ABC vol 38 2 (2015) by Museu Ciències Naturals de Barcelona MCNB

Editor en cap / Editor responsable / Editor in Chief

2Joan 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 y Ecología Evolutiva – SEEEE) 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 Marco Apollonio Univ di Sassari, Sassari, Italy Xavier Bellés Inst. de Biología Evolutiva UPF–CSIC, Barcelona, Spain Salvador Carranza Inst. Biologia Evolutiva UPF–CSIC, Barcelona, Spain Luís Mª Carrascal Museo Nacional de Ciencias Naturales–CSIC, Madrid, Spain 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 Arnaud Faille Museum National histoire naturelle, Paris, France Jordi Figuerola Estación Biológica de Doñana–CSIC, Sevilla, Spain Gonzalo Giribet Museum of Comparative Zoology, Harvard Univ., Cambridge, USA Susana González Univ. de la República–UdelaR, Montivideo, Uruguay Sidney F. Gouveia Univ. Federal de Sergipe, Sergipe, Brasil Gary D. Grossman Univ. of Georgia, Athens, USA Jacob Höglund Uppsala Univ., Uppsala, Sweden Joaquín Hortal Museo Nacional de Ciencias Naturales-CSIC, Madrid, Spain Damià Jaume IMEDEA–CSIC, Univ. de les Illes Balears, Spain Jennifer A. Leonard Estación Biológica de Doñana-CSIC, Sevilla, 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 Estación Biológica de Doñana–CSIC, Sevilla, Spain 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 Diego San Mauro Univ. Complutense de Madrid, Madrid, Spain Constantí Stefanescu Museu de Ciències Naturals de Granollers, Granollers, Spain José L. Tellería Univ. Complutense de Madrid, Madrid, Spain Francesc Uribe Museu de Ciències Naturals de Barcelona, Barcelona, Spain José Ramón Verdú CIBIO, Univ de Alicante, Alicante, 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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Assessorament lingüístic / Asesoramiento lingüístico / Linguistic advisers Carolyn Newey Pilar Nuñez Animal Biodiversity and Conservation 38.2, 2015 © 2015 Museu de Ciències Naturals de Barcelona, Consorci format per l'Ajuntament de Barcelona i la Generalitat de Catalunya Autoedició: Montserrat Ferrer Fotomecànica i impressió: Inspyrame Printing 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 y Ecología Evolutiva i la Sociedad Española de Biología Evolutiva The journal is freely available online at: www.abc.museucienciesjournals.cat Dibuix de la coberta: Canis lupus, llop, lobo, wolf (Jordi Domènech)

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The lion in Ghana: its historical and current status F. M. Angelici, A. Mahama & L. Rossi

Angelici, F. M., Mahama, A. & Rossi, L., 2015. The lion in Ghana: its historical and current status. Animal Biodiversity and Conservation, 38.2: 151–162. Abstract The lion in Ghana: its historical and current status.— Historically, the lion (Panthera leo) population in Ghana has been little studied and its status is poorly documented. Currently, after recent unsuccessful attempts to find signs of the presence of the species, many authors believe that the Ghanaian lion population is most likely extinct. In an attempt to gather more data, since 2005 we have been carrying out lion surveys in the most important parks and other protected areas of Ghana, mainly focusing on Mole National Park (MNP). We have also been extensively reviewing the literature in an attempt to reconstruct the history of the presence of the lion in the country. Although our research has not provided unequivocal evidence of the presence of the lion, we have collected circumstantial evidence that suggests that a small lion population might still be present in MNP and its surrounding areas. Key words: Lion, Panthera leo, Ghana, Status, Mole National Park Resumen El león en Ghana: su situación pasada y presente.— Históricamente, la población de león (Panthera leo) en Ghana ha sido poco estudiada y su situación actual está poco documentada. Tras los últimos intentos infructuosos de encontrar indicios de la presencia de la especie, son numerosos los autores que opinan que la población de león en Ghana está prácticamente extinguida. En un intento por recabar más datos, desde 2005 hemos venido realizando estudios sobre esta especie en los parques más importantes de Ghana y otras zonas protegidas del país, que se han centrado principalmente en el Parque Nacional de Mole (MNP). Asimismo, hemos examinado los datos publicados con el fin de reconstruir la historia de la presencia del león en el país. Si bien nuestra investigación no aportó datos inequívocos, se recabó información circunstancial que sugiere que aún podría existir una pequeña población de león en el MNP y sus zonas circundantes. Palabras clave: León, Panthera leo, Ghana, Situación, Parque Nacional de Mole Received: 23 XII 14; Conditional acceptance:17 II 15; Final acceptance: 23 IV 15 Francesco M. Angelici, FIZV, Via Cleonia 30, Scala C, I–00152 Roma (RM), Italy.– Ali Mahama, �� Wildlife Divi� sion, Staff Headquarters, Mole National Park, Larabanga, Northern Territory, Ghana.– Lorenzo Rossi, Via San Cristoforo 196, I–47522 Cesena (FC), Italy. Corresponding author: F. M. Angelici. E–mail: francescomariaangelici@gmail.com

ISSN: 1578–665 X eISSN: 2014–928 X

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Introduction The first scientific studies of African lion (Panthera leo) populations date back to the 1950s and primarily concern East Africa. Similar studies in West Africa are almost absent. Generally speaking, it is difficult to count individual lions (Myers, 1975). They can cover enormous distances and occupy areas where they have gone unreported for years, sometimes far from their normal range of distribution (Smithers, 1983). Much potential information from local people is likely to remain unknown to science or underestimated, rendering it useless for conservation purposes (Black et al., 2013). As a consequence, detecting the presence of a scarce lion population in large areas is extremely difficult and requires close collaboration with park staff and local people in order to gather as much information as possible. According to Henschel et al. (2014), the West African lion, P. leo senegalensis (i.e., from Senegal to Nigeria), is currently in serious danger of extinction (fig. 1), with about 400 lions in the whole of West Africa, probably representing fewer than 250 mature individuals. In the last twenty years, West African lion populations in Ghana were estimated at a few dozen individuals (Wilson, 1993; Chardonnet, 2002; Bauer & Van Der Merwe, 2004) and the species has recently been considered 'functionally extinct, if not completely eradicated' (Henschel et al., 2010; Burton et al., 2011a). The lion is 'considered absent in Ghana' (Henschel et al., 2014) (fig. 1). In this paper we aim to reconstruct the historical status of the lion in Ghana by reviewing the available literature, unpublished data, and material collected during field expeditions in the Mole National Park (MNP) and other areas between 2005 and 2014. Materials and methods Historical data up to 2010 We reviewed the literature on lion distribution and status in Ghana, including unpublished official reports, and compared all the available data to identify new information and any possible inconsistencies in lion population estimates over the years. We obtained data from scientific articles and books, park documents, unpublished reports provided by the Wildlife Division (Forestry Commission) of the Ghanaian Ministry of Lands and Natural Resources, IUCN and FAO publications and first preliminary expeditions led by the main author since 2005 to 2009 (Angelici, 2006; Angelici & Petrozzi, 2010). Table 1 lists all the unpublished sources we cited. Project 'The Pride of Ghana' (Mole National Park, 2011–2014) The project entitled 'The Pride of Ghana: Local Development and Assistance Toward for the Sustainable Management of the Mole National Park and its Fringe Communities' was officially launched in January 2011. The number of consecutive days spent in the field to date was 36 days in 2011, 22 days in 2012, 52 days in

2013 (and 23 days in 2014, in addition to the constant structured, formal collaboration over the course of the year with the Wildlife Division staff operating in MNP. Protocols used during field activities and habitat suitability model From April 12 to August 8, 2011, 20 digital camera traps (model LTL Welltar 8210A) were positioned. They were placed in different areas of the park, with a total of 2,474 trap days and 24 camera stations, with 1,745 trap days in high–suitability areas (228 on average) and 501 in low–suitability areas (for a definition of suitability levels, see below). A regular transect was not used to position the traps. In 2012, there were only 163 trap days, using 12 camera traps in 14 different positions. From March 2013 to the present (camera trapping is ongoing), up to 22 camera traps have been used over the course of the year, with periods of temporary suspension, especially during the rainy season, for obvious reasons of accessibility. We have collected thousands of photographs, which are still under analysis (fig. 2 for all camera positions; in some places, there is more than one trap position). The choice of where to position the camera traps was based on a habitat suitability model (fig. 3). This model was developed on the basis of lion sightings recorded by park staff between 1968 and 2009 by fitting logistic regression models on the habitat features of lion sightings in MNP. The results of the analysis (significant variables) were used to implement the selected predictive variables in GIS software (ArcGis v. 9.0) to produce a predictive species suitability map, i.e., to map the areas with the greatest probability of lion occurrence in the Park. The maps were built according to three intervals of the probability of occurrence: low–suitability habitat (values from 0 to 0.33); medium–suitability habitat (from 0.34 to 0.66); and high–suitability habitat (from 0.67 to 1). The reliability of the potential distribution model was assessed by AUC criteria using a Jackknife procedure. The main positive feature of AUC is its single threshold–independent measure for model performance. An AUC value can be interpreted as the probability that a presence site, randomly chosen from the dataset, will have a higher predicted value than an absence site. The overall model fitting was good: AUC=0.886. Protocol and statistical analyses We used data recorded over a 41–year time span by park patrols. Each data entry recorded the confirmation of lion presence, including the geographic coordinates of the site. These data were entered into DIVA–GIS software. Around each record of lion occurrence (n = 100), we extracted a circular 100 m radial buffer, and within this buffer we recorded several independent variables: (i) linear distance from the closest road/path (hereby DCR); (ii) rainfall (mm, per year) (RFL); (iii) linear distance from the closest pond/water body (DWB); (iv) linear distance from the closest ungulate prey (Kobus ellipsiprymnus,

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A N W

S Considered absent Potentially present Confirmed present Ghana border 300 km

B E

D F

Fig. 1. Map of West Africa including all protected areas where lion occurrence has been documented according to Henschel et al. (2014). Protected areas in Ghana: A. Gbele Resource Reserve; B. Mole National Park; C. Bui National Park; D. Digya National Park; E. Kogyae Strict Nature Reserve; F. Kalapka Resource Reserve. Fig. 1. Mapa del África occidental con todas las zonas protegidas en las que se ha documentado la presencia del león según Henschel et al. (2014). Zonas protegidas en Ghana: A. Reserva de Recursos de Gbele; B. Parque Nacional de Mole; C. Parque Nacional de Bui; D. Parque Nacional de Digya; E. Reserva Natural de Kogyae Strict; F. Reserva de Recursos de Kalapka.

Hippotragus equinus, Kobus kob, Syncerus caffer, Alcelaphus buselaphus) (DPR); (v) land use (LNU); and (vi) elevation (m a.s.l.) (ELE). In addition, we recorded the same independent variables for 100 random points (also with a 100 m radius) within MNP. A logistic regression modelling approach was applied to lion presence/absence (Hosmer & Lemeshow, 1989) using a backward stepwise model (Luiselli, 2006) and the Von Bertalannfy growth function (Von Bertalannfy, 1934, 1938, 1951, 1964). In these models, lion presence/absence was the dependent variable, and the six above–mentioned variables were the covariates. These techniques are powerful analytical tools that can analyse the effects of one or several independent variables, both discrete and continuous, on a dichotomous dependent variable (Hosmer & Lemeshow, 1989; Teixeira et al., 2001). In addition, logistic regression models rely on fewer statistical assumptions than their alternatives and generally produce robust results (Teixeira et al., 2001). Independence was assessed when r2 < 0.58 (Hosmer & Lemeshow, 1989; Arntzen & Alexandrino, 2004). To determine whether the probability of lion presence in relation to the studied covariates was best described by backward stepwise logistic regression or by the Von Bertalannfy growth function, we relied on a model–selection approach based on the Akaike Information criterion (AIC) (Burnham & Anderson,

2002) according to the formula: AIC = −2 log Likelihood + 2K where n depicts effective sample size, and K is the number of parameters (= number of variables + 1 to include the intercept (Sugiura, 1978). The relative performance of alternative models was measured using the delta AIC: ∆AIC = AICi – min AIC where AICi is the AIC value for model i, and min AIC is the AIC value of the best fitting model. Hence, the differences between the AIC scores of the various models (∆AIC) provides a measure of the relative reliability of the competing models. The advantage of this approach is that it allows the various competing models to be ranked according to their relative likelihood and is not dependent on a threshold value (α–level, Vapnik, 2000). The AIC penalizes the addition of parameters, and thus selects a model using a minimum number of parameters according to the principle of simplicity and parsimony (Akaike, 1973); therefore, the models with the lowest ∆AIC were selected. Starting in 2012, we conducted night sessions (from about 9 pm to 1 am) along some random paths, listening for any possible lion roars. We conducted two night sessions in May 2012, two in March–April 2013, and two in February 2014.

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Table 1. A list of the unpublished sources from which indirect information, and/or references were obtained. Tabla 1. Lista de las fuentes sin publicar de las que se obtuvieron información indirecta y referencias.

Data

Area

Origin

Year

MNP database MNP

Wildlife Division (Forestry Commission) Ghana, compiled by MNP staff

1968–2004

Report by P. J. Pegg on wildlife management

FAO library, Rome

1969

Reports by R. Jamieson Ghana and MNP

Wildlife Division (Forestry Commission), Ghana

1970–72

Reports by Aberdeen MNP University Ghana expeditions to MNP

Wildlife Division (Forestry Commission), Ghana

1974–1977

Wilson (1993)

IUCN library, Gland

1993

MNP

Moreover, as of March 2013, we started to perform linear transects at night (8.30 pm – 2 am) by car along the trails in the park using directional headlights at a constant speed of about 10 km/h. In particular, we travelled along six transects in 2013, and three in 2014. Each transect was 30–35 km long.

Results Historical data up to 2010 Our findings from a review of the historical literature were scattered, sporadic and often inaccurate, testifying

N W

E S

Camera Camera Camera Camera

traps traps traps traps

2011 2012 2013 2014

Fig. 2. Localization of camera–trapping sessions in MNP carried out from 2011 to 2014. Fig. 2. Ubicación de las sesiones de trampeo con cámara en el MNP realizadas entre 2011 y 2014.

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N E

W S

Park border High suitability Medium suitability Low suitability 0

120

160 km

Fig. 3. Lion (Panthera leo) habitat suitability model in MNP and surrounding areas (for methodology, see text). Fig. 3. Mapa de idoneidad del hábitat del león (Panthera leo) en el MNP y las zonas circundantes (véase el texto para obtener más información sobre la metodología).

to the lack of study of this species in Ghana. This is possibly because lions were never particularly abundant or widespread in the region. Grubb et al. (1998) meticulously collected several old records of lions from the late 19th century but even in the map presented in their paper, it is clear that the last strongholds of the species were (supposedly) protected areas, i.e., Bui National Park, Digya National Park, GRR, and MNP, apart from other occasional reports scattered throughout the North and East of the country on the Togo border. Lions were even reportedly seen in Kogyae Strict Nature Reserve and Kalapka Resource Reserve (fig. 1), albeit sporadically, at least until the 1990s (Grubb et al., 1998). For an accurate reference selection, see Grubb et al. (1998). Unfortunately, there are no data on the size of the lion populations in any area, nor are there any even preliminary data on their ecology, with the exception of MNP, as we will see later. In Cansdale’s (1948) provisional checklist of the Gold Coast, he only mentions the lion as being present in the open country in the areas of Togoland, Afram Plains, north–west Ashanti and the Northern Territories. However, as it is a checklist, he does not insert any other data, particularly regarding species abundance or frequency. Only two reports were found that relate to this period, and both were very general and based on rough estimates, not on specific work carried out in the field. Mention should be made of a male lion from Tamale, Doka woodland (about 80 km from MNP),

whose skin is stored in the Natural History Museum in London, no. 394, dated June 1943 (Rosevear, 1974). To date, Wilson's report (1993), based on three months of fieldwork, is the only document which includes data on the distribution and ecology of the species. Wilson confirmed that the lion roar was heard several times during the survey, in particular near the headquarters of the Wildlife Division and along the Lovi River, near Lovi camp (fig. 4). Wilson (1993) claims to have confirmed lion presence in MNP in at least three different locations. In January 1993, a photo was also taken of a lioness in lactation (fig. 5) by John Grainger, near Gbanwele camp (fig. 4). Various lion droppings were also collected in at least five different locations over the 3–month study, specifically in Lovi, Brugbani, Gbanwele, Samole, and Nyanga (fig. 4) (Wilson, 1993). Rangers also collected the skulls of some probable lion kills. Wilson (1993) also states that 'while the lion population in MNP is certainly not high there must be at least sufficient numbers to maintain a breeding population'. Furthermore, the same author reported that in December 1992 the rangers saw prides of up to eight lions all together, including three cubs, in particular at Lovi and Nyanga. He also concluded that lions were more easily encountered at Lovi and Nyanga, along the Lovi River, extending east and south to Brugbani and Samole. While lions could also be found going towards Gbanwele and Konkori, as well as near Kwomwoghlugu. Moreover, they often

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N W

Zanwera

Kulpawn Camp

Sagiya

Belepong Kporio

Holomuni Ducie

Belebile Sabile Camp

Chasla Chasia Camp

Sogsima Camp

Dagbori Camp

Kong

Jang Soma

Jang Camp Nikori Camp

Jelinkan

Lovi River Lovi Camp Brugbani Camp

Jandra Camp

Grupe Sayire

Grubagu

Jinfrono

Konkori Camp

Seriseeni Camp

Bawena Camp

Nyanga Camp

Yazori Camp

Bawena Kpulumbo

Kwomwoghlugu Camp

Yazori Kaden

Muruga Camp

Muruga

Mognori Camp Park H.Q. Kamanto Kabampe Camp Degbere Mognori Grupe Semole Camp Camp Camp Kuboma Camp Camp Kananto Palma Larabanga Camp Kobampe

Wawato

Grubagu Camp

Esalakawu Camp Ducie Camp Gbanwele Camp

Mole River Daborin

Gbantala Camp

10 km

Damongo

Lion sighting/footprints 2009 Lion sighting 2013 Lion sighting 2014 Roars heard 2011 Roars heard 2012 Roars heard 2013 Roars heard 2014 Buffalo carcassa 2013 Camp Park H. Q. Village

Fig. 4. Detailed map of MNP showing data on lions collected from 2005 to 2014. Fig. 4. Mapa detallado del MNP en el que se muestran los datos, obtenidos entre 2005 y 2014, relativos a la presencia de leones.

ventured outside the boundaries of the park, according to rangers' records, particularly those from the camps of Kananto, Jang, Gbanwele and Gbantala (for all localities see figure 4). In Chardonnet's (2002) account of African lions, he estimates that 15 (12–18) lions are present in MNP in the table, while in the text he refers to a range from 15 to 50 lions in MNP according to the estimates of various specialists. Bauer (2003) and Bauer & Van Der Merwe (2004) report on an indirect estimate made by the Ghana Wildlife Society of 20 (12–28) lions in MNP and 10 (6–14) in GRR. In 2002, a lioness was killed by poachers in MNP, and in 2004 a male was shot (fig. 6) very close to

the village of Larabanga (fig. 4). A few days earlier, the same lion had killed several heads of cattle and had also had an aggressive encounter with another man (see Angelici & Petrozzi, 2010). We collected the results of questionnaires administered in both MNP (n = 47) and GRR (n = 6), and only 21.3% of the respondents in MNP reported seeing lions between 2000 and 2009. For further details regarding the methodology applied, see Angelici & Petrozzi (2010). According to MNP records for the 1968–2009 period, the maximum number of sightings of a single lion (of any age) observed on a single occasion was 21, in 1969. No sightings were reported between 1998 and 2008. In 2009, two individuals were spotted

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Fig. 5. Lioness photographed in January 1993 in MNP. Fig. 5. Leona fotografiada en enero de 1993 en el MNP.

(Angelici & Petrozzi, 2010). Our results date back to the end of 2005 when our fieldwork was undertaken, and are intermittent up to 2010. The project was officially launched in 2011. At the same time, the results of research carried out in MNP by C. Burton and collaborators (2006â&#x20AC;&#x201C;2008), i.e., Henschel et al.

(2010) and Burton et al. (2011a, 2011b), begin to emerge. During their research, according to their protocols, the authors did not obtain any findings regarding the lion, whereas they collected a lot of data relating to other mammals. Their conclusion was that in MNP the lion had likely been functionally, if

Fig. 6. Lion shot near Larabanga in August 2004. Fig. 6. LeĂłn abatido cerca de Larabanga en agosto de 2004.

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Table 2. Summary of the results obtained in the literature and in the present work (updated to December 2014) regarding lions in Ghana (for all localities see figures 1 and 4). Tabla 2. Resumen de los resultados obtenidos en las publicaciones y en el presente estudio (actualizados en diciembre en 2014) con respecto a los leones en Ghana (consúltense las figuras 1 y 4 para ver todas las localidades).

Data

Area

Reference

Several anecdotal records of single lion sightings (late 19th C. until the 1960s) General data of occurrence

Bui National Park, Digya Grubb et al. (1998) National Park, GRR, MNP, other occasional reports scattered throughout the North and East Kogyae Strict Nature Reserve and Kalapka Resource Reserve Togoland, Afram Plains, Cansdale (1948) north–west Ashanti and the

Year 1893–1960, some undated

Until 1948

Northern Territories Skin stored in the Natural Tamale, Doka woodland Rosevear (1974) June 1943 History Museum, London, (about 80 Km from MNP) no. 394 (male) Many records (sightings, MNP: Lovi, Brugbani, Wilson (1993) 1992–1993 roars, droppings, Gbanwele, Samole, Nyanga, prey remains, etc.) Konkori, Kwomwoghlugu. Frequently outside the park, e.g. Kananto, Jang, Gbanwele, Gbantala (fig. 5) 15 (12–18) or 15–50 lions MNP Chardonnet (2002) 2002 (estimate) 20 (12–28) lions in MNP, MNP and GRR Bauer & Van 2004 and 10 (6–14) in GRR Der Merwe (2004) Lioness shot by poachers MNP, unknown locality Angelici & Petrozzi (2010) 2002 Lion shot by MNP, Larabanga Angelici & Petrozzi (2010) 2004 Larabanga shepherds surroundings (fig. 6) Several direct observations MNP Angelici & Petrozzi (2010) 1968–2009 made by MNP staff between 1968–2009 (see Discussion) Questionnaires administered MNP, GRR Angelici & Petrozzi (2010) 2000–2009 in both MNP (n = 47) and GRR (n = 6), 21.3% in MNP reported seeing lions, none in GRR Sighting of one lion Digya National Park Henschel et al. (2010) October 2008 Case of human–lion Kalapka Resource Reserve Henschel et al. (2010) February 2009 interaction Sighting of a couple of lions, MNP, Lovi (fig. 7) Angelici & Petrozzi (2010) May 2009 footprints, prey remains, by a park manager

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Table 2. (Cont.) Data

Area

Reference

Year

Roars heard Roars heard Roars heard Sighting of a male lion by a staff guard Possible lion predation of an adult buffalo Roars heard A couple of lions sighted by poachers, roars heard

MNP: staff lodge compound Original data April 2011 MNP: along the road Original data May 2012 Mognori–Lovi MNP: three localities (fig. 4) Original data April 2013 MNP: close to the Original data August 2013 staff quarters MNP: near Brugbani Original data October 2013 camp (fig. 8) MNP: between Lovi and Original data February 2014 Kwomwoghlugu MNP: near Gbantala camp Original data August 2014

not fully, extirpated. The same conclusion was reached regarding lion presence in GRR. Henschel et al. (2010), however, did report some anecdotal local sightings that they considered plausible, although further confirmation and investigation would be needed. One such incidence occurred in Digya National Park where a lone lion was sighted in October 2008 after years with no sightings. An extremely unusual case regarding a human–lion interaction was reported in February 2009 in Kalapka Resource Reserve in south–eastern Ghana near Togo. If confirmed, this would support what was stated in the introduction of this article, that a large predator may unexpectedly 'reappear' in areas where it has been declared extinct despite incomplete knowledge of its status, causing unforeseen problems (Smithers, 1983; Black et al., 2013), as recently happened in Gabon (Anonymous, 2015). In May 2009, park staff in MNP, including an executive manager (Oliver K. Chelewura), clearly sighted a two lions, a male and a female. This event was reported in a paper the following year (Angelici & Petrozzi, 2010). Lion footprints were also observed (fig. 7) at the same sight along with the skull of a hartebeest that the lions abandoned when they saw humans. Unfortunately, the picture contains no elements to estimate footprint size but all the footprints were found in the general area of the sighting and appear convincing. Status 2011–2014

Fig. 7. A lion footprint found immediately after the sighting of two lions in May 2009 in MNP.

Camera trap sessions We have not collected any photos of lions to date, but more than 20 species of mammals have been photographed, as evidenced in more than 6,000 selected pictures and about one hundred short films.

Fig. 7. Un huella de león encontrada inmediata� mente después del avistamiento de dos leones en mayo de 2009 en el MNP.

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Fig. 8. Buffalo carcass shot by poachers in MNP (October 2013) with clear signs of probable attacks by lions. Note, in particular the holes in the hindquarters which are infected and invested with larvae, and the deep and recent long lateral scratches. Fig. 8. Cadáver de búfalo abatido por los cazadores furtivos en el MNP (octubre de 2013) con signos evidentes de un probable ataque de leones. Nótese, en particular las heridas en los cuartos traseros, que están infectados e infestados por larvas, y los largos arañazos laterales, profundos y recientes.

Reports of direct lion sightings At 8:07 pm on the evening of April 13 2011, we distinctly heard a roar from the staff lodge compound. During the night transect sessions in May 2012, a roar was heard twice on one occasion and in April 2013, repeated roars were heard three times in two different areas within the park (for all roars heard, fig. 4). All these roars were heard at night by one of the authors (F. M. A.) along with some Ph D and MSc students and guards Eric Bani and David W. Kabuiri, two of the most experienced members of the staff. We have yet to encounter any lions on any night transects: However, we did encounter and recognized 16 mammal species. An event of greater note was the sighting by a guard (D. W. Kabuiri) of an adult male lion very close to the staff quarters and to the entrance gate to the park in August 2013 (see fig. 4). The following year, in February 2014, we heard the night roarings of male lions in the south–central region of the park on two occasions (fig. 4). In October 2013, we collected some data regarding the possible lion predation of an adult buffalo. The

buffalo was shot by poachers and retrieved by rangers and was severely wounded and limping (see the deep wounds and large, long scratches on the body as well as the holes on its buttocks made by claws that are severely infected and full of blowfly maggots, fig. 8). In August 2014, we obtained information from a poacher who was questioned by rangers. The poacher had seen two adult lions near Gbantala camp (fig. 4), where several roars had also been heard. Further investigation is currently underway in the region and camera traps are being placed throughout the area. All data are summarized in table 2 Discussion Although compelling evidence has yet to be gathered (i.e., clear videos or photographs), the presence of a few individual lions in MNP should not be ruled out a priori. Although the most recent empirical evidence (i.e., a male that was shot) dates back to 2004, the adult lions observed in 2002 and 2004 could have reproduced before being killed. Several previous state-

Animal Biodiversity and Conservation 38.2 (2015)

ments have proven unreliable regarding the extinction of large cats (Black et al., 2013) and the evidence we have collected also challenges such a conclusion. The buffalo shot in October 2013, in addition to the healed injuries probably caused by other buffalos, seems to bear the typical signs of a lion attack. Most of the remaining evidence that has turned up during our work has been from eyewitness. However, on two occasions (May 2009 and August 2013) this evidence was based on reports by qualified and reliable MNP staff. When management resources are scarce, reports of a supposedly extinct species can cause controversy (Roberts et al., 2009), as was the case for lion sightings in the MNP. Monitoring program of the park has been criticized by Burton (2012) as not always reliable. Although often overlooked, the role of parataxonomists (local assistants trained by professional biologists, see Janzen, 1991, 2004) can play a critical role in conservation (Basset et al., 2004), and information provided by trained assistants and the local people can be as accurate as those of field biologists (Danielsen et al., 2014). The May 2009 sighting is supported by additional evidence: footprints that can be attributed to a large cat on the site and the remains of a hartebeest, apparently killed by a large predator. Considering the eyewitnesses’ statements, the footprints and the typology of the prey, the lion is the most likely candidate. Evidence of lion sightings and roars dated August 2014 near Bantala camp is of particular importance as it was provided by poachers, who generally understate the occurrence of wildlife, in particular by not supplying information about lions, for fear of retaliation. The opinion of Henschel et al. (2010) and Burton et al. (2011a), which in our view is perhaps too hasty, was later accepted by many authors and authorities and reiterated in other articles (e.g., Burton et al., 2011b; Henschel et al., 2014). Nevertheless, we believe it is only right to continue to seek out objective data that attests to the continued persistence of a few lions in MNP and the immediately surrounding areas. As pointed out by other authors (e.g., Black & Copsey, 2014), we believe that from the point of view of the governments and park leaders, a mix of incomplete knowledge about the presence of a species is better than assuming its extinction. The possible implications of a wrongful assumption of the extinction of the lion would be so important that in the light of indirect evidence we have collected, we believe it is wiser to apply the Precautionary Principle (Foster et al., 2000) and assume the survival of the species until more scientific data tell us the full picture. If there is a chance that some lions are still present in the MNP, it is essential to avoid Romeo’s Error (Collar, 1998) for which a species is thought to be extinct in an area without assessing all available information. Finally, it should be kept in mind that considering the lion as extinct quickly leads to safety rules concerning a direct encounter with the big cat and all that this implies: danger to livestock but also danger to man himself, being overlooked. Apart from the lion attack on a farmer in 2004 (see above), a good example can be the totally unexpected and reliable report of a lion in Kalapka Resource Reserve,

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Eastern Ghana, towards the border with Togo, where no–one expected such an occurrence. In agreement with the proposal by Chardonnet (2002), we consider the survival of lions in MNP and in Ghana in general could be of strategic importance to having a core group of lions forming a central corridor bridge between Western (Senegal, Guinea) and Eastern (Burkina Faso, Niger, Benin, Nigeria) populations (fig. 1), ideally in the Ivory Coast and in Ghana. Acknowledgements We are grateful to Nana Kofi Adu–Nsiah, Umaru Farouk Dubiure, Enoch Amasa Ashie, Oliver K. Chelewura, (Forestry Commission, Wildlife Division) and all the MNP staff, who assisted us every step of the way, in particular Eric Bani and David W. Kabuiri for their field assistance. We would also like to thank Gianna Da Re (Ricerca & Cooperazione NGO, Italy) for her cooperation and assistance during our visits to Ghana. The institutions sponsoring the project are: the Italian Foreign Affairs Ministry (DGCS), Rome, Italy; the Department of Ecological and Biological Sciences, and the Department of Science and Technology for Agriculture, Forests, Nature and Energy (University of Tuscia, Viterbo, Italy); the Forestry Commission (Wildlife Division), Accra, Ghana; Ricerca & Cooperazione NGO, Rome, Italy. We also thank Mauro Cella for revising and improving the English text. We want to thank John Grainger for giving us a copy of his photo and for allowing us to publish it in our article. We would also like to thank Damiano Luchetti, Marco Signore, Dario M. Soldan, Alberto Zilli and the libraries of IUCN (Gland) and FAO (Rome) for providing us with documents and publications that were difficult to find. Louise Tomsett (Natural History Museum, London) has provided data on museum collections. The following people have collaborated in the field research: Andrea Caboni, Massimiliano Di Vittorio, Stefania Gentili, Edoardo Mastrandrea, Fabio Petrozzi. Luca Luiselli and Massimiliano Di Vittorio contributed to fitting the habitat suitability model. Last, but not least, many thanks to Emmanuel Do Linh San for having encouraged us to submit the manuscript, and for his useful advice, and thanks to the Editor and two anonymous referees for their useful comments on the previous draft of this manuscript. References Akaike, H., 1973. Information theory as an extension of the maximum likelihood principle. In: Second International Symposium on Information Theory: 267–281 (B. N. Petrov & F. Csaki, Eds.). Akademiai Kiado, Budapest. Angelici, F. M., 2006. The Mole National Park. Internal Report (November–December 2005). UNEP, Seniores Italia. Angelici, F. M. & Petrozzi, F., 2010. Lions in the Mole National Park in Ghana, Northern Region. Cat News, 53: 28–31. Anonymous, 2015. Press Release: Lions Making a

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Comeback in Gabon. http://www.panthera.org/blog/ press–release–lions–making–comeback–gabon [Accessed on 12 March 2015]. Arntzen, J. W. & Alexandrino, J., 2004. Ecological modelling of genetically differentiated forms of the Iberian endemic golden–striped salamander, Chioglossa lusitanica. Herpetological Journal, 14: 137–141. Basset, Y., Novotny, V., Miller, S. E., Weiblens, G. D., Missa, O. & Stewart, A. J. A., 2004. Conservation and biological monitoring of tropical forests: the role of parataxonomists. Journal of Applied Ecol� ogy, 41: 163–174. Bauer, H., 2003. Lion conservation in West and Central Africa. Integrating Social and Natural Sci� ence for Wildlife Conflict Resolution around Waza National Park, Cameroon. Leiden University, The Netherlands. Bauer, H. & Van Der Merwe, S., 2004. Inventory of free–ranging lions Panthera leo in Africa. Oryx, 38: 26–31. Black, S. A. & Copsey, J. A., 2014. Does Deming’s 'System of Profound Knowledge' Apply to Leaders of Biodiversity Management? Open Journal of Leadership, 3(2): 53–65. DOI:10.4236/ojl.2014. Black, S. A., Fellous, A., Yamaguchi, N. & Roberts, D. L., 2013. Examining the Extinction of the Barbary Lion and Its Implications for Felid Conservation. Plos One, 8(4): e60174. DOI:10.1371/journal. pone.0060174. Burnham, K. P. & Anderson, D. R., 2002. Model selec� tion and multimodel inference: A practical informa� tion–theoretic approach. New York: Springer–Verlag. Burton, A. C., 2012. Critical evaluation of a long– term, locally–based wildlife monitoring program in West Africa. Biodiversity and Conservation, 21: 3079–3094. Burton, A. C., Buedi, E. B., Balangtaa, C., Kpelle, D. G., Sam, M. K. & Brashares, J. S., 2011a. The decline of lions in Ghana’s Mole National Park. African Journal of Ecology, 49: 122–126. Burton, A. C., Sam, M. K., Kpelle, D. G., Balangtaa, C., Buedi, E. B. & Brashares, J. S., 2011b. Evaluating persistence and its predictors in a West African carnivore community. Biological Conservation, 144: 2344–2353. Cansdale, G. G., 1948. Provisional Check List of Gold Coast Mammals. Printing Department, Accra Government. Chardonnet, P., 2002. Conservation of the African lion: contribution to a status survey. International Foundation for the Conservation of Wildlife, Paris. Collar, N. J., 1998. Extinction by assumption: or, the Romeo error on Cebu. Oryx, 32: 239–244. Danielsen, F., Jensen, P. M., Burgess, N. D., Coronado, I., Holt, S., Poulsen, M. K., Rueda, R. M., Skielboe, T., Enghoff, M., Hemmingsen, L. H., Sørensen, M. & Pirhofer–Walzl, K., 2014. Testing focus groups as a tool for connecting indigenous and local knowledge on abundance of natural resources with science– based land management systems. Conservation Letters. DOI: 10.1111/conl.12100.

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Foster, K. R., Vecchia, P. & Repacholi, M. H., 2000. Science and the Precautionary Principle. Science, 288: 979–981. Grubb, P., Jones, T. S., Davies, A. G., Edberg, E., Starin, E. D. & Hill, J. E., 1998. Mammals of Gha� na, Sierra Leone and the Gambia. The Trendine Press, St. Ives, Zennor. Henschel, P., Azani, D., Burton, C., Malanda, G., Saidu, Y., Sam, M. & Hunter, L. T. B., 2010. Lion status updates from five range countries in West and Central Africa. Cat News, 52: 34–39. Henschel, P., Coad, L., Burton, C., Chataigner, B., Dunn, A., MacDonald, D., Saidu, Y. & Hunter, L. T. B., 2014. The lion in West Africa is Critically Endangered. PLoS ONE, 9(1): e83500. DOI:10.1371/ journal.pone.0083500. Hosmer, D. W. & Lemeshow, S., 1989. Applied logistic regressions. Wiley, New York. Janzen, D. H., 1991. How to save tropical biodiversity. American Entomologist, 37: 159–171. – 2004. Setting up tropical biodiversity for conservation through non–damaging use: participation by parataxonomists. Journal of Applied Ecology, 41: 181–187. Luiselli, L., 2006. Ecological modelling of convergence patterns between European and African 'whip' snakes. Acta Oecologica, 30: 62–68. Myers, N., 1975. The Silent Savannahs. International Wildlife, 5: 5–10. Roberts, D. L., Elphick, C. S. & Reed, J. M., 2009. Identifying anomalous reports of putatively extinct species and why it matters. Conservation Biology, 24: 189–196. Rosevear, D. R., 1974. The Carnivores of West Africa. Trustees of The British Museum (Natural History), London. Smithers, R. H. N., 1983. The Mammals of the Southern African Subregion. University of Pretoria, Pretoria. Sugiura, N., 1978. Further analysis of the data by Akaike’s information criterion and the finite corrections. Communications in Statistics – Theory and Methods, A7: 13–26. Teixeira, J., Ferrand, N. & Arntzen, J. W., 2001. Biogeography of the golden–striped salamander Chioglossa lusitanica: a field survey and spatial modelling approach. Ecography, 24: 618–624. Vapnik, V. N., 2000. The nature of statistical learning theory. Springer, Berlin. Von Bertalanffy, L., 1934. Untersuchen über die Gesetzlichkeiten des Wachstums. I. Roux. Archiive Entwicklungsmech, 131: 613–652. – 1938. A quantitative theory of organic growth (Inquiries on growth laws. II). Human Biology, 10: 181–213. – 1951. Theoretische Biologie – Zweiter Band: Stoffwechsel, Wachstum. A. Francke A. G., Bern. – 1964. Basic concepts in quantitative biology of metabolism. Helgolander Wissenschaftliche Mee� resuntersuchungen, 9: 5–37. Wilson, V. J., 1993. Final report. A zoological survey of Mole National Park, north–western Ghana. Part 1. Large mammals. Accra: Forest Resource Management programme, GWD/IUCN Project 9786.

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Reptile assemblages across agricultural landscapes: where does biodiversity hide? M. Biaggini & C. Corti

Biaggini, M. & Corti, C., 2015. Reptile assemblages across agricultural landscapes: where does biodiversity hide? Animal Biodiversity and Conservation, 38.2: 163–174. Abstract Reptile assemblages across agricultural landscapes: where does biodiversity hide?— The transition from traditional to intensive farming, aimed at large–scale production, has rapidly altered agricultural landscapes, leading to the reduction and fragmentation of natural habitats and to the consequent loss of biodiversity. Herpetofauna is seriously threatened by agriculture intensification worldwide, but less is known about its distribution in agro–ecosystems, especially at field scale. We analysed reptile abundance and diversity in eight agricultural and semi–natural land uses, and inside vegetated buffer strips interspersed among fields. Interestingly, most reptiles were recorded in the buffer strips while intensive crops and pastures hosted just one lizard species. Richness of individuals and species increased when strips were connected to semi–natural areas, independently of their width and vegetation structure. In view of our results, that highlight the role of minor landscape features for the presence of vertebrates in intensive agro–ecosystems, we recommend the implementation of buffer strips among the measures for vertebrate conservation in agricultural landscapes. Key words: Agriculture, Biodiversity, Buffer strips, Herpetofauna, Reptiles Resumen Comunidades de reptiles en paisajes agrícolas: ¿dónde se esconde la biodiversidad?— La transición de la agricultura tradicional a la intensiva, orientada a la producción a gran escala, ha alterado rápidamente los paisajes agrícolas, lo que ha conllevado la reducción y fragmentación de los hábitats naturales y la consiguiente pérdida de biodiversidad. La herpetofauna está gravemente amenazada por la intensificación agrícola en todo el mundo, pero se sabe poco acerca de su distribución en los ecosistemas agrícolas, especialmente a escala local. Se analizaron la abundancia y la diversidad de reptiles en ocho usos del suelo agrícolas y seminaturales, así como dentro de parches de vegetación intercalados entre cultivos. Curiosamente, la mayoría de los reptiles se observó en los parches de vegetación, mientras que en los cultivos intensivos y los pastos solo se encontró una especie de lagarto. La riqueza de individuos y de especies aumenta cuando los parches de vegetación están en contacto con zonas seminaturales, independientemente de la anchura y la estructura de la vegetación de estas. En vista de los resultados obtenidos, que ponen de relieve la influencia de las características del paisaje de menor importancia en la presencia de vertebrados en los ecosistemas agrícolas intensivos, recomendamos incluir parches de vegetación como medida de conservación de los vertebrados en los paisajes agrícolas. Palabras clave: Agricultura, Biodiversidad, Parches de vegetación, Herpetofauna, Reptiles Received: 24 II 15; Conditional acceptance: 14 IV 15; Final acceptance: 11 V 15 Marta Biaggini, Claudia Corti, Museo di Storia Naturale dell’Università degli Studi di Firenze, Sez. di Zoologia 'La Specola', Via Romana, 17, 50125 Firenze, Italy. Corresponding author: M. Biaggini. E–mail: marta.biaggini@virgilio.it

ISSN: 1578–665 X eISSN: 2014–928 X

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Introduction Human activities deeply alter the environment, creating novel habitats and inducing reduction, fragmentation and even loss of the pre–existing habitats. These processes clearly have serious consequences on many organisms, and understanding how anthropic pressure influences the distribution, the population dynamics and the ecology of other species is a fundamental step for conservation. Over the last decades, in most agricultural regions in Europe, there has been a transition from local subsistence farming to more industrial cultivation practices aimed at large–scale production. Intensively cultivated fields have rapidly expanded, leading to the drastic reduction and fragmentation of patches of natural and semi–natural vegetation, and to the creation of more uniform landscapes. Intensive agriculture is largely accepted today as one of the major causes of large–scale biodiversity loss (Wake, 1991; Foley et al., 2005). Reptiles are among the taxa that are primarily threatened by land use changes, habitat fragmentation (Heyer et al., 1994; Gibbons et al., 2000) and, particularly, by the expansion of intensive agriculture, worldwide (Glor et al., 2001; Ribeiro et al., 2009). Due to their ecological and physiological features, relatively low dispersal ability, and small home ranges, reptiles are sensitive even to local habitat changes (Díaz et al., 2000; Driscoll, 2004) and they thus suffer from the consequences of landscape changes more than other vertebrates (White et al., 1997). In the Mediterranean regions, reptiles make up a high portion of the vertebrate fauna in terms of biomass, and they play a key role in ecosystem balance because of their intermediate position in the food web (Rugiero & Luiselli, 1995; Martín & López, 1996; Padilla et al., 2005, 2007; Pérez–Mellado et al., 2008). For such reasons, reptiles can be particularly suitable to detect the consequences of human–induced land use changes on biodiversity. Nevertheless, studies on the effects of intensification of agricultural practices on vertebrates rarely focus on reptiles (but see, for example, Driscoll, 2004; Berry et al., 2005; Ribeiro et al., 2009), concentrating mostly on birds (Donald et al., 2001; Verhulst et al., 2004; Atkinson et al., 2005; Wretengerger et al., 2006), or mammals (Smith et al., 2005; Heroldová et al., 2007). On the other hand, herpetofauna is more often embraced in conservation programs and therefore, knowledge of the distribution of amphibian and reptile species inside agro–ecosystems is key to designing effective conservation strategies and agronomic measures aimed at mitigating the effects of intensive management. In this paper we analysed reptile assemblages in an area mainly devoted to agriculture and dominated by intensively cultivated arable lands. To see how reptiles are distributed in such a landscape, we surveyed and compared reptile abundance and diversity in some agricultural and semi–natural land uses, and also inside vegetated buffers interspersed among crops, namely strips of vegetation along ditch banks and field margins (sensu Greaves & Marshall, 1987). Buffer strips may represent key elements in agro–ecosystems because

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of their role in mitigating against intensive management practices, providing multiple services for water and soil quality (Lynch et al., 1985; Osborne & Kovacic, 1993; Marshall & Moonen, 2002; De Cauwer et al., 2005), and also for invertebrate diversity (Sotherton, 1985; Wratten, 1988; Lagerlöf et al., 1992; Blake et al., 2011; Simão et al., 2015). However, little is known about the role of vegetated buffers in the conservation of vertebrate fauna, particularly regarding small mammals and reptiles (Marshall, 2002). Moreover, given their small extension, these linear landscape elements are neglected in most studies on biodiversity, especially when made at a regional scale and based on land cover databases. Material and methods Study area and sampling method The study was performed in central Italy (45° 42' 49.30'' N, 11° 06' 43.46'' E), in an area (of about 400 km2) mainly devoted to agriculture, with non–irrigated arable lands covering about 64% of the surface, and broad–leaved forests covering about 20% (Corine Land Cover categories) (fig. 1). The altitude of study sites varied from 0 to 180 m a.s.l. To detect reptile abundance and diversity, we performed transects in eight different agricultural (Agr, both crops and pastures) and semi–natural (SNat) land uses, distributed in 31 sites (each including only one land use): broadleaved woodlots (Wo), pinewoods (Pw), sand dune habitats (S), olive orchards with intensive (O) and traditional (Ot) managements, arable lands (A), vineyards (V), pastures (Pa) (fig. 1, table 1 for details). We also surveyed vegetated buffer strips (Bs), linear strips of semi–natural, unmanaged vegetation, which cross the matrix of cultivated lands (table 1). Transects are a quick and effective method to survey reptiles (Latham et al., 2005; Urbina–Cardona et al., 2006); they are particularly practical when sampling more sites in a wide area, and in agricultural lands they allow minimum disturbance to management activity (Paggetti et al., 2006). We walked at constant speed along linear paths, recording every reptile encounter within 1 m on both sides of the observer. Transects were 100 m long on average and were at least 20 m away from one another to prevent multiple recording of the same individual; each transect was replicated twice. Sampling was performed during May and June 2009. Statistical analyses In order to analyse the patterns of reptile abundance and diversity across the sampled land uses, we considered three variables: the number of individuals in 100 m (Nind), the number of species in 100 m (Nsp, considered as a rough index of species diversity), and the Shannon–Wiener index of study sites (H, Shannon & Weaver, 1948). To calculate Nind and Nsp, we used data from single transects. To assess H values, for each land use, we considered the total

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O A Ot

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Wo Pa Wo Wo V Pa O

Ot Pw

Pw S S

A Ot

V Ot A V O Ot

Ot A Pw

30 km

100 km

Fig. 1. Location of the 31 study sites (on the left) and position of the study area at national (top right) and regional scale (bottom right). (For abbreviations see Material and methods and table 1.) Fig. 1. Ubicación de los 31 sitios de estudio (a la izquierda) y situación de la zona de estudio a escala nacional (arriba a la derecha) y a escala regional (abajo a la derecha). (Para las abreviaturas ver Material and methods y tabla 1.)

number of individuals observed in the different study sites, while BS transects were grouped in relation to the land uses which they adjoined (i.e., Bs bordering A). We then considered the following environmental variables: study site area (Area); study site edge density (ED, perimeter/area), an indicator of spatial heterogeneity, taking into account the shapes of patches (EEA, 2000; Walz, 2011); and vegetation structure (VEG). Considering that we mainly dealt with ground dwelling reptiles, transects were classified in two VEG categories on the basis of the ground vegetation structure of the land uses they belonged to: i) VEG1, ground vegetation absent or made up exclusively by herbaceous species; ii) VEG2, presence of shrubs. To each buffer strip (Bs) transect we associated three environmental variables: i) vegetation structure (VEG1, VEG2); ii) average width of buffer strips (W, the average value of three measures taken at the beginning, in the middle, and at the end of each transect); iii) degree of connectivity, indicating whether strips were connected to semi–natural areas (i.e., woodlots or wetlands with a minimum area of 300 ha) (CON1, not connected; CON2, connected with one semi–natural area; CON3, in connection with more semi–natural areas). Width of buffer strips, area, edge density, and connectivity of buffer strips were assessed using Geoportale Nazionale orthophotos and tools (www.pcn.minambiente.it).

We tested the number of individuals (Nind) and species (Nsp) for spatial autocorrelation, using Moran’s values obtained at 10 different distance intervals. To verify how abundance of individuals and species varied in the study area, we compared Nind and Nsp, first among Agr and SNat categories (agricultural and semi–natural lands) and Bs (buffer strips), then among the eight land uses and Bs. Owing to the large number of transects with no observations, Nind and Nsp were not normally distributed even after log–transformation (Kolmogorov–Smirnov, n = 204: Nind, x2 = 100.109, p < 0.001; Nsp, x2 = 109.388, p < 0.01). For this reason, we used log–linear models, assuming a Poisson distribution of the data (Sutherland, 2006; Wilson et al., 2007). Multiple comparisons of mean rank were then applied (Siegel & Castellan, 1988). Such analyses were applied to all the following comparisons. To test the possible influence of vegetation structure on reptile abundance, we compared Nind and Nsp between VEG categories in Agr and in SNat. In the eight land uses, we also tested the possible influence of study site Area and ED (edge density) on the mean values of the number of individuals (Nind) and number of species (Nsp), and on the Shannon–Wiener index (H), performing Spearman correlations. We also compared H values among land use categories Agr, SNat, and Bs (while the number of study sites per land use was

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Table 1. List of surveyed land uses (L) and their main environmental features, categories (Categ), VEG category, number of transects (Nt) and study sites (Ns); the total number of recorded species (Nsp) and mean values (± SD) of reptile abundance (Nind), species abundance (Nsp), and diversity (H) are also reported. Land uses: Wo. Broadleaved woodlots; Pw. Pinewoods; S. Sand dune habitats; O. Olive orchards; Ot. Traditional olive orchards; A. Arable lands; V. Vineyards; Pa. Pasture; Bs. Buffer strips. Categories: Agr. Agricultural; SNat. Semi–natural.

Categ

VEG

Descriptive notes

SNat

Mainly oak (Quercus sp.) forests, natural or partially managed

SNat

Pinus pinaster forests along the coast.

A Few herbaceous species under trees

Coastal sand dunes. Sparse vegetation essentially

SNat

made up of grasses

Olive tree (Olea europaea) plantations, intensively managed.

Agr

Use of chemicals and machinery, ploughed soil,

and almost absent grass

Olive tree (O. europaea) plantations with 'traditional'

Agr

management: maintenance of soil cover (mainly herbaceous species

but also sparse bushes), scarce or absent use of machinery

Agr

Mainly cereal and alfalfa fields

Agr

Mainly intensively managed vineyards (Vitis vinifera).

Agr

Use of chemicals and machinery, ploughed soil Lowland meadows (with very low diversity of grass species)

and pastures devoted to pig farming

Semi–natural strips of vegetation bordering agricultural

–

lands: strips of riparian vegetation along ditches and banks of

small rivers and field margins. Mean width ranging from 2 to 19 m;

varying vegetation structure

too low to allow the comparison of H among land uses). Focusing on buffer strips (Bs), we compared Nind and Nsp in CON and VEG categories in order to verify the possible influence of connectivity and vegetation structure on abundance of reptiles and species. Finally, we tested the influence of Bs width (W) on Nind and Nsp using Spearman correlation. We used Statistica 10.0 (StatSoft, Inc., 2011) for all the analyses, except for spatial autocorrelation analysis performed by PAST 2.17b package (Hammer et al., 2001). Results During sampling of transects we recorded a total of 278 individuals belonging to seven reptile species: Chalcides chalcides (23), Lacerta bilineata (15), Podarcis muralis (29), P. siculus (206), Hierophis viridiflavus (4) and Vipera aspis (1).

Spatial autocorrelation did not affect the patterns of abundance observed (of both individuals and species) (fig. 2). Reptile abundance (Nind) differed among land use categories, with agricultural land uses (Agr) hosting the lowest number of reptiles, significantly lower than buffer strips (Bs) (table 2, fig. 3). The comparisons among land uses showed that arable lands and intensively managed olive orchards (A and O) hosted the lowest number of reptiles, significantly differing, in particular, from buffer strips (Bs) (see table 2 for other significant results, figs. 3, 5). Nind also varied significantly between VEG categories: both agricultural and semi–natural land uses with a simplified ground vegetation (absent or made up of just herbaceous species) were significantly poorer in numbers of reptiles than land uses where shrubs were also present (table 2, fig. 3). Finally, in Agr land uses, mean Nind was negatively correlated with field Area (n = 22, r = –0.506, p = 0.019) and increased with increasing edge density (ED) (n = 22, r = 0.528, p = 0.014). In SNat land uses, mean Nind did not co-

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Tabla 1. Lista de los usos agrícolas estudiados (L) y sus principales características ambientales, categorías (Categ), categoría VEG, número de transectos (Nt) y de áreas estudiadas (Ns); también se muestran el número total de especies observadas (Nsp) y los valores medios (± DE) de la abundancia de reptiles (Nind), de la abundancia de especies (Nsp) y de la diversidad (H). Usos del suelo: Wo. Bosques de frondosas; Pw. Pinares; S. Hábitats de dunas; O. Olivares; Ot. Olivares tradicionales; A. Tierras arables; V. Viñedos; Pa. Pastos; Bs. Parches de vegetación. Categorías: Agr. Agrícola; SNat. Seminatural. Nsp

Mean Nsp (± SD)

Mean Nind (± SD)

Mean H (± SD)

0.603 ± 0.682

0.977 ± 1.286

1.068 ± 0.359

0.267 ± 0.438

0.344 ± 0.625

0.615 ± 0.569

1.705 ± 1.307

0.161 ± 0.227

0.077 ± 0.277

0.561 ± 0.677

0.805 ± 1.034

0.215 ± 0.577

0.053 ± 0.275

0.106 ± 0.550

0.343 ± 0.626

0.438 ± 0.729

0.303 ± 0.674

0.455 ± 1.078

0.926 ± 1.122

1.864 ± 2.227

0.888 ± 0.800

rrelate with Area (n = 9, r = 0.350, p = 0.359) and ED (n = 9, r = 0.317, p = 0.385). The analyses of diversity across land uses gave results analogous to those concerning reptile abundance. In agricultural (Agr) land uses, we recorded the lowest number of species (Nsp), significantly different from that recorded in buffer strips (Bs), the land use with the highest Nsp values (table 2, fig. 4). Arable lands and intensively managed olive orchards (A and O) hosted the lowest number of species (table 2, fig. 4). Agricultural land uses with more complex vegetation structure had significantly higher Nsp values, while in SNat we recorded no differences among VEG categories. As observed for reptile abundance, in Agr land uses Nsp was negatively correlated with field Area (n = 22, r = –0.462, p = 0.035), and positively correlated with edge density (ED) (n = 22, r = 0.486, p = 0.026). In SNat land uses, we did not find correlations between Nsp and Area (n = 9, r = –0.050, p = 0.880) and ED (n = 9, r = 0.133, p = 0.708).

The comparison of the Shannon index (H) among land use categories confirmed that levels of biodiversity in Agr land uses were significantly lower than those recorded in Bs (table 2). Considering cultivated lands, H was different from zero only in traditional olive orchards (Ot) (table 1). H was not influenced by site Area and ED either in Agr (n = 22: Area, r = –0.111, p = 0.633; ED, r = –0.110, p = 0.632) or in SNat (n = 9: Area, r = –0.438, p = 0.269; ED, r = 0.310, p = 0.422). Focusing on buffer strips (Bs), we found that abundance of reptiles and species did not differ significantly between vegetation categories (table 2), and they were not correlated with buffer width (n = 61: Nind, R = 0.178, p = 0.177; Nsp, R = 0.147, p = 0.266). On the contrary, both Nind and Nsp varied in relation to connectivity: buffer strips characterized by the lowest connectivity level hosted significantly fewer individuals and species than buffer strips connected with at least one semi–natural area (table 2).

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Nind

0.8

0.6

0.4

0.2

Moran I

0.8

–0.2

–0.4

–0.6

–0.8

–1.0

Nsp

–1.0 12 16 20 24 28 32 36 0 Distance (km)

12 16 20 24 28 32 36 Distance (km)

Fig. 2. Moran’s I correlograms for number of individuals (Nind) and number of species (Nsp). Circles are non–significant values, continuous lines indicate P critical value (0.05). Fig. 2. Correlogramas de Moran I para el número de individuos (Nind) y el número de especies (Nsp). Los círculos representan los valores no significativos y las líneas continuas indican el valor crítico P (0,05).

Discussion In most European countries, traditional agricultural landscapes, typically characterized by mosaic–like structures, intermediate levels of disturbance, and able to host great biodiversity levels (Bignal & McCracken, 1996), have undergone profound transformations in last decades. The rapid extension of intensively cultivated fields and the reduction and fragmentation of the original habitats have led to the creation of more uniform and depleted landscapes, with significant loss of biodiversity (Whittaker, 1975; Bull & Skovlin, 1982; Burel et al., 1998; Zechmeister & Moser, 2001; Moser et al., 2002; Pfiffner & Luka, 2003). In accordance with these changes, our analysis of reptile assemblages in an agricultural area mainly devoted to the intensive cultivation of arable lands found the landscape to be relatively poor in biodiversity. Four out of five agricultural land uses hosted only one species, while most of the recorded reptiles occurred inside a few patches of semi–natural habitats and, above all, in the grid of vegetated buffer strips interspersed among the cultivated lands. In addition, during field activity, we observed only seven of the eighteen species of terrestrial reptiles reported for the province where the sites are located, and most records were ascribable to a single species. Specifically, we recorded an overriding presence of the lacertid lizard Podarcis siculus (Italian wall lizard), which greatly influenced the patterns of reptile abundance and diversity here described. P. siculus was the only species present in all the surveyed land uses, particularly including all the cultivated lands. A large diffusion of this lizard in the study area is partly ascribable to the ecological requirements of the species that prefers flat, relatively open habitats (Corti

et al., 2010). However, the striking differences in the distribution and abundance of P. siculus with respect to all the other species go beyond mere considerations on habitat preference. In general, in the presence of human–induced landscape alterations, most species can be disadvantaged if suitable conditions for their ecological requirements persist only in fragments of natural habitats (Doak et al., 1992; Bender et al., 1998; Laurance et al., 1998). However, it can also happen that some species benefit from the novel habitat matrix (Laurance et al., 2002; Cardador et al., 2011). This could be the case of P. siculus that, at least to a certain extent, is probably able to resist land use transformations or even to take advantage of the expansion of cultivated lands, open and often depleted areas where other species cannot persist. On the other hand, the low number of species that we recorded could also be partly due to the sampling method: transecting is particularly efficient for detecting species like lizards but it could be less suitable for others, also in relation to the surveyed environments (McDiarmid et al., 2012). However, the gap between the species observed and those potentially present in the area is so wide that it could indicate a real lack, likely related to the scarce availability of suitable habitats. With the data at our disposal, we identified two main categories of land uses: those characterized by conditions apparently adverse to reptiles, where just one species could be observed or it clearly prevailed on the others, and the few land uses in which more species occurred and faunal composition was better balanced (table 1, fig. 4). Almost all cultivated plots, that represented the environmental matrix of the area, belonged to the former category: with the only exception of traditionally managed olive orchards, where

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169

Table 2. Results of comparisons of reptile abundance (Nind) and diversity (Nsp and H) among land use categories (Agr and SNat) and buffer strips (Bs), land uses and Bs, vegetation categories (VEG), and connectivity levels (CON) (just for Bs transects). Tabla 2. Resultados de la comparación de la abundancia de reptiles (Nind) y la diversidad (Nsp y H) entre el uso agrícola (Agr), el uso seminatural (SNat) y los parches de vegetación (Bs); entre el conjunto de todos los usos del suelo y los Bs; entre las categorías de estructura de la vegetación (VEG), y entre los grados de conectividad (CON) (solo para los transectos en los Bs). Variable Comparisons

Multiple comparisons

Wald

Agr = 112; SNat = 41 Bs = 61

72.977 < 0.001

Agr < Bs, SNat

All land uses and Bs See table 1

74.248

< 0.001

A, O < Bs, S, V < Bs

VEG in Agr

VEG1 = 51; VEG2 = 51

11.606

< 0.001

–

VEG in SNat

VEG1 = 13; VEG2 = 28

6.484

0.011

–

VEG in Bs

VEG1 = 33; VEG2 = 28

0.165

0.684

–

CON in Bs

CON1 = 13; CON2 = 30; CON3 = 18 12.395

0.002

CON1 < CON2, CON3

Agr/SNat/Bs

Agr = 112; SNat = 41; Bs = 61

Nind Agr/SNat/Bs

Nsp 27.379 < 0.001

Agr < Bs, SNat

All land uses and Bs See table 1

29.085

< 0.001

A, O < Bs

VEG in Agr

VEG1 = 51; VEG2 = 51

9.065

0.002

–

VEG in Snat

VEG1 = 13; VEG2 = 28

1.940

0.164

–

VEG in Bs

VEG1 = 33; VEG2 = 28

0.054

0.816

–

CON in Bs

CON1 = 13; CON2 = 30; CON3 = 18 10.581

0.005

CON1 < CON3

Agr/SNat/Bs

Agr = 22; SNat = 9; Bs = 6

0.008

Agr < Bs

four reptile species were detected, agricultural lands hosted just one lizard species (P. siculus). The most exacerbated situation was observed inside arable lands, where lizards occurred exclusively near field margins and never in the middle of crops, as noticed in other agricultural areas (pers. obs.; Biaggini et al., 2011). Although specific studies should be performed to strengthen these observations (Kéri, 2002), our data suggest that the extension of fields negatively influenced abundance of both individuals and species, while increasing edge density supported higher values of the two variables. Moreover, a more complex vegetation structure enhanced reptile diversity and abundance, as found in other Mediterranean agricultural landscapes (Germano & Hungerford, 1981; Martín & López, 2002). All these observations further stress how the occurrence of very large monocultures (especially of arable lands) can negatively impact on reptiles in agricultural landscapes. Higher complexity of reptile communities, in terms of both diversity and abundance, subsisted in semi–natural patches (especially in broadleaved woodlots). The importance of such patches has

9.715

also been largely demonstrated for vascular plants, birds and arthropods (i.e., Billeter et al., 2008). More interestingly, vegetated buffer strips mostly contributed to enhance fauna richness in the surveyed agricultural landscape, showing the highest levels of reptile diversity and abundance among the analysed land uses. The importance of these linear landscape features for increasing biodiversity in rural landscapes dominated by intensive managements has been already stressed with regards to flora (Barr et al., 1993), invertebrates (Sotherton, 1985; Wratten, 1988; Lagerlöf et al., 1992), mammals (Pollard & Relton, 1970; Boone & Tinklin, 1988; Fitzgibbon, 1997; Verboom & Huitema, 1997), and birds (O'Connor, 1987; Lack, 1992; Vickery & Fuller, 1998), but not for reptiles. Interestingly, neither the complexity of vegetation structure nor strip width were crucial to determining the presence of reptiles inside buffer strips. This was in contrast with results found in both agricultural and semi–natural land uses, where vegetation structure played a role in shaping reptile presence. On the contrary, the factor that significantly influenced abundance of individuals

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Nind

Mean Mean ± SE Mean ± SD

1 0

–1

1 0

–1 SNat Agr Bs Wo Pw S O Ot A V Pa Bs Land use category Land use

Fig. 3. Reptile abundance (Nind) in the surveyed land use categories (left) and land uses (right), and buffer strips (Bs) (box–plots show mean value ± SD). Fig. 3. Abundancia de reptiles (Nind) en las categorías de usos del suelo (izquierda) y los usos del suelo (derecha) estudiados, así como en los parches de vegetación (Bs) (los diagramas de cajas muestran la media ± DE).

and species inside buffer strips was their degree of connectivity. Namely, the presence of reptiles was minimal in those strips that were not in connection with any semi–natural area. All these observations suggest that reptiles do not settle in buffer strips but exploit them as temporary refuges while foraging at crop margins or during displacements (Madsen, 1984; Wisler et al., 2008). Usually, both vegetation

Mean Mean ± SE Mean ± SD

2.2 2.0 1.8 1.6 1.4 1.2 1 0.8 0.6 0.4 0.2 0 –0.2 –0.4 –0.6

Nsp

2.2 2.0 1.8 1.6 1.4 1.2 1 0.8 0.6 0.4 0.2 0 –0.2 –0.4

structure and strip width are key factors for animal groups, like invertebrates, that steadily inhabit these strips of vegetation (De Cauwer et al., 2005). Specific studies are obviously required to disclose the way in which reptiles exploit buffer strips. However, the grid made up of field borders and strips of riparian vegetation along watercourses may allow reptiles to penetrate the 'barrier' of intensive crops and to

SNat Agr Bs Land use category

Wo Pw S O Ot A V Pa Bs Land use

Fig. 4. Number of species (Nsp) in the surveyed land use categories (left) and land uses (right), and Bs (box–plots show mean value ± SD). Fig. 4. Número de especies (Nsp) en las categorías de usos del suelo (izquierda) y los usos del suelo (derecha) estudiados, así como en los Bs (los diagramas de cajas muestran la media ± DE).

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2 1.8 1.6

Nind/100 m

1.4 Snakes

1.2

C. chalcides

L. bilineata

0.8

P. muralis

0.6

P. siculus

0.4 0.2 0

Pw Wo

Fig. 5. Reptile abundance and fauna composition in the surveyed land uses (columns indicate the mean number of reptiles in 100 m, grey tones indicate the different species, and bars indicate 5% errors). Fig. 5. Abundancia de reptiles y composición faunística en los usos del suelo estudiados (las columnas indican el número medio de reptiles en 100 m, los tonos grises indican las diferentes especies y las barras, el 5% de error).

disperse among cultivated areas. Thanks to their relatively thick vegetation, buffer strips probably represent the safest crosswalk available in intensive agricultural landscapes. Accordingly, the role of field margins and riparian strips in supporting fauna movement across cultivated lands has previously been observed for invertebrates (Burel, 1989), bats (Verboom & Huitema, 1997) and birds (Machtans et al., 1996). In such a perspective, buffer strips would play a key ecological function, considering that the presence of a matrix of unsuitable habitats can represent, in some cases, a selective filter for species throughout the landscape (Gascon et al., 1999) and can prevent dispersion of individuals and gene flow (Wilcove et al., 1986). Focusing on the analysis of reptile assemblages, our study confirms that in landscapes dominated by intensive agriculture (mainly arable lands) biodiversity is low and concentrated in a few, less managed, landscape features. In general, analyses made at a regional scale individuate such features in patches of semi–natural vegetation with quite large surfaces (i.e., woodlot, wetlands) or, at least, in wide vegetated river banks. Interestingly, working at field scale allowed us to highlight the key importance of 'minor' landscape features for the presence of vertebrates in agro–ecosystems, namely strips of

riparian vegetation along the banks of ditches and small rivers, and field borders. Even if relatively narrow and simple in their vegetation structure, these linear features can greatly contribute to the presence of reptiles in agro–ecosystems (especially when in connection with semi–natural patches), and they probably play a key ecological role in allowing dispersal of individuals and species across intensive crops. On the other hand, the absence of reptiles inside intensively managed plots clearly points to the need for mitigation measures aimed at enhancing vertebrate diversity in agricultural landscapes. We strongly recommend the implementation of a grid of vegetated buffer strips together with conservation of the remaining semi–natural patches among the measures for biodiversity conservation in agro–ecosystems. Acknowledgements The study was part of a broader research financed by the Dipartimento Protezione della Natura, Ministero dell'Ambiente e della Tutela del Territorio e del Mare (2009). We thank Neftalí Sillero for the translations into Spanish, and three anonymous referees for their useful comments.

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The role of frugivorous birds and bats in the colonization of cloud forest plant species in burned areas in western Mexico J. Rost, E. J. Jardel–Peláez, J. M. Bas, P. Pons, J. Loera, S. Vargas–Jaramillo & E. Santana

Rost, J., Jardel–Peláez, E. J., Bas, J. M., Pons, P., Loera, J., Vargas–Jaramillo, S. & Santana, E., 2015. The role of frugivorous birds and bats in the colonization of cloud forest plant species in burned areas in western Mexico. Animal Biodiversity and Conservation, 38.2: 175–182. Abstract The role of frugivorous birds and bats in the colonization of cloud forest plant species in burned areas in western Mexico.— The extension of montane cloud forests in western Mexico is threatened by several disturbances that limit their extension. In this study we aimed to assess the contribution of birds and bats in the dispersal and colonization of cloud–forest plants in contiguous surface–burned pine forests. We sampled seed rain and sapling establishment over one year in two surface–burned sites, which differed in the size of their closest cloud forest patch. A total of 17 plant species were found, most of which were late–successional trees, shrubs and climbers. Distance influenced the seed rain of only one dispersed taxon (Solanum sp.) and had no effect on the sapling distribution of this or other plants. In turn, marked differences were found between sites, with more seeds dispersed and higher sapling density in the site that was next to the larger cloud forest patch. The role of long–distance dispersers and the existence of seed banks before fire could explain the little importance of distance from seed source on seed dispersal and sapling distribution. Nevertheless, dispersal by birds and bats before or after fire facilitates the regeneration and conservation of cloud forests in disturbed areas formerly occupied by other habitats. Key words: Cloud forest, Colonization, Disturbance, Fire, Seed dispersal, Seed rain, Succession Resumen La función de las aves y los murciélagos frugívoros en la colonización de las especies vegetales del bosque nuboso en zonas afectadas por incendios en el occidente de México.— La extensión de los bosques nubosos de montaña (o bosques mesófilos) del oeste de México se ve limitada por diversos factores. En el presente estudio se pretende evaluar la contribución de las aves y los murciélagos a la dispersión y colonización de las plantas del bosque nuboso en pinares contiguos afectados por incendios de superficie. Se muestreó la lluvia de semillas y el establecimiento de plántulas durante un año en dos sitios afectados por incendios de superficie, que diferían en el tamaño del fragmento de bosque nuboso más cercano. Se encontró un total de 17 especies vegetales, la mayoría de las cuales eran árboles, arbustos y lianas propios de estados avanzados de sucesión. La distancia respecto a la fuente de semillas influyó solamente en la lluvia de semillas de un taxón (Solanum sp.), pero no tuvo efectos en la distribución de las plántulas de esta especie ni de otras. En cambio, se encontraron importantes diferencias entre los sitios, con más semillas dispersadas y una mayor densidad de plántulas en el que estaba situado junto al fragmento más extenso de bosque nuboso. La función de los vectores de dispersión a larga distancia y la existencia de bancos de semillas previos al incendio podrían explicar la escasa importancia de la distancia desde la fuente de semillas para la lluvia de semillas y la distribución de las plántulas. En todo caso, la dispersión mediante las aves y los murciélagos antes o después del incendio facilita la regeneración y la conservación del bosque nuboso en zonas perturbadas que anteriormente hubieran estado ocupadas por otros hábitats. Palabras clave: Bosque nuboso, Colonización, Perturbación, Incendio, Dispersión de semillas, Lluvia de semillas, Sucesión Received: 14 XI 14; Conditional acceptance: 22 I 15; Final acceptance: 9 VI 15 Josep Rost, Dept. Ciències Ambientals i Indústries Alimentàries, Univ. de Vic–Univ. Central de Catalunya, c/ de la Laura 13, 08500 Vic, Barcelona, Spain.– Enrique J. Jardel–Peláez, Juan Loera, Socorro Vargas–Jaramillo & Eduardo Santana, Dept. de Ecología y Recursos Naturales–IMECBIO, Centro Universitario de la Costa Sur, Univ. de Guadalajara, c/Independencia Nacional 151, 48999 Autlán, Jalisco, México.– Josep Rost, Josep M. Bas & Pere Pons, Dept. de Ciències Ambientals, Fac. de Ciències, Univ. de Girona, Campus de Montilivi, 17071 Girona, Spain. Corresponding author: Josep Rost. E–mail: joseprost@gmail.com ISSN: 1578–665 X eISSN: 2014–928 X

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Introduction Vegetation recovery after natural or human–caused disturbances such as wildfires, hurricanes, floods, grazing or logging depends on disturbance characteristics and on the regeneration capability of plants (Runkle, 1985). Regeneration is possible by several strategies, such as on–site regeneration, e.g., by resprouting from undamaged tissues or germinating from a seed bank, and colonization, by seeds newly dispersed from other areas (Turner et al., 1998; Keeley & Fotheringham, 2000; Norden et al., 2009). To interpret vegetation succession and dynamics in disturbed areas, it is therefore crucial to fully understand the way different plant species can regenerate after an adverse event and the mechanisms and factors that interact with such regeneration at different scales (from disturbance severity, frequency and spatial heterogenity to specific seed facilitation, tolerance and inhibition) (Pickett et al., 1987; Reed et al., 2000; Chazdon, 2003; Balke et al., 2014). Montane forest ecosystems in western Mexico are characterized by a complex landscape mosaic, resulting from environmental gradients related to climate, soil and geomorphological conditions, historical disturbance regimes, and a long history of human use (Jardel et al., 2004a). During the 20th century, most forested areas in Mexico were affected by the impact of logging, cattle grazing and disruption of the natural frequent–fire fire regimes (Heyerdahl & Alvarado, 2003), factors that contributed to shaping the current landscape of the region. After severe disturbances such as stand–replacement fires and clearcutting, the new forest gaps are mainly colonized by pines, which are later replaced by shade–tolerant cloud–forest species in advanced successional stages (González–Espinoza et al., 1991; Jardel, 1991). In absence of disturbances such as surface fires, these pine–dominated forests can turn into pine–oak forests in dry hilltops or into cloud forests in concave landforms and humid slopes (Jardel et al., 2004b). This process is specially interesting because cloud forests are a conservation priority in Mexico (Challenger, 1998) as they share plant species of both Holarctic and Neotropical biogeographic origins and host high levels of biodiversity (Miranda & Sharp, 1950; Rzedowski, 2006). Fire recurrence is expected to decrease in the short–tem in certain parts of western Mexico due to the human depopulation in recent decades (Jardel et al., 2004b). Therefore, if fire is currently the most important factor limiting the expansion of cloud forest in such areas, in the near future cloud forests could expand significantly from their current range (restricted to hollows) to other suitable areas, such as mountain slopes. Colonization of open habitats such as abandoned pastures and farmlands by cloud–forest species is mainly possible due to wind and vertebrate (endozoochory) seed dispersal, and to seed banks of propagules that could have arrived previously by these mechanisms (Del Castillo & Pérez–Ríos, 2008; Muñiz–Castro et al., 2006; Zuloaga–Aguilar et al., 2010). The importance of endozoochorous dispersal is suggested by the high number of plants that show this type of dispersal strategy in cloud forests in western Mexico (up to 87 species;

Rost et al.

Orozco, 1999), and the high number of birds, bats and terrestrial mammals that feed on fleshy fruits in this habitat (Hernández–Conrique et al., 1997; Santana, 2000). Apart from diet specialization, seed shadow, i.e., the spatial distribution of dispersed seeds, depends on several factors that influence the spatial signature of the dispersal vector in the landscape (Howe & Smallwood, 1982; Schupp et al., 2002). The abundance of these birds and mammals, their feeding behaviour, roosting habits, habitat preferences, mobility and speed of their digestion process will affect the seed shadow, the quantity of seeds dispersed, and the quality of the area where seeds are dispersed. All these factors have crucial implications for plant populations (Clark et al., 2005; Jordano, 1986; Schupp et al., 2002). In this regard, different frugivores can disperse seeds at different distances from the seed source (Jordano et al., 2006), and distance from this source is an important constraint in the regeneration of tropical forests in disturbed areas (Chazdon, 2003; Souza, 2014). In this study we aimed to assess the influence of distance on the endozoochorous seed rain and shadow —mediated by flying vertebrates (birds and bats)— of plant species being dispersed from cloud forests into surface–burned pine forests in montane ecosystems of western Mexico. In particular, we analyzed whether: (1) frugivorous bats and birds can contribute to cloud–forest regeneration in a disturbed habitat; and (2) whether distance from undisturbed primary forest can affect the seed dispersal carried out by these frugivores and plant establishment. Material and methods Study area The study was carried out in the 1,250–ha Las Joyas Research Station, located in the Sierra de Manantlán Biosphere Reserve, in the state of Jalisco, Mexico (19º 36' N, 104º 16' W). Las Joyas features a mean annual temperature of 15ºC and mean annual rainfall of 1,800 mm. The area has a rugged relief with an altitudinal gradient ranging from 1,500 to 2,242 m a.s.l. The vegetation in the area is basically composed by pine–hardwood and pine–oak forests, with stands of cloud forest in hollow areas, ravines and along streams, as well as patches of secondary vegetation (Jardel et al., 2004a). We chose two study sites within the research station (at approximately 2,000 m a.s.l.), consisting of two pine–forest stands aged between 40 and 50 years old, both located within an area of 303 ha that burned in a surface wildfire in May 2003. These study sites were located adjacent to two streams. One stream acted as a barrier to the fire spread, and consequently separates the burned area from the undisturbed cloud forest. The other stream was completely surrounded by burned vegetation, except for a narrow strip of about 10–30 m wide at each flank, composed of unburned cloud forest vegetation (fig. 1). The abundance of fruits and dispersers could therefore be expected to be higher in the first area than in the second. The post–fire vegetation was

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B A

150

300 m

Fig. 1. Spatial distribution of the seed traps (dots) in the two study sites (A and B). Thin lines show elevation curves, bold grey lines represent water courses, and the grey area shows the surface–burned area. Fig. 1. Distribución espacial de las trampas de semillas (puntos) en los dos sitios de estudio (A y B). Las líneas delgadas muestran las curvas de elevación, las líneas gruesas grises representan los cursos de agua y la superficie gris muestra el área afectada por un incendio de superficie.

composed of charred but live pines (which conserved the canopy intact), a minority of dead burned pines (< 5%), and regenerating shrubs in the understory. In these study sites, several frugivorous birds and bats were observed (through 23 point counts, lasting for 10–min, surveyed in September and October) and/ or trapped by mist–netting (thirteen 12–m long mist– nets were used for three days in October, from dawn to 8 h later). However, accurate monitoring to estimate their occurrence or relative abundance throughout the year was not possible. The detected frugivorous birds were Penelope purpurascens, Trogon mexicanus, Piranga flava, Myadestes occidentalis, Turdus assimilis, Catharus occidentalis, Catharus guttatus, Piranga erythrocephala, Melanotis caerulescens and Icterus pustulatus. In addition, Santana (2000) found up to 31 frugivore and 28 omnivore birds that could potentially be seed dispersers in the same area. Regarding bats, we were only able to detect (by mist–netting on two nights in October, from dusk to 5 h later) the presence of two species, Sturnira ludovici and Dermanura tolteca. Seed rain and sapling establishment We sampled seed rain of fleshy–fruited plant species that occur in the cloud forest (Cuevas–Guzmán et al., 2004) using seed traps. We set up five trap transects separated 50 m from each other in each study site, placing one seed trap at every selected distance (10, 25, 50, and 100 m) from the fire border (40 seed traps in total; fig. 1). Traps consisted of 1 × 1 m squares made

with a permeable cloth that retained all solid material but not water. The cloth was tied to four poles at 1 m high above the forest ground. The traps were covered with a 1–cm wire mesh to keep rodents and granivorous birds out of the collected material and thus avoid seed predation. We gathered the material collected by the seed traps monthly, from September 2008 to August 2009. This material was later analyzed in the laboratory, where seeds were separated from any other debris. Seeds were identified to the most precise taxonomic level possible, using a reference collection and available literature on local and Mexican flora (Carranza, 1992; González–Villareal, 1996; Lozada, 2000; Ramírez, 1999). Only those seeds of cloud forest plants that undoubtedly came from undisturbed areas were considered for the analysis. Therefore, seeds of cloud–forest plants that occurred and bore fruits in the disturbed area were discarded, in order to avoid counting seeds of uncertain origin. On the other hand, saplings were sampled in the same study sites and distance categories used for seed traps. Sapling density was sampled in ten 1 x 2 m2 randomly located along the four distance categories, in each site (i.e., 80 squares). During seed rain and sapling surveys, we also looked for the presence of scats from medium–sized mammals between traps or quadrats, but none were found. Data analysis We used Generalized Linear Models (GLM) to test for a possible effect of distance in the dispersal of seeds

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Table 1. Cloud–forest fleshy–fruited plants whose seeds were dispersed by birds or bats in surface–burned pine forests adjacent to cloud forest patches. The mean density of seeds (seeds/m2) for each distance category (± SE) is shown. We also give information on the type of life form (LF) each species belongs to (H. Herb; S. Shrub; T. Tree; C. Climber), and the frequency of appearance of each species in the seed traps (number of traps with presence divided by total number of traps). Tabla 1. Plantas de frutos carnosos de los bosques nubosos cuyas semillas se dispersaron mediante aves o murciélagos en pinares afectados por un incendio de superficie y adyacentes a fragmentos de bosques nubosos. Se muestra la densidad media de semillas (semillas/m2) para cada categoría de distancia (± EE). También se proporciona información sobre el tipo de forma de vida (LF) al que pertenece cada especie (H. Hierba; S. Arbusto; T. Árbol; C. Liana), así como la frecuencia de aparición de cada especie en las trampas de semillas (el número de trampas con presencia dividido por el número total de trampas). Species Solanum spp. Phytolacca spp.

Mean density (seeds/m2) LF

Frequency

10 m

25 m

H, S, T

0.34

2.3 ± 1.3

0.9 ± 0.9

0.0 ± 0.0 0.3 ± 0.3

50 m

100 m

H, S

0.24

1.3 ± 0.5

0.1 ± 0.1

0.9 ± 0.4 0.1 ± 0.1

Cornus disciflora DC.

0.17

0.3 ± 0.2

0.4 ± 0.2

0.1 ± 0.1 0.9 ± 0.6

Rhamnus hintonii

0.09

0.1 ± 0.1

0.0 ± 0.0

0.4 ± 0.4 0.4 ± 0.3

0.04

0.2 ± 0.2

0.1 ± 0.1 0.0 ± 0.0 0.1 ± 0.1

M. C. Johnst & L. A. Johnst Magnolia iltisana A. Vázquez Persea hintonii C. K. Allen

0.04

0.0 ± 0.0

0.2 ± 0.2

0.2 ± 0.2 0.0 ± 0.0

Vitis blancoi Munson

0.04

0.2 ± 0.1

0.0 ± 0.0

0.2 ± 0.1 0.0 ± 0.0

Cornus excelsa Humb.

0.03

0.1 ± 0.1

0.1 ± 0.1 0.0 ± 0.0

Ilex brandegeana Loes

0.01

0.1 ± 0.1

0.0 ± 0.0

0.0 ± 0.0 0.0 ± 0.0

Myrsine jurgensenii Mez.

0.01

Bonpl. & Kunth

Total

of fleshy–fruited cloud–forest plants. We tested the effect of distance (as fixed effect factor) on the total number of seeds and total number of saplings. In order to consider the differences between the two study sites, we also included site as a fixed effect factor, and we also tested the interaction between site and distance. Moreover, we analyzed the effect of distance and site on the number of seeds and saplings of those species present in at least 5% of the samples. We used Poisson or quasipoisson error distributions (depending on whether the model presented overdispersion or not) and log link. The significance of the effects was tested using likelihood ratio tests, with chi–square tests in the case of Poisson–distributed models and F–tests in the case of quasipoisson models. Statistical significance level was set at P < 0.05. These analyses were carried out using R 3.1.3 software. Results The traps collected 102 seeds belonging to 10 species of cloud–forest fleshy–fruited plants in seed traps (table 1), confirming the hypothesis that frugivorous

0.0 ± 0.0

0.1 ± 0.1 0.0 ± 0.0

4.6 ± 1.4

1.8 ± 1.0

2.0 ± 0.6 1.8 ± 0.8

birds and bats contribute to seed rain from cloud–forest patches into disturbed pine forests. Moreover, we found 163 saplings of 12 plant species of the same kind as those established in the surface–burned pine forest (table 2). Five species were found in both samplings (Solanum spp., Cornus disciflora, Magnolia iltisana, Persea hintonii and Ilex brandegeana). The most frequent plant species was Solanum spp. in both cases. Distance from undisturbed cloud forest had no significant effect on the overall seed number, but there were more dispersed seeds in site A (F1; 38 = 5.11, P = 0.030; fig. 2). However, significantly more seeds of Solanum spp. were found at 10 m than at further distances from the undisturbed area (F1; 35 = 5.61, P = 0.024). No effects of site or distance were observed in Phytolacca sp. propagules. Finally, more seeds were found at site A for Cornus disciflora (F1; 38 = 5.25, P = 0.028) and Rhamnus hintonii (x21; 38 = 12.48, P < 0.001), but seed distribution was not affected by distance in these species. In the case of saplings, there were no differences in their total abundance between distances, but site A held a significantly higher sapling density than B (F1; 78 = 4.56, P = 0.036). The same result was found in Solanum sp saplings

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Table 2. Cloud forest fleshy–fruited plants with established saplings in surface–burned pine forests adjacent to cloud forest patches. The mean density of saplings (saplings/10 m2) for each distance category (± SE) is shown. We also give information on the type of life form (LF) each species belongs to (H. Herb; S. Shrub; T. Tree; C. Climber), and the frequency of appearance of each species in the seed traps. Tabla 2. Plantas de frutos carnosos de los bosques nubosos con plántulas en pinares afectados por un incendio de superficie y adyacentes a fragmentos de bosques nubosos. Se muestra la densidad media de plántulas (plántulas/10 m2) para cada categoría de distancia (± EE). También se proporciona información sobre el tipo de forma de vida (LF) al que pertenece cada especie (H. Hierba; S. Arbusto; T. Árbol; C. Liana), así como la frecuencia de aparición de cada especie en las trampas de semillas. Species

Mean density (saplings/10 m2)

Frequency

H, S, T

0.48

4.3 ± 2.3 7.8 ± 2.3

4.5 ± 1.4 3.3 ± 1.6

Viburnum hartwegii Benth.

0.21

1.3 ± 0.7 3.3 ± 1.2

1.0 ± 0.5 3.3 ± 1.3

Persea hintonii C. K. Allen

0.13

0.5 ± 0.5 1.5 ± 0.6

3.0 ± 0.7 0.3 ± 0.3

Magnolia iltisana A. Vázquez

0.05

0.5 ± 0.3 0.3 ± 0.3

1.0 ± 0.6 0.3 ± 0.3

S, T

0.03

1.0 ± 0.6 0.0 ± 0.0

0.3 ± 0.3 0.0 ± 0.0

Cornus disciflora DC.

0.02

0.3 ± 0.3 0.0 ± 0.0

0.3 ± 0.3 0.3 ± 0.3

Ilex brandegeana Loes

0.02

0.8 ± 0.4 0.0 ± 0.0

0.0 ± 0.0 0.3 ± 0.3

Symplococarpon purpusii

0.02

0.5 ± 0.5 0.0 ± 0.0

0.3 ± 0.3 0.0 ± 0.0

Miconia sp.

0.01

0.0 ± 0.0 0.3 ± 0.3

0.0 ± 0.0 0.0 ± 0.0

Smilax sp.

0.01

0.3 ± 0.3 0.3 ± 0.3

0.0 ± 0.0 0.0 ± 0.0

Parahtesis villosa Lundell

0.01

0.0 ± 0.0 0.3 ± 0.3

0.0 ± 0.0 0.0 ± 0.0

Prunus serotina Ehrh.

0.01

0.0 ± 0.0 0.3 ± 0.3

0.0 ± 0.0 0.0 ± 0.0

Solanum sp.

Cestrum sp.

10 m

25 m

50 m

100 m

(Brandegee) Kobuski

Total

(F1; 78 = 5.48, P = 0.019). The interaction of distance and site was significant in Viburnum hartwegii, whose saplings were more abundant at 100 m in site A but at 25 m in site B (x23; 38 = 9.09, P = 0.003). No effects of distance and site were observed in Persea hintonii, and finally, Magnolia iltisana showed significantly more saplings in site A (x21; 38 = 11.09, P = 0.001). Discussion Our results show that frugivorous vertebrates play an important role in the recovery of cloud forests in montane areas of WM affected by surface fires. These vertebrates disperse the seeds of a variety of fleshy–fruited broadleaf species from undisturbed cloud forest patches into burned areas located nearby these patches. This action could promote the regeneration of this threatened ecosystem and allow the colonization of new areas. Comparing the list of plant species found in seed rain and sapling surveys, we observed that only five plant species were found in both. The absence as saplings of plant species whose seeds had been

9.3 ± 3.0 13.8 ± 2.7 10.3 ± 1.8 7.5 ± 2.3

dispersed could be due to unsuitable conditions for germination (Schupp, 1993). For instance, pokeweeds (Phytolacca sp.) are early–successional plants that germinate abundantly from seed banks (Hyatt & Casper, 2000) but are displaced by more shade–tolerant species under shade conditions (Elliott et al., 1998). Moreover, the plant functional group can also determine establishment success: climbers may find it difficult to establish themselves because of the lack of an adequate substrate (Vitis blancoi was only present as seed, and Smilax spp. presence was incidental). On the contrary, the absence of seeds of species found as saplings could be caused by interannual variation in the fruit production of some plants (Herrera, 1998); these saplings may have grown from seeds dispersed in previous years or dispersed before the fire and occurring in the seed bank. Overall seed dispersal in the burned habitat did not depend on the distance from the edge of undisturbed forests. This finding may contrast with findings from other studies that found little seed dispersal out of natural forest and into altered habitats, such as early–successional forests or cultivated lands (Ingle, 2003; Estrada–Villegas et al., 2007; Del Castillo &

180

2.5

Saplings/m2

2.0

Seeds/m2

10 9 8 7 6 5 4 3 2 1 0

Rost et al.

1.5 1.0 0.5

0 100 10 Distance (m) Site A

100

Site B

Fig. 2. Differences in seed rain and sapling density between distances from the nearest cloud forest patch and between study sites. Fig. 2. Diferencias de lluvia de semillas y densidad de plántulas en función de la distancia del área de bosque nuboso más cercana y entre los sitios del estudio.

Pérez–Ríos, 2008). However, in our study sites, the surface fire did not change the availability of tall perches (most pines survived fire) needed by many arboreal dispersers, such as birds (Shiels & Walker, 2003). This habitat feature could thus enhance colonization of cloud forest plant species into such areas. Besides, medium–sized and large birds can disperse seeds to great distances, up to hundreds of meters from the mother plant (Jordano et al., 2006; Weir & Corlett, 2007). That could be the case of birds that would disperse large seeds like those of Cornus disciflora and Rhamnus hintonii (whose dispersal was not distance–dependent) since larger frugivores, with a higher gape width, can disperse fruits with larger seeds (Wheelwhright, 1983; Jordano, 1986). Medium–sized mammals could have also contributed to this long–distance seed transport (Willson, 1993), although we did not find scats from mammals in the study sites, probably because the high cover of low vegetation (up to 50 m) made them difficult to find. In turn, this does not seem to apply to bats, which are usually not abundant beyond natural forest edges, even when perches are available (Shields & Walker, 2003). Moreover, small bats like those we found in the study area usually have short seed retention times, and this may prevent them from dispersing seeds over long distances (Shilton et al., 1999). These causes could also explain why we found more seeds of Solanum near the forest edge, since their fruits are one of the few consumed by frugivorous bats in the montane forests of western Mexico (Hernández–Conrique et al., 1997; Íñiguez–Dávalos, 2005). However, Solanum saplings were distributed irrespectively to distance from undisturbed forest, which seems to contradict this explanation. Yet Zuloaga–Aguilar et al. (2010) found that Solanum aphyodendron —which

would probably be one of the Solanum species found as seed or sapling in our study although identity of this genus could not be confirmed with certainty to species level— can tolerate high temperatures (e.g., a fire episode), and that heat shocks induce its germination. Therefore, in the case of Solanum, an already existing seed bank before the fire could be a feasible explanation. Our study also revealed that the habitat configuration at a higher scale has a significant role in the numbers of dispersed seeds and sapling density found in the surface–burned area. The study site that was close to the fire perimeter (A), and therefore to a large patch of cloud forest, showed a higher seed rain and higher sapling densities than the site whose closer seed source was a narrow strip surrounded by the burned area. This site (B) probably showed a much lower amount of fruits and dispersers than the other site. The specific case of Viburnum hartwegii, which showed higher sapling densities at different distances in the two areas, could be explained by the particularly high concentration of seed dispersal in such locations, due to the activity of some dispersal vector. This effect of site on seed rain and shadow means that cloud forest regeneration can be also influenced by the sizes and fragmentation of natural habitat patches after disturbances (Del Castillo & Pérez–Ríos, 2008). Fire has been a historical component of the dynamics of Mexican montane forest ecosystems (Jardel et al., 2004b), but there is now a growing concern about the disruption of the fire regimes due to land–use change, fire suppression and the effects of global climate change (Rodríguez–Trejo, 2008; Flannigan et al., 2009; Liu et al., 2010). In such a scenario, the conservation of cloud forests

Animal Biodiversity and Conservation 38.2 (2015)

surrounded by a matrix of fire–prone pine–oak forests emerges as a priority in western Mexico mountains. As we have shown, regeneration of this threatened forest can be enhanced by the activity of frugivorous birds and bats; this can be either after fire, when they can disperse seeds to long distances from undisturbed seed sources, or before fire, by creating seed banks. Other endozoochorous dispersers (e.g., terrestrial mammals) may also play a role but were not investigated in this study. However, since regeneration can be hampered in disturbed areas that are less connected to large cloud forest patches, we encourage managers to actively preserve these patches from activities that could degrade, fragment or isolate them (e.g., farming and cattle grazing) so as to favour cloud forest colonization of adjacent areas after disturbances such as surface fires. Acknowledgements We thank the staff at Las Joyas Research Station (University of Guadalajara), Ramón Cuevas, Erminio Quiñónez and Rubén Ramírez for their collaboration. The Fondo Mexicano para la Conservación de la Naturaleza A. C. (Project F7–06–024), the project 'Investigación y Conservación de aves en la región de la Sierra de Manantlán (DERN–IMECBIO–CUCSUR–UdG)', and the F. P. U. scholarship program (Spanish Education Ministry) provided funding for this study. We also appreciate the comments of the anonymous reviewers on the draft of this manuscript, which contributed to improve its quality. References Balke, T., Herman, T. M. J., & Bouma, T. J., 2014. Critical transitions in disturbance–driven ecosystems: identifying Windows of Opportunity for recovery. Journal of Ecology, 102(3): 700–708. Carranza, E., 1992. Cornaceae. Flora del Bajío y de regiones adyacentes. Instituto de Ecología A.C., Pátzcuaro. Challenger, A., 1998. Utilización y conservación de los ecosistemas terrestres de México. Universidad Nacional Autónoma de México, México DF. Chazdon, R. L., 2003. Tropical forest recovery: legacies of human impact and natural disturbances. Perspectives in Plant Ecology Evolution and Systematics, 6: 51–71. Clark, C. J., Poulsen, J. R., Bolker, B. M., Connor, E. F. & Parker, V. T., 2005. Comparative seed shadows of bird–, monkey–, and wind–dispersed trees. Ecology, 86: 2684–2694. Cuevas–Guzmán, R., Koch, S., García, E., Núñez, N. M. & Jardel, E. J., 2004. Flora vascular de la Estación Científica Las Joyas. In: Flora y vegetación de la Estación Científica Las Joyas: 117–176 (R. Cuevas–Guzmán & E. J. Jardel, Eds.). Universidad de Guadalajara, Guadalajara. Del Castillo, R. F. & Pérez–Ríos, M. A., 2008. Changes in seed rain during secondary succession in

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Spring diet of the pine marten in Sardinia, Italy M. Lombardini, M. Murru, A. Repossi, C. E. Cinerari, A. Vidus Rosin, L. Mazzoleni & A. Meriggi

Lombardini, M., Murru, M., Repossi, A., Cinerari, C. E., Vidus Rosin, A., Mazzoleni, L. & Meriggi, A., 2015. Spring diet of the pine marten in Sardinia, Italy. Animal Biodiversity and Conservation, 38.2: 183–190. Abstract Spring diet of the pine marten in Sardinia, Italy.— Knowledge of a carnivore’s foraging behaviour is essential to understand its ecology. The pine marten Martes martes is commonly defined as an opportunistic generalist predator since its diet in Europe includes mammals, fruits, birds and invertebrates. Based on faecal analyses, we evaluated spring feeding habits and trophic niche breadth of the pine marten in a Mediterranean landscape on the island of Sardinia (Central Italy). The most important source of food for the pine marten was small mammals (mainly Apodemus sylvaticus, Mus musculus and Eliomys quercinus), accounting for 52% of the percent mean volume. Important secondary foods were invertebrates (mainly beetles and insect larvae) and birds (almost exclusively passerines), whereas large mammals, lagomorphs, reptiles and fruits made little con� tribution to the diet. The diet composition and the Levins’ index value suggest that the pine marten in Sardinia behaves as a facultative specialist predator, with a specialization towards small mammals. Key words: Martes martes, Foraging behaviour, Scat analysis, Trophic niche breadth Resumen Alimentación primaveral de la marta en Cerdeña, Italia.— El conocimiento del comportamiento de alimentación de un carnívoro es esencial para entender su ecología. La marta Martes martes se define comúnmente como un depredador generalista oportunista, porque su dieta en Europa incluye mamíferos, frutas, aves y inverte� brados. A partir del análisis de las heces, hemos descrito los hábitos alimenticios en primavera y la amplitud del nicho trófico de la marta en ambiente mediterráneo en Cerdeña (Italia central). Los pequeños mamíferos (sobre todo Apodemus sylvaticus, Mus musculus y Eliomys quercinus) representan la fuente más importante de alimentación de la marta, ya que constituyen aproximadamente el 52% del volumen medio. Otra fuente importante de alimentos secundarios la constituyen los invertebrados (especialmente escarabajos y larvas de insectos) y las aves (paseriformes casi exclusivamente), mientras que los grandes mamíferos, los lagomorfos, los reptiles y las frutas están poco representados en la dieta. La composición de la dieta y el valor del índice de Levins indican que la marta en Cerdeña es un depredador especialista facultativo, con una especialización en la depredación de micromamíferos. Palabras clave: Martes martes, Comportamiento alimenticio, Análisis de heces, Amplitud de nicho trófico Received: 30 IV 15; Conditional acceptance: 9 VII 15; Final acceptance: 24 VII 15 Marco Lombardini, Marco Murru, Ambra Repossi, Claudia E. Cinerari, Anna Vidus Rosin, Linda Mazzoleni & Alberto Meriggi, Dept. of Earth and Environmental Sciences, Univ. of Pavia, Via Ferrata 1, 27100 Pavia, Italy. Corresponding author: Marco Lombardini. E–mail: zarc00@yahoo.it

ISSN: 1578–665 X eISSN: 2014–928 X

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Introduction

Material and methods

The pine marten (Martes martes L., 1758) is a me� dium–sized mustelid found throughout much of Europe and northern and central Asia, from northern Portugal to western Siberia (Ruiz–González et al., 2013). In Italy, the species has a fragmented distribution in the forested areas of the peninsula, but it has recently been detected in the western part of the River Po plain, an intensively cultivated area (Proulx et al., 2004; Balestrieri et al., 2010). Insular populations also occur in Sardinia, Sicily and Elba (Masseti, 1995; Angelici et al., 2009). In Sardinia, the pine marten is widespread, being present both in the northern and southern parts of the island (Murgia et al., 1995). The species might have been introduced on the island during Roman times or even a little earlier, but the exact period of introduction is not clear (Vigne, 1992; Masseti, 1995). At the beginning of the twentieth century, the Sardinian pine marten was described as a Martes martes latinorum subspecies. It was distinguished from the nominal species by the leather–yellow patch on the throat and by a lighter dominating colour; it is the same size as a Italian form of the species, except for a slightly longer tail (Murgia et al., 1995). Recently, Colli et al. (2011) described the genetic variability of Sardinian pine martens and differences between insular and Italian population. They reported two distinct clusters, corresponding to Sardinia and mainland Italy, and hypothesized the distinctiveness of the Sardinian population. Knowledge of a carnivore’s foraging behaviour is essential to understand its ecology, and to elucidate potential competitive interactions and impacts on prey populations (Litvaitis, 2000; Caryl et al., 2012). The pine marten is commonly defined as an opportunistic generalist predator. Its diet in Europe is very varied, including small and large mammals, fruits, birds and invertebrates, but its main year–round prey is gener� ally small mammals (Helldin, 2000; Zalewski, 2004; Rosellini et al., 2008a; Balestrieri et al., 2011; Caryl et al., 2012). In the Mediterranean region, fruit, plant material and insects are important components of the diet, being consumed more frequently than in central and northern Europe (De Marinis & Masseti, 1995; Zalewski, 2004). F�� eeding habits of �� pine martens on Mediterranean is� lands are poorly known. Research has been carried out only on Mallorca, Minorca and Elba islands (De Marinis & Masseti, 1993, 1995; Clevenger, 1993, 1995, 1996). In these insular environments, the pine marten shows some peculiarities in its diet: the absence of voles, for example, the main mammalian prey throughout Europe, has determined a shift towards mice (Apodemus sylvaticus, Mus sp. and Rattus sp.) (De Marinis & Masseti, 1995; Zalewski, 2004), and in the Balearic islands, the species shows a high level of frugivory (Clevenger, 1995, 1996) and a wider trophic niche breadth than continental populations (Clevenger, 1993). In this paper, we provide the first description of the spring diet and trophic niche breadth for the pine mar� ten in Sardinia and compare our results with those of Clevenger (1995) on the island of Mallorca.

Study area The study was carried out in the province of Olbia– Tempio (NE Sardinia, Central Italy), which extends for 3,404 km2 and has an altitude ranging from sea level to 1,359 m a.s.l. (Mount Limbara) (fig. 1A). The climate is Mediterranean, with the most abundant rainfalls occurring in December and the highest temperatures occurring in July. Vegetation is typically Mediterranean; the area is dominated by garrigue and low maquis with Phillyrea sp., lentisk (Pistacia lentiscus), cistus (Cistus spp.) and heather (Erica arborea), and deciduous forests, mostly including oak (Quercus ilex, Q. suber). Inland flat areas are characterized by arable lands and pastures. Other mesocarnivores in the study area besides the pine marten are the red fox (Vulpes vulpes), the weasel (Mustela nivalis) and the wildcat (Felis silvestris). Lago� morphs present are the Sardinian hare (Lepus capensis mediterraneus) and the wild rabbit (Oryctolagus cuniculus). The community of small mammals is composed of eight species: the hedgehog (Erinaceus europaeus), the Etruscan shrew (Suncus etruscus), the North African white–toothed shrew (Crocidura pachyura), the wood mouse (Apodemus sylvaticus), the house mouse (Mus musculus), the garden dormouse (Eliomys quercinus) and rats (Rattus rattus and Rattus norvegicus). A total of 101 species of birds nest in the study area (Trainito, 2009). Field surveys Diet composition of the pine marten was studied by scat analysis. Scats were collected in spring (March–May) 2012 and 2013 along 80 linear transects (total length = 115.5 km, mean ± standard deviation SD = 1.4 ± 0.5 km, min. = 0.4 km, max. = 2.7 km), distributed in 34 sampling areas (15 protected areas and 19 hunting preserves) selected within the whole study area and representative of all habitat types (fig. 1B and appendix 1). Each itinerary was walked twice a season, with a 25–30 day interval between the two surveys. Faecal samples were assigned to the pine marten according to size, shape and odour (Lanszki et al., 2007; Rosellini et al., 2008b; Barrull et al., 2014). Moreover, we evaluated the proximity of scats to marten tracks and foraging signs. Scats were individually bagged, labelled with number and date of collection, and stored in a freezer at –20°C until analysis. Diet analysis Faecal samples were washed through two sieves of 0.5 and 0.1 mm mesh, and food remains were inspec� ted to estimate the total numbers of each kind of food. Prey items were categorized into eight food classes: small mammals (rodents and insectivores), lagomorphs, large mammals (wild and domestic ungu� lates eaten as carrion), birds, reptiles, invertebrates, fruit (berries and large fruits) and other (garbage or non–natural foods). Prey remains were identified to the lowest taxonomic level possible.

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33 5

Ligurian Sea

29 18

26 24 28 31

Olbia

4 0

10 1 32

27 16 23

Tyrrhenian Sea

Tempio 20

20 km

Fig. 1. Location of the study area (Province of Olbia–Tempio, Sardinia, Central Italy) (A) and location of the sampling sites (B). For the identification of sampling areas, see appendix 1: n Sampling areas with presence of marten scats; o Sampling areas with no marten scats. Fig. 1. Localización de la zona de estudio (provincia de Olbia–Tempio, Cerdeña, Italia central) (A) y localización de las áreas de muestreo (B). Para la identificación de las áreas de muestreo, véase el apéndice 1: n Áreas con heces de marta; o Áreas sin heces de marta.

Mammal hairs were compared at × 10 and × 40 magnifications with the keys of Debrot et al. (1982), Teerink (1991), De Marinis & Agnelli (1993) and De Marinis & Asprea (2006). Bird feathers were identi� fied with reference to MicrolabNW Photomicrograph Gallery (http://microlabgallery.com/Feathers.aspx). Wild or cultivated fruits (seeds) were identified with reference to Ferrari & Medici (2003) and using per� sonal collections. Diet composition was expressed in two ways: percent frequency of occurrence (%FO = number of faecal samples containing a specific food item/total number of faecal samples × 100) and percent mean volume (%mV = total estimated volume of each food item as ingested/total number of faecal samples). The percent volume of each food was estimated visually (Clevenger, 1995). Based on percent mean volumes, we calculated the trophic niche breadth in accordance with Levins (Krebs, 1989): B = 1/Σpi2 where pi is the relative frequency of the ith food item. We standardized the B index across food items: BA = (B – 1)/(n – 1) This measure ranges from 0 to 1, with higher values indicating a broader dietary niche.

Finally, we compared our results with those obtai� ned in spring by Clevenger (1995) on Mallorca. We evaluated differences in the frequency of occurrence of seven prey categories (small mammals, lagomor� phs, large mammals, birds, reptiles, invertebrates and fruit) using a x2–test for contingency tables (Fedriani et al., 1999). Results A total of 87 scats were used for dietary analysis. Scats were found in 23 of the 34 sampling areas (67.6%, fig. 1B), but about half of the scats were collected in four sampling areas: Sorilis, Monte Olia, Monte Limbara and Filigosu. The marten diet was dominated by small mammals, which represented more than half of the total percent mean volume. Three species occurred in the diet regularly: the wood mouse, the house mouse and the garden dormouse. The only insectivore species eaten by martens was the North African white–toothed shrew (table 1). Invertebrates and avian preys were important secondary foods; invertebrates comprised mainly beetles (Coleoptera) and insect larvae, while birds included predominantly passerines (table 1). Large mammals, lagomorphs, fruits and reptiles made little contribution to the pine marten diet. Both wild and domestic ungulates were

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Table 1. Diet composition of the pine marten revealed through faecal analysis (n = 87): Fr. Frequency of occurrence (%); Pm. Percent mean volume (%). Tabla 1. Composición de la dieta de la marta definida a partir del análisis de las heces (n = 87): Fr. Frecuencia de aparición; Pm. Porcentaje de volumen medio. Food items

Fr (%)

Pm (%)

Small mammals

58.6

Rodents Apodemus sylvaticus

Food items

Fr (%)

Pm (%)

52.2

Birds

17.2

14.0

55.2

49.4

Passeriformes

14.8

12.2

24.2

21.5

Galliformes

1.2

0.6

Mus musculus

16.1

14.5

Falconiformes

Eliomys quercinus

10.3

9.5

Invertebrates

33.3

16.8

Rattus sp.

4.6

3.9

Coleoptera

19.5

9.9

Insectivores

3.5

2.8

Insect larvae

11.5

5.1

1.1

0.5

Crocidura pachyura

3.5

2.8

Hymenoptera

Lagomorphs

1.1

Undetermined

3.5

1.3

Oryctolagus cuniculus

1.1

Fruit

10.3

4.6

Large mammals

10.3

6.9

Juniperus spp.

4.6

1.7

1.1

Ovis aries

3.4

Myrtus communis

Bos taurus

2.3

Cistus spp.

1.1

0.6

Sus scrofa

4.6

1.2

Undetermined

3.5

1.2

Undetermined mammals

4.6

3.7

Reptiles

1.1

0.4

Other

1.1

0.3

consumed, with a slight predominance of domestic animals. Wild rabbits were taken very occasionally, and no Sardinian hares were identified in the scats. Fruits included myrtle (Myrtus communis), junipers (Juniperus spp.) and Cistus spp. berries (table 1). Other foods (i.e. garbage) were scarcely present in the diet. The value of the standardized B index of trophic niche breadth was 0.29. The comparison between Sardinian and Mallorcan feeding habits highlighted some significant differences; Mallorcan martens were more frugivorous (x2 = 67.63, df = 1, p < 0.001) and preyed upon more reptiles (x2 = 6.76, df = 1, p = 0.009) than the Sardinian martens, but consumed fewer small (x2 = 7.25, df = 1, p = 0.007) and large mammals (x2 = 4.89, df = 1, p = 0.03). The consumption of lagomorphs (x2 = 0.39, df = 1, p = 0.53), birds (x2 = 0.26, df = 1, p = 0.61) and invertebrates (x2 = 0.41, df = 1, p = 0.52) did not differ significantly between the two islands (table 2). Discussion Our results show that small rodents were the most important food resource for the pine marten in Nor� th–Eastern Sardinia. They were found to be the main prey in spring, in agreement with the findings of other

authors across the European distribution range of the species (e.g. Jędrzejewski et al., 1993; Ruiz–Olmo & López–Martín, 1996; Helldin, 2000; Lanszki et al., 2007; Rosellini et al., 2008a; Caryl et al., 2012). The predator mainly focused on Apodemus sylvaticus, an important resource for this and other medium–sized carnivores in the study area, such as the red fox (Meriggi et al., 2013). Two factors could explain the importance of the wood mouse in the marten diet; its abundance and the similar habitat requirements of both predator and prey species. In the study area, the pine marten positively selects woodlands and shrublands (Lombardini et al., 2015). Similarly, in the Mediterranean region, the wood mouse is basically found in cork oak woodlands and areas covered by trees and shrubs (Cagnin et al., 1998; Rosalino et al., 2011). Amori et al. (2014) indicate that even in Sardinian woodlands, Apodemus sylvaticus is widespread and locally abundant. Furthermore, the exploitation of forested areas by the wood mouse in Sardinia could be favoured by the absence of the forest–dweller Apodemus flavicollis, as reported by Sarà & Casamento (1993) in Sicily. The diet of the pine marten also comprised avian prey and invertebrates. Martens commonly feed on small birds in Europe (Posłuszny et al., 2007; Bales� trieri et al., 2011; Zhou et al., 2011; Caryl et al., 2012).

Animal Biodiversity and Conservation 38.2 (2015)

Table 2. Frequency of occurrence of feeding categories in the spring diet of the pine marten in Sardinia (S, this study) and Mallorca (M, Clevenger, 1995). Tabla 2. Frecuencia de aparición de categorías de alimentos en la dieta primaveral de la marta en Cerdeña (S, este estudio) y en Mallorca (M, Clevenger, 1995). Food items

S (n = 87)

M (n = 130)

58.6

40.0

Lagomorphs

1.1

2.3

Large mammals

10.3

3.0

Birds

17.2

20.0

Invertebrates

33.3

29.2

Plant material

10.3

66.6

Reptiles

1.1

10.0

Other

1.1

1.5

Small mammals

The importance of birds in the spring diet is probably due to their increased vulnerability to predation during hatching and fledgling time (De Marinis & Masseti, 1995) and to the abundance of small mammals be� ing at its lowest in spring (Jędrzejewski et al., 1993; Lanszki et al., 2007; Rosellini et al., 2008a), forcing the pine marten to exploit alternative food, such as birds (Lanszki et al., 2007; Wijsman, 2012). The pine marten preyed mostly on passerines, whereas the predation exerted on game birds was scarce and occasional. The importance of invertebrates as a secondary food resource is typical of lower latitudes. In southern Europe, in fact, characterized by warmer and more stable climates, insects are more abundant than in central and Northern Europe (Ruiz–Olmo & López–Martín, 1996; Zhou et al., 2011). The comparison of the feeding habits of Sardinian and Mallorcan martens stresses the differences bet� ween these two insular populations. In Mallorca, the pine marten exhibited a generalist diet, with a high con� sumption of plant material, while in Sardinia it showed a specialization towards small mammals. In the Balearic islands, martens often eat fruit from orchards close to farmland or disturbed habitats (Clevenger, 1995). The lack of predators, in fact, may favour the exploitation of open habitats for foraging (Clevenger, 1994). On the other hand, in Sardinia, the pine marten coexists with three potential predators: the red fox, the wildcat, and the golden eagle Aquila chrysaetos (Lindström et al., 1995; Pedrini & Sergio, 2002; Moleón & Gil–Sánchez, 2003). For this reason, the species probably avoids open habitats in North–Eastern Sardinia (Lombardini et al., 2015). Diet results support this hypothesis,

187

indicating the prevailing consumption of wild berries, such as myrtle, junipers and Cistus berries, growing in the maquis. Concerning plant foods, Clevenger (1996) indicated a predominance of the carob Ceratonia siliqua, found to be absent in the diet of the species in Sardinia, probably due to the rarity of the carob in the province of Olbia–Tempio (Trainito, 2009). Both in Sardinia and in Mallorca, Apodemus sylvaticus was the main rodent prey in spring, and consumption was simi� lar on both islands (21% and 18% of the total percent mean volume, respectively). The significant difference in the consumption of small mammals highlighted by the analyses is linked to the higher presence in our sample of secondary rodent preys (Mus musculus and Eliomys quercinus), scarcely eaten by Mallorcan martens (Clevenger, 1995). The relatively low Levins' index value suggests that the pine marten in Sardinia adopt an intermediate feeding strategy between that of an opportunist and that of a specialist predator (i.e. a facultative specialist in the predation of small mammals). A similar situation occurs in north–western Spain, where the pine marten shows no reduction in its preference for small mammals even in the seasons when they are scarce (Rosellini et al., 2008a). Future work should test this hypothesis proposed to describe the feeding habits of the pine marten in Sardinia and should analyse the diet of the pine marten all year–round, to evaluate the existence of seasonal feeding patterns. Acknowledgements This study was financially supported by the province administration of Olbia–Tempio. We would like to thank Valentina Dagradi, Olivia Dondina, Nicola Floris, Cami� lle Imbert and Luca Nelli for their help during fieldwork. Giuseppe Lucia kindly revised the Spanish language. Two anonymous reviewers improved our manuscript. References Amori, G., Luiselli, L., Milana, G. & Casula, P., 2014. Distribuzione, diversità e abbondanza di micromammiferi associati ad habitat forestali in Sardegna. Report – Ente Foreste della Regione Sardegna. Angelici, F. M., Laurenti, A. & Nappi, A., 2009. A checklist of the mammals of small Italian islands. Hystrix, 20: 3–27. Balestrieri, A., Remonti, L., Ruiz–González, A., Gómez–Moliner, B. J., Vergara, M. & Prigioni, C., 2010. Range expansion of the pine marten (Martes martes) in an agricultural landscape matrix (NW Italy). Mammalian Biology, 75: 412–419. Balestrieri, A., Remonti, L., Ruiz–González, A., Vergara, M., Capelli, E., Gómez–Moliner, B. J. & Prigioni, C., 2011. Food habits of genetically identified pine marten (Martes martes) expanding in agricultural lowlands (NW Italy). Acta Theriologica, 56: 199–207.

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Appendix 1. Environmental characteristics (% of land use categories) of sampling areas selected for scat surveys in the province of Olbia–Tempio: S. Surface (in Ha); P. Protected: F. Forest; Gg. Garrigue; Gs. Grassland; Ra. Rocky areas; Wb. Water bodies; Al. Arable lands; Ua. Urban areas. Apéndice 1. Características ambientales (% del uso del suelo) de las áreas de muestreo seleccionadas para la recolección de las heces en la provincia de Olbia–Tempio: S. Superficie (en Ha); P. Protegida; F. Bosque; Gg. Garriga; Gs. Pradera; Ra. Áreas rocosas; Wb. Agua; Al. Tierras arables; Ua. Áreas urbanas.

Sampling areas

S (Ha)

Bolostiu (1)

797

Yes

17.3

51.3

3.4

25.8

–

2.2

C. Coda (2)

454

Yes

20.6

14.1

21.4

23.2

3.0

3.5

14.2

Coluccia (3)

490

Yes

11.4

60.3

2.4

4.1

18.2

1.4

2.2

2,164

Yes

25.4

59.6

4.9

7.1

0.3

2.4

Conchedda (4) Costa P. (5)

673

Yes

15.6

53.1

–

7.7

–

0.1

23.5

Figari (6)

1,524

Yes

0.8

77.7

0.3

8.5

0.2

0.6

11.9

Filigosu (7)

4,461

Yes

34.7

44.3

2.1

13.3

0.2

2.9

2.5

M. Limbara (8)

3,900

Yes

15.6

54.2

1.3

26.7

0.1

0.5

1.6

231

Yes

14.1

2.4

65.9

–

10.2

6.0

1.4

M. Olia (10)

2,171

Yes

46.3

30.8

1.3

18.8

0.1

1.0

1.7

P. S. Paolo (11)

1,181

Yes

1.5

64.9

11.6

5.6

–

6.0

10.4

456

Yes

8.8

38.3

25.6

–

19.4

7.9

Saloni (13)

418

Yes

8.0

16.9

52.4

–

5.7

7.7

9.3

Sorilis (14)

1,331

Yes

30.6

60.4

3.5

2.8

0.5

–

2.2

Terranova (15)

2,249

Yes

50.0

39.0

3.4

1.7

0.1

3.1

2.7

Campu N. (16)

654

22.4

2.2

45.7

–

27.3

2.4

Canaili (17)

587

43.9

34.0

9.1

1.0

–

6.3

5.7

Frassiccia (18)

827

22.6

50.8

3.0

17.2

–

2.5

3.9

L’Agnata (19)

551

42.1

49.0

2.0

2.7

–

4.2

Lanzinosa (20)

700

35.7

48.4

3.8

7.5

–

1.8

2.8

Li Parisi (21)

1,674

6.7

23.1

57.4

4.7

–

5.3

2.8

M. Littu (22)

685

4.9

26.2

49.2

0.7

–

10.9

8.1

Locheri (23)

1,034

22.8

14.1

43.8

1.8

–

15.0

2.5

Lu Naracu (24)

649

21.4

13.4

42.8

1.9

0.1

13.8

6.6

Muddizza (25)

1,179

19.7

48,1

19.7

6.5

–

3.2

2.8

Muntagna (26)

978

5.4

35.6

3.4

11.1

0.3

41.6

2.6

Nulvara (27)

824

42.7

23.2

16.3

1.3

–

13.7

2.8

PFG (28)

625

5.4

28.5

53.1

8.0

–

0.8

4.2

M. Russu (29)

1,066

9.5

49.5

32.2

–

6.4

2.4

Sa Matta (30)

2,828

26.4

19.6

44.0

3.4

–

3.6

3.0

S. Biagio (31)

653

22.4

46.5

25.1

0.3

–

2.3

4.6

Liscia (9)

M. Rotu (12)

Ulchi B.C. (32)

895

10.9

54.4

8.3

21.7

0.1

0.3

4.3

1,692

4.0

43.6

6.2

0.9

–

42.6

2.7

Zurria (34)

1,149

14.3

34.6

23.8

8.9

0.1

14.4

3.9

Total

41,750

22.4

41.6

15.7

9.2

0.5

6.6

4.0

Province

340,418

21.2

38.1

13.2

6.0

0.8

15.2

5.5

Vignola (33)

Animal Biodiversity and Conservation 38.2 (2015)

191

Un nuevo Trechus (Coleoptera, Carabidae, Trechini) hipogeo de la Sierra de Parapanda (Andalucía, España): taxonomía, sistemática y biología V. M. Ortuño & P. Barranco

Ortuño, V. M. & Barranco, P., 2015. Un nuevo Trechus (Coleoptera, Carabidae, Trechini) hipogeo de la Sierra de Parapanda (Andalucía, España): taxonomía, sistemática y biología. Animal Biodiversity and Conservation, 38.2: 191–206. Abstract A new hypogean Trechus (Coleoptera, Carabidae, Trechini) from Sierra de Parapanda (Andalucía, España): taxonomy, systematics and biology.— Sampling of arthropod fauna by pitfall traps in the cavern 'Sima de San Rafael' in Íllora (Granada, Spain) has revealed a new carabid beetle species, Trechus parapandus n. sp., with remarkable troglobiomorphic characteristics: eyes visible only as scars, depigmentation, and elongation of antennae and legs. In consonance with these characteristics, this new species, Trechus parapandus n. sp. is absent in the upper region of the cave. The species belongs to the Trechus fulvus species group (that has five species in Andalusia) according to the characteristics of both male and female genitalia. Study of the fauna in the cave suggests that Collembola might be the prey of this new species since they are the most abundant group and have a coincidental phenology. A key for the 11 Trechus species present in Andalusia is provided. Key words: Trechus, Hypogean fauna, Taxonomy, Systematics, Biology, Iberian peninsula Resumen Un nuevo Trechus (Coleoptera, Carabidae, Trechini) hipogeo de la Sierra de Parapanda (Andalucía, España): taxonomía, sistemática y biología.— El muestreo de la artropodofauna de la Sima de San Rafael en Íllora, Gra� nada, mediante trampas de caída, ha permitido el descubrimiento de un nuevo coleóptero carábido, Trechus parapandus sp. n., con notables caracteres troglobiomórficos: reducción ocular quedando relegada a una cicatriz, despigmentación y un ligero alargamiento de apéndices. Para la ubicación de este nuevo taxón dentro del conjunto de especies del género presentes en Andalucía, se ha confeccionado una clave dicotómica que permite identificar las 11 especies del género presentes en este ámbito territorial. Trechus parapandus sp. n. se encuadra dentro del grupo Trechus fulvus debido a la estructura de su genitalia en ambos sexos, el cual se constituye ahora por cinco especies en Andalucía. El estudio faunístico de la sima, sugiere que el grupo presa de esta nueva especie podrían ser los colémbolos, ya que además de su abundancia presentan una distribución temporal coincidente. Trechus parapandus sp. n. se distribuye por la cavidad de acuerdo a su carácter hipogeo, estando ausente en la zona más superficial de la misma. Palabras clave: Trechus, Fauna hipogea, Taxonomía, Sistemática, Biología, Península ibérica Received: 17 VI 15; Conditional acceptance: 23 VII 15; Final acceptance: 24 VII 15 V. M. Ortuño, Grupo de Investigación de Biología del Suelo y de los Ecosistemas Subterráneos, Depto. de Ciencias de la Vida, Fac. de Biología, Ciencias Ambientales y Química, Univ. de Alcalá, A. P. 20, Campus Universitario, E–28805 Alcalá de Henares, Madrid, Spain.– P. Barranco, Depto. de Biología y Geología, Cite II–B, Univ. de Almería, E–04120 Almería, Spain. Corresponding author: Vicente M. Ortuño. E–mail: vicente.ortuno@uah.es

ISSN: 1578–665 X eISSN: 2014–928 X

Ortuño & Barranco

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Introducción El género Trechus Clairville, 1806 tiene una ele� vada diversidad específica, habiéndose descrito más de 600 especies (Moravec et al., 2003). Su distribución es mayoritariamente paleártica si bien se conocen alrededor de 50 especies en la región neártica (Larochelle & Larivière, 2003) y algo más de 25 especies en el este de África (Ortuño & No� voa, 2011). Según el último catálogo de carábidos ibéricos (Serrano, 2013) y considerando algunas enmiendas y adendas (Ortuño & Arribas, 2010; Or� tuño & Barranco, 2013; Ortuño et al., 2014; Toribio, 2014, 2015; Fresneda et al., 2015) se conocen 62 especies de este género en el ámbito íbero–balear, de las que 10 están citadas de Andalucía (ver tabla 1): algunas de estas son claramente euritópicas por lo que exhiben una amplia distribución, como es el caso de Trechus quadristriatus (Schrank, 1781) y Trechus obstusus (Erichson, 1837); otras, endémicas y restringidas a pequeñas áreas (microendemismos sensu Rapoport, 1975) como sucede con la especie orófila Trechus planipennis Rosenhauer, 1856, o con las especies hipogeas Trechus breuili Jeannel, 1913 y Trechus lencinai (Mateu & Ortuño, 2006) (Ortuño, 2008; Ortuño & Barranco, 2013). Andalucía, además cuenta con algunas especies de Trechus

epigeos que muestran hábitos troglófilos (hipogeos facultativos) (ver tabla 1). En el complejo proceso de evolución de los há� bitats subterráneos, es un hecho constatado que, especies epigeas se instalan en estos espacios afóticos prosperando, en muchos casos, poblacio� nes hipogeas que encuentran serias limitaciones para mantener el flujo génico con sus congéneres epigeos. Estos hechos sumados a los factores selectivos que imponen estos singulares hábitats, promueven la divergencia específica y la evolución troglobiomorfa (Christiansen, 1985; Hüppop, 1985; Howarth, 1987; Kane & Culver, 1991; Dethier & Hubart, 2005; entre otros). Ello ha conducido a que organismos geófilos, higrófilos, y en gran medida lucífugos, como son los Trechus, hayan colonizado con éxito el medio subterráneo, dando como resul� tado un elevado número de especies hipogeas con diferente grado de troglobiomorfismo. En el ámbito ibérico son 23 las especies de Trechus que son exclusivas de hábitats hipogeos, lo que supone un alto porcentaje (37%) con respecto al total de especies conocidas. El objetivo de este trabajo es dar a conocer una nueva especie hipogea de Trechus procedente de la Sierra de Parapanda, y discutir su relación taxonómica con otras especies ibéricas.

Tabla 1. Especies de Trechus cuya presencia es conocida en Andalucía. Table 1. Trechus species known in Andalucia.

Hábitat Especie

Corotipo

Corología ibérica

Trechus quadristriatus (Schrank, 1781)

Paleártico

Toda la península

Trechus rufulus Dejean, 1831

Mediterráneo occidental

Por confirmar de Andalucía

Epigeo

y Valencia (citas imprecisas)

Trechus obtusus Erichson, 1837

Europeo–mediterráneo

Toda la península

(Holártico por acción antrópica)

Trechus planipennis Rosenhauer, 1856

Bético (endemismo ibérico) Sierra Nevada

Trechus lallemantii Fairmaire, 1859

Mediterráneo meridional

Punta meridional de Andalucía

Trechus tingitanus Putzeys, 1870

Bético–rifeño

Campo de Gibraltar

Epigeo/hipogeo Trechus fulvus Dejean, 1831

Europeo occidental

Gran parte de la penínsla

Trechus diecki Putzeys, 1870

Bético–mauritánico

Andalucía

Trechus breuili Jeannel, 1913

Bético (endemismo ibérico)

Sierra de Ronda (Málaga)

Trechus lencinai (Mateu & Ortuño, 2006)

Bético (endemismo ibérico)

Hipogeo Sierras de Segura (Jaén) y

Alcaraz (Albacete)

Trechus parapandus sp. n.

Sierra de Parapanda (Granada)

Bético (endemismo ibérico)

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Material y métodos La Sierra de Parapanda, con cotas que oscilan en� tre los 420 y los 1.608 m s.n.m., es una alineación montañosa perteneciente al Sistema Subbético granadino, ubicada en el término municipal de Íllora, provincia de Granada. Parte de esta sierra está coronada por una meseta desnuda en donde predomina el lapiaz y, por toda ella se conocen importantes cavidades subterráneas excavadas en calizas blancas y dolomías, tales como la Sima de los Escorpiones, con más de 100 m de desnivel, o la Sima de San Rafael (Vera Torres et al., 2014). La Sima de los Escorpiones (UTM 30SVG18192969 y a una altitud de 1.575 m s.n.m.) con un desarrollo de 309 m y un desnivel de 103 m, es la cavidad más espectacular de este macizo montañoso, rica en formaciones parietales, presentando bastantes fracturaciones tectónicas sin consolidar en las zo� nas inferiores. En épocas de lluvias hay un intenso goteo en sus zonas profundas, encontrándose las paredes, a partir del tercer pozo, muy húmedas (Santaella Alba & Moreno Espigares, 2006). La

Sima de San Rafael (UTM 30SVG18992897 y a una altitud de 1.305 m s.n.m.) tiene un desarrollo topo� grafiado de 485 m y un desnivel de 54 m. Se abre a consecuencia de varias fracturas interconectadas por una sucesión de pequeños pozos, con recintos cubiertos por grandes bloques que dividen diferentes niveles. En general es muy laberíntica con pasos estrechos y salas o recintos de mediano tamaño. En la cavidad existe una colonia de murciélagos protegida (CMA, 2011). Ambas cavidades fueron prospectadas, si bien sólo la Sima de San Rafael fue muestreada de for� ma sistemática, mediante trampeo con trampas de caída, practicándose cuatro colectas estacionales, desde noviembre de 2012 hasta julio de 2013. Para ello se instalaron 20 trampas de caída, distribuidas por toda la cavidad (fig. 1), enterradas a ras en el sustrato y llenas, hasta la mitad de su capacidad, con una solución compuesta por 2/3 de propilen� glicol y 1/3 de cerveza, y cebadas con sobrasada. Las trampas permanecieron activas alrededor de un mes en cada una de las estaciones del año (10 XI 2012/08 XII 2012; 26 I 2013/09 III 2013;

Sima de San Rafael – GR–232 G.E.G.

± 0

N.M. 10

Proyección N 45º E

–37 m

40 50 –54 m

20 m

–51 m

± 0

2o nivel

3er nivel

Sima de San Rafael – GR–232 N.M.

20 m

Fig. 1. Topografía de la Sima de San Rafael, perfil (A) y planta (B). Ubicación de las trampas de caída (numeradas). Fig. 1. Topography of Sima de San Rafael, profile (A) and floor (B). Location of the pitfall traps (numbered).

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Tabla 2: Inventario de grupos taxonómicos acompañantes de los Trechus colectados en la Sierra de Parapanda, Íllora (Granada, España). Se indica el número total de capturas y porcentaje que representan. Table 2: Checklist of taxonomic groups found with the Trechus collected from Sierra de Parapanda, Íllora (Granada, Spain) and total number of catches and percentages they represent.

Collembola

4.045 37,67

Entomobryidae

138 1,29

Sminthuridae

564 5,25

164 1,53

Onychiuridae Carabidae

Laemostenus sp.

62 0,58

Trechus parapandus sp. n. Staphylinidae

Hymenoptera

Isotomidae

Coleoptera

Parasitica, 1 sp. Psocoptera

3 0,03

Psyllipsocidae Psyllipsocus ramburii 13 0,12 Trichoptera, 1 sp.

3 0,03

Sifonaptera, 1 sp.

4 0,04

Pseudoscorpiones

Neobisiidae Neobisium baenai Acari

Aleocharinae Atheta sp.

9 0,08

Formicidae, 1 sp.

210 1,96

Ac. Trombidiformes, 1 sp.

1.070 9,96

45 0,42

60 0,56

Tachyporinae Sepedophilus sp. 22 0,20

Ac. Oribatida, varias especies 173 1,61

Staphylininae Microsaurus sp.

1 0,01

Ac. Gamasida, varias especies 632 5,89

Pselaphinae Bryaxis sp.

1 0,01

Ac. Rhagidiidae, 1 sp.

Cholevidae

1 0,01 6 0,06

Cryptophagidae Cryptophagus sp. 5 0,05 Diptera

Phoridae, 1 sp.

Anthomyiidae, 1 sp.

47 0,44

Beronium laemostenis

593 5,52

Latridiidae Corticaria sp.

Catops andalusicus

Neothrombidiidae,

Speonemadus angusticollis

3 0,03

16 0,15

Aranea: varias especies

Ixodidae: 1 sp.

19 0,18

Opilionida: 1 sp.

1 0,01

Isopoda: Porcellio sp.

1.141 10,62

935 8,71

Larvas Cholevidae

159 1,48

Larvas Staphylinidae

67 0,62 121 1,13

Sciaridae, 1 sp.

265 2,47

Larvas Carabidae

12 0,11

Tipulidae, 1 sp.

48 0,45

Larvas Diptera

33 0,31

48 0,45

Psychodidae, 1 sp.

20 IV 2013/18 V 2013; 22 VI 2013/20 VII 2013). Paralelamente, se realizaron muestreos por rastreo de las especies que se observaron en el transcurso de la instalación y retirada de las trampas, actividad que permitió la captura de otras especies que no suelen caer en ellas, como sucede con la mayoría de las arañas. Los muestreos fueron realizados por miembros del Grupo Espeleológico de Granada (GEG) dentro del marco del contrato de investigación ''Invertebrados cavernícolas de Andalucía''. Todos los especímenes se conservaron en etanol de 70°. La fauna acompañante fue analizada a partir de un total de 10.739 especímenes de artrópodos colecta� dos en la Sima de San Rafael, siendo posteriormente clasificados en 38 categorías con desigual precisión

10.739

taxonómica dada la imposibilidad de poder llegar a conocer la totalidad de las especies (tabla 2). El estudio del edeago (en posición dorsal y late� ral) y de la genitalia femenina, se ha realizado tras preparar las genitalias en pequeñas láminas transpa� rentes de acetato, utilizando una resina hidrosoluble (dimetil hidantoína formaldehido, DMHF). El complejo espermatecal se tiñó con Negro de Clorazol y pos� teriormente fue aclarado con una solución de KOH, y posteriormente con solución de Scheerpeltz (60% etanol, 39,5% agua destilada, 0,5% ácido acético). Otras estructuras anatómicas han sido deshidratadas y se han metalizado con oro con una capa de 20 nm de grosor (Metalizador Bal–Tec SCD–005) para su observación a microscopía electrónica de barrido

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(Microscopio HITACHI S–3500–N) a alto vacío. Todas las mediciones de los especímenes se han realizado utilizando un estereomicroscopio con ocular calibrado. Como material de comparación se han examinado ejemplares de otras especies ibéricas de Trechus cuya presencia es conocida en Andalucía (tabla 1), dedicando especial atención a aquellas que per� tenecen al ''grupo–Trechus fulvus'': Trechus breuili Jeannel, 1913; Trechus lencinai (Mateu & Ortuño, 2006), Trechus lallemantii Fairmaire, 1859 y Trechus fulvus Dejean, 1831. Resultados Trechus parapandus sp. n. (figs. 2–6) Serie típica: Holotipo: 1♂, Sima de San Rafael, UTM 30SVG18992897, Cerro de la Mesa, Sierra de Parapanda, T. M. Íllora (Granada, España), 10 XI 2012/08 XII 2012, GEG leg. Depositado en la colección del Departamento de Ciencias de la Vida – V. M. Ortuño (UA/VMO), Univer� sidad de Alcalá, Alcalá de Henares (Madrid, España). Paratipos: 42♂♂, 40♀♀ (1♀ metalizada para SEM), Sima de San Rafael, UTM 30SVG18992897, Cerro de la Mesa, Sierra de Parapanda, T. M. Íllora (Granada, España), 10 XI 2012/08 XII 2012, GEG leg.; 18♂♂, 21♀♀, idem, 26 I 2013/09 III 2013; 10♂♂, 10♀♀, idem, 20 IV 2013/18 V 2013; 1♂, 1♀, idem, 22 VI 2013; 24♂♂, 35♀♀, idem, 22 VI 2013/20 VII 2013; 1♂, Sima de los Escorpiones, UTM 30SVG18192969, Cerro de la Mesa, Sierra de Parapanda, T. M. Íllora (Granada, España), 30 IX 2005, GEG leg (A. Santaella leg.). Depositados en la colección del Departamento de Ciencias de la Vida – V. M. Ortuño (UA/VMO), Universidad de Alcalá, Alcalá de Henares (Madrid, España); colección del Museo Nacional de Ciencias Naturales (Madrid, España) con el número de inventario ''MNCN Cat. Tipos Nº 2613''; colección del Centro de Colecciones de la Universidad de Almería (CECOUAL); colección Paolo Magrini (Firenze, Italia); colección Marcos Toribio (Tres Cantos, España). Diagnosis Insecto con el cuerpo alargado y el tegumento despig� mentado (color testáceo) que recuerda a un Trechus fulvus estilizado (figs. 2, 3A). Ojos vestigiales (sólo visible una pequeña cicatriz ocular sin pigmentar) (fig. 3B). Alas rudimentarias reducidas a una pequeña escama alar. Pronoto poco transverso y ligeramente cordiforme. Élitros de lados subparalelos y el disco poco convexo, con las tres estrías externas muy tenues (especialmente la 8ª); margen humeral poco marcado. Antenas y patas largas. Ambos sexos con morfología externa similar, a excepción de los dos primeros protarsómeros que están más ensanchados en los machos. Genitales femeninos según el modelo del ''grupo–Trechus fulvus''. Descripción Longitud del holotipo (desde el ápice de las mandíbu� las hasta el ápice de los élitros): 4,78 mm. Longitud de los paratipos: 4,08–5,00 mm.

1 mm

Fig. 2. Habitus de Trechus parapandus sp. n. Macho. Fig. 2. Habitus of Trechus parapandus n. sp. Male.

Cabeza ligeramente más corta (longitud medida desde el final de la sien hasta el margen anterior del clípeo) que ancha (Lc/Ac ≈ 0,8) y manifiestamente más estrecha que el pronoto. Disco con microescul� tura poligonal poco profunda (fig. 4C). Ojos reducidos a unas pequeñas cicatrices no pigmentadas (fig. 3B). Surcos frontales profundos que se prolongan hasta el clípeo, bordeando el área ocular. Sienes largas (al menos tres veces más largas que la extensión de la cicatriz ocular), muy levemente convexas y micropubescentes. Labro trapezoidal y anchamente escotado en el margen anterior (fig. 3F). Mandíbulas, piezas labiales y maxilares típicas del género. Labio provisto de dos minúsculas foveolas multiperforadas y un diente labial levemente escotado (figs. 3C–3E). Antenas largas, filiformes, cuya longitud alcanza la mitad de los élitros. Quetotaxia cefálica: dos setas supraoculares por cada ojo (la anterior ligeramente retrasada al borde posterior del ojo, y la posterior

196

contigua al tramo basal del surco); dos pares de setas clipeales (fig. 3F); seis setas próximas al margen ante� rior del labro (fig. 3F); una seta en el surco mandibular (fig. 3F); dos setas labiales insertas en la base de la escotadura (fig. 3D); seis setas prebasilares que jalonan la estructura (fig. 3C) (en algún espécimen se ha observado alguna seta supernumeraria); lígula poliqueta; cuatro setas conspicuas en el penúltimo palpómero labial de las cuáles dos se sitúan en la cara interna; antenas pubescentes desde el 2º antenómero hasta el 11º, el 1º glabro pero provisto de una serie de setas próximas al borde anterior. Pronoto apenas convexo, poco transverso (lP/aP ≈ 0,78), ligeramente cordiforme y con su mayor anchu� ra por delante del medio. Borde lateral suavemente sinuado en el tercio basal. Disco con microescultura transversa poco profunda. Canal lateral estrecho en toda su longitud. Surco medio bien indicado y fosetas basales profundas. Ángulos posteriores rectos y vivos. Base sub–recta, no saliente y ligeramente más es� trecha que el borde anterior. Quetotaxia pronotal (fig. 2): dos setas marginales, una en la base del cuarto anterior y otra sobre el ángulo posterior. Élitros estrechos, proporcionalmente largos (aE/ lE ≈ 0,60), manifiestamente subparalelos y levemente más anchos pasada la mitad de su longitud. Disco con microescultura transversa poco profunda (fig. 4D). Margen basal ligeramente inclinado con los hombros redondeados, poco marcados (fig. 2). Reborde basal interrumpido al nivel del origen de la 3ª o 4ª estría. Estrías bien indicadas, aunque superficialmente surcadas y de forma aún más leve en las estrías externas. Interestrías no convexas. Estriola apical recurrente unida a la terminación de la 5ª estría. Estriola yuxtaescutelar bien visible en la base de la 1ª interestría. Quetotaxia elitral (fig. 2): en cada élitro una seta yuxtaescutelar (en la base de la estriola); dos largas setas discales (ambas contiguas a la 3ª estría), la anterior en el quinto basal y la posterior aproximadamente hacia la mitad del élitro; setas del triángulo apical con la anterior larga e inserta junto a la 2ª estría y las otras dos, más cortas, próximas al margen (a nivel de la 1ª y 2ª estría); serie umbilical típica del género, 4 setas humerales, y 4 subapicales separadas en grupos de dos), siendo hipertróficas las setas pares. Patas más largas que en otras especies epigeas o troglófilas. Tibias anteriores surcadas longitudinalmen� te en el margen externo del dorso y manifiestamente setuladas. Órgano limpiador protibial provisto de una o dos setas clip (figs. 4E–4F). Protarsos de los machos con los dos primeros tarsómeros claramente dilatados y dentados en el margen interno. Último esternito abdominal del macho provisto de dos setas cercanas al margen distal; en la hembra con cuatro setas. Anillo edeágico triangular y alar� gado (fig. 5F). Edeago pequeño (≈ 0,74 mm) con el lóbulo medio falciforme y la lámina apical subsimétrica (en visión dorsal), corta, ancha y con el extremo redondeado (figs. 5A, 5C–5E); bulbo basal grande portando en el extremo un gran alerón sagital (fig. 5A); saco interno provisto de dos áreas escamosas (una dorsal y otra

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ventral) y dos piezas esclerotizadas en forma de lámina larga y estrecha (la pieza derecha se aguza hacia el ápice en forma de buril, y la izquierda algo más larga y ancha con el extremo distal redondeado y aspecto de espátula) (figs. 5A, 5C–5E). Parámeros anchos, subiguales (el izquierdo es ligeramente más largo que el derecho) y con cuatro o cinco setas apicales cada uno (fig. 5B). Genitalia femenina con la armadura genital (fig. 6A–B) formada por los gonópodos IX dímeros (go� nocoxito y gonosubcoxito) y los laterotertiguitos IX; gonocoxito unguiforme con 2 setas espiniformes insertas en la superficie dorsal una de ellas cerca del margen interno; cerca del ápice se observa una pequeña foseta que aloja dos pequeñas setas sensoriales; gonosubcoxito ligeramente más ancho que largo, con tres gruesas setas espiniformes alo� jadas en el margen interno y una o dos finas setas próximas al margen distal; lateroterguito IX aliforme, ligeramente esclerotizado con una quincena de setas que festonean el margen distal y otro grupo de setas, de disposición más interna, que no supera la media docena. Complejo espermatecal (figs. 6B–6C) mem� branoso con la vagina corta y ancha que da paso a una bursa copulatrix, larga e invaginada en posición anterior, lo que le confiere cierto aspecto caliciforme; oviducto impar abierto en la invaginación de la bursa en donde también se abre la espermateca que es sacciforme y se estrecha progresivamente. Variabilidad Entre los 204 especímenes que forman parte de la serie típica se ha observado escasa variabilidad mor� fológica, la cual se corresponde, fundamentalmente, con el tamaño; los especímenes de mayor y menor tamaño difieren en casi 1 mm, lo que supone que los individuos más pequeños lo son en un ≈ 18% de la talla de los más grandes. Etimología El epíteto específico ''parapandus'' proviene de la latinización del nombre de la sierra (Sierra de Pa� rapanda), en cuyo karst vive esta especie hipogea. Ecología Trechus parapandus sp. n. que evidencia marcados rasgos troglobiomorfos, no ha sido capturada en las trampas 1 a 4 (fig. 1), por lo que parece que los sec� tores más superficiales de la cueva, cercanos al pozo de entrada, no forman parte del área de confort de esta especie. Esta ha sido hallada mayoritariamente en sectores medios de la cavidad, y también en la zona terminal y más profunda, por tanto distribuida en las áreas de mayor estabilidad ambiental subterránea lo que denota su marcado carácter troglobio. Entre la fauna acompañante, los Collembola son el grupo más numeroso de artrópodos (tabla 2), con un 46% del total, y están representados por cuatro familias, Isotomidae, Sminthuridae, Onychiuridae y Entomobryidae, si bien el 82% de ellos pertenecer a la primera de las familias citadas. Son especialmente numerosos en otoño y verano, al igual que sucede con los Acari (mayoritariamente Gamasida), ambos

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UAL WE 23,9 mm 20,0 kV x90 500 µm

UAL WD 24,2 mm 20,0 kV x120 300 µm

Fig. 3. Detalles de Trechus parapandus sp. n.: A. Espécimen en visión lateral; B. Área ocular; C. SEM de los apéndices bucales en visión ventral; D. SEM del diente labial y área basal; E. SEM de la foveola multiperforada de la superficie labial; F. SEM del área anterior de la cabeza en vista dorsal (clípeo, labro y mandíbulas). Fig. 3. Details of Trechus parapandus n. sp.: A. Specimen in lateral view; B. Ocular area; C. SEM of oral appendages in ventral view; D. SEM of labial tooth and basal area; E. SEM of multiperforated foveola of the labial surface; F. SEM of anterior area of the head in dorsal view (clypeus, labrum and mandibles).

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UAL WD 24,4 mm 20,0 kV x700 50 µm

UAL WD 14,6 mm 20,0 kV x2,0 k 20 µm

UAL WD 23,5 mm 20,0 kV x200 200 µm

UAL WD 24,4 mm 20,0 kV x3,5 k 10 µm

UAL WD 14,3 mm 20,0 kV x2,0 k 20 µm

UAL WD 23,6 mm 20,0 kV x200 200 µm

Fig. 4. Imágenes SEM de detalles de Trechus parapandus sp. n.: A. Ápice del antenómero XI; B. Qui� miorreceptores (tipo ''sensilla ampullacea'') del antenómero XI; C. Microescultura poligonal del disco cefálico; D. Microescultura transversa del disco elitral; E, F. Órgano limpiador protibial en visión dorsal y ventral, respectivamente. Fig. 4. SEM images of details of Trechus parapandus n. sp.: A. Apex of antennomere XI ; B. Chemoreceptors ('sensilla ampullacea' type) of the antennomere XI; C. Polygonal microsculpture of the cephalic disk; D. Transversal microsculpture of the elytral disk; E, F. Protibial cleaning organ in dorsal and ventral view, respectively.

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0,2 mm

B A

0,3 mm

0,2 mm

Fig. 5. Genitalia masculina de Trechus parapandus sp. n.: A. Lóbulo medio en visión lateral izquierda; B. Parámeros izquierdo y derecho; C. Lóbulo medio en visión dorsal; D, E. Saco interno evaginado en visión lateral derecha e izquierda, respectivamente; F. Anillo edeágico. Fig. 5. Male genitalia of Trechus parapandus n. sp.: A. Median lobe in left lateral view; B. Left and right parameres; C. Median lobe in dorsal view; D, E. Internal sac evaginated in right and left lateral view, respectively; F. Aedeagus ring.

0,3 mm

0,1 mm

C B

Fig. 6. Genitalia femenina de Trechus parapandus sp. n.: A. Detalle de la armadura genital en visión ventral; B. Armadura genital y complejo espermatecal en visión ventral; C. Complejo espermatecal en visión lateral. Fig. 6. Female genitalia of Trechus parapandus n. sp.: A. Detail of the genital shield in ventral view; B. Genital shield and spermathecal complex in ventral view; C. Spermathecal complex in lateral view.

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grupos ajustándose, junto a Trechus parapandus sp. n., a una distribución temporal en forma de ''cubeta'' (según gráfico, fig. 7). Esta sincronía en la distribución temporal, sumada a las características biológicas generales de estos artrópodos y su abundancia, su� giere que Collembola y Acari Gamasida podrían ser las principales presas de Trechus parapandus sp. n. El segundo grupo mayoritario lo constituyen los Diptera (tabla 2) que, junto con las larvas colectadas, suponen algo más del 13%, hallándose representa� dos, en orden de abundancia, por especies de las familias Phoridae, Antomyiidae, Sciaridae, Tipulidae y Psychodidae. Dado los hábitos voladores de los imagos hay que descartarlos como presas habituales de Trechus parapandus sp. n., sin embargo las larvas sí podrían formar parte de su dieta. Entre los Coleoptera presentes en la cavidad pre� dominan los Staphylinidae, que incluyendo sus larvas, se sitúan como el tercer grupo más abundante con algo más del 11% (tabla 2). Podría esperarse que, dado el tamaño y forma de vida, los Staphylinidae (fundamentalmente del género Atheta) estuviesen incorporados a la dieta de Trechus parapandus sp. n.

Sin embargo, no parece que esto sea así ya que los tres máximos de capturas de este Trechini, trampas 7, 10 y 17, no han proporcionado ni un solo espéci� men de Staphylinidae y sí de Collembola, el grupo presa más abundante en esos puntos de muestreo (fig. 8). La familia Cholevidae, con algo más del 6% (incluyendo las larvas) está representada por una especie predominante, Speonemadus angusticollis (Kraatz, 1870). También se ha observado Catops andalusicus Heyden, 1870, que se puede considerar accidental en esta cueva (tabla 2). Los Isopoda del género Porcellio Latreille, 1804 es otro de los grupos extensamente distribuido por la cavidad, pues aparece a lo largo de todo el año y en la totalidad de las trampas. Su incidencia representa un importante porcentaje que se eleva hasta casi el 11% (tabla 2). Entre los depredadores destaca la presencia de otro Coleoptera Carabidae, Laemostenus sp. y de un Pseudoscorpion, Neobisium baenai Carabajal, García & Rodríguez, 2011, microendemismo troglobio. El máximo poblacional de este Arachnida se produce en primavera y aparece con mayor frecuencia en la

Clave de identificación de las especies de Trechus de Andalucía. Identification key of Trechus specimens from Andalucia. 1 Base del pronoto con un surco transverso neto, más o menos oblicuo, a cada lado de la base del surco medial. Fosetas basales obsoletas o muy débilmente indicadas 2 Base del pronoto sin surco transverso neto. Fosetas basales netas 5 2 Todas las estrías visibles y punteadas (las estrías externas netas) e interestrías visiblemente convexas. Edeago con el lóbulo medio (fig. 9A) grande, con la lámina apical alargada, afilada y ligeramente caída en el extremo apical; saco interno provisto de dos piezas copulatrices grandes, de desigual longitud (la derecha más larga y de aspecto espatulado; la izquierda aguzada hacia el ápice). Long.: 4,0–4,5 mm Trechus tingitanus Putzeys, 1870 Estrías externas borradas y las internas micropunteadas. Edeago con el saco interno provisto de dos piezas copulatrices más o menos grandes, paralelas y alargadas 3 3 Base del pronoto rectilínea en su tramo central, mientras que los extremos son manifiestamente oblicuos 4 Totalidad de la base del pronoto prácticamente rectilínea. Tegumento rojizo con reflejos azulados irisados. Edeago con el lóbulo medio (fig. 9B) largo y estrechado hacia el ápice para constituir una lámina apical larga (simétrica en visión dorsal); saco interno con dos piezas copulatrices manifiestamente desiguales (la derecha larga, estrecha y con el ápice romo; la izquierda más corta y recurvada en forma de cuchara). Long.: 4,5–5,0 mm Trechus rufulus Dejean, 1831

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4 Élitros de lados más paralelos con las estrías más profundas e interestrías levemente más convexas. Alados. Lóbulo medio (fig. 9C) del edeago con la lámina apical corta y gruesa (en visión dorsal girada hacia el lado izquierdo); saco interno con las dos piezas copulatrices iguales (estrechas, largas y rectas). Long.: 3,2–4,0 mm Trechus quadristriatus (Schrank, 1781) Élitros ligeramente ovalados y, por lo general, con las estrías poco profundas y las interestrías planas. Generalmente braquípteros aunque hay individuos con alas funcionales. Lóbulo medio (fig. 9D) del edeago con la lámina apical más larga y afilada (en visión dorsal simétrica); saco interno con dos piezas copulatrices manifiestamente desiguales (la derecha grande y laminar; la izquierda más corta y recurvada, esbozando una pequeña cuchara). Long.: 3,2–4,0 mm Trechus obtusus Erichson, 1837 5 Élitros pardos con una mácula testácea subhumeral y otra preapical; estrías fuertemente punteadas y bien visibles, salvo las externas que son muy tenues. Alados. Edeago con el lóbulo medio (fig. 9E) manifiestamente curvo y la lámina apical corta y redondeada en su extremo distal; saco interno provisto de una sola pieza copulatriz, estrechada en la parte anterior, a modo de lámina triangular que se comba ligeramente. Long.: 3,2–3,5 mm Trechus diecki Putzeys, 1870 Élitros siempre concolores 6 6 Élitros con las estrías externas más superficiales que las estrías internas. Tegumento rojizo testáceo brillante. Disco de los élitros deprimido. Edeago con el lóbulo medio (fig. 9F) muy arqueado, prolongado en una lámina apical recta y muy larga; saco interno con dos piezas copulatrices (la derecha en su mitad anterior afilada a modo de buril y levemente curvada hacia la parte dorsal; la izquierda constituyendo una lámina triangular). Long.: 3,8–4,0 mm Trechus planipennis Rosenhauer, 1856 Élitros con las estrías externas bien marcadas. Tegumento despigmentado por lo que muestra tonos rojizos o ambarinos. [Especies del grupo–T. fulvus: algunas especies con morfología externa muy semejante, siendo el edeago la estructura anatómica que realmente se muestra inequívocamente resolutiva en la taxonomía clásica] 7 7 Élitros estrechados en la mitad anterior proporcionando un aspecto levemente piriforme; hombros caídos por lo que el margen anterior de los élitros es ligeramente oblicuo respecto al plano sagital. Pronoto manifiestamente cordiforme (con un índice de transversalidad L/A ≈ 0,79). Convexidad ocular muy marcada, con ojos vestigiales y las sienes muy oblicuas. Antenas largas que casi alcanzan la mitad de los élitros. Edeago con el lóbulo medio (fig. 9G) robusto, arqueado formando un ángulo obtuso muy por delante del orificio del bulbo basal, la lámina basal gibosa y la lámina apical corta y truncado–dentada en el extremo distal; saco interno con dos piezas copulatrices laminares subsimétricas. Long.: 3,7–4,8 mm Trechus lencinai (Mateu & Ortuño, 2006) Élitros ligeramente ovalados o de lados subparalelos; hombros más o menos marcados y el margen anterior de los élitros perpendicular (o muy levemente oblicuo) respecto al plano sagital. Edeago con el lóbulo medio (figs. 5A, 9H–9J) provisto de un saco interno con dos piezas copulatrices manifiestamente asimétricas 8

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8 Ojos grandes y salientes por lo que la convexidad ocular está muy marcada de modo que las sienes son muy oblicuas. Élitros de lados subparalelos, con estrías profundas fuertemente punteadas e interestrías visiblemente convexas. Pronoto grande, transverso (con un índice de transversalidad L/A ≈ 0,74~0,79) apenas cordiforme, con el canal lateral ancho y profundo. Edeago con el lóbulo medio (fig. 9H) fuertemente acodado, no girado hacia el lado izquierdo (en visión dorsal) y la lámina apical moderadamente desarrollada; saco interno con la pieza copulatriz derecha laminar, estrechada en su mitad anterior en una laminilla que se curva bruscamente hacia abajo, mientras que la pieza izquierda es más corta y con forma de lámina triangular. Long.: 5,5–5,8 mm Trechus lallemantii Fairmaire, 1859 Ojos vestigiales o funcionales (en cuyo caso la convexidad ocular no es tan acusada) con las sienes menos oblicuas. Élitros con estrías punteadas pero menos profundas y las interestrías convexas, apenas convexas o planas 9 9 Individuos con rasgos troglobiomorfos. Convexidad ocular poco saliente, con ojos vestigiales y con las sienes largas (al menos tres veces más largas que el espacio que ocupa la cicatriz ocular) y prácticamente paralelas al plano sagital. Tegumento muy despigmentado (amarillento). Antenas muy largas, alcanzando la mitad de los élitros. Cuerpo poco convexo. Pronoto poco transverso (con un índice de transversalidad L/A ≈ 0,78) y ligeramente cordiforme. Élitros estrechos y de lados subparalelos con las estrías más superficiales. Edeago con el lóbulo medio (fig. 5A) muy curvado y la lámina apical corta y subsimétrica (en visión dorsal); saco interno con la pieza copulatriz derecha en forma de larga y estrecha lámina que se aguza hacia el ápice (forma de buril), y la izquierda un poco más larga formando una lámina algo más ancha, con el extremo distal redondeado (aspecto de espátula). Long.: 4,1–5,0 mm Trechus parapandus sp. n. Convexidad ocular más saliente, con ojos más o menos desarrollados pero siempre funcionales y con las sienes cortas (con una longitud semejante al espacio que ocupan los ojos) y oblicuas con respecto al plano sagital. Tegumento testáceo rojizo. Élitros más anchos y de lados subovales con las estrías más marcadas 10 10 Apéndices más gráciles. Antenómeros proporcionalmente más largos (ej.: último antenómero A/L ≈ 0,25) constituyendo antenas más largas que casi alcanzan la mitad de los élitros. Pronoto transverso y ligeramente cordiforme (con un índice de transversalidad L/A ≈ 0,74). Edeago con el lóbulo medio (fig. 9I) finalizado en una lámina apical corta y visiblemente girada hacia el lado izquierdo (en visión dorsal); saco interno con la pieza copulatriz derecha aguzada en su extremidad a modo de buril y la izquierda dispuesta más retrasada constituyendo un lámina triangular. Long.: 5,0–5,3 mm Trechus breuili Jeannel, 1913 Apéndices más cortos. Antenómeros proporcionalmente más cortos (ej.: último antenómero A/L ≈ 0,30) constituyendo antenas más cortas que notoriamente no alcanzan la mitad de los élitros. Pronoto transverso y variable en su aspecto más o menos cordiforme (con un índice de transversalidad L/A ≈ 0,70~0,73). Edeago con el lóbulo medio (figs. 9J, 10A) finalizado en una lámina apical corta más o menos simétrica (en visión dorsal); saco interno con la pieza copulatriz derecha estrecha y alargada en forma de espátula y la izquierda en forma de lámina triangular. Long.: 4,8–5,8 mm Trechus fulvus Dejean, 1831

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2.500 2.000

Staphylinidae Trechus parapandus sp. n.

Collembola Acari

1.500 1.000 500 0 Otoño

Invierno

Primavera

Verano

Fig. 7. Frecuencia estacional de algunos artrópodos en la Sima de San Rafael, incluido Trechus parapandus sp. n. Fig. 7. Seasonal frequency of some arthropods in the Sima de San Rafael, including Trechus parapandus n. sp.

Gamásidos

Isotómidos

Atheta sp.

Trechus parapandus sp. n.

200 180 160 140 120 100 80 60 40 20 0 t1

t2 t3 t4 t5 t6 t7 t8 t9 t10 t11 t12 t13 t14 t15 t16 t17 t18 t19 t20

Fig. 8. Número de capturas de Trechus parapandus sp. n. por trampas de caída y capturas de los grupos taxonómicos que pueden ser ''presas potenciales''. Fig. 8. Number of catches of Trechus parapandus n. sp. by pitfall traps and catches of taxa that are 'potential prey'.

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Fig. 9. Lóbulo medio del edeago y piezas copulatrices, en visión lateral izquierda, de las especies de Trechus conocidas de Andalucía: A. Trechus tingitanus; B. Trechus rufulus; C. Trechus quadristriatus; D. Trechus obtusus; E. Trechus diecki; F. Trechus planipennis; G. Trechus lencinai; H. Trechus lallemantii; I. Trechus breuili; J. Trechus fulvus. No están dibujados a escala. Fig. 9. Median lobe of the aedeagus and copulatory piece, in left lateral view, of the Trechus species known from Andalucía: A. Trechus tingitanus; B. Trechus rufulus; C. Trechus quadristriatus; D. Trechus obtusus; E. Trechus diecki; F. Trechus planipennis; G. Trechus lencinai; H. Trechus lallemantii; I. Trechus breuili; J. Trechus fulvus. These are not drawn to scale.

zona superior de la cavidad; ambas circunstancias apuntan hacia una segregación espacial y temporal que minimiza el encuentro con Trechus parapandus sp. n. Otros depredadores importantes son los Ara� nea, destacando por sus rasgos troglobiomorfos un Linyphiidae anoftalmo que posiblemente se trate de una nueva especie (actualmente en estudio).

Discusión Los caracteres externos de Trechus parapandus sp. n. así como la configuración de las genitalias masculina y femenina, sugieren la ubicación de esta especie en el ''grupo–Trechus fulvus''. Este conjunto de especies parecen tener su origen a comienzos

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0,2 mm

Fig. 10. Lóbulo medio del edeago de Trechus fulvus: A. Lóbulo medio en visión lateral iz� quierda; B. Saco interno evaginado en visión lateral izquierda. Fig. 10. Median lobe of the aedeagus of Trechus fulvus: A. Median lobe in left lateral view; B. Internal sac evaginated in left lateral view.

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las especies. De este modo, se observa que Trechus fulvus, Trechus lallemantii y Trechus breuili cuentan con una de las piezas con aspecto de lámina trian� gular (figs. 9H–9J, 10), y Trechus parapandus sp. n. cuenta con una larga lámina acintada con el extremo anterior redondeado (figs. 5A, 5C–5E). Ubicar Trechus lencinai en el ''grupo–Trechus fulvus'' supone un mayor problema debido a la especial configuración de las dos piezas esclerotizadas del saco interno del edeago, estructuras que se muestran laminares y subsimétricas (Mateu & Ortuño, 2006; Ortuño & Barranco, 2013), alejadas del modelo de Trechus fulvus y otras especies próximas (figs. 9G, 9J, 10). Sin embargo, estudios moleculares confirmaron su emplazamiento en dicho grupo (Faille et al., 2014). La forma del saco interno de Trechus parapandus sp. n. con dos largas láminas, de longitud parecida aun� que manifiestamente asimétricas (figs. 5A, 5C–5E), recuerdan más al modelo de saco interno que muestra Trechus lencinai (fig. 9G), por lo que cabría esperar que en el seno de ''grupo–Trechus fulvus'' ambas especies pudiesen tener un ancestro común ajeno al de las demás especies ya citadas. Agradecimientos

del Mioceno (Faille et al., 2014) y, en la actualidad, se extiende por el norte de África, península ibérica y, debido a la amplia distribución de T. fulvus, tam� bién por las costas atlánticas del norte de Europa (Jeannel, 1927), y en el que las especies del antiguo género Antoinella Jeannel, 1937 se han integrado recientemente (Casale, 2011; Faille et al., 2014). En lo que respecta a la fauna de Andalucía (ta� bla 1), hasta el momento, sólo cuatro especies de este grupo han sido reconocidas: Trechus lallemantii, Trechus breuili, Trechus fulvus y Trechus lencinai. Las dos primeras son fácilmente asignables al ''gru� po–Trechus fulvus'' ya que muestran su morfología externa y la disposición de las piezas esclerotizadas del saco interno del edeago claramente semejante a las de Trechus fulvus (figs. 9H–9J, 10A). De hecho, ambas especies fueron consideradas inicialmente por Jeannel (1920) como subespecies de Trechus fulvus. Además, en el caso de Trechus breuili este fue ubicado en el seno del ''linaje–Trechus martinezi'', constituido por un grupo de especies del ''grupo– Trechus fulvus'' con distribución en el noreste de las Cadenas Béticas (Ortuño & Arillo, 2005) y el norte de Argelia (Ortuño, 2008). Trechus parapandus sp. n. pese a formar parte del ''grupo–Trechus fulvus'' mues� tra notables diferencias con Trechus fulvus, Trechus lallemantii y Trechus breuili, no sólo por los rasgos troglobiomorfos que exhibe sino también por la con� figuración de las piezas del saco interno del edeago (figs. 5A, 5C, 5D, 5E). Mientras que los caracteres que se derivan del modelado troglobiomorfo no son útiles para establecer afinidades entre las especies, la configuración del saco interno del edeago sí parece mostrar estructuras conservadoras, que vendría a informar sobre el mayor o menor parentesco entre

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Adephaga: Carabidae). Journal of Natural History, 39: 3483–3500. Ortuño, V. M. & Arribas, O., 2010. Clarification of the Status of Trechus comasi Hernando from the Iberian Peninsula, and Its Taxonomic Position (Coleoptera: Carabidae: Trechini). The Coleopterists Bulletin, 64(1): 73–74. Ortuño, V. M. & Barranco, P., 2013. Duvalius (Duvalius) lencinai Mateu & Ortuño, 2006 (Coleoptera, Carabidae, Trechini) una especie hipogea del sur de la península ibérica. Morfología, reubicación taxonómica, sistemática y biología. Animal Biodiversity and Conservation, 36(2): 141–152. Ortuño, V. M., Cuesta, E., Gilgado, J. D. & Ledesma, E., 2014. A new hypogean Trechus Clairville (Coleoptera, Carabidae, Trechini) discovered in a non–calcareous Superficial Subterranean Habitat of the Iberian System (Central Spain). Zootaxa, 3802(3): 359–372. Ortuño, V. M. & Novoa, F., 2011. A new species of Trechus from the Ethiopian Highlands (Coleoptera: Carabidae: Trechinae) and key to the Trechus species of Ethiopia. Annals of the Entomological Society of America, 104(2): 132–140. Rapoport, E. H., 1975. Aerografía. Estrategias Geográficas de las Especies. Fondo de Cultura Económica, México DF. Santaella Alba, A. & Moreno Espigares, J., 2006. La Sima de los Escorpiones (Sierra de Parapanda) Íllo ra – Granada. Andalucía Subterránea. Revista de Espeleología y descenso de cañones, 17: 16–18. Serrano, J., 2013. New catalogue of the family Carabidae of the Iberian peninsula (Coleoptera). Universidad de Murcia, Servicio de Publicaciones, Murcia. Toribio, M., 2014. Una nueva especie hipogea del género Trechus del Macizo del Sueve, Asturias, norte de España (Carabidae, Trechinae, Trechini). Bulletin de la Société entomologique de France, 119(2): 229–233. – 2015. Datos sobre algunos Trechus Clairville, 1806 del norte de España (Coleoptera: Carabidae: Trechinae). Revista gaditana de Entomología, 6(1): 49–55. Vera Torres, J. A., Santaella Alba, A., González–Ríos, M., Pedregosa Megías, R., Gómez Fontalva, J. M. & Martín Negro, J. C., 2014. Por las cuevas y simas de Íllora y Montefrío (Granada). Sierras de Parapanda – Madrid – Peñas de los Gitanos. Ed. Grupo de Espeleólogos Granadinos, Granada.

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Environmental factors influencing butterfly abundance after a severe wildfire in Mediterranean vegetation A. Serrat, P. Pons, R. Puig–Gironès & C. Stefanescu Serrat, A., Pons, P., Puig–Gironès, R. & Stefanescu, C., 2015. Environmental factors influencing butterfly abundance after a severe wildfire in Mediterranean vegetation. Animal Biodiversity and Conservation, 38.2: 207–220. Abstract Environmental factors influencing butterfly abundance after a severe wildfire in Mediterranean vegetation.— Despite the attention given to the ecology of butterflies, little is known about their community response to wildfires in the Mediterranean region. Here, we evaluated the butterfly assemblage two years after a severe, 13,000 ha wildfire in Catalonia (NE Spain) in relation to the surrounding unburned habitat. Using visual transect censuses we assessed community parameters such as abundance, diversity, species richness and equitability in burned and unburned areas. Correspondence analysis was used to analyse specific composition and relative abundance of species in the community. The influence of environmental variables on the abundance of some common species was analysed using generalized linear mixed models, taking spatial effects into account. No significant differences were found between areas for any of the community parameters, and no dominance was detected in the burned area. The structure of the vegetation and the geographical distribution of transects influenced the ordination of species and transects on the correspondence analysis plot. Generalized linear mixed models (GLMM) results underscored the role of nectar availability, fire and vegetation structure on the abundance of most species studied. Key words: Wildfires, Butterfly populations, Species composition, GLMM, Nectar availability Resumen Factores ambientales que influyen en la abundancia de mariposas después de un gran incendio forestal en la vegetación mediterránea.— A pesar de la atención prestada a la ecología de los lepidópteros, en la región mediterránea poco se sabe acerca de las respuestas de sus comunidades a los incendios forestales. Aquí, evaluamos la comunidad de mariposas dos años después de un gran incendio forestal que afectó 13.000 ha en Cataluña (NE de España) en relación con el hábitat circundante no quemado, mediante transectos para censos visuales. Se examinaron varios parámetros de la comunidad, como la abundancia, la diversidad, la riqueza de especies y equitatividad, comparando las áreas quemadas y no quemadas. Se utilizó el análisis de correspondencias para analizar la composición específica y abundancia relativa de las especies en la comunidad. La influencia de las variables ambientales sobre la abundancia de algunas especies comunes se analizó con modelos mixtos lineales generalizados, teniendo en cuenta los efectos espaciales. No se encontraron diferencias significativas entre los tratamientos en los parámetros de la comunidad y no se detectó dominancia en la zona quemada. La estructura de la vegetación y la distribución geográfica de los transectos influyó en la ordenación de las especies y los transectos en el análisis de correpondencias, peró no se encontró ningún efecto evidente del fuego. Los resultados de los modelos lineales generalizados mixtos (GLMM) señalaron la importancia de la disponibilidad de néctar, el fuego y estructura de la vegetación para explicar la abundancia poblacional de la mayoría de las especies modelizadas. Palabras clave: Incendio forestal, Poblaciones de mariposas, Composición específica, GLMM, Disponibilidad de néctar Received: 16 XII 14; Conditional acceptance: 17 III 15; Final acceptance: 4 VIII 15 Alba Serrat, Pere Pons, Roger Puig–Gironès, Dept. de Ciències Ambientals, Univ. de Girona, Campus de Montilivi, 17071 Girona, Spain.– Constantí Stefanescu, Museu de Ciències Naturals de Granollers, Francesc Macià 52, 08402 Granollers, Spain; and CREAF–Centre for Ecological Research and Forestry Applications, 08193 Cerdanyola del Vallès, Spain. Corresponding author: Alba Serrat. E–mail: albaserrat@hotmail.com ISSN: 1578–665 X eISSN: 2014–928 X

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Introduction Wildfires are a major ecological disturbance, affecting ecosystem functioning and species composition in forests around the world (Bond et al., 2005; Blondel et al., 2010; Mateos et al., 2011). In the Mediterranean region in particular they are considered an ecological factor that shapes the landscape and the ecosystems (Lloret, 1996). The risk, frequency and intensity of forest fires have increased, however, in recent decades largely due to land use changes. Pastures and agricultural land, for example, have been abandoned in mountain areas, and tree plantations have been established for commercial use. Furthermore, the cessation of traditional forestry has led to fuel accumulation over large areas (Debussche et al., 1999; Feranec et al., 2010) Climate change has also contributed to an increase in the frequency of fires, and is expected to cause even greater impact in the future (Piñol et al., 1998). In a context of global change, studying the response of species to environmental disturbances has become crucial and one of the main goals for conservation and landscape management (Bengsston et al., 2000). More particularly, in the Mediterranean and other regions, the responses of invertebrates to fire have been examined in diverse taxonomical groups as a way to quantify the effects of fire on species distribution and abundance, and also on changes occurring at the community level (Swengel, 2001; Kiss & Magnin, 2003; Moretti et al., 2004; Santos et al., 2009; Mateos et al., 2011). However, quite surprisingly, little is known about how butterfly communities are affected by wildfires, even though this group is considered an excellent indicator of biodiversity trends in terrestrial ecosystems (Thomas et al., 2004). Butterflies have highly specific requirements in terms of feeding resources in both the larval and adult stages (Erhardt & Mevi–Schutz, 2009; Munguira et al., 2009) and regarding the microclimatic conditions needed for the viability of populations (Thomas et al., 1999; Roy & Thomas, 2003). Some species have limited mobility and live in meta–populations, being strongly and negatively affected by habitat destruction and landscape fragmentation (Hanski & Thomas, 1994; Steffan–Dewenter & Tscharntke, 2000; Bergman et al., 2004). All these features and the ease with which they can be monitored make them an ideal target to explore the effects of forest fires on terrestrial insects. In this paper we report a study carried out to document the response of butterfly communities in an area in NE Spain that was severely affected by a forest fire. The analysis is presented at two levels: first, at the community level, to describe the effects of fire on the composition of the communities studied, and second, at the species level, to investigate the main factors affecting the relative abundance of some common species within these communities. A large wildfire like the one we examined is a first–order disturbance on the flora and fauna as it drastically reduces food resources and causes massive mortalities of organisms. It can also have indirect

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effects on the structure and species composition of plant and herbivore communities. It has been reported that various insect species (including some butterflies) decrease sharply in numbers in the early stages after a fire (Swengel, 2001). Because of their sensitivity to environmental alterations and changes in vegetation structure, butterfly communities are strongly affected by forest fire. We predicted a decrease in butterfly diversity in burned areas due to the local extinction of some species and a greater dominance by opportunistic species able to recolonize the primary successional stages after such a severe disturbance (Odum, 1969; Steffan–Dewenter & Tscharntke, 1997). The abundance of a butterfly species is determined by a combination of environmental and biological factors. We expected to find a strong influence of factors related to the availability of food resources (nectar availability and abundance of larval host plants), vegetation structure (i.e., cover of different vegetation layers) and the fire effect itself. The importance of these factors depends on specific biological characteristics that determine the sensitivity of species and the resilience of populations. In connection with these predictions, the goals set in this study were: (1) to assess the modification of the butterfly community two years after a large wildfire in relation to control areas, using general descriptors such as species richness, abundance and dominance; and (2) to analyze the abundance of some common species and determine the most influential environmental factors in the recovery of butterfly populations. Material and methods Study area The study area is located in the county of Alt Empordà (Girona province, NE Spain) (fig. 1). The region has a rugged relief and the climate is subhumid and humid Mediterranean, with the strong influence of the northern wind, known as the tramuntana. Although the potential vegetation is holm oak (Quercus ilex) forests, current landscape is a mosaic resulting from historical and current land use, with a dominance of Aleppo Pine (Pinus halepensis) and abandoned cork forests and crops (vineyards and olives) that have now turned into Mediterranean shrub land. The fire in the region occurred on 22nd July 2012 and lasted six days. Driven by the tramuntana, 13,963 ha (according to the Forest Division of the Catalan Government Fire Service 2013) were burned. The region has a long history of wildfires but this was the largest since 1986 when 26,000 ha burned. Post–fire management in the 2012 fire consisted mostly of logging, with timber being removed or made into chips for use as fuel for power plants. Sampling design We selected seven sampling localities. To reduce environmental variability, all localities were situated in pine forests and shrublands on the western part of the burned area and its nearby unburned area, on

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J1C J1F

J2C J3C

J2F

J3F J4F J4C J5F J5C Figueres J6F

1.5

J6C

6 km

J7F J7C

Fig. 1. Study area and sampling localities: Grey polygon, burned area; ¡Unburned transects (control); l Burned transects. Burned area perimeter defined by the Forest Division of the Catalan Government Fire Service (Àrea Forestal de Bombers de la Generalitat de Catalunya, 2013). Fig. 1. Área de estudio y localidades de muestreo: polígono gris, área quemada; ¡ Transectos no quemados (control); l Transectos quemados. El perímetro del área quemada ha sido definido por la Dirección Forestal del Servicio de Bomberos del Gobierno Catalán (Àrea Forestal de Bombers de la Generalitat de Catalunya, 2013).

limestone substrate. Two 200–m long transects were set up in each location: a control transect in the unburned area at 200 to 1,000 m from the fire perimeter and a burnt transect in the burned area, 200–700 m away from the fire perimeter. Transect characteristics are shown in appendix 1. Butterfly sampling Butterflies were sampled every two weeks from the beginning of April 2014 until the end of June 2014, with a total of five visits to each transect. The sampling method consisted of counting adult butterflies following the standards of the Butterfly Monitoring Scheme (BMS) method (Pollard & Yates, 1993). Transects were walked at a constant speed, between 11am and 5pm, under appropriate weather conditions (> 50% sun, wind ≤ force 3 in the Beaufort scale, temperature ≥ 17ºC). We counted only those butterflies that were 5 m or less in front and 2.5 m to the sides of the recorder. When identification to species level was not possible at first sight, butterflies were captured with an entomological net for close inspection and then released. Butterflies were identified consulting Tolman & Lewington (2011) field guide.

Species abundance modelling A subset of 15 butterfly species was selected to study the influence of environmental and biological factors on their relative abundance. Species were selected mainly based on their commonness at the study sites, and also because their flight period coincided with our sampling period. Table 1 gives a list of the 15 species with some basic information on their natural history in the study area and the response of their host plants to fire disturbance. We also measured the availability of food resources (nectar for adults and host plants for larvae) and vegetation structure (foliage cover of the different strata) to model butterfly abundance in each locality (cf. Dennis, 2010). Foliage cover was estimated once for three vegetation strata (grass, shrub and tree) comparing it to a standard template (Prodon & Lebreton, 1981). For each transect, foliage cover was estimated at three equidistant sites and the mean was calculated. Nectar availability was estimated in each census using a semi–quantitative scale of flower abundance. The following four categories were distinguished: 0. No flowers; 1. Few flowers; 2. Moderate number of flow-

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ers; and 3. High abundance of flowers. To model the abundance of butterfly species, nectar availability was calculated as the average of the estimates corresponding to the counts on which the species was detected. The availability of larval host plants for the modelled species was also measured. The list of potential host plants was obtained from García–Barros et al. (2013), complemented with observations by C. Stefanescu in the study area (see table 1). Measures were taken for Prunus spp., Crataegus monogyna, Rhamnus alaternus, Lonicera implexa, Quercus coccifera, Cistus salviifolius, Cistus monspeliensis, Dorycnium pentaphyllum, Thymus vulgaris and Biscutella laevigata. For the first four plant species, abundance was directly measured by counting the number of individual plants along the transect. This was not possible for the other plant species due to their high density. Abundance was then estimated from their cover using the semiquantitative scale: 0. Species absent (0%); 1. Low cover (< 25%); 2. Moderate cover (25–50%) and 3. High cover (> 50%). Data analysis Butterfly data were summarized in a specific composition table showing total abundance of each species for burned and unburned control transects. Data were analysed using the R statistical software (R CoreTeam, 2014). First, we conducted a comparative analysis of the structure of the community in burned and unburned areas. A correspondence analysis (CA) implemented with the statistical package BiodiversityR (Kindt & Coe, 2005) was used to assess the structure of the community and to explore the main gradients influencing butterfly composition. The input for the CA was the total count of each species per transect. In this analysis, we excluded species whose total count was less than five individuals. Second, to highlight differences between the two treatments (fire vs. control), we also calculated the following community descriptors: richness (S), abundance, diversity (Shannon–Wiener index, H) and equitativity (Evenness Index, E): S

H' = S i = 1 pi lnpi

E = exp (H’)/S

Normality and homoscedasticity of the descriptors was tested using Shapiro–Wilks and Bartlett tests. Differences in the descriptors between the two treatments were tested using ANOVA. The abundance of the most common species was modelled using generalized linear mixed models (GLMM) to assess the importance of four environmental variables: fire effect, vegetation structure, and availability of food resources for adults, and for larvae. Six potential explanatory variables were selected as descriptors of these environmental variables and were used as fixed factors on the GLMM: fire (Fire), percentage of foliage cover of the three vegetation layers: herbaceous (Herb), shrub (Shrub) and tree (Canop), host plant availability (Host Pl) and nectar availability (Nectr). We used locality as a random factor to account for our particular sampling design

(with burned and control transects paired in specific localities) while controlling for possible site–based differences. Box–plots were used to check the potential influence of fire on the environmental variables. Variables highly influenced by fire were excluded to avoid redundancy (CANOP and some host plants: Quercus coccifera, Dorycnium pentaphyllum, Cistus sp.). To obtain the model that could best explain the abundance of selected species (response variable) we performed multiple GLMM for each species. Competing models were compared using Aikake’s Information Criteria (AIC) based on maximum likelihood. The model with the lowest AIC value was the approximation that best fitted the data. Differences (Ai) between the AIC value of the best model and the AIC value for each other model were used to assess model performance. Models with Ai values lower than two are considered to be essentially as good as the best approximating model (Symonds & Moussalli, 2011). Analyses were performed with the statistical package lme4 (Bates et al., 2014), using the loglink function and structure of negative binomial residues. Results Community level A total of 918 butterflies belonging to 47 species were observed in the censuses. We observed 398 butterflies belonging to 39 species in burned transects, and 520 individuals belonging to 37 species in control transects (table 2). Regarding the similarity of species between treatments, 28 were common to the burned and control transects, eight were found only in the control transects (e.g., Pararge aegeria and Melanargia lachesis) and 10 were found only in the burned transects (e.g., Vanessa cardui and Coenonympha dorus). The most abundant species in the control transects was Pyronia bathseba, representing almost 16% of the total individuals. Other abundant butterflies were Gonepteryx cleopatra (14.6%) and Lysandra hispana (10.8%). These three species represented 41.4% of abundance in controls, and the 10 most abundant species attained a figure of 78.3%. In burned transects, G. cleopatra was the most common species (16.3%), followed by L. hispana (11.3%) and Satyrium esculi (9.3%). These three species represented 36.9% of the total number in the burned transects, and the 10 most abundant species represented 72.6%. Correspondence analysis (CA) graphics show the ordination of the butterfly species (fig. 2A) and the 14 transects (fig. 2B) based on specific composition and relative abundance of species in transects. The first three axes explained 57% of the variance in the dataset. Despite some overlap in the centre of the graph, burned and control polygons were segregated along a diagonal gradient in the biplot of the first two axes, going from Vanessa cardui —a species only found in burned transects and with extreme negative coordinates— to P. aegeria —a species only found in control transects and showing extreme positive coordinates.

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Table 1. Main butterfly and host plant species and their biological and ecological characteristics (own elaboration based on information from Stefanescu et al. [2011], Garcia–Barros et al. [2013], Paula & Pausas [2013] and Stefanescu [pers. observ.]): I. January; II. February; III. March; IV. April; V. May; VI. June; VII. July; VIII. August; IX. September; X. October; XI. November; XII. December; E. Eggs; L. Larvae; Lhp. Larvae on host plants; P. Pupae; A. Adults; S. Seeders; R. Resprouters. Tabla 1. Principales especies de mariposas y planta huésped y sus características biológicas y ecológicas (elaboración propia en base a información de Stefanescu et al. [2011], García–Barros et al. [2013], Paula & Pausas [2013] y Stefanescu [observ. pers.]+). I. Enero; II. Febrero; III. Marzo; IV. Abril; V. Mayo; VI. Junio; VII. Julio; VIII. Agosto; IX. Septiembre; X. Octubre; XI. Noviembre; XII. Deciembre; E. Huevos; L. Larvas; Lhp. Larvas en las plantas huésped; P. Pupas; A. Adultos; S. Sembradoras; R. Rebrotadoras.

Flight period

Phase during fire event

Callophrys rubi III–IV P (buried) Pseudophilotes panoptes III–IV

Larvae host plants

Regenerative strategy of host plants

Cistus salviifolius, S (Cistus sp.) and C. monspeliensis, R (D. pentaphyllum) Dorycnium pentaphyllum

P (buried)

Thymus vulgaris

E on the litter

Brachypodium phoenicoides

Euphydryas aurinia IV–VI

Lhp (hibernation nests)

Lonicera implexa, Lonicera etrusca

Satyrium esculi VI–VII

Ehp (thin branches)

Quercus coccifera, Quercus ilex

Anthocharis cardamines III–V

P (on the litter)

Several Cruciferae

Pyronia bathseba

V–VII

Gonepteryx cleopatra III–X A Rhamnus alaternus

R, smoke inhibits germination

Gonepteryx rhamni III–X A Rhamnus alaternus

R, smoke inhibits germination

Iphiclides podalirius IV–IX Lhp and A Brintesia circe VI–IX A Colias crocea III–IX Lhp and A Leptidea sinapis Lysandra hispana

Gramineae Mostly R (Brachypodium, Festuca, Bromus, Elymus, Arrhenatherum) Leguminosae (Lotus, Medicago, Trifolium, Vicia, etc.)

R (Trifolium) and S (Vicia, Medicago)

V & VII–VIII

Lhp and A

Dorycnium pentaphyllum

R and S

V–VIII

Lhp and A

Hippocrepis comosa

Gramineae (Brachypodium phoenicoides, Bromus spp., Dactylis glomerata, Festuca spp., Poa trivialis)

Several Gramineae

Mostly R

Melanargia lachesis V–IX Lhp Pararge aegeria

Prunus spp., Crataegus monogyna

III–X

Lhp and A

This ordination can be interpreted as the effect of fire on the butterfly community structure, but it does not show a strong influence of this factor. The proximity of paired transects (corresponding to the same locality)

indicates the influence of the spatial distribution. This was true for all sites except localities 4 and 6, which were characterised by dense forest and had very poor butterfly communities, strongly dominated by P.

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Table 2. Species abundance in each transect and in total (in order of abundance in the control transects): (a) Species only found in burned transects; (b) Species only found in control transects; C. Control; B. Burned; T. Total. Tabla 2. Abundancia de especies en cada transecto y en total (en orden de abundancia en los transectos de control): (a) Especies que sólo se encuentran en los transectos quemados; (b) Especies que sólo se encuentran en los transectos de control; C. Control; B. Quemado; T. Total.

Species

Pyronia bathseba

Species

Glaucopsyche alexis

Gonepteryx cleopatra

141

Maniola jurtina

Lysandra hispana

101

Melitaea didyma

2(b)

Satyrium esculi

Plebejus argus

2(b)

Gonepteryx rhamni

Melitaea phoebe

Pararge aegeria

Polyommatus escheri

Gonepteryx sp.

Issoria lathonia

Pieris brassicae

Vanessa atalanta

Leptidea sinapis

Pieris napi

Iphiclides podalirius

Cupido osiris

1(b)

Polyommatus thersites

Inachis io

1(b)

Colias crocea

Lysandra bellargus

1(b)

Anthocharis cardamines

Vanessa cardui

5(a)

Limenitis reducta

Ceononympha dorus

4(a)

Melanargia occitanica

Eucloe crameri

3(a)

Brintesia circe

Clossiana dia

2(a)

Polyommatus icarus

Celastrina argiolus

1(a)

Melanargia lachesis

7(b)

Charaxes jassius

1(a)

Papilio machaon

Hipparchia semele

1(a)

Lasiommata megera

Lycaena phlaeas

1(a)

Pieris rapae

Pontia daplidice

1(a)

Melitaea deione

Thymelicus acteon

1(a)

Pseudophilotes panoptes

Abundance

520

398

918

Callophrys rubi

Richness

Euphydryas aurinia

(b)

aegeria, a forest species. This particularity increased the distance of both sites from their paired burned transect in the ordination plot. The polarity of the burned transects polygon was given by another diagonal gradient (perpendicular to the first one, referred above) on which the association between paired transects and the differences related to the geographical distance are most obvious. At the bottom of the gradient we find localities 2, 3, 4 and 5 (at a close distance to each other and characterised by mixed forests with predominance of pine), and at the top we find localities 1, 6 and 7 (more isolated from the rest and characterised by mixed forests

2 (b)

with predominance of oak). The three localities at the end of the gradient are also those close to hill tops where P. machaon, I. podalirius and M. occitanica abound. These three species typically show hill–topping behaviour, by which males congregate in topographically elevated points to where females fly for mating. The ordination of species along the first axis appears to be determined by the structure of the vegetation (i.e., open vs. close habitats). Species with highly negative CA1 values (e.g., Melanargia occitanica, Papilio machaon, Iphiclides podalirius, Callophrys rubi) were only found in open habitat transects (either burned or

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CA2

–1

PARAEG

PAPMAC MELOCC IPHPOD CALRUB PSEPAN BRICIR SATESCPOLICA LYSHIS LASMEG MELLAC

213

PYRBAT

COLCRO

POLTHE PIERAP MELPHO GONSP LIMRED ANTCAR EUPAUR PIEBRA GONRHA GONCLE LEPSIN VANCAR

Burned Unburned

MELDEI

–2 –1

1 CA2

1 7 6

1 5 5 2

–1

7 2

4 3

3 Burned Unburned

–2 –1

CA1 Fig. 2. Ordination of butterfly community (A) and transects (B) on the biplot of the first two axes of the correspondence analysis (CA): ANTCAR. Anthocharis cardamines; BRICIR. Brintesia circe; CALRUB. Callophrys rubi; COLCRO. Colias crocea; EUPAUR. Euphydryas aurinia; GONCLE. Gonepteryx cleopatra; GONSP. Gonepteryx sp.; GONRHA. Gonepteryx rhamni; IPHPOD. Iphiclides podalirius; LASMEG. Lasiommata megera; LEPSIN. Leptidea sinapis; LIMRED. Limenitis reducta; LYSHIS. Lysandra hispana; MELLAC. Melanargia lachesis; MELOCC. Melanargia occitanica; MELDEI. Melitaea deione; MELPHO. Melitaea phoebe; PAPMAC. Papilio machaon; PARAEG. Pararge aegeria; PIEBRA. Pieris brassicae; PIERAP. Pieris rapae; POLICA. Polymmatus icarus; POLTHE. Polymmatus thersites; PSEPAN. Pseudophilotes panoptes; PYRBAT. Pyronia bathseba; SATESC. Satyrium esculi; VANCAR. Vanessa cardui. Fig. 2. Ordenación de la comunidad de mariposas (A) y transectos (B) en el diagrama de dispersión biespacial de los dos primeros ejes del análisis de correspondencias (CA). (Para las abreviaturas de las especies, véase arriba.)

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Table 3. Community parameters: abundance, species richness, Shannon Index and Evenness (mean ± SD). Tabla 3. Parámetros de la comunidad: abundancia, riqueza de especies, índice de Shannon y uniformidad (media ± DE).

Abundance

Richness

Shannon Index

Evenness

Burned

56.9 ± 42.9

16.0 ± 5.0

2.35 ± 0.33

0.69 ± 0.69

Unburned

74.3 ± 20.5

13.9 ± 8.2

1.94 ± 0.98

0.64 ± 0.64

control), whereas species with highly positive CA1 values (e.g., P. aegeria, Euphydryas aurinia, Gonepteryx rhamni, Limenitis reducta) were only found, or found in higher abundance, in control transects with dense forest. Species located in the plot centre do not seem to have any distribution pattern associated with the degree of opening of the habitat. Despite these differences in the composition of butterfly communities, we did not find significant differences in the general community descriptors between burned and unburned areas (table 3): ANOVA tests for abundance (F = 0.94, df = 1,12; p = 0.351), species richness (F = 0.348, df = 1,12; p = 0.566), diversity (F = 0.471, df = 1,12; p = 0.505), and evenness (F = 1.527, df = 1,12; p = 0.24). Species level Table 4 summarizes the models that best explained the abundance of the most common species in the study area following AIC criteria (AIC values on appendix 2). Models were obtained for 14 of the 15 species initially considered (table 1). It was not possible to build a model for M. lachesis due to an error related to covariance of data as the result of the multiple zeros in this species. Results included only a single best model for 12 species. However, in three cases, the best model was found within a set of two (Gonepteryx spp. and Gonepteryx cleopatra) or three models (Lysandra hispana). In total, 19 models were considered. Concerning the importance of fire as the only factor influencing butterfly populations, just in two species (G. cleopatra and L. hispana) this variable was sufficient to explain the abundance of butterfly species. In the other two cases with the same result (C. rubi and E. aurinia), this was the only model that could be performed due to an error related to the large amount of zeros when adding more variables. It should be noted that another model is to be considered in L. hispana (Fire + Nectar). In most species (10 out of 14), the best model (or one of the best models) was that which included the effect of the fire and nectar availability. In four species, the model including fire effect, nectar availability and structure of vegetation was the best model. In one species (I. podalirius), the best model also included larval host plant.

The effect of these variables differed depending on the species of butterfly. Fire had a negative influence on 15 of 19 models (12 species). Nectar availability had a positive effect in 11 of the 13 species where it was present. Related to the structure of vegetation, herbaceous cover had a positive influence on three of five models and shrub cover in four of five. Canopy cover was not included in the models as it showed a high negative correlation with fire, explaining the high negative coefficient of fire in the model for P. aegeria, a forest species. Lastly, host plant abundance was irrelevant except for I. podalirius, for which it had an expected positive effect. It should be noted, however, that for nine out of the 15 species of butterflies, this variable was not used to construct the model because we did not have direct measures of host plant abundance, or it showed a high influence of fire. Discussion Butterfly communities in the burned area Our results did not show any significant difference between control and burned transects in terms of mean abundance, species richness, diversity, and evenness of the butterfly community. This finding contradicts our predictions of a decline in diversity in the severely disturbed burned areas (cf. Odum, 1969; Caswell, 1976), presumably associated with the local extinction of some species (i.e. those most vulnerable because of their low mobility or because the fire occurred when they were in the critical egg or larval stages) and an increase in the dominance by rapidly colonizing species. On the contrary, our data showed that many species were able to recolonize the burned area (or resisted in it) within two years of the fire. Moreover, in our study region, butterfly communities in burned and unburned sites were both dominated by a few species, leading to similar evenness values. Thus, at the unburned sites, P. bathseba —a sedentary species— was exceedingly abundant, while at burned sites G. cleopatra —a highly mobile species— reached comparable high numbers. The speed at which butterfly communities can recover from a forest fire was noted by Nel (1986),

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Table 4. Generalized linear mixed models (all including Area as random factor). Only the models or set of models that best explain the abundance of the most common species (following AIC criteria) are detailed. Estimate values (E) are the estimate coefficients of environmental variables: SE. Standard error; Fire. Fire effect; Nectr. Nectar availability; Herb. Herbaceous foliage cover; Shrub. Shrub foliage cover; Hp. Host plant availability. (Only significant relationships are shown.) Tabla 4. Modelos lineales mixtos generalizados (incluyendo el área como factor aleatorio). Se muestran únicamente los modelos o conjunto de modelos que mejor explican la abundancia de las especies más comunes (siguientes criterios AIC). Los valores estimados (E) son los coeficientes de la estimación de las variables ambientales: SE. Error estándar; Fire. Efecto de fuego; Nectr. Disponibilidad de néctar; Herb. Cubierta de follaje herbáceo; Shrub. Cubierta de follaje arbustivo; Hp. Disponibilidad de la planta huésped. (Sólo se muestran las relaciones significativas.)

Fire E ± SE

Nectr E ± SE

4E–05 ± 0.56

–3E–04 ± 0.51

Brintesia circe

–1.385 ± 0.825

3.178 ± 1.103

Callophris rubi

0.693 ± 0.674

Anthocharis cardamines

Colias crocea Gonepteryx cleopatra

–0.237 ± 0.488

Herb E ±SE

Shrub E ± SE

Hp E ± SE

1.402 ± 0.480

2.32E–06 ± 0.569

–0.571 ± 0.650

0.775 ± 0.665

Gonepteryx rhamni

–0.893 ± 0.335

0.945 ± 0.235

Gonepteryx sp.

–0.606 ± 0.289 0.587 ± 0.230

–0.632 ± 0.293

Iphiclides podalirius

–0,678 ± 0.656 –0.707 ± 1.062 0.139 ± 0.040 0.133 ± 0.029 0.190 ± 0.062

Leptidea sinapis

–0,818 ± 0.650

1.521 ± 0.521

Lysandra hispania

0.474 ± 0.608

1.117 ± 0.538

–0.219 ± 0.729

–0.621 ± 0.334

0.221 ± 0.310 0.032 ± 0.011 –1.010 ± 0.012

Pyronia bathseba

–2.491 ± 0.474

3.302 ± 0.945 0.007 ± 0.021 0.045 ± 0.013

Pararge aegeria

0.877 ± 0.294 –0.013 ± 0.010 0.010 ± 0.011

–28.142 ± 2,048.0 2.026 ± 1.541 –0.092 ± 0.046 0.046 ± 0.037 0.131 ± 0.736

0.155 ± 0.654

Pseudophilotes panoptes –0.823 ± 0.846

2.790 ± 0.976

Satyrium esculi Euphydryas aurinia

–0.406 ± 1.291

who monitored butterfly assemblages in an area in southern France that was devastated by a wildfire of characteristics similar to those studied here. Nel (1986) recorded species occurrence during the four years after the fire and compared the changing butterfly assemblages to that known to occur in the area prior to the disturbance (Nel, 1982). The fire occurred at the end of July, and two months later the number of butterfly species was very low (six species, 10% of the initial number). However, in the following years, the recolonization process was fast, with 60% of species recorded in the second year and 80% in the third year. Nel’s results coincide with ours in the Alt Empordà, as we recorded a similarity of 62% in specific composition between burned and

control transects two years after the disturbance (see below). Moreover, preliminary observations from another burned area in Les Gavarres (Girona, NE Spain) suggest a similar pattern of recolonization. In this latter case, the fire took place in March 2014 and burned 359 ha; as in Nel’s (1986) study, our data show an initial phase lasting a few months with no butterflies at all followed by the appearance of six species (10% of species richness in control transects) by the end of the first summer. However, although the differences in community descriptors between burned and non–burned transects were non–significant due to the speed of the recolonization process, the direct or indirect effects of fire were detected in the species composition

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and their relative abundances. Of 47 species found, 10 were found only in burned transects, eight only in control transects, and 27 were common to both, with a similarity of 62% in terms of specific composition. These differences also became evident in the plot of the first two CA axes, which showed some segregation of burned and control areas. Some of these differences relate to large differences in habitat structure in sampling sites, as was the case between paired transects in localities 4 and 6. Controls 4 and 6 sampled dense forest, which resulted in low densities of a few forest specialists. Closed forest in the Mediterranean region typically holds low density and species richness of arthropods (e.g., Mateos et al., 2011; Verdasca et al., 2012), in contrast with more open areas that provide a high concentration of nectar sources and attract adults of most butterflies and other insects (Jubany & Rovira, 2000). Although the sampling design sought to reduce environmental variability of the study areas, habitat heterogeneity, the need to keep a short distance between transect pairs, and the availability of severely burned areas, meant that this was not always possible. The position of our transects was the result of the trade–off between proximity (to reduce environmental variability) and distance (to avoid the border effect) between paired (burned and unburned) transects, resulting in a distance of 200–700 m from the fire perimeter. However, differences between burned and unburned sites may also be related to the functional groups in each area. For example, Kwon et al. (2013) analysed the Lepidoptera communities for five years after a fire in Korea. At first they noted a reduction in the number of specialists in the disturbed areas compared to nearby unaffected areas, but they observed that this difference disappeared by the end of the study period when the proportion of butterfly functional guilds (i.e., generalist and specialist, based on larval host plant use and adults habitat) had returned to original levels. Nel (1986) and Cleary & Genner (2004) obtained similar results. They also noted that during the process of butterfly recovery in burned areas, the first butterflies to arrive were generalist species and that these were replaced by specialist species over the following years. Swengel (1996) and Vogel et al. (2010) also found that butterfly specialists took three to four years to recover from fire disturbances in several open areas, possibly linked to the process of colonization of the area after local extinction. This was not the focus of the present study so we are unable to exclude a role of local resistance in addition to colonization after fire to explain the results found. Several recent works point to a correlation of various life history traits that allows a species to allocate along an axis from extreme generalism to extreme specialism (Carnicer et al., 2013; Dapporto & Dennis, 2013). In this context, the work by Carnicer et al. (2013) —based on data from the Catalan Butterfly Monitoring Scheme which includes several sampling sites in Alt Empordà— offers an excellent framework to investigate this issue further.

Serrat et al.

Main factors affecting the abundance of the most common species To model the butterfly abundance of particular butterfly species in burned and unburned areas, we constructed generalized linear mixed models that took into account those environmental factors that presumably had the strongest effects on the populations. Besides the availability of feeding resources and the general habitat structure, we explicitly tested the importance of fire disturbance in explaining butterfly abundance. Models showed the outstanding importance of nectar availability, which had a significant positive influence in 11 out of 15 successfully modelled species. This result is not surprising, as many studies have shown the key role of nectar availability in explaining butterfly distribution and abundance (e.g., Loertscher et al., 1995; Schneider et al., 2003) in temperate areas. In this respect, fire may have indirect positive effects on some mobile butterfly species, as massive blooms of some flower species (e.g., Galactites tomentosa, Cistus monspeliensis, etc.) are highly characteristic in our study area one or two years after a fire disturbance (Pons & Prodon, 1996). The dominance of the highly mobile Gonepteryx spp. species at the burned sites, where its host plant Rhamnus alaternus was found in low abundance or completely absent (in agreement with Paula & Pausas, 2013), was probably explained by this fact, as population movements in search of nectar sources are a common phenomenon in our region (García–Barros et al., 2013). However, the importance of nectar availability was not only detected in well–known highly mobile species, such as Anthocharis cardamines and Colias crocea (e.g., Stefanescu, 2000; Kuussaari et al., 2014), but also in sedentary species such as Pseudophilotes panoptes where it appeared quite unexpectedly as the single determinant of butterfly abundance. On the other hand, host plant abundance entered as an explanatory variable in only one species, Iphiclides podalirius, with the expected positive effect. For S. esculi and C. rubi, the main host plants (Quercus coccifera and Cistus spp., respectively) were not included in the model even if we had measurements of their abundance because they were strongly and positively associated with fire. These two butterfly species were the only ones showing a positive effect by fire, possibly explained by the high densities that their host plants attain in burned areas as a result of their quick resprouting (Paula & Pausas, 2013). Interestingly, for all other species, fire showed an invariably negative effect, which was significantly detected in almost all the modelled species. This result indicates that the recovery of butterfly populations after a wildfire event may take, in many cases, more than two years. This seems to be specially the case of Pyronia bathseba, the dominant species in unburned areas, which was markedly rarer at the burned sites, and P. aegeria, a forest species which was only found at the unburned transects. Finally, it is worth mentioning that some models included the factors related to the structure of vegeta-

Animal Biodiversity and Conservation 38.2 (2015)

tion (cover of herbaceous plants and shrubs), a result consistent with the many studies pointing to the key role of habitat structure in determining butterfly preferences (see Dennis, 2010, and references therein). To conclude, generalized linear mixed model results evidenced the influence of the availability of trophic resources and habitat structure on butterfly abundance, but also the importance of fire as a depressor of population levels in many species. However, as suggested by Cleary & Grill (2004), many other factors play direct or indirect roles in determining the presence and abundance of butterflies after a fire. These factors can be purely environmental (e.g., changes in humidity that affect sensitive species such as some Satyrines in forest habitats: Hill, 1999), or ecological (such as alterations of complex interactions between various species as in the case of myrmecophilous lycaenids). Undoubtedly, these additional factors influence the recovery of butterfly populations after a wildfire event and may account for the relatively low explanatory power of our models. Acknowledgements Thanks are due to Miguel Clavero for advice on data analysis and to an anonymous referee for suggestions on an early draft. This work was partly supported by CGL2014–54094–R. References Àrea Forestal de Bombers de la Generalitat de Catalunya, 2013. Informe d’incendi forestal de La Jonquera del 22/07/2012. Lo Forestalillo, 155: 3–6. Bengtsson, J., Nilsson, S. G., Franc, A. & Menozzi, P., 2000. Biodiversity, disturbances, ecosystem function and management of European forests. Forest Ecology and Management, 132: 39–50. Bergman, K. O., Askling, J., Ekberg, O., Ignell, H., Wahlman, H. & Milberg, P., 2004. Landscape effects on butterfly assemblages in an agricultural region. Ecography, 27: 619–628. Blondel, J., Aronson, J., Bodiou, J. Y. & Boeuf, W., 2010. The Mediterranean Region. Biological Biodiversity in Space and Time. Oxford University Press, Oxford. Bond, W. J., Woodward, F. I. & Midgley, G. F., 2005. The global distribution of ecosystems in a world without fire. New Physiologist, 165: 525–538. Carnicer, J., Stefanescu, C., Vila, R., Dincă, V., Font, X. & Peñuelas, J., 2013. A unified framework for diversity gradients: the adaptive trait continuum. Global Ecology and Biogeography, 22: 6–18. Caswell, H., 1976. Community Structure: A Neutral Model Analysis. Ecological Monographs, 46: 327–354. Cleary, D. F. R. & Genner, M. J., 2004. Changes in rain forest butterfly diversity following major ENSO–induced fires in Borneo. Global Ecology and Biogeography, 13: 129–140.

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Cleary, D. & Grill, A., 2004. Butterfly response to severe ENSO–induced forest fires in Borneo. Ecological Entomology, 29: 666–676. Crawley, M. J., 2002. Statistical Computing: An Introduction to Data Analysis using S–Plus. John Wiley & Sons, New York. Dapporto, L. & Dennis, R. L. H., 2013. The generalist–specialist continuum: Testing predictions for distribution and trends in British butterflies. Biological Conservation, 157: 229–236. Debussche, M., Lepart, J. & Dervieux, A., 1999. Mediterranean landscape changes: evidence from old postcards. Global Ecology and Biogeography, 8: 3–15. Dennis, H. L., 2010. A resource–based habitat view for conservation. Butterflies in the ritish landscape. Willey–Blackwell. Erhardt, A. & Mevi–Schutz, J., 2009. Adult food resources in butterflies. In: Ecology of butterflies in Europe: 9–11 (J. Settele, T. Shreeve, M. Konvicka & H. Van Dyck, Eds.). Cambridge University Press, Cambridge. Feranec, J., Jaffrain, G., Soukup, T. & Hazeu, G., 2010. Determining changes and flows in European landscapes 1990–2000 using CORINE land cover data. Applied Geography, 30: 19–35. Force, D. C., 1981. Postfire insect succession in southern California chaparral. American Naturalist, 117: 575–582. García–Barros, E., Munguira, M. L., Stefanescu, C. & Vives Moreno, A., 2013. Lepidoptera: Papilionoidea. Fauna Ibérica, 37. Museo Nacional de Ciencias Naturales, CSIC, Madrid. Hanski, I. & Thomas, C. D., 1994. Metapopulation dynamics and conservation: a spatially explicit model applied to butterflies. Biological Conservation, 68: 167–180. Hill, J. K., 1999. Butterfly spatial distribution and habitat requeriments in a tropical forest: impacts of selective logging. Journal of Applied Ecology, 36: 564–572. Jubany, J. & Rovira, S., 2000. Butterfly monitoring scheme (pla de seguiment de ropalòcers) en el parc metropolità de Collserola. I Jornades sobre la Recerca en els sistemes naturals de Collserola: aplicacions a la gestió del Parc. Consorci Parc de Collserola. Kiss, L. & Magnin, F., 2003. The impact of fire on some mediterranean land snail communities and patterns of post–fire recolonisation. Journal of Molluscan Studies, 69: 43–53. Kuussaari, M., Saarinen, M., Korpela, E. L., Pöyry, J. & Hyvönen, T., 2014. Higher mobility of butterflies than moths connected to habitat suitability and body size in a release experiment. Ecology and Evolution, 4: 3800–3811. Kwon, T. S., Kimb, S. S., Leea, C. M. & Junga, S. J., 2013. Changes of butterflies communities after forest fire. Journal of Asia–Pacific Entomology, 16: 361–367. Lloret, F., 1996. El foc en un context mediterrani. In: Ecologia del foc: 41–45 (J. Terrades, Ed.). Proa, Barcelona.

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Appendix 1. Transect description and code. Transects in hills are indicated due to the possible influence of hill–topping on butterfly numbers. Apéndice 1. Descripción del transecto y el código. Se indican los transectos en colinas debido a la posible influencia del comportamiento "hill–topping" en la abundancia de mariposas. Code Fire

Transect location

Description

J1F

Burned

Path

Scrub with scattered trees (hill)

J1C

Unburned

Track

Scrub with scattered trees

J2F

Burned

Track

Pine forest, recently salvage logged

J2C

Unburned

Track & Path

J3F

Burned

Track

Dense formation of pine forest, recently salvage logged

J3C

Unburned

Track

Dense pine and oak mixed forest

J4F

Burned

Track

Dense mixed forest of pine and oak, salvage logged

J4C

Unburned

Track

Dense pine and oak mixed forest

J5F

Burned

Path

Pine forest with some oak, recently salvage logged

J5C

Unburned

Path

Pine and oak mixed forest

J6F

Burned

Path

Scrub with scattered trees

J6C

Unburned

Path

Closed forest of pine and oak Oak and pine mixed forest, thickets and little meadows

J7F

Burned

Path

J7C

Unburned

Track & Path

Pine and oak mixed forest, moderately closed

Dense thicket (hill)

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Appendix 2. Akaike Information Criteria (AIC) for several generalized linear mixed models explaining the abundance of the most common butterflies in the study area. Each column adds one or two environmental variables: (a) Best models (lowest AIC value, AICb); (b) Other models to be considered (AICi–AICb < 2): np. Model not possible to perform; nd. No data available. Apéndice 2. Criterios de Información de Akaike (AIC) para varios modelos mixtos lineales generalizados que explican la abundancia de las mariposas más comunes en el área de estudio. Cada columna añade una o dos variables ambientales: (a) Los mejores modelos (valor AIC más bajo, AICb); (b) Otros modelos a tener en cuenta (AICi–AICb < 2): np. Modelo que no es posible llevar a cabo; nd. No hay datos disponibles.

Treatment (Fire)

Adults food availability (+ Nectr)

Structure of vegetation (+ Herb + Shrub)

Larvae food availability (+ Hp)

Anthocharis cardamines

46.2

34.2(a)

37.6

Brintesia circe

47.2

40.3

43.5

Callophris rubi

36.8(a)

Colias crocea

53.4

41.9

45.5

101.9

102.6

105.2

106.1

Gonepteryx rhamni

77.8

70.8

73.7

74.2

Gonepteryx sp.

73.1

68.1(a)

69.6(b)

71.4

Iphiclides podalirius

58.9

58.4

57.7

54.4(a)

Leptidea sinapis

73.1

66.3(a)

70.2

74.2

Lysandra hispana

91.6

(a)

90.7

91.4

Pyronia bathseba

93.7

70.6

64.5(a)

Pararge aegeria

47.3

49.3

42.9

Satyrium esculi

150.9

70.2

82.5

Pseudophilotes panoptes

28.5

19.4

23.1

29.2(a)

Gonepteryx cleopatra

Euphydryas aurinia

(a)

(b)

(a)

(b)

(a)

(a) (a)

(b)

(a)

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Cultural transmission and its possible effect on urban acoustic adaptation of the great tit Parus major J. Bueno–Enciso, D. Núñez–Escribano & J. J. Sanz

Bueno–Enciso, J., Núñez–Escribano, D. & Sanz, J. J., 2015. Cultural transmission and its possible effect on urban acoustic adaptation of the great tit Parus major. Animal Biodiversity and Conservation, 38.2: 221–231. Abstract Cultural transmission and its possible effect on urban acoustic adaptation of the great tit Parus major.— Urban great tits (Parus major) sing with a higher minimum frequency than their forest conspecifics. Cultural processes may account at least in part for the song divergence in city birds as great tits learn their repertoire from conspecifics and switch to high pitch song types in presence of background noise. However, in small cit� ies, this process of cultural divergence could be constrained because it is likely that these birds have a greater exchange of song types with the outside. We tested this prediction by recording great tit songs in a small city (Toledo, central Spain) and in a nearby forest. We found that background noise and the peak and the maximum frequency of songs were higher in the city but the minimum frequency did not differ. The pause length was also longer in forest birds. Seventy percent of the song types were shared between Toledo and the nearby forest. These results suggest that the small size of Toledo allows a homogenized cultural wealth, preventing the devel� opment of a high pitch song as observed in larger cities. Key words: City size, Anthropogenic noise, Cultural evolution, Meme, Song divergence, Frequency Resumen Transmisión cultural y su posible efecto en la adaptación acústica urbana del carbonero común Parus major.— El carbonero común (Parus major) urbano canta con una frecuencia mínima mayor que sus conspecíficos forestales. Detrás de esta divergencia acústica podrían estar algunos procesos culturales, ya que dichas aves aprenden sus cantos de los vecinos y cambian a tipos de canto con una frecuencia alta en presencia de ruido de fondo. Sin embargo, en las ciudades pequeñas este proceso de divergencia cultural podría estar limitado, ya que en dichas ciudades es esperable un alto grado de intercambio de tipos de canto con el exterior. Nosotros testamos esta predicción grabando cantos de carbonero común en una ciudad pequeña (Toledo, España) y en un bosque cercano. El ruido de fondo fue más alto en la ciudad, al igual que la frecuencia "pico" y máxima de los cantos, pero la frecuencia mínima no difirió. La longitud de la pausa fue mayor en el bosque. El setenta por ciento de los tipos de canto se compartieron entre Toledo y el bosque cercano. Estos resultados sugieren que el pequeño tamaño de Toledo impide el establecimiento de una tradición de cantos particular con una frecuencia alta como se observa en ciudades más grandes. Palabras clave: Tamaño de ciudad, Ruido antropogénico, Evolución cultural, Meme, Divergencia de canto, Frecuencia Received: 8 VI 15; Conditional acceptance: 28 VII 15; Final acceptance: 25 VIII 15 Javier Bueno–Enciso & Daniel Núñez–Escriban o, Inst. de Ciencias del Medio Ambiente (ICAM) de Castilla–La Mancha, Univ. de Castilla–La Mancha, Av. Carlos III s/n., 45071 Toledo, Spain.– Juan José Sanz, Depto. de Ecología Evolutiva, Museo Nacional de Ciencias Naturales (CSIC), c/ José Gutiérrez Abascal 2, 28006 Madrid, Spain. Corresponding author: Javier Bueno–Enciso. E–mail: jbuenoenciso@gmail.com

ISSN: 1578–665 X eISSN: 2014–928 X

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Introduction Like many other urban bird species, urban great tits, Parus major, sing with a higher minimum frequency than their forest conspecifics (Slabbekoorn & den Boer–Visser, 2006) (see table 1), highlighting the role of environmental conditions on sound production (i.e. fig. 1 in Laiolo, 2010). Urbanization produces extreme novel habitat conditions for animals in and also near cities (Warren et al., 2006). Although urban and rural habitats differ in many aspects that could influence animal acoustic communication (Shochat et al., 2006), the loud low–frequency background noise in cities (hereafter referred to as ‘anthropogenic noise’) has shown to be directly related to the elicitation of high minimum frequency city songs (Slabbekoorn, 2013). The mechanism underlying this spectral shift in urban birds remains unclear (Nemeth et al., 2012; Slabbekoorn et al., 2012). It has been hypothesized that urban birds increase the minimum frequency of their songs to avoid overlap with anthropogenic noi� se. This acoustic shift could be a microevolutionary response or a short–term response (Brumm, 2006; Patricelly & Blickley, 2006; Slabbekoorn & Ripmees� ter, 2008). This is an important topic, because if the acoustic shift is a microevolutionary response, and because frequency features of the song may be under sexual selection (Halfwerk et al., 2011; 2012; Des Aunay et al., 2013; but see Eens et al., 2012), the spectral shift would entail genetic inheritance and could promote speciation through reproduction isolation (Slabbekoorn & Smith, 2002; Slabbekoorn & Ripmeester, 2008). However, to be selected, the urban acoustic shift should imply fitness benefits to the singer, but empirical data is lacking (Nemeth et al., 2012; Halfwerk & Slabbekoorn, 2013; but see Slabbekoorn, 2013). Experimental procedures have shown that birds can adjust their acoustic signals as a short–term response to increasing noise levels due to vocal plasticity. These plastic adjustments could be mediated by an active frequency shift (Gross et al., 2010; Bermúdez–Cuamatzin et al., 2011; Hanna et al., 2011; Potvin & Mulder, 2013), by switching song types (Cardoso & Atwell, 2011a; Halfwerk & Slabekoorn, 2009), by increasing the duration of the vocalizations (Montague et al., 2012; Potvin & Mulder, 2013), or by increasing the amplitude of the songs (Brumm, 2004; Schuster et al., 2012). These plastic adjustments are not mutually exclusive and two or more may operate concurrently or vary among species (Slabbekoorn et al., 2012). Such phenotypic plasticity would impede habitat–dependent selection and therefore speciation (Baker, 2006) Great tits are a good model to study this topic because they are a successful urban songbird and one of the best–studied species in terms of acoustics (Slabbekoorn, 2013). It is a closed–ended learning species (Rivera–Gutierrez et al., 2011) that achieves a discrete, crystalized repertoire of songs learned from parents and neighbours (McGregor & Krebs, 1982; Franco & Slabbekoorn, 2009). Thus, the different song types sung by each great tit can be considered cultural traits, referred to hereafter as 'memes' (Baker et al.,

Bueno–Enciso et al.

2001; Baker & Gammon, 2008; Cardoso & Atwell, 2011a) and the learning process in this species is the basis of their 'cultural transmission' (Slabbekoorn, 2013). Among the plastic adjustments cited earlier, great tits have been seen to actively switch to a different meme that transmits better in presence of anthropogenic noise (Halfwerk & Slabekoorn, 2009; Slabbekoorn, 2013). Thus, cultural evolution (i.e. chan� ges in the expression frequencies of the meme pool over time) may play an important role in city–forest song divergence in this species (Cardoso & Atwell, 2011a; Slabbekoorn, 2013). Indeed, the process of cultural transmission has led to the establishment of population–specific repertories or 'dialects', separa� ted by either geographical distance or barriers that prevent dispersal (O´Loghlen et al., 2011; Potvin & Parris, 2012). The mechanisms behind this cultural diversity can be classified into two broad categories: (i) stochastic factors such as meme mutation, drift or immigration, and (ii) selective pressures that favour directional changes in the meme pool frequencies such as choice of female, morphological adaptations (e.g., bill/body size), and environmental conditions that affect sound production (Lynch, 1996; Cardoso & Atwell, 2011a; O´Loghlen et al., 2011; Xing et al., 2013). In relation to the latter, and because the anthropogenic noise in cities mask birdsongs (Sla� bbekoorn & Peet, 2003), acoustic signals can reach the receivers degraded in cities and consequently do not fulfil their biological function of communica� tion (Brumm & Naguib, 2009). The risk of losing the acoustic message may lead birds to try to match their song with the acoustic properties of the environment to enhance transmission as proposed by the Acoustic Adaptation Hypothesis (AAH) (Morton, 1975). The way great tits seem to do this is by switching me� mes to a higher minimum frequency or singing them for longer periods (Halfwerk & Slabbekoorn, 2009). This acoustic shaping to the environment influences cultural transmission and therefore cultural evolution as birds copy undegraded songs differentially during their sensitive period of learning (Peters et al., 2012). Thus, under anthropogenic noise, only those memes from the whole repertoire that escape from masking or transmit and reach the receiver more clearly will be correctly and frequently copied (Slabbekoorn, 2013). Over time, this may lead to changes in the expression frequencies of the memes sung in a city, increasing those that transmit better under anthropogenic noise, i.e., those with a high minimum frequency (Luther & Baptista, 2010) because of their better acquisition by young birds during their sensitive period. This cultural evolution may be responsible, at least in part, for the city–forest song divergence found in great tits. Over time, it could be culturally and genetically fixed (Price et al., 2003; Slabbekoorn, 2013). Furthermore, such phenotypic plasticity allows acoustic shifts in frequen� cies and may account, to some degree, in urban song divergence (Nemeth et al., 2012). The aim of the present study was to compare the typological and spectral song characteristics (Baker, 2006) in two great tit populations, one in a small city, Toledo (central Spain), surrounded by a large rural and

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Table 1. Summary of the studies that assess the effect of the anthropogenic noise in the minimum frequency of songbirds. All shifts are toward high frequencies in urban songs in relation to forest songs. Types of study: O. Observational; E. Experimental; O/E. Both. References: 1. Slabbekoorn & Peet (2003); 2. Fernández–Juricic et al. (2005); 3. Slabbekoorn & der Boer–Visser (2006); 4. Bermúdez–Cuamatzin et al. (2009); 5. Parris & Schneider (2008); 6. Halfwerk & Slabbekoorn (2009); 7. Mockford & Marshall (2009); 8. Nemeth & Brumm (2009); 9. Ripmeester et al. (2010); 10. Gross et al. (2010); 11. Hu & Cardoso (2010); 12. Mendes et al. (2011); 13. Salaberria & Gil (2010); 14. Bermúdez–Cuamatzin et al. (2011); 15. Hanna et al. (2011); 16. Hamao et al. (2011); 17. Potvin et al. (2011); 18. Montague et al. (2012); 19. Potvin & Parrish (2012); 20. Nemeth et al. (2013); 21. Potvin & Mulder (2013). Tabla 1. Recopilación de los estudios que evalúan el efecto del ruido antropogénico en la frecuencia mínima de los pájaros cantores. En los cantos urbanos todos los cambios son hacia frecuencias altas en comparación con los cantos del bosque. Tipos de estudio: O. Observacional; E. Experimental; O/E. Ambos. (Para las abreviaturas de las referencias, véase arriba.) Song birds

Minimum frequency shift

Type of study

References

Great tit

Parus major

Yes

O/E

1, 3, 6, 7, 13, 16

House finch

Carpodacus mexicanus

Yes

O/E

2, 4, 14

Grey shrike–thrush

Colluricincla harmonica

Yes

Grey fantail

Rhipidura fuliginosa

Common blackbird

Turdus merula

Yes

O/E

Reed bunting

Emberiza schoeniclus

Yes

Rainbow lorikeet

Trichoglossus haematodus

Yes

Eastern rosella

Platycercus eximius

Yes

Red wattlebird

Anthochaera carunculata

Yes

Noisy miner

Manorina melanocephala

Bell miner

Manorina melanophrys

Yes

Pied currawong

Strepera graculina

Australian magpie

Gymnorhina tibicen

Grey butcherbird

Cracticus torquatus

Magpie–lark

Grallina cyanoleuca

Willie wagtail

Rhipidura leucophrys

Common myna

Acridotheres tristis

Red–winged blackbirds

Agelaius phoeniceus

Yes

O/E

Silvereyes

Zosterops lateralis

Yes

O/E

17, 19, 21

European robin

Erithacus rubecula

Yes

O/E

forest area, and the other in a nearby forest. The small size of Toledo distinguishes this study from the former studies performed in large cities (i.e., Slabbekoorn & den Boer–Visser, 2006; Hamao et al., 2011 or Sala� berria & Gil, 2010). If cultural evolution is responsible, at least in part, for the city song divergence found in this species, the fact that Toledo great tits may still be learning 'forest memes' (i.e. with relative low minimum frequency) from surrounding rural and forest areas may constrain the plastic cultural response of choosing 'urban memes' (i.e., with high minimum frequency) from their repertories; as many crystalized memes in their

8, 9, 11, 12, 20

repertories will be those learned from outside of the city. This possible high exchange of memes between Toledo and forest may allow a cultural convergence. We predicted that the differences in the minimum frequency and the typological song characteristics (i.e. percentage of meme type used) between Toledo and the forest would be less marked than those described in previous studies in city–forest pairs in large cities. This comparison of the song features between a re� latively small city and the nearby forest could help us to understand the mechanisms involving urban song divergence in great tits.

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Methods

Song analysis

Study areas

We randomly selected one song recording for each male to analyse the same number of songs for each male. Audio tracks were first converted from stereo to mono with the Audacity 1.3.14–beta program. Five strophes of the song were then selected and exported to the program RavenPro 1.4 (Charif et al., 2010), where, before analysis, a band filter was used to remove the typical low frequency background noise after visual inspection of the spectrogram to prevent the unintentional removal of any song element. For spectral description (Baker, 2006), we measured: average minimum, maximum and peak frequency (Hz), frequency bandwidth (difference between the maximum and the minimum frequency of the song, Hz), mean strophe length, mean pause length and average number of notes per phrase (Nº notes). Peak frequency was measured automatically; the other variables were measured manually using a Hann window and a fast Fourier transformation (FFT), length of 1,024, resulting in a spectral resolution of 43.1 Hz. Minimum and maximum frequencies were measured by precisely placing a selection box in the spectrogram view, similar to Francis et al. (2011). Because this manual methodology can entail a bias in the frequency measure of birdsongs (Zollinger et al., 2012), we tested the reliability of our measures by rea� nalysing a subset of song recordings (15 belonging to forest sites and 15 from Toledo) with a power spectra using Avisoft SAS Lab Pro 4.15 (Avisoft Bioacoustics, Schönfließer Str. 83, 16548 Glienicke, Germany). Concretely, we recalculated the minimum and maxi� mum frequency of this subset of song recordings by subtracting 20 dB from the peak amplitude value in the power spectrum. Neither minimum nor maximum frequency differed significantly between the two me� thodologies used (Student t–test: t = 0.01; d.f. = 58; p = 0.99; manually = 3,397.67 ± 66.26 and automa� tically = 3,398.34 ± 71.54; Student t–test: t = –0.51; d.f. = 58; p = 0.61; manually= 5,571.48 ± 136.95 and automatically = 5,459.90 ± 172.25, for the minimum and the maximum frequency respectively). We also calculated the song–rate (relation between the total number of phrases recorded and the total length of the strophe). For the typological level of description (Baker, 2006), we noted the note type following the criteria of McGregor & Krebs (1982) by visual ins� pection of the sound spectrograms, and we classified each meme–type depending on the number of notes, type, and order in the phrase. All song analyses were performed by DNE and JBE. Meme–type classifica� tions of both observers were compared to obtain a single classification.

Great tit songs were recorded in two, well–differentiated habitats: city (Toledo) and nearby forest (Montes de Toledo). Toledo is a relatively small city (17 km2) located 71 km south of Madrid (central Spain, 529 m above sea level), with a population of 80,000 inhabitants. Forest songs were recorded in two nearby areas of the Toledo mountains (Montes de Toledo, Toledo province): Quintos de Mora (39º 25' N, 4º 04' W) and San Pablo de los Montes (39º 31' N, 4º 21' W), located at 80 and 60 km south of Toledo, respectively (average elevation of both forest areas is 908 m a.s.l.; see Ferrer et al., 2012). Both forest areas comprise deciduous forests dominated by Pyrenean Oak Quercus pyrenaica, usually relegated to the shady area of the mountains and ravine funds, accompanied by Mediterranean scrublands. Both study areas have a continental Me� diterranean climate, with mean values of annual rainfall of 350–450 mm for Toledo and 700–800 mm for the forests, concentrated in the months of autumn and spring. Summer drought and daily thermal oscillation are marked in the whole area. Song recording We recorded great tit songs between March 5 and April 27 in 2011. Day 1 = March 1st. All song recordings were made between 900 and 1,500 hours. To ensure the independence of our samples, we only recorded lone individuals located at least 100 m apart, a dis� tance greater than that considered to be the territory size in this species (Naef–Daenzer & Keller, 1999). Urban great tits began to sing 22 days earlier than their forest conspecifics, possibly due to the 'heat island' effect of cities (Shochat et al., 2006) and food availability (Isaksson & Andersson, 2007). To correct the differences in the breeding cycle advancement between forest and urban great tits, we considered the day when the first male was heard singing in each study area as the first day. We recorded 56 songs from 37 different great tit males in Toledo. Four males were recorded three times, eleven males were recorded two times, and the remaining males were recorded once. We recorded 106 songs from 63 males in the forest area; three males were recorded four times, nine males were recorded three times, and 16 males were recorded two times. All songs were recorded using an EDIROL R– 09HR digital recorder equipped with a Sennheiser unidirectional microphone and headphones, pointing directly toward the singing individual. Individuals were recorded at a distance of between 10 and 15 m and at least 10 strophes per song were recorded. After each recording, background noise data (dB) were measured every second for 5 minutes with a mul� tidirectional sound level meter (SLM, A–weighted, reference level 20 µPa) positioned in a fixed location (1.5 m above the ground) with the aid of a tripod. As a measure of background noise, we used the mean value over this period.

Statistical analyses We analysed the acoustic habitat–dependent diver� gences between study areas using General Linear Models (GLMs), with the spectral song characteristics as dependent variables and study area as a cate� gorical predictor with two levels (Toledo and nearby forest). Standardized date was incorporated in all

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models as a covariate. The interaction between study area and date was also included in the full models. Number of notes was analysed using a Generalized Linear Model (GLZ) with a poisson distribution. For post hoc analyses, the Tukey HSD test was used. Differential note type used between Toledo and Forest was analysed using other GLZs with a bino� mial distribution. In this case, the response variable was the ratio between the 'number of one note type used' (numerator) divided by the 'total number of note types used' (denominator) (Zuur et al., 2009), with the study area as categorical predictor and the standardized date as the covariate. The interaction between study area and date was also included in the full models. In the present study, when the interaction between study area and date was not significant, it was eliminated from the final presented models. All these analyses were performed with R (R Core Team, 2014) and the function 'glm' of the package lme4 (Bates et al., 2014). Assumptions of the homoscedasticity, proper distribution used and independence were verified graphically with the re� siduals of the model, following the recommendations of Zuur et al. (2010).

In relation to the typological level of description, a total of 16 memes were recorded in the two stu� dy areas: 13 in Toledo, two of which (15%) were exclusive, and 14 in the forest, three of which (21%) were exclusive. Eleven of the 13 memes were common to both areas (70% of the memes recorded), but only three o were sufficiently sampled to make comparisons. The remaining eight memes were recorded for only one or two males in each study area and therefore not considered suitable for comparison. These shared memes were 'two–note type' memes and they were named arbitrarily as meme 'A' (composed of the notes 'a' and 'b', see below), shared by seven males in Toledo and by 11 in the forest, meme 'B' (composed of the notes 'a' and 'c') shared by three males in Toledo and 31 in the forest, and meme 'C' (composed of the notes 'b' and 'c') shared by seven males in both areas (see fig. 1). Apart from presenting a shorter pause length and a quicker song–rate in Toledo than in the forest (GLM: F1, 64 = 5.89; p = 0.02 and F1, 64 = 5.49; p = 0.02, respectively), these shared memes did not differ in minimum frequency (GLM: F1, 64 = 2.96; p = 0.09) or in peak frequency (GLM: F1, 64 = 2.56; p = 0.11); only the maximum frequency differed, being higher in Toledo than in the forest (GLM: F1, 64 = 19.82; p < 0.001; Toledo = 5,483 ± 117 Hz and Forest = 4,956 ± 57 Hz). The difference in maximum frequency of these three shared memes between study areas (527 Hz) was slightly smaller than the difference in maximum frequency when all memes were taken into account (598 Hz, see table 2). Thus, these three shared memes account for 88% of the magnitude of the overall maximum frequency diver� gence between the Toledo–Forest pair.

Results Table 2 shows mean ± SE values of background noise and spectral song characteristics of great tits recorded in the study areas. Some of the spectral song varia� bles showed significant differences between areas, with the exception of song–rate, strophe length and minimum frequency (tables 3). Neither spectral song variable varied significantly with the date (table 3).

Table 2. Mean ± SE (n) and range of the spectral song characteristics of male great tit (Parus major) songs in Toledo and in a nearby forest. Tabla 2. Media ± ES (n) y rango de las características espectrales del canto del macho de carbonero común (Parus major) en Toledo y en un bosque cercano.

Toledo

Mean ± SE (n)

Background noise (dB) 54.13 ± 0.84 (37)

Nearby forest Range

Mean ± SE (n)

Range

45.10–63.66

38.97 ± 0.55 (63)

30.89–48.90

Nº notes

2.27 ± 0.1 (37)

2–4

2.19 ± 0.07 (63)

2–4

Song–rate (St/s)

0.19 ± 0.03 (37)

0.03–0.37

0.15 ± 0.02 (63)

0.01–0.35

Pause length (s)

3.14 ± 0.38 (36)

1.30–6.80

4.98 ± 0.35 (61)

1.02–12.51

Strophe length (s)

2.83 ± 0.26 (37)

1.61–8.03

2.67 ± 0.14 (63)

1.26–5.64

Min. Freq. (Hz)

3,389.27 ± 82.41 (37) 2,495.1–4,179.6

3,256.19 ± 45.29 (63) 2,288.1–3,978.1

Peak Freq. (Hz)

4,413.24 ± 109.13 (37) 3,375.0–5,564.6

4,198.82 ± 62.17 (63) 3,140.6–4,952.4

Max. Freq. (Hz)

5,813.36 ± 146.16 (37) 4,570.3–7,944.9

Bandwidth (Hz)

2,483.35 ± 124.99 (37) 1,275.6–4,878.5

5,215 ± 89.78 (63)

4,318.4–7,244.3

1,925.81 ± 78.77 (63) 1,039.1–3,732.8

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Table 3. Summary of the GLMs analysing the effect of study area (Factor) and date (Covariate) on song characteristics of male great tits (Parus major). Estimate is the slope of the relationship between the covariate and the dependent variable. Significant results are highlighted in bold. Tabla 3. Resultados de los GLM que analizan el efecto del área de estudio (factor fijo) y la fecha (covariable) en las características del canto de los machos de carbonero común (Parus major). ''Estimate'' es la pendiente de la relación entre la covariable y la variable dependiente. Los resultados significativos se destacan en negrita.

Test Factor

Estimate

Test Covariate

Background noise (dB)

F1, 98 = 283.7

< 0.001

0.02

F1, 97 = 0.16

0.69

Nº notes

Z1, 98 = 0.48

0.57

0.001

Z1, 97 = 5.88

0.89

Song–rate (St/s)

F1, 98 = 3.16

0.08

< 0.001

F1, 97 = 0.38

0.54

Pause length (s)

F1, 98= 9.23

0.003

–0.003

F1, 97 = 0.89

0.35

Strophe length (s)

F1, 98 = 0.78

0.40

0.014

F1, 97 = 0.73

0.40

Min. Freq. (Hz)

F1, 98 = 1.13

0.29

2.86

F1, 97 = 0.29

0.59

Peak Freq. (Hz)

F1, 98 = 7.81

0.006

–8.42

F1, 97 = 1.36

0.25

Max. Freq. (Hz)

F1, 96 = 17.32

< 0.001

35.3

F1, 96 = 1.67

0.20

Bandwidth (Hz)

F1, 96 = 15.39

< 0.001

30.84

F1, 96 = 1.30

0.26

The 16 memes observed in great tit songs were made up of a combination of 11 different note types (table 4). To test city–forest differences, we used note types that represent at least 10% of the total (table 4). Three note types were therefore considered and they were named arbitrarily as: 'a' (Peak frequen� cy = 3,500 Hz), 'b' (Peak frequency = 4,500 Hz) and 'c' (Peak frequency = 5,500 Hz). The use of note type 'a' and 'c' differed significantly between Toledo and forest (GLZ. Note type 'a': Z1, 98 = –13.77; p < 0.001 and Note type 'c': Z1, 98 = 12.91; p < 0.001), while the use of note type 'b' did not differ between study areas (GLZ. Note type 'b': Z1, 98 = 0.65; p = 0.52). Note type 'a' was less frequently used in Toledo than in forest (Tukey test, p < 0.001; fig. 2A), contrary to note type 'c' used, which was significantly more frequently used in Toledo (Tukey test, p < 0.01; fig. 2C). Discussion Great tits in Toledo started singing before those in forest. This advancement in the phenology of urban great tits has been reported in other city–forest pair comparisons (Partecke & Gwinner, 2007; Chamberlain et al., 2009). Analysis of song recordings revealed di� fferences in the spectral song characteristics between city and forest birds, probably due to the constraints of background noise in the city, as observed in many other studies (reviewed in Laiolo, 2010). The diffe� rences observed include the temporal features of the song, such as shorter pauses in Toledo, which help to create a contrast with the background noise and

enhance the detectability of the signal (Warren et al., 2006; Hanna et al., 2011). Beyond this temporal feature of the song, the main difference between city and forest songs in other city–forest pair studies is the shift of the minimum frequency toward high frequencies (table 1). Surpris� ingly, the minimum frequency of great tit songs did not differ between Toledo and forest, but the maximum frequency did so significantly, making bandwidth in Toledo wider. The peak frequency was also signifi� cantly higher in Toledo. This is in contrast with the strategy of urban great tits reported previously, which were shown to increase the minimum frequency of their songs (Slabbekoorn & den Boer–Visser, 2006; Mockford & Marshall, 2009; Montague et al., 2012) even in a very close urban great tit population (Sal� aberria & Gil, 2010), a mechanism that has also been reported in other urban bird species (reviewed in Brumm & Zollinger, 2011; table 1). In our study, the same memes sung in Toledo and forest showed the same frequency shift in the peak and maximum frequency mentioned earlier (fig. 1), suggesting that the singing strategy in Toledo males is a short–term phenotypic response to increase the SNR in presence of anthropogenic noise (Halfwerk et al., 2011). This is also seen in other species (Gross et al., 2010; Verzijden et al., 2010; Potvin & Mulder, 2013). This variation in Toledo songs could be a consequence of an active frequency shift (McMullen et al., 2014), possibly because these frequency features of the great tit songs (peak and maximum frequencies) can be more plastically modulated; or it could be a side effect of louder singing, i.e. the Lombard ef�

Animal Biodiversity and Conservation 38.2 (2015)

Toledo

Nearby forest

kHz

227

4 3

kHz

4 3

kHz

4 3

3 Time (s)

2 Time (s)

Fig. 1. Sonograms of the shared 'memes' most frequently used, Toledo and nearby forest. Each level represents the same 'meme' elicited in Toledo and nearby forest (A, B, or C). Fig. 1. Sonograma de los "memes" más frecuentemente utilizados en Toledo y en un bosque cercano. Cada nivel representa el mismo "meme" cantado en Toledo y en el bosque, etiquetado con el nombre dado a los memes compartidos (A, B o C).

fect (Zollinger & Brumm, 2011). It has been shown that the Lombard effect increases the bandwidth of the signal and, to a lesser extent, peak frequency. Changes in minimum frequency, in contrast, are highly independent of the Lombard effect (Cardoso & Atwell, 2011b). However, we could not confirm this mechanism as we did not accurately measure song amplitude (Brumm, 2004). Both active frequency shift and increased amplitude are short–term plastic adjustments that improve detection and increase the SNR in presence of anthropogenic noise (Nemeth & Brumm, 2010; Halfwerk et al., 2011). The low minimum frequency found in Toledo great tit songs, however, is a surprising result and the main 'urban song difference' reported in this study (table 1). This lack of significant difference in

the minimum frequency between Toledo and forest could be a side effect of the small metropolitan area of Toledo and the process of cultural transmission (O´Loghlen et al., 2011; Potvin & Parris, 2012). Anthropogenic noise might influence cultural trans� mission by favouring songs that propagate better in the acoustic environment and reach the receiver more clearly (Luther & Baptista, 2010), particularly when birds differentially copy undegraded songs in their learning process (Peters et al., 2012). Thus, a phenotypic plastic response like sing with a higher minimum frequency in cities (Slabbekoorn & Peet, 2003), can be culturally fixed in a scenario where most of the population is under the selection pressure (Price et al., 2003). This is also propitiated because high minimum frequency songs transmit better and

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Note type

Toledo

Nearby forest

18.44

38.66

36.18

35.53

29.40

17.75

3.11

3.62

1.31

0.82

2.68

0.79

0.81

1.80

3.97

2.01

3.11

40 Note type 'a'

Tabla 4. Porcentaje de ocurrencia (%) de los 11 tipos de notas diferentes usadas por el carbonero conún en las dos áreas de estudio.

A 35 30 25 20 15 B 40 Note type 'b'

Table 4. Percentage of occurrence (%) of the 11 different note types used by great tits in the two study areas.

35 30 25 20 15

reach the receiver neatly (Des Anuay et al., 2013; Potvin & Mulder, 2013) and great tits choose these high memes from their repertoire and sung them for longer duration (Halfwerk & Slabbekoorn, 2009). In big cities, with a large enough surface area to harbour an almost isolated resident population throughout its life span, this scenario may occur. It can therefore be expected that within a few generations, isolated populations present culturally different sets of memes as the result of differences in background noise. This has been observed in San Francisco (Califor� nia) with the white–crowned sparrow Zonotrichia leucophrys (Luther & Baptista, 2010) and seems to be occurring in Europe with the great tit, where a particular set of 'urban memes' are sung in cities (Slabbekoorn & den Boer–Visser, 2006). However, the small metropolitan area of Toledo may allow a higher meme exchange between great tits in Toledo and those in rural and forest areas outside the city, as the high ratio of shared memes suggests; 70% in contraposition with the low ratio of shared memes in other city–forest pair studies (17% in Slabbekoorn & den Boer–Visser 2006 or 8% in Cardoso & Atwell, 2011a). In this scenario, Toledo chicks would learn both types of memes: the low minimum frequency memes from outside the city and the high minimum frequency memes developed in the city. This may lead to a homogenized cultural wealth between To� ledo and forest, constraining the development of a particular cultural song tradition in Toledo. However, a long–term study marking individuals in Toledo and the surrounding rural areas would be necessary to assess dispersal between these populations.

Note type 'c'

40 35 30 25 20 15 Toledo

Nearby forest

Fig. 2. Percentage of note type used by great tits in Toledo and nearby forest. (Vertical bars indicate the standard error.) Fig. 2. Porcentaje del uso de cada tipo de nota del carbonero común en Toledo y en un bosque cercano. (Las barras verticales indican el error estándar.)

Although Toledo and forest share 70% of their cultural wealth, a differential note type was used in both study areas (fig. 2). The proportional use of 'high–frequency notes' was higher in Toledo while the proportional use of 'low–frequency notes' was lower (Slabbekoorn & den Boer–Visser, 2006; see fig. 2). This suggests that memes that were less masked by noise were sung more frequently, as proposed by Halfwerk & Slabbekoorn, 2009), and could be a cultural plastic response of Toledo males. The effect of this potential cultural plastic response, however,

Animal Biodiversity and Conservation 38.2 (2015)

was not large enough to differentiate the minimum frequency between study areas. This study suggests that the degree of isolation of a population could influence the city–forest song divergence in great tits. The size of a city may be an important feature in song divergence, as birds in smaller cities likely have a higher exchange of me� mes with those in areas outside of the city. In these reduced urban areas, even though great tits are under high anthropogenic acoustic pressure, the minimum frequency shift could be partially constrained due to cultural wealth with a high proportion of low frequency memes from outside the city. It could be that in this situation, great tits sing the same memes louder or shift other frequency features of their songs to enhan� ce sound transmission. Further studies comparing city and forest pairs in different sized cities with different city–forest distances may help to clarify these findings. Acknowledgments We thank the board of the Centro Quintos de Mora and the Council of San Pablo de los Montes for the facilities offered during the field work. J. Bueno was supported by a predoctoral fellowship (Junta de Comunidades de Castilla–La Mancha, European Social Fund). This study was funded by grants POIC10–0269–7632 (Junta de Comunidades de Castilla–La Mancha, Eu� ropean Social Fund) and GCL2010–21933–C02–01 (Ministerio de Ciencia e Innovación). References Baker, M. C., 2006. Differentiation of mating vocaliza� tions in birds: Acoustic features in mainland and island populations and evidence of habitat–depend� ent selection on songs. Ethology, 122: 757–771. Baker, M. C. Esther, M. B. & Merrill, S. A. B., 2001. Island and island–like effects on vocal repertoire of singing honeyeaters. Animal Behaviour, 62: 767–774. Baker, M. C. & Gammon D. E., 2008. Vocal memes in natural populations of chickadees: why do some memes persist and others go extinct? Animal Behaviour, 75: 279–289. Bates, D. Maechler, M. Bolker, B. M. & Walker, S., 2014. lme4: Linear mixed–effects models using Eigen and S4. Journal of Statistical Software. Url: http://arxiv.org/abs/1406.5823. Bermúdez–Cuamatzin, E., Rios–Chelen, A. A. Gil, D. & Garcia, C. M., 2009. �� Strategies of song adapta� tion to urban noise in the house finch: syllable pitch plasticity or differential syllable use? Behaviour, 146: 1269–1286. – 2011. Experimental evidence for real–time song frequency shift in response to urban noise in a passerine bird. Biology Letters, 7: 36–38. Brumm, H., 2004. The impact of environmental noise on song amplitude in a territorial bird. Journal of Animal Ecoogy, 73: 434–440. – 2006. Animal Communication: City Birds Have

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Use of water troughs by wild rabbits (Oryctolagus cuniculus) in a farmland area of north–west Spain J. A. Armenteros, C. Sánchez–García, M. E. Alonso, R. T. Larsen & V. R. Gaudioso

Armenteros, J. A., Sánchez–García, C., Alonso, M. E., Larsen, R. T. & Gaudioso, V. R., 2015. Use of water troughs by wild rabbits (Oryctolagus cuniculus) in a farmland area of north–west Spain. Animal Biodiversity and Conservation, 38.2: 233–240. Abstract Use of water troughs by wild rabbits (Oryctolagus cuniculus) in a farmland area of north–west Spain.— Installation of water troughs is a common approach to increase densities of small game species in the Iberian peninsula but little is known about the watering patterns of target species, such as the wild rabbit (Oryctolagus cuniculus). Using camera trapping, we monitored the use of water troughs by wild rabbits over 228 weeks in three consecutive periods, from June to October in 2008, 2009 and 2010, on farmland in north–west Spain. Wild rabbits used 43% of the water troughs. A significantly higher number of rabbits were observed drinking at troughs surrounded by shrub cover than at those in open fields. Most drinking events were recorded from July to September (98%), though the use of water troughs was not clearly related to weather. Wild rabbits drank mainly during the morning (52% of rabbits), less so in the evening and at night, and rarely in the afternoon. Wild rabbits were photographed together with red–legged partridges (Alectoris rufa) in 6% of photographs. These findings suggest water troughs are useful for species such as wild rabbits and should be allocated close to shrub areas. Key words: Camera traps, Cover, Game management, Water trough, Wild rabbit Resumen Uso de bebederos por parte del conejo de monte (Oryctolagus cuniculus) en un paisaje agrícola en el noroeste de España.— En la península Ibérica, los bebederos son una herramienta de gestión de hábitat muy frecuente para incrementar las densidades de especies de caza menor, aunque el comportamiento de ingestión de agua de las especies "diana" no se ha estudiado en profundidad, como es el caso del conejo de monte (Oryctolagus cuniculus). Estudiamos el uso de bebederos por parte de conejos de monte durante 228 períodos de una semana en tres períodos consecutivos (junio–octubre) de 2008, 2009 y 2010 en un paisaje agrícola de noroeste de España, utilizando cámaras de fototrampeo. Los conejos utilizaron el 43% de los bebederos y se observó un número significativamente mayor de conejos bebiendo en bebederos rodeados por cobertura vegetal en comparación con bebederos situados en campos abiertos sin dicha cobertura vegetal. La mayoría de los conejos que bebieron fueron fotografiados de julio a septiembre (98%), si bien la utilización de bebederos no se relacionó claramente con la climatología. Los conejos bebieron principalmente durante la mañana (52% de los conejos) no tanto durante la tarde y noche, y raramente durante el mediodía. Los conejos se fotografiaron junto con perdices rojas (Alectoris rufa) en el 6% de las fotografías. Estos hallazgos sugieren que los bebederos son útiles para el conejo y otras especies con necesidades hídricas similares y que debieran ser colocados cerca de zonas con cobertura vegetal arbustiva. Palabras clave: Cámaras de fototrampeo, Cobertura vegetal, Gestión cinegética, Bebedero, Conejo de monte Received: 5 V 15; Conditional acceptance: 9 VII 15; Final acceptance: 31 VIII 15 José A. Armenteros, Marta E. Alonso & Vicente R. Gaudioso, Grupo de Producción y Gestión Cinegética, Depto. Producción Animal, Univ. de León, 24071, León, Spain.– Carlos Sánchez–García, The Game & Wildlife Conservation Trust, SP6 1EF, Fordingbridge, UK.– Randy T. Larsen, Dept. of Plant and Wildlife Sciences and the Monte L. Bean Life Science Museum, Brigham Young Univ., 407 WIDB Provo, Utah 84602, USA. Corresponding author: C. Sánchez–García. E–mail address: dp2csg@gmail.com ISSN: 1578–665 X eISSN: 2014–928 X

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Introduction During the last decades, considerable efforts have been made in the Iberian Peninsula to improve the management of small game species. These animals play a key role in ecosystems and hunting such species is a crucial economic activity in many rural areas (Arroyo et al., 2012; Ferreira et al., 2014). One of the most commonly used approaches to preserve populations of small game species in Spain is the installation of water troughs (Arroyo et al., 2012; Ferreira et al., 2014), defined as free–standing water supplied artificially for wildlife. One key–stone species in the Mediterranean Basin is the wild rabbit (Oryctolagus cuniculus, hereafter rabbit), (Delibes–Mateos et al., 2008). Although rabbits can endure extended periods of drought (Hayward, 1961; Cooke, 1982) and distance to drinking water does not seem to affect their abundance in any season in Central Spain (Rueda et al., 2008), their reproduction has been correlated with environmental temperature and water content of the vegetation (Gonçalves et al., 2002). Furthermore, many gamekeepers and wildlife managers claim that rabbits frequently use water troughs when water is scarce. In the past, water troughs mainly targeted red–legged partridge (Alectoris rufa) populations and they were not aimed at rabbits as water availability is not generally considered a major constraint for this species (Hayward, 1961). However, in their study about the frequency of use and cost–effectiveness of habitat management techniques in rabbit populations in Iberia, Ferreira et al. (2014) found that water troughs also targeted rabbits. They suggested that supplying water would maximise returns from a set budget. Studies in the red–legged partridge have shown that water troughs surrounded by shrub cover are used to a greater extent than those in open fields. It has also been observed that weather plays a role, with a higher number of visits to troughs in summer months (Sánchez–García et al., 2012b). A better understanding of rabbit watering patterns could optimize the installation and use of water troughs, not only for the benefit of endangered predators that rely on rabbits, such as the Iberian lynx (Lynx pardinus) (Ferreira et al., 2010), but also for reared rabbits released for both hunting and conservation purposes (Sánchez–García et al., 2012a). Aiming to produce guidelines for best watering practices for small game species, we evaluated the use of water troughs by rabbits to (1) confirm their use by this species, (2) to investigate the effects of trough location in shrub cover or open fields on use by rabbits, (3) to assess whether drinking behaviour is affected by weather, and (4) to study the daily watering patterns of the species. Material and methods Study area This study was conducted from late spring to early autumn (June–October) over three consecutive

years (2008/09/10) on private farmland of 308 ha in the province of Valladolid in north–west Spain (lat 41° 53′ 45″ N, long 4° 52′ 50″ W, 'Finca Coto Bajo de Matallana'). The area is a typical pseudo–steppe of northern Spain with mostly flat terrain (altitude range 771–820 m a.s.l.). The climate is dry Continental Mediterranean, characterized by harsh winters and hot and dry summers, with an annual mean rainfall of 455 mm (AEMT, 2014). Historically, extended periods of drought have been recorded in the area from June to September. The study area has two small streams but these dry up in summer months and only flow after occasional storms. Shrub areas accounted for 38% of the area, cultivated fields for 37%, arboreal plants for 23%, and the remaining 2% for uncultivated areas, farm buildings and tracks. The main cultivated species were sunflowers (Helianthus annuus), lucerne (Medicago sativa) and winter cereal (mainly barley, Hordeum distichon). Game management The area was actively managed for small game densities and other wildlife from 1996 to 2008. A full time gamekeeper was employed and hunting was not allowed. Legal control of predators was conducted all year round following regional law —especially during the game bird breeding season (February–May)— and included magpies (Pica pica, L.), carrion crows, (Corvus corone L.), foxes (Vulpes vulpes, L.), feral dogs (Canis lupus familiaris, L.) and cats (Felis catus, L.), and brown rats (Rattus norvegicus, L.). Fifteen strips of mixes of barley, wheat (Triticum spp. L.), lucerne and common vetch (Vicia sativa L.) (0.2 ha average size), were distributed throughout the area. Management practices included the restricted use of herbicides, no livestock grazing at potential nesting habitats for farmland birds, and harvesting delays until early June and only in daytime. Additionally, 16 feeding troughs with wheat grain were distributed every 15–20 ha in autumn and winter. Due to repeated outbreaks of myxomatosis and rabbit haemorrhagic disease, numbers of rabbits were dramatically reduced in late 1980s (Olmedo, pers. comm.). A re–establishment programme was carried out from 1996 to 2002, using translocated rabbits established in artificial warrens (Díez, 2005; Fernández–Olalla et al., 2010). In 2008, 27 active warrens were detected and rabbits occupied 70% of the study site. Using the Kelker method, rabbit density (rabbits/ha) was estimated at 2.1 in 2008, 3.4 in 2009 and 1.36 in 2010 (Lacasa et al., 2010). Experimental design We used the methodology established in a previous project on water troughs for small game species (Lacasa et al., 2010). Water was supplied from a fibre cement water tank connected to the concrete water troughs by means of plastic pipes. The tank was refilled at the beginning of each summer (400 l of capacity). All troughs were surrounded by a 1.5 m

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2 3 4

Fig. 1. A. Example of a 'protected' water trough: 1. Metal fence; 2. Trough; 3. Plastic pipe; 4. Water tank; and 5. Camera trap. B. Images of three rabbits drinking at the same time; C. Two drinking rabbits (possibly one juvenile); and D. A rabbit and a red–legged partridge. Fig. 1. A. Ejemplo de un bebedero con cobertura vegetal: 1. Valla metálica; 2. Bebedero propiamente dicho; 3. Tubería plástica; 4. Depósito de agua; 5. Cámara de fototrampeo. B. Imágenes de tres conejos bebiendo al mismo tiempo; C. Dos conejos bebiendo (posiblemente uno de ellos juvenil); D. Un conejo y una perdiz roja.

high metal fence with 20 cm2 entrances, a design shown to have no effect on use by small game species (Lacasa et al., 2010) (fig. 1). For practical reasons, the location of troughs was not changed during the study periods. The estimated cost of the water supply, not including maintenance, was 3.5 €/ha/year. We studied five pairs of water troughs (total 10 troughs), located in five plots of 50–60 ha. All troughs were placed at a minimum distance of 500 m from the streams. Aiming to assess whether shrubs surrounding water troughs had any effect on the number of rabbits using them, in each plot we placed one of the water troughs among shrub cover (referred to as protected troughs) and the other at a distance of 50 m in an open field (referred to as open troughs), not surrounded by shrub cover. At the protected troughs, the shrubs were kept at a distance of 3–5 m. Average shrub and tree height across plots was 2.2 m ± 0.8 SE. Specific species were brambles (Rubus spp.), broom (Cytisus spp.) and pine trees (Pinus spp.). To assess the effects of weather on weekly drinking patterns of rabbits, data collection was carried out from June 15th to October 10th of each study year. Based

on data from an on–site weather station (Urbaso S.L., Spain), this was the driest period of the year, with limited rainfall. Maximum temperatures were above 35ºC, and relative humidity was under 60%, though conditions were hottest and driest in July and August. The water troughs were designed for use by the main small game species in the Iberian peninsula: red–legged partridge, rabbit and Iberian hare (Lepus granatensis). As red–legged partridges and rabbits were present in all the plots during the study period (Iberian hares were found at much lower densities), and as previous research has shown small game species use the same water troughs (Lacasa et al., 2010), we expected partridges and rabbits to be photographed together at the plots and possibly drinking at the same time. Data collection We used digital motion–sensing cameras (Bushnell Trailscout Pro©, Bushnell Trophy©) to photograph wildlife visiting water troughs. We started with six cameras in 2008, but had eight in place in 2009 and

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10 in 2010. During the first two years (2008 and 2009), cameras were alternated between plots, sampling a total of 8 and 10 camera trapping weeks per trough, respectively. In 2010, the troughs were monitored over 20 camera trapping weeks. We thus aimed to monitor a total of 380 camera trapping weeks (80 in 2008, 100 in 2009 and 200 in 2010). The two water troughs at each plot were monitored simultaneously. Cameras were placed at a distance of 2–3 m from the trough, and correct camera triggering was ensured. The time lapse between consecutive photographs was 30 seconds. Rabbits were not marked, so although it was assumed that the same animals would be photographed more than once, each photograph was considered a separate event. Time and date stamps, total number of rabbits, and drinking behaviour of each individual were recorded for each photograph. In a previous study conducted in captive–reared rabbits at the same study site, it was observed that daily and seasonal activity patterns were affected by air temperature (ºC), relative humidity (%) and wind speed (meters/second) (Díez, 2005). In the present study, we therefore downloaded weekly average values of these variables together with cumulative rainfall from the weather station located in the study area (www.fieldclimate.com) to assess their influence on the use of water troughs. Statistical analysis The unit of analysis used was the number of rabbits photographed at drinking each trough per week (hereafter camera–trapping week). As we expected rabbits to visit the plots but not to use the water troughs, we considered that the trough was used when any number of rabbits (one or more) was photographed drinking. Although rabbits were subject to different weather conditions as summer progressed, the same water troughs at similar locations were studied each year, so we were unable to rule out the possibility that the same rabbits visited the troughs. Hence, the x2–test, (Canavos, 1986) was used to test whether location of water troughs had an effect on the number of rabbits photographed drinking. We pooled data from all years from water troughs where rabbits were photographed to assess whether the number of rabbits photographed drinking per trough per week was related to mean values of air temperature, relative humidity, wind speed or cumulative rainfall. To do this, we used Pearson’s correlation coefficient. We used sunrise and sunset times of the central date (day 15) for each month and grouped photographs of drinking rabbits into four periods of the day: two in the morning (three hours before sunrise and three hours after sunrise, six hours in total), and two in the evening (three hours before sunset and three hours after sunset, six hours in total). Hence, afternoon and night periods were defined as the remaining time between morning and evening periods of time (Lee et al., 2010). To assess the possible effects of the time of day on the number of rabbits photographed drinking, we fitted a GLMM (Agresti,

2007) of the number of rabbits photographed drinking per month, with Poisson error, logarithmic link, ln(possible number of photographs taken for each period of the day), month, period of time (morning, afternoon, evening and night) as fixed factors and year as random factor. Differences with p < 0.05 were considered significant and all tests were carried out using SPSS© (v. 17.0 for WINDOWS©, IBM Corporation©). Results After subtracting incidents of cameras and trough malfunctioning (n = 152), we monitored troughs during 228 camera–trapping weeks, with a total of 1,546 camera–trapping days. The most common problems were camera malfunctions, photographs not valid for analysis (over–exposure), failures in the water supply, damage to a water system by wild boar (Sus scrofa) or stray dogs (Canis lupus familiaris), and combinations of these problems. Rabbits were photographed during 134 camera–trapping weeks (59%, n = 228), taking 3,359 photographs depicting 5,738 rabbits (table 1). From these, 599 rabbits were photographed drinking in 57 camera–trapping weeks (43%). These rabbits were mainly single individuals (95% of photographs), although we also observed rabbits in pairs (4.9% of photographs) and one group of three rabbits (0.1% of photographs) (fig. 1). The number of rabbits photographed drinking was significantly higher at protected troughs than in open fields in all years (p < 0.001; table 1). In 2008 and 2010, no rabbits were photographed drinking at open troughs. In protected troughs, 10.5% of the rabbits were seen drinking, while in open troughs, 0.03% of rabbits were seen drinking (table 1). Owing to the few rabbits photographed at troughs in open fields (n = 4), we pooled all these troughs for the remaining analysis. The number of rabbits photographed drinking per trough per week was not significantly correlated to any of the weather variables recorded: air temperature (r = 0.08, p = 0.38), relative humidity (r = – 0.13, p = 0.15), wind speed (r = – 0.03, p = 0.67), or rainfall (r = 0.02, p = 0.75). No month*period interaction effects were found for the number of rabbits photographed drinking per week (table 2). When pooling all years, drinking rabbits were photographed from July to September (98% of the rabbits), though a higher number were photographed in August (54%). Morning was the period of the day with the highest number of rabbits photographed drinking (52%), followed in order by evening (20%), night (19%) and afternoon (9%) (fig. 2). We observed rabbits and red–legged partridges together in 202 photographs (2008, n = 35; 2009, n = 55; 2010, n = 112), drinking at the same time on three occasions (fig. 1). Most of these photographs (88%) were taken in the morning and early afternoon (7 am–12 am). The remaining photos (12%) were taken in the afternoon–evening (2 pm–8 p m). No photographs of Iberian hares were taken.

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Table 1. Number of camera–trapping weeks (N), photos, rabbits, drinking rabbits per year and location, and comparison of the number of drinking rabbits photographed between open and protected water troughs using the x2–test. (The number of camera–trapping weeks in which rabbits were photographed is given in brackets.) Tabla 1. Número de períodos de una semana (N), fotos, conejos, conejos fotografiados bebiendo por año y tipo de localización junto con la comparación del número de conejos fotografiados bebiendo entre los bebederos situados a campo abierto y los rodeados por cobertura vegetal utilizando el test x2. (El número de períodos de una semana en los que se fotografiaron conejos está entre paréntesis.)

Number of Number of rabbits photographed drinking rabbits photographed

Year

Photos

Open

Protected

2008

42 (24)

269

1,042

110 (58)

716

1,446

76 (52)

2,374

228 (134)

3,359

120

138

135

x = 21.31, p < 0.001

3,130

322

x2 = 11.19, p < 0.001

2010

Protected

x2 = 15.55, p < 0.001

2009

Open

5,618

595

Discussion We confirmed that rabbits tended to use protected troughs over those in open spaces. Use of troughs was highest from July to September, and most photographs were taken during the morning. Although rabbits survive conditions of water restriction (Hayward, 1961; Cooke, 1982), we observed drinking

behaviour in 43% of the troughs where rabbits were detected. Our results confirm moderate use of troughs by rabbits, a finding in agreement with a previous study in the same area using different methodology (Lacasa et al., 2010). However, we did not study rabbit activity at plots without troughs so we are unable to determine whether water was the main reason for visiting the plots. At troughs

Nº of drinking rabbits photographed (± SE)

120 100

80 60 40 20 0

June Morning

July

August

Afternoon

September

Evening

October

Night

Fig. 2. Mean number of drinking rabbits photographed (± SE) per month during the four periods of the day (morning, afternoon, evening and night). Fig. 2. Media del número de conejos fotografiados bebiendo (± ES) por mes durante los cuatro períodos del día estudiados (mañana, mediodía, tarde y noche).

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Table 2. Results of the GLMM models explaining variation in number of drinking rabbits photographed in relation to month, to period of the day, and to interaction between month and period of the day: W. Wald statistic; F. F statistic. Tabla 2. Resultados de los modelos lineales generalizados mixtos para explicar la variación en el número de conejos fotografiados bebiendo en relación al mes, al período del día y a la interacción del mes con el período del día: W. Estadistico W; F. Estadistico F. Fixed term

Month

46.91

F4, 38 = 11.73 < 0.001

Period

30.97

F3, 38 = 10.32 < 0.001

Month*period

12.76

F12, 38 = 1.06 0.416

where no drinking behaviour was recorded, rabbits may have been attracted by other reasons, such as cooler temperatures around shrub cover. Our results should be considered with caution as the location of troughs did not change during the summer. Further studies are needed to address individual drinking patterns and to evaluate drinking behaviour at Iberian locations with a different climate. As expected, most drinking rabbits were detected at protected troughs (only two open troughs were used), also in agreement with the previous study conducted in the same area (Lacasa et al., 2010). These findings suggest that protected troughs offer safer conditions for watering, in accordance with the anti–predator strategy of this species (Moreno et al., 1996). The number of rabbits drinking at open troughs was very low, suggesting rabbits prefer closed troughs more than other trough visitors such as red–legged partridges and chukar (Alectoris chukar) partridges (Larsen et al., 2007; Larsen et al., 2009; Sánchez– García et al., 2012b). These findings suggest the location of troughs should therefore be considered in management strategies targeting rabbits. Our results are in line with other studies showing the high need for cover and refuge for this species (Moreno et al., 1996; Lombardi et al., 2007). The number of rabbits photographed drinking per trough per week was not clearly associated with the weather variables studied, though the number of drinking rabbits was highest in August. It may be that rabbits did not need water when food moisture content met their needs, a notion demonstrated in small species of birds (Degen et al., 1984). We did not study food moisture, however, and this is one of the main limitations of this study. It is likely that the strips of game crops distributed throughout the property and delayed crop harvesting (end of June) resulted in quality herbaceous communities with adequate water content. This might explain the higher

number of drinking rabbits photographed from July to September, especially in August, when food moisture may have been depleted. We cannot rule out the possible effects of different rabbit density and activity across months on the number of rabbits observed drinking, but it is known that rabbits reduce breeding activity in summer (Gonçalves et al., 2002) and the activity rate observed in captive–reared rabbits at the same site was very low during the second half of the year (Díez et al., 2013). Our study was carried out from late spring to early autumn only. It could be of interest for future research to investigate water needs throughout the whole year. Rabbits drank mainly during the morning in all months, followed by evening and night drinking behaviour recorded during the afternoon was very low. The drinking pattern was similar to the general activity pattern observed in wild populations (Villafuerte et al., 1993) and in the pattern for reared rabbits in the same area (Díez, 2005), both of which showed lower drinking activity in the afternoon and evening. Rabbits and other lagomorph species are known to be inactive during harsh climatic conditions (Mykytowycz & Rowley, 1958; Wallage–Drees, 1989), so it is possible that rabbits were reluctant to use the troughs during the hottest and driest periods of the day, which in our study site were late morning to early evening. Rabbits may have conducted other activities in the evening and at night, such as foraging (Mykytowycz & Rowley, 1958; Villafuerte et al., 1993), watering then during the morning before resting. Neither can we rule out effects of the predator community on rabbit watering behaviour (at the study site mainly foxes and raptors, see Lacasa et al., 2010). In the grey partridge (Perdix perdix), the use of feeders is higher at dusk and dawn when diurnal and nocturnal predators are respectively less active or not yet active (Potts, 2012). As expected, we photographed rabbits and red– legged partridges around troughs at the same time, mainly during the morning and early afternoon, the time of the day when partridges concentrate their visits to the troughs (Sánchez–García et al., 2012b). The small number of photographs showing the two species drinking simultaneously could be attributed to the size of the trough, which allowed a limited number of animals at the same time. We did not observe any antagonistic behaviour between species and in situ observations suggest that rabbits and partridges shared the same trough. Further studies at other sites using models with a larger water area (such as small ponds) are needed to confirm the simultaneous use of watering sites rabbits and partridges. Owing to rabbits’ frequent use of water troughs in our study and the possible effectiveness of this strategy when compared to other techniques (Ferreira et al., 2014), game managers and practitioners aiming to favour rabbit populations through water supply should allocate troughs close to shrub cover and ensure that water is supplied during the summer. For management or research purposes, human visits to water troughs should be conducted between 11 and 16 h in areas with a similar climate as this is the period of the day when rabbits are less active and drink less.

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The use of water troughs for wildlife has been widely questioned as it can be viewed as a disturbance of natural ecological processes, but available research does not always support negative effects (Simpson et al., 2011). Although water troughs may help to increase rabbit density, further research is needed to understand the effects of water supply on rabbit ecology and to determine the factors triggering the use of troughs by rabbits (such as lack of food moisture). Such knowledge could help adapt management decisions to different scenarios. Acknowledgements This study is part of the first author´s doctoral thesis. We wish to thank the Junta de Castilla y León, Excma. Diputación de León and Caja España for partial funding of the field study. We also thank J. A. Olmedo, Kaat Brulez, the Excma. Diputación de Valladolid and Fundación Caja Madrid. Special thanks are extended to the reviewers for their comments, E. J. Tizado for statistical advice and J. M. Lomillos and F. Escalera for their patience and sense of humour during data collection. References AEMT, 2014. Valores climatológicos aeropuerto de Valladolid 1971–2000. Available at http://www.aemet.es/es/serviciosclimaticos/datosclimatologicos/ valoresclimatologicos?l= 539&k=cle [Accessed on July 2015]. Agresti, A., 2007. An introduction to categorical data analysis. Wiley & Sons, Inc., New Jersey. Arroyo, B., Delibes–Mateos, M., Díaz–Fernández, S. & Viñuela, J., 2012. Hunting management in relation to profitability aims: red–legged partridge hunting in central Spain. European Journal of Wildlife Research, 58: 847–855. Canavos, G. C., 1986. Probabilidad y estadística. Aplicaciones y Métodos. Ed. Mc–Graw Hill, Madrid. Cooke, B., 1982. Reduction of food intake and other physiological responses to a restriction of drinking water in captive wild rabbits Oryctolagus cuniculus (L.). Australian Wildlife Research, 9: 247–252. Degen, A., Pinshow, B. & Shaw, P., 1984. Must desert chukars (Alectoris chukar sinaica) drink water? Water influx and body mass changes in response to dietary water content. Auk, 101: 47–52. Delibes–Mateos, M., Delibes, M., Ferreras, P. & Villafuerte, R., 2008. Key Role of European rabbits in the conservation of the Western Mediterranean Basin hotspot. Conservation Biology, 22: 1106–1117. Díez, C., 2005. Análisis de los biorritmos de actividad del conejo de monte (Oryctolagus cuniculus. L.) en condiciones de semi–libertad y su relación con determinados parámetros ambientales. Ph D Thesis, University of León, Spain. Díez, C., Sánchez–García, C., Pérez, J. A., Bartolomé, D. J., González, V., Wheatley, C., Alonso, M. E. & Gaudioso, V. R., 2013. Behavioural activity of wild

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rabbits (Oryctolagus cuniculus) under semi–natural rearing systems: establishing a seasonal pattern. World Rabbit Science, 21: 263–270. Fernández–Olalla, M., Martínez–Jauregui, M., Guil, F. & San Miguel–Ayanz, A., 2010. Provision of artificial warrens as a means to enhance native wild rabbit populations: what type of warren and where should they be sited? European Journal of Wildlife Research, 56: 829–837. Ferreira, C. & Delibes–Mateos, M., 2010. Wild rabbit management in the Iberian Peninsula: state of the art and future perspectives for Iberian lynx conservation. Wildlife Biology in Practice, 6: 48–66. Ferreira, C., Touza, J., Rouco, C., Díaz–Ruiz, F., Fernandez–de–Simon, J., Ríos–Saldaña, C. A., Ferreras, P., Villafuerte, R. & Delibes–Mateos, M., 2014. Habitat management as a generalized tool to boost European rabbit Oryctolagus cuniculus populations in the Iberian Peninsula: a cost–effectiveness analysis. Mammalian Review, 44: 30–43. Gonçalves, H., Alves, P. C. & Rocha, A., 2002. Seasonal variation in the reproductive activity of the wild rabbit (Oryctolagus cuniculus algirus) in a Mediterranean ecosystem. Wildlife Research, 29: 165–173. Hayward, J. S., 1961. The ability of the wild rabbit to survive conditions of water restriction. CSIRO Wildlife Research, 6: 160–175. Lacasa, V. R., García–Abad, C., Martín, R., Rodríguez, D. J., Garrido, J. A. & De La Varga, M. E., 2010. Small game water troughs in a Spanish agrarian pseudo steppe: visits and water site choice by wild fauna. European Journal Wildlife Research, 56: 591–599. Larsen, R. T., Bissonette, J. A., Flinders, J. T., Hooten, M. B. & Wilson, T. L., 2009. Summer spatial patterning of chukars in relation to free water in western Utah. Landscape Ecology, 25: 135–145. Larsen, R. T., Flinders, J. T., Mitchell, D. L., Perkins, E. R. & Whiting, D. G., 2007. Chukar watering patterns and water site selection. Rangeland Ecology and Management, 60: 559–565. Lee, J. E., Larsen, R. T., Flinders, J. T. & Eggett, D. L., 2010. Daily and seasonal patterns of activity at pygmy rabbit burrows in Utah. Western North American Naturalist, 70: 189–197. Lombardi, L., Fernández, N. & Moreno, S., 2007. Habitat use and spatial behaviour in the European rabbit in three Mediterranean environments. Basic Applied Ecology, 8: 453–463. Moreno, S., Delibes, M. & Villafuerte, R., 1996. Cover is safe during the day but dangerous at night: the use of vegetation by European wild rabbits. Canadian Journal of Zoology, 74: 1656–1660. Mykytowycz, R. & Rowley, I., 1958. Continuous observations of the activity of the wild rabbit, Oryctolagus cuniculus (L.), during 24 hour periods. CSIRO Wildlife Research, 3: 26–31. Potts, G. R., 2012. The Partridges: Countryside Barometer. Collins, London. Rueda, M., Rebollo, S., Gálvez–Bravo, L. & Escudero, A., 2008. Habitat use by large and small herbivores in a fluctuating Mediterranean ecosystem: Implications of seasonal changes. Journal of Arid

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Environment, 72: 1698–1708. Sánchez–García, C., Alonso M. E., Díez, C. & Gaudioso, V. R., 2012a. An approach to the statistics of wild lagomorph captive rearing for releasing purposes in Spain. World Rabbit Science, 20: 49–56. Sánchez–García, C., Armenteros, J. A., Alonso, M. E., Larsen, R. T., Lomillos, J. M. & Gaudioso, V. R., 2012b. Water–site selection and behaviour of red–legged partridge Alectoris rufa evaluated using camera trapping. Applied Animal Behaviour Science, 137: 86–95.

Simpson, N. O., Stewart, K. M. & Bleich, V. C., 2011. What have we learned about water developments for wildlife? Not enough! California Fish and Game, 97(4): 190–209. Villafuerte, R., Kufner, M. & Delibes, M., 1993. Environmental factors influencing the seasonal daily activity of the European rabbit (Oryctolagus cuniculus) in a Mediterranean area. Mammalia, 57: 341–347. Wallage–Drees, J., 1989. A field study on seasonal changes in the circadian activity of rabbits. Zeitschrift fuer Saeugetierkunde, 54: 22–30.

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A new species of Bryconamericus (Characiformes, Stevardiinae, Characidae) from the Pacific coast of northwestern Ecuador, South America C. Román–Valencia, R. I. Ruiz–C., D. C. Taphorn B., P. Jiménez–Prado & C. A. García–Alzate

Román–Valencia, C., Ruiz–C., R. I., Taphorn B., D. C., Jiménez–Prado, P. & García–Alzate, C. A., 2015. A new species of Bryconamericus (Characiformes, Stevardiinae, Characidae) from the Pacific coast of northwestern Ecuador, South America. Animal Biodiversity and Conservation, 38.2: 241–252. Abstract A new species of Bryconamericus (Characiformes, Stevardiinae, Characidae) from the Pacific coast of northwestern Ecuador, South America.— A new species of Bryconamericus (Characiformes, Characidae, Stevardiinae) is described from the Pacific coast of northwestern Ecuador, South America. The new species is distinguished from all congeners by the presence in males of bony hooks on the caudal fin rays (vs. absence). The different layers of pigment that constitute the humeral spots have differing degrees of development and structure that are independent of each other. Brown melanophores are distributed in a thin, vertical, superficial layer of the epithelium (layer 1) and in another deeper (layer 2) that overlaps the first and is centered over the lateral–line. B. ecuadorensis has a horizontally oval or elliptical shape layer 2 pigment in the anterior humeral spot (vs. a rectangular or circular layer 2). The new species further differs in having an anterior extension of the caudal peduncle spot (vs. no anterior extension of the caudal peduncle spot) and by having a dark lateral stripe overlaid by the peduncular spot and by the regularly distributed pigmentation on scales on the sides of the body (vs. peduncular spot and other body pigments not superimposed over a dark lateral stripe). Hooks present on all fins of males (vs. hooks present only on anal and pelvic fins of males) distinguishes the new species from B. dahli, the only sympatric congener. Seven other diagnostic characters separating the new taxon from B. dahli are reported. We also include physical, chemical and biological habitat parameters and analyse the impacts from mining on this new species and other organisms present at the type locality. Key words: Conservation, Taxonomy, Tropical fish, New taxon, Bryconamericus ecuadorensis n. sp. Resumen Una nueva especie de Bryconamericus (Characiformes, Stevardiinae, Characidae) de la costa pacífica al noroeste de Ecuador, América del Sur.— Se describe una nueva especie de Bryconamericus (Characiformes, Stevardiinae, Characidae) de la costa Pacífica al noroccidente de Ecuador, América del Sur. La nueva especie se distingue de todos sus congéneres por la presencia en machos de espinas sobre los radios de la aleta caudal (vs. ausencia). Observamos que las diferentes capas de pigmentos que conforman la mancha humeral registran diferentes grados de desarrollo y estructura que son independientes una de otra. Los melanóforos marrones se distribuyen en una capa delgada, vertical superficial de epitelio (capa 1), y en otra capa más oscura y profunda (capa 2) centrada sobre el canal latero sensorial, ambas sobrepuestas. B. ecuadorensis tiene la capa 2 horizontalmente ovalada o de forma elíptica en la mancha humeral anterior (vs. capa 2 de la región humeral rectangular o circular). La nueva especie difiere también por tener una extensión anterior de la mancha en el pedúnculo caudal (vs. sin extensión anterior de la mancha peduncular), por presentar una banda lateral oscura sobrepuesta por la mancha peduncular y por una distribución regular de los pigmentos en las escamas de los lados del cuerpo (vs. mancha peduncular y otros pigmentos del cuerpo no superpuestos sobre la banda lateral oscura). La nueva especie se distingue de B. dahli, el único congénere simpátrico, por la presencia de ganchos en todas las aletas de los machos (vs. presencia de ganchos solamente en los radios de las aletas anal y pélvicas de los machos). Además, se reportan siete caracteres diagnósticos adicionales que separan el nuevo taxón de B. dahli. Se incluyen también datos sobre los parámetros físicos, químicos y biológicos del hábitat de la nueva especie y un análisis de los impactos causados por la minería sobre esta nueva especie y otros organismos que comparten su hábitat. ISSN: 1578–665 X eISSN: 2014–928 X

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Palabras clave: Conservación, Taxonomía, Pez tropical, Nuevo taxón, Bryconamericus ecuadorensis sp. n. Received: 26 II 15; Conditional acceptance: 11 V 15; Final acceptance: 15 IX 15 C. Román–Valencia, R. I. Ruiz–C., D. C. Taphorn B., P. Jiménez–Prado & C. A. García–Alzate, Lab. de Ictiología, Univ. del Quindío, A. A. 2639, Armenia, Quindío, Colombia.– D. C. Taphorn B., 1822 N. Charles St., Belleville, Illinois, USA.– P. Jiménez–Prado, Pontificia Univ. Católica del Ecuador sede Esmeraldas, espejo y subida a Santa Cruz.– C. A. García–Alzate, Univ. del Atlántico, Programa de Biología, km 7 antigua vía a Puerto Colombia, Barranquilla, Colombia. Corresponding author: C. Román–Valencia: ceroman@uniquindio.edu.co, zutana_1@yahoo.es

Animal Biodiversity and Conservation 38.2 (2015)

Introduction Bryconamericus found on the Pacific coasts of Central and South America are Bryconamericus brevirostris, B. bucayensis, B. dahli, B. emperador, B. guaytarae, B. miraensis, B. peruanus, B. simus and B. terrabensis (Román–Valencia, 2000, 2002, 2003; Román–Valencia et al., 2011, 2013; Román–Valencia & Vanegas–Ríos, 2009). �� From the Pacific versant of Ecuador, four spe� cies have been recorded: B. brevirostris, B. bucayensis, B. dahli and B. simus (Jiménez–Prado et al., 2015). The discovery of a new species of Bryconamericus from the Pacific versant of western Ecuador is a result of the ongoing revision of Bryconamericus (Román–Va� lencia, 2000, 2002; Román–Valencia et al., 2011, 2013, 2014) by the first author and further demonstrates the undocumented biodiversity of the genus. Material and methods Fish were collected using seine nets, preserved in the field in 10% formalin, and later stored in 70% ethanol. Measurements were taken using digital calipers, recorded to hundredths of millimeters and usually expressed as percentages of standard length (SL) or head length (HL) (table 1). Counts were made using a stereoscope with a dissection needle to extend the fins. Counts and measurements were taken from the left side of specimens when possible, according to guidelines in Vari & Siebert (1990). Observations of bones and cartilage were made on cleared and stained adult specimens (C&S) prepared according to techniques outlined in Taylor & van Dyke (1985) and Song & Parenti (1995). Bone nomenclature follows Weitzman (1962) and Vari (1995). In the lists of para� types, the number of individuals is given immediately after the catalog number, which is followed by the range of standard length in mm (SL) for each lot. For example: MEPN 4004 (2) 63.5–73.9 mm SL indicates two individuals in lot MEPN 4004. The smallest fish was 63.5 mm SL and the largest was 73.9 mm SL. In reporting counts, the values for the holotype are indicated with an asterisk (*). All collections were made in Ecuador. Acronyms used follow Sabaj–Pé� rez (2010) except CEMZ–p– (Pontificia Universidad Católica del Ecuador sede Esmeraldas, colección de peces). Meters above sea level is abbreviated as m a.s.l. Departamento is translated as Province. Municipio is translated as Canton. Comparative material Bryconamericus oroensis (see Román–Valencia et al., 2013) Bryconamericus dahli (see Román–Valencia, 2000; Román–Valencia et al., 2013), all from Ecuador: CEMZ–p–107 (3), Esmeraldas Province, San Lorenzo Canton, Wimbí community, 00° 57' 23.4'' N–78° 46' 17.9'' W, 5 VI 2011. CEMZ–p–220 (28), Esmeraldas Province, Atacames Canton, Las Mareas locality, 00° 50' 25.0'' N– 79° 50' 01.2'' W, V 2012. CEMZ–p–129 (41), Esmer� aldas Province, Atacames Canton, Las Brisas local�

243

ity, 00° 50' 31.1'' N–79° 51' 55.5'' W, VI 2012. CEMZ– p–309(48), Esmeraldas Province Atacames Canton, Súa locality, Súa River (medio), 00° 46' 43.4'' N– 79° 53' 26.6'' W, VII 2013. CEMZ–p–260 (25), Es� meraldas Province, Atacames Canton, Aguafría community, 00° 43' 08.0'' N–79° 51' 26.5'' W, VI 2012. CEMZ–p–303 (118), Esmeraldas Province, San Lorenzo Canton, Estero Sabalera, 01° 13' 56.5'' N– 78° 45' 20.8'' W, 21 VI 2013. CEMZ–p–165 (2), Esmeraldas Province, Atacames Canton, Boca de Tazones community, 00° 44' 35.5'' N–79° 50' 56.4'' W, 7 VII 2012. CEMZ–p–257 (72), Esmeraldas Province, Cumba community, 00° 48' 44.3'' N–79° 51' 02.8'' W, VI 2012. CEMZ–p–222 (6), Esmeraldas Province, 'Eloy Alfaro' Canton, Maldonado locality, 01° 04' 32.8'' N –78° 54' 30.3'' W, 21 m a.s.l., 15 III 2012. CEMZ–p–315 (37), Esmeraldas Province, Atacames Canton, Súa River (alto), Súa community, 00° 43' 00.1'' N–79° 53' 07.7'' W, VII 2013. CEMZ–p–101 (12), Esmeraldas Province, Muisne Canton, Mompiche River, near the community of the same name, 00° 29' 57.4'' N–80° 01' 00.5'' W, May 2012. CEMZ–p–151 (1), Esmeraldas Province, San Lorenzo Canton, Cayapas River, Zapallo Grande community, 00° 49' 35.9'' N–78° 55' 52.9'' W, X 2011. CEMZ–p–223 (18), Esmeraldas Province, Atacames Can� ton, Puente Taseche River, 00° 52' 16.7'' N–79° 49' 18.7'' W, VI 2012. CEMZ–p–215 (9), Esmeraldas Province, San Lorenzo Canton, Playa de Oro community, Concepción locality, 01° 02' 16.6'' N–78° 49' 51.6'' W, VI, X, XI 2011. CEMZ–p–217(113), Esmeraldas Province, Atacames Canton, La Unión locality, 00° 48' 48.0'' N–79° 52' 01.8'' W, VI 2012. CEMZ–p–153(12), Esmeraldas Province, ''Eloy Alfaro Canton'', Maldonado locality, 01° 04' 32.8'' N –78°�� 54'�� 30.3''�� W, 15 III 2012. CEMZ–p–216 �� (58), Es� meraldas Province, Muisne Canton, Mompiche River, near community of the same name, 00° 29' 53.0'' N –80° 00' 52.7'' W, V 2012.CEMZ–p–213 (27), Esmer� aldas Province, San Lorenzo Canton, San Javier de Cachaví locality, 00° 58' 05.1'' N–78° 39' 08.9'' W, V, X, XI 2011. CEMZ–p–182(8), Esmeraldas Prov� ince, San Lorenzo Canton, Selva Alegre local� ity, Santiago River, 00° 55' 53.1'' N–78° 51' 28.1'' W, II 2012. CEMZ–p–297(69), Esmeraldas Province, San Lorenzo Cantón, Estero Sabalera, 01° 13' 56.5'' N– 78° 45' 20.8'' W, 23 V 2013.CEMZ–p–115(2), Esmeral� das Province, San Lorenzo Cantón, Wimbí community, 00° 57' 23.4'' N–78° 46' 17.9'' W, 5 V 2011. CEMZ– p–169(16), Esmeraldas Province, Atacames Cantón, Repartidero locality, 00° 42' 22.6'' N–79° 51' 07.6'' W, VI 2012. CEMZ–p–205 (12), Esmeraldas Province, Atacames Canton, Playa Grande locality, 00° 46' 38.2'' N– 79°�� 49'�� 41.5''�� W, VI 2012. CEMZ–p–239 (36), Esme� raldas Province, Atacames Cantón, Puente Ataca� mes locality, 00° 51' 03.9'' N–79° 50' 56.2'' W, VI 2012. CEMZ–p–223 (22), Esmeraldas Province, Atacames Can� ton, Puentetaseche River, 00° 52' 16.7'' N–79° 49' 18.7'' W, VI 2012. CEMZ–p–134(14), Esmeraldas Province, Atacames Canton, Pato locality, 00° 44' 00.7'' N– 79° 50' 58.6'' W, VI 2012. IUQ 3804(1), Esmeraldas Province, Estero María. Canton 'Eloy Alfaro'. San Agustín community, close to the main road. 'Esme� raldas–San Lorenzo, 02° 21' 2,8'' N–78° 55' 21,21'' W; 21 m a.s.l., 29 IX 2014. �� IUQ3138 (1, C&S), Mata�

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jita River, Mataje tributary one half hour from the Mataje community center, on the road to Rio Mira and Pensamiento Lagoon, 800 m downstream, Pan River locality. IUQ 3140 (1 C&S), 58.32 mm SL, Esmeraldas Province, Pistalosa creek, half hour downstream of“Vargas Torres'', IV 1994. IUQ 3806 (33), Esmeraldas Province, Estero Las Antonias, 00° 58' 50,2'' N–78° 51' 56'' W, 28 m a.s.l., 27 IX 2014. IUQ 3808 (9), Esmeraldas Province, Santiago River, by Estero Las Antonias, 'Eloy Alfaro' Canton, 00° 58' 54,2'' N– 78° 51' 54,3'' W, 24 m a.s.l., 27 IX 2014.IUQ 3809 (51), Esmeraldas, Estero María. 'Eloy Alfaro' Canton, San Agustín community, on main road. Esmeral� das–San Lorenzo, 01° 02' 32,8'' N–78° 55' 21,2'' W; 21 m a.s.l., 27 IX 2014. IUQ 3811 (75), Esmeraldas, Sabalera River, Reserve La Chiquita, San Lorenzo Canton, 01° 14' 33,2'' N–78° 45' 05,5'' W, 36 m a.s.l. 27 IX 2014. Bryconamericus brevirostris (see Román–Valencia et al., 2011, 2013), B. bucayensis (see Román–Va� lencia et al., 2013). Bryconamericus simus (see Román–Valencia et al., 2013).

Results Bryconamericus ecuadorensis n. sp. (figs. 1–6; tables 1, 2) Holotype: IUQ 3813, male, 77.1 mm SL, Ecuador, Esmeraldas Province, Santiago River, 78° 51' 28.1'' W, 00° 55' 53.1'' N, 24 m a.s.l., VI 2014. Paratypes: MEPN 11129 (1), 62.8 mm SL, male, Santo Domingo de los Tsáchilas Province, Alluriquín River, km 28, south of Santo Domingo, 79º 08' 25'' W, 00º 18' 29'' N, 680 m a.s.l., 12 II 1985. MEPN 4004 (5) 63.5–73.9 mm SL, Esmeraldas Province, Estero Tatica, 2 km from Timbre community, 79º 37' 07'' W, 00º 50' 26'' N, 18 m a.s.l., 20 III 1985. MEPN 3979 (15), 60.5–70.9 mm SL, Esmeraldas Province, Estero Boca del Ónzole, right bank of Guayllabamba River, 450 m from Golondrina Hill, 14 III 1985. IUQ 3141 (1 C&S), 66.1 mm SL, Esmeraldas Province, Estero La Bocana del Cupa 100 m from Puerto Chupa Hill, 11 III 1985. IUQ �� 3136 (1 C&S), 54.9 mm SL, Esme� raldas Province, Estero Sabalera, 60 m from Chiquita camp, 22 X 1985. IUQ 3147 (1 C&S), 65.2 mm SL,

A 1 2

1 2 1

Fig. 1. Distribution of pigment in the humeral region of: A. Bryconamericus dahli, paratype IUQ 219 (80.1 mm SL); B. B. ecuadorensis n. sp. The anterior humeral spot consists of two overlapping layers of pigment (indicated by numbers 1 and 2) and a third configuration defined as the posterior humeral spot. Fig. 1. Distribución del pigmento en la región humeral de: A. Bryconamericus dahli, paratipo IUQ 219 (80,1 mm LS); B. B. ecuadorensis sp. n. La mancha humeral anterior consiste en dos capas superpuestas de pigmento (indicado con los números 1 y 2) y una tercera configuración definida como la mancha humeral posterior.

Animal Biodiversity and Conservation 38.2 (2015)

B 1 mm

Fig. 2. Bryconamericus ecuadorensis n. sp., IUQ 3141 (66.1 mm SL), upper and lower jaws: A. Premaxilla, B. Maxilla, C. Lower jaw. Fig. 2. Bryconamericus ecuadorensis sp. n., IUQ 3141 (66,1 mm LS), mandíbula superior e inferior: A. Premaxilar; B. Maxilar; C. Mandíbula inferior.

245

Esmeraldas Province, estero Tatica, 2 km from Tim� bre, 20 III 1985. MEPN 4023 (15), 31.7–65.3 mm SL, Esmeraldas Province, Estero Sabalera 600 m from La Chiquita forestry camp, 78º 45' 19'' W, 01º 13' 47'' N, 36 to 61 m a.s.l., 22 X 1985. MEPN 11155 (9), 51.2–83.2 mm SL, Santo Domingo de los Tsachilas Province, Río Alluriquín, km 28 on road to Santo Domingo, 79º 08' 25'' W, 00º 18' 29'' N, 680 m a.s.l., 12 II 1984. MEPN 87158 (20), 38.0– 69.4 mm SL, Esmeraldas Province, San Marcos Creek half hour from town, Mira–Mataje River drainage, 78º 32' 12'' W, 01º 20' 24'' N, 580 m a.s.l., 8 II 1987. MEPN 4042 (1), 51.2–83.2 mm SL, Esmeraldas Province, Estero la Comunidad up� stream of confluence with Ónzole River, tributary of the Cayapas River, 79º 00' 48'' W, 0º 57' 54'' N, 95 m a.s.l., 20 VIII 1985. MEPN 11131 (1), 75.4 mm SL, El Oro Province, Huertas locality, tributary of the Jubones River, 79º 40' 20'' W, 03º 35' 40'' S, 1,450 m a.s.l., 30 VIII 1978. IUQ 3142 (1 C&S), 54.7 mm SL, Esmeraldas Province, Estero Boca del Ónzole, right bank of Guayllabamba River, 14 III 1985. CEMZ–P–291 (4), 60.5–72.4 mm SL Esmeraldas Province, San Lorenzo, Estero Sabal� era, 78° 44' 59,26'' W, 01° 15' 39,87'' N, 23 V 2013. CEMZ–P–405 (14), 38.5–54.9 mm SL, Esmeral� das Province, Santiago River, 78° 51' 28.1'' W, 00° 55' 53.1'' N, 24 m a.s.l., VI 2014.

1 cm

Fig. 3. Sexual dimorphism of Bryconamericus ecuadorensis n. sp.: A. Holotype, male, IUQ 3813 (77.1 mm SL); B. Paratype, female. Fig. 3. Dimorfismo sexual de Bryconamericus ecuadorensis sp. n.: A. Holotipo, macho, IUQ 3813 (77,1 mm LS); B. Paratipo, hembra.

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Table 1. Morphometric and meristic data of Bryconamericus ecuadorensis n. sp. (standard length and total length in mm; SD. Standard deviation). Tabla 1. Datos morfométricos y merísticos de Bryconamericus ecuadorensis sp. n. (longitud estándar y longitud total en mm; SD. Desviación estándar). Morphometry Standard length Total length Percentages of SL Body depth Snout–dorsal fin distance Snout–pectoral fin distance Snout–pelvic fin distance Dorsal–pectoral fin distance Snout–anal fin distance Dorsal fin–hypural distance Dorsal–anal fin distance Dorsal–fin length Pectoral–fin length Pelvic–fin length Anal–fin length Caudal peduncle depth Caudal peduncle length Head length Percentages of HL Snout length Orbital diameter Postorbital distance Maxilla length Interorbital distance Meristics Lateral–line scales Scale rows between dorsal–fin origin and lateral line Scale rows between anal–fin origin and lateral line Scale rows between pelvic–fin and lateral line Predorsal median scales Dorsal–fin rays Anal–fin rays Pelvic–fin rays Pectoral–fin rays Teeth on maxilla

Diagnosis Bryconamericus ecuadorensis n. sp. is distinguished from congeners by the presence in males of bony

Paratypes (n = 90)

77.1 93.82

22.56–79.74 28.56–101.1

0.74 3.18

30.97 30.97 25.48 43.41 39.77 59.07 51.82 30.41 23.88 18.81 13.44 14.81 12.63 11.27 24.23

25.45–44.19 49.8–56.37 24.36–28.1 42.0–49.6 31.91–44.99 45.72–66.92 33.58–56.58 24.57–38.89 15.28–33.64 12.93–30.10 6.18–17.03 11.03–39.42 8.0–14.76 6.53–14.86 19.17–28.50

5.09 0.65 0.94 0.68 3.78 0.8 0.13 6.81 5 4.05 0.74 5.27 0.84 1.83 2.55

26.5 32.87 37.53 31.37 40.9

17.0–31.72 23.10–45.76 31.17–51.98 21.86–46.16 25.27–45.27

8.22 4.3 9.27 1.91 0.33

40 6

36–40 5–7

6 6 12 iii, 9 vi, 27 i, 7 i, 11, i

5–8 5–6 10–12 iii, 9 vi, 25–29 i,7 i, 10–12, i

1–3

Holotype

hooks on the caudal fin rays (vs. absent). The di� fferent layers of pigment that conform the humeral spot(s) have differing degrees of development and structure that are independent of each other. Brown

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Table 2. Physicochemical and biological variables in habitat of Bryconamericus ecuadorensis n. sp.: 1. Estero Sabalera; 2. Quebrada San Marcos; 3. Estero La Comunidad en boca del Onzole; 4. Rio Santiago; a Mosquera (2014); b Almeida (2014); c CID–PUCESE & PRAS–MAE (2012); d CID–PUCESE & PRAS–MAE (2014); e Texto Unificado de Legislación Ambiental Secundaria del Ministerio de Ambiente (TULAS): anexo 1, tabla 3; f Secretaria Nacional del Agua y PUCESE (2011). Tabla 2. Variables fisiccoquímicas y biológicas del hábitat de Bryconamericus ecuadorensis sp. n. (Para las abreviaturas, véase arriba.) Locality

3c,d

4c,d

m a.s.l.

580

Water temperature ºC

27.1

19.8

26.9

23.5

Natural condition + 3ºC

Air tempearture ºC

28.4

21.6

27.3

25.4

Natural condition + 3ºC

Dissolved oxygen (mg/l) pH Turbidity (NTU)

Limit allowed*

–

7.66

8.59

7.1

8.52

7.26

7.14

5a9

11.63

639

4.67

2,12

100

Conductivity (μs/cm)

430

29.5

No records

Total solid (mg/l)

–

288

5.33

No records

Visible light (cm)

–

27.5

No records

1.05

0.38

No records

Hardness

–

< 0.0001

–

0.13

0.08

0.1

–

0.19

0.09

0.05

–

0.04

1.18

0.21

0.02

–

0.62

0.05

0.01

–

16.3

0.99

0.13

0.3

Zinc

–

0.52

0.18

–

0.64

0.28

0.2

0.1

Fecal coliforms (nmp/100 ml)

3,500

79f

350f

200

melanophores are distributed in a thin vertical super� ficial layer of the epithelium (layer 1), and another deeper layer (layer 2, see fig. 1) centered over the lateral–line canal overlaps the first. In B. ecuadorensis the anterior humeral spot has a horizontally oval or elliptical shape for pigment layer 2 (vs. layer 2 of anterior humeral spot rectangular or circular) (fig. 1). The new species further differs by having an anterior extension of the caudal–peduncle spot (vs. no anterior extension of caudal peduncle spot) and a dark lateral stripe overlaid by the peduncular spot and by the regularly distributed pigmentation on scales on the sides of the body (vs. peduncular spot and other body pigments not superimposed over a dark lateral stripe), except for B. oroensis, from which it differs by the number of unbranched anal–fin rays (vi vs. iii–iv) and by the distribution and number of hooks on the anal and pelvic fin of sexually mature males. B. oroensis is found in the Amazon basin in the states of Loja, El Oro and Azyay (Román–Valencia et al., 2013).

< 0.0001 < 0.0001

0.0002

Description Table 1 shows morphometric and meristic data. Body somewhat elongate. Area above orbits flat. Dorsal profile of head and body oblique from supraoccipital tip to dorsal–fin origin and from last dorsal–fin ray to base of caudal fin. Ventral profile of body rounded from snout to base of anal fin. Caudal peduncle laterally compressed. Snout pointed. Head and snout short, jaws equal; mouth terminal, lips soft and flexible, covering the outer row of premaxilla teeth; ventral border of upper jaw not straight; posterior edge of maxilla reaching anterior edge of orbit; opening of posterior nostrils vertically ovoid; opening of anterior nostrils with membranous flap. First dorsal–fin ray vestigial not emerged from skin tissue. Distal tip of pectoral fin surpassing pelvic–fin insertion. Distal tip of pelvic fin not reaching anal–fin origin. Premaxilla with two rows of teeth and a short lateral process where the maxilla inserts. Five to six teeth of outer row tricuspid, arranged in zigzag. Internal

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80º 0'0''W

79º40'0''W

1º30'0''N

79º20'0''W

79º0'0''W

78º40'0''W

108ºW 90ºW 72ºW 54ºW 36ºW 18ºW

B. dahli B. ecuadorensis n. sp.

9ºN 0º 9ºS 18ºS

27ºS 36ºS 45ºS 54ºS

1º0'0''N 108ºW 90ºW 72ºW 54ºW 36ºW 18ºW 0 83ºW 81ºW 80ºW

78ºW

39,600 m 77ºW

75ºW 2ºN 0º

0º30'0''N

2ºS 3ºS 5ºS 83ºW 81ºW 80ºW

0º0'0''N

0 miles 80º 0'0''W

79º40'0''W

79º20'0''W

79º0'0''W

3,000

78ºW 0

77ºW

75ºW 400,000 m

78º40'0''W

Fig. 4. Distribution and location of B. ecuadorensis n. sp. and B. dahli in western Ecuador. Fig. 4. Distribución y localización de B. ecuadorensis sp. n y B. dahli en Ecuador occidental.

row with four pentacuspid teeth that do not diminish gradually in size. Posterior tip of maxilla surpasses anterior half of second infraorbital; its anterior margin continuous, with one or two pentacuspid teeth (fig. 2). Lateral ethmoid lateral surface covered by cartilage. Dentary with four large pentacuspid teeth with the central cusp largest, followed by six or eight small conical and tricuspid teeth, the anterior tooth the largest. Six infraorbitals present, the first long and with sensorial canal running its entire length, reaching posterior margin of antorbital. Second infraorbital short and wide, covering the dorsal part of the angulo–articular. Third infraorbital the widest and longest, its ventral border in contact with the sensorial canal of preopercle. Fourth and fifth infraorbitals short and narrow, covering the dorso–posterior margin of the hyomandibular. Sixth infraorbital covers anterior half of sphenotic foramen over the postero–lateral tip of frontal. Supraorbital absent. Rhinosphenoid present and cartilaginous along border, attached to orbitosphenoid and extending to vomer. Orbitosphenoid wide, short and united to pterosphenoid with or without a band of cartilage. The frontal sensory canal not extended to reach parietal. Lateral line complete, perforated scales 36–40 (40*, n = 90). Scales rows between dorsal–fin origin and lateral line 5–7 (6*, n = 90); scale rows between

lateral line and anal–fin origin 5–8 (6*, n = 90); scale rows between lateral line and pelvic–fin insertion 5–6 (6*, n = 90). Anal–fin rays vi, 25–29 (vi, 27*, n = 90); anal–fin origin posterior to vertical through base of first dorsal–fin ray. Dorsal–fin ray iii, 9 (iii, 9*, n = 90). Pectoral–fin rays i, 10–12,i (i, 11, i*, n = 90). Pelvic– fin rays i, 7 (n = 90); pelvic origin anterior to vertical through dorsal–fin origin. Caudal fin not scaled, forked with short pointed lobes, principal rays 1–9/8–1 with 6/9–10 procurrents. Total number of vertebrae 36–37. Secondary sexual dimorphism Sexually mature males have rows of hooks on branched anal–fin rays 1 to 26; each simple ray has 15–22 hooks located along the entire length of rays. There are also from 14–16 small hooks on all branched pelvic–fin rays, located on both branches of the rays, also extending along the entire length of rays. Small hooks are also present on pectoral–fin rays located on all branched rays with 10–15 hooks extending on the middle and most posterior portions. Dorsal fin with hooks located on branched rays, with 10–18 hooks found on both branches of rays; short hooks present on caudal–fin rays with 1–12 hooks located on posterior part of 8–10 middle rays. Males have more prominent, darker lateral stripes, deeper and wider caudal peduncles, and thicker caudal–fin

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1 cm B 1 cm B

1 cm Fig. 5. Pectoral girdle of B. dahli (A) and B. ecuadorensis n. sp. (B).

1 cm

Fig. 5. Cintura escapular de B. dahli (A) y B. ecuadorensis sp. n. (B). Fig. 6. Orbito–rhinosphenoid of B. dahli (A) and B. ecuadorensis n. sp. (B). rays than females. Male B. ecuadorensis show darker colors, related to higher concentration of melanophores along posterior margins of scales that are strongly concentrated along the lateral stripe and also along the infraorbitals, opercle and lateral surface of the cranium, whereas females have a lighter pigmentation pattern, with only the humeral and peduncular regions pigmented (fig. 3). Thus, males and females have sexually dimorphic pigmentation patterns, a condition rarely observed in other species of Bryconamericus, with the exception of B. oroensis. Color in alcohol Dorsum dark brown. Body with silvery lateral stripe from posterior edge of opercle to base of caudal fin. Anterior humeral spot consists of two overlapping layers of pigment: one layer runs transverse from ventral margin of second layer to posterior margin of opercle and crosses lateral line. The second pigment layer is a horizontal ellipse that extends over the series of scales above lateral line. Posterior humeral spot is present over lateral stripe. The caudal peduncle spot extends over the lateral stripe and continues on the middle caudal–fin rays. Sides and ventral region are yellow from tip of snout to caudal peduncle. Fins hyaline (figs. 1, 3).

Fig. 6. Órbito y rinosfenoide de B. dahli (A) and B. ecuadorensis sp. n. (B).

Distribution and ecological notes This species is known from the La Bocana de Cupa, Boca del Ónzole, Blanco, Sabalera and Atacames Rivers; and the Mataje, Santiago and Súa Rivers, in Esmeralda Province, Pacific versant, northwestern Ecuador and is sympatric with B. dahli (fig. 4). Bryconamericus ecuadorensis n. sp. was captured in rivers and creeks in lentic habitats over detritus and decomposing vegetal material with low transparency. The pH was near neutral, dissolved oxygen values, conductivity and total solids were high; turbidity and hardness were low (table 2), as is typical of eutrophic environments. These para� meters indicate a highly disturbed environment that affects the survival of this new species and those that share its habitat. Etymology Bryconamericus ecuadorensis is named for the country of Ecuador, where the type series was collected.

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Discussion The new species presented here differs from the sympatric B. dahli by the presence of hooks on all fins in males (vs. hooks present only on anal and pelvic fins). In B. dahli, there is a foramen over the lateral tips of the premaxillary that is absent in B. ecuadorensis; in B. dahli, the teeth of the outer pre� maxillary row stick out, but in B. ecuadorensis they are covered or not visible. An ongoing analysis has shown seven informative character–states that distinguish Bryconamericus ecuadorensis n. sp. from B. dahli: (1) margin of pig� ment layer two of the anterior humeral spot is well defined (vs. irregular in B. dahli) (fig. 1); (2) layer two of the anterior humeral spot is horizontally oval or elliptical (vs. rectangular or circular in B. dahli); (3) anterior extension of caudal peduncle spot extended over the lateral stripe of the body to reach an imaginary vertical through the last anal–fin pterygiophore l (vs. spot restricted to caudal peduncle); (4) area of dermal tissue beneath pectoral fin narrow, the area narrower than one scale in diameter (vs. area of dermal tissue beneath pectoral fin wider than one scale in diameter) (fig. 5); (5) sheath of scales over pectoral–fin origin consisting of more than four scales (vs. two or three scales) (fig. 5); (6) maxillary teeth orientation vertical, not inclined (vs. maxillary teeth inclined internally); and (7) anterior process of orbitosphenoid not narrowed to filament at union with rhinosphenoid (vs. anterior process of orbitosphenoid narrowed to filament at union with rhinosphenoid) (fig. 6). Males of some species of Characidae usually have hooks on the anal and pelvic fins but less frequently on the dorsal and caudal fins, a character that has been used as a synapomorphy for several genera of Characidae (Malabarba & Weitzman, 2003). The pre� sence of bony hooks on all fins including the caudal fin is not common for species of Knodus, Hemibrycon Hyphessobrycon or Tyttocharax; hooks on all fins of males have only been reported diagnostic in B. ecuadorensis described herein, Hemibrycon brevispini (Román–Valencia & Arcila–Mesa, 2009), Hyphessobrycon natagaima (Garcìa–Alzate et al., 2015), H. togoi (Miquelarena & Lopez, 2006), H. taguae (García–Alzate et al., 2008) and Tyttocharax metae (Román–Valencia et al., 2012). Moreover, the presence of bony hooks on the rays of the caudal fin of males (vs. absent) is diagnostic for Acrobrycon (Arcila et al., 2013) and has been used as a phylogenetically informative character by Mirande (2010) (coded as state 1 but only in some species of Characidae). Conservation status of the ichthyofauna and in particular Bryconamericus from the Pacific slope of northwestern Ecuador Calvo (2008) listed, for low and highlands in the An� des, common causal elements that typically result in loss of biodiversity and habitat quality: (1) increased human population leading to intensification of resource exploitation; (2) climate change; and (3) pollution from mining activities. While these impacts increase in spatial

extension and local intensity we continue to confront vast information gaps for Neotropical fish distribution, biology and taxonomy (Sarmiento & Barrera, 2008). Gold mining activities have completely destroyed natural waterways in some areas and dangerously increased mercury levels in many species of fish and tadpoles (Carnegie Institution for Science, 2013; Hernández et al., 2013). For the particular case of the fishes in the Pacific versant of Ecuador, and especially those in Esmeraldas Province (the type locality of the new species) in the rivers of the Santiago–Cayapas drainage: Tululbí, Cachaví, Bogotá, Wimbí, Santiago, Estero María, Zabaleta and Zapallito (Rebolledo–Mon� salve & Jiménez–Prado, 2013), current conditions are regrettably negative for the survival of fishes and health of human communities, having been affected by both legal and illegal mining activities, uncontrolled tourism, destruction of mangroves and the shrimp cultivation in� dustry (Jiménez Prado, 2012a, 2012b; Rebolledo–Mon� salve & Jiménez–Prado, 2013). Levels of Cr, N, V, Co, Hg and S in soils and sediments exceed established maximum limits and heavy metals have been found in the tissues of fishes and other organisms. One example that stands out is the presence of elevated levels of Zn and Cu in tissues of the Cagua (Gobiomorus maculatus), in more than ten localities examined from the middle and lower sections of the Santiago–Cayapas River, and also the presence of Hg, Al, As, Cd or Cr, in some of them. The elements analyzed also included Fe, Zinc, Mn, and fecal coliforms were found to exceed the legal limits permitted (table 2). The type of exploitation observed along the length of the Santiago–Cayapas is not so different from that recorded in other drainages of the Pacific coast, where fluvial gold mining occurs (Idárraga et al., 2010; Gon� zález, 2013; Hernández et al., 2013). The process is very similar: a cut is made to expose the hillside, and then the area is flooded by blasts of water from high pressure hoses to flush the soil into the sluice where a machine is used to separate gold particles from the remaining matrix of soil and stone. This gravity washing system accelerates erosion and liberates toxic materials from the matrix that then are flushed into the river. This explains why high levels of arsenic have been found in some rivers (CID–PUCESE & PRAS–MAE, 2012, 2014), even though arsenic is not used in the mining process. The huge amounts of sediments dangerously increase the turbidity of the water and downstream transport of materials that directly affect water quality downstream (table 2). This activity has greatly increased since 2008, when there were just a few mines, to over 200 in 2012, along almost the entire length of the Santiago River (Lapierre–Robles, 2012), and the practice shows no signs of diminishing, since gold prices remain high, a fact that underlies the great increase observed since 2008. Acknowledgements The authors (CR–V, RI–RC, CAG–A, and PJ–P) wish to thank the Centro de Investigación y Desarrollo of the Pontificia Universidad Católica del Ecuador,

Animal Biodiversity and Conservation 38.2 (2015)

(Esmeraldas Campus) for financing an expedition to Esmeraldas Province in the Pacific versant of Ecuador, and the organization of (PJ–P) and participation in (CR–V, RI–RC, CAG–A) the First National Meeting of Ichthyology that took place in Esmeraldas, Ecuador from 23–28 September 2014. We thank Eduardo Re� bolledo (CEMZ–P) for help with sampling and Ramiro Barriga (MEPN) for the loan of specimens. Cristian Román–P (UV & IUQ) made figure 4. We thank the editor of ABC, Rafael Zardoya, and Marcos Mirande and an anonymous referee for their comments and corrections which helped to improve this paper. References Almeida, B. L., 2014. Una revisión de la evaluación de la calidad del agua en los ríos de la provincia de Imbabura. Trabajo de fin de titulación. Universidad Técnica Particular del Loja, Ecuador. Arcila, D., Vari, R. P. & Menezes, N. A., 2013. Revision of the neotropical genus Acrobrycon (Ostariophysi: Characiformes: Characidae) with description of two new species. Copeia, 4: 604–611. Calvo, L. M., 2008. Problemas y amenazas en pueblos indígenas, originarios y comunidades locales para la conservación y uso sostenible de la biodiversidad In: Biodiversidad: La riqueza de Bolivia. Estado de conocimiento y conservación: 190–204 (L. P. Ibisch & G. Mérida, Eds.). Ministerio de Desarrollo Rural, Agropecuario y Medio Ambiente. Editorial FAN, Santa Cruz de la Sierra, Bolivia. Carnegie Institution for Science, 2013. Gold Mining Ravages Perú. Informe final de Monitoreo de cali� dad ambiental de ríos de la cuenca del Santiago afectados por la Actividad minera aurífera entre el periodo noviembre del 2011 a noviembre del 2012. Informe Técnico, Centro de Investigación y Desarrollo de la Pontificia Universidad Católica del Ecuador Sede Esmeraldas; programa de reparación ambiental y social del Ministerio del Ambiente. [Accessed on January 2015] http:// carnegiescience.edu/news/gold–mining–ravages– per%C3%BACID–PUCESE & PRAS–MAE.2012 CID–PUCESE & PRAS–MAE (Centro de Investigación y Desarrollo de la Pontificia Universidad Católica del Ecuador Sede Esmeraldas, Programa de reparación ambiental y social del Ministerio del Ambiente), 2012. Monitoreo de calidad ambiental de rios de la cuenca del Santiago afectados por la actividad minera aurífera entre el periodo noviembre de 2011 a noviembre del 2012. Informe técnico (no publicado). [Accessed on January 2015] https:// www.academia.edu/5777633/informe_final_de_ monitoreo_de_calidad_ambiental_de_rios_de_ la_cuenca_del_santiago_afectados_por_la_ac� tividad_minera_aurifera_entre_el_periodo_noviem� bre_del_2011_a_noviembre_del_2012 – 2014. Consultoría para la continuación de análisis de impactos de la minería aurífera en los cantones ''Eloy Alfaro'' y San Lorenzo de la Provincia de Esmeraldas. Producto 8.7. Informe final de obser� vación de calidad de agua en los cantones ''Eloy

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Animal Biodiversity and Conservation 38.2 (2015)

Animal Biodiversity and Conservation

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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. Comunicacions breus Les comunicacions breus seguiran el mateix proce� diment que els articles y tindran el mateix procés de revisió. No excediran de 2.300 paraules incloent–hi títol, resum, capçaleres de taula, peus de figura, agraïments i referències. El resum no ha de passar de 100 paraules i el nombre de referències ha de ser de 15 com a màxim. Que el text tingui apartats és op� cional i el nombre de taules i/o figures admeses serà de dos de cada com a màxim. En qualsevol cas, el treball maquetat no podrà excedir de quatre pàgines.

Animal Biodiversity and Conservation 38.2 (2015)

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, Discusión, Agradecimientos y Referencias) no se numerarán. No utilizar más de tres niveles de títulos. Los autores procurarán que sus trabajos originales no excedan las 20 páginas incluidas figuras y tablas. Si en el artículo se describen nuevos taxones, es imprescindible que los tipos estén depositados en alguna institución pública. Se recomienda a los autores la consulta de fascícu� los recientes de la revista para seguir sus directrices. Comunicaciones breves Las comunicaciones breves seguirán el mismo pro� cedimiento que los artículos y serán sometidas al mismo proceso de revisión. No excederán las 2.300 palabras, incluidos título, resumen, cabeceras de tabla, pies de figura, agradecimientos y referencias. El resumen no debe sobrepasar las 100 palabras y el número de referencias será de 15 como máximo. Que el texto tenga apartados es opcional y el número de tablas y/o figuras admitidas será de dos de cada como máximo. En cualquier caso, el trabajo maque� tado no podrá exceder las cuatro páginas.

Animal Biodiversity and Conservation 38.2 (2015)

Animal Biodiversity and Conservation

Manuscripts

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 regis� tered in all principal data bases and is freely available online at www.abc.museucienciesjournals.cat assur� ing 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, Pala� bras 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.

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 related studies should be discussed. Suggestions for future research may be given at the end of this section. Acknowledgements (optional). References. All manuscripts must include a 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. Brief communications Brief communications should follow the same proce� dure as other articles and they will undergo the same review process. They should not exceed 2,300 words including title, abstract, figure and table legends, ack� nowledgements and references. The abstract should not exceed 100 words, and the number of references should be limited to 15. Section headings within the text are optional. Brief communications may have up to two figures and/or two tables but the whole paper should not exceed four published pages.

Animal Biodiversity and Conservation 38.2 (2015)

VII

Welcome to the electronic version of Animal Biodiversity and Conservation

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Animal Biodiversity and Conservation joins the worldwide Open Access Initiative of providing a permanent online version free of charge and access barriers This is the result of the growing consensus that open access to research is essential for efficient and rapid scientific communication ABC alert, a free alerting service, provides eâ&#x20AC;&#x201C;mail information on the latest issue To sign on for this service, please send an eâ&#x20AC;&#x201C;mail to: abc@bcn.cat

Ruiz–García & Ferreras–Romero

Animal Biodiversity and Conservation 38.2 (2015)

Arxius de Miscel·lània Zoològica vol. 12 (2014) Museu de Ciències Naturals de Barcelona ISSN: 1698–0476 www.amz.museucienciesjournals.cat

Índex / Índice / Contents Borredà, V. & Martínez–Ortí, A., 2014. Babosas terrestres (Mollusca, Gastropoda) de la Región de Murcia (SE España). Arxius de Miscel·lània Zoològica, 12: 1–12. Abstract Terrestrial slugs (Mollusca, Gastropoda) from Murcia (SE Spain).— This study examines the few citations about terrestrial slugs from Murcia (SE Spain) and adds data from our own surveys. We provide a systematic checklist of the species of terrestrial slugs known from Murcia (SE Spain) and comment on the most significant species: Deroceras nitidum, Arion gilvus and A. lusitanicus s. l. Key words: Molluscs, Slugs, Biodiversity, Murcia, Spain Viñolas, A. & Masó, G., 2014. The collection of type specimens of the family Carabidae (Coleoptera) deposited in the Natural History Museum of Barcelona, Spain. Arxius de Miscel·lània Zoològica, 12: 13–82. Abstract The collection of type specimens of the family Carabidae (Coleoptera) deposited in the Natural History Museum of Barcelona, Spain.— The type collection of the family Carabidae (Coleoptera) deposited in the Natural History Museum of Barcelona, Spain, has been organised, revised and documented. It contains 430 type specimens belonging to 155 different taxa. Of note are the large number of hypogean species, the species of Cicindelidae from Asenci Codina’s collection, and the species of Harpalinae extracted from Jacques Nègre’s collection. In this paper we provide all the available information related to these type specimens. We therefore provide the following information for each taxon, species or subspecies: 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: Collection type, Coleoptera, Carabidae taxonomic revision family, Ground beetles III Encuentro Ibérico de Biología Subterránea Naranjo, M., Moreno, Á. C. & Martín, S., 2014. ¿Dónde buscar troglobiontes? Ensayo de una cartografía predictiva con MaxEnt en Gran Canaria (islas Canarias). Arxius de Miscel·lània Zoològica, 12: 83–92.

Abstract

Where should we search for troglobionts? A study of predictive cartography using MaxEnt in Gran Canaria (Canary Islands, Spain).— A total of, 160 terrestrial troglobiont species are known to date in the Canary Islands. These species are mainly located on the youngest islands that have abundant volcanic tubes (Tenerife, La Palma and El Hierro). On Gran Canaria, an older island with few volcanic caves, the hypogean fauna was considered poor until recent explorations in the mesocavernous shallow substratum and water mines were carried out, with remarkable results. The island covers an area of 1,560 km² and has a maximum height of 1,949 m. As ecological diversity is high, it is fundamental to identify the best areas to conduct effective sampling. Software for species habitat modelling such as MaxEnt, can be useful to select such areas, creating species distribution maps from environmental data. We obtained a potential distribution map using MaxEnt and including all the known troglobiont species of the island. This predictive map is statistically significant and accurately classifies a high percentage of the observed data. Lithology and average rainfall are the two variables that best predict the presence of these species. Basaltic materials —preferably modern— and gravitational landslides are good places for finding subterranean fauna. Locations with high average rainfall appear to be the most appropriate for this purpose. Key words: Troglobionts, Fauna, Modelling, SDM, Habitat, Gran Canaria, Canary Islands ISSN: 1578–665 X eISSN: 2014–928 X

Animal Biodiversity and Conservation 38.2 (2015)

Carrillo, M., Alcántara, E., Taverna, A., Paredes, R. & Garcia–Franquesa, E., 2014. Descripción osteológica del rorcual común (Balaenoptera physalus, Linnaeus, 1758) del Museo de Ciencias Naturales de Barcelona. Arxius de Miscel·lània Zoològica, 12: 93–123. Abstract Description of the skeleton of the fin whale (Balaenoptera physalus, Linaeus, 1758) at the Natural History Museum of Barcelona.— We describe the osteology of the fin whale (Balaenoptera physalus, L., registration code MZB 83–3084) at the Natural History Museum of Barcelona (MCNB). The specimen was stranded and died on a beach in Llançà (Girona, Spain) in 1862. The skeleton weighed 1,161.59 kg and measured 14.6 m, although the live animal would have been longer because the invertebral tissue was not included in the length originally stated. The newly reconstructed skeleton includes the invertebral discs and is 17.35 m long. The skull weighs 484 kg and the condilobasal length is 431 cm (24.84% of the total length), corresponding to the mean length of specimens in the Mediterranean. The vertebral column has 58 vertebra structured following the formula: C7 T14 L14 Cd23. It is 10.29 m long and weighs 470.95 kg. All the cervical vertebra are free and show dorsoventral compression, differing from the thoracic, lumbar and first caudal vertebra that are relatively uniform, and rounded. From Cd14 se onwards, the vertebra vary in shape and the relationship between width and height is greater than 1, indicating lateral compression. The lack of suture lines in the epiphysis of the ulna and radius indicates the specimen is an adult that has reached ossification maturation. Key words: Fin whale, Skeleton, Osteometry, Mediterranean Jawad, L. A., Al–Mukhtar, M. & Faddagh, M. S., 2014. First record of Heniochus acuminatus Linnaeus, 1758) (Chaetodontidae) and Pomacanthus maculosus (Forsskål, 1775) (Pomacanthidae) in Iraqi marine waters, Arabian Gulf. Arxius de Miscel·lània Zoològica, 12: 124–129. Abstract Confirmation of the presence of Heniochus acuminatus (Linnaeus, 1758) (Chaetodontidae) and first record of Pomacanthus maculosus (Forsskål, 1775) (Pomacanthidae) in Iraqi marine waters, Arabian Gulf.— Two specimens (116,119 mm TL) of Heniochus acuminatus (Linnaeus, 1758) and four specimens (171–190 mm TL) of Pomacanthus maculosus (Forsskål, 1775) were collected from Iraqi marine waters of the Arabian Gulf. These findings confirm the presence of H. acuminatus and establish the first record of P. maculosus from Iraqi waters. The samples were captured by hook and line off the coasts of Al–Fao City Peninsula, southern Iraq. Arabian Gulf. morphometric and meristic data are provided and compared with data from other parts of the world. Key words: Heniochus acuminatus, Pomacanthus maculosus, Basrah, Range extension, Iraq Viñolas, A., Caballero–López, B. & Masó, G., 2014. The collection of type specimens of the families Dytiscidae, Histeridae, Hydraenidae and Staphylinidae (Coleoptera) hosted in the Natural History Museum of Barcelona, Spain. Arxius de Miscel·lània Zoològica, 12: 130–161. Abstract The collection of type specimens of the families Dytiscidae, Histeridae, Hydraenidae and Staphylinidae (Coleoptera) hosted in the Natural History Museum of Barcelona, Spain.— The type collection of the families Dytiscidae, Histeridae, Hydraenidae and Staphylinidae (Coleoptera) deposited in the Natural History Museum of Barcelona, Spain, has been organised, revised and documented. It contains 130 type specimens belonging to 66 different taxa. Of note is the presence of a considerable number of hypogean species of the Staphylinidae family thanks to the descriptions by Enri Coiffait and Francesc Español. 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 following the description of the taxa. Key words: Collection type, Coleoptera, Dytiscidae, Histeridae, Hydraenidae and Staphylinidae taxonomic revision families

ISSN: 1578–665 X eISSN: 2014–928 X

Animal Biodiversity and Conservation 38.2 (2015)

III Encuentro Ibérico de Biología Subterránea Mederos, J. & Chandler, P. J., 2014. A constellation of fungus gnats (Diptera: Keroplatidae and Mycetophilidae) from caves of the Parc Natural dels Ports, Tarragona, Western Catalonia. Arxius de Miscel·lània Zoològica, 12: 163–173. Abstract A constellation of fungus gnats (Diptera: Keroplatidae and Mycetophilidae) from caves of the Parc Natural dels Ports, Tarragona, western Catalonia.— A preliminary approximation of the fungus gnat fauna (Diptera: Keroplatidae and Mycetophilidae) is presented, captured in caves of the Parc Natural dels Ports (Tarragona, Catalonia) following surveys conducted in the massif in 2012. We report a total of 11 species from ten prospected caves and provide data on the biology and images where possible. Exechiopsis coremura (Edwards) stands out due to few previous records of this species in the Iberian peninsula. Key words: Cave Habitats, Diptera, Mycetophilidae, Keroplatidae, Ports, Catalonia

ISSN: 1578–665 X eISSN: 2014–928 X

Ruiz–García & Ferreras–Romero

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Índex / Índice / Contents Animal Biodiversity and Conservation 38.2 (2015) ISSN 1578–665 X eISSN 2014–928 X

151–162 Angelici, F. M., Mahama, A. & Rossi, L. The lion in Ghana: its historical and current status 163–174 Biaggini, M. & Corti, C. Reptile assemblages across agricultural landscapes: where does biodiversity hide? 175–182 Rost, J., Jardel–Peláez, E. J., Bas, J. M., Pons, P., Loera, J., Vargas–Jaramillo, S. & Santana, E. The role of frugivorous birds and bats in the colonization of cloud forest plant species in burned areas in western Mexico 183–190 Lombardini, M., Murru, M., Repossi, A., Cinerari, C. E., Vidus Rosin, A., Mazzoleni, L. & Meriggi, A. Spring diet of the pine marten in Sardinia, Italy 191–206 Ortuño, V. M. & Barranco, P. Un nuevo Trechus (Coleoptera, Carabidae, Trechini) hipogeo de la Sierra de Parapanda (Andalucía, España): taxonomía, sistemática y biología

221–231 Bueno–Enciso, J., Núñez–Escribano, D. & Sanz, J. J. Cultural transmission and its possible effect on urban acoustic adaptation of the great tit Parus major 233–240 Armenteros, J. A., Sánchez–García, C., Alonso, M. E., Larsen, R. T. & Gaudioso, V. R. Use of water troughs by wild rabbits (Oryctolagus cuniculus) in a farmland area of north–west Spain 241–252 Román–Valencia, C., Ruiz–C., R. I., Taphorn B., D. C., Jiménez–Prado, P. & & García–Alzate, C. A. A new species of Br yconameric us (Characiformes, Stevardiinae, Characidae) from the Pacific coast of northwestern Ecuador, South America IX–XI Abstracts del volum 12 (2014) d'Arxius de Miscel·lània Zoològica Abstracts del volumen 12 (2914) de Arxius de Miscel·lània Zoològica Abstracts of volume 12 (2014) of Arxius de Miscel·lània Zoològica

207–220 Serrat, A., Pons, P., Puig–Gironès, R. & Stefanescu, C. Environmental factors influencing butterfly abundance after a severe wildfire in Mediterranean vegetation

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