackson, M. McCallum, V. Resh, J.–C. Senar, J. Schaefer and J. Rohr for thoughtful comments on the manuscript, and my family for their ever present support. In addition, the reviewers: Mario Diaz and Sara Schroter provided insightful commentary on the ms. Conceptual stimulation for this paper was aided by Jittery Joe's and Two Story. Finally, the Warnell School of Forestry and Natural Resources provided material support for this paper. References Albuquerque, U. P., 2011. The tragedy of the common reviewers – the peer review process. Rev. Bras. Farmacogn. Braz. J. Pharmacogn., 21: 1–3. Arnqvist, G., 2013. Editorial rejects? Novelty, schnovelty! Trends Ecol. Evol., 28: 448–449. Bohannon, J., 2013. Who's Afraid of Peer Review?

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Science, 342: 60–65. Broad, W. J., 1981. The publishing game: getting more for less. Science, 211: 1137–1139. Clavero, M., 2010. 'Awkward wording, rephrase': linguistic injustice in ecological journals. TREE, 25: 552–553. DeVries, D. R., Marschall, E. A. & Stein, R. A., 2009. Exploring the peer review process: what is it does it work and can it be improved? Fisheries, 34: 270–279. Duffy, D. C., 2013. Reviewing reviewers. The Scientist http://www.the–scientist.com/?articles.view/articleNo/36575/title/Opinion––Reviewing–Reviewers/ Eyre–Walker, A. & Stoletzki, N., 2013. The Assessment of Science: The Relative Merits of Post–Publication Review, the Impact Factor, and the Number of Citations. PLoS Biol, 11(10): e1001675. Doi:10.1371/ journal.pbio.1001675. Fox, J. & Petchey, O. L., 2010. Pubcreds: fixing the peer review process by 'privatizing' the reviewer commons. Bull. Ecol. Soc. Am., 91: 325–333. Grod, O. N., Lortie, C. J. & Budden, A. E., 2010. Behind the shroud: a survey of editors in ecology and evolution. Front. Ecol. Environ., 8: 187–192. Hauser, M. & Fehr, E., 2007. An incentive solution to the peer review problem. PLoS Biol., 5: 703. Hochberg, M. E., Chase, J. M., Gotelli, N. J., Hastings, A. & Naeem, S., 2009. The tragedy of the reviewer commons. Ecol. Lett., 12: 2–4 Lyman, R. L., 2013. Three–Decade History of the Duration of Peer Review. J. Scholarly Pub., 44: 211–220. McPeek, M. A., DeAngelis, D. L, Shaw, R. G., Moore, A. J., Rausher, M. D., Strong, D. R., Ellison, A. M., Barrett, L., Rieseberg, L., Breed, M. D., Sullivan, J., Osenberg, C. W., Holyoak, M. & Elgar, M. A., 2009. The golden rule of reviewing. Am. Nat., 173(5): E155–E158. Mesnard, L., 2010. On Hochberg et al.'s 'the tragedy of the reviewer commons'. Scientometrics, 84: 903–917. Montesinos, D., 2012. Type I error hinders recycling: a response to Rohr and Martin. Trends Ecol. Evol., 27: 311–312. Pautasso, M. & Schaefer, H., 2010. Peer review delay and selectivity in ecology journals. Scientometrics, 84: 307–315. Rohr, J. R. & Martin, L. B., 2012a. Reduce, reuse, recycle scientific reviews. Trends Ecol. Evol., 27: 192–193. – 2012b. Type I error is unlikely to hinder review recycling: a reply to Montesinos. Trends Ecol. Evol., 27: 312–313. Smith, R., 2006. The trouble with medical journals. J. R. Soc. Med., 99: 115–119. Statzner, B. & Resh, V. H., 2010. Negative changes in the scientific publication process in ecology: potential causes and consequences. Freshwat. Biol., 55: 2639–2653. Tite, L. & Schroter, S., 2007. Why do peer reviewers decline to review? A survey. J. Epidem. Comm. Health, 61: 9–12. VanValen, L. & Pitelka, F., 1974. Intellectual Censorship in Ecology. Ecology, 55: 925–926.

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Animal Biodiversity and Conservation

Manuscrits

Animal Biodiversity and Conservation és una revista interdisciplinària publicada, des de 1958, pel Museu de Ciències Naturals de Barcelona. Inclou articles d'investigació empírica i teòrica en totes les àrees de la zoologia (sistemàtica, taxonomia, morfologia, biogeografia, ecologia, etologia, fisiolo� gia 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 com� pilacions 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 execu� tiu, un editor i dos revisors independents, triats d'una llista internacional, a fi de garantir–ne la qualitat. El procés de revisió és ràpid i constructiu. La publicació dels treballs acceptats es fa normalment dintre dels 12 mesos posteriors a la recepció. Una vegada hagin estat acceptats passaran a ser propietat de la revista. Aquesta es reserva els drets d’autor, i cap part dels treballs no podrà ser reproduïda sense citar–ne la procedència.

Els treballs seran presentats en format DIN A–4 (30 línies de 70 espais cada una) a doble espai i amb totes les pàgines numerades. Els manuscrits han de ser complets, amb taules i figures. No s'han d'enviar les figures originals fins que l'article no hagi estat acceptat. El text es podrà redactar en anglès, castellà o català. Se suggereix als autors que enviïn els seus treballs en anglès. La revista els ofereix, sense cap càrrec, un servei de correcció per part d'una persona especialitzada en revistes científiques. En tots els casos, els textos hauran de ser redactats correctament i amb un llenguatge clar i concís. La redacció del text serà impersonal, i s'evitarà sempre la primera persona. Els caràcters cursius s’empraran per als noms científics de gèneres i d’espècies i per als neologis� mes intraduïbles; les cites textuals, independentment de la llengua, seran consignades en lletra rodona i entre cometes i els noms d’autor que segueixin un tàxon aniran en rodona. Quan se citi una espècie per primera vegada en el text, es ressenyarà, sempre que sigui possible, el seu nom comú. Els topònims s’escriuran o bé en la forma original o bé en la llengua en què estigui escrit el treball, seguint sempre el mateix criteri. Els nombres de l’u al nou, sempre que estiguin en el text, s’escriuran amb lletres, excepte quan precedeixin una unitat de mesura. Els nombres més grans s'escriuran amb xifres excepte quan comencin una frase. Les dates s’indicaran de la forma següent: 28 VI 99 (un únic dia); 28, 30 VI 99 (dies 28 i 30); 28–30 VI 99 (dies 28 a 30). S’evitaran sempre les notes a peu de pàgina.

Normes de publicació Els treballs s'enviaran preferentment de forma electrònica (abc@bcn.cat). El format preferit és un document Rich Text Format (RTF) o DOC que inclogui les figures i les taules. Les figures s'hauran d'enviar també en arxius apart en format TIFF, EPS o JPEG. Cal incloure, juntament amb l'article, una carta on es faci constar que el treball està basat en investigacions originals no publicades anterior ment i que està sotmès a Animal Biodiversity and Conservation en exclusiva. A la carta també ha de constar, per a aquells treballs en que calgui manipular animals, que els autors disposen dels permisos necessaris i que compleixen la normativa de protecció animal vigent. També es poden suggerir possibles assessors. Quan l'article sigui acceptat, els autors hauran d'enviar a la Redacció una còpia impresa de la versió final acompanyada d'un disquet indicant el progra� ma utilitzat (preferiblement en Word). Les proves d'impremta enviades a l'autor per a la correcció, seran retornades al Consell Editor en el termini de 10 dies. Aniran a càrrec dels autors les despeses degudes a modificacions substancials introduïdes per ells en el text original acceptat. El primer autor rebrà una còpia electrònica del treball en format PDF. ISSN: 1578–665X eISSN: 2014–928X

Format dels articles Títol. Serà concís, però suficientment indicador del contingut. Els títols amb designacions de sèries numèriques (I, II, III, etc.) seran acceptats previ acord amb l'editor. Nom de l’autor o els autors Abstract en anglès que no ultrapassi les 12 línies mecanografiades (860 espais) i que mostri l’essència del manuscrit (introducció, material, mètodes, resultats i discussió). S'evitaran les especulacions i les cites bibliogràfiques. Estarà encapçalat pel títol del treball en cursiva. Key words en anglès (sis com a màxim), que orientin sobre el contingut del treball en ordre d’importància. Resumen en castellà, traducció de l'Abstract. De la traducció se'n farà càrrec la revista per a aquells autors que no siguin castellanoparlants. Palabras clave en castellà. Adreça postal de l’autor o autors. (Títol, Nom, Abstract, Key words, Resumen, Pala� bras clave i Adreça postal, conformaran la primera pàgina.)

Introducción. S'hi donarà una idea dels antecedents del tema tractat, així com dels objectius del treball. Material y métodos. Inclourà la informació perti� nent de les espècies estudiades, aparells emprats, mètodes d’estudi i d’anàlisi de les dades i zona d’estudi. Resultados. En aquesta secció es presentaran úni� cament les dades obtingudes que no hagin estat publicades prèviament. Discusión. Es discutiran els resultats i es compa� raran amb treballs relacionats. Els suggeriments de recerques futures es podran incloure al final d’aquest apartat. Agradecimientos (optatiu). Referencias. Cada treball haurà d’anar acompanyat de les referències bibliogràfiques citades en el text. Les referències han de presentar–se segons els models següents (mètode Harvard): * Articles de revista: Conroy, M. J. & Noon, B. R., 1996. Mapping of spe� cies richness for conservation of biological diversity: conceptual and methodological issues. Ecological Applications, 6: 763–773. * Llibres o altres publicacions no periòdiques: Seber, G. A. F., 1982. The estimation of animal abundance. C. Griffin & Company, London. * Treballs de contribució en llibres: Macdonald, D. W. & Johnson, D. P., 2001. Dispersal in theory and practice: consequences for conserva� tion biology. In: Dispersal: 358–372 (T. J. Clober, E. Danchin, A. A. Dhondt & J. D. Nichols, Eds.). Oxford University Press, Oxford. * Tesis doctorals: Merilä, J., 1996. Genetic and quantitative trait vari� ation in natural bird populations. Tesis doctoral, Uppsala University. * Els treballs en premsa només han d’ésser citats si han estat acceptats per a la publicació: Ripoll, M. (in press). The relevance of population studies to conservation biology: a review. Animal Biodiversity and Conservation.

La relació de referències bibliogràfiques d’un tre� ball serà establerta i s’ordenarà alfabèticament per autors i cronològicament per a un mateix autor, afegint les lletres a, b, c,... als treballs del mateix any. En el text, s’indicaran en la forma usual: "... segons Wemmer (1998)...", "...ha estat definit per Robinson & Redford (1991)...", "...les prospeccions realitzades (Begon et al., 1999)...". 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 ressen� yades 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 au� tors. 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 (Intro� ducción, 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.

Animal Biodiversity and Conservation 37.1 (2014)

Animal Biodiversity and Conservation Animal Biodiversity and Conservation es una revista interdisciplinar, 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, taxonomía, morfología, biogeografía, ecología, etología, fisiología y genéti� ca) 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 biodi� versidad 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 re� gistrada 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 garan� tizar 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 propie� dad de la revista. Ésta se reserva los derechos de autor, y ninguna parte del trabajo podrá ser reprodu� cida sin citar su procedencia.

Normas de publicación Los trabajos se enviarán preferentemente de forma electrónica (abc@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. Cuando el trabajo sea aceptado los autores de� berán enviar a la Redacción una copia impresa de la versión final junto con un disquete del manuscrito preparado con un procesador de textos e indicando el programa utilizado (preferiblemente Word). Las pruebas de imprenta enviadas a los autores deberán

ISSN: 1578–665X eISSN: 2014–928X

III

remitirse corregidas al Consejo Editor en el plazo máximo de 10 días. Los gastos debidos a modifica� ciones sustanciales en las pruebas de imprenta, intro� ducidas por los autores, irán a cargo de los mismos. El primer autor recibirá una copia electrónica del trabajo en formato PDF. Manuscritos Los trabajos se presentarán en formato DIN A–4 (30 lí� neas de 70 espacios cada una) a doble espacio y con las páginas numeradas. Los manuscritos deben estar completos, con tablas y figuras. No enviar las figuras originales hasta que el artículo haya sido aceptado. El texto podrá redactarse en inglés, castellano o catalán. Se sugiere a los autores que envíen sus trabajos en inglés. La revista ofrece, sin cargo ningu� no, un servicio de corrección por parte de una persona especializada en revistas científicas. En cualquier caso debe presentarse siempre de forma correcta y con un lenguaje claro y conciso. La redacción del texto deberá ser impersonal, evitándose siempre la primera persona. Los caracteres en cursiva se utilizarán para los nombres científicos de géneros y especies y para los neologismos que no tengan traducción; las citas textuales, independientemente de la lengua en que estén, irán en letra redonda y entre comillas; el nombre del autor que sigue a un taxón se escribirá también en redonda. Al citar por primera vez una especie en el trabajo, deberá especificarse siempre que sea posible su nombre común. Los topónimos se escribirán bien en su forma original o bien en la lengua en que esté redactado el trabajo, siguiendo el mismo criterio a lo largo de todo el artículo. Los números del uno al nueve se escribirán con letras, a excepción de cuando precedan una unidad de medida. Los números mayores de nueve se escribirán con cifras excepto al empezar una frase. Las fechas se indicarán de la siguiente forma: 28 VI 99 (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 designacio� nes 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 esen� cia del manuscrito (introducción, material, métodos, resultados y discusión). Se evitarán las especulacio� nes 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.

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

Uppsala University. * Los trabajos en prensa sólo se citarán si han sido aceptados para su publicación: Ripoll, M. (in press). The relevance of population studies to conservation biology: a review. 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 & 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, Dis� cusión, Agradecimientos y Referencias) no se nume� rarán. No utilizar más de tres niveles de títulos. Los autores procurarán que sus trabajos originales no excedan las 20 páginas incluidas figuras y tablas. Si en el artículo se describen nuevos taxones, es imprescindible que los tipos estén depositados en alguna institución pública. Se recomienda a los autores la consulta de fascículos recientes de la revista para seguir sus directrices.

Animal Biodiversity and Conservation 37.1 (2014)

Animal Biodiversity and Conservation

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Animal Biodiversity and Conservation is an inter� disciplinary journal published by the Natural Science Museum of Barcelona since 1958. It includes empiri� cal and theoretical research from around the world that examines any aspect of Zoology (Systematics, Taxonomy, Morphology, Biogeography, Ecology, Ethol� ogy, 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.museucienciesjour� nals.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 pro� perty of the journal, which reserves copyright, and no published material may be reproduced or cited without acknowledging the source of information.

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 untrans� latable 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. 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 consecu� tive days, e.g. 28th to 30th). Footnotes should not be used.

Information for authors 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 considera� tion 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. Once an article has been accepted, authors should send a paper copy and an electronic copy of the final version. Please identify software (preferably Word). Proofs sent to the authors for correction should be returned to the Editorial Board within 10 days. Expenses due to any substantial alterations of the proofs will be charged to the authors. The first author will receive electronic version of the article in PDF format. ISSN: 1578–665X eISSN: 2014–928X

Formatting of articles 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. (Title, Name, Abstract, Key words, Resumen, Palabras clave and Address should constitute the first page.) Introduction. Should include the historical back� ground of the subject as well as the aims of the paper. © 2014 Museu de Ciències Naturals de Barcelona

Material and methods. This section should provide relevant information on the species studied, materi� als, 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 re� lated 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 bibliog� raphy of the publications cited in the text. References should be presented as in the following examples (Harvard method): * Journal articles: Conroy, M. J. & Noon, B. R., 1996. Mapping of spe� cies richness for conservation of biological diversity: conceptual and methodological issues. Ecological Applications, 6: 763–773. * Books or other non–periodical publications: Seber, G. A. F., 1982. The estimation of animal abundance. C. Griffin & Company, London. * Contributions or chapters of books: Macdonald, D. W. & Johnson, D. P., 2001. Dispersal in theory and practice: consequences for conserva� tion biology. In: Dispersal: 358–372 (T. J. Clober, E. Danchin, A. A. Dhondt & J. D. Nichols, Eds.). Oxford University Press, Oxford. * Ph. D. Thesis: Merilä, J., 1996. Genetic and quantitative trait variation in natural bird populations. Ph. D. Thesis, Uppsala University. * Works in press should only be cited if they have been accepted for publication: Ripoll, M. (in press). The relevance of population studies to conservation biology: a review. Animal Biodiversity and Conservation. References must be set out in alphabetical and chrono�

logical order for each author, adding the letters a, b, c,... to papers of the same year. Bibliographic citations in the text must appear in the usual way: "...according to Wemmer (1998)...", "...has been defined by Robinson & Redford (1991)...", "...the prospections that have been carried out (Begon et al., 1999)..." Tables. 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, photo� graphs) 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 tridimen� sional. 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 Referen� ces) should not be numbered. Do not use more than three levels of headings. Manuscripts should not exceed 20 pages including figures and tables. If the article describes new taxa, type material must be deposited in a public institution. Authors are advised to consult recent issues of the journal and follow its conventions.

Animal Biodiversity and Conservation 37.1 (2014)

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Animal Biodiversity and Conservation 37.1 (2014)

Arxius de Miscel·lània Zoològica vol. 11 (2013) Museu de Ciències Naturals de Barcelona ISSN: 1698–0476 www.amz.museucienciesjournals.cat

Índex / Índice / Contents Viñolas, A. & Masó, G., 2013. The collection of type specimens of the family Ptinidae (Coleoptera) deposited in the Natural History Museum of Barcelona, Spain. Arxius de Miscel·lània Zoològica, 11: 1–79. Abstract The collection of type specimens of the family Ptinidae (Coleoptera) deposited in the Natural History Museum of Barcelona, Spain.— The collection of the Ptinidae family (Coleoptera) deposited in the Natural History Museum of Barcelona, Spain has been organised, revised and documented. This collection belonged to Francesc Español Coll, a world specialist in this family, and it has become an important reference point for research on that family. The collection is composed of a total of 8,854 specimens, 498 of which are type species. In this paper we provide all the available information related to these type specimens, so for any single taxon, species or subspecies, the following information is given: the original and current taxonomic status, original citation of type materials, exact transcription of original labels, and preservation condition of specimens. Moreover, the differences between original descriptions and labels are discussed. When a taxonomic change has occurred, the references that examine those changes are included at the end of the taxa description. Key words: Español collection, Ptinidae taxonomic revision, Death–watch, Spider beetles. Prieto, M. & Háva, J., 2013. Aportaciones corológicas de la colección del Museu de Ciències Naturals de Barcelona a la fauna iberobalear del género Dermestes Linnaeus, 1758 (Coleoptera, Dermestidae). Arxius de Miscel·lània Zoològica, 11: 80–116. Abstract Chorological contributions from the Natural History Museum of Barcelona collection to the Ibero–Balearic fauna of the genus Dermestes Linnaeus, 1758 (Coleoptera, Dermestidae).— Chorological information of 15 Ibero–Balearic species of the genus Desmestes Linnaeus, 1758 is provided. Data were obtained from the review of 635 specimens housed in the entomological collection of the Natural History Museum of Barcelona. All examined material is listed, and the distributions are discussed taking into account the collection records and bibliographical sources. The distribution areas for most species have been extended. Key words: Coleoptera, Dermestidae, Dermestes, Chorology, Iberian peninsula, Balearic Islands, Museum collection, Natural History Museum of Barcelona. Bros, V., 2013. Contribució a l’estudi dels mol·luscs terrestres (Mollusca, Gastropoda) del Parc de la Serralada Litoral (Barcelona). Arxius de Miscel·lània Zoològica, 11: 117–133. Abstract Contribution to the study of terrestrial molluscs (Mollusca, Gastropoda) Serralada Litoral Park (Barcelona).— The population of molluscs in Serralada Litoral Park (Barcelona, NE Iberian Peninsula) was assessed, contributing to the conservation plan. A wildlife inventory was conducted based on fieldwork and literature. Gastropod communities in various natural environments and conchological species of interest for conservation are also described. Several environmental aspects that determine the distribution and abundance of snails are discussed and management measures to meet their ecological requirements are suggested. Key words: Biodiversity, Snails, Slugs, Gastropods, Conservation priorities, Management.

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Kazerani, F., Khaghaninia, S. & Grichanov, I., 2013. Diversity of the genus Dolichopus Latreille in three different habitats of East Azerbaijan Province, with new records for Iran. Arxius de Miscel·lània Zoològica, 11: 134–152. Abstract Diversity of the genus Dolichopus Latreille in three different habitats of East Azerbaijan Province, with new records for Iran.— The present study is a survey of species diversity of the genus Dolichopus in East Azerbaijan province, Iran. The species were collected using a standard entomological net from three habitats (forest, grassland and wetland areas) in north–west Iran in 2013. Based on the data collected, the forest area with the highest diversity indices (H' = 2.53, 14 species, and H' = 2.19, 10 species, in Chichakli and Keleybar regions respectively) had the most diverse and abundant species, followed by grassland and wetland areas. The dominant species in the study areas were Dolichopus longitarsis and D. simplex. Besides, three species (D. siculus, D. kiritshenkoi and D. plumipes) were recorded from Iran for the first time. Diagnostic characters and geographical distribution of the species occurring in the studied areas with supplementary figures are provided. Key words: Dolichopus, Species diversity, New records, East Azerbaijan Province, Iran. Vila–Farré, M., Álvarez–Presas, M. & Achatz, J. G., 2013. First record of Oligochoerus limnophilus (Acoela, Acoelomorpha) from British waters. Arxius de Miscel·lània Zoològica, 11: 153–157. Abstract First record of Oligochoerus limnophilus (Acoela, Acoelomorpha) from British waters.— We report the occurrence of the acoel Oligochoerus limnophilus (Acoelomorpha) from the British Islands, based on specimens captured in the river Thames (locally known as the river Isis) in Oxford, England, thereby considerably widening the distributional range of the species that had formerly been reported only from continental Europe. We further present live images and CLSM–projections of systematically informative structures that corroborate a close relationship with the genus Convoluta Ørsted, 1843. Key words: Acoela, Oligochoerus, Limnic, Thames. Jawad, L. A., 2013. Confirmed record of Monodactylus argenteus (Linnaeus, 1758) (Family Monodactylidae) from Jubail, Saudi Arabia, Arabian Gulf. Arxius de Miscel·lània Zoològica, 11: 158–162. Abstract Confirmed record of Monodactylus argenteus (Family Monodactylidae) from Jubail, Saudi Arabia, Arabian Gulf.— The first record of M. argenteus from the Arabian Gulf coasts of Saudi Arabia is confirmed, based on a sample measuring 158 mm in SL. Morphometric and meristic data are provided for this specimen. Key words: Arabian Gulf, Saudi Arabia, New record, Monodactylus argenteus, Range distribution. Ripoll Rodríguez, J., De las Heras Carmona, M., Moreno Benítez, J. M., Prunier, F. & Solano, F., 2013. Grandes branquiópodos (Crustacea, Branchiopoda, Anostraca, Notostraca) en la provincia de Málaga, España (año hidrológico 2012/2013). Arxius de Miscel·Lània Zoològica, 11: 163–177. Abstract Large branchiopods (Crustacea, Branchiopoda, Anostraca, Notostraca) from Málaga province, Spain (2012/2013 hydrological year).— This paper presents the occurrence of the large branchiopods detected during a survey carried out in the province of Malaga (Andalusia, southern Spain). Five species (Branchipus cortesi, Chirocephalus diaphanus, Streptocephalus torvicornis, Triops mauritanicus aggr. and Phallocryptus spinosa) were recorded at 90 sampled wetlands. Key words: Crustacea, Large branchiopods, Branchiopoda, Temporary pools, Malaga, Spain. Bros, V., 2013. Data paper: Land snails and slugs at the Natural Park of the Serralada Litoral (Barcelona, Spain). Arxius de Miscel·lània Zoològica, 11: 178–180. Abstract Land snails and slugs at the Natural Park of the Serralada Litoral (Barcelona, Spain).— The population of molluscs at the Natural Park of the Serralada Litoral (Barcelona, Spain, NE Iberian peninsula) was assessed, contributing to the conservation plan. A wildlife inventory was conducted based on fieldwork. Key words: Occurrence dataset, Biodiversity, Terrestrial gastropods, Catalonia, Protected areas. ISSN: 1578–665 X eISSN: 2014–928 X

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Belda, A., Arques, J., Peiró, V., Martínez–Pérez, J. E. & Zaragozí, B., 2013. Abundancia y distribución de la liebre ibérica (Lepus granatensis Rosenhauer, 1856) en el Parque Natural de la Sierra de Mariola (Alicante– Valencia). Arxius de Miscel·lània Zoològica, 11: 181–195. Abstract Abundance and distribution of the Iberian hare (Lepus granatensis Rosenhauer, 1856) in the Sierra de Mariola Natural Park (Alicante–Valencia).— The Iberian hare (Lepus granatensis Rosenhauer, 1856) is a species of great value in Spanish Mediterranean ecosystems for several reasons, such as its interest to hunters, its contribution to soil fertility and plant diversity, and its role as prey. However, factors such as fragmentation, degradation and loss of habitat and diseases, predation and high pressure hunting are having a detrimental effect on the conservation of the species. It is therefore of interest to determine the abundance and distribution of the Iberian hare in areas of the peninsula where there are insufficient data to establish guidelines for conservation and sustainable management of hare populations in the peninsula. Our goal was to assess the abundance and distribution of the Iberian hare in the most widely used areas of a mountainous Mediterranean landscape in the Iberian peninsular (Mariola Mountain Park, located between the provinces of Alicante and Valencia). Data obtained from studies conducted in transects from 2008 to 2010 showed that intra–annual abundance was highest in spring (KIA half of 0.26 hares/km) and lowest in winter (average 0.075 KIA hares/km). As to their preferences in relation to land use, abundance was highest in the matrix of dry groves (KIA half of 0.33 hares/km) and irrigated groves (average 0.2 KIA hares/km). The matrix of natural vegetation and agricultural abandonment had low numbers of hares, with values of 0.083 and 0.033 hares/km, respectively. Key words: Abundance, KAI (Kilometric Abundance Index), Lepus granatensis, Landscape matrix, Sierra de Mariola. Chakrabarty, M. & Homechaudhuri, S., 2013. Fish guild structure along a longitudinally–determined ecological zonation of Teesta, an eastern Himalayan river in West Bengal, India. Arxius de Miscel·lània Zoològica, 11: 196–213. Abstract Fish guild structure along a longitudinally–determined ecological zonation of Teesta, an eastern Himalayan river in West Bengal, India.— The Eastern Himalaya Biodiversity Hotspot contains exceptional freshwater biodiversity and ecosystems that are of vital importance to local and regional livelihoods, but these are under threat from the developmental and anthropogenic pressures arising from the 62 million people living in the area. Therefore, monitoring the riverine health and considering future conservation approach, the study of fish biodiversity plays a significant role in this region. The River Teesta in the Brahmaputra basin in India forms one of the major rivers in the Eastern Himalayas. In the present investigation, we studied ecological fish guilds as they can enhance the usefulness of fish zonation concepts and serve as tools to assess and manage the ecological integrity of large rivers. We classified fish species according to their water flow preference and spawning substrate preference. Ten spawning habitats were identified, occurring in three water flow guilds. The most widely preferred habitat in upstream zones was lithophils while in lower stretches it was lithopleagophils. On applying predictions of the River Continuum Concept, our results indicated the presence of a zonation pattern based on fish species assemblage and their ecological attributes along the longitudinal stretch of the Teesta River in west Bengal. Along the longitudinal stretch of the river, species richness increased downstream, with maximum richness in the mid–reaches. However, species richness decreased further downstream. The number of ecological guilds also increased downstream, and there were clear shifts in the structure of the guilds. Key words: Eastern Himalayas, Teesta, Lotic water, Biodiversity, Flow–preference guild, Altitudinal gradient.

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Les cites o els abstracts dels articles d’Animal Biodiversity and Conservation es resenyen a / Las citas o los abstracts de los artículos de Animal Biodiversity and Conservation se mencionan en / Animal Biodiversity and Conservation is cited or abstracted in: Abstracts of Entomology, Agrindex, Animal Behaviour Abstracts, Anthropos, Aquatic Sciences and Fisheries Abstracts, Behavioural Biology Abstracts, Biological Abstracts, Biological and Agricultural Abstracts, BIOSIS Previews, CiteFactor, Current Primate References, Current Contents/Agriculture, Biology & Environmental Sciences, DIALNET, DOAJ, DULCINEA, e–revist@s, Ecological Abstracts, Ecology Abstracts, Entomology Abstracts, Environmental Abstracts, Environmental Periodical Bibliography, FECYT, Genetic Abstracts, Geographical Abstracts, Índice Español de Ciencia y Tecnología, International Abstracts of Biological Sciences, International Bibliography of Periodical Literature, International Developmental Abstracts, Latindex, Marine Sciences Contents Tables, Oceanic Abstracts, RACO, Recent Ornithological Literature, Referatirnyi Zhurnal, Science Abstracts, Science Citation Index Expanded, Scientific Commons, SCImago, SCOPUS, Serials Directory, SHERPA/ RoMEO, Ulrich’s International Periodical Directory, Zoological Records.

Consorci format per / Consorcio formado por / Consortium formed by:

Índex / Índice / Contents Animal Biodiversity and Conservation 37.1 (2014) ISSN 1578–665 X eISSN 2014–928 X

1–12 M. Cortés–Marcial, Y. M. Martínez Ayón & M. Briones–Salas Diversity of large and medium mammals in Juchitan, Isthmus of Tehuantepec, Oaxaca, Mexico

59–67 F. Lisón, Á. Haz & J. F. Calvo Preferencia de hábitat del murciélago hortelano meridional Eptesicus isabellinus (Temminck, 1840) en ambientes mediterráneos semiáridos

13–21 A. Farashi, M. Kaboli, H. Reza Rezaei, M. Reza Naghavi & H. Rahimian Plankton composition and environmental parameters in the habitat of the Iranian cave barb (Iranocypris typhlops) in Iran

69–76 M. S. Khan, N. K. Dimri, A. Nawab, O. Ilyas & P. Gautam Habitat use pattern and conservation status of smooth–coated otters Lutrogale perspicillata in the Upper Ganges Basin, India

23–33 P. Sarmento, J. Cruz, C. Eira & C. Fonseca A spatially explicit approach for estimating space use and density of common genets

77–88 R.–R. Ray, H. Seibold & M. Heurich Invertebrates outcompete vertebrate facultative scavengers in simulated lynx kills in the Bavarian Forest National Park, Germany

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