ABC 41.1 (2018) by Museu Ciències Naturals de Barcelona MCNB

en J. Hatchwell, Univ. of Sheffield, UK 2

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

Animal Biodiversity and Conservation 41.1, 2018 © 2018 Museu de Ciències Naturals de Barcelona, Consorci format per l'Ajuntament de Barcelona i la Generalitat de Catalunya Autoedició: Montserrat Ferrer Fotomecànica i impressió: 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 / Animal Biodiversity and Conservation se publica con el apoyo de / Animal Biodiversity and Conservation is published with the support of: Asociación Española de Ecología Terrestre – AEET Sociedad Española de Etología y Ecología Evolutiva – SEEEE Sociedad Española de Biología Evolutiva – SESBE Disponible gratuitament a internet / Disponible gratuitamente en internet / Freely available online at:

Animal Biodiversity and Conservation 41.1 (2018)

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

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Dispersal of the monarch butterfly (Danaus plexippus) over southern Spain from its breeding grounds R. Obregón, D. Jordano, M. Cuadrado, J. M. Moreno–Benítez, J. Fernández Haeger

Obregón, R., Jordano, D., Cuadrado, M., Moreno–Benítez, J. M., Fernández Haeger, J., 2018. Dispersal of the monarch butterfly (Danaus plexippus) over southern Spain from its breeding grounds. Animal Biodiversity and Conservation, 41.1: 1–8. Abstract Dispersal of the monarch butterfly (Danaus plexippus) over southern Spain from its breeding grounds. From 2000–2016, monarch butterflies were detected at 127 locations away from their usual coastal breeding areas in the south of the Iberian peninsula. These findings were recorded in the summer–autumn period, coinciding with the highest abundance of individuals and the highest proportion of patches occupied in their reproduction areas near the Strait of Gibraltar. These dispersing individuals have no chance of successfully establishing new colonies at these sites because the food plants for egg laying do not grow in the localities where they were detected. However, these dispersive movements could be the source of their successful colonisation on food plants growing in other areas of the Iberian Peninsula and in other Mediterranean countries. Key words: Danaus plexippus, Monarchs, Dispersal, Southern Spain Resumen Dispersión de la mariposa monarca (Danaus plexippus) en el sur de España desde las zonas de apareamiento. Durante el período comprendido entre los años 2000 y 2016, se detectaron mariposas monarca en 127 lugares fuera de las zonas costeras donde se reproducen habitualmente en el sur de la península ibérica. Estos datos se obtuvieron en verano e invierno, coincidiendo con la máxima abundancia de individuos y la mayor proporción de sitios ocupados en sus zonas de reproducción cercanas al estrecho de Gibraltar. Los individuos que se dispersan no tienen ninguna posibilidad de establecer nuevas colonias en estos sitios porque las plantas en las que ponen los huevos no crecen en las localidades en las que fueron detectados. Sin embargo, estos movimientos de dispersión podrían ser la causa de la colonización de plantas alimentarias que crecen en otras zonas de la península ibérica y en otros países del Mediterráneo. Palabras clave: Danaus plexippus, Monarcas, Dispersión, Sur de España Received: 22 II 17; Conditional acceptance: 8 V 17; Final acceptance: 21 V 17 Rafael Obregón, Diego Jordano, Juan Fernández Haeger, Dept. of Botany, Ecology and Plant Physiology, Univ. of Cordoba, E–14071 Cordoba, Spain.– José Manuel Moreno–Benítez, c/ Larga del Palmar 34, E–29650 Mijas, Malaga, Spain.– Mariano Cuadrado, ZooBotánico de Jerez, c/ Madreselva s/n., E–11408, Jerez de la Frontera, Cadiz, Spain.

ISSN: 1578–665 X eISSN: 2014–928 X

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Introduction In butterflies, the fundamental functions of imagos are reproduction and dispersion (Gilbert and Singer, 1975). Dispersion refers to the spatial separation between individuals of a population (Begon et al., 1999) and is a basic process in ecology, evolution and conservation biology (Clobert et al., 2001; Bullock et al., 2002). Numerous empirical studies have studied the differences in the processes of dispersion between species and between sexes, and in different types of landscapes (Fernández et al., 2016). These processes are especially relevant to the understanding of population dynamics in fragmented landscapes (Hanski and Gilpin, 1997; Hanski and Gaggiotti, 2004). An attempt has been made to determine whether dispersion is a consequence of the routine movements carried out by the species in question (for example, in the search for food) or whether, on the contrary, it is the result of specific movements made for that purpose (Van Dyck and Baguette, 2005). Dispersion in fragmented landscapes can be considered a three–step process that involves the decision to migrate (abandoning a favourable habitat), the crossing through an unfavourable matrix, and migration to another favourable habitat fragment (Clobert et al., 2004). In dispersion studies, the spatial arrangement of these fragments combined with the ability to move between them is critical to understanding the success of the process. In general, in long–distance dispersal through unfavourable habitats (matrices), the trajectories of the butterflies tend to be faster and rectilinear, leading to a decrease in the mortality while crossing the matrices. In contrast, more–sinuous trajectories are characteristic of an exploratory search for resources (nectar, plants for laying eggs, etc.) within the favourable habitat fragment (Fernández et al., 2016). However, the matrix should not be considered an empty habitat that is hostile to the dispersal of butterflies, since the possibility of exploiting it depends on the biology of each species (Zalucki et al., 2015b). As discussed by Dennis (2004), the matrix can provide many sources of nectar and other resources for butterflies. However, in multivoltine species living in environments with strong seasonal fluctuations, the availability of these resources and the quality of this habitat can vary significantly throughout the year (Fernández Haeger et al., 2011b). The monarch butterfly (Danaus plexippus) is probably one of the best–studied insects in the world (Oberhauser et al., 2015, and references therein). It is of American origin that was first described in the mid–nineteenth century in New Zealand (1840), and a little later Australia (1870), where it seems to have arrived by migrating from island to island across the Pacific (Zalucki and Clarke, 2004). Around this time it was also recorded in various Atlantic archipelagos —Madeira (1860), the Azores (1864), the British Isles (1876) and the Canary Islands (1880)— and in the Iberian Peninsula (Portugal and Gibraltar), in 1886. In 1988 it was first recorded in North Africa (Fernández Haeger et al., 2015). Although

the first monarch butterfly in the Campo de Gibraltar (Spain) was recorded in 1886, it did not appear again in the scientific literature (Gonella, 2001) until 1963. In the mid–1980s monarchs again attracted the attention of entomologists, although these occasional sightings in the south of Spain have been interpreted as migrating or wandering individuals from Atlantic archipelagos. The existence of stable populations in the vicinity of the Strait of Gibraltar has been documented at least since 1994; in this location, they reproduce on species of the Asclepiadaceae (Fernández Haeger and Jordano Barbudo, 2009 and references therein). In this vicinity of the Strait of Gibraltar, the monarchs are centred around their fundamental food plants, Asclepias curassavica, Gomphocarpus fruticosus and G. physocarpus. These plants have a fragmented distribution as they form stands in frost–free coastal areas in the southern Iberian Peninsula and North Africa (Fernández Haeger et al., 2015) (fig. 1). The presence of stands of their food plants is favoured by moist soils (edges of streams, ponds, springs, wet meadows, etc.) throughout the year, and by intense livestock pressure. The butterflies appear in flight throughout the year in a succession of generations. They do not have a migratory behaviour equivalent to that of the monarchs of North America, but movements of marked individuals between vegetation fragments separated by more than 2 km have been recorded (Fernández Haeger et al., 2011b). In this way the butterflies persist in this area, maintaining the dynamics of a patchy population structured on the basis of the fragmented populations of its food plants. Asclepiadaceae flower from March to December and produce large amounts of nectar (Wyatt and Broyles, 1994). Since they are in full bloom during the summer, when most herbaceous plant species of the Mediterranean have dried out, they are doubly attractive to butterflies at this time of year because they provide both nectar for the adults and abundant food for the caterpillars (Fernández Haeger et al., 2011a; 2011b). In recent years, and especially in 2016, numerous imagos have been recorded in flight or feeding on nectar far from the areas where the food plants of their caterpillars grow. These individuals can be considered vagrants and could contribute to the expansion of the range of the species in the Iberian Peninsula. The objective of this work was to produce a map of all the sightings of monarch butterflies recorded outside their known breeding areas from 2000 to 2016 in southern Spain, and also to interpret these dispersive movements in relation to the possibilities of success for their expansion in the south of the Iberian Peninsula. Material and methods The observations of monarch butterflies considered here have various origins. First, 10 citations are from bibliographic references published by different authors over the study period (Gonella, 2001; Gil–T., 2006; Huertas–Dionisio, 2007; Laffitte et al., 2010; Moreno– Benítez, 2015). Three further citations come from the

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Spain

Se Hu Ma

Ca Reproductive patches

Dispersal observations

180 km

Extinct reproductive patches

Fig. 1. Spatial distribution in the south of the Iberian Peninsula and in the north of Africa of the localities where the Monarch butterfly reproduces and the location of dispersal individuals during 2000–2016: Hu, Huelva; Se, Sevilla; Co, Cordoba; Ca, Cadiz; Gi, Gibraltar; Ta, Tangier; Ma, Malaga. Fig. 1. Distribución espacial en el sur de la península ibérica y en el norte de África de las localidades en las que la mariposa monarca se reproduce y localización de los individuos que se dispersaron durante 2000–2016: Hu, Huelva; Se, Sevilla; Co, Córdoba; Ca, Cádiz; Gi, Gibraltar; Ta, Tánger; Ma. Málaga.

registers in the collection of the Andalusian Society of Entomology of the Museum of Natural Sciences of Guadalcázar (Cordoba), and 31 sightings were provided by individuals on the Facebook social network groups Mariposas diurnas de Andalucía and Sociedad Andaluza de Entomología. Another 53 sightings were provided by naturalists in the area and 30 more were provided by ourselves. Some of these latter observations were obtained during the periodic censuses that are carried out within the project BMS Spain, in the province of Cadiz. Since identification of the monarch butterfly is unmistakable within the butterfly fauna of southern Spain, all observer records were treated as valid. Results Figure 1 shows the geographical locations of the areas where reproduction of the species has been verified, along with the sightings corresponding to vagrant individuals collected during the study period (2000–2016, n = 127). The areas of reproduction include the Portuguese Algarve, a small coastal area between the towns of Chipiona and Rota (recently discovered), the coastal zone near the Strait of Gibraltar, some enclaves of the Costa del Sol, and North Africa. Other enclaves where the butterfly has reproduced in the past (Doñana, Torrox or Almeria) are also represented on the map, although these have

now disappeared due to human intervention (Tarrier, 1994; Fernández Haeger et al., 2009). As shown in the map, most of the observations of dispersive butterflies are near the coast, although some inland localities are especially notable, such as Güejar–Sierra (Granada), Ronda (Málaga), Cabra (Córdoba), and Córdoba. The observation at highest altitude was at Güejar Sierra (964 m a.s.l.) and that furthest from the coast was in the city of Córdoba. It is of note that the food plants (A. curassavica, G. fruticosus and G. physocarpus) of the caterpillars were not noted in any of these localities, so all these individuals can be considered vagrants in search of new areas where they can successfully establish themselves. Some sightings refer to consecutive days in the same locality. Although they have been considered as different individuals, we do not rule out the possibility that they are the same individuals attracted to the nectar of plants in bloom. In most cases it was not possible to determine the sex of the individuals since they were observed in flight at a certain distance. However, sex was determined in 17 cases, corresponding to 12 males and five females. The large majority of sightings were obtained in the summer and early autumn (July–October, fig. 2). It seems, therefore, that these butterflies tend to fly outside their breeding grounds in the second half of the year (summer–autumn). From January to May, only one specimen was observed, this being in March.

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Percentage of observations

35 % 30 % 25 % 20 % 15 % 10 % 5 % 0 %

Mar

Jun

Jul

Aug

Sep Month

Oct

Nov

Dec

Fig. 2. Monthly distribution of the number of sightings of individuals recorded outside the breeding habitat, from 2000 to 2016 (n = 127). Fig. 2. Distribución mensual del número de avistamientos de individuos registrados fuera del hábitat de reproducción, entre los años 2000 y 2016 (n = 127).

Discussion In the south of the Iberian Peninsula, the monarch butterfly behaves as a non–migratory, multivoltine species, showing no winter diapause and forming local populations associated with stands of A. curassavica, G. fruticosus and G. physocarpus (Fernández Haeger et al., 2011b). Although, we have occasionally detected some eggs on Cynanchum acutum (Asclepiadaceae), this plant species has a very restricted distribution in our geographical area and lacks the biomass to maintain a local population of monarchs. In addition, it does not maintain aerial biomass during the winter (John et al., 2015). Over the year, both the occupancy of patches and the abundance of individuals vary markedly, with minimum values occurring in winter and maximum values in late summer and early autumn (Fernández Haeger et al., 2015). Hence, during the winter, the butterflies are concentrated in a few highly favourable patches. Over the course of the year they are able to colonise other patches, and both the occupation of patches and the maximal abundance of individuals occur in late summer and early autumn. In other species of butterflies, the abundance of local populations is related to the size of the fragment, abundance of food plants for larvae, and abundance of nectar sources (Krauss et al., 2004; Lenda and Skórka, 2010). These first two factors are also key to the establishment of new populations (Kuussaari et al., 2015). In the case of D. plexippus in southern Spain, the largest local populations are generally associated with stands where A. curassavica is abundant, as opposed to those with only G. fruticosus (Fernández Haeger et al., 2015). In the most favourable fragments, the local populations reach high densities, so that after the development of successive generations of caterpillars, the

food plants are practically defoliated by late summer/ early autumn (fig. 3). As a result, the quality of the stand drops sharply, forcing the imagos to abandon it in search of more favourable stands where they can find nectar and lay their eggs. The sexual behaviour of the monarchs means that, in many cases, the females try to avoid fragments that have a high density of males so as to avoid harassment and to devote more time to egg–laying (Frey et al., 1998), as occurs in other species (Shapiro, 1970). We observed that the sex ratio (males/females) in a set of 478 individuals captured randomly throughout the year (2016) in two fragments in the Tarifa area was 2.04. This ratio rose to 3.6 in October 2016 in one of the sampled fragments (unpublished data), suggesting that females try to find new places for egg–laying, especially during autumn. Therefore, during late summer and autumn in the south of the Peninsula, recurrent conditions that promote the dispersal of the Danaus are repeated. Accordingly, we recorded diverse observations of dispersive individuals in different localities and at varying distances with respect to the coastal areas of reproduction. The observations of dispersive individuals occurred mainly between June and December, coinciding to a great extent with the time when population numbers in breeding areas are higher (Fernández Haeger et al., 2015). Leaving the stand during the summer and crossing the matrix of unfavourable habitats is complicated for butterflies because most herbaceous plants have dried up and nectar sources are scarce (personal observation; Zalucki and Lammers, 2010). Although no specific observations for monarchs are available, it is possible that butterflies move along river courses to reach other favourable fragments, as described in Viejo et al. (1992) for other species of butterflies.

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A B

Fig. 3. Caterpillars in the last stage of development devouring apical parts of A. curassavica and G. fruticosus (A and B). Both photos were taken in October 2011. As a result of such herbivore pressure, stands of hundreds of plants can be completely defoliated (C). The final outcome is a sudden loss of stand quality, as seen in the third image, and the dispersal of the butterflies. Fig. 3. Orugas en el último estadio de desarrollo alimentándose de las partes apicales de A. curassavica y G. fruticosus (A y B). Ambas fotografías fueron tomadas en octubre de 2011. Como consecuencia de esta presión herbívora, los tallos de cientos de plantas pueden defoliarse por completo (C). El resultado final es la pérdida repentina de calidad del tallo, tal como se aprecia en la tercera imagen, y la dispersión de las mariposas.

Such movement along the riparian forests offers greater possibilities to find nectar source plants, a more favourable microclimate and occasionally refuge from the wind. The present study also reveals the appearance of the species during 2016 in southern areas of the Iberian Peninsula where the species was thought to be absent or rare. This expansion has been especially striking in two specific areas: the Bahía de Cádiz and inland areas of the province of Málaga. The monitoring of the butterflies carried out by one of the authors in different plots of the same area shows that: (1) from 2011 to 2015, only two specimens were detected in an intensely–sampled plot in Jerez; (2) the species was not previously detected in the BMS censuses carried out intensively in six plots of the same area throughout the year; and (3), in 2016, the species was detected for the first time (a total of six individuals) in three BMS transects, in San Fernando, El Marquesado and Jerez. The species was recorded in 2016 in various mountainous areas in the interior of Málaga and even in the province of Córdoba. It appears, therefore, that the dispersal of the monarchs from their coastal breeding areas was especially important in 2016, coinciding with an exceptional abundance in these areas (unpublished data). The dispersive movements reported in this work suggest that the monarch butterflies could extend their permanent distribution area to other zones in the south of the Iberian Peninsula. At a global scale, the potential–distribution model developed by Zalucki et al. (2015a) predicts the existence of favourable coastal zones for the monarch in North Africa, southern Europe, and the Middle East —which it could certainly reach due to its great capacity for flight. This capacity has enabled it to cross the Pacific and the Atlantic successfully. However, the colonisation of these zones depends critically on the presence of A. curassavica and/or G. fruticosus. When butterflies crossing the Atlantic reach the coastal areas of the south of the Iberian Peninsula they find stands of their food plants (Fernández Haeger et al., 2009, 2015). This is not the case for those arriving at the European coasts at more northern latitudes (such as northern parts of the Iberian Peninsula, France, Ireland, United Kingdom) where the food plants do not exist and monarchs are unable to reproduce (Asher et al., 2001). In the south of the Iberian Peninsula, A. curassavica, G. fruticosus and G. physocarpus are naturalised species. The first of these was introduced from Central America by the Spanish and the other two were brought from southern Africa by the Portuguese (Fernández Haeger et al., 2015). Their presence in the Doñana National Park and the work carried out to eradicate them probably determined their inclusion in the list of invasive species in Andalusia (Dana et al., 2005), although they are not currently listed in the Catálogo Español de Especies Exóticas Invasoras–Flora, of the Ministerio de Agricultura, Pesca, Alimentación y Medio Ambiente (http://www.mapama.gob.es/es/ biodiversidad/temas/conservacion–de–especies/especies–exoticas–invasoras/ce_eei_flora.aspx), where

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other species of the Asclepiadaceae (Araujia sericifera and Calotropis procera), considered invasive in the Canary Islands, do appear. Although of native origin and despite the effective long–range airborne dispersal of their seeds, the plants that the monarch caterpillars feed on do not have the characteristics of invasive plants. This is because they are pioneer species that only thrive in disturbed sites, and because they compete poorly with various native species (rushes, brambles, reeds, oleanders, etc.) that displace them easily in secondary succession processes. They are also very sensitive to frost, especially A. curassavica, and need a lot of moisture in the soil. Therefore, their distribution is limited to enclaves that maintain high soil moisture throughout the year, they are subject to frequent disturbance (e.g. extensive cattle ranching, clearing, flooding), and they are located in low–lying, frost–free coastal areas with high annual sunshine hours (Valdés et al., 1987; Blanca et al., 2011; Fernández Haeger et al., 2010). The observations of monarch butterflies outside their breeding areas documented in this work correspond to dispersive individuals with very little or no chance of success in these enclaves due to the absence or scarcity of their food plants. Since the climatic conditions are adequate in many other areas of the Mediterranean zone (Zalucki et al., 2015a), the previous spreading of these plants —in a natural or artificial way— is the key factor that would facilitate the expansion of monarch butterfly territories. Finally, during the last few years, exotic butterflies —in many cases including monarchs— have been released in ceremonial events (weddings, baptisms, funerals). This seems to be the reason for the presence of monarchs on islands as far away as Ibiza and Cyprus (John et al., 2015). During the summer of 2015, several specimens of Hypolimnas misippus and D. plexippus appeared in gardens near the River Guadalquivir in Córdoba. The origin of both species was a ceremonial release in memorial of deceased children and they disappeared a few days later. These monarch records have not been considered in this work, but an isolated specimen detected in 2016 has been included due to the high abundance of monarchs in southern Spain during that year and no evidence of more ceremonial releases. Although escaped butterflies from insect houses cannot be discounted, the presence of monarchs in the south of the Iberian Peninsula (Málaga and Cádiz) predates the existence of these facilities (Torres Mendez, 1979; Verdugo Páez, 1981; Tapia Domínguez, 1982). For southern Portugal, the first colonies were described in 2003 (Palma and Bivar de Sousa, 2003), although they must have been established long before this date. The insect house located in Benalmádena (Málaga) began its activity in 2011 but monarchs were detected in this area more than 30 years earlier (Torres Mendez, 1979). In its vicinity, in the park of La Paloma of Benalmádena, there is a flourishing colony of monarchs that persists due to an artificial flower bed of A. curassavica. Some years, there is a phenomenon equivalent to that described by ourselves in the south of Cádiz: the density of caterpillars is such that they

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consume the total biomass of A. curassavica available and there is a high mortality of caterpillars due to starvation. At such times, most imagos are forced to leave this enclave to find other, more–favourable sites. Finally, since G. fruticosus is widespread in coastal areas of other countries across the Mediterranean basin to Anatolia, including some of the main islands (e.g. Balearics, Corsica, Sardinia, Crete), the dispersal movements of monarch butterflies from their breeding areas in the south of the Iberian Peninsula, documented in this work, could be the origin of their successful establishment in other countries of the Mediterranean basin. Acknowledgements We wish to express our gratitude to Asunción Gómez and Elena Escribano of the butterfly house at Benalmádena. We also thank the many naturalists, friends and collaborators of the Facebook groups of the Sociedad Andaluza de Entomología and the Mariposas diurnas de Andalucía: Francisco Cobos, Adrià Miralles, Antonio Verdugo, M. Huertas–Dionisio, Andrés Rodríguez, Gonzalo Astete, José M. Ayala, Juan A. García, Juan Quetglas, Enrique Sánchez–Gullón, Álvaro Pérez, J. M. Mateo–Lozano, Dolores Cabrera, Carlos García, Clara M. Sánchez, Francisco Solano, Inmaculada Prieto, Emiliano Ruiz, José M. Canca, Antonio J. Plaza, Carlos Guerrero, Francisco M. Sánchez, Gaspar Mena, Enrique Coto, Elena Gallego, Salvador Naranjo, Clara Freiin, Brenda Jones, Pedro Gómez, Carlos Tapia and Antonio A. Jiménez, for their invaluable observations. Isabel Fernández García supplied us with a valuable reference from Cabra (Córdoba) for the summer of 2015. We also express our gratitude to Fabiola Lloret for the valuable revision of the final version of this paper. Two anonymous referees also improved this manuscript. References Asher, J., Warren, M., Fox, R., Harding, R., Jeffcoate, G., Jeffcoate, S., 2001. The Millennium Atlas of Butterflies in Britain and Ireland. Oxford University Press, Oxford. Begon, M., Harper, J. L., Townsend, C. R., 1999. Ecología. Individuos, poblaciones y comunidades. Omega Ed., Barcelona. Blanca, G., Cabezudo, B., Cueto, M., Salazar, C., Morales Torres, C., 2011. Flora Vascular de Andalucía Oriental. Universidades de Almería, Granada, Jaén y Málaga, Granada. Bullock, J. M., Kenward, R. E., Hails, R. S., 2002. Dispersal Ecology. Blackwell Science, Malden. Clobert, J., Danchin, E., Dhondt, A. A., Nichols, J. D., 2001. Dispersal. Oxford University Press, New York. Clobert, J., Ims, R. A., Rousset, F., 2004. Causes mecahnisms and consequences of dispersal. In: Ecology, genetics and evolution of metapopulations: 307–335 (I. Hanski, O. E. Gaggiotti, Eds.). Academic Press, London.

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Identifying earthworms (Oligochaeta, Megadrili) of the Southern Kuril Islands using DNA barcodes S. V. Shekhovtsov, Yu. N. Sundukov, R. J. Blakemore, K. B. Gongalsky, S. E. Peltek

Shekhovtsov, S. V., Sundukov, Yu. N., Blakemore, R. J., Gongalsky, K. B., Peltek, S. E., 2018. Identifying earthworms (Oligochaeta, Megadrili) of the Southern Kuril Islands using DNA barcodes. Animal Biodiversity and Conservation, 41.1: 9–17. Abstract Identifying earthworms (Oligochaeta, Megadrili) of the Southern Kuril Islands using DNA barcodes. The Kuril Islands are a volcanic archipelago located between Hokkaido and Kamchatka. In this study we investigated earthworm fauna of three of the Southern Kuril Islands, Kunashir, Shikotan, and Yuri, using both morphological analysis and DNA barcoding. Our results highlight the potential of DNA barcoding for studying earthworm fauna: while previous studies reported only six earthworm species and subspecies on the Southern Kurils, we detected 15 genetic clusters. Six of them correspond to European cosmopolites; six, to Asian species; and three, to unidentified species. While no European earthworms were found on Yuri that is uninhabited since WWII, they dominated on larger and inhabited Kunashir and Shikotan, suggesting that they are recent invaders. Of the six Asian species, five had cox1 sequences identical or very closely related to published sequences from the mainland or the Japanese islands and thus are recent invaders. Key words: Earthworms, Megadrili, cox1, Barcoding, Kuril islands Resumen Identificación de las lombrices (Oligochaeta, Megadrili) del sur de las Islas Kuriles utilizando códigos de barras de ADN. Las Islas Kuriles forman un archipiélago volcánico situado entre Hokkaido y Kamchatka. En este estudio analizamos las lombrices de tres de las islas Kuriles meridionales: Kunashir, Shikotan y Yuri, utilizando el análisis morfológico y los códigos de barras de ADN. Nuestros resultados ponen de relieve el potencial de los códigos de barras de ADN para estudiar las lombrices: si bien en estudios anteriores solo se habían registrado seis especies y subespecies de lombriz en las islas Kuriles meridionales, nosotros detectamos 15 grupos genéticos. Seis de ellos son especies cosmopolitas europeas; seis, especies asiáticas; y tres, sin determinar. A pesar de que no se encontraron lombrices europeas en Yuri, que está deshabitada desde la Segunda Guerra Mundial, estas especies dominaron en las islas Kunashir y Shikotan, que son más grandes y están habitadas, lo que sugiere que se trata de especies invasoras recientes. De las seis especies asiáticas, cinco tenían secuencias cox1 idénticas o muy emparentadas con las secuencias publicadas encontradas en el continente o en las islas del Japón y, por tanto, se trata de invasoras recientes. Palabras clave: Lombrices, Megadrili, cox1, Código de barras de ADN, Islas Kuriles Received: 19 V 17; Conditional acceptance: 26 V 17; Final acceptance: 5 VI 17 S. V. Shekhovtsov, S. E. Peltek, Inst. of Cytology and Genetics SB RAS, Siberian Branch of the Russian Academy of Sciences, Novosibirsk, 630090, Russia.– Yu. N. Sundukov, State Nature Reserve 'Kurilskiy', Zarechnaya str. 5, Yuzhno–Kuril’sk, Sakhalinskaya oblast, 694500, Russia.– R. J. Blakemore, VermEcology, Zama, Japan and Lake Biwa Museum, Kusatsu, Shiga–ken, 525–0001, Japan.– K. B. Gongalsky, A. N. Severtsov Inst.of Ecology and Evolution, Russian Academy of Sciences, Leninsky Pr. 33, Moscow, 119071, Russia. Corresponding author: S. V. Shekhovtsov. E–mail: shekhovtsov@bionet.nsc.ru ISSN: 1578–665 X eISSN: 2014–928 X

Shekhovtsov et al.

Introduction DNA barcoding is a powerful tool for identification of organisms (Hebert et al., 2003). It allows one to work with specimens that cannot be identified using their morphology, e.g., juvenile or degraded ones (Hajibabaei et al., 2006). In addition, DNA barcoding is promising for groups with problematic taxonomy, e.g. earthworms (Decaëns et al., 2013). Although earthworms are fairly well studied, recent molecular research demonstrated outstanding cryptic diversity within most earthworm species, which is exacerbated by the paucity of diagnostic morphological characters in this group (King et al., 2008; Rougerie et al., 2009; James et al., 2010; Marchán et al., 2017). Earthworms are important components of soil biota, generally believed to increase general ecosystem productivity (Darwin, 1881; Lee, 1985). However, human–mediated earthworm invasions are becoming problems in some regions (Hendrix et al., 2008): they may alter community composition, promote colonization of ecosystems by other invasive species (Roth et al., 2015), and even decrease ecosystem productivity by depleting available nutrients (Resner et al., 2015). The Southern Kurils is an archipelago located to the north of Hokkaido. There are several studies on the earthworm fauna of Southern Kurils (Perel, 1979; Marusik, 2002; Gongalsky et al., 2014). Six earthworm species and subspecies were reported: Metaphire hilgendorfi (Michaelsen, 1982) (= Pheretima hilgendorfi = Amynthas hilgendorfi), Eisenia japonica (Michaelsen, 1892), Lumbricus rubellus Hoffmeister, 1843, Dendrobaena octaedra (Savigny, 1826), Dendrodrilus rubidus tenuis (Eisen, 1874), and Allolobophora parva (Eisen, 1874). While the latter four are cosmopolites of European origin (Hendrix et al., 2008), the former two are local Asian species from the Japanese archipelago. The aim of this study was to investigate earthworm samples from three Southern Kuril Islands, Kunashir, Shikotan and Yuri. These islands with areas of 1,500, 225, and 10 km2, respectively, are characterized by moderate marine climate (Krestov et al., 2009; Pietsch et al., 2003; Razzhigaeva et al., 2014). Shikotan is covered by meadows and dark coniferous, birch and alder forests. Kunashir contains diverse biotopes with a mixture of boreal and subtropical species. Yuri is completely devoid of trees and is covered by meadows and bogs. We used DNA barcoding in order to update species composition of these islands and attempt to detect species origins. We wanted to clarify the origin of its Asian earthworm fauna (ancient vs. recent) and to examine the patterns of their coexistence with European cosmopolites, in terms of genetic diversity and relative abundance. Material and methods Specimens were collected in 2012–2016 (fig. 1, table 1). A part of the material from the south of Kunashir (13 individuals) was taken from the study

of Gongalsky et al., 2014). Taxonomic nomenclature follows Blakemore, 2003). In addition to morphological identification, a set of specimens, including juvenile individuals, was taken for DNA barcoding. DNA isolation, amplification, and sequencing were performed as described in Shekhovtsov et al., 2013). All obtained cox1 sequences were deposited in GenBank under accession numbers KX400579–KX400731 and KY750693–KY750709 (see table 1). Nucleotide and gene diversity values and mismatch distributions were calculated using Arlequin v.3.5.2.2 (Excoffier and Lischer, 2010). Divergence times for the E. japonica dataset were calculated by BEAST v.1.7.2 (Drummond et al., 2012) and Tracer v.1.6.0 (Rambaut et al., 2014). It is problematic to calibrate molecular clock for earthworms (Fernández et al., 2016), as earthworm fossils are almost absent and cannot be attributed to any particular taxon. There are calibration methods based on outgroup clitellate fossils (Marchán et al., 2017), however, they seem to be more applicable to deeper time spans and high–level taxa. Here we used the molecular clock rate of 2.5 % sequence divergence per million years obtained using vicariance dating (Pérez–Losada et al., 2011) as a rough estimate. Phylogenetic trees were built using MrBayes v.3.2.0 (Ronquist and Huelsenbeck, 2003) and MEGA v.5.0 (Tamura et al., 2011). For Bayesian analysis, the HKY+I+G model was suggested by MrModeltest (Nylander, 2004). Two independent analyses were performed using ‘metropolis coupled Monte Carlo’ simulations for 10 million generations, sampling a tree every 10,000 generations. For minimum evolution and maximum likelihood algorithms, bootstrapping was performed with 1,000 replications. Results A total of 170 individuals from Kunashir, Shikotan, and Yuri were barcoded using the cox1 gene. We detected 15 clusters corresponding to distinct Operational Taxonomic Units that may represent different earthworm species or mitochondrial genetic lineages (fig. 2). Six of these OTUs corresponded to well–known European cosmopolites: L. rubellus, A. caliginosa (Savigny, 1826), Aporrectodea rosea (Savigny, 1826), Octolasion tyrtaeum (Savigny, 1826) (also incorrectly referred to as Octolasion lacteum (Örley, 1885)), D. octaedra, and D. r. tenuis. Three of those, O. tyrtaeum, A. caliginosa, and A. rosea were reported from Southern Kurils for the first time. Lumbricus rubellus was prevalent on Kunashir, making up 47 % of the total sample (and all European cosmopolites accounted for about 60%). We should also note that D. octaedra and D. r. tenuis were relatively under–represented in this study due to sampling bias as we took few litter samples. All mtDNA haplotypes of European cosmopolites were identical to the representatives of these species found in European populations (Fernández et al., 2012, 2016; Porco et al., 2013; Shekhovtsov et al., 2014a, 2014b, 2016). For the two most frequent

Animal Biodiversity and Conservation 41.1 (2018)

Russia

Iturup

Kamchatka Peninsula

Sea of Okhotsk China

K ut

la n

Sakhalin

Sea of Japan Japan

Pacific Ocean

Hokkaido

Kunashir

Shikotan

Yuri 0

50 km

Fig. 1. Locations sampled in this study. Upper left, small scale map of the region; location of the studied islands is shown by a square. (For abbreviations of locations see table 1). Fig. 1. Lugares muestreados en este estudio. En la esquina superior izquierda, mapa de la región en pequeña escala; la ubicación de las islas estudiadas se indica con un cuadrado. (Para las abreviaturas de los lugares muestreados, véase tabla 1).

European species found in this study, L. rubellus and A. caliginosa, we detected high genetic diversity (table 2), similar to that found in native and introduced populations by Porco et al. (2013) and Martinsson and Erséus (2017). Other European cosmopolites were represented only by a few individuals and lacked genetic diversity. Another five OTUs could be identified as species of the family Megascolecidae: M. hilgendorfi (Michaelsen, 1892), Amynthas agrestis (Goto and Hatai, 1899), Amynthas phaselus (Hatai, 1930), A. vittatus (Goto and Hatai, 1898), and A. tokioensis (Beddard, 1892). For the latter four this is the first report for the Kuril islands, as well as for Russia as a whole. The Asian M. hilgendorfi was represented by two haplotypes, one of which was identical to AB542629 GenBank from Hokkaido, and the second one differed from those by 2 %. All sequences of A. phaselus were identical to a specimen collected on the Ulleung–do island (South Korea) by Blakemore (2013a) and 99 % similar to an individual from northern Honshu (AB542519). Amynthas vittatus from Kunashir was

identical to the AB542566 accession from Hokkaido. Amynthas agrestis haplotypes were closely (99 %) related to those from Shikoku and Honshu (AB542597, AB542599–AB542602). Two individuals of A. tokioensis differed by one nucleotide substitution, and by two substitutions from AB542557 GenBank entry from Sendai (Honshu). Our specimens of E. japonica from the Kuril Islands were unrelated to ones from Honshu (Blakemore, 2012b), South Korea (Blakemore, 2013b), or Shikoku (GenBank accessions AB542698, AB543188, AB543237) (fig. 3A). In contrast to other Asian species, we detected several closely related haplotypes of E. japonica that could have diverged on the Kuril islands. The estimate of divergence time for our E. japonica sample was 7,580 years (95 % CI, 3,750–11,800 years). Mismatch distribution data (fig. 3B) suggests a recent demographic expansion event for E. japonica. There were another three clusters on the tree (fig. 2). Two of those (shown as E2716 and E2727 on figure 2) were small unpigmented juvenile worms, and E2727

Shekhovtsov et al.

Table 1. Locations sampled in this study: N, number of individuals; r, river; v, village; p, peninsula; l, lake; t, town; GenBank, GenBank accession numbers. Tabla 1. Lugares muestreados en este estudio: N, número de individuos; r, río; v, aldea; p, península; l, lago; t, ciudad; GenBank, números de accesión en la base de datos GenBank. Location

Species

GenBank

M. hilgendorfi, E2954

KX400703–KX400704

Kunashir 1

Mouth of r. Golovnina

Watershed of r. Golovnina

12 M. hilgendorfi, L. rubellus KX400720–KX400731

and r. Khlebnikova

Cape Ivanovskii

20 A. phaselus, L. rubellus, KX400646–KX400649,

Caldera of Golovnina volcano

A. caliginosa, E. japonica KX400659–KX400672

30 M. hilgendorfi, M. agrestis, KX400652–KX400658,

E. japonica, A. vittatus,

KX400680–KX400690,

L. rubellus

KX400701–KX400702,

KX400705–KX400714

KX400613–KX400614,

V. Alekhino

M. hilgendorfi, E. japonica

Cape Znamenka

KX400631–KX400632

M. hilgendorfi, L. rubellus, KX400615–KX400623

E. japonica

West bank of l. Pechianoye

M. hilgendorfi, M. agrestis KX400697–KX400700

r. Andreevka

L. rubellus KX400609–KX400612

Mouth of r. Belkina

A. caliginosa, L. rubellus KX400624–KX400628

10 Mouth of r. Ilyushina

A. caliginosa, A. phaselus KX400650–KX400651

E. japonica KX400629–KX400630

Lower reaches of r. Philatova

12 Lower reaches of r. Saratovskaya 6

A. phaselus, E. japonica,

D. rubidus, D. octaedra

KX400640–KX400645

13 Mouth of r. Tyatina

11 L. rubellus KX400633–KX400639

14 Lovtsova p., shore

of Spokoistvia bay

M. hilgendorfi, L. rubellus,

KX400673–KX400679

E. japonica

15 Lovtsova p., near l. Nadya

L. rubellus, D. octaedra KX400715–KX400719

16 Lovtsova p., lighthouse

L. rubellus KX400691–KX400696

Shikotan 17 N43.8100 E146.7941

A. caliginosa, L. rubellus KX400579–KX400584

18 N43.8828 E146.8447

L. rubellus, E2716, E2717

KX400585–KX400592

19 Krabozavodskoy t., courtyard

E. japonica, A. caliginosa,

KX400593–KX400600,

L. rubellus, A. rosea, O. tyrtaeum KX400608

20 Northwest of Tserkovnaya bay, 7

E. japonica, A. caliginosa,

A. rosea, E2727

100 m from the sea

KX400601–KX400607

Yuri 21

17 E. japonica, M. hilgendorfi, KY750693–KY750709

M. agrestis, A. phaselus,

A. tokioensis, E2954

Total

170

Animal Biodiversity and Conservation 41.1 (2018)

100/100/1.0

L. rubellus

100/100/1.0

A. caliginosa D. rubidus tenuis 100/100/1.0 A. rosea 100/100/1.0 O. tyrtaeum 100/100/1.0 Unidentified specimen E2716 Unidentified specimen E2727

95/98/1.0

95/96/0.99

89/96/1.0

68/71/0.99

100/100/1.0

D. octaedra A. vittatus 99/98/1.0 A. tokioensis 100/100/1.0 A. phaselus

100/100/1.0

100/100/1.0 72/76/1.0

E. japonica

A. agrestis

100/100/1.0

Unidentified specimen E2954

M. hilgendorfi

0.02 Fig. 2. Minimum evolution phylogenetic tree of the obtained cox1 sequences. Numbers near branches indicate bootstrap support for minimum evolution/maximum likelihood/Bayesian posterior probabilities. Species are collapsed into triangles; the base of the triangle is proportional to the number of sequences and its height, to nucleotide diversity within the sample. Fig. 2. Árbol filogenético de las secuencias cox1 obtenidas con el método de la mínima evolución. Los números cercanos a las ramas indican la valoración de bootstrap para los algoritmos de mínima evolución, máxima probabilidad y probabilidad a posteriori de Bayes. Las especies convergen en triángulos cuya base es proporcional al número de secuencias y su altura, a la diversidad nucleotídica de la muestra.

was found alongside A. caliginosa and A. rosea and was provisionally identified as one of those species. E2954 was found together with M. hilgendorfi and generally resembled its morphology. All these three clusters undoubtedly represent distinct species, as their cox1 sequences had approximately 10 % nucleotide differences from any GenBank or BOLD entries. Discussion Our results indicate that DNA barcoding is a promising tool for studying earthworm biodiversity: all previous studies reported six species and subspecies from the Kuril islands, while our rather limited sample yielded fifteen. There are two explanations for this. First, barcoding allows one to identify juvenile or poorly preserved specimens. Second, many earthworm species look similar, and so the species representing small minority of the sample tend to be grouped with more numerous species and overlooked. An example

to this was discovered by Chang et al., 2016): in the USA, three Megascolecidae species, M. hilgendorfi, A. agrestis, and A. tokioensis often occur together, but are often identified as one of the two former species. Co–occurrence of these species was also detected in this study. Multiple studies demonstrated very high genetic differences between even closely located populations in endemic earthworm species (e.g., Novo et al., 2009, 2015; Shekhovtsov et al., 2013, 2015). On the other hand, in the case of cosmopolitan earthworms a singe cox1 haplotype may be spread on multiple continents (e.g., Fernández et al., 2011, 2016; Porco et al., 2013; Martinsson et al., 2015). Based on these contrasting patterns one can hypothesize on the origin of the studied populations. There is almost 400 species of European Lumbricidae endemic to Southern Europe (Hendrix et al., 2008), while only 20–30 of them managed to colonize other regions, and about a half of those are cosmopolites widespread throughout the world. Six of them

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Table 2. Earthworm species detected in this study: N, number of individuals; n, number of unique haplotypes; GD, gene diversity; pi, nucleotide diversity; SD, standard deviation. Tabla 2. Especies de lombriz detectadas en este estudio: N, número de individuos; n, número de haplotipos únicos; GD, diversidad génica; pi, diversidad nucleotídica; SD, desviación estándar.

GD (SD)

pi (SD)

European cosmopolites Lumbricus rubellus

0.776 (0.035)

0.0223 (0.0112)

0.833 (0.060)

0.0126 (0.0070)

Aporrectodea caliginosa caliginosa

A. rosea

3 1

Octolasion tyrtaeum

2 1

Dendrobaena octaedra

4 1

Dendrodrilus rubidus tenuis

1 1

–

Asian species Eisenia japonica

0.949 (0.028)

0.0093 (0.0051)

Metaphire hilgendorfi

0.519 (0.028)

0.0087 (0.0048)

M. agrestis

4 1

Amynthas phaselus

15 1

A. vittatus

2 1

A. tokioensis

1.000 (0.500)

0.0015 (0.0021)

Unidentified specimens E2727

1 1

–

E2954

1 1

–

E2716

2 1

Total 170

were found by us on Kunashir and Shikotan, and all had haplotypes identical to those found in European populations, which confirms recent introduction. The same was true for five species of the family Megascolecidae, which also indicates that these species were introduced from the Japanese Archipelago. Our results indicate that the earthworms of the Southern Kurils are mostly of recent invasive origin, and the native earthworm fauna of the Southern Kurils failed to survive the Last Glacial Period (with the only possible exception of E. japonica). A similar pattern was observed for terrestrial isopods, which find suitable habitats across both islands but do not host any indigenous species (Gongalsky et al., 2014). All terrestrial isopods were either the cosmopolitan Porcellio scaber Latreille, 1804 or amphibiotic species occurring on the littoral zone of the islands. The latter undoubtedly have much better dispersal ability and may have arrived more recently. It should be noted that in many Asian regions, invasive earthworms of European origin often seem to be more vigorous than local species, either replacing them in many habitats (Tiunov et al., 2006; Hendrix et al.,

2008) or co–habiting (Blakemore, 2009, 2012). Asian species are also important as invasive earthworms, e.g., in North America (Hendrix and Bohlen, 2002; Callaham et al., 2003; Blakemore, 2009; Snyder et al., 2011; Greiner et al., 2012). The Southern Kurils is a region where both groups meet. Judging by our collection sample, European and Asian earthworms successfully coexist on Kunashir and Shikotan. It is also interesting that only Asian earthworms were found on Yuri (we managed to collect only a small sample, but the island is also small and ecologically homogeneous). Yuri hosted a Japanese village before the WWII but was uninhabited since that time (apart from military units). On the contrast, Kunashir and Shikotan have functioning Russian settlements. These facts may suggest that Asian earthworms were introduced by Japanese settlers prior to WWII, while European cosmopolites are the results of post–WWII introduction. The only possibly native species on the studied Southern Kuril islands is E. japonica. The obtained cox1 haplotypes were unrelated to those from the Japanese Archipelago or Southern Korea, which

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100/100/1.0

E. japonica Kuril Islands E. japonica Korea E. japonica Honshu

92/94/0.97 93/94/0.99 78/90/0.96 100/100/1.0

E. japonica Shikoku 99/99/1.0

99/99/1.0

E. japonica Honshu E. japonica hiramoto Honshu

99/99/1.0

E. japonica Korea

99/99/1.0

E. japonica vaga Korea E. fetida

0.02

80 70

Frequency

60 50 40 30 20 10 0

3 4 5 6 7 Number of substitutions

Fig. 3. A, minimum evolution phylogenetic tree constructed for E. japonica sequences; branches are collapsed into triangles; the base of the triangle is proportional to the number of sequences, and its height to nucleotide diversity within the sample; B, mismatch distribution for E. japonica sample from the Kuril Islands: horizontal axis, number of pairwise differences; vertical axis, total number of substitutions; solid line, observed distribution; dotted lined, modeled distribution; gray shading, 95 % confidence interval for the modeled distribution. Fig. 3. A, árbol filogenético construido con el método de la mínima evolución para las secuencias de E. japonica; las ramas convergen en triángulos cuya base es proporcional al número de secuencias y su altura, a la diversidad nucleotídica de la muestra; B, distribución de las diferencias entre pares de secuencias para la muestra de E. japonica de las islas Kuriles: eje horizontal, número de diferencias por pares; eje vertical, número total de sustituciones; línea continua, distribución observada; línea punteada, distribución obtenida por el modelo; sombreado gris, intervalo de confianza del 95 % para la distribución obtenida por el modelo.

indicates that this species may actually represent a species complex like the congeneric E. nordenskioldi (Eisen, 1879) (Blakemore, 2013c; Shekhovtsov et al., 2013). Currently available data imply that E. japonica populations from the Kurils underwent a recent demographic expansion after the end of the Last Glacial Period, although further studies of other islands of the archipelago are required to make strong conclusions.

Acknowledgements This study was supported by the MK–6685.2015.4 Grant of the President of the Russian Federation, and the State Assignment no. 0324–2015–0003. We are grateful to L. A. Zelenskaya for her help with the collection of specimens and to N. E. Bazarova for technical assistance.

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models. Bioinformatics, 19: 1572–1574. Roth, A. M., Whitfeld, T. J. S., Lodge, A. G., Eisenhauer, N., Frelich, L. E., Reich, P. B., 2015. Invasive earthworms interact with abiotic conditions to influence the invasion of common buckthorn (Rhamnus cathartica). Oecologia, 178: 219–230. Rougerie, R., Decaëns, T., Deharveng, L., Porco, D., James, S. W., Chang, C. H., Richard, B., Potapov, M., Suhardjono, Y., Hebert, P. D. N., 2009. DNA barcodes for soil animal taxonomy. Pesquisa Agropecuária Brasileira, 44: 789–801. Shekhovtsov, S. V., Berman, D. I., Peltek, S. E., 2015. Phylogeography of the earthworm Eisenia nordenskioldi nordenskioldi (Lumbricidae, Oligochaeta) in Northeastern Eurasia. Doklady Biological Sciences, 461: 1–4. Shekhovtsov, S. V., Golovanova, E. V., Peltek, S. E., 2013. Cryptic diversity within the Nordenskiold’s earthworm, Eisenia nordenskioldi subsp. nordenskioldi (Lumbricidae, Annelida). European Journal of Soil Biology, 58: 13–18. – 2014a. Genetic diversity of the earthworm Octolasion tyrtaeum (Lumbricidae, Annelida). Pedobiologia, 57: 245–250. – 2014b. Invasive lumbricid earthworms of Kamchatka (Oligochaeta). Zoological Studies, 53: 52. – 2016. Different dispersal histories of lineages of the earthworm Aporrectodea caliginosa (Lumbricidae, Annelida) in the Palearctic. Biological Invasions, 18: 751–761. Snyder, B.A., Callaham, Jr. M. A., Hendrix, P. F., 2011. Spatial variability of an invasive earthworm (Amynthas agrestis) population and potential impacts on soil characteristics and millipedes in the Great Smoky Mountains National Park, USA. Biological Invasions, 13: 349–358. Tamura, K., Peterson, D., Peterson, N., Stecher, G., Nei, M., Kumar, S., 2011. MEGA5: Molecular evolutionary genetics analysis using maximum likelihood, evolutionary distance, and maximum parsimony methods. Molecular Biology and Evolution, 28: 2731–2739. Tiunov, A. V., Hale, C. M., Holdsworth, H. M., Vsevolodova–Perel, T. S., 2006. Invasion patterns of Lumbricidae into the previously earthworm–free areas of northeastern Europe and the western Great Lakes region of North America. Biological Invasions, 8: 1223–1234.

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Short–term effects of horse grazing on spider assemblages of a dry meadow (Western France) J. Pétillon, A. François, D. Lafage

Pétillon. J., François, A., Lafage, D., 2018. Short–term effects of horse grazing on spider assemblages of a dry meadow (Western France). Animal Biodiversity and Conservation, 41.1: 19–32. Abstract Short–term effects of horse grazing on spider assemblages of a dry meadow (Western France). In this study, the biodiversity impacts of a little studied herbivore, the horse, were assessed in a high conservation value habitat of dry meadows in Brittany (Western France). Spiders, a diversified and abundant group of predators, were used as bioindicators. Three complementary sampling techniques were used to assess changes in spider assemblages in both soil and vegetation strata, over time (diachronic comparison of managed unit before vs. after management) and space (synchronic comparison of managed vs. control units). Few effects of grazing, i.e. only one significantly indicative species, were found on assemblage composition (ANOSIM), and none on abundances, α– and β– diversities (GLM on pitfall trap data). On the contrary, important differences were found between units before management took place. The main effects of grazing management were revealed over time (after one year), and not between managed and control units (CCA on pitfall trap data and x²–tests on guilds from each sampling method), showing the relevance of a diachronic approach more than a synchronic approach in such management monitoring. Grazing by horses could be relevant to manage meadows because it creates a high spatial heterogeneity, but further (long–term) studies including other model groups are required. Key words: Indicators, Management, Synchronic and diachronic approaches, Araneae, Brittany Resumen Efectos a corto plazo del pastoreo de equinos en las comunidades de arañas de una pradera seca (Francia occidental). En este estudio se evaluaron los efectos de un herbívoro poco estudiado, el caballo, en la biodiversidad de un hábitat de alto valor de conservación en las praderas secas de la Bretaña (Francia occidental). Se utilizaron como bioindicadores las arañas, que constituyen un grupo de depredadores diversificado y abundante. Se emplearon tres técnicas complementarias de toma de muestras para evaluar los cambios en las comunidades de arañas en el estrato edáfico y en la vegetación a lo largo del tiempo (comparación diacrónica de la unidad gestionada antes y después de la gestión) y del espacio (comparación sincrónica de las unidades gestionadas y las de control). Se observaron pocos efectos del pastoreo, esto es, solo una especie significativamente indicativa, en la composición de las comunidades (ANOSIM) y ninguno en la abundancia ni en la diversidad α y β (modelo linear general en los datos obtenidos mediante trampas de caída). Por el contrario, se encontraron diferencias importantes entre las unidades antes de que se llevara a cabo la gestión. Los principales efectos de la ordenación del pastoreo se revelaron con el tiempo (un año después) y no entre las unidades gestionadas y las de control (análisis de correspondencias restringido de los datos obtenidos en la trampa de caída y pruebas de la x² de los gremios obtenidos con cada método de muestreo), lo que pone de manifiesto la importancia de utilizar un planteamiento diacrónico más que uno sincrónico en este tipo de seguimiento de la ordenación. El pastoreo de equinos podría revestir interés para gestionar las praderas porque crea una elevada heterogeneidad espacial; no obstante, es necesario realizar más estudios (a largo plazo) que comprendan otros grupos de modelos. Palabras clave: Indicadores, Ordenación, Planteamientos sincrónico y diacrónico, Araneae, Bretaña Received: 10 VIII 16; Conditional acceptance: 4 XI 16; Final acceptance: 9 VI 17 Julien Pétillon, Denis Lafage, Univ. de Rennes 1, EA 7316, 263 Av. du Général Leclerc, CS 74205, 35042 Rennes Cedex, France.– Denis Lafage, Dept. of Environmental and Life Sciences / Biology, Karlstad Univ., Sweden.– Alexandre François, Emirates Center for Wildlife Propagation, PB 47, 33250 Missour, Maroc. Corresponding author: Julien Pétillon. E–mail: julien.petillon@univ–rennes1.fr ISSN: 1578–665 X eISSN: 2014–928 X

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Introduction Management is frequently carried out in grasslands to simulate past, naturally occurring disturbances such as grazing by large herbivores (Bakker, 1989) or fire (Valkó et al., 2014). The impact of management practices is usually monitored using plant diversity (e.g. Kahmen et al., 2002). However, plant species diversity seems to be a poor predictor for the diversity of other groups, such as arthropods (Kirby, 1992; Morris, 2000; Van Klink et al., 2015). It is therefore necessary to use various groups of organisms to evaluate the potential of a management practice to conserve overall biodiversity. Grazing affects arthropods both directly and indirectly. Direct effects of grazing concern trampling and accidental predation of insects (as well as scavengers and dung feeders: Lumaret et al., 1992), whereas indirect effects are more complex and mainly encompass vegetation and soil–mediated changes (Van Klink et al., 2015). In general, grazing by large herbivores has mostly a negative impact on species richness and abundances of arthropods as it reduces plant cover and biomass. Many large herbivores produce similar effects on the diversity of phytophagous (including flower–visiting) arthropods (Van Klink et al., 2015). Some studies linked changes in arthropod diversity with changes in plant diversity (e.g. Foote and Hornung, 2005), height (Ausden et al., 2005; Ryder et al., 2005) or biomass (Kruess and Tscharntke, 2002), but shifts in functional plant groups are believed to have a higher explanative power (Van Klink et al., 2015). On the contrary, measured indirect effects are likely more dependent on the target taxa or metrics considered, and not consistently reported for all herbivores (Read and Andersen, 2000). As an example, arthropods have been shown to be more diverse and more abundant under ungulate grazing, but arthropod biomass overall is reduced (González–Megías et al., 2004). The effects on predatory arthropods depends on the intensity of grazing (e.g. Dennis et al., 1998; Dennis et al., 2002; Pétillon et al., 2007; Rosa García et al., 2009; Van Klink et al., 2013). Although several studies did not find any effect on arthropod diversity, some found effects on their species composition (Gardner et al., 1997; Woodcock et al., 2005; Fadda et al., 2008). Reported effects of management on arthropod diversity are overall negative, but they differ depending on the types of herbivores and grazing intensity, and are likely context–dependent. Conclusions are also dependent on the taxa and metrics used, and thus on the objectives of management (e.g. Leroy et al., 2014; Török et al., 2016). In this study, we assessed the impact of horse grazing on spiders in a high conservation value habitat in Brittany, dry meadows. Despite its high potential for conservation management (mainly because of an increased plant consummation, for example, higher than similar numbers of cows grazing (Ménard et al., 2002), the effects of horses on arthropod diversity remains poorly documented (Bell et al., 2001; Joern and Laws 2013). Spiders were used because of their bioindicative values (Maelfait and Baert, 1988; Marc et al., 1999), such as to monitor the effects of biological invasions

(Pétillon et al., 2005; Mgobozi et al., 2008), success in habitat restoration (Cristofoli et al., 2010; Pétillon et al., 2014), and changes in land use (Schmidt et al., 2005; Prieto–Benítez and Méndez, 2011). We conducted a pre–management inventory so both synchronic and diachronic approaches could be used. Changes in assemblages of both epigeic and vegetation–dwelling spider assemblages were assessed using (three) complementary sampling methods. Taxonomic and functional changes in community structure were evaluated in the short–term to test the following hypotheses: i) grazing alters the functional composition of a spider community as it will select aeronautic and disturbance–resistant species (Bell et al., 2001) as well as hunting spider species that are less dependent on the plant physiognomy (e.g. Churchill and Ludwig, 2004); ii) abundance and alpha diversity are both expected to be lower in the grazed unit than in the non–grazed unit because of reduced plant biomass and cascading effects (e.g. Kruess and Tscharntke, 2002) whereas β–diversity is expected to be higher in the grazed unit than in the non–grazed unit due to an increased heterogeneity of vegetation (Loucougaray et al., 2004). Material and methods Study site The study site is located in Brittany (Western France), 20 km south–west of Rennes, near the city of Guichen (47° 97' 16 N–1° 89' 20 W). The natural reserve, Vallée du Canut, is a public land of 147 ha, part of a larger Natura 2000 site (total 427 ha) that encompasses meadows, heathlands, forests and a dense network of hedgerows. The dry meadows are located on Cambrian outcrops of red shale on the slopes of a small valley, Le Canut River. The meadows have a N–S slope, with the lower part dominated by bracken (Pteridium aquilinum). Before horses were introduced, the upper part, where the sampling took place, had a mean vegetation height of 1.3 m (visual estimation), and the following dominant plants: Dactylus glomerata, Rumex spp., Stellaria holostea, Plantago lanceolata, Holcus mollis, and different species of Apiaceae, Geranium, Ranunculus, Centaurea and Trifolium. The sampled area was 4 ha in total, subdivided into two experimental units (hence referred to as 'units'), grazed and non–grazed (during the second year of the study). Two horses were then introduced in October 2003 after the first year of sampling. Although the impact of grazers on vegetation cover and diversity was obvious (see fig. 1s in supplementary material), data on habitat changes were not used here because our hypotheses relate to changes in spider assemblages using both diachronic and synchronic approaches (effects of time vs. management). Sampling design and methods In order to sample spiders both at ground level and in the different strata of vegetation, we conducted three complementary sampling methods over the two

Animal Biodiversity and Conservation 41.1 (2018)

years of the study: pitfall trapping, hand–collecting and sweep–netting. Traps to catch ground–active spiders consisted of polypropylene cups (diameter: 12 cm, height: 15 cm) filled with ethanol 70° and covered with a wooden roof to prevent overflow by rainfall. Four traps arranged in a square grid were set up in each unit, and spaced 10 m apart (Topping and Sunderland, 1992; Churchill and Arthur, 1999; Ward et al., 2001), each trap representing one sample (for a total of four replicates per treatment, 16 samples in total). Traps were active on 15 days in June and 15 days in September (i.e. the most favorable periods for spiders in this region: Varet et al., 2013) in 2003 and in 2004. The numbers of individuals caught in traps were divided by the number of days traps were active (Luff, 1975; Curtis, 1980). Hand collections were carried out to sample all visible ground–living spiders, including some less mobile species that would have been few or not sampled by traps (Churchill, 1993). Hand collections were time–standardized (two collectors for 10 min in each unit represent one sample, for a total of eight samples). Vegetation–dwelling spiders were collected using sweep–netting (40 cm diameter) along two parallel transects (20 m + 10 m and 45 sweeps, which represents one sample altogether), two times in each unit (June and September 2003 and 2004, for a total of eight samples). Identification and conservation of specimens Spiders were sorted and stored in tubes with 70 % ethanol (University of Rennes 1, France). Adults were identified to species using Roberts (1987), Heimer and Nentwig (1991) and Roberts (1995). Data analysis Differences in species composition between years and management were visualized by a Venn diagram of shared vs. exclusive species of grazed and un–grazed units before and after the management took place, and tested using an analysis of similarity (ANOSIM) completed by IndVal calculations (Dufrêne and Legendre, 1997) on species abundances using individuals sampled using pitfall traps. A detrended correspondence analysis (DCA) was done on individual counts as response variables and year/management as predictors. A redundancy (RDA) or a constrained correspondence analysis (CCA) was then chosen according to the axis length of the DCA, < 3 or > 4, respectively (Legendre and Gallagher, 2001). Here, a CCA (first axis of the DCA = 1.40) was performed using vegan R package (Oksanen et al., 2013). Monte Carlo tests with 999 permutations were carried out to test the significance of the two factors and constrained analyses axes. Hunting guilds were defined according to Uetz et al. (1999), and assigned to species according to the families they belong to. Guild composition differences between years and management were studied using x²–tests for each sampling method. As the same pitfall traps were operative in 2003 and 2004, abundances and species richness were pair–matched over time and consequently com-

pared using repeated analysis of variance (Pétillon et al., 2010; Lafage and Pétillon, 2014). Tests were performed using spider activity–density and species richness as dependent variables, management (grazed or non–grazed) as a fixed factor, and date (pre– vs. post–grazing) as a within subject effect. If the interaction between fixed factors was not significant (in model 1), a second GLM with logit link and negative binomial distribution (model 2, see O’Hara and Kotze, 2010) was used to test significant effects of separated fixed factors, without their interaction. If the interaction was significant, t–tests were used to detect significant differences between sampling periods. In the case of grazing effects, a significant interaction between management and date was expected (i.e. the within subject factor being expressed differentially for the two units due to grazing effects in one of them). For each analysis, the level of statistical significance used was α = 0.05. β–diversity was estimated through a dissimilarity matrix (corresponding to Sørensen pair–wise dissimilarity), then partitioned into its two components —pecies turnover (βt) and nestedness (βn)— following Baselga (2010) and using the betapart R package (Baselga and Orme, 2012). Only data from pitfall traps were used for the β–diversity. All statistical analyses were performed using R 3.2.3 (R Core Team, 2015). Results A total of 1990 spiders belonging to 55 species were collected in 2003 and 1,040 spiders belonging to 66 species were collected in 2004, respectively (see table 1s in supplementary material). Most spiders were sampled using pitfall traps (1,801 individuals), followed by sweep–net and hand–collection (702 and 537 individuals, respectively). There was a slight increase in the number of species shared between the two units from 2003 to 2004 (22 to 24 species: fig. 1). Overall, no difference in species composition was found between experimental units in 2003 (R = –0.02, P = 0.537) and 2004 (R = 0.31, P = 0.056). Only Pachygnatha degeeri had a significant IndVal, with a preference for the grazed vs. non–grazed unit (IndVal = 4.24 vs. 2.49 respectively, P = 0.028). Furthermore, no significant difference was found in species composition of the non–grazed unit between 2003 and 2004 (R = 0.47, P = 0.057). Conversely, a significant difference in species composition of the grazed unit was found between 2003 and 2004 (R = 0.84, P = 0.026); this was confirmed by the CCA (significant effect of year along the axis 1, F1,13 = 2.17, P < 0.001; fig. 2). CCA was significant (F2,13 = 1.76, P < 0.001), explaining 34.5 % of the observed variance with the first axis being significant (F1,13 = 2.29, P = 0.003). No significant difference was found in guild frequencies (table 1) for individuals sampled by pitfall trap (fig. 3), sweep net (fig. 4) or hand collecting (fig. 5). A significant interaction between 'management' and 'date' effects was found for total activity–den-

Pétillon et al.

Discussion

2003 Before management: 22 shared 18 exclusive species

species

Horse grazing

Control:

15 exclusive species

No management

2004 After management: 24 shared 21 exclusive species

species

Control:

11 exclusive species

Fig. 1. Venn diagram of exclusive vs. shared species of grazed and non–grazed units, before (2003) and after (2004) management took place. Fig. 1. Diagrama de Venn de las especies exclusivas y las compartidas de las unidades con y sin pastoreo, antes (2003) y después (2004) de que se realizara la ordenación.

sity (F1,6 = 7.33, P = 0.035). Total activity–density decreased in the two units (non–grazed: t = 5.56, df = 3, P = 0.011; grazed: t = 8.85, df = 3, P = 0.003) but the decrease was more important in the grazed unit (fig. 6). No significant interaction between 'management' and 'date' effects was found for total species richness (F1,6 < 0.001, P > 0.999). No significant effect of year (F1,13 = 2.049, P = 0.387) or management was found on species richness (F1,13 = 2.049, P = 0.176). β–diversity was nearly constant over time in each unit, yet the nestedness increased in the non–grazed unit (table 2), whereas its species turnover decreased. Differences between the units were similar in the two years of sampling, i.e. before and after management by grazing took place.

Table 1.

Species composition Considering the species that were not shared between the two units in 2004 but that were shared in 2003, the most numerous species was Linyphiids (e.g. Bathyphantes gracilis or Lepthyphantes ericaeus), a family well known for its long–distance dispersal abilities, at both young and adult stages (e.g. Blandenier, 2009; Simonneau et al., 2016). This change is therefore probably due to inter–annual variations in ballooning and/or under–sampling due to their small size. Almost all new shared species in 2004 were also Linyphiids (e.g. Erigone atra, Meioneta mollis, M. beata and Pelecopsis radicicola), and therefore unlikely attributable to any management effect. Among the species that disappeared after the management took place, some are strongly dependent on vegetation structure (a key factor in shaping spider assemblages: e.g. Hatley and Macmahon, 1980), and were thus likely disfavored by the grazing. This could be the case of several ambush hunters (Xysticus cristatus, X. erraticus and X. tortuosus) and web–builders (e.g. Enoplognatha ovata). It is harder to link the disappearance of the other species from the grazed unit to the grazing, because they were not initially in high numbers and/or they are not directly linked to the vegetation. However, the appearance of several thermophilous species (either at ground level or at vegetation level: Myrmarachne formicaria and Tegenaria agrestis, Cyclosa oculata, Dictyna latens and Hypsosinga albovittata respectively; Harvey et al., 2002) can be explained by a more open micro–habitat under grazing (e.g. Gibson et al., 1992). This tendency should be verified by longer–term monitoring. Overall, few, if any (see e.g. the preference of Pachynatha degeeri, an ubiquist species, for the grazed meadow), significant change in species composition were observed after horses were introduced on the dry meadow, a finding in agreement with a few other studies on the impact of grazing on spider composition (Pozzi et al., 1998; Dennis et al., 2001; Pétillon et al. 2007). No significant changes were found after the grazing took place in spider hunting guilds, although web–builders tended to be reduced in

x²–tests on hunting guilds (2df).

Tabla 1. Pruebas de la x² en los gremios de especies cazadoras (2gl). Comparison 2003/2004

Pitfall trap Non–grazed Grazed

Non–grazed/grazed 2003

2004

x² = 7.52, P = 0.023 x² = 37.35, P < 0.001 x² = 5.94, P = 0.051 x²² = 4.76, P = 0.092

Sweep net

x² = 14.97, P < 0.001 x² = 39.04, P < 0.001 x² = 1.02, P = 0.600 x² = 3.94, P = 0.139

Hand collection

x² = 14.79, P < 0.001 x² = 8.77, P = 0.012 x² = 0.89, P = 0.639 x² = 0.16, P = 0.921

Animal Biodiversity and Conservation 41.1 (2018)

Hemm and Höffer, 2012; Ford et al., 2013), mainly explained by a reduced vegetation complexity (Churchill and Ludwig, 2004; Kovac and Mackay, 2009).

Table 2. Partition of beta–diversity into species turnover and nestedness in both grazed and non–grazed units, before (2003) and after (2004) management took place (data from pitfall traps: βt, β turnover; βn, β nestedness).

Structural diversity A decrease in total activity–density was observed in both units over time, but the rate was higher in the grazed unit. This is likely the result of the deleterious effects of grazing on ground–dwelling spiders (see above) and inter–specific competition–mediated changes (Wise, 2006; Van Klink et al., 2015). Competition among spider species is indeed higher in simple habitats (Hurd and Fagan, 1992; Marshall and Rypstra, 1999). Closer attention to change in abiotic and biotic conditions would be necessary to disentangle direct and indirect effects of habitat change on spider abundance (local factors are likely more important here, in an extensively managed land: Horváth et al., 2015). Contrary to several previous studies on other large herbivores (e.g., Pozzi and Borcard, 2001; Pétillon et al., 2007), no significant effect of grazing was found on spider species richness. Indeed, as for many other taxa, and as with many other disturbances, grazing by large herbivores is usually reported to be negative on spiders (Bell et al., 2001; see also the meta–analysis of Prieto–Benítez and Méndez, 2011). Mechanisms usually involved in such a rich-

Tabla 2. Partición de la diversidad beta en el anidamiento y la renovación de especies en unidades con y sin pastoreo, antes (2003) y después (2004) de que se realizara la ordenación (datos de trampas de caída: βt, β renovación; βn, β anidamiento). Year Unit

β (Sørensen) βt

2003 Grazed Non–grazed 2004 Grazed Non–grazed

βn

0.561

0.449

0.112

0.461

0.429

0.033

0.539

0.458

0.082

0.431

0.282

0.148

the grazed unit (visible for sweep–net sampling only). Hunting guilds are usually affected by grazing, with several studies showing a decrease in web–builders (Gibson et al., 1992; Kirby, 1992; Bell et al., 2001;

ovata experta vittata tortuosus

cristatus clercki erraticus neglecta mordax vagans spinimana antica leopardus atra

proxima

CCA2

retusus pullata

Grazed

palitans mollis

ruricola

2003

prativaga simplex latitans sanctuaria degeeri gracilispraeticus terricola 2004 tenuis Ungrazed ericaceus

rurestris palustris dentipalpis agrestis

beata

subaequalis

–2

acerbus cuneata

pulverulenta nigriceps juncea albimana

–4

pallidus atrotibialis insignis arundineti pumilla

–4

–2

CCA1

Fig. 2. Projection of significant variables from the CCA on spider species. Pitfall traps are represented by circles and species by crosses. Fig. 2. Previsión de las variables significativas a partir del análisis de correspondencias restringido en especies de arañas. Las trampas de caída se representan con círculos y las especies, con cruces.

Pétillon et al.

100

Grazed

Non–grazed

Percent

75 Ambush hunters 50

Ground runners Web–builders

2003

2004

Year

2003

2004

Fig. 3. Percentage of activity density of each hunting guild sampled by pitfall trapping in 2003 and 2004. Fig. 3. Porcentaje de la densidad de actividad de cada gremio de especies cazadoras recogidas en trampas de caída en 2003 y 2004.

Non–grazed

Grazed

100

Percent

75 Ambush hunters

Ground runners Web–builders

2003

2004

Year

2003

2004

Fig. 4. Percentage of abundance of each hunting guild sampled by sweep netting in 2003 and 2004. Fig. 4. Porcentaje de la abundancia de cada gremio de especies cazadoras recogidas en red de barrido en 2003 y 2004.

Animal Biodiversity and Conservation 41.1 (2018)

Non–grazed

Grazed

100

Percent

75 Ambush hunters 50

Ground runners Web–builders

2003

2004

Year

2003

2004

Fig. 5. Percentage of abundance of each hunting guild sampled by hand collecting in 2003 and 2004. Fig. 5. Porcentaje de la abundancia de cada gremio de especies cazadoras recogidas manualmente en 2003 y 2004.

Log (AD + 1)

5.25

5.00

4.75

4.50

4.25

Non–grazed Grazed 2003

Year

2004

Fig. 6. Activity–density of spiders in the grazed and non–grazed units in 2003 and 2004 (data from pitfall traps). Fig. 6. Actividad–densidad de arañas en las unidades con y sin pastoreo en 2003 y 2004 (datos obtenidos de las trampas de caída).

ness decrease encompass local habitat simplification (litter and vegetation strata: Dennis et al., 2001) and related food reduction (Kruess and Tscharntke, 2002; for soil functioning: Koppel et al., 1997). However, Lafage et al. (2014) recently demonstrated that spiders were not influenced by the plant biomass (either in terms of abundance or species richness). Finally, it is possible that the study period was too short to show effects. Our hypothesis about increased β–diversity was not verified. This negative result has two main explanations: the study period on the effects of grazing was too short to be visible (although arthropods, and especially spiders, are known to quickly react to changes in habitat structure: Pétillon et al., 2014), and/or the effects of grazing were counterbalanced by inter–annual variations. We consider our second explanation is the most likely because it would explain the increased nestedness in the ungrazed unit. Plant heterogeneity is usually higher under grazing treatment (e.g. Gallet and Rozé, 2001; Van Klink et al., 2015), which reinforces the idea that plants and arthropods, here spiders, react in a different way to management practices (probably due to differences in mobility: Lafage et al., 2015; Lafage and Pétillon, 2016). We should stress that our sampling design did not necessarily encompass the spatial heterogeneity resulting from grazing in general (e.g. for spiders: Bonte et al., 2000), and especially by horses (Loucougaray

et al., 2004), and would deserve a higher replication for all the sampling methods to properly assess spider diversity (including beta–diversity, see e.g. Klimek et al., 2008; Báldi et al., 2013) Concluding remarks In this study, despite the existence of true replicates within each unit, units were confounded with the management treatment, which can be considered as a case of pseudoreplication in the sense of Hurlbert (1984). Comparing stations between different sites is likely to increase inter–class variance through the existence of other co–varying factors (Oksanen, 2001). Here, even at a small scale, we showed a high variance between stations, with differences between units before the grazing took place. This underlines the importance of carrying out diachronic approaches, also because the effects of management were sometimes only visible over time, and not when comparing grazed vs. non– grazed units. Although spiders present high dispersal ability and high mobility (Lafage and Pétillon, 2014) that may hide or decrease the effects of management (Pech et al., 2015), they can be considered a relevant group for monitoring biodiversity consequences of management, bringing complementary information to changes in vegetation structure. Acknowledgements We are grateful to the land manager Jean–François Lebas (Conseil Général d’Ille–et–Vilaine) for funding and providing access to the site, Dr. Jesse Eiben and two anonymous referees for useful comments, and GRETIA (Armorican Entomological Society) for technical support. References Ausden, M., Hall, M., Pearson, P., Strudwick, T., 2005. The effects of grazing on tall–herb fen vegetation and molluscs. Biological Conservation, 122: 317–326. Bakker, J. P., 1989. Nature Management by Grazing and Cutting. Kluwer Academic Publishers, Dordrecht. Báldi, A., Batáry, P., Kleijn, D., 2013. Effects of grazing and biogeographic regions on grassland biodiversity in Hungary –analysing assemblages of 1200 species. Agriculture, Ecosystems and Environment, 166: 28–34. Baselga, A., 2010. Partitioning the turnover and nestedness components of beta diversity. Global Ecology and Biogeography, 19: 134–143. Baselga. A., Orme, C. D. L., 2012. betapart : an R package for the study of beta diversity. Methods in Ecology and Evolution, 3: 808–812. Bell, J. R., Wheater, C. P., Cullen, W. R., 2001. The implications of grassland and heathland management for the conservation of spider communities. Journal of Zoology, 255: 377–387.

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Supplementary material

Table 1s. Taxonomic list of the species collected in grazed and non–grazed units (2003–2004), all sampling methods considered. Tabla 1s. Lista taxonómica de las especies recogidas en las unidades con y sin pastoreo (2003 y 2004), se han tenido en cuenta todos los métodos de muestreo.

2003 Species

Non–grazed Grazed

2004 Non–grazed

Grazed

Agalenatea redii (Scopoli, 1763)

Agroeca proxima (O. Pickard–Cambridge, 1871)

Agroeca sp.

Agyneta affinis (Kulczyński, 1898)

Agyneta mollis (O. Pickard–Cambridge, 1871)

Agyneta rurestris (C. L. Koch, 1836)

Alopecosa cuneata (Clerck, 1757)

Alopecosa pulverulenta (Clerck, 1757)

Alopecosa sp.

Anelosimus vittatus (C. L. Koch, 1836)

Araneus diadematus Clerck, 1757

Araneus quadratus Clerck, 1757

Arctosa leopardus (Sundevall, 1833)

Argiope bruennichi (Scopoli, 1772)

Aulonia albimana (Walckenaer, 1805)

Bathyphantes approximatus (O. Pickard–Cambridge, 1871)

Bathyphantes gracilis (Blackwall, 1841)

Brigittea latens (Fabricius, 1775)

Cercidia prominens (Westring, 1851)

Clubiona neglecta O. Pickard–Cambridge, 1862 Clubiona sp.

2 7

Cyclosa oculata (Walckenaer, 1802)

Drassodes sp.

Drassyllus praeficus (L. Koch, 1866)

Enoplognatha mordax (Thorell, 1875)

Enoplognatha ovata (Clerck, 1757)

Enoplognatha sp.

2 1

Eratigena agrestis (Walckenaer, 1802)

Erigone atra Blackwall, 1833

Erigone dentipalpis (Wider, 1834)

Ero cambridgei Kulczyński, 1911 Genus sp.

Heliophanus sp.

1 397

119

172 2

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Table 1s. (Cont.)

2003 Species

2004

Non–grazed Grazed

Non–grazed

Grazed

Hypsosinga albovittata (Westring, 1851) Larinioides cornutus (Clerck, 1757)

Lepthyphantes sp.

Mangora acalypha (Walckenaer, 1802)

Micrargus subaequalis (Westring, 1851)

Myrmarachne formicaria (De Geer, 1778)

Neoscona adianta (Walckenaer, 1802)

Neottiura bimaculata (Linnaeus, 1767)

Neriene clathrata (Sundevall, 1830)

Oedothorax fuscus (Blackwall, 1834)

Oedothorax retusus (Westring, 1851)

Ozyptila sanctuaria (O. Pickard–Cambridge, 1871)

Ozyptila simplex (O. Pickard–Cambridge, 1862)

Ozyptila sp.

Pachygnatha clercki Sundevall, 1823 Pachygnatha degeeri Sundevall, 1830 Palliduphantes ericaeus (Blackwall, 1853)

252 8

4 239

151

133

Palliduphantes insignis (O. Pickard–Cambridge, 1913)

Palliduphantes pallidus (O. Pickard–Cambridge, 1871)

Pardosa nigriceps (Thorell, 1856)

Pardosa palustris (Linnaeus, 1758)

Pardosa prativaga (L. Koch, 1870)

Pardosa proxima (C. L. Koch, 1847)

Pardosa pullata (Clerck, 1757)

117

Pardosa sp.

131

Pelecopsis radicicola (L. Koch, 1872)

Pardosa vittata (Keyserling, 1863) Pirata sp.

8 6

Piratula latitans (Blackwall, 1841)

Pisaura mirabilis (Clerck, 1757)

Pocadicnemis juncea Locket & Millidge, 1953

Pocadicnemis pumila (Blackwall, 1841)

Prinerigone vagans (Audouin, 1826) Robertus arundineti (O. Pickard–Cambridge, 1871)

Scotina celans (Blackwall, 1841)

Sintula corniger (Blackwall, 1856)

Stemonyphantes lineatus (Linnaeus, 1758)

Tallusia experta (O. Pickard–Cambridge, 1871)

Tapinopa longidens (Wider, 1834)

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Table 1s. (Cont.)

2003 Species Tenuiphantes tenuis (Blackwall, 1852)

2004

Non–grazed Grazed 3

Non–grazed

Grazed 8

Tibellus oblongus (Walckenaer, 1802)

Tibellus sp.

14 17

13 12

Trochosa ruricola (De Geer, 1778)

Trochosa sp.

Trochosa terricola Thorell, 1856

Walckenaeria acuminata Blackwall, 1833

Walckenaeria antica (Wider, 1834)

Walckenaeria atrotibialis (O. Pickard–Cambridge, 1878) Xysticus acerbus Thorell, 1872

Xysticus erraticus (Blackwall, 1834)

Xysticus sp.

Xysticus cristatus (Clerck, 1757) Xysticus ferrugineus Menge, 1876

1 115

Xysticus tortuosus Simon, 1932

25 1

Zelotes sp.

Zora armillata Simon, 1878 3 1 Zora sp. Zora spinimana (Sundevall, 1833)

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Fig. 1: 16 VI 2003

Fig. 2: 24 VII 2003

Fig. 3: 16 VI 2003

Fig. 4: 24 VII 2003

Fig. 5: 24 VII 2003

Fig. 6: 05 IX 2003

Fig. 7: 05 IX 2003

Fig. 8: 20 IX 2003

Fig. 1s. Change in grazed vs. non–grazed vegetation over time. Fig. 1s. Cambio en el tiempo de la vegetación en las unidades con y sin pastoreo.

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Long–term variation of demographic parameters in four small game species in Europe: opportunities and limits to test for a global pattern A. Gée, M. Sarasa, O. Pays

Gée, A., Sarasa, M., Pays, O., 2018. Long–term variation of demographic parameters in four small game species in Europe: opportunities and limits to test for a global pattern. Animal Biodiversity and Conservation, 41.1: 33–60. Abstract Long–term variation of demographic parameters in four small game species in Europe: opportunities and limits to test for a global pattern. For decades, decreases in several populations of some small sedentary game species have been reported in Europe. From the literature, we extracted mortality and reproductive rates that were available for European populations in four iconic species, the grey partridge (Perdix perdix), the black grouse (Tetrao tetrix), the capercaillie (T. urogallus) and the brown hare (Lepus europaeus), to examine how demographic parameters vary with time. Our study revealed the need to consider many confounding factors (age, sex, origin of studied individuals, season, country and methods) and the scarcity of recent demographic studies. Statistical analyses showed contrasted patterns of demographic traits with time within and between species. Our results highlighted that there may be consistency with a population decrease in grey partridge and black grouse that has been reported in the literature. However, analyses in capercaillie and brown hare showed less support for a population decrease at the European scale. The significant effects of interactions between time and age (in grey partridge, capercaillie and brown hare), method or origin of individuals on demographic traits and the emergence of contrasted patterns between short, intermediate and long monitoring periods (in grey partridge and black grouse) suggested that further studies should pay particular attention to potential confounding factors. Finally, the lack of recent data and doubts about the relative importance of reported causal factors indicate the need for further studies on the links between demographic traits, densities and environmental changes in the long term, and particularly on the role of predation and habitat change. Key words: Grey partridge, Black grouse, Capercaillie, Brown hare, Demography, Population monitoring Resumen Variación a largo plazo de los parámetros demográficos en cuatro especies de caza menor en Europa: oportunidades y limitaciones. Durante décadas, se ha registrado la disminución de varias poblaciones de algunas especies sedentarias de caza menor. A partir de los estudios publicados, se extrajeron las tasas de mortalidad y reproducción documentadas para las poblaciones europeas de cuatro especies icónicas: la perdiz pardilla (Perdix perdix), el gallo lira (Tetrao tetrix), el urogallo (T. urogallus) y la liebre europea (Lepus europaeus), para examinar la variación de los parámetros demográficos con el tiempo. Nuestro estudio reveló la necesidad de considerar muchos factores que pueden confundir (edad, sexo, origen de los individuos estudiados, estación, país y métodos) y la escasez de estudios demográficos recientes. Los análisis estadísticos mostraron diferentes patrones de variación de los parámetros demográficos con el tiempo dentro de cada especie y entre ellas. Nuestros resultados subrayaron que estos patrones pueden ser coherentes con la disminución de las poblaciones de perdiz pardilla y gallo lira que han sido documentadas en las publicaciones especializadas. No obstante, los análisis sobre el urogallo y la liebre europea mostraron menos indicios que apuntaran a una disminución de las poblaciones a escala europea. El efecto significativo de las interacciones entre el tiempo y la edad (en la perdiz pardilla, el urogallo y la liebre europea), los métodos o el origen de los individuos en los parámetros demográficos y la aparición de patrones distintos entre los períodos de seguimiento a corto, medio y largo plazo (en la perdiz pardilla y el gallo lira) sugieren que los nuevos estudios deberían dedicar una atención especial a los factores que pueden confundir. Por último, la falta de datos recientes sobre la ISSN: 1578–665 X eISSN: 2014–928 X

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importancia relativa de los factores causales conocidos y las dudas que ello genera indican la necesidad de realizar más estudios sobre los vínculos entre los parámetros demográficos, las densidades y los cambios ambientales a largo plazo, en particular sobre la función de la depredación y el cambio del hábitat. Palabras clave: Perdiz pardilla, Gallo lira, Urogallo, Liebre europea, Demografía, Seguimiento de poblaciones Received: 4 V 16; Conditional acceptance: 25 X 16; Final acceptance: 13 VI 17 Alexandre Gée and Olivier Pays, UMR 6554 CNRS, LETG–Angers, Université d’Angers, 2 Bd Lavoisier, 49045 Angers, France.– Mathieu Sarasa, Fédération Nationale des Chasseurs, 13 rue du Général Leclerc, 92136 Issy les Moulineaux, France (former address); BEOPS, 1 Esplanade Compans Caffarelli, 31000 Toulouse, France (present address). Corresponding author: M. Sarasa. E–mail: msarasa@beops.fr; contact@beops.fr

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Introduction For decades, decreases in several populations of farmland and woodland species have been highlighted in different areas of geographical Europe by various stakeholders (Comolet–Tirman et al., 2015), including hunting associations (Tapper, 2001; Vallance et al., 2008) and academic researchers (Sumption and Flowerdew, 1985; Smith et al., 2005; Storch, 2007; Inger et al., 2014). Wide–scale analyses controlling for the effects of, for instance, heterogeneity in studied areas and methods and individuals used as biological models would be useful to investigate whether demographic parameters might be used as proxy for variation in the abundance of small game species in Europe. Direct testing of populations decreases from abundance estimates might appear to be an intuitive approach to examine the temporal variations of small game species in Europe (European Environment Agency, 2014). However, reliable long–term series of abundance (or density) are available from very few study sites (Potts and Aebischer, 1995) and short–term estimates of abundance or density may be misleading because of the diversity of methods or the uneven probability of detection across habitats (Thompson, 2002). Furthermore, the restocking of game species might be a confounding factor of abundance estimates (Díaz–Fernández et al., 2013). An overall population decrease should theoretically be associated with a decrease in demographic traits, including survival and reproduction. However, a decrease in only one trait, particularly reproduction (Panek, 1992b; Lindström et al., 1997), might require complementary information as potential demographic trade–offs or density–dependent processes might be involved in population dynamics (Panek, 1992b; Lindström et al., 1997; Grimm and Storch, 2000; Sachot et al., 2006). Variations in mortality and reproductive rates over the last half century might be difficult to highlight as some factors might have changed concomitantly, such as monitoring and estimation methods and the origin of birds. Captures, animal handling, and tracking equipment (including radio–transmitters or GPS) might affect animal mobility and behaviour and also reproductive and mortality rates (Bro et al., 1999; Mech and Barber 2002; Barron et al., 2010). The number of breeding black grouse hens (Tetrao tetrix), for instance, can be halved when they are equipped (Caizergues and Ellison, 1998). Furthermore, for restocking purposes, hunting associations have made many attempts to release captive–reared individuals. However, various studies have shown that these birds often have different foraging and anti– predator behaviour, and poorer body condition than wild birds (Parish and Sotherton, 2007; Rantanen et al., 2010; Rymešová et al., 2013), affecting mortality and reproductive rates (Birkan, 1977a, 1977b). The mortality of reared birds can be ten times greater than that of wild birds (Putaala et al., 2001), reaching 90% in one study (Rymešová et al., 2013). Age and sex might also have a strong effect on mortality and breeding success. Yearling black grouse hens, for

instance, might have a lower clutch size and might be more likely to lose their entire brood than adults (Willebrand, 1992; Marjakangas and Törmälä, 1997). In the same way, mortality in summer in grey partridge males (Perdix perdix) has been reported to be lower than in females (Birkan et al., 1992) while sex and age effects have been reported in other grouse and age, sex and origin should thus be controlled when analysing patterns of mortality and reproductive estimates over time. From the data available in the scientific literature, our aim was to explore whether estimates of mortality and reproductive rates might be used as empirical support of the demographic declines in four small game species in Europe: the grey partridge, the black grouse, the capercaillie (Tetrao urogallus), and the brown hare (Lepus europaeus). We assumed that an increase in mortality estimates associated with a decrease in reproduction estimates suggests population declines (Bro et al., 2000; Ludwig et al., 2006). However, uncertainty may exist if only one proxy (reproduction or survival) varies with time or if potential factors, such as methods, origin, or sex, significantly affect estimates of demographic traits. This question was studied in small game species by testing for the effect of year on estimates of reproductive and mortality rates. When possible, we controlled for the effect of the following confounding factors: sex, age and origin of individuals, method, season and country of monitoring. Material and methods Data collection We identified all articles available up to May 2015 from the Web of Science and Science Direct reporting demographic traits (mortality and reproductive rates) in grey partridge, black grouse, capercaillie and brown hare. These articles were collected through a systematic search that included key–words related to demographic traits (e.g. survival, mortality, reproduction) and names of species. Articles on wild rabbit (Oryctolagus cuniculus), pheasant (Phasianus colchicus), red–legged partridge (Alectoris rufa), hazel grouse (Tetrastes bonasia) and rock ptarmigan (Lagopus muta) were not considered here as they were too rare to allow statistical analyses on the variation of estimates of demographic parameters with time. For a few studies, data were not directly accessible (e.g. studies in non–English languages) but were available from scientific reviews or comparative studies. Data were not considered when: 1) the estimation value revealed confounding factors (e.g. when estimates of winter mortality from counts showed a 'negative' mortality, meaning that the number of migrants was greater than the number of deaths and emigrants); and 2) data were missing for several factors that we wanted to test (e.g. the year in which rates were estimated) or control for (e.g. the country where the study was performed). Some data sets can be found in several articles (e.g. the Game and Wildlife Conser-

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vation Trust’s Count Scheme in England); these were therefore gathered only once to avoid double counting and were completed when additional information was available from a different article with the same data set. When data were presented as averages over several years, special attention was paid to using the raw data whenever possible. From all indicators of mortality rates that were used in the literature, we extracted an individual´s probability of dying during the monitoring period considered in the study. Back–transformation from partial estimates (e.g. for data presented in a fragmented format linked to predation) to overall mortality was carried out when possible (i.e. when the total population size was given) (discarded otherwise). Similarly, several reproductive traits were reported in the literature, including clutch and brood size, the proportion of nests in which at least one egg hatched or in which at least one young survived, the proportion of females with a nest or with young, the probability of re–nesting, the number of chicks, young or fledglings per hen, per adult, or per couple. The estimates of reproductive parameters are presented as rates (e.g. proportion of breeding hens) or real numbers (e.g. brood size, clutch size). We studied the variation in mortality rates and reproduction through the post–hatching brood size, the number of fledglings per hen, and the proportion of hens with fledglings; this provided enough data from the literature to run statistical analyses. Finally, we noted the year and season of the study. For studies exceeding one year and presenting pooled data, the median year was considered. We dealt with the marked heterogeneity in the monitoring periods by considering three classes: (1) one season or less; (2) from one season to less than one year; and (3) one year or more. To control for potentially confounding effects, we also noted the age, sex and origin (reared or wild) of the studied individuals, the monitoring method, and the country of the study, although a detailed presentation of their effects is beyond the scope of this study. Data analysis For each species, we ran linear models to test for the long–term effect of time on mortality and reproductive rates (Tx) that were logit–transformed to fulfil statistical requirements related to the homogeneity of variances and linearisation. We started our procedure by performing a model (Eq. 1) including time (x1) and its square (i.e. time², x2) to investigate the nonlinear relationship between time and demographic traits: Logit(Tx) = β0 + β1 x1 + β2 x2 + ε

(Eq. 1)

with β, the model parameters and ε, the residual error. As it is hazardous to compare estimated mortality rates from 3–month studies with those calculated over several years, we ran different sets of separate models for each species when the duration of the study was: (1) one season or less, (2) from one season to less than one year, and (3) one year or more. To

include a reasonable number of degrees of freedom in each procedure, we tested controlling factors such as method, sex, age, origin and country separately. Within each model set, each controlling factor (xn) was included in the procedure (i.e. in Eq. 1) and the deviance between the model with and without the controlling factor was tested using F–tests. Two–way interactions were also tested with the same procedure. When the deviance between the two models was significant, this factor (or the interaction) was kept throughout the procedure. Once all significant controlling factors and interactions were determined, we tested the effect of time and time2. Concerning mortality in capercaillie and brown hare, the limiting number of data required a specific statistical procedure. The effect of time was tested using a forward selection procedure that compared deviance between a null model and a model including time only. This was then compared with a model that included time and time² using the F–test. Then, we tested for the effect of the controlling factors using the same procedure, and we also tested two–way interactions. The structural limits of the data set hindered mixed– effect models, including the identity of the publication as a random effect. However, we checked that residuals fulfilled statistical requirements including a normal distribution of residuals, a lack of heterogeneity of variance when plotting residuals against fitted values, and a lack of temporal correlation of residuals when using a partial autocorrelation function (pacf). All statistical procedures were performed using R software (R Development Core Team, 2015). Results Data collection The available data highlighted a large difference between the four studied species in the number of studies in the literature. The number of studies on grey partridge was considerably higher than that for the other three species (tables 1 and 2). Grey partridge For grey partridge we collected data for mortality rates from 30 papers and data for reproductive traits from 26 papers. The lowest mortality rate was 0.012 (in a short monitoring period; Kaiser, 1998) while the highest rate was 0.99 (in a long monitoring period; Buner et al., 2011). Variation in reproduction with time was analysed using post–hatching brood size, which ranged from 0.66 to 20.00 (Rands, 1985; Rymešová et al., 2013), and the proportion of hens with fledglings, which ranged from 0.17 to 0.66 (Buner et al., 2011). Mortality rates Analysis showed a significant association between time and mortality rate estimates in grey partridge, but this link differed between the temporal scales at which it was assessed (table 3).

Animal Biodiversity and Conservation 41.1 (2018)

Table 1. Mortality data collected from the scientific literature. Tabla 1. Datos de mortalidad recolectados en publicaciones científicas.

Species Numer of studies Number of data Period References Grey partridge 30 403 1917–2011 Middleton (1935), Jenkins (1961), Birkan et al. (1975), Potts (1980), Montagna and Meriggi (1991), Birkan et al. (1992), Nösel (1992); Panek (1992a), Potts and Aebischer (1995), Boatman and Brockless (1998), Kaiser (1998), Sotherton (1998), Bro et al. (2000, 2001), Faragó (2001), Putaala et al. (2001), Reitz and Mayot (2001), Aebischer et al. (2002), Panek (2002), Aebischer and Ewald (2004), Panek (2005, 2006), Parish and Sotherton (2007), Watson et al. (2007), Aebischer and Baines (2008), Besnard et al. (2010), Rantanen et al. (2010), Buner et al. (2011), Draycott (2012), Rymešová et al. (2012) Black grouse 8 77 1933–2008 Angelstam (1984), Baines (1991), Caizergues and Ellison (1997), Warren and Baines (2002), Baines et al. (2007), Bowker et al. (2007), Wegge and Rolstad (2011), Pekkola et al. (2014) Capercaillie 7 49 1984–2002 Gjerde and Wegge (1989), Schroth (1991), Storch (1994), Moss et al. (2000), Wegge and Kastdalen (2007), Wegge and Rolstad (2011), Åhlen et al. (2013) Brown hare 7 (+ 1*) 52 1959–2008 Broekhuizen (1979), Frylestam (1979), Wasilewski (1991), Hansen (1992), Marboutin and Peroux (1995), Marboutin and Hansen (1998), Misiorowska and Wasilewski (2012) * Abildgard et al. (1972) in Broekhuizen (1979), Marboutin and Hansen (1998)

Controlling for the effects of age, sex, origin, method, country and season, mortality rates from short monitoring periods (equal to or less than one season) showed a significant decrease between 1917 and 2010 (time: β ± SE = –0.02 ± 0.01; fig. 1A; table 3), while this link was affected by age, method and origin (fig. 1s in supplementary material). Adult mortality showed a high increase with time (β ± SE = 0.77 ± 0.38) while chick mortality decreased (β ± SE = –0.72 ± 0.34; fig. 1s in supplementary material). Mortality increased with time when the methods used in the studies were by radio–collar and census (β ± SE = 0.77 ± 0.38), while the link of the modelling method was relatively stable (fig. 1s in supplementary material). Controlling for the effects of age, method, season, sex, origin and country, mortality rates from intermediate monitoring periods showed a significant increase between 1933 and 2011 (time: β ± SE = 0.02 ± 0.01; fig. 1B; table 3) while this link was affected by the age of birds (fig. 2s in supplementary material). Mortality of adults showed the highest increase with time (β ± SE = 0.03 ± 0.01) while the dynamic of chick mortality did not differ significantly over time for the sample at hand (fig. 2s in supplementary material). For long monitoring periods (greater than or equal to one year) and controlling for the effects

of method, age, sex, origin and country, mortality rates increased significantly with time from 1978 to 2010 (time: β ± SE = 0.04 ± 0.14; fig. 1C; table 3). Thus, the apparent decrease in mortality rates between 1975 and 2010 from short monitoring studies contrasted with an apparent increase in mortality rates from longer monitoring studies. Adjusted R² of the three models (for short, intermediate and long–monitoring periods) was 0.33, 0.39 and 0.77, respectively. Reproduction Controlling for the effects of country, method and season, we found a strong overall decline in brood size estimates between 1910 and 2010 (β ± SE = –0.11 ± 0.01; fig. 2A; table 3) while this pattern interacted with method and season (fig. 3s in supplementary material). The few data in June suggested an increasing pattern in brood size (β ± SE = 3.60 ± 1.30) while the prevailing data in non– determined months suggested an overall decrease. Similarly, our results showed that the proportion of hens with fledglings decreased linearly between 1968 and 2005 (time: β ± SE = –0.03 ± 0.01; fig. 2B; table 3) when the effects of country, method, season and origin were controlled for. Adjusted R² of the models of brood size and proportion of hens with fledglings were 0.80 and 0.23, respectively.

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Table 2. Reproductive data collected from the scientific literature: * numbers in brackets (black grouse and capercaillie) correspond to data gathered in Jahren (2012). Tabla 2. Datos de reproducción recolectados en publicaciones científicas: * los números entre paréntesis (en el gallo lira y el urogallo) corresponden a los datos recopilados en Jahren (2012).

Species Number of studies Number of data Period References Grey partridge 26 580 1910–2010 Middleton (1935, 1936, 1937), Jenkins (1961), Potts (1970, 1980), Rands (1985); Montagna and Meriggi (1991), Birkan et al. (1992), Nösel (1992), Panek (1992a), Thomaides and Papageorgiou (1992), Potts and Aebischer (1995), Tapper et al. (1996), Boatman and Brockless (1998), Bro et al. (2000), Aebischer and Ewald (2004), Bro et al. (2004), Panek (2006), Parish and Sotherton (2007), Besnard et al. (2010), Buner et al. (2011), Rymešová et al. (2013) Black grouse 16 (+ 7*) 304 (+ 7*) 1975–2009 Angelstam et al. (1984), Moss (1986), Baines (1991), Willebrand (1992), Caizergues and Ellison (1997), Kurki et al. (1997), Marjakangas and Törmälä (1997), Caizergues and Ellison (1998), Warren and Baines (2002), Summers et al. (2004), Baines et al. (2007), Bowker et al. (2007), Grant et al. (2009), Ludwig et al. (2010), Summers et al. (2010), Wegge and Rolstad (2011), Jahren (2012) * Siivonen (1953), Myrberget and Hagen (1974), Lindén (1981), Ellison et al. (1982), Angelstam (1983), Storaas and Wegge (1984), Brittas and Willebrand (1991) in Jahren (2012) Capercaillie 17 (+ 9*) 247 (+ 10*) 1975–2009 Moss (1986, 1987), Leclercq (1988), Storch (1994), Kurki et al. (1997), Picozzi et al. (1999), Moss et al. (2000), Proctor and Summers (2002), Saniga (2002), Summers et al. (2004), Wegge and Kastdalen (2007), Summers et al. (2009, 2010), Wegge and Rolstad (2011), Jahren (2012) * Höglund (1953), Siivonen (1953), Myrberget and Hagen (1974), Wegge and Grasaas (1977), Lindén (1981), Jones (1982), Klaus (1984), Spidsø et al. (1984), Storaas and Wegge (1984) in Jahren (2012)

Black grouse Mortality rates in black grouse were collected from eight papers and reproductive traits were collected from 23 papers. The lowest mortality rate was 0.003 (in a short monitoring period; Baines et al., 2007) and the highest was 0.97 (in a long monitoring period (Bowker et al., 2007)). The variation in reproductive rates over time was analysed using the post–hatching brood size, which ranged from 2.0 to 4.82 (Kurki et al,. 1997; Caizergues and Ellison, 1998), and the number of fledglings per hen, which ranged from 0 to 6.3 (Moss, 1986; Caizergues and Ellison, 1998; Summers et al., 2010). Mortality rates Estimates of mortality rate decreased linearly with time (time: β ± SE = –0.31 ± 0.04; fig. 3A; table 4) for short monitoring periods between 1985 and 2005 when controlling for the effects of method, season, age, and country. The mortality rates appeared stable between 1980 and 2008 for intermediate monitoring periods, controlling for the effects of method, sex, age and country (fig. 3B). Estimates of mortality rate from

long monitoring studies increased between 1980 and 2008 (time: β ± SE = 0.05 ± 0.03; fig. 3C; table 4) when controlling for the effect of method, sex, age and country. Thus, the temporal pattern of black grouse mortality rates differed according to the temporal scale at which the rates were assessed. Adjusted R² of the three models was 0.77, 0.63 and 0.94, respectively. Reproduction We detected a significant nonlinear relationship between time and brood size. Controlling for the effects of country, method, and season, we found that average brood size increased from 1988 to 1995 (time: β ± SE = 50.78 ± 30.58; time²: β ± SE = 0.01 ± 0.01; fig. 4A; table 4). Nevertheless, this pattern seemed to depend greatly on the data heterogeneity with time, including one point at earlier and later periods. When controlling for the effect of country, method, season and age, our data indicated that the number of fledglings per female decreased linearly with time between 1975 and 2009 (time: β ± SE = –0.02 ± 0.01; fig. 4B; table 4). Adjusted R² of the two models was 0.57 and 0.04, respectively.

Animal Biodiversity and Conservation 41.1 (2018)

Table 3. Results of the links between demographic traits and time in grey partridge Perdix perdix: factors, F–values and p–values. Tabla 3. Resultados relativos a la relación entre los parámetros demográficos y el tiempo para la perdiz pardilla Perdix perdix: factores y valores F y p.

Mortality rates

Reproduction

Short monitoring periods

Brood size

Time

F1,115 = 5.81, p = 0.02

Time

Time x age

F1,115 = 5.26, p = 0.02

Time x method

F1,156 = 8.70, p < 0.01

Time x method

F2,115 = 3.55, p = 0.03

Time x season

F1,156 = 5.70, p = 0.02

F1,156 = 175.25, p < 0.001

Time x origin of birds F1,115 = 5.26, p = 0.02

Method

F1,156 = 79.64, p < 0.001

Country

F5,156 = 22.47, p < 0.001

Season

F2,156 = 46.96, p < 0.001

Age

F2,115 = 51.61, p < 0.001

Sex

F2,115 = 1.34, p = 0.27

Origin of birds

F2,115 = 6.87, p < 0.01

Proportion of hens with fledglings

F2,115 = 42.69, p < 0.001

Time

Country

F6,115 = 35.20, p < 0.001

Method

F1,24 = 2.50, p = 0.13

Season

F1,115 = 16.79, p < 0.001

Country

F1,24 = 2.50, p = 0.13

Season

F2,24 = 1.39, p = 0.27

Origin of birds

F2,24 = 3.41, p = 0.05

Method

Intermediate monitoring periods Time Time x age

F2,144 = 5.54, p < 0.01 F2,144 = 5.54, p < 0.01

Age

F3,144 = 6.13, p < 0.001

Sex

F2,144 = 0.60, p = 0.55

Origin of birds

F1,24 = 5.26, p = 0.03

F2,144 = 11.47, p < 0.001

Method

F2,144 = 0.53, p = 0.59

Country

F6,144 = 4.87, p < 0.001

Season

F1,144 = 0.52, p = 0.47

Long monitoring periods Time

F1,89 = 8.71, p < 0.01

Age

F3,89 = 7.25, p < 0.001

Sex

F2,89 = 10.88, p < 0.001

Origin of birds

F2,89 = 24.02, p < 0.001

Method

F1,89 = 13.52, p < 0.001

Country

F3,89 = 16.07, p < 0.001

Capercaillie In capercaillie, mortality rates and reproductive traits were collected from seven and 26 articles respectively. The lowest mortality rate was 0.01 (in an intermediate monitoring period (Wegge and Rolstad, 2011) and the highest was 0.93 (in a long monitoring period; Moss et al., 2000). The proportion of hens with fledglings and the number of fledglings were considered to analyse the variation in reproductive rates with time. The former rose from 0.02 to 0.77 (Leclercq, 1988) and the latter from 0 to 7 (Leclercq, 1988; Saniga, 2002; Summers et al., 2004, 2010).

Mortality rates Controlling for the effect of age, country, sex, and season, we found that estimates of mortality rate for intermediate monitoring periods decreased for a linear way with time (time: β ± SE = –0.10 ± 0.05; fig. 5A; table 5); this link was affected by the age of birds (fig. 4s in supplementary material). The few data on young birds suggested an increase in mortality with time (β ± E = 1.78 ± 0.56; fig. 4s in supplementary material) while data on adults suggested an overall decrease. For long monitoring periods, mortality rates appeared stable between 1995 and 2002 (fig. 5B), and none of the control factors seemed to have a

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A Logit (mortality rate)

3 2 1 0 –1 –2 –3 –4 –5 1910

1930

1950 1970 Year

1990

2010

Logit (mortality rate)

4 3 2 1 0 –1 –2 –3 –4 1930 1940 1950 1960 1970 1980 1990 2000 2010 Year

C Logit (mortality rate)

5 4 3 2 1 0

–1 1975 1980 1985 1990 1995 2000 2005 Year

2010

Fig. 1. Estimates of mortality rate (logit transformed) of grey partridge over time (average years) in Europe for a: A, short (one season or less); B, intermediate (from one season to less than one year); and C, long monitoring periods (greater than or equal to one year). The line indicates a significant correlation with time (see results). See figures 1s and 2s in supplementary material for the significant effects of age, methods and origin. Fig. 1. Estimación de la tasa de mortalidad (transformada logarítmicamente) de la perdiz pardilla a lo largo del tiempo (promedio de los años) en Europa para periodos de seguimiento: A, cortos (una estación o menos); B, intermedios (más de una estación, pero menos de un año); y C, largos (un año o más). La recta indica una correlación significativa con el tiempo (véanse los resultados). En las figuras 1s y 2s del material suplementario se ilustran los efectos significativos de la edad, los métodos y el origen de los individuos.

Animal Biodiversity and Conservation 41.1 (2018)

Brood size

15 10 5 0 1905 B

1925

1945 1965 Year

1985

2005

Logit (proportion of hens with fledglings)

–1

–2 1965

1975

1985 Year

1995

2005

Fig. 2. Estimates in grey partridge in Europe of: A, average brood size; and B, proportion (logit transformed) of hens with fledglings over time. The line indicates a significant correlation with time (see results). See figure 3s in supplementary material for the significant effect of season of study and monitoring method Fig. 2. Estimación, relativa a la perdiz pardilla en Europa de: A, el tamaño medio de puesta; B, la proporción (transformada logarítmicamente) de hembras con pollos a lo largo del tiempo. La recta indica una correlación significativa con el tiempo (véanse los resultados). En la figura 3s del material suplementario se ilustra el efecto significativo de la temporada de estudio y del método de seguimiento.

significant effect (table 5). Adjusted R² of the model on mortality rate was 0.78. Reproduction Controlling for the effect of country, we observed that the proportion of hens that raised fledglings increased with time in a nonlinear way from 1976 to 2000 (time²: β ± SE = 0.003 ± 0.004; time: β ± SE = –12.26 ± 16.42; fig. 6A; table 5). Controlling for the effects of country, method and season, we found that the overall decrease in the number of fledglings per female with time from 1975

to 2009 (time: β ± SE = –0.06 ± 0.01; fig. 6B; table 5) was affected by country (fig. 5s in supplementary material). The number of fledglings per female increased with time in Norway (β ± SE = 0.19 ± 0.09), but decreased in France, Scotland and Slovakia (fig. 5s in supplementary material). Adjusted R² of the two models was 0.47 and 0.63, respectively. Brown hare Mortality rates for brown hare were collected from seven papers that referred to intermediate and long

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A Logit (mortality rate)

1 0 –1 –2 –3 –4 –5 –6 1985

1990

1995 Year

2000

2005

Logit (mortality rate)

2 1 0 –1 –2 –3 –4 1980

1990

Year

2000

2010

Logit (mortality rate)

4 3 2 1 0 –1 –2 1975

1985

1995 Year

2005

Fig. 3. Estimates of mortality rate (logit transformed) of black grouse over time (average years) in Europe for: A, short (one season or less); B, intermediate (from one season to less than one year); and C, long monitoring periods (greater than or equal to one year). The line indicates a significant correlation with time (see results). Fig. 3. Estimación de la tasa de mortalidad (transformada logarítmicamente) del gallo lira a lo largo del tiempo (promedio de los años) en Europa para periodos de seguimiento: A, cortos (una estación o menos); B, intermedios (más de una estación, pero menos de un año); y C, largos (un año o más). La recta indica una correlación significativa con el tiempo (véanse los resultados).

Animal Biodiversity and Conservation 41.1 (2018)

A 5

Brood size

4 3 2 1 1980 B

1985

1990 Year

1995

2000

Fledglings per hen

6 5 4 3 2 1 0 1975

1985

1995 Year

2005

Fig. 4. Estimates, in black grouse in Europe of: A, average brood size; and B, average number of fledglings per hen over time. The line indicates a significant correlation with time (see results). Fig. 4. Estimación, relativa al gallo lira en Europa, de: A, el tamaño medio de puesta; y B, la media de pollos por hembra a lo largo del tiempo. La recta indica una correlación significativa con el tiempo (véanse los resultados).

monitoring periods. The lowest and highest mortality rates were 0.12 (Frylestam, 1979) and 0.94 (Wasilewski, 1991) (both in intermediate monitoring periods). Controlling for age, season and country, we found that mortality rates of brown hare in intermediate monitoring periods seemed stable between 1966 and 1989 (fig. 7A, table 5) although this link was affected by age. The mortality of young hare increased over time (β ± SE = 0.13 ± 0.03) while the mortality of adults decreased (β ± SE = –0.05 ± 0.03) (fig. 6s in supplementary material). Mortality rates in long monitoring periods appeared stable between 1959 and 2007 (fig. 7B) while none of the control factors appeared significant (age, sex, country, method; table 5). Adjusted R² of the model on mortality rates was 0.71.

Discussion Variation of demographic traits over time Grey partridge and black grouse were the species for which data most consistently indicated an alteration in their demographic traits over time. In grey partridge, the decrease in reproductive traits (fig. 2) and the increase in annual mortality over time from long monitoring studies (fig. 1) supports the observed European decrease in population reported in the literature (e.g. Kuijper et al., 2009). The results showed that post–hatching brood size appeared stable between 1910 and 1980 but that reproductive success seemed to collapse from the 1980s (be discussed below). In black grouse, annual mortality rates from long monitoring studies increased between 1980 and 2008

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Table 4. Results regarding the links between demographic traits and time in Black grouse Tetrao tetrix: factors, F–values and p–values. Tabla 4. Resultados relativos a la relación entre los parámetros demográficos y el tiempo para el gallo lira Tetrao tetrix: factores y valores F y p. Reproduction

Mortality rates

Brood size

Short monitoring periods

F1,53 = 12.63, p < 0.001

Time

F1,7 = 46.84, p <0.01

Time

Age

F2,7 = 3.16, p = 0.09

Method

F1,53 = 4.92, p = 0.03

Method

F1,7 = 1.55, p = 0.24

Country

F4,53 = 19.43, p < 0.001

Country

F2,7 = 2.43, p = 0.14

Season

F2,53 = 27.84, p < 0.001

Season

F2,7 = 0.87, p = 0.37

Number of fledglings per female

Intermediate monitoring periods

Time

F1,91 = 5.22, p = 0.02

Time

F1,33 = 0.30, p = 0.59

Method

F1,91 = 2.32, p = 0.13

Age

F2,33 = 5.56, p < 0.01

Country

F6,91 = 1.83, p = 0.10

F2,33 = 2.45, p = 0.10

Season

F3,91 = 1.71, p = 0.17

Age

F2,91 = 1.52, p = 0.23

Sex Method

F1,33 = 11.00, p < 0.01

Country

F5,33 = 1.26, p < 0.001

Long monitoring periods Time

F1,3 = 11.46, p = 0.01

Age

F2,3 = 2.81, p = 0.11

Sex

F2,3 = 23.65, p < 0.001

Method

F2,3 = 0.79, p = 0.48

Country

F4,3 = 21.21, p < 0.001

(fig. 3). Moreover, post–hatching brood size seemed to increase between 1989 and 1995, contrasting with a decrease in the average number of fledglings per female (fig. 4), suggesting a reduction in chick survival. Overall, these results agree with the past population decreases reported for this species (Storch 2007; Ludwig et al., 2008; Sim et al., 2008). The other species, capercaillie and brown hare, gave fewer clues about population decreases. Capercaillie annual mortality seemed stable between 1985 and 2002 (fig. 5), although the small amount indicated the observed pattern should be interpreted with caution. The proportion of females with fledglings may have increased between 1977 and 1999, while this pattern was balanced by a decrease in the number of fledglings per female from 1975 to 2009 (fig. 6). With such contrasting results, it is hard to highlight an overall European pattern in capercaillie populations. Previous studies highlighted predictions that depended on the structural assumptions of the models in this species in Finland (Kangas and Kurki, 2000). Nevertheless, some authors have claimed from population modelling that populations might decline when the mortality rates does not compensate for low reproductive success or recruitment (Fernández–

Olalla et al., 2012). This hyppothesis is supported by recent studies in a population of capercaillie affected by marten species (Martes martes and M. foina) (Moreno–Opo et al. 2015). In brown hare, we were unable to analyse the importance of changes in reproductive traits. Some studies have reported that adult survival is relatively constant (in non–hunted areas in the absence of disease outbreaks) and that breeding success can vary between sites and years (Marboutin and Peroux, 1995; Marboutin and Hansen, 1998; Marboutin et al., 2003). However, the low amount of data, the apparent stability of mortality, and the age–dependent pattern suggested by the results (fig. 6s in supplementary material) prevent clear conclusions on the strength of the decline reported since the 1960s (Edwards et al., 2000). In addition, studies reporting an overall population decline in brown hare rely on hunting bags as a proxy of population abundance, which may be rather hazardous. Indeed, hunting bag dynamics do not necessarily match abundance dynamics in game species because of, for instance, voluntary hunting restraint, saturation effects or hunting legislation changes (Sarasa and Sarasa, 2013; Kahlert et al., 2015).

Animal Biodiversity and Conservation 41.1 (2018)

Table 5. Results of the links between demographic traits and time in capercaillie Tetrao urogallus and in Brown hare Lepus europaeus: factors, F–values and p–values. Tabla 5. Resultados relativos a la relación entre los parámetros demográficos y el tiempo para el urogallo Tetrao urogallus y la liebre europea Lepus europaeus: factores y valores F y p.

Mortality rates of capercaillie Intermediate monitoring periods

Reproduction of capercaillie Proportion of hens that raised fledglings

Time

F1,18 = 8.06, p < 0.01

Time

Time x age

F1,18 = 9.98, p < 0.01

Time²

Age

F1,18 = 21.38, p < 0.001

Sex

F1,18 = 0.44, p = 0.52

Country

F1,18 = 18.83, p < 0.001

Season

F1,18 = 7.98, p = 0.08

Country

F1,49 = 1.96, p = 0.17 F1,49 = 12.41, p = < 0.001 F3,49 = 11.69, p < 0.001

Number of fledglings per female Time Time x country

F1,122 = 12.33, p < 0.001 F3,122 = 6.12, p < 0.001

Method

F1,122 = 0.45, p = 0.51

Time

F1,9 = 0.29, p = 0.60

Country

F3,122 = 39.34, p < 0.001

Age

F1,9 = 0.00, p = 0.97

Season

F3,122 = 13.19, p < 0.001

Sex

F1,9 = 0.03, p = 0.87

Country

F1,9 = 1.51, p = 0.26

Long monitoring periods

Mortality rates of brown hare Intermediate monitoring periods Time Time x age Age

F1,31 = 0.01, p = 0.91

Mortality rates of brown hare Long monitoring periods Time

F1,1 = 3.09, p = 0.11;

F2,31 = 9.86, p < 0.001

Age

F1,1 = 1.64, p = 0.23;

F2,31 = 28.11, p < 0.001

Sex

F2,1 = 1.99, p = 0.19 F3,1 = 1.72, p = 0.24

Season

F1,31 = 8.06, p < 0.01

Country

F2,31 = 0.73, p = 0.49

Method v3,1 = 1.72, p = 0.24

Explanatory factors The decrease in annual survival rates in grey partridge and black grouse might be triggered by habitat changes (Berg 1997; Stoate et al., 2001; Panek 2002; Benton et al., 2003). The observed patterns with time are consistent with the homogenisation reported in farmland systems during agriculture intensification (Benton et al., 2002; Wretenberg et al., 2006). Reduction or removal of hedges that formerly fragmented the fields has been reported as an important factor in grey partridge population decline, by increasing predation rates of nests, chicks and adults, and reducing arthropod density and thus chick food availability (Potts, 1980, 1986; Rands, 1985, 1987; Kaiser, 1998; Šálek et al., 2004; Holland et al., 2006). In addition, direct and indirect effects of pesticides (e.g. neonicotinoids) on wildlife (Goulson, 2013; Gibbons et al., 2015; Pisa et al., 2015; Van der Sluijs et al., 2015) including farmland birds (Boatman et al., 2004; Lopez–Antia et al.,

2013; Hallmann et al., 2014; Lopez–Antia et al., 2015) have been reported particularly regarding to the massive use of neonicotinoids after 1990 (Simon–Delso et al., 2015). The drastic drop in post–hatching brood size in grey partridge in the 1980s (fig. 2A) points out the sudden appearance of a major destabilising factor rather than to progressive changes in the ecological interactions. This might be a consequence of an abrupt intensification of agriculture (wide re–parcelling, scrapping of edges and scattered trees, rise of oilseed rape, use of pesticides and inorganic fertilizers) occurring in the 1980s in Europe (Chamberlain, 2000; Plieninger, 2012). This illustrative example should be considered in further agri–environmental schemes. For instance, with adaptive measures (restriction or banning of problematic pesticides, increase in edge abundance, artificial feeding and predator control), the density of grey partridge in August can even be quadrupled (Faragó, 2001). On the contrary, studies have pointed out that too small cover trips or an

Gée et al.

Logit (mortality rate)

2 1 0 –1 –2 –3 –4 –5 –6 1980

1985

1990 1995 Year

2000

2005

1985

1990 1995 Year

2000

2005

Logit (mortality rate)

3 2 1 0 –1 –2 1980

Fig. 5. Estimates of mortality rate (logit transformed) of capercaillie over time (average years) in Europe for: A, intermediate (from one season to less than one year); and B, long monitoring periods (greater than or equal to one year). The line indicates a significant correlation with time (but see figure 4s in supplementary material for the age effect). Fig. 5. Estimación de la tasa de mortalidad (transformada logarítmicamente) del urogallo a lo largo del tiempo (promedio de los años) en Europa para periodos de seguimiento: A, intermedios (más de una estación, pero menos de un año); y B, largos (un año o más). La recta indica una correlación significativa con el tiempo (en la figura 4s del material suplementario se ilustra el efecto de la edad).

inadequate amount of cover trips at the landscape scale might be associated with an increase of predation rates (Bro et al., 2004). These contrasted results in several countries (Faragó, 2001; Bro et al., 2004; Ewald et al., 2010) suggest that further studies are needed to understand ecological and landscape drivers of populations dynamics in grey partridge. Black grouse and capercaillie seem to be highly influenced by land cover (Storaas and Wegge, 1987; Wegge and Rolstad, 2011; Seibold et al., 2013; White et al., 2013) (including via resource availability), especially chick survival which is particularly dependent on the abundance of insects (Picozzi et al., 1999; Ludwig et al., 2010). Nevertheless, their populations appear to respond differently to habitat changes, which could

explain the difference in observed patterns between the two species in our results. The clear–cutting of forest and its replacement by young plantations may have contrasted and un–expected effects on abundance and sex–ratio of black grouse and capercaillie in August, whereas similar positive effects on the number of young per hen were observed in Sweden in the two species (Wegge and Rolstad, 2011). Habitat homogenisation might also have a strong impact on populations. For instance, black grouse selects transitional habitats (moors, heaths, meadows, young and open forests), which have been massively converted into farmland and mature forest during recent decades (Pearce‐Higgins et al., 2007; Ludwig et al., 2009a, 2009b). Black grouse may benefit from habitat

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Logit (proportion of hens with fledgings)

1 0 –1 –2 –3 –4 –5 1975

1980

1985 Year

1990

1995

2000

Fledgings per hen

7 6 5 4 3 2 1 0 1975

1985

Year

1995

2005

Fig. 6. Estimates related to capercaillie in Europe regarding: A, proportion of hens with fledglings (logit transformed); and B, average number of fledglings per hen over time. (See figure 5s in supplementary material for the significant effect of country). Fig. 6. Estimación, relativa al urogallo en Europa, de: A, la proporción (transformada logarítmicamente) de hembras con pollos; y B, la media de pollos por hembra a lo largo del tiempo. (En la figura 5s del material suplementario se ilustra el efecto significativo del país).

changes as early stages of trees provide food and nest sites (Scridel et al., 2017). In contrast, capercaillie selectes for mature forest although the importance of habitat matrix including key ressources should not be neglected (Quevedo et al., 2006a, 2006b; Teuscher et al., 2011). The reported agricultural abandonment and reforestation in mountain areas during the last century, such as in the Alps and the Pyrenees (MacDonald et al., 2000; Lasanta–Martínez et al., 2005; Caplat et al., 2006; Agnoletti, 2007; Kulakowski et al., 2011), have probably disadvantaged populations of black grouse while favouring capercaillie. Incidentally, we did not find any clear support of an overall decline in this latter species. Thus, demographic traits and abundance estimates need to be updated for the reappraisal of the dynamics of forest grouse, particularly capercaillie, in European countries.

The opposing patterns in mortality rates for different lengths of monitoring period in grey partridge and black grouse might appear contradictory at first sight. However, the three monitoring periods underline different information. Mortality data from short monitoring periods corresponded mostly to the breeding season and thus chick mortality. Chick survival seemed unable to explain the observed decline between 1968 and 1993 in England (Potts and Aebischer, 1995), while other studies reported that it was crucial in the understanding of the population dynamics in grey partridge (Potts, 1980, 1986; Topping et al., 2010). This could agree with the decrease in mortality rates in shorter monitoring periods (fig. 1A) while populations declined. However, this finding may also suggest that the main studied season (spring–summer) does not have a major effect on the decline and that processes

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A Logit (mortality rate)

3 2 1 0 –1 –2 –3 1960 1965 1970 1975 1980 1985 1990 1995 Year

Logit (mortality rate)

1.5 1 0.5 0 –0.5 –1 1950

1960

1970

1980 Year

1990

2000

2010

Fig. 7. Estimates of mortality rate (logit transformed) of brown hare over time (average years) in Europe for: A, intermediate (from one season to less than one year); and B, long monitoring periods (greater than or equal to one year). (See figure 6s in supplementary material for the significant effect of the interaction time x age). Fig. 7. Estimación de la tasa de mortalidad (transformada logarítmicamente) de la liebre europea a lo largo del tiempo (promedio de los años) en Europa para periodos de seguimiento: A, intermedios (más de una estación, pero menos de un año); y B, largos (un año o más). (En la figura 6s del material suplementario se ilustra el efecto significativo de la interacción entre el tiempo y la edad).

occurring in autumn–winter should be studied further. In forest grouse, most studies attributed the population decline to a reduced recruitment of young (Kurki et al., 1997; Storch, 2007; Ludwig et al., 2009a; Summers et al., 2010), a mortality of adults (Caizergues and Ellison,1997) or both adults and fledged young (Wegge and Rolstad, 2011). Our study suggests that reproduction is the demographic trait that is the most consistent with the demographic decline of grouse. The two proxies suggesting reproductive patterns indicated a decrease in black grouse (fig. 4), whereas they were contrasted in capercaillie (fig. 6), for which it is hazardous to infer a demographic evolution. Mortality patterns in black grouse also support the idea of a population decrease, as mortality seemed to increase

for longer monitoring periods (fig. 3C). However, this was not the case in capercaillie (fig. 5B). During the last decades, hunting management has become increasingly restrictive (Angulo and Villafuerte, 2004; Čas, 2008; Ministère de l’Écologie du Développement durable des Transports et du Logement, 2012; Ikonen et al., 2014) and abundance of small game hunters was reported as rather decreasing (Lecocq and Meine, 1998). Thus, absolute hunting pressure on small game species is expected to have decreased during this recent period. Hunting bans have already been highlighted to limit incentives from hunters and landowners for wildlife monitoring although habitat preservation is ultimately needed for sustainability of small game species (Storch, 2007; Čas, 2008). Thus,

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this overall situation and our results might emphasis the urgency of addressing further the predominant factors in these species, such as habitat (Evans, 2004; Whittingham and Evans, 2004; Scridel et al., 2017).

2005 in these two species. This lack of information prevents an accurate, updated analysis of recent population dynamics.

Limitations of the analysis

Conclusion

Several limitations might affect the present study. First, if articles that reported past declines (Sirkiä et al., 2010) rather than stable or recent increasing populations (Aebischer and Ewald, 2012; Scridel et al., 2017) were more likely published, a pessimistic appraisal of European dynamics in wild species might be highlighted. In addition, data from studies that use successive counts of the number of individuals are prone to the effects of confounding factors (e.g. detectability problems, migration events) (Thompson, 2002). Morover, the use of radio–collars makes individual–based studies possible but might affect survival and reproduction (Putaala et al., 1997; Caizergues and Ellison, 1998; Mech and Barber, 2002; Barron et al., 2010; Gibson et al., 2013). Both monitoring techniques have a different bias, but radio–collars have mainly been used since the 1980s, implying a possible shift in bias over time (Barron et al., 2010; Naef–Daenzer and Grüebler 2014). Miniaturisation in biotelemetry is expected to provide more accurate estimates of demographic traits in future studies that will use lighter transmitters (< 3 % of the bird’s weight) (Casas et al., 2015). Concerning reproductive traits, the number of fledglings per female is reported heterogeneously across the literature (chicks almost old enough to fly and young), which might also involve a bias. Data on mortality and reproduction in four small game species including wild rabbit, hazel grouse, pheasant and red–legged partridge are rare and thus prevent robust statistical analyses. We attempted to control for a part of the variability in mortality and reproduction data including factors such as age, method, season, sex, origin and country in our models. Efforts to homogenise demographic traits in further studies, particularly proxies of reproductive success, could strengthen future reappraisals of the literature. Our data sets included spatial heterogeneity. For example, in grey partridge, most of data in reproduction traits (n = 580) came from studies investigating populations in England (73.4 %) and actually few articles existed from populations in France (11.0 %), Poland (10.3 %), or other countries (5.2 % together). Such disparities also occurred in black grouse, capercaillie and brown hare and should alert on the challenge to address scenario on expected future decline at the scale of distribution area. In addition, a temporal heterogeneity is present between and within species. For instance, data available on the mortality in grey partridge ranged from 1917 to 2011; however, there were few data before 1970 for intermediate and longer monitoring periods. An insufficient number of data related to reproduction was available in grey partridge after 1995. Capercaillie and black grouse seem to be rather recently studied species with no data on reproduction and mortality rates before 1975 while data are rather scarce after

The present study suggests that data from demographic parameters (mortality and reproductive traits) extracted from published articles might be used to investigate population trends with time. The effects of age, sex, origin of birds, season, country and method on demographic traits should encourage further studies to consider these factors in their experimental design to avoid misleading findings regarding potential patterns. For instance, methodological issues might arise from animal equipment that would affect survival and reproduction rate. Similar comments can be addressed for the effect of the monitoring period, particularly when investigating mortality rates. Although it is not the case in capercaillie and brown hare, our study seems to support the overall decrease that has been reported in grey partridge and black grouse during the last decades. Although black grouse, capercaillie, brown hare and grey partridge were relatively well–studied species, studies from the last 10–20 years on demographic traits are scarce or lacking, hampering the understanding of ongoing demographic processes and current status of species. Thus, long–term studies including recent publications in peer–reviewed journals are needed to investigate the links between mortality, reproduction and density, predation, and habitat changes. Acknowledgements We thank the editor and the anonymous reviewers for helpful comments on an earlier version and Carol Robins for editing the English of the manuscript. This study was made possible thanks to the financial support provided by the Fédération Nationale des Chasseurs (FNC) (FNC–PSN–PR3–2013 and FNC– PSN–PR15–2014). References Abildgard, F., Andersen, J., Barndorff–Nielsen, O. 1972. The hare population (Lepus europaeus Pallas) of Illumö island, Denmark: a report on the analysis of the data from 1957–1970. Danish Review of Game Biology, 6(5), 32pp. Aebischer, N., Ewald, J., 2012. The grey partridge in the UK: population status, research, policy and prospects. Animal Biodiversity and Conservation, 35: 353–362. Aebischer, N., Ewald, J., Potts, G., 2002. Preliminary results from using GIS to examine spatial variation in grey partridge demography over 30 years in Sussex, UK. In: 24 th International Congress of the International Union of Game Biologists: 23–33 (C. Thomaides, N. Kypridemos, Eds.). Hellenic Hunting

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Individual and environmental determinants of daily black grouse nest survival rates at variable predator densities. Annales Zoologici Fennici, 47: 387–397. Ludwig, T., Storch, I., Gärtner, S., 2009a. Large–scale land use change may explain bird species declines in semi–natural areas: the case of black grouse population collapse in Lower Saxony, Germany. Journal of Ornithology, 150: 871–882. – 2009b. Historic landscape change and habitat loss: the case of black grouse in Lower Saxony, Germany. Landscape Ecology, 24: 533–546. Ludwig, T., Storch, I., Wübbenhorst, J., 2008. How the black grouse was lost: historic reconstruction of its status and distribution in Lower Saxony (Germany). Journal of Ornithology, 149: 587–596. MacDonald, D., Crabtree, J. R., Wiesinger, G., Dax, T., Stamou, N., Fleury, P., Gutierrez Lazpita, J., Gibon, A., 2000. Agricultural abandonment in mountain areas of Europe: environmental consequences and policy response. Journal of Environmental Management, 59: 47–69. Marboutin, E., Bray, Y., Péroux, R., Mauvy, B., Lartiges, A., 2003. Population dynamics in European hare: breeding parameters and sustainable harvest rates. Journal of Applied Ecology, 40: 580–591. Marboutin, E., Hansen, K., 1998. Survival rates in a nonharvested brown hare population. Journal of Wildlife Management, 62: 772–779. Marboutin, E., Peroux, R., 1995. Survival pattern of European hare in a decreasing population. Journal of Applied Ecology, 32: 809–816. Marjakangas, A., Törmälä, L., 1997. Female age and breeding performance in a cyclic population of black grouse Tetrao tetrix. Wildlife Biology, 3: 195–203. Mech, L. D., Barber, S. M., 2002. A critique of wildlife radio–tracking and its use in national parks. Biological Resources Management Division, US National Park Service, Fort Collins, CO Technical Report. Middleton, A., 1935. The population of partridges (Perdix perdix) in 1933 and 1934 in Great Britain. Journal of Animal Ecology, 4: 137–145. – 1936. The population of partridges (Perdix perdix) in Great Britain during 1935. Journal of Animal Ecology, 5: 252–261. – 1937. The population of partridges (Perdix perdix) in Great Britain during 1936. Journal of Animal Ecology, 6: 318–321. Ministère de l’Écologie du Développement durable des Transports et du Logement, 2012. Stratégie nationale d’actions en faveur du grand tétras Tétrao urogallus major 2012–2021: 172. Misiorowska, M., Wasilewski, M., 2012. Survival and causes of death among released brown hares (Lepus europaeus Pallas, 1778) in Central Poland. Acta Theriologica, 57: 305–312. Montagna, D., Meriggi, A., 1991. Population dynamics of grey partridge (Perdix perdix) in northern Italy. Italian Journal of Zoology, 58: 151–155. Moreno–Opo, R., Afonso, I., Jiménez, J., Fernández– Olalla, M., Canut, J., García–Ferré, D., Piqué, J., García F., Roig J., Muñoz–Igualada J., González L. M., López–Bao, J. V., 2015. Is it necessary managing carnivores to reverse the decline of

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Supplementary material

A Logit (mortality rate)

3 2 1 0

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–2 –3 –4 –5 1910

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Fig. 1s. Estimates of mortality rate (logit transformed) of grey partridge over time in shorter monitoring periods in relation to: A, age; B, method; and C, origin of birds. ND, not determined. Fig. 1s. Estimación de la tasa de mortalidad (transformada logarítmicamente) de la perdiz pardilla a lo largo del tiempo (año promedio) para periodos cortos en relación con: A, la edad; B, el método de estimación; y C, el origen de las aves. ND, no determinado.

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Logit (mortality rate)

3 2 1

Adult

Young

–1

Chick ND

–2 –3 –4 1930 1940 1950 1960 1970 1980 1990 2000 2010 Year

Fig. 2s. Estimates of mortality rate (logit transformed) of grey partridge over time in intermediate monitoring periods in relation to age of birds: ND, not determined. Fig. 2s. Estimación de la tasa de mortalidad (transformada logarítmicamente) de la perdiz pardilla a lo largo del tiempo (año promedio) para periodos intermedios en relación con la edad de las aves: ND, no determinado.

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Brood size

15 August

June ND

0 1900 B

1950

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15 Census

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0 1900

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Fig. 3s. Estimates, in grey partridge of average brood size over time and in relation to: A, season of study; and B, monitoring method. ND, not determined. Fig. 3s. Estimación, relativa a la perdiz pardilla, del tamaño medio de puesta a lo largo del tiempo y según: A, la temporada de estudio; y B, el método. ND, no determinado.

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Logit (mortality rate)

2 1 0 Young

–1

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1985

1990

1995 Year

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Fig. 4s. Estimates of mortality rate (logit transformed) of capercaillie over time (average years) in intermediate monitoring periods in relation to age. Fig. 4s. Estimación de la tasa de mortalidad (transformada logarítmicamente) del urogallo a lo largo del tiempo (promedio de los años) en periodos intermedios y según la edad.

Fledging per hen

6 5

France

Norway Scotland

Slovakia

2 1 0 1975

1985

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2005

Fig. 5s. Estimates, in capercaillie, of average number of fledglings per hen over time in different countries. Fig. 5s. Estimación, relativa al urogallo, de la media de pollos por hembra a lo largo del tiempo y según el país.

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Logit (mortality rate)

3 2 1 Adult

ND Young

–1 –2 –3 1960

1970

1980 Year

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2000

Fig. 6s. Estimates of mortality rate (logit transformed) of brown hare over time (average years) for intermediate monitoring periods in relation and according to age: ND, not determined. Fig. 6s. Estimación de la tasa de mortalidad (transformada logarítmicamente) de la liebre europea a lo largo del tiempo (promedio de los años) para periodos intermedios y según la edad: ND, no determinado.

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Genetic structure and demographic history of the endemic Mediterranean scallop Pecten jacobaeus inferred from mitochondrial 16s DNA sequence analysis K. Telahigue, T. Hajji, M. Cafsi, C. Saavedra Telahigue, K., Hajji, T., El Cafsi, M., Saavedra, C., 2018. Genetic structure and demographic history of the endemic Mediterranean scallop Pecten jacobaeus inferred from mitochondrial 16s DNA sequence analysis. Animal Biodiversity and Conservation, 41.1: 61–73. Abstract Genetic structure and demographic history of the endemic Mediterranean scallop Pecten jacobaeus inferred from mitochondrial 16s DNA sequence analysis. Understanding the genetic population structure of species going through population decline is primordial in implementing a management plan. In the case of Pecten jacobaeus, previous genetic studies have been limited to populations in the western Mediterranean (Spain) and the Adriatic Sea (Italy). To check the presence of phylogeographic breaks between the two Mediterranean basins, we scored the variability of the mitochondrial 16S rRNA gene in two populations from the eastern basin (Tunisia and Greece) and pooled them with those cited above. The two newly analyzed populations shared the most frequent haplotypes with the other populations and showed no evidence of phylogeographic breaks. We found lower levels of genetic variability in the Adriatic and the Aegean populations, but not in Tunisia, with respect to the Western Mediterranean. Significant differences in pooled haplotype frequencies indicated some genetic differentiation between the pooled Chioggia and Vouliagmeni populations and the other pooled populations. Key words: Scallop, Sequence, Genetic diversity, Population structure Resumen Estructura genética e historia demográfica de la vieira endémica del Mediterráneo Pecten jacobaeus inferidas a partir del análisis de la secuencia del ADN mitocondrial que codifica la subunidad 16S del ARNr. Con vistas a implementar un plan de gestión para las especies cuyas poblaciones están menguando, es fundamental comprender la estructura genética de dichas poblaciones. En el caso de Pecten jacobaeus, los estudios genéticos previos se han limitado a analizar poblaciones situadas en el Mediterráneo occidental (España) y en el mar Adriático (Italia). Para comprobar la presencia de discontinuidades filogeográficas entre las dos cuencas del Mediterráneo, hemos estudiado la variabilidad del gen mitocondrial del ARNr 16S en dos poblaciones de la cuenca oriental (Túnez y Grecia) y la hemos analizado junto con la de las mencionadas anteriormente. Las dos poblaciones estudiadas recientemente compartieron los haplotipos más frecuentes con las otras y no se encontraron indicios de que exista una discontinuidad filogeográfica. Se observó un grado menor de variabilidad genética en relación con el Mediterráneo occidental en las poblaciones del Adriático y el Egeo, pero no en Túnez. Las diferencias significativas observadas cuando se agruparon los datos sobre las frecuencias haplotípicas indicaron la existencia de una cierta diferenciación genética entre las poblaciones de Chioggia (Italia) y Vouliagmeni (Grecia) y las de las otras poblaciones. Palabras clave: Vieira, Secuencia, Diversidad genetica, Estructura de la población Received: 23 II 17; Conditional acceptance: 28 IV 17; Final acceptance: 13 VI 17 Khaoula Telahigue, M’hamed El Cafsi, Physiology and Aquatic Environment, Dept. of Biology, Fac. of Sciences of Tunis, Univ. of Tunis, El Manar, 2092 Tunis, Tunisia.– Tarek Hajji, Higher Inst. of Biotechnology–Sidi Thabet, Univ. Manouba, BVBGR–LR11ES31 Biotechpole, Sidi Thabet, 2020 Ariana, Tunisia.– Carlos Saavedra, Inst. de Acuicultura de Torre la Sal–CSIC, Ribera de Cabanes, 12595 Castellon, Spain. Corresponding author: Khaoula Telahigue. E–mail: k_telahigue@yahoo.fr ISSN: 1578–665 X eISSN: 2014–928 X

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Introduction Scallops (family Pectinidae) are filter–feeding bivalve molluscs that live mainly on sandy gravel or gravel seabed and colonize all waters from the northern to the southern hemispheres (Gosling, 2003). Due to its high nutritional values, this group is highly prized as a food source (Caers et al., 1999; Palacios et al., 2005; Telahigue et al., 2010). Currently, more than 40 commercial species of scallop are exploited worldwide. The global production from fishing and aquaculture combined reached 2.5 million tonnes (FAO, 2012). Among pectinids, the scallops of the genus Pecten are the most appreciated in Europe. Two Pecten taxa can be found on European coasts. The king scallop Pecten maximus (Linneaeus, 1758) occurs in the Atlantic, reaching the coasts of northern Africa. It enters the western end of the Mediterranean Sea, but not further than the Almeria–Oran Oceanographic front, which appears to be a barrier to dispersal of this and many other species in the region (Wilding et al., 1999; Rıo ́ s et al., 2002; Saavedra and Peña, 2004, 2005; Morvezen et al., 2016). The king scallop is replaced in the remaining part of the Mediterranean Sea by the great Mediterranean scallop Pecten jacobaeus (Linneaeus, 1758) (Rombouts, 1991). The two taxa can be easily distinguished by the shell morphology. However, genetic studies carried out with different types of genetic markers, such as allozymes, mitochondrial DNA restriction fragment–length polymorphism, mitochondrial sequences and microsatellites, suggest that the two scallops could be races or subspecies (Wilding et al., 1999; Rıo ́ s et al., 2002; Saavedra and Peña, 2004, 2005; Morvezen et al., 2016). For clarity we will use P. jacobaeus to refer to the Mediterranean populations throughout the paper. The great Mediterranean scallop P. jacobaeus occurs in exploitable quantities only in the northern Adriatic Sea. Fishing activities using dredges and trawls came into use in the early 1980s, and the total annual landing has since declined. After 30 years of exploiting the natural beds of this bivalve (particularly in the Adriatic Sea), stocks are now severely depleted (Pranovi et al., 2001; Katsanevakis, 2005). To manage and conserve the species, a better understanding of the genetic variability and genetic population structure is essential (Moritz, 1994; Mahidol et al., 2007). Several studies have shown that Mediterranean marine species often exhibit important genetic differentiation between western and eastern Mediterranean populations. Such differentiation includes phylogeographic breaks, attributed to the isolation of stocks due to sea level changes during the Pleistocene glaciations (Patarnello et al., 2007; Borrero–Pérez et al., 2011; Bowen et al., 2016). Quite often, genetically differentiated populations are restricted to specific regions such as the Adriatic, the Aegean, or the Gulf of Gabès (Rabaoui et al., 2011; Cordero et al., 2014). In the case of P. jacobaeus, the data available about the genetic population structure of the species are very limited. Previous studies in this species showed relatively little genetic differentiation between populations and an absence of phylogeographic subdivision

(Wilding et al., 1999; Rıo ́ s et al., 2002; Saavedra and Peña, 2004, 2005; Morvezen et al., 2016). However, most studies were based on only a few populations from the west Mediterranean Sea. Only two genetic studies in this species have considered samples from the eastern Mediterranean Sea, and these were limited to a single population from the northern Adriatic (Rıo ́ s et al., 2002; Saavedra and Peña, 2005). This population showed less genetic variability than other P. jacobaeus populations from the western Mediterranean, but it was not possible to state if this was a local effect or a feature shared by all eastern Mediterranean populations. Moreover, no statistically significant genetic differentiation between all populations was found. But since a large part of the distribution area of the species in the Eastern Mediterranean was not sampled, a conclusion of no genetic subdivision in the species would be premature. Clearly, a genetic study based on a wider sampling of P. jacobaeus in the eastern Mediterranean is desirable to clarify the genetic structure of this taxon. Here we present a study of the genetic structure of P. jacobaeus over a broader geographical scale than previous studies. We sampled two further populations in the eastern Mediterranean, one in the Aegean Sea (Greece) and the other in the Gulf of Tunis (Tunisia). These new samples expand the geographic area covered to the south and the east. We analyzed these populations together with those studied previously by Saavedra and Peña (2005), thus covering the majority of the range of P. jacobaeus. We used partial sequences of the same mitochondrial 16S ribosomal RNA gene studied by the previous authors. This genetic marker is a powerful tool for measuring genetic variation and gene flow among populations (Duran et al., 2009). Material and methods Sample collection A total of 35 P. jacobaeus adults of different sizes were sampled from the eastern basin (fig. 1); 22 individuals were dredged from the open sea near Kelibia in the northeastern coast of Tunisia, and 13 specimens were caught in the marine Lake Vouliagmeni, a lagoon located in the Korinthiakos Gulf, on the Aegean coast of Greece. Tissue samples (adductor muscle) from each individual were soaked in 90 % ethanol and sent to the IATS–CSIC laboratory (Castellón, Spain) where the molecular analyses procedures were carried out. Samples with ID numbers 'KEL 1 to 22' and 'VOU 1 to 13' are kept at the IATS–CSIC laboratory and are available upon request to the corresponding author. DNA extraction, amplification and sequencing Total DNA was extracted by a salt extraction protocol (Miller et al., 1988). We used primer Pec16S–F1 (5'GTTTTAAGGTCGGGGAAAG–3') (Saavedra and Peña, 2005) designed from the complete P. maximus 16S rRNA sequence deposited in GenBank (accession no. X92688) and reverse primer 16Sbr of Palumbi

Animal Biodiversity and Conservation 41.1 (2018)

France

CHI Italy

Spain

CAS VAL

Western Mediterranean

Algeria

Greece

KEL

Tunisia

Morocco N O

E S

Adriatic Sea

200 km

Ionian Sea

VOU

Aegean Sea

Siculo–Tunisian Strait Eastern Mediterranean

Libya

Egypt

H13

H14

H15

H16

H17

H18

H19

H20

H21

H10

H22

H11

H23

H12

H24

Fig. 1. Geographic locationn of the two populations of Pecten jacobaeus sampled for this study (stars), and of the three previously populations sequenced by Saavedra and Peña (2005) that were included in the analysis (dots): VAL, Valencia; CAS, Castellon; KEL, Kelibia; CHI, Chioggia; VOU, Vouliagmeni. Pie charts show the frequencies of the 16S rRNA haplotypes detected. Fig. 1. Localización geográfica de las dos poblaciones de Pecten jacobaeus muestreadas para este estudio (estrellas), y de las tres poblaciones secuenciadas previamente por Saavedra y Peña (2005) que se han incluido en el análisis (puntos): VAL, Valencia; CAS, Castellón; KEL, Kelibia; CHI, Chioggia; VOU, Vouliagmeni. Los gráficos circulares muestran las frecuencias de los haplotipos detectados en el gen del ARNr 16S.

(1996) (5'CCGGTCTGAACTCAGATCACGT–3') to amplify a fragment of 512 base pairs (bp) of the mitochondrial 16S rRNA gene. Amplifications were done in 20 µL reaction volume containing 1 µL template DNA (~ 500 ng), 0.2 mM of each dNTP, 0.8 µM of each primer, 1.5 mM of MgCl2, and 0.15 U Taq polymerase (GIBCO–Life Technologies) in the buffer supplied by the manufacturer. PCR was performed using the following parameters: after an initial denaturation at 95 °C for 4 min, the PCR mix was subjected to 35 cycles consisting of a 1 min step at 95 °C, a 30 s step at 55 °C followed by a 30 s step at 72 °C, and a final extension of 3 min at 72 °C. PCR products were purified with the QiaQuick PCR kit (QIAGEN), and sequenced in an ABI 377 automatic sequencer by using the Big Dye Terminator chemistry in the DNA Sequencing Service of the University of Valencia (Spain). Sequences were edited in BioEdit 7.2.5 (Hall, 1999) and aligned with Clustal W (Thompson et al., 1994), as implemented in BioEdit. Genetic analyses Our data were analyzed together with sequences from three other populations, two from the western Mediterranean (CAS and VAL) and one from the Adriatic Sea (CHI) (Saavedra and Peña 2005). Genetic diversity

was estimated for each population with the software DnaSP 5.10.01 (Librado and Rozas, 2009) using several indices, such as the number of haplotypes (h), haplotype diversity (H), number of segregating sites (SS), and nucleotide diversity (θπ). A median–joining (MJ) network was constructed to visualize genetic relationships between mtDNA haplotypes with Network 5.0 software (Bandelt et al., 1999). Differences in haplotype frequencies between populations were tested using the x²–test (Preacher, 2001). For population structure analysis, we used the Arlequin version 3.5.2.2 program (Fu, 1997) to estimate pairwise FST–values (Ramos–Onsins and Rozas, 2002) for each population pair using haplotype frequencies. Significance of all pairwise values was measured using 10,000 permutations and assessed using correction for multiple tests under Arlequin. We also applied the same software to examine the distribution of genetic variability into hierarchical levels through Analysis of Molecular Variance (AMOVA) within and between P. jacobaeus populations. Nei’s genetic distances (Nei, 1972, 1979) were computed from haplotype frequencies and used for constructing a tree showing the similarities between population samples by the Neighbor Joining Method (Saitou and Nei, 1987). To infer the historical demography, DnaSP was used to conduct neutrality tests of Tajima’s D (Tajima,

Telahigue et al.

1989), Fu's FS (Fu, 1997) and Ramos–Onsins and Rozas'R2 (Ramos–Onsins and Rozas, 2002) to detect whether there was any deviation from the assumption of neutrality, which would indicate a recent population expansion. Tajima's D is widely used in neutrality tests. It is known that 'D' is based on the comparison of two different estimates of genetic diversity (Ɵ and π). 'D' is expected to be equal to 0 under the neutral equilibrium model. A significant negative value indicates an excess of rare variants as expected under positive selection. It can also reflect a demographic event such as population expansion. A significant positive 'D' value, in contrast, indicates an excess of intermediate frequency variants reflecting balancing selection or diversifying selection. It can also reflect population structure or bottleneck events (Oleksyk et al., 2010; Camus–Kulandaivelu et al., 2014) Fu’s FS and the Ramos–Onsins and Rozas R2–test were confirmed to be the most powerful tools in examining population growth (Ramos–Onsins and Rozas, 2002). The latter was proven to be particularly sensitive for limited sample sizes (Ramos–Onsins and Rozas, 2002). P–values for neutrality statistics were obtained by coalescent simulations with 10,000 replicates. A goodness of fit test was performed to test the validity of the sudden expansion model using a parametric bootstrap approach based on the sum of square deviations (SSD) between the observed and expected mismatch distributions. The raggedness index, which measures the smoothness of the mismatch distribution, was calculated for each distribution. Small raggedness values represent a population that has experienced sudden expansion whereas higher values of the raggedness index suggest a stationary population or a population that has experienced a bottleneck. The demographic expansion parameters tau (τ), Ɵ0 and Ɵt based on the mismatch distribution outputs from Arlequin were estimated under a demographic expansion hypothesis by a generalized non–linear least–square approach (Schneider and Excoffier, 1999). The time since the expansion (t) was estimated using the Li's formula (Li, 1977): t = τ/2u, where u is the mutation rate per sequence per generation (Schneider and Excoffier, 1999; Rogers, 1995). We adopted the mutation rate of 1.22 x 10–6 proposed by Saavedra and Peña (2005) for the same species and same gene fragment, which is based on fossil calibration. Finally, mismatch distributions of P. jacobaeus populations were performed using DnaSP to test whether demographic processes were consistent with the mismatch distribution test statistics. The distribution is usually multimodal in samples drawn from populations at demographic equilibrium, whereas populations that have gone through a recent demographic expansion are expected to be unimodal (Harpending, 1994). Results Sequence variation and haplotype diversity Sequences of a 512 base pair fragment of the 16S rRNA gene were obtained for 27 individuals from VOU and KEL. The obtained sequences were pooled

with 48 sequences from the populations VAL, CAS and CHI from the study of Saavedra and Peña (2005). Polymorphism was detected at 25 sites: 24 substitutions and two indels (insertion/deletion). Among the substitutions, we identified four transversions and 21 transitions, two sites showed multiple substitutions that allowed detection in twenty–four haplotypes (table 1). Table 2 shows haplotype frequencies. Sixteen haplotypes were previously found by Saavedra and Peña (2005), and eight are new sequences found in KEL and VOU. The sequences of the new haplotypes have been deposited in GenBank under accessions MF183948–MF183955. Among the scored haplotypes, only two (H1 and H4) were common to all studied populations, together representing 60 % of the total number of obtained sequences. Among these common haplotypes, haplotype H1 was the most frequent. The remaining haplotypes (71 % of total number of haplotypes) were present in a single population (private), with the exception of haplotypes H3, H6 and H10, which appeared in four (VAL, KEL, CHI, VOU), three (VAL, CAS, CHI) and two (CAS, KEL) populations, respectively. Figure 1 shows the geographical distribution of all recorded haplotypes and their respective frequencies per site . The median joining haplotype network (fig. 2) showed two central haplotypes (H1 and H4) linked by four low frequency haplotypes and 17 other haplotypes differing from the central ones by 1–4 mutations. No obvious pattern of haplotype distribution across geographical locations is apparent. Table 3 shows estimates of gene diversity. Overall, all populations displayed high values of haplotype diversity and low values of nucleotide diversity, as is typical in this species (Saavedra and Peña, 2005). The two populations from the eastern Mediterranean (CHI and VOU) presented the lowest values of haplotype diversity (0.717–0.691) and nucleotide diversity (0.00223–0.00164), while the highest values were observed in KEL (0.925 and 0.0044). Population genetic differentiation We quantified the inter–population genetic divergence in haplotype frequencies using the pairwise population differentiation statistic (FST). The estimated FST–values varied between 0.006 and 0.117, and were not significant (even with correction for multiple tests), except for the comparison between VOU and VAL (table 4). Since low sample sizes and high haplotype numbers could interfere with statistical detection of genetic differentiation, we performed 2 x 2 contingency x2–test after pooling the less frequent haplotypes all together, and pooling populations according to their location on the Western or on the Eastern Mediterranean. We carried out two tests that differed in pooling of the Tunisian population of KEL alternatively with the western or with the eastern Mediterranean populations. The rationale for this is that this population is geographically located in the Siculo–Tunisian Strait, just in the area of separation of the two Mediterranean basins. In studies of other species the populations from this region turned out to

Animal Biodiversity and Conservation 41.1 (2018)

Table 1. Variant sites of 16S rRNA haplotypes in Pecten jacobaeus scallop: H, haplotype Tabla 1. Sitios variables de los haplotipos de ARNr 16S en la vieira Pecten jacobaeus: H, haplotipo.

2 2 2 2 2 2 2 2 2 2 2 2 2 3 3 3 3 3 3 3 3 3

2 2 2 5 5 6 6 6 7 7 7 7 7 9 9 9 0 0 0 0 0 1 3 6 9

6 7 9 8 9 2 6 7 0 1 3 6 8 1 6 7 0 3 4 5 6 3 3 4 0

C – A G A T A A C T – G G A C A A A C T T T C C A

– – – – – – – – – – – – – – – – – G – – – – – – –

– – – – – – – – – – – T – – – – – – – – – – – – –

– – – – – – – – – – – – A – – – – – – – – – – – –

T – – – G – – – – – – – – – – – – – – – – – – – –

T – – – – – – – – – – – – – – – – – – – – – – – –

– – – – – – – – – – – – – – T – – – – C – – – – –

– – G – – – – – – – – – A – – – – – – – – – – – G

– – – – – – – – – C – – – – – – – G – – – – – – –

H10

– – – T – – – – – – – – – – – – – – – – – – – – –

H11

– – – – – – – – – – – A A – – – – – – – – – – – –

H12

– – – – – – – – – – – – A – – – – – – – – – T – –

H13

– – – – – – G – – – – – – G – – – – – – – – – – –

H14

– – – – – – – – – – – – A – – – – G – – – – – – –

H15

– C – – – – – G – – – – A – – – G – – – – – – – –

H16

– – – – – – – – – – – – – – – – – – – – – – T – –

H17

– – – – – – – – – – – T – – – – – G – – – – – – –

H18

– – G – – – – – – – – – A – – – – – – – – – – – –

H19

– – – – – – – – – – – – A – – G – – – – – C – – –

H20

– – – – – – – – – – – – – – – – – – – – – – – T –

H21

– – – – – – – – – C – T – – – – – – – – – – – – –

H22

– – – – – C – – – – – – A – – – – – T – – – – – –

H23

T – – – – – – – – – – – – – – – – – G – – – – – –

H24

– – – – – – – – T – – – A – – – – – – – A – – – –

be genetically more similar to eastern Mediterranean or to western Mediterranean populations (Patarnello et al., 2007), but in our case data are not conclusive and therefore the two possibilities should be considered. When KEL was pooled with the eastern Mediterranean populations (CHI and VOU) a non–significant result was obtained (x2 = 1.5, df = 2, P = 0.22). In the test in which KEL was pooled with the two western Mediterranean populations (CAS and VAL) we obtained a significant x2 of 4.6 (P = 0.032, df = 2). Moreover, we performed a test between western and eastern populations, excluding KEL. This test was significant (x2 = 8.11, df = 2, P = 0.004). To conclude our analysis of genetic differentiation we performed a hierarchical analysis of molecular variance (AMOVA) taking nucleotide variability into account in addition to

haplotype frequencies. This analysis did not find any significant genetic divergence either among (P = 0.10) or within (P = 0.84) the western (VAL, CAS and KEL) and eastern (CHI and VOU) Mediterranean basins. Genetic distances between pairs of populations varied between 0.139 and 0.767. Figure 3 shows the NJ tree obtained from the distance matrix . It reflects the close similarity of CHI and VOU, and the higher divergence of VAL and KEL. Demographic history of the populations Results from neutrality tests showed negative values for Tajima’s D and Fu's Fs in all studied populations (table 5). Fu's test resulted in highly significant negative Fs values only for KEL (–6.8; P < 0.001)

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Table 2. The haplotype frequencies of 16S rRNA gene in five Pecten jacobaeus populations: h, haplotype; VAL, Valencia; CAS, Castellon; KEL, Kelibia; CHI, Chioggia; VOU, Vouliagmeni. Tabla 2. Frecuencias haplotípicas del gen del ARNr 16S en cinco poblaciones de Pecten jacobaeus: h, haplotipo; VAL, Valencia; CAS, Castellón; KEL, Kelibia; CHI, Chioggia; VOU, Vouliagmeni.

Haplotype frequency VAL

CAS

KEL

CHI

VOU Total

H1 1 8

4 8 6 27

H2 – –

– – 2 2

H3 1 –

1 1 1 4

H4 3 6

3 4 2 18

H5 – –

– 1 – 1

H6 1 1

– 1 – 3

H7 – –

– 1 – 1

H8 – –

1 – – 1

H9 – –

1 – – 1

H10 –

1 – –

H11 – –

1 – –

H12 –

–

1 – –

H13 –

–

1 – –

H14 –

–

1 – –

H15 –

–

1 – –

H16 –

– – –

H17 –

– – –

H18 –

– – –

H19 –

– – –

H20 –

– – –

H21 –

– – –

H22 –

– – –

H23 1

–

– – –

H24 1

–

– – –

and CAS (–5.9; P < 0.001) populations. However, Tajima’s test showed no significant negative D values for all populations except for the KEL sample (1.8; P < 0.001). The overall negative values resulting from both tests pointed towards an excess of low frequency polymorphisms in relative to expectation. Since the western Mediterranean populations were genetically homogeneous, we carried out the same tests pooling the three samples from this region. The results show a clear deviation from neutrality (D and Fs) and demographic instability (R2). Potential explanations include population size expansion, positive selection

(Tajima, 1989), purifying selection, and a recruitment of rare alleles from the western or eastern basins. Figure 4 shows the mismatch frequency spectra for the two populations (KEL, VOU). The studied populations showed a positive skewed unimodal distribution and supported the hypothesis of the sudden expansion model. None of the sums of squared deviations (SSD) of mismatch distribution (table 6) was significant, indicating that the curves fit the sudden expansion model tested. The significant fit between the observed and the expected distributions was also confirmed by the low and not significant raggedness index values for all studied populations. We noted that the VOU population was distinguished from the other populations with a higher but non–significant Rg value (0.23702). These results were further confirmed by the Ramos–Onsins and Rozas test R2. This test gave significant results, rejecting the null hypothesis of constant size and supporting a recent demographic expansion. The tau value (τ), which reflects the location of the mismatch distribution crest, provided a rough estimation of the time when rapid population expansion started. The observed values of the age expansion parameter (τ) were very close between KEL and VAL (2.250 and 2.695 respectively) suggesting a similar timing of demographic events. In comparison, CHI and VOU showed the lowest τ (0.71). Estimates of ϴ0 and ϴ1 indicated that all studied populations expanded from a very small (close to 0 in almost cases) to a very large size (table 6). The results for the pooled western Mediterranean samples are in the same line. Assuming the mutation rate of 1.22 x 10–6 for 16S gene and the equation τ = 2ut, the time of expansion for P. jacobaeus populations likely occurred approximately between 1.10 myr and 0.29 myr before present for VAL and CHI respectively. Discussion By sampling two populations of P. jacobaeus from Tunisia and Greece, we have increased the number of populations available for the genetic study in the eastern Mediterranean from one to three, and we have also extended the sampled area by ~1,000 km. We studied a total of five populations, two in the west Mediterranean and three in the eastern Mediterranean. This allowed a comparison between the two basins with acceptable rigor. The first interesting result of our study is that the 16S haplotype network did not show new clades, in contrast with the network described by Saavedra and Peña (2005), after adding the new populations. This confirms the absence of a phylogeographic break between the two basins that was suggested in the previous study by Saavedra and Peña (2005); however, we note that this does not completely eliminate the possibility of a phylogeographic subdivision in the Mediterranean scallop. For example, in the carpet–shell clam, Cordero et al. (2014) found two mitochondrial clades, but the less frequent cladewas restricted to the northern Aegean

Animal Biodiversity and Conservation 41.1 (2018)

H21

H13

Kelibia Vouliagmeni Castellon Chioggia Valencia

H20

H10

H6 H5 H12

H23

H17

H16

H15

H1 H11

H4 H18

H14 H19 H24

H9 H22

Fig. 2. Median joining haplotype network of the 24 haplotypes determined in Pecten jacobaeus specimens at the 16S mitochondrial gene. The circle areas are proportional to haplotype frequency and the lines connecting haplotypes represent one step mutation. The locations of the haplotypes are indicated in color according to the codes in the legend. Fig. 2. Red de unión de medianas de los 24 haplotipos determinados en los especímenes de Pecten jacobaeus en el gen mitocondrial del ARNr 16S. La superficie de los círculos es proporcional a la frecuencia de los haplotipos y las líneas que conectan los haplotipos representan una mutación. La localización de los haplotipos se indica con un código de colores (véase la leyenda).

and Turkey. Therefore, sampling of the easternmost coasts of the Mediterranean, especially the Aegean Sea, would be necessary to be conclusive in the case of the scallop. Saavedra and Peña (2005) also found no significant differentiation in haplotype frequencies between the three P. jacobaeus populations that they studied. We found essentially the same result after adding two eastern Mediterranean populations, as indicated by the non–significant pairwise FST estimates. However, there was an exception in the comparison of VOU with respect to VAL, which gave a significant FST value of 0.11. This is a relatively high value for a marine species with a planktonic larval stage that lasts for several weeks. Genetic divergence in this case could be the result of the long geographic distance (3,000 km) that separates these two localities, but the lack of significant differentiation between CAS (which is very close to VAL) and VOU casts

doubt on this explanation, and other explanations should be sought. The VOU sample was taken from Lake Vouliagmeni, which originated ca. 2,000 years ago (Papapetrou–Zamanis, 1969). The lake was originally brackish, as indicated by the fossil fauna recovered, but a channel was open some 100 years ago to connect the lake with the sea and the fauna of the lake was replaced by typically marine species (Vardala–Theodorou and Nicolaidu, 2007). It is at that time when the origin of the VOU scallop population can be established. Since the area of the lake is small (1.8 km long), it is most likely that the population size is small. Consequently, genetic drift should be an important factor acting on the genetic pool of this population and could increase its genetic differentiation over levels typical of open sea populations. On the other hand, the VOU and VAL samples were the smallest in our study, and therefore another possibility is that these samples were

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Table 3. Sample size (N), number of haplotypes (h), haplotype diversity (H), segregating sites (SS) and nucleotide diversity based on the pairwise difference between sequences (π) of 16S mitochondrial gene for five Pecten jacobaeus populations: VAL, Valencia; CAS, Castellon; KEL, Kelibia; CHI, Chioggia; VOU, Vouliagmeni. Tabla 3. Tamaños de muestra (N), número de haplotipos (h), diversidad haplotípica (H), sitios segregantes (SS) y diversidad nucleotídica basada en las diferencias entre pares de secuencias (π) del gen del ARNr 16S en cinco poblaciones de Pecten jacobaeus: VAL, Valencia; CAS, Castellón; KEL, Kelibia; CHI, Chioggia; VOU, Vouliagmeni. Population

H ± SD

π ± SD

VAL

0.893 ± 0.111

0.00394 ± 0.00088

CAS

0.841 ± 0.056

0.00336 ± 0.00057

KEL

0.925 ± 0.050

0.00444 ± 0.00080

CHI

0.717 ± 0.099

0.00223 ± 0.00056

VOU

0.691 ± 0.128

0.00164 ± 0.00042

Total

0.814 ± 0.035

0.00310 ± 0.00033

Table 4. Pairwise genetic differentiation statistics (FST) (below diagonal) and their associated probabilities (P) for the null hypothesis of FST = 0 (above diagonal) among the studied populations of Pecten jacobaeus, based on 16S sequences. The bold P–value remained significant after multitest correction: VAL, Valencia; CAS, Castellon; KEL, Kelibia; CHI, Chioggia; VOU, Vouliagmeni; * significant at p = 0.05. Tabla 4. Parámetros estadísticos de diferenciación genética (FST) (debajo de la diagonal) y sus probabilidades asociadas (P) para la hipótesis nula de FST = 0 (encima de la diagonal), entre parejas de las poblaciones estudiadas de Pecten jacobaeus, basados en secuencias parciales del gen del ARNr 16S. El valor de P en negrita se mantuvo significativo tras la corrección multitest: VAL, Valencia; CAS, Castellón; KEL, Kelibia; CHI, Chioggia; VOU, Vouliagmeni; * valores significativos para p = 0,05. VAL

VAL –

CAS

KEL

CHI

VOU

0.47273 0.58182 0.19091 0.02727

CAS 0.01195

–

0.94545 0.47273 0.60909

KEL

0.01193 0.01998

–

CHI

0.04851 0.00766 0.02316

0.17273 0.23636 –

VOU 0.1174* 0.02933 0.00989 0.00617

affected by larger sampling variances of haplotype frequencies, which could have resulted in biased, significant FST estimates (Kitada et al., 2007). Alternatively, we could be experiencing type II errors in the pair–wise comparisons due to the lack of power to detect small genetic differences among populations. This is suggested by the results of the chi–squared tests with pooled samples and haplotypes. Pooling samples can help to overcome the power problem due to small sample sizes. In this case, the results of the x2–tests after pooling samples gave statistical

0.40909 –

support to a west–east differentiation when the KEL population was considered genetically similar to the western Mediterranean, and when KEL was excluded from the tests. Therefore, these tests suggest that the western and eastern basins could exhibit a lower connectivity than the populations located within each of the two basins. A number of studies have shown a decrease in gene flow across the Siculo–Tunisian Strait, which has been recognized as a barrier to gene flow between eastern and western groups for several benthic species with long pelagic larval

Animal Biodiversity and Conservation 41.1 (2018)

CHI 100 VOU

100 CAS KEL

0.05 VAL

Fig. 3. Unrooted neighbor joining tree showing the genetic differences between populations, based on Nei's genetic distances computed from haplotype frequencies. Significant bootstrap values are also indicated. Fig. 3. Árbol de unión de vecinos sin raíz que muestra las diferencias genéticas entre poblaciones, basado en las distancias genéticas de Nei calculadas a partir de las frecuencias haplotípicas. Se indican también los valores de bootstrap significativos.

stages (Nikula and Vainola, 2003; Zitari–Chatti et al., 2009; Tir et al., 2014). However, the NJ tree based on Nei distances shows that the genetic differentiation increases on approaching the Strait of Gibraltar.

This leads to the hypothesis that the influence of the Atlantic populations contributes to the observed patterns. The Atlantic influence can result from gene flow facilitated by the current that enters the western Mediterranean from the Atlantic and flows eastward along the northern coasts for Africa, past the Almeria–Oran front (Millot and Taupier–Letage, 2005). The second interesting result of our study regards the levels of intra–population genetic variability. Saavedra and Peña (2005) described lower levels of genetic variability in the Adriatic population (CHI) than in the western Mediterranean populations (VAL and CAS). We found that the Aegean population (VOU) shares this feature with CHI, but that the Tunisian population (KEL) does not, and that it currently shows the highest levels of variability. The high level of genetic variability observed in KEL contrasts with the other two eastern Mediterranean populations, which show only half or a quarter of the nucleotide variability found in the Tunisian population. A number of factors could contribute to this result. One factor could be a gene diversity input from the Atlantic and western Mediterranean favored by the North African current, which flows from the Atlantic to the eastern Mediterranean along the North African coast (Patarnello et al., 2007). This possibility is supported by the x2–test of genetic differentiation with pooled populations and haplotypes, which suggest a higher similarity of KEL to the western Mediterranean populations. In other species, the populations of the northern Tunisian coasts often show similarities to the western Mediterranean populations (Bahri–Sfar et al., 2000; Cordero et al., 2014). Furthermore, the apparent west–east cline for the two most common haplotypes, the highest amount of private haplotypes and genetic variability recorded for the western populations may suggest that the western populations

Table 5. Statistical tests of neutrality and estimates of demographic parameters in populations of Pecten jacobaeus, based on partial sequences of the 16S rRNA gene: VAL, Valencia; CAS, Castellon; KEL, Kelibia; CHI, Chioggia; VOU, Vouliagmeni; WM, West Mediterranean (pooled VAL, CAS, KEL) (* P < 0.05 ** P < 0.001). Tabla 5. Pruebas estadísticas de neutralidad y estimaciones de los parámetros demográficos en poblaciones de Pecten jacobaeus, basados en secuencias parciales del gen del ARNr 16S: VAL, Valencia; CAS, Castellon; KEL, Kelibia; CHI, Chioggia; VOU, Vouliagmeni; WM, Mediterráneo occidental (agrupación VAL, CAS, KEL) (* P < 0,05 ** P < 0,001).

Mismatch

SSD (P) Tajima's D

Rg (P)

–0.63262

–2.050 (0.057)

0.03444 (0.940)

0.177**

VAL

2.500

0.033 (0.35)

CAS

1.834

0.002 (0.65)

–1.76498

–5.866 (0.001)**

0.04513 (0.280)

0.123**

KEL

2.258

0.001 (0.97)

–1.80726**

–6.830 (0.001)**

0.05201 (0.230)

0.141**

CHI

1.367

0.006 (0.53)

–1.27752

–1.742 (0.089)

0.04611 (1.000)

0.137**

VOU

0.836

0.035 (0.21)

–0.62785

–1.116 (0.079)

0.23702 (0.440)

0.155**

1.151

0.000 (0.70)

–1.92370*

–16.942 (0.00)**

0.03651 (0.550)

0.106**

Telahigue et al.

KEL

VOU 0.3

Frequency

0.5 0.4

0.2

0.3 Exp. 0.1 Obs.

0.2

Exp. Obs.

0.1 0

0 5 10 15 20 0 Pairwise differences

5 10 15 Pairwise differences

Fig. 4. Observed and expected pairwise mismatch distributions in the two Pecten jacobaeus populations (VOU and KEL) under the sudden population expansion model. The number of pairwise nucleotide differences between haplotypes is represented on the abscissa whereas their frequencies are represented on the ordinate axis: KEL, Kelibia; VOU. Vouliagmeni. Fig. 4. Distribuciones de desajuste observadas y esperadas en las dos poblaciones de Pecten jacobaeus muestreadas (VOU y KEL), utilizando el modelo de expansión súbita. El número de diferencias nucleotídicas entre pares de haplotipos se representa en el eje de abscisas mientras que sus frecuencias se representan en el eje de ordenadas: KEL, Kelibia; VOU, Vouliagmeni.

are ancestral and are the source of the eastern populations, with KEL representing an intermediate transition sample between the twoh basins. Another factor could be that the Adriatic and Aegean populations have been affected by population–specific events that resulted in lower effective population size (Ne) and subsequently in the loss of intrapopulation diversity. The smaller Ne in the Adriatic could be caused by a restriction of available habitat during the glacial epoch, present–day geographic isolation, or historical demographic instability. The Adriatic Sea disappeared almost completely during the last glacial maximum (ca. 18.00 years BP) (Lambeck and Purcell, 2005), and the scallop populations now living in the northern Adriatic certainly derive from a recolonization process which might have resulted in lower genetic variability. Finally, fishing could affect variability levels through its effect on the effective population size. The Adriatic scallop population is the most exploited in the Mediterranean Sea. Several authors have reported large fluctuations of populations and landings of this scallop over the past 50 years (Orel et al., 1993; Mattei and Pellizzato, 1996). Overexploitation activities could cause bottleneck effects and alter genetic diversity within populations, as recorded in several commercially exploited scallops (Saavedra and Peña, 2004; Gaffney et al., 2010; Bert et al., 2011). Low variability in the VOU population could have resulted from a founder effect and a bottleneck occur-

rind at the moment the population became established a century ago, and could have subsequently reduced Ne. Several studies have recorded low intrapopulation diversity in bivalves that have experienced severe genetic drift due to bottlenecking and/or founder events (Jordaens et al., 2000; Tarnowska et al., 2010; Barbieri et al., 2015). According to Hanzawa et al. (2012), the founder events when the marine lakes were formed must have affected the genetic diversity of the present populationy. The higher value of R2 and lower t showed by VOU in the demographic analysis support this view. In conclusion, our study has clearly determined the absence of a phylogeographic subdivision between the western and the eastern Mediterranean populations of P. jacobaeus, and has shown that the populations from the northern coasts of the eastern Mediterranean have lower genetic variability levels due to local historical events that wouldd have affected their Ne. The possibility of a phylogeographic subdivision in the easternmost populations (Aegean Sea, Levantine Sea) remains to be explored. The use of 16S rRNA or other mitochondrial marker could be a nice tool for this. However, the increasing availability of nuclear markers in this species will be of great help for this purpose. In their study on the carpet–shell clam, Cordero et al. (2014) observed a high homogeneity across the Atlantic, the western Mediterranean and the eastern Mediterranean at the mitochondrial marker. However, the study of

Animal Biodiversity and Conservation 41.1 (2018)

References Table 6. Parameters for Pecten jacobaeus populations obtained from mismatch distribution analyses: age of expansion in units of mutational time (τ), population size before (ϴ0) and after (ϴ1) the expansion in units of mutational time and time since expansion expressed in million years t: VAL, Valencia; CAS, Castellon; KEL, Kelibia; CHI, Chioggia; VOU, Vouliagmeni; WM, West Mediterranean (pooled VAL, CAS, KEL). Tabla 6. Parámetros de los análisis de la distribución del desajuste de secuencias entre poblaciones de Pecten jacobaeus: edad de la expansión en unidades de tiempo mutacional (τ), tamaño de la población antes (ϴ0) y después (ϴ1) de la expansión, en unidades de tiempo mutacional, y tiempo desde la expansión expresados en millones de años t: VAL, Valencia; CAS, Castellón; KEL, Kelibia; CHI, Chioggia; VOU, Vouliagmeni; WM, Mediterráneo occidental (agrupación VAL, CAS, KEL). Populations

ϴ0

ϴ1

VOU

1.000 0.000 3407 0.41

CAS

1.875 0.000 3444 0.77

KEL

2.250 0.000 3469 0.92

VAL

2.695 0.000 27 1.10

CHI

0.7111 0.072 7400 0.29

1.956 0.083 65042 0.78

six nuclear intronic markers revealed that three of these showed clear differentiation between the three basins. In the case of scallops, the microsatellite data reported by Morvezen et al. (2016) suggest that there is greater genetic differentiation for nuclear markers than for mitochondrial DNA between the Atlantic and the western Mediterranean populations, so it is possible that the same could be happening to the genetic differentiation between the western and the eastern basins, or even among other smaller regions within the Mediterranean Sea. Finally, the use of nuclear markers and more extensive sampling will be necessary to determine whether there is a reduced genetic connectivity between scallop populations living in the eastern and western Mediterranean basins. Acknowledgements We are indebted to Juan Peña (retired) for his invaluable help during the preparation of this study, and to S. Katsanevanis for providing the samples from Vouliagmeni Lake. We also thank an anonymous reviewer for kind suggestions to improve the work.

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Animal Biodiversity and Conservation 41.1 (2018)

Applying IUCN Red List criteria to birds at different geographical scales: similarities and differences M. Charra, M. Sarasa

Charra, M., Sarasa, M., 2018. Applying IUCN Red List criteria to birds at different geographical scales: similarities and differences. Animal Biodiversity and Conservation, 41.1: 75–95. Abstract Applying IUCN Red List criteria to birds at different geographical scales: similarities and differences. Extinction risk and conservation status of species are assessed at the global scale by the International Union for Conservation of Nature (IUCN). To ensure objectivity, repeatability and traceability, assessments follow a standardized process that uses reliable and verifiable information. Assessments are synthesized according to guidelines, which have recently been adjusted for application at sub–global scales. Nevertheless, species may have several, different or overlapping conservation status. To quantitatively compare assessments from global to sub–national scales, in this study we analyzed 15 assessment lists for 66 game bird species in France. Assessments were made following IUCN guidelines. Overall, our results reveal that (1) assessments at large spatial scales tend to give lower threat status than small–scale assessments; (2) large–scale assessments made it possible to formally verify information whereas smaller–scale assessments usually did not; (3) large–scale assessments are more likely to be based on standardized evidence of reduction in population size and are less exposed to 'scale–effects' and 'edge–effects'; (4) large–scale assessments are also more often based on scientific literature sensu stricto; and (5) sources are more accurately synthesized than Red Lists at small spatial scales. Our results suggest that small–scale Red Lists do not fully match IUCN guidelines and differ significantly in their assessment processes when compared to global standards. The use of subjective and unreliable data in small–scale Red Lists (above all in national and sub–national lists) may jeopardise the original aim of IUCN Red Lists to provide comprehensive and scientifically rigorous information, and could thus compromise the credibility and prestige of IUCN Red Lists in the eyes of researchers, the general public, and other stakeholders. Key words: Biodiversity assessment, Game bird species, Conservation status, Information–based management, IUCN Red Lists, Regional assessment Resumen Aplicación de los criterios de la Lista Roja de la Unión Internacional para la Conservación de la Naturaleza en diferentes escalas geográficas a las aves. La Unión Internacional para la Conservación de la Naturaleza (UICN) se encarga de evaluar a escala mundial el riesgo de extinción y el estado de conservación de las especies. Para garantizar su objetividad, repetibilidad y trazabilidad, en las evaluaciones se sigue un proceso estandarizado que hace uso de información fiable y verificable. Asimismo, las evaluaciones se sintetizan de acuerdo con determinadas directrices, que se han ajustado recientemente para su aplicación a escalas inferiores. No obstante, la misma especie puede clasificarse en varios estados de conservación, distintos o superpuestos. Para comparar cuantitativamente las evaluaciones de escala mundial a escala subnacional, analizamos 15 listas de evaluación relativas a 66 especies de aves cinegéticas en Francia; según se había declarado, dichas evaluaciones se realizaron en consonancia con las directrices de la UICN. En general, nuestros resultados ponen de manifiesto que (1) las evaluaciones a gran escala espacial tienden a dar como resultado estados de peligro inferiores que las de pequeña escala; (2) las evaluaciones a gran escala permitieron comprobar de forma oficial la información en que se basan, mientras que las evaluaciones a menor escala no; (3) las evaluaciones a gran escala son más propensas a basarse en pruebas estandarizadas de una reducción significativa del tamaño de la población y a estar menos expuestas a efectos de 'escala' o de 'borde de distribución’; (4) ISSN: 1578–665 X eISSN: 2014–928 X

Charra & Sarasa

las evaluaciones a gran escala también se basan más frecuentemente en publicaciones científicas en sentido estricto; y (5) las fuentes se sintetizan con mayor exactitud en comparación con las Listas Rojas a escalas espaciales pequeñas. Por lo tanto, nuestros resultados sugieren que las Listas Rojas a pequeña escala no coinciden plenamente con las directrices de la UICN y que difieren de forma significativa con respecto a sus procesos de evaluación en comparación con los estándares mundiales. El uso de información subjetiva y poco fiable en las Listas Rojas en pequeña escala (sobre todo en las listas nacionales y subnacionales) puede poner en peligro el objetivo original de las Listas Rojas de la UICN de proporcionar información completa y científicamente rigurosa y, por lo tanto, podría comprometer la credibilidad y el prestigio de las Listas Rojas de la UICN a los ojos de los investigadores, del público en general y de otras partes interesadas. Palabras clave: Evaluación de la biodiversidad, Especies de aves cinegéticas, Estado de conservación, Gestión basada en información, Listas Rojas de la IUCN, Evaluación regional Received: 8 X 16; Conditional acceptance: 9 VI 17; Final acceptance: 24 VII 17 Former address: Margaux Charra, Mathieu Sarasa, Fédération Nationale des Chasseurs, 13 rue du Général Leclerc, 92136 Issy les Moulineaux Cedex, France. Present address: M. Charra, Lab. des Sciences du Climat et de l’Environnement (LSCE), CEA–CNRS–UVSQ, L'Orme des Merisiers CEA Saclay, bat 701, 91191 Gif–sur–Yvette, France.– M. Sarasa, BEOPS, 1 Esplanade Compans Caffarelli, 31000 Toulouse, France. Corresponding author: Mathieu Sarasa. E–mail: msarasa@beops.fr; contact@beops.fr

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Introduction Evidence–based wildlife management (Sutherland et al., 2004) requires reliable information on, above all, the conservation status and the extinction risk of species. The most widely recognized assessment of the conservation status of species is the Red List of Threatened Species, established by the International Union for Conservation of Nature (IUCN) (de Grammont and Cuarón, 2006; Rodrigues et al., 2006; Szabo et al., 2012; Maes et al., 2015). The great value of the IUCN Red Lists is derived from their original aim to represent a comprehensive source of scientifically rigorous information (Rodrigues et al., 2006). IUCN assessments have to be objective, transparent, repeatable and traceable (Fitzpatrick et al., 2007; Miller et al., 2007). To this aim, Red Lists are derived from assessments that use data published in a searchable format (Rodrigues et al., 2006). Nevertheless, assessments also routinely use expert knowledge (McBride et al., 2012), so that guidelines for the reliable integration of such knowledge are under development (McCarthy et al., 2004; McBride et al., 2012; Drescher et al., 2013; Drolet et al., 2015). Publication of both data and assessments is now consolidated at global scale through to the on–line searchable IUCN databases accessible via the Internet at http://www.iucnredlist.org. Assessments are constructed using explicitly defined categories and quantitative criteria that are applicable and valid at global scales (Akçakaya et al., 2000; IUCN, 2001). Over the past two decades, these criteria and categories have been revised, thresholds have been adjusted and new categories created (Gärdenfors, 2001; IUCN, 2013). The robustness of assessments has been consolidated through the standardization of data–driven procedures and the use of objective criteria that no longer depend on approaches that entail risks of subjectivity (e.g. threat categorizations based directly on expert opinions) (Mace and Lande, 1991; Rodrigues et al., 2006). Nevertheless, some of the issues that still need to be resolved have been underlined by, for instance, the application of IUCN criteria at sub–global scales (Gärdenfors, 2001; Mace et al., 2008). Hereafter we use 'sub–global’ as a synonym of ‘regional’ sensu lato to avoid potential confusions with political districts (regions sensu stricto), that in several countries such as France correspond to sub–national administrative territories. Given that the majority of conservation actions take place at sub–global scales, and that the most influential institutions working on conservation legislation and action are national and regional governments, the concern for the spatial sub–structuring of the threat status of species is increasing (Gärdenfors, 2001; Gärdenfors et al., 2001; Miller et al., 2007). The regional concept (sensu lato) implies a geographically–defined sub–global area, which could be a continent, a country, a state or a province (IUCN, 2012). Guidelines for the application of IUCN Red List criteria at regional levels were published (Gärdenfors, 2001; IUCN, 2012) and assessment of the conservation status of species at sub–global scales were developed

(Miller et al., 2007; Azam et al., 2016). These updated guidelines represent the standardized processes that must be applied (without deviation or modification) if regional Red List authorities wish to state that their assessments follow the IUCN system (IUCN, 2012: p. 3). Nevertheless, a risk of subjectivity in the regional adjustment process has been identified, along with the need for more complementary information for identifying national priorities and responsibilities with regards to species' conservation (Keller and Bollmann, 2004; Rodríguez et al., 2004; Keller et al., 2005). Despite efforts to create objective processes for assessing species' extinction risks at sub–global scales, some problems still persist (Gärdenfors, 2001; Martín, 2009; Seoane et al., 2011). Natural scarcity or rarity at local scales may result in the overestimation of threat levels and so the ecological bases of their rarity should be taken much more into account (Martín, 2009; Seoane et al., 2011). Moreover, the second step of the IUCN regional guidelines, which consists of adapting categories according to vaguely formulated terms such as the level of contact with neighbouring populations (Keller et al., 2005: p. 1828), leaves room for interpretation by assessors and a degree of subjectivity (Eaton et al., 2005; Keller et al., 2005). This step is particularly important when assessing very mobile species, such as birds at small landlocked sites (Keller and Bollmann, 2004; Keller et al., 2005) since it requires a lot of accurate data and knowledge on species that is not always available and can be difficult to obtain (Keller and Bollmann, 2004; Eaton et al., 2005; Keller et al., 2005). Indeed, available data for regional assessments are sometimes limited to local populations, and obtaining data for cross–boundary populations is often difficult (Keller et al., 2005). Conversely, in some cases available data may be accurate for large–scale assessments but unsuitable or unreliable for local scales due to limits imposed by their resolution (Hurlbert and Jetz, 2007) and may be negatively affected by confounding methodological factors such as missing data or low number of counts per site (Atkinson et al., 2006). In addition to these yet unresolved issues, recent concerns about threatened species have led to an increase of regional and global Red Lists, sometimes reflecting different conservation statuses of the same species at different scales. France is a particularly interesting example (Azam et al., 2016), as the conservation status of its birds has been characterized at the global level by IUCN and Birdlife International (IUCN, 2015b), at European level by Birdlife International (BirdLife International, 2015), at the national level by the French Committee for IUCN (UICN France et al., 2011), and at the sub–national level (French regions) by local organizations (Flitti and Vincent–Martin, 2013; LPO Alsace, 2014). As a result, each bird species may be classified under five different conservation statuses in France (detailed below). Thus, in light of the increasing number of Red Lists referring to the same species, the simple question 'what is the conservation status of the considered bird species?' has become far more complex for all concerned. The choice to use one or other of these different classifica-

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tions according to the particular situation may rest on the understanding of their respective characteristics and limits. A quantitative comparative analysis of their characteristics is thus required, but to our knowledge is not yet available. In this study, we used a set of game bird species as a case study to compare several Red Lists (from sub–national to global scale) that classify the conservation status of birds in France. Previous studies in other countries have compared global and regional Red Lists or analyzed the regional assessment of certain taxa. Their results highlight the fact that regional lists tend to lead up to higher threat statuses than the global IUCN Red Lists due, for instance, to the scale–dependent chance of meeting Red List criteria (hereafter referred to as 'scale effect') and to the 'edge effect' of small–scale assessments of the conservation status of species (Keller et al., 2005; Milner–Gulland et al., 2006; Brito et al., 2010). Theoretically, different assessments should agree if they use (for instance, for endemic species) a common methodology, identical species and the same information. Nevertheless, in light of the results from other countries (Milner–Gulland et al., 2006), we expect that in this study the assessments of the status of birds (including numerous non–endemic species) at smaller scales would result in higher threat statuses compared to larger scale assessments. As mentioned above, the variability in conservation status between lists at different scales may be due to 'scale–' or 'edge–effects', in particular in regional assessments, when criterion D of the IUCN is used, which considers small population size (Eaton et al., 2005; Keller et al., 2005). To evaluate whether the variability in conservation statuses between lists at different scales in France is due to the 'population size effect', we compared the proportions of criteria used in assessments. On the basis of the associations reported in previous studies (Eaton et al., 2005; Keller et al., 2005), we expected that a higher proportion of species would be classified as threatened based on criterion D at small–scale assessments compared to large–scale lists. The differences in status due to the scale of the assessment could also be associated with disparities in the type of information used to evaluate species’ conservation status (de Grammont and Cuarón, 2006). One hypothesis suggests that some lists and evaluations may depend primarily on data from grey literature, which conflicts with the comprehensive, scientifically rigorous and transparent nature of Red Lists (Mrosovsky and Godfrey, 2008). To evaluate whether variability in conservation status between different lists is linked to the differences in the type of information that were synthethized, the proportions of categories of information used in French Red Lists were compared. On the basis of the reported greater likelihood of subjectivity in small–scale assessments (Eaton et al., 2005; Keller et al., 2005), we predicted that grey literature would play a more important role in small–scale Red Lists than in large–scale assessments. The distinction between scientific and grey literature has been widely debated (Schöpfel, 2006; Mrosovsky and Godfrey, 2008) and is detailed below

in the methods section. Among other characteristics, grey literature is usually less available, less reliable and more incomplete than scientific literature (Conn et al., 2003; Schöpfel, 2006; Mrosovsky and Godfrey, 2008). Moreover, literature unavailability has been proposed as a factor that might be associated with unreliable citations (Todd and Ladle, 2008). Thus, according to this hypothesis, we predicted that grey literature might be more associated with cases of no supported citations than scientific literature. Material and methods Study species Highly mobile species may be more affected by the challenges posed by regional adaptations to the IUCN system, particularly when applied at small geographical scales and when data on across–boundary population dynamics are required (Akçakaya et al., 2000; Keller and Bollmann, 2004; Keller et al., 2005; IUCN, 2012). To ensure comparability, analyses should be based on overlapping sets of species; as well, data sets should contain well–studied species in order to make quantitative analyses possible. Many birds are widely distributed mobile species that represent a well–studied taxonomic group (van Jaarsveld et al., 1998; Butchart et al., 2004; Fazey et al., 2005). However, it is a vast taxonomic group and so we chose to analyse a subset for potential heterogeneity between Red Lists. Among birds, game species may be the object of multiple and additive conservation actions, such as monitoring programs conducted by wildlife recreationists with a variety of different motivations (Cooper et al., 2015) and hence these species may be better studied than others. Thus, we focused our analyses on 66 game bird species in France (table 1). IUCN–type Red Lists France is a biogeographically diverse country crossed by many migratory flyways and there are several Red Lists for birds to assess their conservation status. To ensure comparability, in this study we only analyzed lists based on the IUCN classification system, as these lists are expected to follow the standardized processes detailed in the IUCN guidelines (IUCN, 2001, 2012). The global IUCN Red List was updated during the development of this study at the end of 2015. This allowed us to verify the potential effects of this update on the potential scale–dependent heterogeneity in Red Lists. As a result, we compared 15 lists in this study (table 2): two versions (a pre–update version from October 2015 and an updated version from December 2015) of the IUCN global Red List of Birds (IUCN, 2015a, 2015b); the European Red Lists of Birds (BirdLife International, 2015), available at regional levels for (i) geographical Europe (Europe) and (ii) the member states of the European Union in 2012 (EU27); the national (French) Red List of Birds (UICN France et al., 2011); and 10 sub–national

Animal Biodiversity and Conservation 41.1 (2018)

Red Lists of Birds for Île–de–France (IDF) (Birard et al., 2012), Limousin (Roger and Lagarde, 2015), Pays de la Loire (PaysLoire) (Marchadour et al., 2014), Midi–Pyrénées (MidiPyr) (Fremaux, 2015), Provence–Alpes–Côte–d’Azur (PACA) (Flitti and Vincent–Martin, 2013), Languedoc–Roussillon (LangRou) (Meridionalis, 2015), Centre (Nature Centre, 2013), Alsace (LPO Alsace, 2014), Bretagne (Bretagne Environnement, 2015) and Bourgogne (Bourg) (Abel et al., 2015). All these assessments declared to have followed the IUCN system; the considered sub–national lists were approved and labelled by the UICN French committee, a process designed to guarantee that sub–national Red Lists follow IUCN guidelines (UICN France, 2011; http://uicn.fr/ etat–des–lieux–listes–rouges–regionales/). Compiled data Conservation status and criteria For all the species on each Red List, we compiled the conservation status (LC, least concern; NT, near threatened; VU, vulnerable; EN, endangered; CR, critically endangered; RE, regionally extinct) and, whenever possible, the criteria underlying it (A, population reduction; B, geographic range; C, small population size and decline; D, very small or restricted regional population) (IUCN, 2001, 2012). Criterion E (based on quantitative analyses that estimate the probability of extinction; IUCN, 2012) was not present in our sample. In a few cases (6.56 %), the species were qualified as Threatened based on multiple criteria. In such cases, we used the first used criterion according to the classification E > A > B > C > D because small regional populations or local rarities do not necessarily imply a high risk of extinction (Harnik et al., 2012) while criterion A deals with species that are at risk because of a steep rate of decline (Collen et al., 2016). Bibliographical categories The references cited in the Red Lists were compiled and classified in four categories on the basis of the following definitions. The first category (A) is 'scientific literature' sensu stricto, that is, work published in scientific journals that is indexed in scientific data sources (Björk et al., 2010) and peer reviewed (Steven et al., 2011). Scientific literature meets methodological standards (Conn et al., 2003), is easily available (Pyšek et al., 2008) and represents, notwithstanding certain flaws, a widely accepted strategy for ensuring quality control in scientific research (Ferreira et al., 2015). The second category (B) includes 'referenced books', in particular, books identified by an International Standard Book Number (ISBN) and referenced academic publications, such as PhD theses. These data sources may be accessible yet allow some freedom from the peer–review process, and thus they summarize science from a personal perspective to present ideas in a liberating manner (McWilliams and Bauchinger, 2012). The third category (C) is 'grey literature' that includes publications that are not peer– reviewed (Conn et al., 2003) and articles that appear in non–indexed journals. Such articles are difficult to

identify and to access through classical routes and often lack robust methodology and traceability (Corlett, 2011; Friess and Webb, 2011). Finally, the fourth category (D), 'expert opinion', includes estimates based on empirical knowledge or even field experience that is not to be found even in grey literature. Citation categories We compiled and classified the way in which data, citations and sources were included in the IUCN– based assessments First, we analysed whether it was possible to link the detailed information in assessments to citations and sources. Next, the reliability of the cited information was classified through consensus between the two authors (MC and MS) into one of four categories as defined in Todd et al. (2010): (1) There is 'Clear support', when the cited article provides unequivocal support of the assertion via either statements in the text or the data presented. (2) 'Ambiguous', when the material (either text or data) in the cited article has been interpreted one way, but could also be interpreted in other ways, including the opposite. The assertion in the primary article is supported by a portion of the cited article, but that portion runs contrary to the overall thrust of the cited article. The assertion includes two or more components, but the cited article only supports one of them. (3) 'No support', when the cited article does not in any way substantiate the assertion via either statements in the text or the data presented. The cited article may even contradict the assertion in the primary article. (4) 'Empty citation', when the cited article simply cites other articles that support the assertion made in the primary article. Citing a review article is acceptable if the support for the assertion is, for example, a new insight or opinion offered by the author(s) of the review. As in Todd et al. (2010), if the cited article was classified as 'empty citation' plus 'no support', ‘no support’ took precedence. If the cited article was classified as 'empty citation' plus 'ambiguous', 'ambiguous' took precedence. Another citation category ('unverifiable') was created for expert opinions and for cases in which the lack of published or available documents make the assessment of the links to the information source impossible. Statistical analysis We used Fisher's tests (Millot, 2011) to test if the proportion of threatened categories was significantly different between Red Lists for the overlapping sets of species, taking into account the small sample size in some tests and the need for standardized analyses for comparisons. Additionally, the odds ratio (ranging from 0 to infinity) in bilateral tests on contingency tables was used to analyze the direction of detected differences (Millot, 2011). The further away the odds ratio was from 1 towards infinity, the more the first list in the test was characterized by the considered factor. The more the second list in the test was characterized by the considered factor, the closer the odds ratio was to 0.

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Table 1. Bird species and conservation status according to IUCN–based Red Lists at different scales. Global (IUCNo, IUCN version from October 2015; IUCNd, IUCN version from December 2015). European (Eur, geographical Europe; EU27, European politico–economic Union). National (UICN–F, French Committee of the International Union for Conservation of Nature). Sub–national (SN: IDF, Île–de–France; Ce, Centre; Al, Alsace; PL, Pays de Loire; Br, Bretagne; Bo, Bourgogne; LaR, Languedoc–Roussillon; MiP, Midi–Pyrénées; PACA, Provence–Alpes–Côte–d'Azur; Li, Limousin. Conservation status (LC, least concern; NT, near threatened; VU, vulnerable; EN, endangered; CR, critically endangered; RE, regionally extinct). DD, data deficient; NA, not applicable; NE, not evaluated.

UICN–F

PACA

MiP

LaR

IDF

LC LC LC LC LC LC LC LC LC LC LC LC LC LC LC

EU 27

Corvus corone

Eur

Scientific name

IUCNo

IUCNd

Tabla 1. Especies de aves y estado de conservación según las Listas Rojas basadas en el método UICN a distintas escalas: Global (IUCNo, versión IUCN de octubre de 2015; IUCNd, versión IUCN de diciembre de 2015). Europea (Eur: Lista Roja europea para Europa geográfica; EU27, Lista Roja para la unión político– económica europea. Nacional (UICN–F, Lista Roja nacional de Francia). Subnacional (SN: IDF, Isla de Francia; Ce, Centro; Al, Alsacia; PL, País de Loira; Br, Bretaña; Bo, Borgoña; LaR, Languedoc–Rosellón; MiP, Mediodía–Pirineos; PACA, Provenza–Alpes–Costa Azul; Li, Lemosín). Estatus de conservación: LC, preocupación menor; NT, casi amenazada; VU, vulnerable; EN, en peligro; CR, en peligro crítico; RE, extinto a nivel regional. DD, datos insuficientes; NA, no aplicable; NE, no evaluada.

Garrulus glandarius LC LC LC LC LC LC LC LC LC LC LC LC LC LC LC Pica pica

LC LC LC LC LC LC LC LC LC LC LC LC LC LC LC

Corvus frugilegus

LC LC LC LC LC LC LC LC LC LC LC LC VU NT LC

Sturnus vulgaris

LC LC LC LC LC LC LC LC LC LC LC LC LC LC LC

Anas platyrhynchos

LC LC LC LC LC LC LC LC LC LC LC DD LC LC

Anas strepera

LC LC LC LC LC NA EN CR NT CR EN NT CR VU CR

Anas clypeata

LC LC LC LC LC CR EN NA LC EN CR DD

Anas acuta

LC LC LC VU LC

Anas penelope

LC LC LC VU LC

Anas crecca

LC LC LC LC VU CR EN CR CR CR CR NA NA CR

Anas querquedula

LC LC LC VU VU CR CR NA VU CR CR DD

Aythya ferina

LC VU VU VU LC EN NT CR LC CR VU EN

NA CR

Aythya marila

LC LC VU VU NT

Aythya fuligula

LC LC LC LC NT

EN NA

Bucephala clangula

LC LC LC LC NA NA

CR EN

NA VU

NT VU VU NT CR VU

EN NA

NA NA

Netta rufina

LC LC LC LC LC VU VU

Melanitta fusca

EN VU VU VU EN

Melanitta nigra

LC LC LC LC LC

Somateria mollissima LC NT VU EN CR

CR CR

NA NA

Clangula hyemalis

VU VU VU VU NA

Anser anser

LC LC LC LC VU

Anser fabalis

LC LC LC LC VU

Anser albifrons

LC LC LC LC NA

NA EN

VU NT

NA CR

VU NA

EN NA

Gallinago gallinago LC LC LC LC EN RE CR RE CR RE CR CR

Lymnocryptes minimus LC LC LC LC NA

Vanellus vanellus

LC NT VU VU LC VU VU EN LC VU EN EN CR EN EN

Pluvialis apricaria

LC LC LC LC LC

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MiP

PACA

LaR

EN VU

IDF

UICNâ&#x20AC;&#x201C;F

LC LC LC LC LC

EU 27

Pluvialis squatarola

Haematopus ostralegus LC NT VU VU LC

Eur

Scientific name

IUCNo

IUCNd

Table 1. (Cont.)

EN NA

Numenius arquata NT NT VU VU VU NA EN CR EN EN VU CR CR CR Numenius phaeopus LC LC LC LC VU

Limosa limosa

NT NT VU EN VU

Limosa lapponica

LC NT LC LC LC

Tringa totanus

LC LC LC VU LC

Tringa nebularia

LC LC LC LC LC

Tringa erythropus

RE NA VU RE

RE LC EN

EN NA

LC LC LC NT NA

Philomachus pugnax LC LC LC EN NT

Calidris canutus

LC NT LC LC NT

Fulica atra

LC LC NT LC LC

LC LC LC LC LC LC VU LC EN

Gallinula chloropus LC LC LC LC LC LC LC LC LC LC LC LC LC NT Rallus aquaticus

LC LC LC LC DD VU VU VU DD EN DD LC EN LC EN

Scolopax rusticola LC LC LC LC LC NT NT LC NT LC VU DD NT DD DD Columba palumbus LC LC LC LC LC LC LC LC LC LC LC LC LC LC Columba oenas

LC LC LC LC LC LC LC LC LC LC DD VU VU VU VU

Columba livia

LC LC LC LC EN

NE LC LC DD

DD RE RE NA

Streptopelia turtur LC VU VU NT LC NT LC NT NT LC VU LC LC VU Streptopelia decaocto LC LC LC LC LC LC LC LC LC LC LC LC LC LC Turdus viscivorus

LC LC LC LC LC LC LC LC LC LC LC LC LC LC

Turdus philomelos LC LC LC LC LC LC LC LC LC LC LC LC LC LC Turdus pilaris

LC LC LC VU LC NA NA VU

DD EN VU CR LC LC

Turdus iliacus

LC NT NT VU LC

Turdus merula

LC LC LC LC LC LC LC LC LC LC LC LC LC LC

Coturnix coturnix

LC LC LC LC LC NT LC NT LC LC DD NT VU NT

Alauda arvensis

LC LC LC LC LC LC NT NT NT LC NT LC LC LC

Tetrao urogallus

LC LC LC LC VU

EN VU

Tetrao tetrix

LC LC LC LC LC

Lagopus muta

LC LC NT VU LC

VU NT VU

Tetrastes bonasia

LC LC LC LC VU

Alectoris graeca

NT NT NT VU NT

Perdix perdix

LC LC LC LC LC LC NT EN NE DD DD CR RE NA DD

Alectoris rufa

LC LC LC LC LC DD LC NA NE DD DD DD

Phasianus colchicus LC LC LC LC LC Syrmaticus reevesii

VU VU

LC NE LC NE DD LC NA

NA NA NA

Callipepla californica LC LC

NT NT

Colinus virginianus

VU RE

VU DD LC DD NA NA

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Table 2. Characteristics of the 15 compared lists of birds: Y, year; Sc, scale (G, global; E, European; N, national; Sn, sub–national); Gs, game species; Ca, IUCN categories; Cr, detailed IUCN criteria; Sr, sources; As, assessment–sources link; T, total number of sources useful for assessment; Pr, proportion of examined sources; 1 proportion of references obtained among all sources useful for assessment; 2 proportion of references obtained among cited and referenced references used in the status assessment. (For other abbreviations see table 1). Tabla 2. Características de las 15 listas de aves comparadas: Y, año; Sc, escala (G, mundial; E, europea; N, nacional; Sn, subnacional); Gs, especies de caza; Ca, categorias UICN; Cr, criterios UICN detallados; Sr, fuentes; As, enlace entre dato de evaluación y fuente; 1 proporción de referencias obtenidas entre todas las fuentes útiles para la evaluación; 2 proporción de referencias obtenidas entre referencias citadas y referenciadas utilizadas en la evaluación de estatus. (Para otras abreviaturas véase la tabla 1). IUCN–based Red Lists of Birds

IUCN (version from October 2015)

2015

IUCN (version from December 2015)

2015

Europe (geographical Europe, EU27) 2015

P 1,182

UICN French Committee

2011

P – – 0

Bourgogne

2015 Sn 35 P

P – 97

Limousin

2015 Sn 59 P

P – 5

IDF

2012 Sn 34 P

P – 23

Alsace

2014 Sn 40 P

P – – 0

Centre

2013 Sn 38 P

P – – 0

Pays de Loire

2013

P – – 0

Languedoc–Roussillon

2015 Sn 40 P

P – – 0

Midi–Pyrénées

2015 Sn 17 P – P – 3

PACA

2013 Sn 40 P – P – 6

Bretagne

2015 Sn 51 P – – – 0

Sc Gs Ca Cr

Using Fisher's tests and based on the odds ratio we also examined (i) the proportion of adduced criteria for threatened categories, (ii) the proportion of bibliographical categories and (iii) the proportion of citation categories in assessments among the different scales of Red Lists. Multiple estimation of significance values can increase type I errors (i.e., rejecting the null hypothesis H0 when H0 is true). Thus, in tables 3–6 we also present p–values corrected using a Benjamini–Hochberg procedure (BH; Benjamini and Hochberg, 1995) to control for potentially false discovery rates (FDR), the expected proportion of 'discoveries' (rejected null hypothesis H0) that might be false (incorrect rejection). Nevertheless, this type of correction incurs reduction in power. Using this kind of procedure for the more detailed studies would have implied a lower probability of finding significant results, increasing the risk of type II errors, sometimes to an unacceptable level (not rejecting H0 when H0 is false) (Nakagawa, 2004). Ecological results suggested by BH p–values were generally similar to those indicated by the p–value of Fisher's tests. Thus, in the text we only detail and discuss the results from Fisher’s tests.

P 133 36.8 %1/72.1 %2 8.50 %

Results Traceability Although all 66 species from our sample were included in the global and national Red Lists, only 17–63 of these species were included in the European and sub–national lists (table 1). The classification criteria were clearly presented in almost all studied lists, the exceptions being the lists for Midi–Pyrénées, Provence–Alpes–Côte–d’Azur and Bretagne. Sources and clear links between data in assessments and sources were simultaneously available for global and European Red Lists but not for national and sub–national lists. Nevertheless, sources (but not links) were given for the Bourgogne, Limousin, Île–de–France, Midi–Pyrénées and Provence–Alpes–Côte–d’Azur sub–national lists (table 2). Accessibility and language constrained the literature review. The analysis of the types of information sources was based on 76.1 % of sources (303 identifiable sources out of 398) for the global Red List, 91.2 % of sources (1078/1182) for the European Red List and 100 % of sources for sub–national lists (3–97 sources, depending on the

Animal Biodiversity and Conservation 41.1 (2018)

100 %

n 28

90 % 80 %

70 % 60 %

50 %

40 %

30 % 20 %

10 %

RE Al

PACA

LaR

MiP

IDF

UICN–F

EU 27

Eur

IUCNd

IUCNo

0 %

Fig. 1. Proportions of Red Lists status of birds. (For abbreviations see table 1). Fig. 1. Proporciones de las categorías de conservación en las Listas Rojas de aves. (Para las abreviaturas véase la tabla 1).

list). The links between information and sources were analyzed on the basis of (1) the traceable and accessible sources that determined the conservation status for the global Red List and (2) the relevant sources for France that were traceable and accessible in the European Red List. In the current global Red List, 48.9 % (65/133) of citations were untraceable or not reported in the bibliography section. Of the 68 traceable sources (51.1 % of sources) we managed to obtain 49 (72.1 %), thereby allowing us to analyze 51 % (177/347) of all links between information and source used in the evaluation of the conservation status of the species from our sample. For the European Red List, the sources cited for France represent 11.2 % (132) of the whole bibliography for the study species. In this sample, 3.8 % (5/132) of citations were untraceable. Of the 127 sources of identifiable literature (96.2 % of sources), 111 were accessible and obtained (87.4 %), thereby allowing us to analyze 83.2 % (559/672) of the links between information and source cited for France. These inequalities in information availability determined which comparisons between lists and scales could be performed. Conservation status The proportion of conservation statuses attributed varied between the Red Lists at different spatial scales (fig. 1, table 3). Overall, Red Lists at larger scales

were associated with more 'least concern' statuses and fewer 'threatened' statuses. Nevertheless, European and national lists did not differ significantly and proportions of statuses between lists at the same spatial scale did not show any significant differences when examined directly. Nonetheless, the comparisons of the two global lists, before and after the update at the end of 2015, with the sub–global lists did not give identical results. Overall, the pre–update global Red List exhibited fewer threatened statuses than the European and national lists; however, these differences disappeared when we compared them with the post–update global Red List. In addition, the pre–update global Red List exhibited few more differences than the post–update global Red List in comparison with the sub–national Red Lists. Red List criteria The proportion of the Red List criteria that was applied varied between the lists at different spatial scales (fig. 2, table 4). Criterion 'A' (reduction in population size) was used significantly more in the assessment of species at larger scales (global and European lists) than in national and sub–national lists. On the other hand, criterion 'D' (small regional population) was used more on sub–national lists than in national, European and global Red Lists.

Charra & Sarasa

Table 3. Results of the comparison of the proportions of different IUCN statuses on Red Lists of birds. N, number of species; p, p–values of Fisher’s test; BHp, p–values corrected using the Benjamini–Hochberg procedure; OR, odds ratio; BL, European Red List from Birdlife International, SN.SN, all comparisons of sub–national Red Lists. (For other abbreviations see taable 1). Tabla 3. Resultados de las comparaciones de proporciones de los estados de conservación de la UICN entre las Listas Rojas de aves: N, número de especies; p, valor p de la prueba de Fisher; BHp, valor p corregido según el procedimiento de Benjamini–Hochberg; OR, razón de momios; BL, Lista Roja europea de Birdlife International; SN.SN, todas las comparaciones entre Listas Rojas subnacionales. (Para las otras abreviaturas véase tabla 1).

LC status

NT status

BHp

p BHp OR

IUCNo–IUCNd 66 –

–

VU status

EN status

BHp

p BHp OR

– – – – – – –

CR status p BHp OR

– – – – – –

IUCNo–BL.Eur 63 0.004 – 3.284 – – – 0.009 0.018 0.087 – – – – – – IUCNo–BL.EU27 63 0.001 0.004 5.326 – – – 0.000 0.000 0.053 – – – – – – IUCNd–BL.Eur 63 –

–

– – – – – – –

IUCNd–BL.EU27 63 –

–

– – – – 0.011 0.015 0.219 – – – – – –

– – – – – –

IUCNo–UICN 66 0.002 0.004 6.023 – – – 0.003 0.006 0.000 – – – – – – IUCNd–UICN

66 –

–

IUCNo–Al

34 0.000 0.000 23.557 – – – – – –

IUCNo–Bo

28 0.000 0.000 25.502 – – – 0.023 – 0.000 – – – – – –

IUCNo–Br

41 0.000 0.000 20.588 – – – – – –

– – – 0.025 – 0.000

IUCNo–Ce

32 0.000 0.000 23.557 – – – – – –

– – – – – –

IUCNo–IDF

28 0.000 0.000 Inf – – – – – –

– – – – – –

IUCNo–PL

33 0.002 0.004 9.739 – – – – – –

– – – – – –

IUCNo–PACA

34 0.000 0.000 28.693 – – – 0.002 0.040 0.000 – – – – –

IUCNo–Li

31 0.000 0.000 31.453 –

IUCNo–MiP

17 0.004 0.007 15.311 – – – – – –

IUCNo–LaR

32 0.000 0.000 13.812 –

IUCNd–Al

34 0.011 0.015 5.401 – – – – – –

– – – 0.024 – 0.000

IUCNd–Bo

28 0.009 0.013 5.799 – – – – – –

– – – – – –

IUCNd–Br

41 – – – 0.025 – Inf – – – – – – 0.025 – 0.000

IUCNd–Ce

32 0.011 0.014 5.401 – – – – – –

– – – – – –

IUCNd–IDF

28 0.027 0.030 5.434 – – – – – –

– – – – – –

IUCNd–PL

33 –

– – – – – –

IUCNd–PACA

34 0.003 0.006 6.616 – – – 0.013 – 0.086 – – – – – –

IUCNd–Li

31 0.005 0.008 5.581 –

IUCNd–MiP

17 0.004 0.007 15.311 – – – – – –

IUCNd–LaR

32 0.026 0.031 4.060 –

–

– – – – – – –

–

– –

– – – – – – – –

–

– –

– – – – – – – – – 0.024 – 0.000

0.021 – 0.000 0.021 – 0.000 – – – – – – 0.020 – 0.000 –

–

0.021 – 0.000 0.021 – 0.000 – – – – – – 0.020 – 0.000 –

–

BL.Eur–BL.EU27 63 –

–

– – – – – – –

– – – – – –

BL.Eur–UICN 63 –

–

– – – – – – –

– – – – – –

BL.EU27–UICN 63 –

–

– – – – – – –

– – – – – –

BL.Eur–Al

34 0.026 – 4.179 – – – – – –

– – – 0.024 – 0.000

BL.Eur–Bo

28 0.023 – 4.469 – – – – – –

– – – – – –

BL.Eur–Br

41 –

– – – 0.025 – 0.000

BL.Eur–Ce

32 0.026 – 4.179 – – – – – –

– – – – – –

BL.Eur–IDF

28 0.027 – 5.434 – – – – – –

– – – – – –

–

– – – – – – –

Animal Biodiversity and Conservation 41.1 (2018)

Table 3. (Cont.)

LC status N BL.Eur–PL

33 –

BHp –

NT status

p BHp OR

VU status p

BHp

EN status p BHp OR

CR status p BHp OR

– – – – – – –

– – – – – –

BL.Eur–PACA 34 0.018 – 4.139 – – – – – –

– – – – – –

BL.Eur–Li

31 0.026 – 4.116 – – – – – –

– – – – – –

BL.Eur–MiP

17 –

–

– – – – – – –

– – – – – –

BL.Eur–LaR

32 –

–

– – – – – – –

– – – – – –

BL.EU27–Al

34 0.026 – 4.179 – – – – – –

– – – 0.024 – 0.000

BL.EU27–Bo

28 0.050 – 3.578 – – – – – –

– – – – – –

BL.EU27–Br

41 –

–

– – – – – – –

– – – 0.025 – 0.000

BL.EU27–Ce 32 0.026 – 4.179 – – – – – –

– – – – – –

BL.EU27–IDF 28 –

–

– – – – – – –

– – – – – –

BL.EU27–PL

–

– – – – – – –

– – – – – –

BL.EU27–PACA 34 0.038 – 3.427 – – – – – –

– – – – – –

BL.EU27–Li

33 – 31 –

–

– – – – – – –

– – – – – –

BL.EU27–MiP 17 –

–

– – – – – – –

– – – – – –

BL.EU27–LaR

32 –

–

– – – – – – –

– – – – – –

UICN–Al

34 –

–

– – – – – – –

UICN–Bo

28 0.023 – 4.469 – – – – – –

– – – – – –

UICN–Br

41 –

– – – – – – –

– – – – – –

UICN–Ce

32 0.047 – 3.736 – – – – – –

– – – – – –

UICN–IDF

28 –

–

– – – – – – –

– – – – – –

UICN–PL

33 –

–

– – – – – – –

– – – – – –

UICN–PACA

34 0.003 0.010 6.776 – – – 0.043 – 0.175 – – – – – –

UICN–Li

31 0.003 0.015 8.251 – – – – – –

– – – – – –.

UICN–MiP

17 –

– – – – – – –

– – – – – –

UICN–LaR

32 0.003 0.030 8.251 – – – – – –

– – – – – –

SN.SN

13–29 –

–

– –

– – – – – – –

Some significant differences appeared between Red Lists in adducing criteria B and C, no clear trends emerged when examining spatial scale. Overall, the proportions of adduced Red List criteria did not vary between lists at comparable spatial scales (global and European), although a few exceptions did occur between sub–national lists (table 4). Bibliographical categories The Red Lists at different scales used different literature. For instance, Red Lists at the global scale were more based on scientific literature sensu stricto (about 50 %) than European (about 10 %) and subnational Red Lists (0–6 %), which were, in turn, more based on grey literature (fig. 3, table 5). The Red List for geographical Europe was more based on scientific literature (12 %) and less on grey literature (53 %) than the list for the EU27 (7 % and 60 %, respectively). The

– – – 0.024 – 0.000

– – – – – –

European lists were more based on expert opinion (16–18 %) and less on books (15–17 %) than the global lists (0.4–1.6 % and 24.5–26.1 %, respectively). At comparable spatial scales, the updated global Red List was based on a smaller proportion of scientific literature (47 %) than the previous version (56 %) (fig. 3). There were also a few other differences between sub–national Red Lists related to their use of books and grey literature as sources (table 5). Citation categories The global and European Red Lists (the only ones in which the links between information and source could be analysed) had different proportions of citation categories (fig. 4, table 6). The results revealed that 'clearly supported' assertions were significantly more common in the global Red List (83 %) than in the

Charra & Sarasa

100%

n 14

90% 80% 70% 60% 50% 40%

30%

20%

10% Li

LaR

IDF

UICN–F

EU 27

Eur

IUCNd

D IUCNo

Fig. 2. Proportions of classification criteria on Red Lists for birds: A, population reduction; B, geographic range; C, small population size; D, very small or restricted regional population. (For other abbreviations see table 1). Fig. 2. Proporciones de los criterios de clasificación en las Listas Rojas de aves: A, reducción de población; B, distribución geográfica; C, tamaño de población pequeño y en decrecimiento; D, población local muy pequeña o limitada. (Para las otras abreviaturas véase tabla 1).

European list (32 %). 'Ambiguous' and 'not supported' assertions were both significantly more abundant in the European Red List (both 34 %) than in the global Red List (13 and 4 %, respectively) (fig. 4; table 6). Comparisons of citations in Red Lists according to the bibliographic category of the original sources highlighted the fact that 'clearly supported' assertions were more abundant in citations of books (86.5 %) and grey literature (79.1 %) from the global Red List while 'no supported' assertions were more abundant in citations of books (30.4 %) and grey literature (31.1 %) of the European Red List. In the global Red List, there were no significant differences in citation categories between the citations from different bibliographical category. However, in regard to the European List, 'ambiguous' assertions were more frequently linked to grey literature than to books and 'not supported' assertions were more frequently linked to books than to scientific articles and grey literature. Discussion We conducted this study for game bird species in France and therefore the applicability of the results to other species or other geographical areas still

remains an open question. Despite the huge volume of work that was required for this study, the sample size was still on occasions a limiting factor when attempting to unravel some of the less evident differences, for instance in comparisons at equivalent geographic scales. Notwithstanding this limitation, this study highlights clear trends in scale–dependent patterns. IUCN standards and transparency We found clear differences in the transparency and the traceability of the assessment processes used for Red Lists at different geographic scales. Although all the Lists considered in this study were certified by logos and labels that are directly linked to the IUCN guidelines (or indirectly through the IUCN French committee), national and sub–national lists in France were the product of data and processes that could not be verified and were not presented in a transparent and accessible way. Thus, the national and sub–national Red Lists in France do not fully comply with the standardized processes 'to be applied without deviation or modification, if regional Red List authorities wish to state that their assessment follows the IUCN system' (IUCN, 2012). Thus, we need additional information to be able to fully understand

Animal Biodiversity and Conservation 41.1 (2018)

Table 4. Results of comparison between Red Lists for the criteria leading to threatened status of birds: N, number of criteria; p, p–values of Fisher’s test; BHp, p–values corrected using the Benjamini–Hochberg procedure; OR, Odds ratio; A, population reduction; B, geographic range; C, small population size and decline; D, very small or restricted regional population; BL, European Red List from Birdlife International. (For other abbreviations see table 1). Tabla 4. Resultados de las comparaciones entre Listas Rojas de los criterios que determinan que las aves están amenazadas: N, cantidad de criterios; p, valor p de la prueba de Fisher; BHp, valor p corregido según el procedimiento de Benjamini–Hochberg; OR, razón de momios; A, reducción de población; B, distribución geográfica; C, tamaño de población pequeño y en decrecimiento; D, población local muy pequeña o limitada; BL, Lista Roja europea de Birdlife International. (Para las otras abreviatures véase tabla 1).

Criteria A N

BHp

Criteria B p

BHp

Criteria C p

BHp

Criteria D

BHp

IUCNo–IUCNd 3–5 – – – – – – – – – – – – IUCNo–BL.Eur 3–10 – – – – – – – – – – – – IUCNo–BL.EU27 3–18 – – – – – – – – – – – – IUCNd–BL.Eur 5–10 – – – – – – – – – – – – IUCNd–BL.EU27 5–18 – – – – – – – – – – – – IUCNo–UICN 3–13 0.001 0.002 Inf – – – – – – – – – IUCNd–UICN 5–13 0.015 0.015 Inf – – – – – – – – – IUCNo–Boe

3–13 0.015 0.021 Inf – – – – – – 0.015 0.042 0.000

IUCNo–PL

3–8 0.005 0.012 Inf – – – – – – 0.005 0.035 0.000

IUCNo–IDF

3–7 0.003 0.011 Inf – – – – – – – – –

IUCNo–Ce

3–10 0.001 0.005 Inf – – –

IUCNo–Al

3–11 0.004 0.011 Inf – – – – – – – – –

IUCNo–LaR

3–11 0.001 0.007 Inf – – – – – – 0.005 0.023 0.000

– – – 0.015 – 0.000

IUCNo–Li

3–13 0.000 0.000 Inf – – – – – – 0.002 0.028 0.000

IUCNd–Bo

5–13 – – – – – – – – – – – –

IUCNd–PL

5–8 0.036 0.042 Inf – – –

IUCNd–IDF

5–7 0.028 0.036 Inf – – – – – – – – –

IUCNd–Ce

5–10 0.015 0.023 Inf – – – – – – – – –

IUCNd–Al

5–11 0.038 0.041 Inf – – – – – – – – –

– – – 0.036 – 0.000

IUCNd–LaR

5–11 0.015 0.026 Inf – – –

IUCNd–Li

5–13 0.010 0.020 Inf – – –

– – – 0.045 – 0.000 – – 0.029 – 0.000

BL.Eur–BL.EU27 10–18 – – – – – – – – – – – – BL.Eur–UICN 10–13 0.000 0.000 Inf – – – – – – – – – BL.EU27–UICN

18–13 0.000 0.000 Inf 0.037 – 0.000 0.010 0.020 0.000 0.037 – 0.000

BL.Eur–Bo

10–13 0.000 0.000 Inf – – – – – – 0.000 0.000 0.000

BL.Eur–PL

10–8 0.000 0.000 Inf – – – – – – 0.000 0.000 0.000

BL.Eur–IDF

10–7 0.000 0.000 Inf 0.015 – 0.000 – – – – – –

BL.Eur–Ce

10–10 0.000 0.000 Inf – – – – – – 0.001 0.001 0.000

BL.Eur–Al

10–11 0.000 0.000 Inf – – – – – – 0.004 0.005 0.000

BL.Eur–LaR

10–11 0.000 0.000 Inf – – – – – – 0.000 0.000 0.000

BL.Eur–Li

10–13 0.000 0.000 Inf – – – – – – 0.000 0.000 0.000

BL.EU27–Bo

18–13 0.000 0.000 Inf – – – – – – 0.000 0.000 0.000

BL.EU27–PL

18–8 0.000 0.000 Inf – – – – – – 0.000 0.000 0.000

BL.EU27–IDF

18–7

0.000 0.000 Inf 0.003 0.042 0.000

–

0.015 0.016 0.000

BL.EU27–Ce 18–10 0.000 0.000 Inf – – – – – – 0.000 0.000 0.000

Charra & Sarasa

Table 4. (Cont.)

Criteria A N

BL.EU27–Al

BHp

Criteria B

18–11 0.000 0.000 Inf

p –

BHp OR –

–

Criteria C p

BHp

Criteria D

BHp

0.045 – 0.000 0.000 0.000 0.000

BL.EU27–LaR 18–11 0.000 0.000 Inf – – – – – – 0.000 0.000 0.000 BL.EU27–Li

18–13 0.000 0.000 Inf – – – – – – 0.000 0.000 0.000

UICN–Bo

13–13 –

UICN–PL

13–8 – – – – – – – – – 0.011 0.039 0.000

UICN–IDF

13–7 – – – – – – – – – – – –

UICN–Ce

13–10 – – – – – – – – – – – –

UICN–Al

13–11 – – – – – – – – – – – –

– – – – – 0.024 – Inf 0.040 – 0.143

UICN–LaR

13–11 – – – – – – – – – 0.020 0.047 0.058

UICN–Li

13–13 – – – – – – – – – 0.006 0.042 0.044

Bo–PL

13–8 – – – – – – – – – – – –

Bo–IDF

13–7 – – – 0.007 – 0.000 – – – – –

Bo–Ce

13–10 – – – – – – – – – – – –

Bo–Al

13–11 – – – – – – – – – – – –

Bo–LaR

13–11 – – – – – – – – – – – –

Bo–Li

13–13 – – – – – – – – – – – –

PL–IDF

8–7 – – – – – – – – – – – –

PL–Ce

8–10 – – – – – – – – – – – –

PL–Al

8–11 – – – – – – – – – – – –

PL–LaR

8–11 – – – – – – – – – – – –

PL–Li

8–13 – – – – – – – – – – – –

IDF–Ce

7–10 – – – 0.015 – Inf – – – – – –

IDF–Al

7–11 – – – 0.011 – Inf – – – – – –

IDF–LaR

7–11 – – – 0.015 – Inf – – – – – –

IDF–Li

7–13 –

Ce–Al

10–11 – – – – – – – – – – – –

– – 0.007 – Inf –

–

– 0.031 – 0.075

Ce–LaR

10–11 – – – – – – – – – – – –

Ce–Li

10–13 – – – – – – – – – – – –

Al–LaR

11–11 – – – – – – – – – – – –

Al–Li

11–13 – – – – – – – – – – – –

LaR–Li

11–13 – – – – – – – – – – – –

how lists including national and sub–national lists in France that deviate from the IUCN standards on transparency benefit from the labels that link them, either directly or indirectly via national committees, to the IUCN system. Likewise, on–line searchable databases including all the details of the data and assessments accessible via the Internet (e.g. http:// www.iucnredlist.org, available for the global Red List) should be required for all sub–global Lists. This improvement in transparency should be a priority for national and sub–national lists that are, at least in France, the less transparent ones.

Conservation status Our results for game birds in France agreed with the predictions derived from previous studies (Gärdenfors et al., 2001; Keller et al., 2005) and highlight the fact that Red Lists at smaller geographical scales frequently give higher threatened statuses than those at larger scales. The reason for this might be local variability in the status of species when compared to conservation status at larger scales. Species may exibit a threatened status first at a local level prior to exibiting threatened status at a global level or even, in occasion,

Animal Biodiversity and Conservation 41.1 (2018)

100 %

237

303

1,078

615

n 97

90 % 80 % 70 % 60 % 50 % 40 %

PACA

D IDF

0 %

MiP

C Br

10 % EU 27

Eur

20 %

IUCNd

IUCNo

30 %

Fig. 3. Proportions of bibliographical categories on Red Lists of birds: A, scientific literature sensu stricto; B, books and referenced academic publications; C, grey literature; D, expert opinion. (For other abbreviations see table 1). Fig. 3. Proporciones de las categorías bibliográficas en las Listas Rojas de aves: A, publicaciones científicas sensu stricto; B, libros y publicaciones académicas referenciadas; C, literatura gris; D, opinión de expertos. (Para las otras abreviaturas véase la tabla 1).

several species may be less threatened at local level than globally (Szabo et al., 2012). Nevertheless, the averagely higher threatened statuses of birds on Red Lists at smaller scales might also be linked to the risk of pessimistic assessments at regional level owing to over–narrow scale–dependent geographical focus (Keller et al., 2005; IUCN, 2012). As seen above, a risk of subjectivity in the regional adjustment process has already been identified (Keller and Bollmann, 2004; Rodríguez et al., 2004; Keller et al., 2005), in addition to risks of 'scale–' and 'edge–effect' (Keller et al., 2005; Milner–Gulland et al., 2006; Brito et al., 2010). Thus, these potential risks and the observed not supported data that was included in Red Lists does not rule out the possibility that the higher threatened statuses of Red Lists at smaller scales in France might be due, at least in part, to methodological factors or potential bias. Thus, methodological improvements aimed at reducing the risk of subjectivity in the second step of the IUCN regional guidelines and at avoiding the 'scale–' and 'edge–effect' in assessments are needed to strengthen the robustness of the Red Lists. Red List criteria Our results also match the predictions derived from previous articles on regional IUCN Red List assess-

ments (Eaton et al., 2005; Keller et al., 2005), and underscore the fact that assessments of threatened status in lists at smaller scales in France were predominantly based on the criterion 'D' (small regional population), while criteria linked to reductions of population (criterion 'A') were predominant in Red Lists at larger (European and global) scales. This scale–dependent characteristic of the assessment process may highlight and emphasize 'scale–' or 'edge–effects' in regional Red Lists, as has previously been reported (Eaton et al., 2005, Keller et al., 2005). Variation in the most commonly adduced criteria may reflect what data are available to assess species at the scale in question, and so data availability may hamper the feasibility and reliability of assessments at spatial scales that are too small. Bibliographical categories Our results concur with the predictions for bibliographical categories in Red Lists at different geographic scales and reveal that Red Lists at smaller scales were in general based on grey literature, while global Red Lists were more based on scientific articles. These results underlined the greater risk of subjectivity in Red Lists at smaller geographical scales. This risk is even greater in the European Red

Charra & Sarasa

Table 5. Results of comparisons between Red Lists of birds for bibliographical categories: N, number of references; p, p–values of Fisher’s test; BHp, p–values corrected using the Benjamini–Hochberg procedure; OR, odds ratio; A, scientific literature sensu stricto; B, books and referenced academic publications; C, grey literature; BL, European Red List from Birdlife International. (For ther abbreviations see tabe 1). Tabla 5. Resultados de las comparaciones entre Listas Rojas de aves para las categorías bibliográficas: p, valor p de la prueba de Fisher; BHp, valor p corregido según el procedimiento de Benjamini–Hochberg; OR, razón de momios; A, publicaciones científicas sensu stricto; B, libros y publicaciones académicas referenciadas; C, literatura gris; BL, Lista Roja europea de Birdlife International. (Para otras abreviaturas véase tabla 1).

Categoria A N

BHp

IUCNo–IUCNd 237–303 0.037 0.037 1.449

Categoria B p –

BHp –

Categoria C

OR –

p –

BHp OR –

–

Categoria D p –

BHp –

OR –

IUCNo–BL.Eur 237–1,078 0.000 0.000 8.689 0.009 0.009 1.584 0.000 0.000 0.198 0.000 0.000 0.022 IUCNo–BL.EU27 237–615 0.000 0.000 16.514 0.003 0.004 1.795 0.000 0.000 0.158 0.000 0.000 0.020 IUCNd–BL.Eur 303–1,078 0.000 0.000 5.997 0.001 0.002 1.724 0.000 0.000 0.287 0.000 0.000 0.088 IUCNd–BL.EU27 303–615 0.000 0.000 11.410 0.000 0.000 1.953 0.000 0.000 0.229 0.000 0.000 0.078 IUCNo–Bo

237–97 0.000 0.000 19.247 0.000 0.000 4.152 0.000 0.000 0.040 –

–

IUCNo–PACA

237–6

0.008 0.016 Inf

–

0.002 0.004 0.048 –

–

IUCNo–IDF

237–23 0.000 0.000 Inf

–

0.000 0.000 0.084 –

–

IUCNo–MiP 237–3 – – – 0.016 – 0.000 – – – – – IUCNo–Li

237–5 0.018 0.030 Inf

IUCNd–Bo

303–97

–

– 0.006 0.010 0.060 –

–

0.000 0.000 13.310 0.000 0.000 4.521 0.000 0.000 0.058

IUCNd–PACA

303–6

0.033 0.047 Inf

–

– 0.006 0.009 0.069 –

–

IUCNd–IDF

303–23 0.000 0.000 Inf

–

0.000 0.000 0.121 –

IUCNd–MiP 303–3 – – – 0.019 0.048 0.000 – – – – – – IUCNd–Li

303–5 – – – – – – 0.018 0.023 0.086 – – –

BL.Eur–BL.EU27 1,078–615 0.000 0.000 1.904 BL.Eur–Bo

1,078–97

–

0.028 0.028 0.797 –

–

0.009 0.030 2.627 0.000 0.000 0.200 0.000 0.000 18.207

BL.Eur–PACA 1,078–6 – – – – – – – – – – – – BL.Eur–IDF 1,078–23 – – – – – – – – – – – – BL.Eur–MiP 1,078–3 – – – 0.005 0.025 0.000 – – – – – – BL.Eur–Li

1,078–5 – – – – – – – – – – – –

BL.EU27–Bo

615–97

–

0.040

–

2.318 0.000 0.000 0.252 0.000 0.000 20.642

BL.EU27–PACA 615–6 – – – – – – – – – – – – BL.EU27–IDF 615–23 – – – – – – – – – – – – BL.EU27–MiP 615–3 – – – 0.004 0.040 0.000 – – – – – – BL.EU27–Li 1,078–5 – – – – – – – – – – – – Bo–PACA

97–6 – – –– – – – – – – – – –

Bo–IDF

97–23 – – – – – – – – – – – –

Bo–MiP

97–3

Bo–Li

97–5 – – – – – – – – – – – –

PACA–IDF

6–23 – – – – – – – – – – – –

–

0.001 0.010 0.000 0.004 0.040 Inf –

0.048

–

0.000 0.048

–

Inf

–

PACA–MiP

6–3

PACA–Li

6–5 – – – – – – – – – – – –

–

IDF–MiP

23–3 – – – 0.022 – 0.000 0.032 – Inf – – –

IDF–Li

23–5 – – – – – – – – – – – –

MiP–Li

3–5 – – – – – – – – – – – –

Animal Biodiversity and Conservation 41.1 (2018)

100 %

166

174

357

557

90 % 80 % 70 % 60 % 50 % 40 % 30 % 20 % 10 %

0 %

A B C Total Global IUCN Red List of birds Clear support

A B C Total European Red List of birds

Ambigous

No support

Fig. 4 . Proportions of citation categories on global and European Red List of birds: A, scientific literature sensu stricto; B, books and referenced academic publications; C, grey literature. Fig. 4. Proporciones de las categorías de citas en las Listas Rojas mundial y europea de aves: A, publicaciones científicas sensu stricto; B, libros y publicaciones académicas referenciadas; C, literatura gris.

List than in global Red Lists due to its higher reliance on expert opinions, even though the revisions of the IUCN assessment criteria in recent decades have been explicitly oriented towards reducing subjectivity (Mace and Lande, 1991; Rodrigues et al., 2006). Thus, these differences may weaken the reliability of regional Red Lists compared to global Red Lists. Furthermore, the greater dependence on scientific literature in the previous global Red List than in the current global Red List highlight the need for greater attention to be paid (1) to ensuring that the IUCN system, designed to provide comprehensive, scientifically rigorous information, is reliably applied at global and regional levels (IUCN, 2012), and (2) to preventing the current uncertain assessment processes at regional level, which use predominantly grey literature and expert opinions, from being increasingly applied at a global scale. If society collectively wants science– based and reliable Red Lists at small spatial scales and if peer–reviewed literature is (as it currently is) the widely accepted strategy for ensuring quality control in scientific research (Ferreira et al., 2015), the publication of small–scale studies in peer–reviewed journals will be necessary even despite the inherent difficulties of the publication process. The further integration of these potential needs by the editors of scientific journals might help promote more reliable Red Lists at regional scales in the future.

Citation categories Finally, our results agreed with the predictions on the citation categories and show that 'not supported' assertions were frequently linked to grey literature and books. Nevertheless, our results also highlighted the high degree to which the type of Red List affects these results. The global Red List is predominantly based on assertions 'clearly supported' by the cited references. 'Ambiguous' and 'not supported' citations were a minority (less than 20 %) in the global Red List but were a majority (more than 65 %) in our sample from the European Red List. Numerous citations of books in the European Red List were particularly questionable, for instance, old references that were cited (e.g. from 1964, 1977 or 1994) as support for assertions regarding recent short–term population trends (e.g. for Alauda arvensis, Aythya ferina, Gallinago gallinago, Limosa limosa, Numenius arquata, Tetrao urogallus). These results underline the fact that numerous assertions that were either ambiguous or not supported by the cited books and grey literature were included in regional assessments. Thus, the studied Red Lists, focused at different geographic scales, exhibit significant heterogeneity in both fundaments and reliability. This highlights the need for additional reviewing processes for sub–global assessments and, in particular, checks of

Charra & Sarasa

Table 6. Results of comparisons of citation categories between the Global and European Red Lists of birds: Bc, bibliographical category; Ns, number of sources; p, p–values of Fisher’s test; BHp, p–values corrected using the Benjamini–Hochberg procedure; OR, odds ratio; IUCN: Global Red List from the International Union for Conservation of Nature; BL, European Red List from Birdlife International; A, scientific literature sensu stricto; B, books and references academic publications; C, grey literature. Tabla 6. Resultados de las comparaciones de las categorías de citas entre las Listas Rojas de aves a escala mundial y europea: Bc, categoría bibliográfica; Ns, número de fuentes; p, valor p de la prueba de Fisher; BHp, valor p corregido según el procedimiento de Benjamini–Hochberg; OR, razón de momios; IUCN, Lista Roja mundial de la Unión Internacional por la Conservación de la Naturaleza; BL, Lista Roja europea de Birdlife International; A, publicaciones científicas sensu stricto; B, libros y publicaciones académicas referenciadas; C, literatura gris.

Clear support p

BHp

Ambiguous

BHp

No support p

BHp

IUCN–BL All

166–547

A oct–26

0.000 0.000 10.629

0.000 0.000 0.296

– – –

0.000 0.000 0.072 – – –

89–164

0.000 0.000 14.483

0.002 0.003 0.314

0.000 0.000 0.046

67–357

0.000 0.000 8.343

0.001 0.002 0.328

0.000 0.000 0.103

IUCN A.B oct–89

– – –

B.C 89–67

– – –

A.C oct–67

– – –

– – –.

BL A.B 26–174

– – –

B.C

–

174–357

A.C 26–357

–

– – –

the accuracy of links between citations and primary sources. Furthermore, the relative lack of data and the difficulties of the publication process may potentially increase the temptation to use grey literature in Red List assessments. Nevertheless, our results reveal that such practices increase the risk of inclusion of ambiguous and not supported data in regional assessments. Compensating for a lack of data (Butchart and Bird, 2010) by using grey literature may misrepresent the situation of species that would otherwise be regarded as 'data deficient' according to IUCN guidelines (DD conservation status), and thus may reduce the visibility of the ignored information that could justify financial support for additional research and species monitoring. Consequently, conservation decisions may be associated with an 'assessment dilemma': should we report data deficiencies strictly to highlight the need for further research and thus have to confront potential delays in evidence–based management?, or should we use all available information to promote reactive management, albeit at the risk of using not–supported and/or unreliable data in assessments, thereby jeopardizing the credibility

– – – 0.009 0.027 0.241 0.011 0.033 0.599 – – –

0.009 0.014 1.656 – – –

of Red Lists and reducing the visibility of needs to improve species monitoring? Further studies directly focused on this dilemma could have constructive implications for the monitoring and consensual conservation of species. Conclusion This study mainly revealed information about monitoring schemes, data collection and availability at different spatial scales in birds from Europe and France in particular. Sub–national to global Red Lists differed in regard to (i) the reported conservations status, (ii) the transparency and traceability of assessments, (iii) the most commonly adduced criteria, (iv) the categories of the sources synthetized during assessments, (v) and the reliability of assertions compared to cited references. Such variability between lists in terms of both data and transparency confirms that the sources used in the global Red List were cited as reliably as usual in ecological sciences (Todd et al., 2007, 2010). However, there were many ambiguous

Animal Biodiversity and Conservation 41.1 (2018)

and not supported citations on the European Red List and unverifiable assessments on the national and sub–national lists. These results thus open the door for further analysis and improvements of the reliability of Red Lists at regional levels and for other taxa to strengthen evidence–based wildlife management and avoid the decrease in the credibility and prestige of IUCN–based Red Lists (Mrosovsky, 1997; Mrosovsky and Godfrey, 2008) in the eyes of researchers, the general public and other stakeholders. Acknowledgements We would like to thank Mario Díaz, Judit Szabo and two anonymous reviewers for helpful comments on an earlier version of the manuscript. The Fédération Nationale des Chasseurs supported this study (n° FNC–PSN–PR12–2013). We would also like to thank Michael Lockwood and Judit Szabo for editing the English of the manuscript. References Abel, J., Babski, S.–P., Bouzendorf, F., Brochet, A. L., 2015. Liste rouge régionale des oiseaux nicheurs menacés en Bourgogne. LPO côte–d’Or. Akçakaya, H. R., Ferson, S., Burgman, M. A., Keith, D. A., Mace, G. M., Todd, C. R., 2000. Making consistent IUCN classifications under uncertainty. Conservation Biology, 14: 1001–1013. Atkinson, P. W., Austin, G. E., Rehfisch, M. M., Baker, H., Cranswick, P., Kershaw, M., Robinson, J., Langston, R. H. W., Stroud, D. A., Van Turnhout, C., Maclean, I. M. D., 2006. Identifying declines in waterbirds: The effects of missing data, population variability and count period on the interpretation of long–term survey data. Biological Conservation, 130: 549–559. Azam, C.–S., Gigot, G., Witte, I., Schatz, B., 2016. National and subnational Red Lists in European and Mediterranean countries: current state and use for conservation. Endangered Species Research, 30: 255–266. Benjamini, Y., Hochberg, Y., 1995. Controlling the False Discovery Rate: a practical and powerful approach to multiple testing. Journal of the Royal Statistical Society, Series B, 57: 289–300. Birard, J., Zucca, M., Lois, G., Natureparif, 2012. Liste rouge régionale des oiseaux nicheurs d’Île– de–France. Natureparif, Paris. BirdLife International, 2015. European Red List of Birds. Office for Official Publications of the European Communities. Luxembourg. Björk, B.–C., Welling, P., Laakso, M., Majlender, P., Hedlund, T., Guðnason, G., 2010. Open access to the scientific journal literature: situation 2009. PLoS ONE, 5: e11273. Bretagne Environnement, 2015. Liste rouge régionale, responsabilité biologique régionale, oiseaux nicheurs, oiseaux migrateurs de Bretagne. Bretagne vivante, Groupe ornithologique breton, Office

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Animal Biodiversity and Conservation 41.1 (2018)

Secondary habitats are important in biodiversity conservation: a case study on orthopterans along ditch banks A. Torma, M. Bozsó, R. Gallé

Torma, A., Bozsó, M., Gallé, R., 2018. Secondary habitats are important in biodiversity conservation: a case study on orthopterans along ditch banks. Animal Biodiversity and Conservation, 41.1: 97–108. Abstract Secondary habitats are important in biodiversity conservation: a case study on orthopterans along ditch banks. It has been shown that native biota can survive in secondary habitats such as road verges, dikes and hedges. We aimed to assess the conservation value of ditch banks for orthopterans in an agricultural landscape in Hungary, based on the analyses of species richness and abundance data using mixed–models. We did not find any differences in the species richness between isolated ditch banks, semi–isolated ditch banks and control meadows. The extent of isolation had a significantly negative effect, however, on the abundance of sedentary species. We found that the density of woody vegetation along ditch banks had a negative effect on the total abundance and the abundance of mobile species. Positive relationships were found between the width of ditch bank vegetation and the abundance of Caelifera, mobile, xerophilous and mesophilous species. Our results suggest that the density of orthopterans may be a more sensitive measure for habitat quality than their species richness. We concluded that ditch banks are a suitable habitat for the majority of orthopterans, including rare and endangered species, emphasizing that ditch banks and similar linear habitats should receive more attention and should be given a more prominent role in invertebrate conservation. Key words: Invertebrate diversity, Species traits, Linear habitat, Agricultural landscape Resumen Los hábitats secundarios son importantes en la conservación de la biodiversidad: un estudio práctico sobre los ortópteros en orillas de acequias. Se ha demostrado que la biota autóctona puede sobrevivir en hábitats secundarios como cunetas, diques y setos. La finalidad de este estudio es evaluar el valor de las orillas de acequias para la conservación de los ortópteros en un paisaje agrícola en Hungría, a partir del análisis de los datos relativos a la riqueza y la abundancia de especies utilizando modelos mixtos. No encontramos ninguna diferencia en cuanto a la riqueza de especies entre las orillas de acequias aisladas, semiaisladas y en praderas de control. Sin embargo, el grado de aislamiento tuvo un efecto negativo significativo en la abundancia de especies sedentarias. Constatamos que la densidad de vegetación leñosa junto a las orillas de las acequias tenía un efecto negativo en la abundancia total y la abundancia de especies móviles. Se observó la existencia de una relación positiva entre la anchura de las orillas de acequias que estaba cubierta por vegetación y la abundancia de especies del suborden Caelifera y de especies móviles, xerófilas y mesófilas. Nuestros resultados sugieren que la densidad de ortópteros puede ser una medida más sensible de la calidad del hábitat que la riqueza de especies. Concluimos que las orillas de las acequias son un hábitat adecuado para la mayoría de ortópteros, incluidas las especies raras o en peligro de extinción, lo que pone de relieve que debería prestarse más atención a estos y otros hábitats lineales parecidos y que se les debería dar más importancia en la conservación de invertebrados. Palabras clave: Diversidad de invertebrados, Características de las especies, Hábitat lineal, Paisaje agrícola Receiced: 9 IX 16; Conditional acceptance: 12 XII 16; Final acceptance: 20 VI 17 Attila Torma, Róbert Gallé, Dept. of Ecology, Univ. of Szeged, Közép fasor 52, Szeged, H–6726, Hungary.– Miklós Bozsó, Plant Health and Molecular Biology Lab., Directorate of Plant Protection, Soil Conservation and Agri–environment, National Food Chain Safety Office, Budaörsi u. 141–145, Budapest H–1118, Hungary. Corresponding author: Attila Torma. E–mail: torma_a@yahoo.com ISSN: 1578–665 X eISSN: 2014–928 X

Torma et al.

Introduction Natural and semi–natural grasslands in Europe still contain a diverse fauna and flora, but recent studies (e.g. Hernández–Manrique et al., 2012; Torma and Bozsó, 2016) conclude that existing conservation strategies based mainly on the protection of areas of high natural value may be insufficient to ensure conservation of the invertebrate species pool at landscape scale. Conservation of the invertebrate diversity thus needs a landscape perspective. In Europe, habitat destruction and deterioration caused by the intensification of agriculture and the change in landscape patterns such as increasing fragmentation and isolation of habitats have been shown to result in a decline of biodiversity (Kruess and Tscharntke, 1994; Stoate et al., 2001; Jongman, 2002). As conservation strategies, agri–environmental schemes aim to reduce the impact of agricultural activities on species that inhabit the agricultural landscape. However, these programs have only a limited effect on European agriculture due to land–owners’ reluctance to participate (Espinosa– Goded et al., 2010), and their efficiency in biodiversity conservation is under debate. Tscharntke et al. (2005) suggested that agri–environmental programs may be effective in simple, but not in complex landscapes where a biodiversity is already likely to be higher. In contrast, Duelli and Obrist (2003) highlighted that these programs have a major chance of success in complex landscapes where arthropods can also survive in nearby habitats. To avoid a decrease in the diversity of arthropods and thus, in the ecosystem services and functions provided by them, we urgently need to seek possibilities for proper conservation strategies adapted to the regional landscape features and history (Tscharntke et al., 2005; Batáry et al., 2015). Many recent studies have highlighted the importance of linear secondary habitats such as road verges (e.g. Saarinen et al., 2005; Söderström and Hedblom, 2007), dikes (e.g. Torma and Császár, 2013; Bátori et al., 2016), and hedges (e.g. Ernoult et al., 2013; Morandin and Kremen, 2013) in biodiversity conservation. It has been shown that native biota can survive in these habitats. Such anthropogenic habitats often have a long history, facilitating development of species–rich habitats (Musters et al., 2009), and they may provide resources for populations of rare and endangered species (Torma and Bozsó, 2016). In contrast, newly established sawn grass strips and abandoned field margins are comparatively species poor and are beneficial particularly for common species (Musters et al., 2009; Ernoult et al., 2013). If they remain intact for the long term, it is possible they will develop to a species rich secondary habitat, similarly to road verges, dikes, etc. The goal of our study was to assess the ecological value of ditch banks as secondary habitats for invertebrate conservation in an agricultural landscape. While the remaining natural and semi–natural habitats within arable fields are generally regarded as crucial for wildlife, the value of ditch banks for providing habitats and refugia remains an open question (Herzon and Helenius, 2008; Musters et al., 2009). We studied species richness and abundance of orthopterans at

ditch banks in the Tisza–Maros angle in the southern part of the Great Hungarian Plain. We chose to study orthopterans because they are among the most important consumers and abundant prey sources for many vertebrates (Rodríguez and Bustamante, 2008; Kiss et al., 2014), and their diversity is currently declining in many temperate regions (Berg and Zuna–Kratky, 1997; Maas et al., 2002; Reinhardt et al., 2005; Krištin et al., 2007; Holuša et al., 2012). The sensitivity of species to environmental conditions is a function of their ecological and life history traits. In the present study, we considered dispersal ability, habitat affinity and reproduction strategy traits because they are hypothesized to be key determinants of species persistence (Kotiaho et al., 2005). The dispersal ability and habitat affinity of species highly influences their responses to landscape features (Joern and Laws, 2013). Sedentary species are generally more affected by fragmentation and isolation of habitats than mobile species that can (re)colonize relatively distant habitat patches (e.g. Marini et al., 2010, 2012). Similarly, generalist species are more likely to find suitable habitat patches in a fragmented landscape than specialist species (e.g. Collinge, 2000). Besides the number of offspring, reproduction strategy can influence species persistence in various manners. For instance, Ensifera species usually produce larger eggs then Caelifera, and lay those individually in plants or under tree bark, and this can increase, for example, the chance of hydrochory (Dziock et al., 2011). We also focused on immature orthopterans as they are usually sedentary and a large number of immature specimens indicates reproductive sites. We addressed the following questions: (1) are there significant differences between isolated ditch banks, semi–isolated ditch banks and control meadows in species richness and abundance of orthopterans? (2) are there significant relationships between the width of ditch bank vegetation and Orthopteran species richness and abundance? and (3) do the presence and density of woody vegetation along ditch banks influence species richness and abundance of orthopterans? Material and methods Study region The study was carried out in an approximately 150 km2 area close to the confluence of the Maros and Tisza rivers in Csongrád County, Hungary (fig. 1). As a part of the Great Hungarian Plane, the area is characterized by dry continental climatic conditions. The annual mean temperature is 10.5–10.6 °C and the average annual rainfall is 570 mm. Before the rivers were regulated, the area was frequently flooded and characterized by wet grasslands (Bátori et al., 2016). After river regulation and drainage works, which were typical in the 19th and 20th centuries, the lowered water levels and desiccation of habitats induced secondary salt accumulation in higher soil layers, especially in former wet, non–alkali meadows (Molnár and Borhidi, 2003). Although most grasslands

Animal Biodiversity and Conservation 41.1 (2018)

Hungary

River Tisza

River Maros 5 km

Fig. 1. On the schematic map of Hungary (upper left corner), the empty square represents the locality of the study area. The satellite imagines show the study area with the drainage system and the localities of sampling sites. Black circles and half–black circles represent isolated ditch bank sections and connected ditch bank sections, respectively. Empty circles represent control meadows. Fig. 1. En el mapa esquemático de Hungría (esquina superior izquierda), el cuadrado vacío representa la localidad de la zona de estudio. La imagen por satélite muestra la zona de estudio con el sistema de drenaje y la ubicación de los sitios de muestreo. Los círculos negros y los que tienen una mitad de color negro representan las secciones de orillas de acequias aisladas y conectadas, respectivamente. Los círculos vacíos representan las praderas de control.

were transformed into arable fields, alkaline grassland patches were not cultivated because their poor soil quality was unsuitable for intensive agriculture and forestry (Bátori et al., 2016). Currently, the area is dominated by arable fields with a considerable drainage system that encloses the grassland remnants. Sampling design We applied a nested balanced design. Five sites were sampled within each c.a. 500 m long selected section of ditches and within each control meadow. Sections of ditches were selected according to isolation treatment i.e., isolated or semi–isolated. Sections of ditches were

considered isolated when running through arable fields with no meadows in their surroundings, such as in a buffer of 1,000 m radius. Sections connected with meadows were considered as semi–isolated sections. For controls, we chose meadows because they are presumably the preferred habitats for orthopterans in the landscape. Arable fields were not targeted in the present study because they generally provide a poor habitat for most orthopterans (e.g. Marshall et al., 2006). Four replicates were selected for each treatment and control, and they were located at least two kilometers apart from each other. Minimal distance between sites within each section and within control meadows was 100 m. Orthopterans were sampled by

100

sweep netting. In each site 50 sweeps were performed four times (25 VI, 28 VII, 29 VIII and 27 IX) in 2012. Since our focus was not on the seasonal dynamics of the orthopterans, species–abundance data matrix were pooled according to sampling periods. At each site, the width of the strip–like vegetation was measured, and the density of woody vegetation was assessed by visual observation. We used three categories: absent (no woody vegetation), present (a single tree or one–two single bushes), dense (more than one tree and/or more than three bushes). Species traits Based on the mobility index as a measure of dispersal ability (Reinhardt et al., 2005), two mobility classes (sedentary and mobile species) were analyzed (Marini et al., 2012). The specific preferences for humidity were used to group them in relation to their habitat specialization, and they were sorted into xerophilous, mesophilous and hygrophilous species groups (cf. Fartmann et al., 2012). We distinguished Ensifera and Caelifera groups to represent the differences between them e.g. in reproductive potential and egg deposition of females (Torma and Bozsó, 2016). Based on the mean number of ovarioles (Reinhardt et al., 2005), Ensifera species are usually considered to have a high reproductive potential compared to Caelifera. Statistical analyses According to the nested design, generalized linear mixed models (GLMM, Poisson and negative binomial errors, maximum likelihood fit) were applied and the effect of sites nested within sections was used as random effect. First, we analyzed the species richness and abundance of orthopterans in relation to the treatment, that is, isolated, connected and control. Pairwise comparisons were carried out with the help of 'relevel' function and Bonferroni corrections were applied. In a second set of models, we analyzed the species richness and abundance data in relation to the width of ditch bank vegetation and the density of woody vegetation along ditch banks. Since hygrophilous species were represented by very restricted numbers of species and individuals, we analyzed their presence / absence using a binomial model. All statistical analyses were carried out in an R Statistical Environment (R Core Team, 2013), using lme4 package (Bates et al., 2013). Results Altogether, we collected 4,212 and 940 adult individuals of 17 Caelifera and 13 Ensifera species, respectively (table 1). Immature specimens of Acrididae were also collected in a high number (table 1). Therefore, their abundance was only considered in the analyses. According to the mobility of species, 18 mobile and nine sedentary species were distinguished; 19 and nine

Torma et al.

species were sorted into the categories of xerophilous and mesophilous species, respectively; however only two hygrophilous species were collected. The most abundant species were Euchorthippus declivus (Brisout de Barneville, 1849) (with a frequency of 29.3 %), Omocestus haemorrhoidalis (Charpentier, 1825) (16.5 %), Chorthippus brunneus (Thunberg, 1815) (10.6 %) and Oecanthus pellucens (Scopoli, 1763) (10.2 %). Species with a high natural value were also collected. However, Gampsocleis glabra (Herbst, 1786) and Modicogryllus frontalis (Fieber 1844), for example, were represented by only one specimen. Epacromius coerulipes (Ivanov, 1887) was collected in only one ditch bank section beside control meadows, whereas e.g. Ruspolia nitidula (Scopoli, 1786) was collected only along ditch banks. Acrida ungarica Herbst 1786 and Tessellana veyseli (Koçak, 1984) were collected in almost all sites. Species richness and abundance pattern of Orthoptera assemblages The results of GLMM did not show any significant differences in the species richness of orthopterans between isolated ditch banks, semi–isolated ditch banks and control meadows; nearly significant differences were found in the species richness of mobile and mesophilous species between control meadows and isolated ditch banks (table 2). The extent of isolation had a significant effect on the abundance of sedentary species (table 2). The highest and lowest abundances of sedentary species were found in control meadows and isolated ditch banks, respectively (fig. 2). No other significant differences in the abundance of orthopterans were found between isolated ditch banks, semi–isolated ditch banks and control meadows. We analyzed the presence / absence of hygrophilous species using a binomial model and we did not find any significant effects (control vs. connected: z = 1.135, p = 0.257; control vs. isolated: z = 0.624, p = 0.532; isolated vs. connected: z = 0.550, p = 0.582). Neither the width of ditch bank vegetation nor the density of woody vegetation had any significant effects on the species richness of orthopterans, but both had effects on their abundance pattern according to the results of the GLMM (table 3). Presence of dense woody vegetation had a significant negative effect on the total number of individuals and the number of mobile individuals (fig. 3), and had a marginally significant negative effect on the abundance of Caelifera and xerophilous species. Significant positive relationships were found between the width of ditch bank vegetation and the abundance of Caelifera, mobile, xerophilous and mesophilous species (fig. 4). We also found a marginally significant effect of the width of ditch bank vegetation on the total individual number of orthopterans. We did not find any significant effects of ditch bank vegetation on the presence ⁄absence of hygrophilous species (width of vegetation: z = –0.423, p = 0.679; woody vegetation: z = –0.064, p = 0.949; dense woody vegetation: z = 0.045, p = 0.964).

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Table 1. Collected species of Orthoptera: Trait 1, mobility (Mob, mobile; Int, intermediate; Sed, sedentary); Trait 2, humidity preference (Xero, xerophilous; Mezo, mesophilous; Hygro, hygrophilous); Crl, control meadows; Con, connected ditch banks; Iso, isolated ditch banks. Tabla 1. Especies de ortópteros recogidas: Trait 1, movilidad (Mob, móvil; Int, intermedia; Sed, sedentaria); Trait 2, preferencia por la humedad (Xero, xerófilas; Mezo, mesófilas; Higro, higrófilas); Crl, praderas de control; Con, orillas de acequias conectadas; Iso, orillas de acequias aisladas. Taxa

Trait 1

Trait 2

Ensifera

Crl

Con

Iso

Total

Conocephalus discolor Thunberg, 1815

Mob Hygro 22 32 52 106

Gampsocleis glabra (Herbst, 1786)

Sed

Xero

Leptophyes albovittata (Kollar, 1833)

Sed

Xero

111

Leptophyes discoidalis (Frivaldszky, 1868)

Sed

Mezo

Metrioptera bicolor (Philippi, 1830)

Mob

Xero

Metrioptera roeselii (Hagenbach, 1822)

Int

Mezo

Oecanthus pellucens (Scopoli, 1763)

Mob

Xero

104

354

524

Phaneroptera nana Fieber, 1853

Mob

Xero

Platycleis affinis Fieber, 1853

Int

Xero

Platycleis grisea (Fabricius, 1781)

Int

Tessellana veyseli (Koçak, 1984)

Sed Xero 61 22 29 112

Modicogryllus frontalis (Fieber 1844)

Sed

Mezo

Ruspolia nitidula (Scopoli, 1786)

Mob

Hygro

Caelifera

Acrida ungarica Herbst, 1786

Mob

Xero

Calliptamus barbarus (Costa, 1836)

Mob

Xero

Calliptamus italicus (Linnaeus, 1758)

Mob

Xero

Chorthippus oschei Helversen, 1986

Mob

Mezo

Chorthippus dichrous (Eversmann, 1859)

Mob

Mezo

143

Chorthippus dorsatus (Zetterstedt, 1821)

Mob

Mezo

169

254

Chorthippus parallelus (Zetterstedt, 1821)

Mob

Mezo

111

Chorthippus brunneus (Thunberg, 1815)

Mob

Xero

197

137

216

550

Chorthippus mollis (Charpentier, 1825)

Mob

Xero

134

Chorthippus vagans (Eversmann, 1848)

Sed

Xero

Chorthippus biguttulus (Linnaeus, 1758)

Mob

Xero

Epacromius coerulipes (Ivanov, 1887)

Mob

Xero

Euchorthippus declivus (Brisout de Barneville, 1849) Mob

Xero

621

646

247 1,514

Omocestus haemorrhoidalis (Charpentier, 1825) Sed

Mezo

476

277

851

Omocestus petraeus (Brisout de Barneville, 1855) Sed

Xero

Omocestus rufipes (Zetterstedt, 1821)

Sed

Xero

Pezotettix giornae (Rossi, 1794)

Mob

Xero

140

235

452

Acrididae nymph

694

485

504 1,683

Catantopidae nymph

Conocephalidae nymph

Tettigonidae nymphs

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Table 2. The effects of the extent of isolation of ditch banks on species richness (left side) and abundance (right side) of orthopterans delineated by mixed models (GLMM). Poisson and negative binomial error terms were used to analyse species richness and abundance data, respectively. Pairwise comparisons were carried out with the help of the 'relevel' function in R, and Bonferroni corrections were applied: Crl, control meadows; Iso, isolated ditch banks; Con, connected ditch banks. Tabla 2. Los efectos del grado de aislamiento de las orillas de las acequias en la riqueza (izquierda) y la abundancia (derecha) de especies de ortópteros definidos por los modelos mixtos (GLMM). Se utilizaron los términos de error que siguen una distribución de Poisson y binomial negativa para analizar los datos relativos a la riqueza y la abundancia de especies, respectivamente. Se realizaron comparaciones por pares con la ayuda de la función de reordenación de niveles (relevel) en R y se aplicaron las correcciones de Bonferroni: Crl, praderas de control; Iso, orillas de acequias aisladas; Con, orillas de acequias conectadas.

Treatment

Orthoptera

Parameter Parameter estimation (± SE) z p estimation (± SE) 0.669

–0.128 (0.248)

–0.517

0.605

Crl vs. Con

0.041 (0.095)

Crl vs. Iso

0.125 (0.093)

1.350

0.178

–0.041 (0.247)

–0.166

0.868

Iso vs. Con –0.085 (0.092)

–0.920

0.357

–0.087 (0.248)

–0.351

0.726

Caelifera

0.430

Crl vs. Con

0.056 (0.118)

0.475

0.635

–0.189 (0.323)

–0.586

0.558

Crl vs. Iso

0.142 (0.116)

1.218

0.223

–0.234 (0.323)

–0.725

0.468

Iso vs. Con –0.085 (0.114)

–0.744

0.457

0.045 (0.323)

0.139

0.889

Ensifera

Crl vs. Con –0.150 (0.213)

–0.703

0.482

–0.155 (0.517)

–0.300

0.764

Crl vs. Iso

0.148 (0.202)

0.731

0.465

0.540 (0.515)

1.050

0.294

Iso vs. Con –0.298 (0.208)

–1.430

0.153

–0.695 (0.516)

–1.348

0.178

Mobile species Sedentary species

Crl vs. Con

0.076 (0.117)

0.647

0.517

0.056 (0.386)

0.146

0.884

Crl vs. Iso

0.253 (0.113)

2.237

0.075

0.358 (0.385)

0.930

0.352

Crl vs. Con

–0.177 (0.111)

–1.597

0.165

–0.302 (0.385)

–0.785

0.433

Crl vs. Iso

–0.197 (0.182)

–1.085

0.278

–0.544 (0.230)

–2.362

0.027

Crl vs. Con –0.127 (0.178)

–0.712

0.476

–0.991 (0.233)

–4.239

< 0.001

Crl vs. Iso

–0.070 (0.187)

–0.375

0.708

0.447 (0.236)

–2.029

0.049

–0.119 (0.118)

–1.006

0.314

–0.198 (0.277)

–0.717

0.473

0.013 (0.114)

0.115

0.909

–0.169 (0.277)

–0.610

0.542

–1.121

0.262

–0.029 (0.277)

–0.107

0.915

0.860

0.390

0.101 (0.440)

0.229

0.818

Xerophilous Crl vs. Con species Crl vs. Iso

Iso vs. Con –0.132 (0.118)

Mesophilous Crl vs. Con species Crl vs. Iso Acrididae nymphs

0.175 (0.203) 0.426 (0.193)

2.201

0.067

0.889 (0.404)

2.199

0.083

Iso vs. Con –0.180 (0.160)

–1.119

0.263

–0.758 (0.434)

–1.743

0.122 0.257

Crl vs. Con

–0.779 (0.688)

–1.133

Crl vs. Iso

–0.950 (0.686)

–1.383

0.167

Iso vs. Con

0.170 (0.690)

0.247

0.805

Discussion To assess the ecological value of ditch banks, we compared species richness and abundance of orthopterans between isolated ditch banks, semi–isolated ditch banks and control meadows. Species richness did not differ between ditch banks and control meadows, but significant differences were found in the abundance pattern of orthopterans. Braschler et al.

(2009) suggested that fragmentation and isolation may have a stronger effect on the abundance of orthopterans than on their species richness. Similarly, farming practices are also known to particularly influence the density of orthopterans (Badenhausser and Cordeau, 2012). It seems that the density of orthopterans is a more sensitive measure of the quality of grassy habitats than their species richness, as was previously concluded by Báldi and Kisbenedek (1997).

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Table 3. The effects of vegetation in ditch banks on species richness (left side) and abundance (right side) of orthopterans delineated by mixed models (GLMM). Poisson and negative binomial error terms were used to analyze species richness and abundance data, respectively: Width, width of ditch bank vegetation; Present, presence of woody vegetation; Dense, presence of dense woody vegetation. Tabla 3. Los efectos de la vegetación de las orillas de las acequias en la riqueza (izquierda) y la abundancia (derecha) de especies de ortópteros definidos por los modelos mixtos. Se utilizaron los términos de error que siguen una distribución de Poisson y binomial negativa para analizar los datos relativos a la riqueza y la abundancia de especies, respectivamente: Width, anchura de la orilla de la acequia cubierta por vegetación; Present, presencia de vegetación leñosa; Dense, presencia de vegetación leñosa densa.

Variable Orthoptera

Parameter Parameter estimation (± SE) z p estimation (± SE)

Width

–0.003 (0.035)

–0.071

0.943

0.105 (0.062)

1.680

0.093

Present

0.047 (0.109)

0.437

0.662

–0.025 (0.128) –0.195

0.845

Dense

–0.110 (0.155)

–0.711

0.477

–0.559 (0.191) –2.921

0.003

Width

0.036 (0.043)

0.839

0.402

0.246 (0.078)

3.127

0.002

Present

0.037 (0.136)

0.270

0.788

–0.087 (0.109) –0.791

0.428

Dense

–0.184 (201)

–0.917

0.359

–0.289 (0.174) –1.661

0.096

Caelifera

Ensifera

Width

0.005 (0.087)

0.065

0.947

0.065 (0.139)

0.467

0.640

Present

–0.048 (0.207)

–0.234

0.814

0.273 (0.233)

1.172

0.241

Dense

0.017 (0.311)

0.054

0.956

–0.353 (0.361) –0.978

0.328

Width

0.017 (0.041)

0.410

0.682

0.177 (0.066)

2.670

0.007

Mobile species Sedentary species

Present

0.088 (0.129)

0.681

0.496

0.060 (0.103)

0.580

0.561

Dense

–0.171 (0.192)

–0.889

0.374

–0.416 (0.164) –2.535

0.011

Width

0.051(0.071)

0.726

0.468

0.003 (0.098)

0.037

0.970

Present

–0.161(0.235)

–0.687

0.492

–0.013 (0.196) –0.070

0.944

Dense

–0.091 (0.314)

–0.290

0.771

0.063 (0.314)

0.203

0.839

Xerophilous species

Width

0.002 (0.045)

0.053

0.958

0.168 (0.007)

2.389

0.016

Present

0.085(0.139)

0.612

0.541

–0.011 (0.116)

0.098

0.922

Dense

–0.067 (0.197)

–0.341

0.733

–0.342 (0.181) –1.885

0.059

Mesophilous species

Width

0.095 (0.065)

1.456

0.145

0.251 (0.009)

2.575

0.010

Present

–0.086 (0.215)

–0.403

0.686

0.013 (166)

0.080

0.936

Dense

–0.335 (0.332)

–1.009

0.312

–0.185 (0.276) –0.670

0.503

0.029 (0.073)

0.692

Acrididae nymphs

Width

0.396

Present

–0.135 (0.196) –0.690

0.490

Dense

–0.046 (0.322) –0.144

0.885

Based on the analyses of the trait groups separately, we showed that the mobility of species has a prominent role in shaping the abundance pattern of orthopterans, and sedentary species are presumably not able to build viable populations along ditch banks. This is in accordance with numerous studies highlighting the importance of dispersal ability of orthopterans in agricultural landscapes (Dziock et al., 2011; Marini et al., 2010, 2012; Torma and Bozsó,

2016; Poniatowski and Fartmann, 2010). In general, low mobility of insects is linked to their increased vulnerability to extinctions in a fragmented landscape since sedentary species, for instance, are less able to (re)colonize remaining suitable habitats in the unsuitable matrix (Braschler et al., 2009; Bommarco et al., 2010; Habel et al., 2016). Linear habitats in the agricultural matrix, however, have shown to be preferred for insect dispersal (Berggren et al., 2002;

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250 Nunber of individuals

60 40 20 0

Fig. 2. Box plots represent the differences in the abundance of sedentary species of Orthoptera between isolated ditch banks, connected ditch banks and control meadows, delineated by the GLMM: * P < 0.05; ** P < 0.01; *** P < 0.001. Circles mark outlier data. (Further details are given in table 2). Fig. 2. El diagrama de caja representa las diferencias en la abundancia de especies sedentarias de ortópteros entre orillas de acequias aisladas, orillas de acequias conectadas y praderas de control, definidas por el modelo lineal generalizado mixto (GLMM): * P < 0,05; ** P < 0,01; *** P < 0,001. Los círculos indican los datos atípicos. (En la tabla 2 pueden consultarse más detalles).

Saarinen et al., 2005; Söderström and Hedblom, 2007), even for flightless and sedentary species (Poniatowski and Fartmann, 2010), suggesting the importance of such habitats in connecting populations. Besides their corridor function, linear habitats in the agricultural matrix can also have an important role in foraging and reproduction of animals (Huusela–Veistola and Vasarainen, 2000; Downs and Racey, 2006; Marshall et al., 2006). As immature orthopterans were present along ditch banks in a similar number to that in control meadows, ditch banks presumably provide suitable conditions for reproduction, particularly for grasshoppers. This is an important issue considering that different ecological conditions are often required for larval development and for spreading and foraging of adults (e.g. Hodek, 2003). In strips of mowed grass, for instance, high grasshopper (Gomphocerinae) densities consisted of a high density of adults but not of immature grasshoppers (Badenhausser and Cordeau, 2012). The width of the vegetation and the presence of dense woody vegetation along ditch banks affected the orthopterans more than the extent of isolation of ditch banks. Woody vegetation is known to influence

200 150 100 50 0

Control Connected Isolated

Absent

Present

Dense

200 Nunber of individuals

Nunber of individuals

150 100 50 0

Absent

Present

Dense

Fig. 3. Box plots show the differences in the total number of individuals (A) and in the number of mobile individuals (B) in relation to the density of woody vegetation: Absent, no woody vegetation; Present, a single tree or one–two single bushes; Dense, more than one tree and/or more than three bushes; * P < 0.05; ** P < 0.01; *** P < 0.001. (Further details are given in table 2). Fig. 3. Los diagramas de caja muestran las diferencias en el número total de individuos (A) y en el número de individuos móviles (B) en relación con la densidad de vegetación leñosa. Abreviaciones: Absent, sin vegetación leñosa; Present, un único árbol o uno o dos arbustos individuales; Dense, más de un árbol o más de tres arbustos: * P < 0,05; ** P < 0,01; *** P < 0,001. (En la tabla 2 pueden consultarse más detalles).

arthropod communities via alternating nearby environmental conditions such as soil water content, microclimate, vegetation, light regime, etc. (Sparks and Greatorex–Davis, 1992; Entling et al., 2007; Gossner, 2009; Torma and Gallé, 2011). The negative effect of woody vegetation on orthopterans has been shown in previous studies (Samways and Moore, 1991; Bieringer and Zulka, 2003). However, grasshoppers

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Nunber of individuals

250

200

150

100

50 0

8 C

250 200

250

200

Nunber of individuals

105

8 D

40 30

150

100 10

50 0 0

0 6 8 0 2 Width of ditch bank vegetation (m)

Fig. 4. The relationship between the width of ditch bank vegetation and the abundance of: A, Caelifera; B, mobile species; C, xerophilous species; D, mesophilous species. (Further details are given in table 2). Fig. 4. La relación entre la anchura de la orilla de la acequia cubierta por vegetación y la abundancia de: A, Caelifera; B, especies móviles; C, especies xerófilas; D, especies mesófilas. (En la tabla 2 pueden consultarse más detalles).

seem to be more affected by the presence of dense woody vegetation and the width of grassy vegetation than Ensifera species. Most grasshoppers prefer open habitats whereas Ensifera species often require habitats consisting of both grassy and shrubby vegetation patches (Schirmel et al., 2010). As artificial strip–like habitats generally have a quasi–constant width, the variation in their width is generally too low to detect effects on the distribution of species (Badenhausser and Cordeau, 2012). In the present study, the width of vegetation along ditches was more variable, resulting in significant effects on orthopterans. This variation in the width of vegetation was presumably due to the differences in the ditches (e.g. the steepness of bank slope, water regime, etc.) and in the surrounding land use. In some cases, arable fields or dirt roads were situated as close to ditches as is physically possible, reducing the width of ditch bank vegetation. Reduced width of vegetation can reduce humidity in ditch banks, whose condition is preferred

by certain species (Herzon and Helenius, 2008). Soil moisture also influences the larval development of orthopterans (Hodek, 2003). However, we did not find differences in the distribution of hygrophilous species and immature orthopterans in relation to the width of vegetation. Presumably, a narrower vegetation–strip along ditches gained fewer resources for foraging and fewer resting and hiding places, causing a lower abundance in general. In a linear habitat it is crucial whether it is functioning as a suitable habitat (provides resources needed for survivorship, reproduction, and movement), a corridor (provides some resources, especially for movement, but not necessarily for reproduction) or an ecological trap or sink for animals (Chetkiewicz et al., 2006). The role of linear grassy habitats as corridors for orthopterans was highlighted by previous studies in the region (Gausz, 1969; Krausz et al., 1995; Kisbenedek et al., 2010). Our findings suggest that ditch banks, like dikes (Torma and Bozsó, 2016),

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can be a suitable habitat, providing resources for survivorship, reproduction and movement for most orthopterans including rare and endangered species. Numerous collected species e.g., T. veyseli, G. glabra, E. coerulipes, M. frontalis, R. nitidula, C. italicus and A. ungarica are included in National Red Lists as endangered or critically endangered species in surrounding countries (e.g., Berg et al., 2005; Maas et al., 2002; Liana, 2007; Holuša et al., 2013). T. veyseli is suggested to be close to extinction at the edge of Pannon region (Holuša et al., 2012), while R. nitidula is currently spreading (Krištin et al., 2007; Holuša et al., 2013). Some species such as M. frontalis and C. italicus are locally common in Hungary as well as in eastern and southern countries in Europe respectively, but they are endangered and declining in Central Europe (Liana, 2007). The decline of these species is often considered a consequence of the loss and destruction of their habitats (Liana, 2007; Rada and Trnka, 2016; Holuša, 2012; Holuša et al., 2012). Considering that the above species generally occur along linear secondary habitats in the region (Gausz, 1969; Krausz et al., 1995; Torma and Bozsó, 2016), and further endangered species such as the endemic Isophya costata Brunner von Wattenwyl, 1878 and Isophya stisy Cejchan, 1957 were also detected (Kisbenedek et al., 2010), we highlight the importance of linear secondary habitats for orthopterans and presumably for other arthropod groups even in countries where a considerable area of natural, semi–natural grasslands still harbor rich invertebrate fauna. References Badenhausser, I., Cordeau, S., 2012. Sown grass strip – A stable habitat for grasshoppers (Orthoptera: Acrididae) in dynamic agricultural landscapes. Agriculture, Ecosystems and Environment, 159: 105–111. Báldi, A., Kisbenedek, T., 1997. Orthopteran assemblages as indicators of grassland naturalness in Hungary. Agriculture, Ecosystems and Environment, 66: 121–129. Batáry, P., Dicks, L. V., Kleijn, D., Sutherland, W. J., 2015. The role of agri–environment schemes in conservation and environmental management. Conservation Biology, 29: 1006–1016. Bates, D., Maechler, M., Bolker, B., Walker, S., 2013. lme4: Linear mixed–effects models using Eigen and S4. R package version 1.0–5. http:// CRAN.R– project.org/package=lme4 [Accessed on 19 March 2014]. Bátori, Z., Körmöczi, L., Zalatnai, M., Erdős, L., Ódor, P., Tölgyesi, C., Margóczi, K., Torma, A., Gallé, R., Cseh, V., Török, P., 2016. River Dikes in agricultural landscapes: the importance of Secondary habitats in maintaining landscape–scale diversity. Wetlands, 36: 251–264. Berg, H. M., Bieringer, G., Zechner, L., 2005. Rote Liste der Heuschrecken (Orthoptera) Österreich. In: Rote Listen gefahrdeter Tiere Österreichs. Checklisten, Gefahrdungsanalysen, Handlungsbedarf

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The light–dark cycle of Desmoulin’s whorl snail Vertigo moulinsiana Dupuy, 1849 (Gastropoda, Pulmonata, Vertiginidae) and its activity patterns at different temperatures Z. Książkiewicz–Parulska Książkiewicz–Parulska, Z., 2018. The light–dark cycle of Desmoulin’s whorl snail Vertigo moulinsiana Dupuy, 1849 (Gastropoda, Pulmonata, Vertiginidae) and its activity patterns at different temperatures. Animal Biodiversity and Conservation, 41.1: 109–115. Abstract The light–dark cycle of Desmoulin’s whorl snail Vertigo moulinsiana Dupuy, 1849 (Gastropoda, Pulmonata, Vertiginidae) and its activity patterns at different temperatures. Vertigo moulinsiana is a minute land snail species which requires high humidity conditions and is found in wet, temporarily inundated habitats. The species is listed in the IUCN Red List of Threatened Species under the VU (vulnerable) category and is considered a high conservation priority. It is also mentioned in Annex II of the EU Habitat Directive, which imposes the obligation to monitor the species in member countries. The monitoring of V. moulinsiana is based on counting individuals attached to plants in the field, and thus any results may only be properly evaluated when the behavior of the species is understood. Therefore, the aim of this study was to investigate the light–dark cycle of both adults and juveniles within the species as well as to compare activity patterns of both age groups in dark conditions in temperatures of 6 ºC and 21 ºC. Observations were carried out under laboratory conditions, at a high and constant humidity (humidity was at or nearly 100 %). It was shown that juveniles were more active than adults during the day, at night, and at 6 ºC and 21 ºC. In addition, both age groups of V. moulinsiana were more active at 21 ºC than at 6 ºC, and their activity was higher at night than during the day. Such behavior may have an impact on monitoring results based on visual examination and should be taken into account when the data are interpreted. Key words: Land snails, Behavior, Adults, Juveniles, Activity, Conservation Resumen El ciclo de luz y oscuridad del caracol Vertigo moulinsiana Dupuy, 1849 (Gastropoda, Pulmonata, Vertiginidae) y su patrón de actividad a diferentes temperaturas. Vertigo moulinsiana es una especie de diminutos caracoles terrestres que necesitan condiciones de humedad elevada y que se encuentran en hábitats húmedos y temporalmente inundados. La especie figura en la Lista Roja de Especies Amenazadas de la UICN con la categoría VU (vulnerable) y se considera una prioridad alta de conservación. Asimismo, se menciona en el anexo II de la Directiva Hábitat de la Unión Europea, en la que se aconseja a los Estados miembros que hagan un seguimiento de la especie. Como el seguimiento de V. moulinsiana se basa en el conteo de los individuos adheridos a plantas en el campo, los resultados solo se podrán evaluar debidamente cuando se comprenda el comportamiento de la especie. Por consiguiente, el propósito de este estudio es analizar el ciclo de luz y oscuridad de adultos y juveniles, y comparar los patrones de actividad de ambos grupos de edad en condiciones de oscuridad a temperaturas de 6 ºC y 21 ºC. Las observaciones se realizaron en condiciones de laboratorio y con humedad elevada constante (del 100 % o casi). Se observó que los juveniles eran más activos que los adultos durante el día, por la noche y tanto a 6 ºC como a 21 ºC. Además, los dos grupos de edad de V. moulinsiana eran más activos a 21 ºC que a 6 ºC y su actividad era más elevada por la noche que durante el día. Estos resultados indican que el comportamiento puede incidir en los resultados del seguimiento basado en el examen visual y que debería tenerse en cuenta a la hora de interpretar los datos. Palabras clave: Caracoles terrestres, Comportamiento, Adultos, Juveniles, Actividad, Conservación Received: 10 IV 17; Conditional acceptance: 9 VI 17; Final acceptance: 21 VII 17 Zofia Książkiewicz–Parulska, Dept. of General Zoology, Fac. of Biology, Adam Mickiewicz Univ. in Poznań, Umultowska 89, 61–614 Poznań, Poland. E–mail: ksiazkiewicz@amu.edu.pl ISSN: 1578–665 X eISSN: 2014–928 X

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Introduction Desmoulin’s whorl snail Vertigo moulinsiana Dupuy 1849 is an Atlantic–Mediterranean species (Pokryszko, 1990) but it is distributed across most of Europe (Pokryszko, 2003; Killeen et al., 2012). This minute land snail (its shell does not exceed 2.7 mm in height) requires high and constant humidity and inhabits marshes dominated with Reed sweet grass (Glyceria maxima), Common reed (Phragmites australis) and sedges (Pokryszko, 1990; Killeen, 2003; Książkiewicz, 2010). It feeds on microorganisms related to decaying matter and growing on marsh plants (Steusloff, 1937; Bondesen, 1966). The occurrence of V. moulinsiana is strongly associated with local hydrological conditions (e.g. Tattersfield and McInnes, 2003; Książkiewicz et al., 2015) and it occurs in temporarily flooded areas (Killeen, 2003). The importance of the water level, however, seems to be indirect since it affects air humidity, which is important for V. moulinsiana (Killeen, 2003). The snail climbs up tall vegetation (mainly monocots), where it may be found in large numbers in late summer, in autumn and in winter (Killeen, 2003; Książkiewicz–Parulska and Pawlak, 2016). A part of the population, however, is always found within the litter layer (Killeen, 2003; Książkiewicz et al., 2013). V. moulinsiana has been in decline throughout much of its range (Seddon, 1997). This loss of range is evident in Ireland, Germany and France but may be equally severe in other European countries although this cannot be proven due to inadequate monitoring (Killeen et al., 2012). The principal causes of the decline are the physical disturbance of habitats (Pokryszko, 2003) such as wetland drainage or changes in agricultural and land management practices (Killeen, 2003). Thus V. moulinsiana is highly dependent on conservation and listed in the IUCN Red List of Threatened Species in the VU category (vulnerable) (Killeen et al., 2012). Furthermore, it is included in the Annex II of the EU Habitat Directive which imposes an obligation on EU members to designate sites where this species is found as protected areas and to properly manage as well as regularly monitor its populations. The monitoring of V. moulinsiana populations is traditionally based on a visual examination, i.e. counting individuals which are attached to plants (Moorkens and Killeen, 2011; Lipińska et al., 2012). To understand data collected using this method, a good knowledge of the behavior of V. moulinsiana in response to changing environmental conditions is needed. The previous studies on this species, however, have focused mainly on its habitat and microhabitat requirements (Cameron et al., 2003; Killeen, 2003; Tattersfield and McInnes, 2003; Jankowiak and Bernard, 2013; Książkiewicz et al., 2013) or on population dynamics (e.g. Killeen, 2003; Książkiewicz–Parulska and Ablett, 2016). Moreover, data on the activity of Vertigo species in general are scarce (e.g. Boag, 1985), and only a few publications on the behavior of V. moulinsiana in respect to changing temperature conditions have been published so far (Książkiewicz–

Parulska and Pawlak, 2017; Książkiewicz–Parulska, 2017, in press). Thus, this study aims to contribute to the knowledge on the behavior of V. moulinsiana and investigate a light–dark cycle of adults and juveniles of the species as well as activity patterns of these age groups at temperatures of 6 ºC and 21 ºC. Material and methods Individuals used for this study were acquired from a laboratory breeding group which was kept in a room temperature in a 24 hr light–dark cycle (LD 12:12). To make sure that only alive individuals were used in the experiments, I placed individual snails from the breeding container into a plastic box with high–humidity conditions. I selected only those snails which extended their body outside of the shell. Individuals were divided into adults and juveniles based on shell development. The snail was considered an adult when the aperture was fully developed (Pokryszko, 1990). In the experiments, two–whorled juveniles of V. moulinsiana were used. Each snail was placed in a separate 2 ml test tube and individually numbered. The experiments started 24hrs after the snails were put in the test tubes to exclude any false readings that could occur due to distress caused by the relocation of the animal. Each tube had a ca. 0.7 mm hole on the top for oxygenation. Animals were supplied with food, i.e. the decaying leaf of a sedge harvested from a site where V. moulinsiana was present. At the bottom of each tube I placed cotton wool bud saturated with water, but no standing water was present in the tubes. The tubes were placed in a lockable plastic box and were sprinkled with water once a day. Relative humidity inside the tubes was at or nearly 100% and condensation formed on the sides of the tubes. A snail was only regarded as active if it was crawling, or if, although immobile, its body was extended and the tentacles fully everted (Cameron, 1970). Activity of adults and juveniles of Vertigo moulinsiana at 6 ºC and 21 ºC The observations of activity of adults and juveniles of V. moulinsiana were carried out at two different temperatures: at 6 ºC and 21 ºC. The temperature of 6 ºC was maintained in a refrigerated room while the temperature of 21 ºC was in the heated room. Since data on the impact of the temperature on V. moulinsiana are scarce (see Książkiewicz–Parulska, in press), I chose the aforementioned temperatures since I was able to easily maintain these particular conditions at a constant level. The impact of a photoperiod was excluded from the experiment and both temperature groups were stored in dark conditions. Snails were, however, exposed to the light of a flashlight for a short time, twice a day, when the activity of the snails was recorded, e.g. at 11 a.m. and 11 p.m. The experiment in both temperatures was carried out for 32 juveniles and 32 adults of V. moulinsiana. Observations were carried out for 14 days.

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Table 1. Mean activity of snails during experiments: adults and juveniles at 6 ºC and 21 ºC (mean of 28 observations over 14 days) as well as activity of adults and juveniles in light and dark conditions at 21 ºC (mean of 20 observations over 20 days): N, individual number. Tabla 1. Actividad media de los caracoles durante los experimentos: adultos y juveniles a 6 ºC y 21 ºC (media de 28 observaciones durante 14 días) y actividad de los adultos y juveniles en condiciones de luz y oscuridad a 21 ºC (media de 20 observaciones durante 20 días): N, número ide individuos.

Adults

Juveniles

Adults

Juveniles

6 ºC

21ºC

6 ºC

21 ºC

Light

Dark

Light

Dark

0.08

0.04

0.08

0.82

0.05

0.1

0.45

0.2

0.19

0.36

0.19

0.82

0.1

0.05

0.75

0.04

0.07

0.08

0.18

0.1

0.45

0.5

0.7

0.12

0.04

0.12

0.04

0.05

0.25

0.1

0.6

0.12

0.04

0.12

0.46

0.1

0.65

0.45

0.65

0.12

0.07

0.50

0.21

0.15

0.5

0.25

0.85

0.04

0.07

0.08

0.79

0.1

0.4

0.05

0.85

0.08

0.50

0.19

0.36

0.1

0.7

0.2

0.9

0.04

0.14

0.04

0.50

0.1

0.15

0.55

0.12

0.00

0.29

0.05

0.1

0.08

0.36

0.04

0.25

0.05

0.4

0.00

0.25

0.12

0.96

0.05

0.2

0.08

0.57

0.15

0.21

0.3

0.1

0.45

0.05

0.15

0.04

0.27

0.68

0.35

0.1

0.55

0.2

0.19

0.11

0.19

0.54

0.05

0.1

0.25

0.04

0.11

0.15

0.71

0.65

0.55

0.7

0.85

0.04

0.00

0.79

0.3

0.6

0.35

0.95

0.12

0.29

0.04

0.79

0.5

0.55

0.9

0.19

0.54

0.04

0.29

0.15

0.35

0.1

0.45

0.15

0.57

0.19

0.25

0.1

0.45

0.2

0.04

0.43

0.15

0.18

0.2

0.05

0.4

0.00

0.07

0.23

0.11

0.15

0.4

0.6

0.75

0.27

0.00

0.23

0.04

0.25

0.3

0.25

0.12

0.07

0.31

0.18

0.15

0.2

0.12

0.14

0.00

0.25

0.4

0.1

0.55

0.04

0.11

0.08

0.43

0.15

0.1

0.55

0.08

0.21

0.08

0.21

0.15

0.05

0.15

0.19

0.18

0.46

0.18

0.25

0.55

0.15

0.85

0.08

0.25

0.35

0.00

0.2

0.05

0.3

0.08

0.21

0.04

0.46

0.05

0.3

0.15

0.75

0.08

0.39

0.46

0.14

0.35

0.45

0.1

0.08

0.57

0.12

0.11

0.25

The light–dark cycle of adults and juveniles of Vertigo moulinsiana in 21 ºC The observations of the light–dark activity of adults and juveniles of V. moulinsiana were carried out at

21 ºC, in a 24–hr light–dark cycle (LD 12:12). The light conditions were acquired with a daylight lamp AQUAEL Decolight LT (6W LED). The light/ dark shifts had been set for 3 a.m. and 3 p.m., controlled with an automatic light switch. The activity of individuals was

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checked at 11 a.m. in light conditions and at 11 p.m in dark conditions. The snails were exposed to the light of a flashlight in dark conditions for a short time when the snails' activity was recorded, i.e. at 11 p.m. The experiment was carried out for 32 juveniles and 32 adults of V. moulinsiana and observations lasted 20 days. Statistical analyses To compare the activity of Vertigo moulinsiana in the case of (1) adults at 6 ºC and 21 ºC (2) juveniles at 6 ºC and 21 ºC (3) adults and juveniles in 6 ºC (4) adults and juveniles at 21 ºC (5) activity of adults and juveniles in light conditions at 21 ºC (6) activity of adults and juveniles in dark conditions at 21 ºC, I performed a one–way ANOVA test, randomized version using RundomPro 3.14 software. To compare the activity of (1) adults in light conditions and in dark conditions at 21 ºC and (2) juveniles in light conditions and in dark conditions at 21 ºC, I performed Wilcoxon matched– pairs test using Past3 software (I used non–parametric methods because of the lack of normality). The calculations were made for the mean activity of particular individuals (i.e. total number of times that particular individual was active divided by a total number of observations; see table 1). I considered p < 0.05 as the minimum level determining significance. Results Activity of adults and juveniles of Vertigo moulinsiana at 6 ºC and 21ºC The one–way ANOVA showed that both adults and juveniles of V. moulinsiana were more active at 21 ºC than at 6 ºC (adults: F = 9.474, P = 0.003, fig. 1A; juveniles F = 17.180, P < 0.001, fig. 1B). On the other hand, juveniles were more active than adults at 6 ºC (F = 5.367; P = 0.020, fig. 1C) and at 21 ºC (F = 8.109; P = 0.006, fig. 1D).

The light–dark cycle rhythm of adults and juveniles of Vertigo moulinsiana in 21 ºC Both juveniles and adults of V. moulinsiana were more active in dark conditions than in light conditions (juveniles: z = 4.724, P < 0.001, fig. 1E; adults: z = 3.367, P < 0.001, fig. 1F). Furthermore, juveniles of the species were more active than adults in light (F = 4.127, P = 0.042, fig. 1G) and dark conditions (F = 15.240, P < 0.001, fig. 1H). Discussion The activity of terrestrial snails appears to be determined partially by environmental factors (such as moisture and/or temperature), and partially by an internal rhythms (e.g. Wells, 1944; Cook, 2001; Attia, 2004). Circadian activity may be modified by different factors including temperature, wind speed, and the relative humidity of litter moisture (see Cook, 2001). In general, however, it has been shown that many of the terrestrial snails are crepuscular (e.g. the Roman snail H. pomatia Linnaeus 1758, see Stępczak et al., 1982), while others, such as the giant African snail Lissachatina fulica (Bowdich 1822) and the Bush snail Fruticicola fruticum (O. F. Müller 1774), are nocturnal (Butler, 1965; Kuźnik–Kowalska et al., 2013). Furthermore, Baur and Baur (1988) did not find any differences in activity between day and night in the case of the minute Dwarf snail Punctum pygmaeum Draparnaud 1801 when kept at a constant humidity and temperature. These authors suggested that it may be a trait characteristic for snails inhabiting the litter. This conclusion is supported by Boag (1985) who observed some litter dwelling snails, including representatives of the family Vertiginidae, namely Variable vertigo Vertigo gouldi (Binney 1843) and the Cross vertigo Vertigo modesta (Say 1824). Although the study presented here does not give a conclusive result on the endogenous character of the circadian activity of V. moulinsiana, it shows that

Fig. 1. Box plot shows data on mean activity of Vertigo moulinsiana individuals: A, adults in temperatures of 6 ºC and 21 ºC (F = 9.474, P = 0.003); B, juveniles in temperatures of 6 ºC and 21 ºC (F = 17.180, P < 0.001); C, adults and juveniles in temperature of 6 ºC (F = 5.367; P = 0.020); D, adults and juveniles at a temperature of 21 ºC (F = 8.109; P = 0.006); E, juveniles in light and dark conditions at 21ºC (z = 4.724, P < 0.001); F, adults in light and dark conditions at 21 ºC (z = 3.367, P < 0.001); G, juveniles and adults in light conditions at 21 ºC (F = 4.127, P = 0.042); H, juveniles and adults in dark conditions at 21 ºC (F = 15.240, P < 0.001). Middle line, mean; box range, standard error; whiskers, standard deviation. Fig. 1. En los diagramas de caja se muestran los datos relativos a la actividad media de los individuos de Vertigo moulinsiana: A, adultos a temperaturas de 6 ºC y 21 ºC (F = 9,474; P = 0,003); B, juveniles a temperaturas de 6 ºC y 21 ºC (F = 17,180; P < 0,001); C, adultos y juveniles a una temperatura de 6 ºC (F = 5,367; P = 0,020); D, adultos y juveniles a una temperatura de 21 ºC (F = 8,109; P = 0,006); E, juveniles en condiciones de luz y oscuridad a 21 ºC (z = 4,724; P < 0,001); F, adultos en condiciones de luz y oscuridad a 21 ºC (z = 3,367; P < 0,001); G, juveniles y adultos en condiciones de luz a 21 ºC (F = 4,127; P = 0,042); H, juveniles y adultos en condiciones de oscuridad a 21 ºC (F = 15,240; P < 0,001). Línea media, promedio; ancho de caja, error estándar; bigotes, desviación estándar.

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Mean activity

A B 0.57 0.51 0.46 0.40 0.34 0.29 0.23 0.17 0.11 0.06 0.00 6 ºC

0.96 0.87 0.77 0.67 0.56 0.48 0.39 0.29 0.19 0.10 0.00

21 ºC

6 ºC

21 ºC

0.50 0.45 0.40 0.35 0.30 0.25 0.20 0.15 0.10 0.05 0.00

Mean activity

C D

Adults

0.97 0.87 0.77 0.67 0.58 0.48 0.39 0.29 0.19 0.10 0.00

Juveniles

Adults

Juveniles

Light

Dark

1.00 0.90 0.80 0.70 0.60 0.50 0.40 0.30 0.20 0.10 0.00

Mean activity

E F

Light

Dark

0.70 0.63 0.56 0.49 0.42 0.35 0.28 0.21 0.14 0.07 0.00

Mean activity

G H

Adults

Juveniles

1.00 0.90 0.80 0.70 0.60 0.50 0.40 0.30 0.20 0.10 0.00 Adults

Juveniles

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at a constant temperature and in conditions of high humidity, the impact of light and dark is significant for the species. Despite the fact that the life cycle of V. moulinsiana is integrally linked to the litter layer (place of egg laying and where the snail may overwinter; Killeen, 2003) and in some habitats, a significant part of its population may reside within the litter during the whole growing season (see Książkiewicz et al., 2013), V. moulinsiana may be numerously found on plants (e.g. Pokryszko, 1990; Cameron et al., 2003). In this way, individuals of this species are frequently exposed to the sunlight, which might have contributed to the development of a behavior protecting the snail against UV radiation (see Olson and Barbieri, 2014; these authors have shown that shell provides photoprotection and proved the behavioral avoidance of UV radiation in European physa Physa acuta Draparnaud 1805). It may explains why both adults and juveniles were less active (stayed hidden in its shells) in light conditions than in darkness. It should be also considered that such behavior may protect snails against desiccation when exposed to daylight. On the other hand, all experiments (observations of activity patterns in the case of both: light and dark conditions) were carried out in high humidity conditions where desiccation was not an issue. Boag (1985) suggested that small, litter dwelling Vertigo species are most active in temperatures between 6 ºC and 15 ºC, and the proportion of observable snails diminished at higher temperatures, regardless of humidity. The experiments carried out for V. moulinsiana, at the high and constant humidity levels show that the highest activity of the species was noted at 11 ºC, lower activity at 21 ºC (see Książkiewicz–Parulska, in press) and the lowest at 6 ºC. On the other hand, the results presented here suggest that V. moulinsiana shows various responses to the different temperatures, regardless of humidity, as was also concluded by Boag (1985). Comparison of the activity of particular age groups showed that juveniles of V. moulinsiana were more active than adults in the introduced experimental conditions (i.e. when exposed to light and darkness as well as at 6 ºC and 21 ºC). The higher activity of juveniles has previously been observed for some other species of terrestrial snails. For example, juveniles of F. fruticum are more active than adults in the spring, summer and autumn as well as at night, irrespective of the season (Kuźnik–Kowalska et al., 2013). Also, Pollard (1974) concluded that the activity of H. pomatia decreases with age. Due to the fact that juveniles of Vertigo species (similarly to juveniles of other snail species, e.g. Cowie, 1985) are more prone to desiccation than adults (Pokryszko, 1990), it may be suspected that completing growth as soon as it possible is an advantage as it makes them more robust. Thus juveniles usually dwell in a litter (instead of climbing up plants) that ensures high humidity conditions as well as rich and easily accessible source of food (see Cowie, 1985).To confirm this suspicion, however, further studies are needed. Our knowledge on terrestrial snail activity patterns is still scarce while extremely necessary, especially for the conservation of rare and endangered species.

In the case of the V. moulinsiana, the monitoring and determination of conservation activities is based on the visual assessment of population abundance (Moorkens and Killeen, 2011; Lipińska et al., 2012). This is why understanding the behavior of this species in response to changing environmental conditions is essential to properly interpret the data gathered in the field. This study shows that adults and juveniles of V. moulinsiana react differently to diverse temperatures and act differently in light and dark conditions. Such behavior may have an impact on the monitoring results based on visual examination and should be taken into account when the data are interpreted. Acknowledgements The author would like to thank Jon Ablett (Natural History Museum London) for the English correction References Attia, J., 2004. Behavioral rhythms of land snails in the field. Biological Rhythm Research, 35: 35–41. Baur, A., Baur, B., 1988. Individual–movement patterns of the minute land snail Punctum pygmaeum (Draparnaus) (Pulmonata: Endodontidae). The Veliger, 30: 372–376. Boag, D. A., 1985. Microdistribution of three genera of small terrestrial snails (Stylommatophora: Pulmonata). Canadian Journal of Zoology, 63: 1089–1095. Bondesen, P., 1966. Population studies of Vertigo moulinsiana Dupuy in Denmark. Natura Jutlandica, 12: 240–251. Butler, G. D., 1965. Observation on the movement and diurnal activity of the Giant African Snail in Hawaii (Pulmonata: Achatinidae). Proceedings, Hawaiian Entomological Society, XIX: 83–86. Cameron, R. A. D., 1970. The effect of temperature on the activity of three species of helicid snail (Mollusca: Gastropoda). Journal of Zoology, 162: 303–315. Cameron, R. A. D., Colville, B., Falkner, G., Holyoak, G. A., Hornung, E., Killeen, I. J., Moorkens, E. A., Pokryszko, M. B., von Proschwitz, T., Tattersfield, P., Valovirta, I., 2003. Species Accounts for snails of the genus Vertigo listed in Annex II of the Habitats Directive: V. angustior, V. genesii, V. geyeri and V. moulinsiana (Gastropoda, Pulmonata: Vertiginidae). Heldia, 5: 151–170. Cook, A., 2001. Behavioral Ecology: On Doing the Right Thing, in the Right Place at the Right Time. In: The Biology of Terrestrial Molluscs: 447–487 (G. M. Barker, Ed.). CABI Publishing, Wallingford. Cowie, R. H., 1985. Microhabitat choice and high temperature tolerance in the land snail Theba pisana. Journal of Zoology, London (A), 207: 201–211. Jankowiak, A., Bernard, R., 2013. Coexistence or spatial segregation of some Vertigo species (Gastropoda: Vertiginidae) in a Carex rich fen in central Poland. Journal of Conchology, 41: 399–406.

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Killeen, I. J., 2003. Ecology of Desmoulin’s Whorl Snail. Conserving Natura 2000 Rivers Ecology Series 6. English Nature, Peterborough. Killeen, I. J., Moorkens, E., Seddon, M., 2012. Vertigo moulinsiana. The IUCN Red List of Threatened Species 2012: e.T22939A16658400. http://dx.doi.org/10.2305/IUCN.UK.2012–1.RLTS. T22939A16658400.en. Downloaded on 08 February 2017. Książkiewicz, Z., 2010. Higrofilne gatunki poczwarówek północno–zachodniej Polski. Wydawnictwo Klubu Przyrodników, Świebodzin. Książkiewicz, Z., Biereżnoj–Bazille, U., Krajewski, Ł., Gołdyn, B., 2015. New records of Vertigo geyeri Lindholm, 1925, V. moulinsiana (Dupuy, 1849) and V. angustior Jeffreys, 1830 (Gastropoda: Pulmonata: Vertiginidae) in Poland. Folia Malacologica, 23: 121–136. Książkiewicz, Z., Kiaszewicz, K., Gołdyn, B., 2013. Microhabitat requirements of five rare vertiginid species (Gastropoda, Pulmonata, Vertiginidae) in wetlands of western Poland. Malacologia, 56: 95–106. Książkiewicz–Parulska, Z., in press. The impact of temperature on activity patterns of two Vertiginid micro–molluscs (Mollusca: Gastropoda) in conditions of high, constant humidity. American Malacological Bulletin. Książkiewicz–Parulska, Z., Ablett, J. D., 2016. Investigating the influence of habitat type and weather conditions on the population dynamics of land snails Vertigo angustior Jeffreys, 1830 and Vertigo moulinsiana (Dupuy, 1849). A case study from western Poland. Journal of Natural History, 50: 1749–1758. Książkiewicz–Parulska, Z, Pawlak, K., 2016. Rare species of micromolluscs in the city of Poznań (W. Poland) with some notes on wintering of Vertigo moulinsiana (Dupuy, 1849). Folia Malacologica, 24: 97–101. – 2017. The influence of temperature on the hibernation patterns and activity of Vertigo moulinsiana (Dupuy, 1849) (Gastropoda: Pulmonata: Vertiginidae). Turkish Journal of Zoology, 41: 370–374. Kuźnik–Kowalska, E., Lewandowska, M., Pokryszko, B. M., Proćków, M., 2013. Reproduction, growth and circadian activity of the snail Bradybaena fruticum (O. F. Müller, 1774) (Gastropoda: Pulmonata:

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Bradybaenidae) in the laboratory. Central European Journal of Biology, 8: 693–700. Lipińska, A., Książkiewicz, Z., Zając, K., Barga–Więcławska, J., 2012. Poczwarówka jajowata Vertigo moulinsiana (Dupuy, 1849). In: Monitoring gatunków zwierząt. Przewodnik metodyczny: 463–481 (M. Makomaska–Juchiewicz, P. Baran, Eds.). GIOŚ, Warszawa. Moorkens, E. A., Killeen, I. J., 2011. Monitoring and Condition Assessment of Populations of Vertigo geyeri, Vertigo angustior and Vertigo moulinsiana in Ireland. Irish Wildlife Manuals, No. 55. National Parks and Wildlife Service, Department of Arts, Heritage and Gaeltacht, Dublin, Ireland. Olson, M. H., Barbieri, N. E., 2014. Mechanisms of ultraviolet radiation tolerance in the freshwater snail Physa acuta. Freshwater Science, 33: 66–72. Pokryszko, B. M., 1990. The Vertiginidae of Poland (Gastropoda: Pulmonata: Vertiginidae) – a systematic monograph. Annales Zoologici, 43: 133–257. – 2003. Vertigo in Continental Europe– autecology, threats and conservation status (Gastropoda, Pulmonata: Vertiginidae). Heldia, 5: 13–25. Pollard, E., 1974. Aspects of the ecology of Helix pomatia L. Journal of Animal Ecology, 44: 305–329. Seddon, M. B., 1997. Distribution of Vertigo moulinsiana (Dupuy, 1849) in Europe. In: Vertigo moulinsiana: Surveys and studies commissioned in 1995–96: 56–68 (C. M. Drake, Ed.). English Nature Research Report 217. Peterborough, England. Stępczak, K., Ławniczak, H., Wieland, A., 1982. Dobowa i sezonowa aktywność ślimaków winniczków (Helix pomatia L.) – biologia, morfologia muszli, problemy przechowalnictwa i transportu w punktach skupu i bazach eksportowych. Prace Komisji Biologicznej PTPN, 66: 13–35. Steusloff, U., 1937. Beitraege zur Molluskenfauna der Niederrheingebietes: Lebensraum und Ernaehrung von Vertigo moulinsiana in Mitteleuropa. Decheniana, 94: 30–46. Tattersfield, P., McInnes, R., 2003. Hydrological requirements of Vertigo moulinsiana on three candida the Special Areas of Conservation in England (Gastropoda, Pulmonata: Vertiginidae). Heldia, 5: 135–147. Wells, G. P., 1944. The water relations of snails and slugs III. Factors determining activity in Helix pomatia L. Journal of Experimental Biology, 20: 79–87.

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Brief communication 117

Local habitat disturbance increases bird nest predation in the Brazilian Atlantic rainforest V. B. Rodrigues, F. M. Jesus, R. I. Campos

Rodrigues, V. B., Jesus, F. M., Campos, R. I., 2018. Local habitat disturbance increases bird nest predation in the Brazilian Atlantic rainforest. Animal Biodiversity and Conservation, 41.1: 117–120. Abstract Local habitat disturbance increases bird nest predation in the Brazilian Atlantic rainforest. We evaluated the effect of anthropogenic disturbance on nest predation in Brazilian Atlantic forest. Artificial nests were distributed in fragments with distinct degrees of anthropogenic disturbance. We found a higher proportion of egg predation on the ground and in the fragments classified as 'high' and 'medium' disturbance than in the fragments classified as 'low' degree of disturbance. The higher egg predation is probably linked to low structural complexity of vegetation and high accessibility of these areas to opportunistic predators. We suggest that forest fragments with high vegetation complexity and low human activity should be preserved in order to maintain the biodiversity of bird species. Key words: Artificial nest, Forest fragments, Habitat conservation, Vegetation structure Resumen La perturbación del hábitat local aumenta la depredación de nidos de aves en la pluviselva atlántica del Brasil. Evaluamos los efectos de la perturbación antropógena en la depredación de nidos en el bosque atlántico del Brasil. Se distribuyeron nidos artificiales en fragmentos con distintos grados de perturbación antropógena. Observamos una mayor proporción de depredación de huevos en el suelo y en los fragmentos clasificados como de perturbación alta y media que en los fragmentos con un bajo grado de perturbación. La mayor depredación de huevos está probablemente relacionada con una vegetación de complejidad estructural baja y con la elevada accesibilidad de estas zonas para depredadores oportunistas. Con vistas a mantener la biodiversidad de especies de aves, proponemos que se conserven los fragmentos forestales con vegetación de complejidad estructural elevada y escasa actividad humana. Palabras clave: Nido artificial, Fragmentos de bosque, Conservación del hábitat, Estructura de la vegetación Received: 2 III 17; Conditional acceptance: 17 V 17; Final acceptance: 26 VII 17 Vinícius B. Rodrigues, Programa de Pós–Graduação em Entomologia, DBE, UFV, Viçosa–MG, Brazil.– Fabiene M. Jesus, Ricardo I. Campos, Programa de Pós–Graduação em Ecologia, DBG, UFV, Viçosa–MG, Brazil.– Ricardo I. Campos, Depto. de Biologia General, UFV, Viçosa–MG, Brazil. Corresponding author: Vinícius B. Rodrigues. E–mail: viniciusbrbio@gmail.com

Introduction The Brazilian Atlantic Forest is considered one of the most important areas in the world for species conservation (Ribeiro et al., 2009). In the case of birds, this biome has the highest conservation priority in the ISSN: 1578–665 X eISSN: 2014–928 X

country once it harbors 75.6 % of threatened and endemic bird species of Brazil (Aleixo, 2001). Despite its importance, only 12 % of the original Atlantic rainforest area remains intact and it continues to be disturbed by human activity, such as cattle grazing, monocultures, and urban expansion (Ribeiro et al., 2009). © 2018 Museu de Ciències Naturals de Barcelona

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On a landscape scale, habitat fragmentation has extremely negative effects on biodiversity, even for groups with notorious dispersal abilities, such as birds (Wang et al., 2015). The loss and isolation of natural habitats promoted by fragmentation alters bird movements and consequently decreases gene flow. Previous studies have shown that fragmented landscapes have high rates of bird nest predation and it is probably related to the increased abundance of generalist predators in smaller fragments (Melo and Marini, 1997). Nest predation is considered a major cause of change in abundance, richness, and species composition in bird communities (Evans, 2004). However, the effect of local habitat disturbance on bird nest predation is still poorly studied, especially in the neotropics (but see Borges and Marini, 2010). We consider local disturbance as any human influence that modifies local habitat conditions. For instance, a change in vegetation structure (e.g. a decrease in foliage cover and/or tree density) and a reduction in distance from areas of intense human activity can increase bird nest vulnerability to predators (Borges and Marini, 2010; Nana et al., 2015). Increased predation, in turn, can also affect the distribution of nest sites among vertical strata (Silvia and Voltolini, 2015). A couple of studies have shown that bird nests on the ground were more vulnerable to predators than nest in low vegetation (Castro–Caro et al., 2014). However, we have not found any studies evaluating the effect of local habitat disturbance and vertical stratification on bird nest predation in Atlantic rainforest. In the present study we experimentally evaluated the effect of habitat strata (ground and low vegetation) on artificial nest predation in three disturbed Atlantic Forest fragments inside the Parque Nacional do Caparaó (PNC). We tested two hypotheses: i) artificial nest predation is higher in habitats with increased disturbance and ii) predation on artificial nests is higher on the ground than in low vegetation strata. It is important to highlight that an artificial nest will not present the same visual and olfactory signals as a natural nest. Artificial nests, for example, do not include any parental activities and they might present differences in egg size and exposure (Mezquida and Marone, 2003). Although they have a certain degree of bias, quail eggs can be considered a good model owing to their average size and color matching that of other common species of forest birds that lay eggs in open nests (Marini and Melo, 1998), but all potential predators, both introduced and native, can break the quail egg shells (Marini and Melo, 1998). Artificial nests, on the other hand, can provide key information (such as probabilities of predation) that would otherwise be very difficult to obtain in natural conditions (Nana et al., 2015). Material and methods The study was conducted in the Parque Nacional do Caparaó (PNC), one of the largest reserves of Atlantic Forest (32,000 ha) in southeastern Brazil along the states of Minas Gerais and Espírito Santo (20º 19′–20º

37' S, 41º 43'–41º 53′ W). The field sampling took place in December 2004. Within the PNC, we selected three forest fragments based on distance from the intensive use area (administrative and touristic centers) and degrees of anthropogenic disturbance (table 1). Based on their past logging history we classified the three fragments according to their degree of disturbance as follows: low (selective logged at least 100 years ago); medium (selective logged at least 30 years ago) and high (completely logged at least 30 years ago). All three forest fragments were around 20 ha in size and are surrounded by roads and other forest fragments. See table 1 for vegetation description. We handmade 138 artificial nests from filter paper and vegetation materials. We placed a single quail egg (Coturnix coturnix L.) in each nest. We chose quail eggs because they are commonly used model in experiments of artificial nest predation (Marini and Melo, 1998). Throughout the experiment, we used rubber boots and gloves to minimize the effect of human scent. Inside each of the three forest fragments (at least 500 m far from the border), we placed 46 artificial nests with a quail egg approximately 3 m from each other, alternating one nest on the ground and one on a tree branch (about 1.8 m high). All the nests were left in the field for four days and after that they were checked for prey marks. When pierced, fragmented, destroyed or missing, the egg was considered as preyed upon. To determine the effects of the categorical explanatory variables 'forest fragment' (with three levels of disturbance: low, medium and high) and 'habitat strata' (ground and trees) on the response variable 'proportion of eggs prayed' (calculated from a binomial variable: 0 for non–preyed and 1 for preyed eggs), we first built a Generalized Linear Mixed Model (GLMM) (Crawley, 2012). For this model, to deal with our low number of real replicates (one fragment per treatment) the explanatory variable 'forest fragment' was included as a random effect and 'habitat strata' as a fixed factor. To analyze the differences between each pair of disturbance levels (low, medium and high) and 'habitat strata' (ground and trees) we used model selection. This procedure was followed using Akaike’s information criterion with second order bias correction (AICc) due to our limitations in sample size (Burnham et al., 2011). In this procedure we first built a global model and then compared this model with all candidate variables. Finally, we selected the best model based on AICc weights (Wagenmakers and Farrell, 2004). All analyses were performed in R using the package MuMIn. Results From the 138 eggs placed in artificial nests, 35 were preyed upon, 2 in a low degree of disturbance, 15 in a medium degree, and 18 eggs in a high degree. Regarding the habitat strata, 21 eggs were preyed upon on the ground and 14 on tree branches. Corroborating our first hypothesis, our analysis showed that a higher proportion of eggs were preyed upon in fragments classified under a 'high' and 'medium' degree of disturbance (AICc = 147.4; ∆ = 0.00; weight = 0.748)

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Table. 1. Degree of disturbance (Dd), description of vegetation (visual estimate), distance to the intensive use area (D, in m, administrative and touristic centers), human activity (H, based on interviews with five rangers) and geographic coordinates from the three forest fragments sampled inside the PNC. Tabla. 1. Grado de perturbación (Dd), descripción de la vegetación (estimación visual), distancia a la zona de uso intensivo (D, en m, centros administrativos y turísticos), actividad humana (H, basada en entrevistas con cinco guardabosques) y coordenadas geográficas de los tres fragmentos de bosques muestreados dentro del Parque Nacional de Caparaó. Dd

Vegetation

Low

Canopy approximately 30 m high

Coordinates

620

Rare

20º 25' 7.28 4" S

with 90 % of foliage cover

41º 50' 45.74 4" W

Medium

Canopy approximately 25 m high

20º 25' 8.83 2" S

340

Low

with 80 % of foliage cover.

High

Canopy approximately 20 m high

with 70 % of foliage cover

when compared to the 'low' degree (AICc = 149.6; ∆ = 2.18; weight = 0.252) (fig. 1). As expected, we also found that the proportion of preyed eggs was higher on the ground than in trees (AICc = 149.6; ∆ = 0.00; weight = 0.503). Discussion We found that artificial nest predation was greater in more disturbed forest fragments. Despite the limita-

Mean proportion of eggs preyed upon

1.0 0.8

41º 50' 55.42 8" W

High

20º 25' 13.512" S 41º 51' 7.92" W

tions of the artificial nests (Mezquida and Marone, 2003), our result supports the hypothesis that habitat degradation has a negative effect on reproductive success of birds (e.g. Borges and Marini, 2010). The higher predation observed in the disturbed fragments here might be explained by three major factors. First, the smaller tree density and foliage cover found in highly disturbed area (table 1) might enhance the transmission of acoustic, chemical, or visual signals from nests to predators, increasing the predation rate (Hazler et al., 2006). Second, areas with a high

Ground Understory

B n.s.

0.6

B n.s.

0.4 0.2 0.0

A n.s. Low

Medium Degrees of disturbance

High

Fig. 1. Mean (+ SE) proportion of eggs preyed upon per fragment and habitat strata (ground and trees), considering non–preyed eggs as 0 and preyed upon eggs as 1. Fig. 1. Proporción media (+ DE) de huevos depredados por fragmento y estrato del hábitat (suelo y árboles), considerando los huevos no depredados como 0 y los depredados, como 1.

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degree of disturbance might have greater accessibility to opportunistic and small predators, such as rodents, lizards, and small mammals (Melo and Marini, 1997). Finally, humans live and work in the park facilities, close to the study area classified as 'high degree of disturbance' and therefore we can speculate that the presence of humans avoids large predators, allowing an increase in the activity of small wild predators, or at least of dogs and feral cats. We also corroborate our second hypothesis as we found that egg predation was higher on the ground than in trees. The high predation rates on the ground can be related to the greater abundance of terrestrial than arboreal predators (Castro–Caro et al., 2014). It is also plausible that predation of nests by non–avian species, such as rodents and lizards, may increase in fragments near urban areas, as in our forest fragments. Two studies have previously reported that bird nests were more vulnerable to predators on the ground than in low vegetation (Castro–Caro et al., 2014; Silvia and Voltolini, 2015). However, this is the first study that evaluates the effect of local habitat disturbance and vertical stratification on bird nest predation in Atlantic rainforest. We showed here that local habitat disturbance can increase bird nest predation. Although our findings do not differ from previous studies (Melo and Marini, 1997; Hazler et al., 2006; Borges and Marini, 2010) they should be interpreted with caution due to our low number of true replicates. However, we consider our analytical approach robust enough to remedy this methodological limitation. Since nest predation is considered a key factor affecting bird richness, abundance, and distribution (Evans, 2004), the present study might be considered as a fast and low–cost alternative to indicate high–priority areas for bird species conservation in forest fragments. Based on our results, we suggest that forest fragments with high structural complexity and low human activity should be preserved in order to maintain bird species biodiversity (Nana et al., 2015). Acknowledgements We are grateful to 'Parque Nacional do Caparaó' for their support during the fieldwork. This work was supported by fellowships and grants awarded by the Ecology graduate program (PPG–Eco) at the Federal University of Viçosa and the following Brazilian scientific promoting agencies: CAPES, CNPq, and FAPEMIG. References Aleixo, A., 2001. Conservação da avifauna da Floresta Atlântica: efeitos da fragmentação e a importância de florestas secundárias. In: Ornitologia e Conser-

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vação: da ciência às estratégias: 199–206 (J. L. B. Albuquerque, J. F. Cândido Jr., F. C. Straube, A. L. Roos, Eds.). Sociedade Brasileira de Ornitologia, Editora Unisul, Tubarão. Borges, F. J. A., Marini, M. A., 2010. Birds nesting survival in disturbed and protected Neotropical savannas. Biodiversity and Conservation, 19: 223–236. Burnham, K. P., Anderson, D. R., Huyvaert, K. P., 2011. AIC model selection and multimodel inference in behavioral ecology: some background, observations, and comparisons. Behavioral Ecology and Sociobiology, 65: 23–35. Crawley, M., 2012. The R Book. 2nd. Ed. John Wiley & Sons Ltd, London, UK. Castro–Caro, J. C., Carpio, A. J., Tortosa, F. S., 2014. Herbaceous ground cover reduces nest predation in olive groves. Bird Study, 61: 537–543. Evans, K. L., 2004. The potential for interaction between predation and habitat change to cause population declines of farmland birds. Ibis, 146: 1–13. Hazler, K. R., Amacher, A. J., Lancia, R. A., Gerwin, J. A., 2006. Factors influencing Acadian flycatcher nesting success in an intensively managed forest landscape. Journal of Wildlife Management, 70: 532–538. Marini, M. A., Melo, C., 1998. Predators of quail eggs, and the evidence of the remains: implications for nest predation studies. Condor, 100: 395–399. Melo, C., Marini, M. A., 1997. Predação de ninhos artificiais em fragmentos de Matas do Brasil Central. Ornitología Neotropical, 8: 7–14. Mezquida, E. T., Marone, L., 2003. Are results of artificial nest experiments a valid indicator of success of natural nests? Wilson Bulletin, 115: 270–276. Nana, D., Sedláček, O., Doležal, J., Dančák, M., Altman, J., Svoboda, M., Majeský, L., Hořák, D., 2015. Relationship between survival rate of avian artificial nests and forest vegetation structure along a tropical altitudinal gradient on Mount Cameroon. Biotropica, 47: 758–764. Ribeiro, M. C., Metzger, J. P., Martensen, A. C., Ponzoni, F. J., Hirota, M. M., 2009. The Brazilian Atlantic Forest: How much is left, and how is the remaining forest distributed? Implications for conservation. Biological Conservation, 142: 1141–1153. Silvia, A. S. A., Voltolini, J. C., 2015. Predação de ninhos artificiais em dois fragmentos urbanos de Mata Atlântica no Sudeste do Brasil. Revista Biociências, 21: 29–37. Wagenmakers, E.–J., Farrell, S., 2014. AIC model selection using Akaike weights. Psychonomic Bulletin and Review, 11(1): 192–196. Wang, Y., Thornton, D. H., Ge, D., Wang, S., Ding, P., 2015. Ecological correlates of vulnerability to fragmentation in forest birds on inundated subtropical land–bridge islands. Biological Conservation, 191: 251–257.

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Effect of temperature and type of diet on the metamorphosis of Pleurodema thaul (Lesson, 1826) in a population of south–central Chile H. Díaz–Páez, C. Canales–Arévalo

Díaz–Páez, H., Canales–Arévalo, C., 2018. Effect of temperature and type of diet on the metamorphosis of Pleurodema thaul (Lesson, 1826) in a population of south–central Chile. Animal Biodiversity and Conservation, 41.1: 121–130. Abstract Effect of temperature and type of diet on the metamorphosis of Pleurodema thaul (Lesson, 1826) in a population of south–central Chile. The effect of the environment on the vital cycle of amphibians has been shown in diverse studies, indicating that diet and temperature affect the duration of the larval period, and the size of the newly metamorphosed. We analyzed the effect of temperature and quality of diet on the duration of the larval period and size reached by larvae up until metamorphosis in the species Pleurodema thaul. We used an experimental design with two temperatures (15 ºC and 25 ºC) and two types of diet, one rich in proteins (RP) and one low in proteins (LP). We evaluated body size (cm) and body mass (g), and staged the larval development according to Gosner (1960). Our results indicate that temperature is crucial for the larval development, affecting its duration, whereas diet has a secondary effect on size and mass of larvae, always depending on the temperature of development. Key words: Anuran, Diet, Metamorphosis, Temperature Resumen Efecto de la temperatura y el tipo de alimentación en la metamorfosis de Pleurodema thaul (Lesson, 1826) en una población del centro sur de Chile. El efecto del ambiente en el ciclo vital de los anfibios ha sido puesto de manifiesto en diversos estudios. Estos han indicado que tanto la alimentación como la temperatura afectan a la duración de las etapas larvales y al tamaño de los recién metamorfoseados. En el presente estudio se analizó el efecto de la temperatura y la calidad de la alimentación en la duración del periodo larval y el tamaño alcanzado por las larvas hasta la metamorfosis en Pleurodema thaul. Para ello se utilizó un diseño experimental con dos temperaturas (15 ºC y 25 ºC) y dos tipos de alimentación, una con un alto contenido proteico (RP) y otra con bajo (LP). Se evaluó el tamaño (cm) y la masa corporal (g), y se identificó el estadio larval de acuerdo con Gosner (1960). Los resultados indican que el efecto de la temperatura resulta crucial para el desarrollo de las larvas, ya que afecta a la duración del mismo, mientras que la alimentación ejerce un efecto secundario en el tamaño y la masa de las larvas, dependiendo siempre de la temperatura de crianza. Palabras clave: Anuro, Alimentación, Metamorfosis, Temperatura Received: 21 VII 16; Conditional acceptance: 22 IX 16; Final acceptance: 27 VII 17 Helen Díaz–Páez, Carla Canales–Arévalo, Depto. de Ciencias Básicas, Campus Los Ángeles, Univ. de Concepción, Juan Antonio Coloma 0201, Los Ángeles (región del Biobío), Chile. Corresponding author: hediaz@udec.cl

ISSN: 1578–665 X eISSN: 2014–928 X

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Introduction

Material and methods

Amphibians are the group of vertebrates whose complex life cycle involves sequential aquatic and terrestrial stages. To survive in these diverse environments they have particular physiological mechanisms (Hickman et al., 2002). In ecological terms, they undergo high mortality during their aquatic phase, basically due to predation and desiccation of the temporary ponds where they breed (Wilbur, 1980; Newman, 1988). Hence, fast development or shortening of the larval phase may be beneficial for their subsistence (Newman, 1989). Conditions of larval development thus create a critical period for survival, where the need to reach certain stages of development may be fundamental to the persistence of the newly metamorphosed (Arnold and Wassersug, 1978; Wilbur, 1980). However, a delay in metamorphosis may also have a positive effect, as it can result in a larger body size, giving advantages to individuals in the transition to land. It has been reported that large metamorphs may have a greater ability to resist starvation and desiccation, as well as better abilities to escape predators (Tracy et al., 1993; Semlitsch, 1993). This delay in metamorphosis, therefore, could impose a trade–off between pre– and post–metamorphic survival. Environment and especially diet and temperature are vital factors for both development and metamorphosis of the larvae (Naya et al., 2008). Different studies have shown that larvae of anurans exhibit a high phenotypic plasticity in response to variations of temperature and food, which is expressed in aspects such as growth and rates of development (Álvarez and Nicieza, 2002a, 2002b; Benavides, 2003; Benavides et al., 2005). Temperature appears to be the environmental variable with the most pervasive effect on this taxon because it can be a substantial source of physiological stress on tadpoles and may cause selection pressure, favoring adaptive evolution in thermal tolerance and sensitivity (Tejedo et al., 2012). On the other hand, temperature has shown to affect developmental timing, reducing larval stage at higher temperatures (Atkinson, 1996). Diet quality also appears to play a role in this taxon. Álvarez and Nicieza (2002a, 2002b) reported that it affects both duration and size reached by newly metamorphosed individuals, establishing that diets rich in nutrients generate larger organisms. Amphibians represent a small fraction of the endemic fauna in Chile. Interestingly, although they are the vertebrates with least diversity, they exhibit the highest degree of endemism (Spotorno, 1995; Vidal et al., 2009). One of the most abundant species of amphibians in Chile is Pleurodema thaul (Cei, 1962), a small from that lives in multiple climates (Iturra–Cid, 2007). Its life cycle has been described by Díaz–Páez and Ortiz (2001), who determined that the species breeds in winter–spring and that larvae remain in temporary pools until their metamorphosis. However, information on larval growth and development and the environmental variables affecting it are lacking. The present work thus aims to analyze the effect of temperature and the type of diet on larval growth and developmental rates in P. thaul.

In August 2014, we collected five egg clutches of Pleurodema thaul from a temporary pool in the Commune of Santa Barbara (37 º 36' 42.1'' S–72 º 07' 41.6'' W) in the Biobío Region, Chile. The clutches were taken to the Herpetological Ecophysiology Laboratory at the University of Concepción, Los Angeles Campus where they were separated in aquariums provided with 2 L of dechlorinated of water and equipped with an aeration pump. The aquariums were maintained at a temperature of 20 ºC with a 12/12 hrs light/darkness photoperiod. Water was changed each week. After hatching, larvae were maintained in these conditions until they reached Gosner stage 25 (Gosner, 1960). A 2 x 2 factorial design was used to analyze the effect of the type of food, with two treatments: low protein content (LP) and high protein content (RP). Similarly, the effect of the temperature was analyzed, selecting two treatments representing the temperature extremes of the optimal range for larval development, 15 ºC (± 1 ºC), and 25°C (± 1 ºC) (Alvarez and Nicieza, 2002a, 2002b; Benavides, 2003; Benavides et al., 2005; Sanuy et al., 2008). In accordance with Benavides (2003), we used boiled lettuce in the diet of low protein content (1.3 % protein and 0.3 % fat per 100 g of lettuce, Granval and Gaviola, 1991); and for the rich diet, we used a commercial product of the Micron brand because of its high protein content (54.7 % protein and 2.6 % fat per 100 g of Micron). Larvae were fed ad libitum every two days and as their development progressed, the amount of food was increased in order to meet their growing demands. Excess food was removed along with water changes. A total of 180 larvae were randomly placed in each of the four experimental treatments (n = 15 per treatment) and three replicas per each treatment were established. Each recipient maintained two l of water in recirculation through an air pump, which helped to reduce thermal heterogeneity. To identify the different larval stages, we followed Gosner´s table of stages in Duellman and Trueb, 1986). Containers were checked every day to measure the larval stage, size (± 0.1 cm) and mass (± 0.1 mg) until full larval tail resorption, that is, Gosner stage 46. Four size classes were selected for further analyses: initial larval (stage 25–30), larval (stage 31–35), metamorphic (stage (36–40), and climax metamorphic (stage 41–46) (fig. 1). To estimate the averages of size, mass and stage per aquariums in the different treatments, we performed multiple variance analysis (Manova) and non–parametric tests using SPSS version 13.0 software (Blair and Taylor, 2008). Results We found a significant effect of temperature on the mass of the larvae (F1, 90 = 2.822; P < 0.001) and on their body size (F1, 35 = 4.410; P < 0.001). Larvae exposed to high temperatures showed faster development and growth, reaching greater masses (fig. 2A) and sizes (fig. 2B) in less time. In contrast, larvae exposed

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10 mm

Fig. 1. Larval stages of development in Pleurodema thaul: A, stage 25, no budding of hind limbs; B, detail of stage 30 with the appearance of buds that give rise to the hind limbs; C, stage 34 with the appearance of the second and third fingers attached; D, stage 41 with the protuberances of the forelimbs and marked disappearance of anal fold; E, stage 45, almost complete metamorphosis with much of the tail reabsorbed; F, complete metamorphosis, stage 46. Fig. 1. Estados de desarrollo larval en Pleuroderma thaul: A, estadio 25 sin brotes de extremidades posteriores; B, detalle del estadio 30 con la aparición de los brotes que darán origen a las extremidades posteriores; C, estadio 34 con la aparición del segundo y tercer dedo unidos; D, estadio 41 con las protuberancias de los miembros anteriores y la marcada desaparición del pliegue anal; E, estadio 45, metamorfosis casi completa con gran parte de la cola reabsorbida; F, metamorfosis completa, estadio 46.

to low temperatures showed prolonged duration in development (fig. 2). We observed no effect of diet on the mass of the larvae (F1, 90 = 1.210; P = 0.095). However, we found that diet had a noticeable effect on size (F1, 35 = 1.634; P = 0.012), so that larvae fed RP reached larger sizes than those fed LP (table 1). Similarly, temperature was the variable that most affected larvae mass and body size, with a significant effect on both variables (F1, 380 = 1.429; P < 0.001).

This indicates that temperature is a crucial factor on size and mass of P. thaul in its larval development, minimizing the impact of diet on these parameters in the development of this species P. thaul (F1, 380 = 0.808; P = 0.993) (table 2). Larvae increased in both size and mass in all treatments (fig. 2). Temperature had a strong effect on the duration of the larval stage (F4, 1201 = 29.47; P = 0.003), causing larvae exposed to 25 ºC to have

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1,000

Diet

800 600

25 Temperature (ºC)

Mass (mg)

400 200

0 1,000 800 600

400 200 0

0 1 2 3 4 5 6 7 8 9 10 11 12 13 0 1 2 3 4 5 6 7 8 9 10 11 12 13 Time (weeks) LP

5 4 25

Temperature (ºC)

Total size (cm)

1 5 4

3 2 1 0 1 2 3 4 5 6 7 8 9 10 11 12 13 0 1 2 3 4 5 6 7 8 9 10 11 12 13 Time (weeks) Fig. 2. Effect of temperature and type of diet on: A, larval size; and B, larval mass.

Fig. 2. Efecto de la temperatura y el tipo de alimentación en: A, el tamaño larval; y B, la masa larval.

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Table 1. Data for mass (M, in mg) and body size (Ts, total size, in cm) in tadpoles of Pleurodema thaul for each stage grouping (Li, larval initial; L, larval; Mt, metamorphic; Cm, climax metamorphic), in two temperature (15 ºC and 25 ºC) and two diet (LP, low protein content; RP, high protein content) treatments (values are given as mean ± SE). Analyses from Mann–Whitney test of differences between the diet treatments at each stage group are given (Z: NS, not significant; * P < 0.05; ** P < 0.005; *** P < 0.001). Analyses from Kruskal–Wallis H tests comparing the different stage group within treatment are presented as the x2–value. Tabla 1. Datos relativos a la masa y (M, en mg) y el tamaño corporal (Ts, tamaño total, en cm) de los renacuajos de Pleurodema thaul agrupados por estado de desarrollo (Li, larval inicial; L, larval; Mt, metamórfico; Cm, clímax metamórfico) en dos tratamientos de temperatura (15 ºC y 25 ºC) y dos de alimentación (LP, contenido porteico bajo; RP, contenido proteico alto) (los valores son indicados como media ± EE). Se utiliza el test de Mann–Whitney para analizar las diferencias entre los tratamientos de alimentación en cada grupo de estados de desarrollo (Z: NS, no significativo; * P < 0,05; ** P < 0,005; *** P < 0,001). Se utiliza el test de Kruskal–Wallis H para comparar los diferentes grupos de estado dentro de los tratamientos, los valores se representan como valor de x2.

Development stage group

Treatment (ºC)

Li (25–30)

L (31–35)

Mt (36–40)

Cm (41–46)

1.89 ± 0.48

3.11 ± 0.39

4.20 ± 0.30

–

270.00 ***

1.99 ± 0.65

3.53 ± 0.42

3.98 ± 0.22

–

470.30 ***

Z M

LP RP

Z Ts

25 25

93.47 ± 65.40

–8.355 ***

–1.355 NS –

339.04 ± 144.90 833.33 ± 145.72

128.16 ± 106.50 479.45 ± 117.36

667.86 ± 130.28

–

195.77 ***

–

465.88 ***

–2.09 –

13.49 ***

–8.29 ***

1.67 ± 0.53

2.86 ± 0.33

3.38 ± 0.28

2.62 ± 0.99

122.96 ***

1.77 ± 0.58

2.88 ± 0.31

3.51 ± 0.33

3.11 ± 0.78

146.08 ***

–2.70 *

–0.84

–3.57 ***

–1.93 *

107.86 ± 95.50

288.85 ± 82.33

516.75 ± 124.64 383.33 ± 115.00 131.15 ***

77.19 ± 66.69

296.57 ± 95.42

563.18 ± 185.26 493.00 ± 114.53 146.58 ***

Z M

NS –1.292

–2.01 *

NS NS 3.13 –0.44 –1.08 NS

shorter times of development than those exposed to 15 ºC. Complete tail reabsorption concluded in five weeks (fig. 1) in larvae exposed to 25 ºC but at 13 weeks in larvae exposed to 15 ºC (fig. 3). Additionally, mortality was higher at the low temperature (90%), and a high percentage of larvae (X = 6.83 ± 3.06) in this 15 ºC treatment group did not complete metamorphosis. To better understanding the treatment variation with larval age we grouped data into four age groups or developmental stage. The results showed that at the lower temperature, development was not affected by diet (table 1), so that the larvae did not increase in size due to diet. The effect was concentrated in the masses during Gosner stages 25 to 35, demonstrating a notable increase in the mass for the larvae fed RP. On the other hand, at higher temperatures (25 ºC), the effects were reverted, so that RP generated significantly larger larvae than those fed LP, while the mass variation was not significant (table 1).

Finally, we found that larvae in the RP/15 ºC treatment reached the largest masses and body sizes (table 1) during all stages of development. Conversely, larvae in the LP/25 ºC group reached the smallest mass and body mass (table 1). Discussion Analysis of ecological physiology in amphibians is essential to understand many aspects of the biology of these organisms. This includes the conditions that define their fundamental niches, geographical distribution and evolutionary dynamics which in turn allow to determine their vulnerability to climate change (Gutiérrez–Pesquera et al., 2016). At the same time, the thermal environment exerts a strong effect on aspects of life history in ectotherms, playing an important role in their growth rate and body size (Angilletta et al., 2004; Angilletta, 2009).

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Table 2. Two–way ANOVA for effect of temperature and diet on larval development: M, mass of tadpole; Ts, body size of tadpole; Diet, diet; Temp, temperature of water in aquarium; Type III MS, type III of sum of squares: (a) R2 = 0.386 (adjusted R2 = 0.115); (b) R2 = 0.467 (adjusted R2 = 0.232); Df, degrees of freedom. (NS, not significant; * P < 0.05; ** P < 0.005; *** P < 0.001). Tabla 2. ANOVA bidireccional para el efecto de la temperatura y la alimentación en las larvas: M, masa del renacuajo; Ts, tamaño corporal de los renacuajos; Diet, alimentación; Temp, temperatura del agua en el acuario; Type III MS, tipo III de suma de cuadrados; (a) R2 = 0,386 (R2 ajustada = 0,115); (b) R2 = 0,467 (R2 ajustada = 0,232); Df, grados de libertad. (NS, no significativo; * P < 0,05; ** P < 0,005; *** P < 0,001). Trait

Source

Model correct Intercept M Ts

Diet Temp Diet Temp

M * Ts Error

Type III MS

159,482 505 0.316 1.430 0.000 20,644.427 (b) 505 40.880 2.011 0.000 900,868 105,359.122

1 1

900.868 4,079.880 0.000 105,359.122

5,182.925

Temp

57.375

2.822

0.000 ***

12,624

0.361

1.634

0.012 *

3,137.374

89.639

4.410

0.000 ***

67,838

380

0.179

0.808

0.993 NS

Temp

11,037.085

380

29.045

1.429

0.000 ***

Diet

254,591

1153

0.221

23,438.322

1153

20.328

Temp Diet

Temp

4,248.000

1659

Temp

465,051.000

1659

Diet

414,074

1658

Temp

44,082.749

1658

Diet

Total corrected

0.000

5,163.739

Total

(a)

NS Diet 24,051 90 0.267 1.210 0.095

Diet

Type III MS

Larvae of aquatic amphibians are an ideal model to analyze thermal adaptations (Gutiérrez–Pesquera et al., 2016) due to their high dependence on temperature in the surrounding environment. These larvae have a low ability to regulate body temperature (Hutchison and Dupré, 1992), making them poikilothermic organisms (Balogová and Gvozdík, 2015). Their search for favorable microhabitats is therefore limited as the temperature of the pool affects their rate of development and growth and the duration of the metamorphosis. Most studies on larvae have shown that development responds to temperature, following Bergmann's Rule (1847) and resulting in large sizes at low temperatures (Ashton, 2002; Álvarez and Nicieza, 2002a; Laugen et al., 2005; Walsh, 2008; Walsh et al., 2008). Álvarez and Nicieza (2002b) reported that hatching temperature had no effect on mass loss during metamorphosis in Discoglossus galganoi, confining its main effect to the size and larval duration at metamorphosis. Our results show that temperature affects the development of larvae of Pleurodema thaul, causing those exposed to higher temperatures to re-

ach metamorphosis in less time than those exposed to low temperatures. This is because of the direct dependence between development and metabolic rate which it is increased with temperature, causing a faster development at higher temperatures (Barja de Quiroga, 1993). This is not surprising, since the hormones that regulate metamorphosis also control development of limbs and are highly sensitive to temperature (Ryan and Winne, 2000; Álvarez and Nicieza, 2002a, 2002b). Kehr (1998) indicated that the tadpole's increase in the rate of growth and development is generally regarded as a mechanism to decrease the risk of mortality through predation or desiccation of the habitat (Goldberg et al., 2012). This is why when individuals develop in an environment with high temperatures, the duration of larval stages are shorter, causing newly metamorphosed individuals to have both smaller body and mass than those organisms that develop at lower temperatures (Benavides, 2003; Benavides et al., 2005). We found a similar trend in P. thaul, where temperature had a direct effect on both mass and body size of the newly metamorphosed

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40 38 36 32

30 Temperature (ºC)

95 % CI development stages

28 26 40 38 36 34

32 30 28 26 1

21 28 35 45 50 57 64 Time of development (day)

Fig. 3. Results of larval development according to Gosner's table (1960) from stage 25 to 46. This indicates the percentage of larvae in each of the stages during the weeks of the experiment for each treatment. The results are presented separately for treatment at 15 ºC and treatment at 25 ºC. Fig. 3. Resultados del desarrollo larval de acuerdo con la tabla de Gosner (1960), desde el estadio 25 al 46. En ella se indica el porcentajes de larvas en cada estadio para cada tratamiento durante las semanas que duró el experimento. Los resultados se presentan separados por tratamiento de temperatura a 15 ºC y tratamiento a 25 ºC.

individuals. This has also been reported by Blouin and Brown (2000) in Rana cascadae. On the other hand, studies by Sanuy et al. (2008) on Epidalea calamita (sin. Bufo calamita; Frost, 2014) suggest that this species requires a minimum size and a larval body mass to complete its metamorphosis; an effect not evidenced in P. thaul because individuals completed their metamorphosis at different sizes and body masses. The study of phenotypic plasticity in amphibian larval development has long been of interest to ecologists. Our results show that temperature is the main factor affecting larval development, whereas the quality of diet presents a secondary effect on larval size and body mass. This is only significant in the initial larval stage (Gosner 25 to 30), probably coincident with the beginning of the metamorphic transformation. By observing the effect produced by the type of diet on larvae, it can be affirmed that there is a relationship between mass and body size of the individual, becau-

se those larvae fed a diet rich in protein (RP) reach a greater mass and body size than those fed a diet poor in protein (LP). This is because the presence of proteins in the diet is a determining factor for growth as they correspond to the main component of organs cell structures of tissues (Pelegrín et al., 2004). In addition, the lack of nutrients in diet constitutes an important selective factor since it influences the functioning of the thyroid gland, which would directly affect the rate of growth and differentiation of organisms (Álvarez and Nicieza, 2002a). We can conclude that both diet and temperature affect the development of larvae of P. thaul, where those larvae exposed to higher temperatures reach lower sizes and body mass than those found at lower temperatures. On the other hand, larvae fed with RP have greater mass and size than those fed with LP. When both factors are combined, diet and temperature, individuals exposed at higher temperatures and fed with LP reach metamorphosis more slowly than

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those fed an RP diet. These results relegate the effect of diet over that of temperature, confirming the main effect of temperature over diet. It should be noted that the temperature of 25 ºC is considered optimum for larval growth in this species, which leads to the fact that larvae can maintain a similar development with lower amounts of proteins (Kupferberg et al., 2011). It would also be expected (similarly to findings in studies on metamorphosis of insects; Stevens, 2004) that temperature has high importance on the rates of development (Gillooly et al., 2002), influencing the duration and size of the amphibians at the end of the metamorphic climax (Walsh, 2008). Thus, the influence of temperature on the development of larvae makes it a selective agent, promoting thermal adaptations (Angilletta, 2009; Bozinovic et al., 2011). Acknowledgements The authors thank the Servicio Agrícola y Ganadero (SAG) for providing collection permit Nº 9411/2014) and the VRID project 213.413.010–1.0. Peter Lewis translated the manuscript and provided helpful comments. References Angilletta, M., 2009. Thermal adaptation a theorical and empirical synthesis. Editorial Oxford, New York, USA. Angilletta, M., Steury, T., Sears, M., 2004. Temperature, Growth Rate, and Body Size in ectotherms: Fitting pieces of a life–history puzzle. Integrative and Comparative Biology, 44: 498–509. Ashton, K. G., 2002. Do amphibians follow Bergmann’s rule?. Canadian Journal of Zoology, 80: 708–716. Álvarez, D., Nicieza, A., 2002a. Effects of induced variation in anuran larval development on postmetamorphic energy reserves and locomotion. Ecophysiology, 131: 186–195. – 2002b. Effects of temperature and food quality on anuran larval growth and metamorphosis. Funcional ecology, 16(5): 640–648 Arnold, S. J., Wassersug, R. J., 1978. Differential predation on metamorphic anurans by garter snakes (Thamnophis): social behaviour as a possible defence. Ecology, 59: 1014–1022. Atkinson, D., 1996. Ectotherm life– history responses to development temperature. In: Animals and Temperature: Phenotypic and Evolutionary Adaptation: 183–204 (I. Johnston, A. Bennett, Eds.). Cambridge University Press, New York. Balogová, M., Gvozdík, L., 2015. Can newts cope with the heat? Disparate thermoregulatory strategies of two sympatric species in water. PLoS ONE 10: 128–155. Barja de Quiroga, G., 1993. Fisiología Animal y Evolución. Editorial Universitaria, España. Benavides, A., 2003. Biología térmica de Bufo spinulosus: Efecto de la temperatura sobre el desarrollo

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larval, Una Comparación Intraespecífica. Tesis doctoral, Universidad de Chile, Santiago, Chile. Benavides, A., Veloso, A., Jiménez, P., Méndez, M., 2005. Assimilation efficiency in Bufo spinulosus tadpoles (Anura: Bufonidae): effects of temperature, diet quality and geographic origin. Revista Chilena de Historia Natural, 78: 295–302. Bergmann, C., 1847. Über die Verhältnisse der Wärmeökonomie der Thiere zu ihrer Grösse. Göttinger Studien, 3(1): 595–708. Blair, R., Taylor, R., 2008. Bioestadística. Editorial Pearson, México. Bozinovic F., Calosi, P., Spicer, J. I., 2011. Physiological correlates of geographic range in animals. Annual Review of Ecology, Evolution, and Systematics, 42: 155–179. Blouin, M., Brown, T., 2000. Effects of temperature–induced variation in anuran larval growth rate on head width and leg length at metamorphosis. Oecologia, 125: 358–361. Cei, J. M., 1962. Batracios de Chile. Ediciones Universidad de Chile, Santiago, Chile. Díaz–Páez, H., Ortiz, J., 2001. The reproductive cycle of Pleurodema thaul (Anura, Leptodactylidae) in Central Chile. Amphibia – reptilian, 22: 431–445. Duellman, W., Trueb, L., 1986. Biology of amphibians, Mc Graw–Hill Press, USA. Frost, D. R., 2014. Amphibian Species of the World: an Online Reference Version 6.0. Electronic Database. American Museum of Natural History, New York, USA. Retrieved from: http://research.amnh.org/vz/ herpetology/amphibia/ [Accessed on May, 2016] Gillooly, J. F., Charnov, E. L., West, G. B., Savage, V. M., Brown, J. H., 2002. Effects of size and temperature on developmental time. Nature, 417: 70–73. Goldberg, T., Nevo, E., Degani, G., 2012. Phenotypic Plasticity in Larval Development of Six Amphibian Species in Stressful Natural Environments. Zoological Studies, 51(3): 345–361. Gosner, K. L., 1960. A simplified table for staging anuran embryos and larvae with notes on identification. Herpetologica, 16: 183–190. Granval, N. I., Gaviola, J. C., 1991. Manual de producción de semillas hortícolas. Instituto Nacional de Tecnología Agropecuaria, Argentina. Gutiérrez–Pesquera, L. M., Tejedo, M., Olalla–Tarraga, M. A., Duarte, H., Nicieza, A., Sole, M., 2016. Testing the climate variability hypothesis in thermal tolerance limits of tropical and temperate tadpoles. Journal of Biogeography, 2016: 1–13. Hickman, C., Roberts, L., Larson, A., 2002. Principios Integrales de Zoología (10ª edic.) Editorial McGraw–Hill Interamericana S. A., Madrid, España. Hutchison, V. H., Dupré R. K., 1992. Thermoregulation. In: Environmental Physiology of the amphibians: 206–249 (M. E. Ferder, W. M. Burggren, Eds.). The University of Chicago Press, Chigago. Iturra–Cid, M., 2007. Efecto de la latitud y altitud en la tasa de crecimiento y edad de madurez sexual en poblaciones de Pleurodema thaul (Amphibia: Leptodactylidae) en Chile. Seminario de Título presentado a la Facultad de Ciencias Naturales y Oceanográficas para optar al título de Biólogo.

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Universidad de Concepción, Concepción. Kehr, A. I., 1998. Applicability of three growth model to tadpoles body size in natural conditions. Physis, 55: 23–27. Kupferberg, S., Catenazzi, A., Power, M., 2011. The importance of water temperature an algal assemblage for frog conservation in northern California rivers whit hydroelectric projects. Final project report. University of California, Berkeley Laugen, A. T., Laurila, A., Jonsson, K. I., Soderman, F., Merila, J., 2005. Do common frogs (Rana temporaria) follow Bergmann’s rule? Evolutionary Ecology Research, 7: 717–731. Naya, D., Bozinovic, F., Sabat, P., 2008. Ecología nutricional y flexibilidad digestiva en anfibios. In: Herpetología de Chile: 427–451 (M. Vidal, A. Labra, Eds.), Science Verlag, Chile. Newman, R. A., 1988. Adaptive plasticity in development of Scaphiopus couchii tadpoles in desert ponds. Evolution, 42: 774–783. – 1989. Developmental plasticity of Scaphiopus couchii tadpoles in an unpredictable environment. Ecology, 70: 1775–1787. Pelegrín, E., Fraga, I., Álvarez, S., Galindo, J., Jaime, B., 2004. Efecto de diferentes niveles de proteína en la dieta de renacuajos de Rana Toro (Rana catesbeiana). Comunicación científica: 557–565. Ryan, T., Winne, Ch., 2000. Effects of Hydroperiod on Metamorphosis in Rana sphenocephala. The American Midland Naturalist, 145: 46–53. Sanuy, D., Oromí, N., Galofré, A., 2008. Effect of temperature on embryonic and larval development and growth in the natterjack toad (Bufo calamita) in a semi–arid zone. Animal Biodiversity and Conservation, 31(1): 41–46. Semlitsch, R. D., 1993. Effects of different predators on the survival and development of tadpoles from

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the hybridogenetic Rana esculenta complex. Oikos, 67: 40–46. Spotorno, M., 1995. Vertebrados. In: Diversidad biológica de Chile: 299–313 (J. Simonetti, M. Arroyo, A. Spotorno, E. Lozada, Eds.). Conicyt, Santiago, Chile. Stevens, D. J., 2004. Pupal development temperature alters adult phenotype in the speckled wood butterfly, Pararge aegeria. Journal of Thermal Biology, 29: 205–210. Tejedo, M., Duarte, H., Gutiérrez–Pesquera, L., Beltrán, J., Katzenberger, M., Rezende, E., Marangoni, F., Richter–Boix, A., Navas, C., Santos, M., Nicieza, A., Simon, M., Relyea, R., Solé, M., 2012. El estudio de las tolerancias térmicas para el examen de hipótesis biogeográficas y de la vulnerabilidad de los organismos ante el calentamiento global. Ejemplos en anfibios. Boletín de la Asociación Herpetológica española, 23(2): 2–27. Tracy, C. R., Christian, K. A., O’Connor, M. P., Tracy, C. R., 1993. Behavioral thermoregulation by Bufo americanus: the importance of the hydric environment. Herpetologica, 49: 375–382. Vidal, M. A., Soto, E. R., Veloso, A., 2009. Biogeography of Chilean herpetofauna: distributional patterns of species richness and endemism. Amphibia–Reptilia, 30:151–171. Walsh, P. T., 2008. The plasticity of life histories during larval development and metamorphosis, using amphibians as study organisms. PhD thesis, Glasgow University, Glasgow. Walsh, P. T., Downie, J.R., Monaghan, P., 2008. Larval over–wintering: plasticity in the timing of life history events in the common frog. Journal of Zoology, 276: 394–401. Wilbur, H. M., 1980. Complex life–cycles. Annual Review of Ecology and Systematics, 11: 67–93.

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Striking resilience of an island endemic bird to a severe perturbation: the case of the Gran Canaria blue chaffinch Á. C. Moreno, L. M. Carrascal, A. Delgado, V. Suárez, J. Seoane

Moreno, Á. C., Carrascal, L. M., Delgado, A., Suárez, V., Seoane, J., 2018. Striking resilience of an island endemic bird to a severe perturbation: the case of the Gran Canaria blue chaffinch. Animal Biodiversity and Conservation, 41.1: 131–140. Abstract Striking resilience of an island–endemic bird to a severe perturbation: the case of the Gran Canaria blue chaffinch. Evidence regarding population trends of endangered species in special protection areas and their recovery ability from catastrophic disturbances is scarce. We assessed the population trend of the Gran Canaria blue chaffinch (Fringilla polatzeki), a habitat specialist endemic to the pine forest of Inagua in the Canary Islands, following a devastating wildfire in July 2007. Using a standardized census program that accounts for detectability, we have monitored the population trend of the species since Inagua was declared a Strict Nature Reserve in 1994. The breeding population density of the blue chaffinch remained stable in Inagua from the beginning of the monitoring program in 1994 until the year before the wildfire. However, in spring 2008, the population density decreased by half with respect to density in the preceding years. Since 2008, the population has gradually increased, reaching its highest recorded density in 2016 (15.8 birds/km2).This represents an average annual increase of 23.7 %, indicating impressive resilience to catastrophic events. The creation of Inagua as a strict nature reserve did not therefore increase the global population or protect the blue chaffinch against a demographic crisis but probably prevented a deepening of the demographic crisis or further declines. Except for the two years immediately after the severe wildfire of 2007, the population density of the blue chaffinch in Inagua has remained relatively stable at around 9–16 birds/km2, the lowest recorded abundance for a small woodland passerine in the Western Palearctic. Key words: Blue chaffinch, Canary Islands, Density, Population trend, Strict nature reserve, Wildfire Resumen Marcada resiliencia de una especie de ave insular endémica después de una perturbación intensa: el caso del pinzón azul de Gran Canaria. Son pocos los datos disponibles sobre la tendencia demográfica de las especies en peligro de extinción en zonas de protección especial y su capacidad de recuperarse de perturbaciones catastróficas. Se estudia la tendencia demográfica del pinzón azul de Gran Canaria (Fringilla polatzeki), un especialista de hábitat endémico de las Islas Canarias, restringido al pinar de Inagua, que sufrió un devastador incendio forestal en julio de 2007. Mediante un programa de censo estandarizado que tiene en cuenta la variación en la capacidad de detección, se ha hecho un seguimiento de la tendencia demográfica de la especie desde la declaración de Inagua como reserva natural integral en 1994. La densidad reproductiva del pinzón azul se mantuvo estable en Inagua desde el inicio del programa de seguimiento en 1994 hasta un año antes del incendio. No obstante, en la primavera de 2008, la densidad de la población se redujo a la mitad en comparación con los años anteriores. A partir de 2008, la población del pinzón azul ha venido aumentando gradualmente hasta alcanzar la densidad más alta jamás registrada en 2016 (15,8 aves/km2), lo que equivale a un incremento anual medio del 23,7 % y pone de manifiesto la resistencia impresionante de estas poblaciones ante catástrofes. Por lo tanto, la creación de la reserva integral de Inagua no promovió el aumento de población ni protegió al pinzón azul frente a una grave crisis demográfica, sino que probablemente evitó que la disminución de la población fuera más profunda o que se produjeran otras reducciones. Aparte de los dos años inmediatamente posteriores al incendio forestal de 2007, la densidad de población del pinzón azul en Inagua se mantuvo relativamente estable alrededor de 9–16 aves/km2, la menor abundancia jamás registrada para un paseriforme forestal de tamaño pequeño en todo el paleártico occidental. ISSN: 1578–665 X eISSN: 2014–928 X

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Palabras clave: Pinzón azul, Gran Canaria, Densidad, Tendencia poblacional, Reserva natural integral, Incendio forestal Received: 12 V 17; Conditional acceptance: 21 VI 17; Final acceptance: 31 VII 17 Ángel C. Moreno, Dirección General de Protección de la Naturaleza, Gobierno de Canarias, c/ Prof. Agustín Millares Carló 18, 35071 Las Palmas de Gran Canaria, Canary Islands, Spain.– Luis M. Carrascal, Dept. of Biogeography and Global Change, Museo Nacional de Ciencias Naturales–CSIC, c/ José Gutiérrez Abascal 2, 28006 Madrid, Spain.– Alejandro Delgado, Víctor Suárez, Javier Seoane, Terrestrial Ecology Group, Depto. de Ecología, Univ. Autónoma de Madrid, 28049 Madrid, Spain. Corresponding author: L. M. Carrascal. E–mail: lmcarrascal@mncn.csic.es

Animal Biodiversity and Conservation 41.1 (2018)

Introduction Resilience against critical events is a scarcely studied but important matter, especially in endangered species. From 1994 to 2004, Butchart et al. (2006) documented the relative success of conservation efforts that prevented sixteen bird species from becoming extinct. Many of them were threatened birds inhabiting oceanic islands, with very low populations restricted to single, discrete sites. The main sources of extinction risk in these circumstances were related to habitat loss and degradation, deleterious effects which were reduced or eliminated through habitat protection, management and restoration, especially inside protected areas. Strict natural reserves are established to protect biodiversity, both as a whole and considering those threatened species that face conservation challenges. Nevertheless, the effectiveness of protected areas is a subject of continuous debate and testing to evaluate its success, poor results, or need for improvement (Martínez et al., 2006; Craigie et al., 2010; Gutiérrez and Duivenvoorden, 2010; Cantú–Salazar et al., 2013; Dunn et al., 2016). This is most notably the case when phenomena and processes occurring outside the limits of the protected areas affect the populations within them (e.g., global warming, changes in rainfall regime, emergent diseases, invasive species), and is of concern for species with very small ranges, and possibly restricted to a single location. Such conditions attract conservation focus and efforts to declare such areas a reserve. It is therefore important to accumulate evidence regarding whether protected areas for endangered species have contributed to the recovery of their populations, in particular the reserves that are the last shelters for the most narrowly–distributed species (Geldmann et al., 2013). Moreover, considering the low amount of detailed information on particular species regarding how extinctions are prevented, it is necessary to increase our knowledge about their recovery ability after drastic population declines. The Gran Canaria blue chaffinch (Fringilla polatzeki, Canary Islands) is a rare, threatened species that occupies an island–habitat within the island of Gran Canaria (Martín and Lorenzo, 2001 for the probable status of the species since the beginning of the 20th century). Currently split from F. teydea according to genetic, morphological and behavioural data (Pestano et al., 2000; Lifjeld et al., 2016; Sangster et al., 2016), it is mainly restricted to the Strict Nature Reserve of Inagua–Ojeda–Pajonales (Inagua, hereafter; 39.2 km2; Moreno and Rodríguez, 2007), although a few pairs have recently established elsewhere as a result of a translocation program (Delgado et al., 2016). The Gran Canaria blue chaffinch is a habitat specialist of the mature Canarian pine forests (Pinus canariensis), likely as a consequence of past competition with other Fringilla species and niche displacement (Illera et al., 2016). It nests in tall trees. Breeding success is low for a Fringillidae, with only ca. 1.5 fledglings per successful nesting attempt, and 1.4 clutches per breeding season (Rodríguez and Moreno, 2008; Delgado et al., 2016). The estimated

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population size (with a previous educated guess at around 300 birds, BirdLife International, 2016) lies within the left tail of the distribution of minimum viable population (MVP) estimates for many species, far from the average MVP of 3,750 individuals for birds (Brook et al., 2006; Traill et al., 2007). This is most notable if we take into account the small size of the species (approx. 30 g), since body mass in birds is usually negatively correlated with abundance or maximum ecological densities in the preferred habitats (Carrascal and Tellería, 1991; Gaston and Blackburn, 2000). The main goal of this study was to analyse the population trend shown by the Gran Canaria blue chaffinch in Inagua since the forest was declared a Strict Nature Reserve in 1994, the only area in the world where the species was present until then as a regular breeder (Martín and Lorenzo, 2001). If the declaration of this area as a reserve contributed to the conservation of the species, we would expect to find a non–decreasing population trend (either positive or stable annual counts). A wildfire in July 2007 that badly damaged the pine forest of the Inagua Reserve provided an opportunity to quantify how severe fire affected the blue chaffinch population and how it recovered in the following years. Material and methods Study area The study area is located in the Inagua pine forest of Gran Canaria (27º 58' N, 15º 35' W), an island of volcanic origin (1,560 km2, maximum altitude of 1,950 m a.s.l.; for more details on the vegetation of the island see Santos, 2000). The Inagua Integral Natural Reserve (39.2 km2, 250–1,550 m a.s.l.; Special Protection Area of the European Union since 1979; see fig. 1) is a mature pine forest that harbours the main extant breeding population of the Canaria blue chaffinch (Moreno and Rodríguez, 2007). This chaffinch is scarce in pinewoods below 1,000 m a.s.l. (Moreno and Rodríguez, 2007). A severe fire in July 2007 badly affected the Inagua Reserve (see fig. 1 in Suárez et al., 2012). The Canary pine has the remarkable characteristic of being able to survive and grow after fire, and thus in most places, pine foliage was partially recovered by June 2008, and the tree foliage showed full growth by the breeding season of 2010. For environmental characteristics of the Inagua pine forest see Rodríguez and Moreno (2008). Bird census Data on bird abundance were obtained through line transect sampling in Inagua during the breeding season of the species (the second fortnight in May and the first fortnight in June; see Rodríguez and Moreno, 2008) over 18 years, from 1994 to 2016. We surveyed a fixed net of trails following a single route of a total length of 22.9 km on adequate habitat over the area with the highest density of the species (see fig. 1). Since 1994 we used the same line–transect method. From 1994 to

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2008, the route was censused only once per year, but from 2009 to 2016, the transect was repeated three times on different days and bird counts were averaged to assess whether more precise results could be obtained. Transects were carried out on rainless days. Researchers walked slowly (1–3 km/h approximately) along small trails in the first four hours after dawn. The censuses were performed by different people: A. C. M. from 1994 to 2004; L. M. C. and J. S. in 2008; and V. S. and A. D., in 2006, 2009–2016. To account for between–observers and between–year variations in detectability while we used distance sampling methods. For each bird heard or seen, we estimated the perpendicular distance to the observer’s trajectory. Previous training helped to reduce between–observer variability in distance estimates. Detection distances were right–truncated as recommended by Buckland et al. (2001), excluding 5 % of birds recorded far away (i.e. beyond 125 m). Four models that are commonly used to explain the loss of detectability as a function of the distance from the transect line were fitted to estimate the probability of detection within strips of width equal to the truncated distance: half–normal and hazard–rate, with the inclusion of polynomial or cosine adjustment terms (Buckland et al., 2007). Models were evaluated according to AICc to obtain model weights. The weighted mean of the probability of detection and the effective strip width were used to estimate population densities from the number of blue chaffinches detected (using Akaike’s weights). Detectability models for the blue chaffinch were built with R version 3.1.2 (R Core Team, 2014) and specialized packages: Distance (Miller, 2016a) and mrds (Miller, 2016b). Population density and trends Population density of the blue chaffinch in Inagua was calculated considering the counts of birds in the 22.9 km census route and the effective strip width (ESW) derived from the probability of detection. The total length of transects were divided into 100 contiguous units of equal length (229 m), to which the detected blue chaffinches were assigned in each year. As these one–hundred units are not truly independent samples, a bootstrapping procedure was carried out to estimate the average density and the proper confidence intervals (Davison and Hinkley, 2007). Density for each year in each randomization trial was estimated considering (1) the total number of chaffinches in the bootstrap sample, (2) a random probability of detection obtained from the corresponding 95 % confidence interval for that year (to account for uncertainty in the probability of detection; see table 1), and (3) a strip width of 125 m on both sides of the 22.9 km route. We carried out 20,000 randomizations to estimate population density in each of the 18 years of study. Confidence intervals were obtained using the percentile method, considering the non–Gaussian distribution of density figures. To assess population trends of blue chaffinch in Inagua, we used the bird counts obtained from 1994 to 2016 within the 100 sample units of 229 m–long transects (we used the counts of one census per

year from 1994 to 2008, and the average of three counts from 2009 to 2016). First, we estimated the between–years population changes (byPC) in any two consecutive years t and t+1 as: byPC = 1 + [(Dt+1 – Dt) / Dt] with D being the average density in the 100 sample units. Second, we randomly assigned the bird density in each one of the 100 sample units between years t and t+1, by shuffling the density figures within rows (with sample units as rows and years as columns), and calculated the null between–year population change as presented in the previous step. Note that this randomization procedure preserves the spatial structure of the data, because the shuffling is limited to rows. And finally, this randomization procedure was repeated 20,000 times to obtain the null distribution of population trend figures between consecutive years. The observed population changes between the two years under comparison were tested against the two–tailed 95 %, 99 % and 99.9 % percentiles of the null distributions. Analyses were carried out using the Bootstrapping, Resampling and Monte Carlo functions of 'PopTools 3.0', http://www.cse.csiro.au/poptools/, run in Microsoft Excel 2010. Results Gran Canaria blue chaffinch counts ranged from 17 to 50 individuals over the years, and probability of detection within the 125–m strip width ranged between 0.52 and 0.71 over study periods (table 1). The width of the confidence intervals of bird counts, relative to the average, was lower in years when three repetitions of the censuses were carried out (2009–2016; average relative width = 54.1 %) than in years when only one census was carried out (1994–2008; average = 74.3 %; p < 0.001 in the t–test comparing the two census periods; table 2). Thus, three repetitions per year of the same census transect increased the precision of the estimates of average density. The population density of the blue chaffinch remained stable from the beginning of the monitoring program in 1994 to one year prior to the devastating forest fire in July 2007 (table 2, fig. 2), with an average density of 9.7 birds/km2 (range of year averages: 8.0–12.7 birds/km2). Pairwise tests comparing counts on all pairs of years showed that even the peak in chaffinch abundance in 2000 was not significantly different from the other density estimates (55 tests using sequential Bonferroni correction for type I error–wise rate at α = 0.05). Population density in spring 2008 (10 months after the forest fire) halved with respect to that measured in 2006 (58 % reduction to 4.8 birds/km2; p = 0.001). From 2008 onwards, the blue chaffinch population gradually increased, with a significant increase from 2009 to 2010 (p = 0.005 that remains significant after a sequential Bonferroni correction of the six tests between consecutive years from 2008 to 2016). The linear correlation between year and population

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135

3 km

0 250 500 km

Lanzarote La Palma Tenerife La Gomera

N W

El Hierro

0 5 10 15 km

Fuerteventura

Gran Canaria 0 50 100 150 km

S Fig. 1. Study area in Gran Canaria island. Black dots show the centre of 100 units of 229 m in length of a census route of 22.9 km repeated from 1994 to 2016. Fig. 1. Zona de estudio en la isla de Gran Canaria. Los puntos negros indican el centro de 100 unidades de 229 m de longitud de un transecto de 22,9 km para elaborar el censo que se ha venido repitiendo desde 1994 hasta 2016.

density was high from 2008 to 2016 (r = 0.886, 99 % bootstrapped confidence interval: 0.696–0.988). Population abundance in the last monitoring year, 2016, was higher than any other previous year, with an average density of 15.8 birds/km2. The percentage of population increase from 2008 to 2016 was 229 %. To summarize, the population density of the blue chaffinch in the Inagua reserve remained stable at around 10 birds/km2 from 1994 to 2006, decreased as a consequence of the devastating forest fire in July 2007, remained low during the subsequent two years, and then showed a clear increasing trend during the following eight years, reaching the highest density ever recorded in 2016. Discussion The endangered blue chaffinch of Gran Canaria Island has shown a remarkably stable long–term population trend over the last 23 years. Given its scarcity in the past and the extremely restricted distribution area of this species (Martín and Lorenzo, 2001), a strict natural reserve was established in 1994 in Inagua. The

devastating forest fire in July 2007 halved the chaffinch population on the island. Nevertheless, it has shown an impressive resilience as the population recovered 3–4 years after the wildfire, reaching the highest population density ever recorded in 2016. Moreover, the demographic bottleneck was not accompanied by a clear genetic erosion, as the blue chaffinch has not experienced a significant decline in allelic richness or an increase in the inbreeding coefficient (Suárez et al., 2012). These results reveal the ability of this endemic chaffinch to survive in these unique forests within the context of the Western Palearctic, and the adaptation of both bird and tree to recovery after wildfires, a common phenomenon in volcanic islands such as the Canary archipelago. The population trend of the species in the Inagua Strict Nature Reserve supports that 'broad and shallow' protection of endangered species, resting only in the passive protection of areas, is less effective than 'narrow and deep' protection, with more financial expenditures, dealing with populations (e.g., Kolecek et al., 2014; Luther et al., 2016), because the creation of Inagua nature reserve did not avoid the wildfire risk for this species. The highly stable population density of the blue chaffinch in the mature

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Table 1. Detectability estimates of the Gran Canaria blue chaffinch carried out for different time periods, each with a different team of observers. Bird counts were obtained by distance–sampling over the same fixed route of 22.9 km: Best model, best fitted model with the lowest AIC figure; HNc, half–normal with cosine adjustment; HNp, half–normal with polynomial adjustment; HRc, hazard–rate with cosine adjustment; HRc, hazard–rate with polynomial adjustment); pDET, probability of detection within 125–m strip width (SE, standard error); ESW, effective strip width (in m); #birds, number of bird contacts (also including other contacts obtained censusing other forest tracts in Inagua in 2008). Tabla 1. Estimaciones de la capacidad de detección del pinzón azul en Gran Canaria realizadas con respecto a períodos distintos, cada una de ellas con un equipo diferente de observadores. Los conteos de aves se obtuvieron mediante un muestreo a distancia a lo largo de la misma ruta establecida de 22,9 km: Best model, mejor modelo ajustado con la menor cifra del AIC; HNc, seminormal con términos de ajuste de coseno; HNp, seminormal con términos de ajuste polinómicos: HNc, tasa de riesgo con términos de ajuste de coseno; HRc, tasa de riesgo con términos de ajuste polinómicos; pDET, probabilidad de detección en una franja de 125 m de ancho (SE, error estándar); ESW, ancho efectivo de la franja (en m); #birds, número de contactos con aves (incluidos otros contactos obtenidos en censos de otros transectos forestales realizados en Inagua en 2008). Years

Best model

#birds

ESW

pDET

SE pDET

1994–2004

HNc

345

79.3

0.634

0.054

2008

HNc

88.3

0.664

0.113

2006, 2009–2011

HNc

265

65.1

0.521

0.047

2013, 2015–2016

HNc

350

69.5

0.556

0.043

pine forest of Inagua may be understood considering long–term stability of this forest habitat, causing places suitable in one year to remain so over many seasons,

and to cross–generational reproducibility of the criteria used by birds in their settlement decisions (see also Wesołowski et al., 2015).

Table 2. Bird counts (Bc), and their 95 % confidence intervals (L, lower 95 %; U, upper 95 %), for the Gran Canaria blue chaffinch population in Inagua pine forest during the second fortnight in May and the first fortnight in June, throughout the 18–year study period, from 1994 to 2016. From 2009 to 2016, three censuses were carried per year on different days, while only one census per year was made in the remaining years; rel. width, width of the confidence intervals of bird counts, relative to the average. Tabla 2. Conteos de aves (Bc) y sus intervalos de confianza del 95 % (L, inferior; U, superior) de la población de pinzón azul en el pinar de Inagua, Gran Canaria, durante la segunda quincena de mayo y la primera de junio de los 18 años del estudio, entre 1994 y 2016. Entre 2009 y 2016 se realizaron tres censos anuales en distintos días, mientras que los demás años solo se realizó un censo anual; rel. width, amplitud del intérvalo de confianza, relativa al valor medio del conteo de aves. Years

Rel. width

Years

Rel. width

1994

20.0

43.0

74.2

2004 38 26.0 51.0 65.8

1995

17.0

43.0

89.7

2006

23.0

46.0

67.6

1996 29 19.0 39.0 69.0

2008 21 12.0 32.0 95.2

1997

69.7

2009

12.0

23.0

64.7

1998 30 19.0 43.0 80.0

2010

21.0

37.3

56.3

1999

75.0

2011

23.7

40.7

53.1

2000 46 34.0 59.0 54.3

2013

21.7

38.3

55.6

2001

26.0

55.0

72.5

2015

28.0

45.7

47.7

2002

21.0

47.0

78.8

2016

38.7

62.3

47.3

33 36

22.0 23.0

45.0 50.0

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Birds/km2

2016

2015

2013

2011

2010

2009

2008

2006

2004

2002

2001

2000

1999

1998

1997

1996

1995

1994

Year Fig. 2. Temporal variation of the blue chaffinch density in Inagua. Dots and continuous line denote average estimations, while shadow area shows the 95 % confidence intervals. Density estimates take into account the probability of detection (within its 95 % confidence interval) and the spatial heterogeneity in bird counts along the 22.9 km of the census trail. Asterisks show significant differences between consecutive density estimations after sequential Bonferroni correction (* P < 0.05; ** P < 0.01). Drawing of blue chaffinch from www.birdlife.org. Fig. 2. La variación temporal de la densidad del pinzón azul en Inagua. Los puntos y la línea discontinua denotan el promedio de las estimaciones, mientras que el área sombreada indica los intervalos de confianza del 95 %. Para calcular las estimaciones de la densidad se tuvo en cuenta la probabilidad de detección (dentro de su intervalo de confianza del 95 %) y la heterogeneidad espacial en los conteos de aves a lo largo de los 22,9 km del transecto. Los asteriscos indican las diferencias significativas entre estimaciones de densidad consecutivas tras una corrección secuencial de Bonferroni (* P < 0,05; ** P < 0,01). Dibujo del pinzón azul extraída del sitio web www.birdlife.org.

Apart from the two years immediately after the severe forest fire of 2007, the population density of the blue chaffinch in Inagua remained relatively stable at around 10 birds/km2 within its well–preserved core area (with a maximum of 15.8 birds/km2). This is one of the lowest ever recorded abundances for a small woodland passerine in the whole Western Palearctic (Hagemaijer and Blair, 1997), and more than four times lower than the maximum densities measured for the other blue chaffinch species in the pine forests of Tenerife Island (Fringilla teydea, 69 birds/km2, Carrascal and Palomino, 2005; 170 birds/km2, García– del–Rey et al., 2010). Similarly, the endemic Azores bullfinch Pyrrhula murina, also an endangered habitat specialist, reaches considerably higher densities of 100–200 birds/km2 (in native laurel forests of São Miguel Island; Ceia et al., 2009, 2011). This recorded low population density suggests important environmental limitations for the blue chaffinch in the Gran Canaria island, even in its emblematic protected core area.

The historic Gran Canaria pine forests (i.e., not derived from recent plantations), despite some relict populations of high haplotypic diversity (Vaxevanidou et al., 2006), are located in the south–eastern distribution limit of the species, and are probably remnants of larger populations severely reduced by human activities and adverse climatic conditions (precipitation decreases from west to east in the Canary Islands; Marzol, 2000). This is particularly evident for the remnant pine forests located around Tauro, where extremely dry conditions are manifested in symptoms of decay in many individuals (Vaxevanidou et al., 2006). Moreover, this situation will likely worsen as a consequence of climate change in the Canary Islands, where models predict increases in temperature and a decrease in precipitation over the next 85 years (Morata, 2014; Expósito et al., 2015). Warming has been more evident at high mountains than at lower altitudes in both Tenerife and Gran Canaria islands since 1970 (0.16 °C/decade; Martín et al., 2012; Luque

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et al., 2014). Recently, Brawn et al. (2016) suggested that the increase in dry season length may threaten populations of tropical birds in protected areas, even without a direct loss of habitat. Such evidence, together with the general biogeographic pattern of a decrease in species richness and abundance of woodland bird species towards the SW of the Western Palearctic (Mönkkönen, 1994; Tellería and Santos, 1994), suggest that the Gran Canaria blue chaffinch is a 'woodland survivor' stranded in a suboptimal habitat, in the eastern limit of the Canary forests of any kind. The South hills crossbill (Loxia sinesciuris of the curvirostra complex), inhabiting only the higher elevations of two small mountain ranges in southern Idaho (Rocky Mountains, USA), poses a similar case of a declining habitat specialist of coniferous forests (Benkman, 2016), where hot events (i.e., more than four hot days > 32 ºC per year) recorded from 2003 to 2011 caused a 20 % annual decline, with a total decline of 80 % of the population. Since one year after the forest fire of July 2007, the blue chaffinch population of Inagua has shown a steady growth until 2016, with an average annual increase of 23.7 %, a figure that is around the upper boundary of other threatened species (Green and Hirons, 1991; Butchart et al., 2006). This increase occurred with minor implementation of conservation actions (these limited to providing water supplies; Pascual Calabuig, pers. com.), leaving the species to its fate and dependent on the natural recovery of the pine forest. Moreover, 15 blue chaffinch juveniles were translocated from Inagua to La Cumbre pinewood forest, 2–4 km away, at the end of the summer 2015 (nine females and six males; this extraction was the most remarkable carried out in any one year from 1994 to 2016; Felipe Rodriguez and María Dolores Estévez, pers. comm.). In spite of this extraction, the population at Inagua did not show any sign of a population decrease, continuing with its steady increase from 2015 to 2016. The positive population trajectory is typical of species living at low densities that often recover after the perturbation that decreased their numbers ceases. This phenomenon is the result of high fidelity to good habitat patches, reduced mortality and increased fecundity and reproductive rate (e.g., Ferrer et al., 2013; Krüger et al., 2010; Le Corre et al., 2015; Smith et al., 2015). The case of the Gran Canaria blue chaffinch is one of those rare examples of how an endangered species recovers from a demographic crisis in the absence of human interventions, when the mere protection of the habitat is sufficient [see also Impey et al. (2002) for the Rodrigues fody, Foudia flavicans; Groombridge et al. (2009) for the Seychelles kestrel, Falco araea; Brooke et al. (2012) for the Raso lark, Alauda razae, confined to the 7 km2 island Raso, Cape Verde; Guevara et al. (2016) for Podiceps juninensis in northern Andes; Burt et al. (2016) for Copsychus sechellarum in the Seychelles]. In conclusion, the Gran Canaria blue chaffinch is a small passerine of the Western Palearctic that attains the lowest population densities for a forest bird, even in the most favourable woodland areas (ca. 10 birds/km2). However, the population has remained relatively sta-

ble during the last twenty–three years. The creation of the Inagua strict reserve and its role as a special protection area for birds was not followed by a population increase and did not protect the species from the demographic crisis associated with a devastating wildfire that halved its population, although the strict protection status of Inagua allowed for a quick recovery of the species. The species showed high resilience and adaptation to wildfires, recovering at a fast rate (24 % average yearly increase) in the following eight years, without human intervention. These results clearly illustrate that an insular endemic species with a population size below the 'average' minimum viable population level may have stable numbers during relatively long periods without becoming extinct in spite of being recognized as endangered (Martín, 2009). Acknowledgements The study was supported by the Conservation Program for the blue chaffinch implemented by the Gobierno de Canarias (1991–2004), Cabildo de Gran Canaria (2005–2015), and was partially funded by the European Union (1995–1996: LIFE94 NAT/E/ 001159; 1999–2002: LIFE98 NAT/E/005354; 2016: LIFE14 NAT/ES/000077) and a research contract between MNCN/CSIC and GESPLAN, S. A. U. (2008). JS currently works within the Madrid’s Government research group network REMEDINAL3–CM (S–2013/ MAE–2719). We thank Joachim Hellmich, Pascual Calabuig, Ruth de Oñate and Felipe Rodríguez for their help and company during different phases of this long work, and to Claire Jasinski and Ana Rey for improving the English of the manuscript. References Benkman, C. W., 2016. The Natural History of the South Hills Crossbill in Relation to Its Impending Extinction. American Naturalist, 188: 589–601. BirdLife International, 2016. Fringilla polatzeki. The IUCN Red List of Threatened Species 2016: e. T103822640A104230366. http://dx.doi.org/10.2305/ IUCN.UK.2016–3.RLTS.T103822640A104230366.en Brawn, J. D., Benson, T. J., Stager, M., Sly, N. D., Tarwater, C. E., 2016. Impacts of changing rainfall regime on the demography of tropical birds. Nature Climate Change, 19: e3183 Brook, B. W., Traill, L. W., Bradshaw, C. J. A., 2006. Minimum viable population sizes and global extinction risk are unrelated. Ecology Letters, 9: 375–382. Brooke, M. D., Flower, T. P., Campbell, E. M., Mainwaring, M. C., Davies, S., Welbergen, J. A., 2012. Rainfall–related population growth and adult sex ratio change in the Critically Endangered Raso lark (Alauda razae). Animal Conservation, 15: 466–471. Buckland, S. T., Anderson, D. R., Burnham, K. P., Laake, J. L., Borchers, D. L., Thomas, L., 2001. Introduction to distance sampling. Oxford University Press, Oxford. – 2007. Advanced distance sampling. Oxford Uni-

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Spatial and environmental variation in phyllostomid bat (Chiroptera, Phyllostomidae) distribution in Mexico J. C. Arriaga–Flores, A. Rodríguez–Moreno, A. Correa–Sandoval, J. V. Horta–Vega, I. Castro–Arellano, C. J. Vázquez–Reyes, C. S. Venegas–Barrera Arriaga–Flores, J. C., Rodríguez–Moreno, A., Correa–Sandoval, A., Horta–Vega, J. V., Castro–Arellano, I., Vázquez–Reyes, C. J., Venegas–Barrera, C. S., 2018. Spatial and environmental variation in phyllostomid bat (Chiroptera, Phyllostomidae) distribution in Mexico. Animal Biodiversity and Conservation, 41.1: 141–159. Abstract Spatial and environmental variation in phyllostomid bat (Chiroptera, Phyllostomidae) distribution in Mexico. Species’ spatial distribution patterns allow us to understand the establishment of different biotic components in different environmental conditions. This study analyzes the spatial distribution of the Phyllostomidae family in Mexico to identify groups of species that occur in similar sites, the environmental conditions associated with species distribution, and the percent of overlap with human–modified areas. The results suggest six groups of sites with particular species composition. The spatial variation in richness pattern was associated with species tolerance to environmental conditions, such as minimum temperature and tree cover. The convergence between species distribution and modified areas varied per species feeding guild. Insectivorous and nectarivorous bats were sensitive species because they occurred in narrow environmental conditions and their distributions overlapped with areas modified by human activities. The approach implemented here analyzes regional species distributions and estimates their environmental requirements, contributing to the development of optimal conservation strategies for susceptible bat species. Key words: Biodiversity conservation, MaxEnt, Multivariate analysis, Niche breadth, Species diversity Resumen Variación espacial y ambiental en la distribución de murciélagos filostómidos (Chiroptera, Phyllostomidae) en México. Los patrones de distribución espacial de las especies permiten comprender el establecimiento de distintos componentes bióticos en diferentes condiciones ambientales. En este estudio se analiza la distribución espacial de la familia Phyllostomidae en México para identificar grupos de especies que están presentes en sitios similares, las condiciones ambientales asociadas a su distribución y el porcentaje de solapamiento con zonas modificadas por el hombre. Los resultados sugieren que existen seis grupos de sitios con una composición de especies particular. La variación espacial en el patrón de riqueza se asoció con la tolerancia de las especies ante condiciones ambientales, como la temperatura mínima y la cobertura arbórea. La convergencia de la distribución de las especies y las zonas modificadas varió según el gremio trófico de las especies. Los murciélagos insectívoros y nectarívoros se consideraron especies sensibles, debido a que se encuentran en un reducido rango de condiciones ambientales y su distribución se solapa con zonas modificadas por actividades humanas. El planteamiento utilizado consistió en analizar la distribución regional de las especies y estimar sus requerimientos ambientales, lo que permite elaborar estrategias de conservación óptimas para las especies susceptibles de murciélagos. Palabras clave: Conservación de la biodiversidad, MaxEnt, Análisis multivariante, Amplitud de nicho, Diversidad de especies Received: 09 IX 16; Conditional acceptance: 27 II 17; Final acceptance: 03 VIII 17

ISSN: 1578–665 X eISSN: 2014–928 X

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J. C. Arriaga–Flores, A. Correa–Sandoval, J. V. Horta–Vega, C. J. Vázquez–Reyes, Crystian S. Venegas– Barrera, División de Estudios de Posgrado e Investigación, Inst. Tecnológico de Ciudad Victoria, Blvd. Emilio Portes Gil 130, 87089 Ciudad Victoria, Tamaulipas, México.– A. Rodríguez–Moreno, Lab. de Sistemas de Información Geográfica, Inst. de Biología, Univ. Nacional Autónoma de México, 04510 Coyoacán, CDMX, México.– Iván Castro–Arellano, Biology Dept., Texas State Univ., 601 University Dr, San Marcos, Texas, USA. Corresponding author: Crystian S. Venegas–Barrera. E–mail: crystian_venegas@itvictoria.edu.mx

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Introduction Species diversity distribution is heterogeneous in space, so understanding and predicting this variation in different environments and taxa is a fundamental goal in ecology and biogeography (Lomolino et al., 2005). Species richness and community assembly variation may occur either gradually or abruptly, depending on the spatial scale analyzed and associated factors (e.g. availability of energy and water; Hawkins et al., 2003; Field et al., 2009). Species’ spatial distribution patterns are useful to understand the establishment of different biotic components in different environmental conditions (Morrone, 2009). However, the persistence of biological diversity and the functionality of ecosystems are at risk due to the increase in anthropogenic activities such as agriculture, livestock, and urban activities (Klein–Goldewijk and Ramankutty, 2004). Changes in native vegetation, for example, have modified 12 % of Earth’s land surface, affecting wildlife habitats and the diversity of species distribution (Ries et al., 2004). Therefore, the patterns of species composition vary depending on the environment, and high biodiversity areas that are threatened by human land cover changes must be evaluated to optimize conservation strategies, a key topic in conservation biology (Margules and Sarkar, 2009). Species potential distribution models are a useful tool to identify environmental conditions related to species presence, high biodiversity areas and zones with similar species composition (Mateo et al., 2013). Despite being locally biased by the number of collections records, the type of sampling, the choice of predictors or the algorithm used (Peterson et al., 2011), the fact that species distribution models provide reliable results at a regional scale (Raxwhorty et al., 2007; Lee et al., 2012) makes this technique appropriate to study bat communities at larger scales. The superposition of information from a distribution model and land–use cartography provides an estimate of the degree of overlap between human–modified areas and can be a useful strategy to identify areas at risk from human activities (Wu et al., 2014). An integrative approach to distribution models allows helps to evaluate ecological suitability in biodiversity areas to identify species that are susceptible to environmental changes and important as conservation targets (Peterson et al., 2015). The Phyllostomidae family is a Neotropical taxa that can be used to identify high biodiversity areas and zones at risk of human activities because it occurs in several different environments (Stevens, 2006) and has different specific feeding preferences (Giannini and Kalko, 2004). The regional distribution of phyllostomids has been associated with oscillations in temperature and humidity (McCain, 2007) and is locally determined by sensitivity to disturbance (Wordley et al., 2015). However, little is known about areas with a similar species composition, species environmental tolerances, and the degree of overlap between the distribution of bats and modified areas (López–González et al., 2011; Razgour et al., 2016). Phyllostomid bats provide ecosystem services as

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agents of seed dispersal, pollination, and pest regulation (Kunz et al., 2011).High diversity areas that may include species susceptible to constant human degradation must therefore be identified. Mexico has a wide range of environmental conditions, produced by altitudinal variations, the influence of two oceans, and the convergence of two biogeographical regions (i.e., the Neotropical and the Nearctic; Morrone, 2005) that offer multiple environments for 60 phyllostomid bat species (Ramirez–Pulido et al., 2014). However, these bat species are at risk from human activities, mainly the conversion of tropical forests (Challenger and Dirzo, 2009). This study analyzes the distribution of Phyllostomidae family species in Mexico, as well their different trophic guilds, since this information is useful to understand the response of the community structure to human disturbance (Klingbeil and Willig, 2009). The objectives of this study were to identify the spatial richness patterns, define groups of sites with a similar species composition, estimate the environmental tolerance of each species, and calculate the convergence between species distribution and human–modified areas. Methods Species potential geographic distribution Potential geographic distribution models for phyllostomid bat species in Mexico were developed using the maximum entropy algorithm (MaxEnt ver 3.3; Phillips et al., 2006, 2016), which has proven to be robust (Elith et al., 2006). The algorithm searches a combination of variables with maximum entropy and estimates the importance of each in species distribution with respect to sites where the species was recorded (Elith et al., 2011). Presence records were obtained from the database provided by the Instituto de Biología of the Universidad Nacional Autónoma de México. The database was complemented with records from the Global Biodiversity Information Facility (www.gbif.org). A total of 64,773 records were compiled, from which doubtful records or those with spatial redundancy were removed, resulting in 11,701 records. The variables for predicting the species distribution included those that limit their presence at regional scales, such as temperature, precipitation, and elevation, and those variables related to niche requirements such as type of vegetation and a regionalization variable, at a spatial resolution of 0.0083º (~0.85 km2; see table 1s in supplementary material). Climatic factors (such as temperature and precipitation) were used as direct variables that have physiological importance for bat species, but are not consumed; vegetation cover variables (such as trees or herbaceous plants) were related to resources used directly (López–González et al., 2011). Elevation was an indirect ecological variable that has no direct physiological relevance for species' persistence in response to future changes in environmental conditions (Guisan and Zimmermann, 2000), while the categorical variable of regionalization represents unique fauna and flora

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of the world’s continents (Olson et al., 2001). The suite of predicted variables allows us to evaluate bat species' responses to the environmental gradient present in Mexico. The layers of climatic variables were modified to generate new layers (see table 2s in supplementary material) using the Image Calculator in IDRISI (edition Selva; Eastman, 2012). However, using a set of variables is biased to autocorrelation (Peterson et al., 2011) and so to reduce information redundancy, we deleted continuous variables with a high correlation (r2 > 0.8) and performed a cluster analysis (1–Pearson r distance and Ward algorithm). The cluster analysis was performed with a sample of 500 random points, generated in ArcGIS (ver. 10.1; ESRI, 2010) using the Random Points Extension. The number of groups of correlated variables was defined from the amalgamation graph (threshold value = 1.1). From each group, we chose one variable and excluded the rest (see fig. 1s in supplementary material). Potential distribution models were performed with one categorical and 10 continuous low correlated variables to represent the environmental heterogeneity in the country (table 1). Models were generated only for species with a minimum of 10 records (53 species; table 2), since MaxEnt is stable with this number of records (Wisz et al., 2008). Default MaxEnt settings were used, with 75 % of the records used to create the model and 25 % to test it. Due to this random component, 10 single models were generated for each species to compile a consensus map via the weighted average method (Marmion et al., 2009). This method provides a continuous map and to obtain the species presence/absence, we calculated the average occurrence threshold from the single models. The threshold was the fixed cumulative value for 10 % probability of occurrence. Species were considered present in values above the threshold and absent in values below the threshold. Thus, a binary consensus map was obtained, increasing accuracy and decreasing the uncertainty of single models. Species richness variation The spatial pattern of richness for phyllostomids in Mexico was obtained from the sum of the consensus presence/ absence (binary) models for each species in IDRISI (Eastman, 2012). Additionally, we generated a richness map for trophic guild classification (Giannini and Kalko, 2004; table 2). Richness maps show the variation in species richness and spatially to identify highly specific richness areas. Sites with similar species composition Binary maps were used to define sites with a similar species composition. We used 500 random points (sites) to extract the presence/absence from each model and to compile a binary matrix that contained all the possible species combinations. We defined groups of sites with similar species composition using a generalized k–means analysis, which finds the optimum 'partition' for dividing a number of objects into k clusters in

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function of categorical variables (presence–absence) in STATISTICA (StatSoft, 1984–2013). The used k groups were defined from the amalgamation graph of a cluster analysis (Sorensen–Dice distance and Ward algorithm). The analysis searches for a combination of sites that maximize significant differences in local species composition between groups based on individual x2–tests for each species. The null hypothesis was that the frequency of sites per group was similar at a probability of 0.05. The result was the assignment of each site in a group as a function of local species composition. We present a map that shows the spatial tendency in the geographical distribution of groups of sites that differ in bat species composition. In addition, we identified species that distinguish each group based on their proportional distribution affinity, obtained as the percentage of sites by group where each species occurs. The affinity values (from 0 to 100 %) indicate the species geographical distribution along the different groups of sites. Species environmental segregation Bat community–level response to environmental conditions was identified via the outlying mean index (OMI) to estimate the preference (marginality) and the amplitude (tolerance) of conditions used by each species (Dolédec et al., 2000). The analysis estimates the distance between the average conditions (centroid) in which each species was present with respect to the average conditions of the study area (center of gravity). The OMI generates canonical variables that account for the highest variability of sites, which were associated with sites where species were recorded by a canonical correspondence. As a result, the variability (inertia) associated with the species distribution is decomposed in three components (Dolédec et al., 2000): (1) marginality or OMI, which is the deviation of average conditions used for the species in the study area. High values for species were found in different conditions of the average evaluated and low values for species were found in the average; (2) tolerance, which is the number of sites with which species are associated and their location in an environmental gradient (i.e., niche breadth). Low values imply that a species occurs in a narrow range of conditions (i.e., specialist) and high values imply that a species occurs in a wide range (i.e., generalist); and (3) residual tolerance, which is the variation in species occurrence not explained by the variables. The OMI analysis was performed in ADE–4 software (Thioulouse et al., 1997) using a faunal matrix with species presence/ absence and an environmental matrix with environmental characteristics; both were extracted from the 500 sites. The faunal matrix was used in the k–means analysis, while the environmental matrix included values of the 10 continuous variables used to generate the models (table 1). Additionally, we included a new column in the faunal matrix containing points where species absence was predicted. The column identifies environmental conditions that limit the species distribution. Statistical significance was estimated with a Monte Carlo permutation test

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Table 1. Variables used to generate species distribution models: Cl, climatic; Tp, topographic; Lc, land cover; Ct. categorical; SD, standard deviation. (Measurement units for each variable type in parentheses). Tabla 1. Variables usadas para generar los modelos de distribución de las especies: Cl, climática; Tp, topográfica; Lc, cubierta de tierra; Ct, categórica; SD, desviación estándar). (Unidades de medida para cada tipo de variable entre paréntesis).

Variables

Code

Maximum of yearly maximum temperatures (ºC) MXXT

Type Source

Mean

Cl www.worldclim.org

22.9

4.6

Mean yearly minimum temperatures (ºC)

MMNT Cl www.worldclim.org

12.1

4.9

Percentage of precipitation in March

PP03

Cl www.worldclim.org

6.4

2.8

Percentage of precipitation in May

PP05

Cl www.worldclim.org

7.7

4.2

Percentage of precipitation in July

PP07

Cl www.worldclim.org

14.3

6.3

Mean yearly evapotranspiration (mm)

MMEV Cl www.worldclim.org

1,562.1 174.1

Elevation (m a.s.l.)

ELEV

Percentage of bare cover

BARP Lc www.glcf.umd.edu

18.5

27.1

Percentage of tree cover

TREP

Lc www.glcf.umd.edu

23.4

24.4

Percentage of herbaceous cover

HERP Lc www.glcf.umd.edu

57.1

22.6

Ecoregions

ECOR Ct Olson et al., 2001

–

(1,000 permutations; Metropolis and Ulam, 1949). We report eigenvalues, factor loadings, and a graph of species centroids. To identify bat species that tend to co–occur in similar sites, we grouped the species according to the environmental conditions where their presence was predicted. The grouping was conducted with a Q agglomerative analysis for the faunal matrix via cluster analysis (Sorensen–Dice distance and Ward algorithm). The groups found are highlighted on the species marginality graph to determine whether a tendency exists for species to be homogeneously dispersed within the available environmental gradient. All cluster analyses were performed in STATISTICA (ver. 12; StatSoft, 1984–2013). Species incidence with anthropic changes The degree to which species occurrence overlapped with human activities was analyzed by estimating the percentage of area in which a species distribution converges with modified areas. Modified areas were obtained using a layer of land use/vegetation in Mexico (INEGI serie V; www.inegi.org.mx). The layer was reclassified into cropland, livestock, and urban zones. We obtained the percent of overlap between family and trophic guild richness using the classified land–use layer. A convergence percentage was calculated from the ratio of overlain pixels to the total pixels in the study area, multiplied by 100 (Venegas–Barrera and Manjarrez, 2011). Additionally, we present the convergence percentage by richness proportions for the family and guilds in groups of sites where species diversity was highest.

Tp https://lta.cr.usgs.gov/hydro1k 744.6 716.2

Results Spatial patterns in phyllostomid bat richness The Phyllostomidae family is potentially distributed in 79.6 % of Mexico’s land surface, and we found that the highest richness areas were in the southern tropical environments (> 40 spp., fig. 1A). Nectarivorous bats were the guild with the widest geographic distribution (73.9 %), whereas hematophagous species were found in 46.2 % of Mexico and frugivorous and insectivorous bats occurred in 51.5 % and 59 % of Mexico, respectively. Nectarivorous bats showed a potential highest richness on the Pacific coast, while the other guilds had a higher richness on the Atlantic coast (fig. 1B–1E). Richness for all groups was lowest in the northern arid environments. Groups of sites with similar phyllostomid bat composition We found six groups of sites that contained a distinctive species (threshold value of 1.5 unities; fig. 2): (1) The Atlantic group, on the coast of the Gulf of Mexico and the Yucatan Peninsula, had the highest potential richness areas (10 to 49 species) with Mimon crenulatum, Lophostoma brasiliense, and Vampyrum spectrum as higher proportionality affinity species (table 2). (2) The Pacific group, located on the Pacific coast and in the Balsas Basin, had areas with potentially 15 to 45 species, such as Uroderma magnirostrum, Musonycteris harrisoni, and Glossophaga morenoi. (3) The Mountain group, in the mainly mountainous ranges of Mexico (Sierra Madre Oriental, Sierra Madre Occidental, Sie-

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Table 2. Phyllostomid bat species included in the study. Trophic guilds (TG: Hem, hematophagous; Nec, nectarivorous; Ins, insectivorous; Fru, frugivorous). Proportional distribution affinity by group of sites (At, Atlantic; P, Pacific; M, mountain; Pl, plateau; Ar, arid; D, desert). Listed with conservation status in: a NOM–059–SEMARNAT (SEMANRNAT, 2010); b IUCN Red List (www.iucnredlist.org). Nicho parameters: OMI, outlying mean index; Tol, tolerance; TolR, residual tolerance. Tabla 2. Especies de murciélagos filostómidos incluidas en el estudio. Gremios tróficos (TG: Hem, hematófago; Nec, nectarívoro; Ins, insectívoro; Fru, frugívoro). Afinidad de distribución proporcional por grupo de sitios (At, Atlántico; P, Pacífico; M, montaña; Pl, altiplano; Ar, árido; D, desierto). Incluida con categoría de conservación en: a NOM–059–SEMARNAT (SEMANART, 2010); b Lista Roja de la UICN (www.iucnredlist.org). Parámetros del nicho: OMI, índice de marginalidad media; Tol, tolerancia; TolR, tolerancia residual.

Proportional distribution affinity

Species

Code TG At

OMI Components

Pl Ar D Inertia OMI Tol TolR

Mimon crenulatum a

Micre Ins 100 0 0 0 0 0 12.45 9.44 0.89 2.12

Lophostoma brasiliense a

Lobra Ins 97.4 0 2.6 0 0 0 11.23 7.29 1.09 2.85

Vampyrum spectrum

a,b

Mimon cozumelae a

Vaspe Ins 90.5 5.7

0 3.9 0 9.39 6.43 0.21 2.75

Micoz Ins 90.3 6.3

0 1.4 0 11.59 7.72 0.98 2.89

Artibeus watsoni

Arwat Fru 90.3 0 5.6 0 4.1 0 7.52 4.47 0.38 2.67

Phylloderma stenops a

Phste Ins 85.5 1.8 11.5 0 1.2 0 11.52 7.20 1.38 2.94

Lonchorhina aurita a

Loaur Ins 84.8 11.3 0

Macrophyllum macrophyllum a Mamac Ins 82.0 13.5 0

0 3.9 0 11.84 8.12 0.67 3.05 0 4.6 0 9.72 6.92 0.18 2.63

Micronycteris schmidtorum a

Misch Ins 77.8 11.6 7.2 2.1 1.3 0 11.07 7.01 1.33 2.73

Vampyrodes major

Vamaj Fru 77.6 13.8 8.6 0 0 0 6.96 3.57 0.29 3.11

Lampronycteris brachyotis a

Labra Ins 77.1 16.8 3.5

0 2.5 0 11.48 7.42 0.92 3.13

Lophostoma evotis a

Loevo Ins 76.7 21.7 0

0 1.6 0 13.00 7.97 0.50 4.54

Chrotopterus auritus a

Chaur Ins 75.1 4.5 16.7 1.6 1.0 1.2 11.30 5.47 1.70 4.13

Carollia perspicillata

Caper Fru 74.5 14.2 9.3

0 1.9 0 10.25 6.12 0.75 3.38

Tonatia saurophila

Tosau Ins 74.5 20.7 0

0 4.7 0 10.85 7.53 0.39 2.94

Urbil Fru 73.1 21.6 3.1

0 2.3 0 11.06 6.78 0.95 3.34

Diaemus youngi

Diyou Hem 67.5 18.4 9.4

0 3.4 1.3 10.09 5.98 0.50 3.62

Carollia sowelli

Casow Fru 66.0 13.2 19.2 1.6 0

Uroderma bilobatum a

0 9.79 4.60 1.07 4.12

Trachops cirrhosus

Trcir Ins 65.2 23.8 8.6

Phyllostomus discolor

Phdis Ins 59.4 32.6 5.4 0 2.6 0 9.48 5.33 0.74 3.41

Diphylla ecaudata

Dieca Hem 57.1

Chiroderma villosum

Chvil Fru 57.0 28.8 12.0 0 2.2 0 10.28 5.50 0.93 3.85

0 2.5 0 10.60 5.92 1.05 3.64

0 37.0 5.1 0.8 0 9.22 3.44 1.98 3.81

Vampyressa thyone

Vathy Fru 50.6 48.0 0

Platyrrhinus helleri

Plhel Fru 46.8 30.5 20.8 0 1.9 0 7.92 3.90 0.46 3.56

0 1.4 0 10.40 5.69 0.60 4.11

Artibeus phaeotis

Arpha Fru 46.1 37.8 6.2 2.0 7.1 0.7 11.14 5.48 0.99 4.68

Micronycteris microtis

Mimic Ins 42.5 31.5 22.3 1.6 2.1 0 9.52 3.98 1.17 4.37

Choeroniscus godmani

Chgod Nec 41.5 34.0 13.5 3.3 7.7 0 9.75 4.28 0.98 4.49

Centurio senex

Cesen Fru 36.2 31.0 18.9 8.1 5.1 0.7 9.83 3.03 1.54 5.26

Glossophaga leachii

Gllea Nec 34.4 32.1 25.1 1.6 6.2 0.6 9.39 3.53 1.18 4.67

Artibeus jamaicensis

Arjam Fru 32.6 31.6 24.9 4.7 5.6 0.5 9.08 3.28 1.16 4.64

Glossophaga soricina

Glsor Nec 31.5 27.8 25.5 9.3 5.5 0.5 8.83 2.76 1.26 4.81

Artibeus lituratus

Arlit Fru 31.3 29.9 26.9 6.3 5.6 0 8.68 2.61 1.24 4.83

Carollia subrufa

Casub Fru 30.6 49.0 15.3 1.7 3.3 0 9.43 3.66 1.38 4.39

Sturnira lilium

Stlil Fru 28.4 28.5 27.8 9.8 5.1 0.4 8.73 2.17 1.19 5.37

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Table 2. (Cont)

Proportional distribution affinity P

OMI Components

Species

Code TG At

Hylonycteris underwoodi

Hyund Nec 24.1 24.5 39.1 5.3 4.2 2.9 7.43 1.59 0.38 5.47

Pl Ar D Inertia OMI Tol TolR

Desmodus rotundus

Derot Hem 23.4 24.9 25.7 22.0 3.7 0.4 8.53 1.54 1.07 5.92

Glyphonycteris sylvestris

Glsyl Ins 23.0 41.2 22.2 7.5 5.4 0.7 9.21 2.53 1.41 5.27

Glossophaga commissarisi

Glcom Nec 19.9 50.7 16.6 6.3 5.6 0.9 9.74 3.23 1.77 4.74

Glossophaga morenoi

Glmor Nec 17.5 76.8 0

0 5.7 0 10.01 6.30 0.89 2.82

Artibeus toltecus

Artol Fru 14.2 32.5 36.5 10.8 5.4 0.6 7.72 1.72 0.45 5.55

Sturnira ludovici

Stlud Fru 13.8 26.1 41.4 14.0 2.8 1.9 7.29 1.26 0.20 5.83

Anoura geoffroyi

Angeo Nec 13.3 30.7 38.4 14.3 2.7 0.6 7.72 1.35 0.19 6.17 Chsal Fru 13.3 30.7 29.3 15.8 6.8 4.1 8.57 1.61 0.44 6.52

Chiroderma salvini Enchisthenes hartii

Enhar Fru 13.0 30.4 38.9 10.9 2.5 4.3 8.43 1.57 0.19 6.67

Leptonycteris yerbabuenae

a,b

Arazt Fru 7.1 18.1 47.3 22.4 1.3 3.7 7.62 1.96 1.02 4.64

Artibeus aztecus Leptonycteris nivalis

Leyer Nec 7.8 23.4 24.1 25.6 14.7 4.3 7.50 0.90 1.33 5.28

a,b

Leniv Nec 5.2

Choeronycteris mexicana

a,b

Chmex Nec 2.5 16.1 22.2 24.4 17.4 17.4 7.83 0.26 1.30 6.27 Arhir Fru 1.7 33.7 20.8 20.9 19.4 3.5 7.88 2.98 0.86 4.04

Artibeus hirsutus Uroderma magnirostrum Musonycteris harrisoni

20 23.0 26.8 7.2 17.9 7.33 0.66 0.44 6.23

a,b

Urmag Fru 0

94.2 0

0 5.8 0 12.46 9.91 0.33 2.22

Muhar Nec 0

85.4 0

0 14.6 0 12.17 9.70 0.59 1.88

Macrotus waterhousii

Mawat Ins 0 40.5 24.7 22.2 10.3 2.2 7.71 2.26 1.28 4.18

Macrotus californicus

Macal Ins 0

rra de Chiapas, and Eje Neovolcánico Transversal), showed a richness area with 10 to 35 species, such as Artibeus aztecus, Sturnira ludovici and Hylonycteris underwoodi. (4) The Plateau group, in the inner slopes of the Neovolcanic Belt in the southern part of the Mexican plateau, had areas with up to 20 species, such as Leptonycteris nivalis, Leptonycteris yerbabuenae and Choeronycteris mexicana. (5) The Arid group, in the Sonoran Desert, included areas with one to 10 species where Macrotus californicus was distinctive. (6) The Desert group, in a portion of the Chihuahuan Desert, had areas where one to six species occurred, such as Choeronycteris mexicana and Leptonycteris nivalis. In a seventh group, in a large part of the Chihuahuan Desert, no species were observed (i.e., the Absence group Environmental segregation of phyllostomid bats The multidimensional distances of environmental conditions in which all species were recorded are significantly different from the study area average (p < 0.05); that is, at least one species occurred in different environmental conditions that averaged all sites (table 2). The first two axes of the OMI analysis accounted for 92.2 % of the total variance. The first axis explained 81.4 % of the variation, mainly due to variations in the yearly mean minimum temperature and the tree–cover percentage.

8.7 3.5 13.4 70.5 3.9 10.03 2.43 0.65 6.95

The second axis explained 10.8 % of the variation due to variations in the precipitation percentage in July and the herbaceous cover percentage (fig. 3A). The first OMI axis separates the environments where the family occurs from those where it is absent. Phyllostomid bats were present in environments that have a yearly minimum mean temperature above 13.5 ºC (from 12 to 20 ºC; fig. 3B) and a mean tree–cover percentage higher than 24 % (from 20 to 80 %; fig. 3C). The second axis revealed an environmental gradient associated with variations in precipitation in which the species are distributed. Most species (e.g., Mimon crenulatum, Phylloderma stenops) occurred in environments where six to 12 % of the annual rainfall occurs in July (fig. 3D) and the herbaceous cover percentage is from 20 to 40 % (fig. 3E). Conversely, several species (e.g., Choeronycteris mexicana, Macrotus californicus) occurred in environments where 12 to 18 % of the annual precipitation occurs in July and the herbaceous cover constitutes 40 to 60 %. We found three groups of bat species that occurred in similar sites (threshold value of 2.5 units, fig. 3F). The groups were differentiated according to species distribution range: (1) Widespread distribution (across 17 to 51 % of Mexico’s surface) was found throughout the entire study area, occurring in groups in all sites. This species group was composed of 24 species with mainly frugivorous or nectarivorous bats (e.g.,

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148

Species 0 1–7 8–14 15–21 22–28 29–35 36–42 43–49

N B

Species 0 1 2 3

Species 0 1–4 5–8 9–12 13–17

0 150 300 600 km Species 0 1–2 3–4 5–6 7–8 9–11

Species 0 1–4 5–8 9–12 13–16 17–20

Fig. 1. Spatial richness pattern in Mexico: A, Phyllostomidae family; B, hematophagous bats; C, nectarivorous bats; D, insectivorous bats; E, frugivorous bats. Fig. 1. Patrón espacial de riqueza en México: A, familia Phyllostomidae; B, murciélagos hematófagos; C, murciélagos nectarívoros; D, murciélagos insectívoros; E, murciélagos frugívoros.

Sturnira lilium, Leptonycteris yerbabuenae) and one hematophagous species (i.e., Desmodus rotundus). (2) Regional distribution (in eight to 21%) was associated with the Neotropical region, with a limited presence in mountain systems, mainly present in the Atlantic group. This species group was composed of 18 species, mainly insectivorous bats (e.g.,

Micronycteris microtis, Phylloderma stenops) and the remaining two hematophagous species. (3) Narrow distribution (in three to 15 %) had an endemic or restricted distribution in the Atlantic or Pacific groups. This species group had 11 species (e.g., Uroderma magnirostrum, Musonycteris harrisoni): six frugivorous, three insectivorous, and two nectarivorous species.

Animal Biodiversity and Conservation 41.1 (2018)

149

Linkage distances

Groups Atlantic Pacific Mountain Plateau Arid Desert Absence

6 3 N 0

0 150 300 600 km

Fig. 2. Groups of sites with a distinctive phyllostomid bat species composition. Fig. 2. Grupos de sitios con una composición distintiva de especies de murciélagos filostómidos.

Overlap between phyllostomid bat distribution and human modified areas Native/secondary vegetation covers 72.2 % of Mexico, and the remaining 27.8 % has been human–modified for crops (17 %), livestock (9.9 %) and urban zones (0.9 %). We found 30.9 % of the entire Phyllostomidae family spatial distribution overlapped with human–modified areas (table 3). Cropland was the most broadly distributed cover type converging with the family and the four feeding guilds. In the Atlantic group, the high richness areas of the Phyllostomidae family and the guilds of hematophagous, insectivorous and frugivorous bats overlapped with livestock (fig. 4). In the Pacific group, the high richness areas for nectarivorous bats converged similarly with all types of human changes. Discussion In this study, we provide a perspective about phyllostomid bat distribution in Mexico. We identified six site groups with distinctive species assembly, three species groups that co–occurred in similar sites, environmental tolerances of each species, and the degree of overlap with human activities. The distribution of phyllostomid bats showed a spatial richness pattern that related negatively to latitude, decreasing the number of species from tropical environments in southern to arid environments in northern Mexico (fig. 1A; Stevens, 2006). The highest species richness areas were in the Tehuantepec Isthmus region, except for nectarivorous bats (figs. 1B–1E). The latitudinal pattern has been reported for groups

such as reptiles (Ochoa–Ochoa and Flores–Villela, 2006), birds (García–Trejo and Navarro, 2004) and terrestrial mammals (Escalante et al., 2007). The

Table 3. Proportional convergence between phyllostomid bat distribution and human–modified areas in Mexico: Cr, cropland; Lv, livestock; Urb, urban; Phy, Phyllostomidae family; Hem, hematophagous bats; Nec, nectarivorous bats; Ins, insectivorous bats; Fru, frugivorous bats. Tabla 3. Convergencia proporcional entre la distribución de murciélagos filostómidos y las zonas modificadas por el hombre en México: Cr, cultivo; Lv, ganado; Urb, urbano; Phy, familia Phyllostomidae; Hem, murciélagos hematófagos; Nec, murciélagos nectarívoros; Ins, murciélagos insectívoros; Fru, murciélagos frugívoros.

Anthropic activities Cr

Urb Total

Phy 18.97 11.04 0.98 30.99 Hem 23.39 15.13 1.21 39.73 Nec 19.81 11.56 1.07 32.44 Ins 18.91 13.15 0.97 33.03 Fru 21.93 14.52 1.03 37.48

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0.4 –0.9 0.8 –0.3

PP07 HERP

ELEV

MXXT

MMEV

Loadings Variables Axis 1 Axis 2 ELEV 0.467 0.211 TREP –0.790 –0.078 MMEV –0.345 0.102 MMNT –0.811 –0.076 MXXT –0.710 0.100 BARP 0.628 –0.191 PP07 0.083 0.312 PP05 –0.150 –0.129 PP03 0.249 –0.231 HERP 0.051 0.308

MMNT TREP

PP05 BARP PP03 MMNT

0–20 20–40 40–60 60–80 80–100

PP07

TREP

0–4 4–8 8–12 12–16 16–20

Linkage distances

0–6 6–12 12–18 18–24 24–30

HERP

0–20 20–40 40–60 60–80

Fig. 3. Factor loadings for environmental variables along the two axes of the OMI analysis (3A), showing the most important occurrences of phyllostomid bats in Mexico (3B–3E). The cluster identified species that have similar distribution ranges (3F): widespread (●), regional (■), and narrow (▲). (For abbreviations see tables 1 and 2). Fig. 3. Carga de factores para las variables ambientales en los dos ejes del análisis OMI (3A), donde se muestran las presencias más importantes de filostómidos en México (3B–3E). El clúster identifica las especies que presentan un rango de distribución similar (3F): amplio (●), regional (■) y estrecho (▲). (Para consultar las abreviaturas, véanse las tablas 1 y 2).

Animal Biodiversity and Conservation 41.1 (2018)

43–47

36–42

29–35

22–28

8–14

15–21

17–20 9–11

13–16

9–12

5–8

13–16

9–11

7–8

5–6

3–4

1–2

9–11

0 7–8

0 5–

3–4

1–2

7–8

17–20

13–16

9–12

5–8

1–4

17–20

0 13–16

9–12

5–8

Species richness

9–12

5–6

13–17

9–12

5–8

1–4

13–17

9–12

5–8

1–4

5–8

3–4

1–4

1–7

43–49

36–42

29–35

22–28

8–14

1–2

50 Fru

1–4

Ins

Nec

Proportional incidence

Urban

Hem

15–21

1–7

0 43–48

36–42

29–35

22–28

15–21

Livestock 50

1–7

Cropland 40

8–14

Phy

151

Fig. 4. Proportional convergence between phyllostomid bat richness and human–modified areas in groups of sites with the highest richness: the Atlantic group for the Phyllostomidae family (Phy) and hematophagous (Hem), insectivorous (Ins), and frugivorous (Fru) bats; and the Pacific group for nectarivorous bats (Nec). Fig. 4. Convergencia proporcional entre la riqueza de especies de filostómidos y las zonas modificadas por el hombre en los grupos de sitios con la mayor riqueza: grupo del Atlántico para murciélagos de la familia Phyllostomidae (Phy) y para murciélagos hematófagos (Hem), insectívoros (Ins) y frugívoros (Fru); y el grupo del Pacífico para murciélagos nectarívoros (Nec).

identified groups of sites with distinctive bat species composition (fig. 2) showed a specific occurrence along with temperature and vegetation cover gradient (mainly tree cover; fig. 3; López–González et al., 2011) in relation to specific environmental tolerance.

The species occurrences with common origin and differential environmental preferences in the same area are related to their evolutionary histories leading to their diversification and current distributions (Dumont et al., 2012).

152

The groups of sites identified as the Neotropic region, the Atlantic and Pacific groups, present distinctive environments (i.e., tropical and subtropical) and are composed of ancestral bat species, such as Micronycteris or Diphylla genus. The Atlantic group containe mostly insectivorous species from the subfamily Phyllostominae (e.g., Mimon crenulatum and Lophostoma brasiliense), with a high marginality (mean OMI = 6; table 2) that occur in specific environmental conditions, such as perennial forests. Insectivorous bats are narrowly distributed (fig. 3) and have been associated with large forest fragments, where a high availability of resources is found (e.g., food and roost; Medellín et al., 2000). On the other hand, the Pacific group is composed principally of frugivorous and nectarivorous species with a narrow or widespread distribution (e.g., Musonycteris harrisoni and Chiroderma salvini), in deciduous forests. However, the species has a higher tolerance (mean Tol ~1) than species in the Atlantic group, allowing their persistence when resources are scarce through the variation in seasons found in deciduous environments (Chávez and Ceballos, 2001). The groups of sites identified in the limits of the Neotropic region or in the Nearctic region present distinctive environments (i.e., temperate, arid) and are composed of the phylogenetically derived bat species, such as Artibeus and Leptonycteris genus (fig. 2). In the Mountain and Plateau groups, an elevational gradient derived from the mountains promotes variable climatic conditions (i.e., temperature and humidity) that results in different vegetation types (e.g., temperate forests and grasslands; Ramanoorthy et al., 1998). The groups are composed of frugivorous and nectarivorous species (e.g., Artibeus aztecus and Leptonycteris nivalis) that have low marginality (mean OMI < 1.3, table 2). However, the species are distributed widely (fig. 3) due to adaptations such as dietary changes and migratory movements that allow them to colonize temperate environments (Fleming et al., 2009). In Northern Mexico, the species most commonly present in the Arid and Desert groups are Macrotus californicus and Leptonycteris nivalis, species that have a widespread distribution and that are tolerant to average conditions in Mexico (mean Tol = 0.5). The species that occur in semi–arid environments, such as Macrotus californicus, tolerate desert conditions through behavioral and physiological adaptations (Bell et al., 1986), while Leptonycteris nivalis uses migration and mutualism with scrubland xeric plants (Fleming et al., 2009). The persistence of these species in arid zones is high. These zones are the best conserved ecosystems in Mexico because the low availability of water limits the establishment and growth of human populations (Challenger and Dirzo, 2009). The potential distribution of the Phyllostomidae family in Mexico showed a 30 % overlap with areas modified by human activities (table 3). The percent of overlap with the modified areas varied among feeding guild species; for example, convergence of higher richness of hematophagous and nectivorous species with modified areas differed with insectivorous guild, so they may be affected differentially (fig. 4). Frugivo-

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rous bats will be less vulnerable to human activities because they have a higher niche tolerance and a lower marginality (e.g., genus Sturnira and Artibeus, table 2). These species are generalist, and they exploit a broad amplitude of resources in both preserved and disturbed environments (Klingbeil and Willig, 2009). Common vampire bats (Desmodus rotundus) have a negative interaction with humans because they cause economic loss to the beef industry and play a role in the epidemiology of bovine paralytic rabies in rural areas (Dantas–Torres, 2008). However, future climate projections predict an increase in the distribution of D. rotudus (Lee et al., 2012). In contrast, other endangered hematophagous species (i.e., Diphylla ecaudata and Diaemus youngi) that do not interfere with human activities are under risk through population control of the common vampire bat (Vercauteren et al., 2012). Insectivorous bats (e.g., Micronycteris microtis and Mimon cozumelae) are vulnerable to agricultural practices because they show a low tolerance and a high marginality (Medellín et al., 2000) and because they converge with croplands where humans regulate pests with chemicals that are toxic to bats (table 3; Lawer and Darkoh, 2016). The nectarivorous bats, due to their widespread distribution and high richness areas, overlap with human–modified areas (fig. 4) and their distribution can indirectly affect their vital biological processes, such as the loss of reproduction areas, mutualistic associations and modified migratory movements (Fleming et al., 2009). The spatial–environmental distribution of the Phyllostomidae family in Mexico reveals a species response to niche requirements and human changes. For example, species in homogeneous environments (e.g., tropical forests) have smaller niche breadth and narrow distributions, being more sensitive to environmental changes (Brown, 2014). On the other hand, species in heterogeneous environments (e.g., deserts) have broad niche breadth and widespread distributions with tolerance to disturbance (Krebs, 2001). In the country, the highest native cover loss occurs in tropical forests of the Atlantic and Pacific coastal plains, because low slopes and precipitation provide optimal conditions for the development of agricultural activities (Challenger and Dirzo, 2009), while the human population is concentrated mostly in Meseta Central (Klein–Goldewijk and Ramankutty, 2004), areas where phyllostomid bat richness is also high (fig. 1). Medium (2023) and long–trend (2033) scenarios predict that if conservation actions are not implemented, an unstable to very critical scenario will occur in these areas (Sánchez–Salazar et al., 2013); in this sense, Protected Areas play a fundamental role in conserving biodiversity (Myers et al., 2000). Mexico has decreed 181 Protected Areas (www. conanp.gob.mx) to maintain the integrity of ecosystems and environmental services. The decrees are supported by laws that regulate anthropogenic activities. However, most of these areas are surrounded by modified zones, are in mountainous regions, do not include high diversity areas, or show some degree of deterioration (Fuller et al., 2006). In the case of phyllostomid bats, nectarivorous and insectivorous bats

Animal Biodiversity and Conservation 41.1 (2018)

are susceptible to constant environmental degradation caused by humans. Thus, conservation strategies may complement the surface of protected areas through the delimitation of areas with suitable environmental conditions to promote the persistence and flow of biodiversity (Nori et al., 2016). The approach used in this work can be extrapolated to other regions and taxa, which in turn will provide a better understanding of regional patterns of richness and species composition (Peterson et al., 2015). Our results showing the areas with the highest bat richness may be useful to propose biological corridors and conservation priority areas using approaches such as systematic conservation planning (Margules and Sarkar, 2009). Acknowledgements We are grateful to the Consejo Nacional de Ciencia y Tecnología for a doctoral fellowship (350170) to JCAF. The authors also wish to thank four anonymous reviewers who provided helpful comments on early drafts of the manuscript. References Bell, G. P., Bartholomew, G. A., Nagy, K. A., 1986. The roles of energetics, water economy, foraging behavior, and geothermal refugia in the distribution of the bat Macrotus californicus. Journal of Comparative Physiology, 156: 441–450. Brown, J. H., 2014. Why are there so many species in the tropics?. Journal of Biogeography, 41: 8–22. Challenger, A., Dirzo, R., 2009. Factores de cambio y estado de la biodiversidad. In: Capital natural de México, Vol. I: 37–73 (R. Dirzo, R. González, J. I. March, Eds.). CONABIO, México. Chávez, C., Ceballos, G., 2001. Diversidad y abundancia de murciélagos en selvas secas de estacionalidad contrastante en el oeste de México. Revista Mexicana de Mastozoologia, 5: 27–44. Dantas–Torres, F., 2008. Bats and their role in human rabies epidemiology in the Americas. Journal of Venomous Animals and Toxins including Tropical Diseases, 14: 193–202. Dolédec, S., Chessel, D., Gimaret–Carpentier, C., 2000. Niche separation in community analysis: a new method. Ecology, 81: 2914–2927. Dumont, E. R., Dávalos, L. M., Goldberg, A., Santana, S. E., Rex, K., Voigt, C. C., 2012. Morphological innovation, diversification and invasion of a new adaptive zone. Proceedings of the Royal Society B, 279: 1797–1805. Eastman, J. R., 2012. IDRISI Selva Tutorial, Manual Version 17.0. Clark University. Available online at: http://uhulag.mendelu.cz/files/pagesdata/eng/gis/ idrisi_selva_tutorial.pdf Elith, J., Graham, C. H., Anderson, R. P., Dudik, M., Ferrier, S., Guisan A, Hijmans, R. J., Huettmann, F., Leathwick, J. R., Lehmann, A., Li, J., Lohmann, L. G., Loiselle, B. A., Manion, G., Moritz, C., Nakamura, M., Nakazawa, Y., Mc Overton, J. C. M.,

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Supplementary material

Table 1s. List of original environmental variables. Tabla 1s. Lista de las variables ambientales originales.

Variable

Code Type Source

Maximum temperature in January

XT01

Climatic

www.worldclim.org

Maximum temperature in February

XT02

Climatic

www.worldclim.org

Maximum temperature in March

XT03

Climatic

www.worldclim.org

Maximum temperature in April

XT04

Climatic

www.worldclim.org

Maximum temperature in May

XT05

Climatic

www.worldclim.org

Maximum temperature in June

XT06

Climatic

www.worldclim.org

Maximum temperature in July

XT07

Climatic

www.worldclim.org

Maximum temperature in August

XT08

Climatic

www.worldclim.org

Maximum temperature in September

XT09

Climatic

www.worldclim.org

Maximum temperature in October

XT10

Climatic

www.worldclim.org

Maximum temperature in November

XT11

Climatic

www.worldclim.org

Maximum yemperature in December

XT12

Climatic

www.worldclim.org

Minimum yemperature in January

NT01

Climatic

www.worldclim.org

Minimum temperature in February

NT02

Climatic

www.worldclim.org

Minimum temperature in March

NT03

Climatic

www.worldclim.org

Minimum temperature in April

NT04

Climatic

www.worldclim.org

Minimum temperature in May

NT05

Climatic

www.worldclim.org

Minimum temperature in June

NT06

Climatic

www.worldclim.org

Minimum temperature in July

NT07

Climatic

www.worldclim.org

Minimum temperature in August

NT08

Climatic

www.worldclim.org

Minimum temperature in September

NT09

Climatic

www.worldclim.org

Minimum temperature in October

NT10

Climatic

www.worldclim.org

Minimum temperature in November

NT11

Climatic

www.worldclim.org

Minimum temperature in December

NT12

Climatic

www.worldclim.org

Precipitation in January

PI01

Climatic

www.worldclim.org

Precipitation in February

PI02

Climatic

www.worldclim.org

Precipitation in March

PI03

Climatic

www.worldclim.org

Precipitation in April

PI04

Climatic

www.worldclim.org

Precipitation in May

PI05

Climatic

www.worldclim.org

Precipitation in June

PI06

Climatic

www.worldclim.org

Precipitation in July

PI07

Climatic

www.worldclim.org

Precipitation in August

PI08

Climatic

www.worldclim.org

Precipitation in September

PI09

Climatic

www.worldclim.org

Precipitation in October

PI10

Climatic

www.worldclim.org

Precipitation in November

PI11

Climatic

www.worldclim.org

Animal Biodiversity and Conservation 41.1 (2018)

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Table 1s. (Cont.)

Variable

Code Type Source

Precipitation in December

PI12

Climatic

www.worldclim.org

Evapotranspiration in January

EV01

Climatic

www.worldclim.org

Evapotranspiration in February

EV02

Climatic

www.worldclim.org

Evapotranspiration in March

EV03

Climatic

www.worldclim.org

Evapotranspiration in April

EV04

Climatic

www.worldclim.org

Evapotranspiration in May

EV05

Climatic

www.worldclim.org

Evapotranspiration in June

EV06

Climatic

www.worldclim.org

Evapotranspiration in July

EV07

Climatic

www.worldclim.org

Evapotranspiration in August

EV08

Climatic

www.worldclim.org

Evapotranspiration in September

EV09

Climatic

www.worldclim.org

Evapotranspiration in October

EV10

Climatic

www.worldclim.org

Precipitation in November

EV11

Climatic

www.worldclim.org

Climatic

www.worldclim.org

Precipitation in December

EV12

Elevation

ELEV Topographic https://lta.cr.usgs.gov/hydro1k

Aridity Index

ARIX

Land Cover

www.glcf.umd.edu

Bare cover percentage

BARP

Land Cover

www.glcf.umd.edu

Tree cover percentage

TREP

Land Cover

www.glcf.umd.edu

Herbaceous cover percentage

HERP

Land Cover

www.glcf.umd.edu

Evergreen vegetation cover percentage

PERP

Land Cover

www.glcf.umd.edu

Deciduous vegetation cover percentage

DECP

Land Cover

www.glcf.umd.edu

Broad–leaf vegetation cover percentage

BROP

Land Cover

www.glcf.umd.edu

Needle–leaf vegetation cover percentage

NEDP

Land Cover

www.glcf.umd.edu

Ecoregions

ECOR

Categorical

Olson et al. (2001)

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Table 2s. List of new variables obtained by climatic variables modification. Tabla 2s. Lista de nuevas variables obtenidas al modificar las variables climĂĄticas.

New variables

Code

Mean yearly maximum temperatures

MXMT

Standard deviation of yearly maximum temperatures

MXSD

Maximum of yearly maximum temperatures

MXXT

Maximum of maximum temperatures in first quarter

MX1Q

Maximum of maximum temperatures in second quarter

MX2Q

Maximum of maximum temperatures in third quarter

MX3Q

Maximum of maximum temperatures in fourth quarter

MX4Q

Maximum of yearly minimum temperatures

MXNT

Maximum of minimum temperatures in first quarter

MN1Q

Maximum of minimum temperatures in second quarter

MN2Q

Maximum of minimum temperatures in third quarter

MN3Q

Maximum of minimum temperatures in fourth quarter

MN4Q

Mean of yearly minimum temperatures

MMNT

Minimum of yearly maximum temperatures

MNXT

Minimum of maximum temperatures in first quarter

NX1Q

Minimum of maximum temperatures in second quarter

NX2Q

Minimum of maximum temperatures in third quarter

NX3Q

Minimum of maximum temperatures in fourth quarter

NX4Q

Minimum of yearly minimum temperatures

YNNT

Standard deviation of yearly minimum temperatures

MTSD

Minimum of minimum temperatures in first quarter

NN1Q

Minimum of minimum temperatures in second quarter

NN2Q

Minimum of minimum temperatures in third quarter

NN3Q

Minimum of minimum temperatures in fourth quarter

NN4Q

Total of yearly precipitation

YTPP

Mean of yearly precipitation

YEPP

Percentage of precipitation in January

PP01

Precipitation percentage in February

PP02

Percentage of precipitation March

PP03

Percentage of precipitation April

PP04

Percentage of precipitation May

PP05

Percentage of precipitation June

PP06

Percentage of precipitation July

PP07

Percentage of precipitation August

PP08

Percentage of precipitation September

PP09

Percentage of precipitation October

PP10

Percentage of precipitation November

PP11

Percentage of precipitation December

PP12

Mean of yearly evapotranspiration

MMEV

Standard deviation of yearly evapotranspiration

EVSD

Maximum of yearly evapotranspiration

MXEV

Minimum of yearly evapotranspiration

MNEV

Animal Biodiversity and Conservation 41.1 (2018)

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MNEV MXNT MN1Q MX1Q MX2Q MMEV MXMT MN4Q PP190 MMNT NX2Q YNNT NN1Q NX1Q NN4Q NN2Q MXEV MXXT MX3Q MN3Q MX4Q MNXT NX3Q NX4Q NN3Q EVSD MXSD MNSD BARP PP01 PP12 PP02 PP03 PP11 YTPP ARIX YEPP DECP BROP TREP PP04 PP05 PP06 PP09 HERP PP07 PP08 ELEV NEDP PERP 0

4 6 Linkage distance

Fig. 1s. Cluster analysis to select the uncorrelated environmental variables (underlined). Fig. 1s. Anรกlisis de conglomerados para seleccionar las variables ambientales no correlacionadas (subrayadas).

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Ethological uniqueness of a damselfly with no near relatives: the relevance of behaviour as part of biodiversity A. Cordero–Rivera, H. Zhang

Cordero–Rivera, A., Zhang, H., 2018. Ethological uniqueness of a damselfly with no near relatives: the relevance of behaviour as part of biodiversity. Animal Biodiversity and Conservation, 41.1: 161–174. Abstract Ethological uniqueness of a damselfly with no near relatives: the relevance of behaviour as part of biodiversity. Taxonomically isolated species may contribute unique characters to biological diversity, particularly at the level of ethodiversity. To test this idea, we analysed the territorial and reproductive behaviour of Pseudolestes mirabilis (Zygoptera, Pseudolestidae), an endemic damselfly from Hainan island, China, and the only representative of its family. Our hypothesis was that the uniqueness of this taxon would be evident in its behaviour. We found that the agonistic encounters between males were usually very short (less than 2 min) and consisted of a face–to–face display with both males maintaining a close distance while flying using only the forewings. No other odonate flies with only two wings in territorial contests. Furthermore, a small proportion of fights were escalated and lasted about one hour, with clear exhibition of the coloured hindwings. Males also confronted wasps (Eustenogaster nigra) that used the same microhabitat in a similar way, albeit for short time. Females were found in low numbers. This limited copulatory frequency and most males did not mate in the whole day. Unexpectedly for a damselfly with coloured wings, precopulatory courtship was almost absent, suggesting that intrasexual selection is behind the evolution of coloured wings in this species. Copulation lasted an average of seven minutes, with a first stage of rivals’ sperm removal (64 % of sperm removed) and a second stage of insemination. In agreement with our initial hypothesis, copulatory behaviour was unique: males did not translocate sperm to their vesicle before each mating but translocated sperm after copulation, a behaviour that cannot be easily explained. These exclusive characteristics point to the relevance of this species as an exceptional taxon that merits high conservation priority. Key words: Ethodiversity, Odonata, China, Pseudolestes mirabilis, Sexual selection, Sperm competition Resumen Singularidad etológica de una damisela sin parientes cercanos: la importancia del comportamiento como parte de la biodiversidad. Las especies taxonómicamente aisladas pueden aportar caracteres únicos a la diversidad biológica, particularmente en el ámbito de la diversidad etológica. Para comprobar esta idea, analizamos el comportamiento territorial y reproductor de Pseudolestes mirabilis (Zygoptera, Pseudolestidae), una damisela endémica de la isla de Hainan, en China, y el único representante de su familia. Nuestra hipótesis era que la singularidad de este taxón sería evidente en su comportamiento. Encontramos que los combates agonísticos entre machos fueron generalmente muy cortos (menos de 2 min) y consistieron en un vuelo de presentación cara a cara a una distancia cercana usando únicamente las alas anteriores. Ningún otro odonato vuela solo con dos alas en luchas territoriales. Además, una pequeña proporción de las peleas se intensificó y duró alrededor de una hora; en estas peleas los machos mostraron de forma evidente sus alas posteriores coloreadas. Los machos también se enfrentaron de manera similar, aunque durante poco tiempo, a las avispas (Eustenogaster nigra) que utilizaron el mismo microhábitat. El hecho de que se encontraran pocas hembras limitó la frecuencia de cópula, lo que determinó que la mayoría de los machos no se apareara durante todo el día. A diferencia de lo que cabría esperar para una libélula con alas coloreadas, el cortejo anterior a la cópula fue prácticamente inexistente, lo que sugiere que la selección intrasexual está detrás de la evolución de las alas coloreadas en esta especie. La cópula duró un promedio de siete minutos y estuvo compuesta por una primera etapa de eliminación del esperma de los rivales (64 % de esperma retirado) y una segunda ISSN: 1578–665 X eISSN: 2014–928 X

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etapa de inseminación. De acuerdo con nuestra hipótesis inicial, el comportamiento de cópula fue único: los machos no transfirieron el esperma a su vesícula antes de cada apareamiento, sino que lo hicieron después de la cópula, un comportamiento que no se puede explicar fácilmente. Estas características exclusivas apuntan a la importancia de esta especie como taxón excepcional que merece una alta prioridad de conservación. Palabras clave: Diversidad etológica, Odonata, China, Pseudolestes mirabilis, Selección sexual, Competencia espermática Received: 27 IV 17; Conditional acceptance: 3 VIII 17; Final acceptance: 8 VIII 17 Adolfo Cordero–Rivera, ECOEVO Lab, Escola de Enxeñaría Forestal, Campus Universitario, 36005 Pontevedra, Galiza, Spain.– Haomiao Zhang, State Key Lab. of Genetic Resources and Evolution, Kunming Inst. of Zoology, Chinese Academy of Sciences; and Kunming Natural History Museum of Zoology, Kunming Inst. of Zoology, Chinese Academy of Sciences, Yunnan, R. P. China. Corresponding author: A. Cordero–Rivera. E–mail: adolfo.cordeo@uvigo.es

Animal Biodiversity and Conservation 41.1 (2018)

Introduction There is an emerging consensus in the scientific community that characterises our time as an era of fast human–driven extinctions (Levinton, 2001). Many species are becoming rare and others have already disappeared, and this is a general process, affecting vertebrates (e.g. Galán Regalado, 2015) and invertebrates (e.g. Gibbs, 1998), and terrestrial (Ceballos and Ehrlich, 2002) and marine ecosystems (Roberts, 2003), generating a truly irreversible change (Dirzo and Raven, 2003). The word 'biodiversity' has been instrumental in promoting species conservation efforts (Takacs, 1996). However, much biodiversity is currently not evaluated in conservation programs, because it is related to levels of organic complexity which usually escape monitoring. Among these neglected levels of biodiversity, ethodiversity is prominent (Cordero–Rivera, 2017). Some behaviours vanish before being scientifically studied (Caro and Sherman, 2012), although species may persist. Furthermore, when species with unique behaviours are lost, these elements of biodiversity disappear forever. Therefore, documenting species– specific, population–specific and individual–specific behaviours is a priority to fully embrace biodiversity conservation (Cordero–Rivera, 2017). Freshwater systems are among the most threatened ecosystems in the world due to overexploitation (water extraction), pollution, and, in the case of forest streams, intensification of forestry, with substitution of forests by plantations of exotic trees (Cordero–Rivera et al., 2017) or their transformation to agricultural systems (Revenga, et al., 2005). Mitigation of these negative effects is of primary importance, particularly because freshwater systems are home to 6 % of world species, even though they represent only 0.8 % of the world surface (Dudgeon et al., 2006). Odonates are a key taxon in the ecological networks of small forested streams, where they may dominate the food webs (e.g. Yule, 1996). They are sensitive to changes in land use, and to the loss of forests (Cordero–Rivera, 2006). The behaviour and ecology of odonates is well studied, particularly in temperate regions, and for this reason this order has been the focus of much research (Corbet, 1999; Córdoba–Aguilar, 2008). Odonates have become text–book examples of sexual selection and sperm competition since the discovery of the dual function of male genital ligula as a device for rivals’ sperm removal and transfer (Waage, 1979). Nevertheless, the diversity of reproductive strategies is understudied, and we lack information on the basic ethology of many taxa, especially those confined to tropical forests (Córdoba–Aguilar and Cordero–Rivera, 2008). Among these understudied taxa, the Hainan endemic Pseudolestes mirabilis Kirby, a species so unique that its taxonomic position has been the centre of much controversy (Yu and Bu, 2011), is now proposed as a monotypic family (Dijkstra et al., 2014). It shows several morphological and ethological particularities (Reels, 2008). This species is restricted to small forested streams and has its habitat threatened by human activities (Zhu et al., 2015; Xue et al., 2017). The larva, with unique characters such as abdominal gills, is found in moderate numbers

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in these streams (Yu and Bu, 2011). This species has expanded since the last glacial period, but it shows signatures of slight decline in the last centuries, which might be exacerbated by current land use changes (Xue et al., 2017). Our goal was therefore to test the hypothesis that taxonomically isolated species show peculiar and unique behaviours, and are therefore of special interest at the level of ethodiversity (Cordero–Rivera, 2017). We concentrated on the study of territorial and reproductive behaviour, using marked animals in forested streams. Methods Field sites Field work was done at two localities in Hainan island (China) in May–June 2014 (for futher details see Garrison et al., 2015). First, we completed exploratory surveys to become familiar with the species and locate suitable streams for detailed behavioural observations. Between 27 and 30 May we sampled several streams at Diaoluoshan Forest Reserve, in Lingshui county (coordinates 18.727933 ºN, 109.880182 ºE, 900 m altitude). A second period of preliminary sampling was completed between 31 May and 1 June at Shuinan village, Wuzhishan (18.894678 ºN, 109.672557 ºE, 700 m). This second locality was found to have high density of P. mirabilis (in agreement with Xue et al., 2017) and therefore we selected a stream of about 1 m in width and 0.1–0.5 m in depth, which was visited daily between 13 and 23 June. Field methods We studied a sector of 275 m of the stream (of which 70 m were dry) running thru secondary forest and tea plantations, and visited the site to mark and observe the animals, for a mean of 6.23 ± 0.58 hours (± SE), and a total of 68.5 hours. Observations started at around 9:00 h and ended at 17:00 h, except when rainy conditions (common after 13:00 h) shortened field–work. From June 13 to June 15, we concentrated on behavioural observations. From June 16 to June 23, we marked all males found in the stream to estimate movements and survival. On 13 June we installed a datalogger to record air and water temperature every 5 minutes (Gemini Dataloggers, UK). The logger was situated in a small tree at the shore of the stream, about 1 m from the water surface, and a probe was placed in the stream. The logger was covered with leaves to avoid direct sun exposure. Water temperature showed very small changes, with a mean of 22.3 ± 0.005 ºC (N = 2,939 readings; range 21.68–23.54). Air temperature showed repetitive daily changes over the study period, starting at about 25 ºC at 9:00 h, with a maximum of 28–29 ºC at 15:00 h, and a mean of 23.8 ± 0.037 ºC (range 20.45–29.65). From 16 to 23 June, we captured and marked all the males found on the stream, writing a number on their wing using a permanent marker (Faber–Castell Multimark, 1523) (see fig. 3A). They were then released immediately in the same spot. Marked males returned

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to the same place after marking (albeit with around 10 min of delay) and behaved normally. Marked animals were not recaptured and their identification number was read without disturbing them. It was not possible to record data blind because our study involved focal animals in the field. We conducted a focal observation period of 10 min of marked males. If a male was not marked, we observed it and marked it at the end of the observations. For each male we recorded the time observation began, the number of flights, the number of agonistic interactions with other males (or male wasps, see Results), the number of grooming behaviours (cleaning of head, eyes, legs or abdomen) and the number of wing displays. Wing displays are defined as spontaneous movements of the wings, opening and closing them in a fast sequence (video 1 in supplementary material); this makes the males highly conspicuous, at least to human observers. If a focal male was observed in a fight, we measured fight duration and whether the male returned to his previous perch/territory. Some fights were also recorded with a video camera, including high–speed video (300 frames/ second, Casio EX–F1). Given the limited number of males in the stream (see Results), we observed the same marked males at different times of the day, on different days, trying to get one focal observation per male within each hour interval between 9 and 16 h, the period of maximum activity at the stream. Reproductive behaviour When a female was detected near the water, we followed her closely to describe her reproductive behaviour. No females were marked and they were collected at the end of the focal observation. Whenever possible, copulations were video–recorded for further analyses. To study sperm competition mechanisms we collected females after mating (N = 5; presumed to have full sperm load), and females at the end of stage I (N = 5; expecting to have a smaller sperm volume if sperm removal occurs). We also collected four females found alone at different times of the day, to estimate the sperm volume of females arriving at the stream. Sample size was limited by the number of females found and the numbers of copulations observed. The sperm volume of the storage organs was estimated by measuring the area of the sperm mass and multiplying by a uniform thickness (Cordero and Miller, 1992). Analysis of mark–recapture data To analyse the recapture histories of marked males, we used the full time–dependent Cormack–Jolly–Seber model (CSJ), as implemented in Mark 8.1 (White and Burnham, 1999). First, we estimated the degree of fit of the CJS model to the data, by using TEST 2 and TEST 3 of program Release, from within Mark. Although data were limited, there were no significant deviations, indicating a reasonable fit (Goodness of Fit Results TEST 2 + TEST 3; c211 = 5.08, p = 0.927). We then estimated the value of the extra–multinomial variance (c–hat) parameter using two strategies, and

selected the most conservative. We computed the estimated c–hat from the CJS full model and divided this value by the mean c–hat from the bootstrap procedure in Mark. This c–hat indicated underdispersion (0.6780). The second estimation was obtained by dividing the observed deviance of the CJS model by the mean deviance of the bootstrap procedure. The value obtained (1.2489) was used to correct estimates of parameters and confidence intervals. We estimated the number of males found in the transect following the methods of Jolly (1965) and Manly and Parr (1968), using the software Popan 5 (Arnason et al., 1998). Throughout the text, mean values are presented with their standard errors and sample size (in parenthesis). Statistical analyses were performed using xlStat 2017 (www.xlstat.com). Results Demography and general activity Over the study period, we marked 36 males and resighted all but four on subsequent days. On average, males were resighted on 4.6 ± 0.4 different days, and one male, marked on the first day of study, was observed every day until the end of the fieldwork, always at the same spot of the stream. This high site–fidelity resulted in high recapture rates. The best model to explain variability in recapture histories is the simplest model {Phi(.) p(.)}, with constant survival and recapture rates (table 1). The remaining models had low statistical support (delta QAIC > 6.8; table 1). The estimated survival rate was 0.8989 (SE = 0.0272; confidence interval: 0.8318–0.9412), and the recapture rate was 0.9252 (SE = 0.272; CI: 0.8514–0.9639). This survival rate translates into an expected longevity of 9.4 days, using the formula of Cook et al. (1967) [lifespan = −1/ loge(survival)]. The methods of Jolly (1965) and Manly and Parr (1968) yielded almost identical estimates of population size for the period between 16 and 22 June (no valid estimates were obtained before 16 June). The estimated number of males was 17–21 individuals (average 20.4 ± 1.0 (11) males), suggesting stable and low population density. This number is concordant with field observations because we found between 16 and 24 males each day. Males were observed arriving at the stream between 8:50 and 9:30 h and leaving about 16:30– 17:00 h, although some were still at the stream after 17:00 h. They spent most of their time perched. The most frequent behaviour was 'wing display', which peaked between 13 and 15 h, with 5–7 displays per 10 min. This behaviour was rarely observed after 15 h (1 display/10 min; fig. 1; video 1 in supplementary material). The second most common activity was spontaneous flight, also peaking at 13:00–15:00 h, but with less variation over the day (1.4–3.1 flights/10 min). Grooming and agonistic interactions (fighting) were infrequent, with an average of fewer than 0.5/10 min (fig. 1).

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Table 1. Results of model selection for P. mirabilis, using Mark software. The most supported model (in bold) is also the simplest, with only two parameters. Survival (Phi) and recapture rates (p), and model notation follow Lebreton et al. (1992), where t indicates variation of the parameter over time, and a dot indicates constant value. Tabla 1. Resultados de la selección de modelos para P. mirabilis, utilizando el programa informático MARK. El modelo con mayor respaldo estadístico (en negrita) también es el más sencillo, con solo dos parámetros. Índices de supervivencia (Phi) y recaptura (p) y la notación del modelo siguen lo establecido en Lebreton et al. (1992), donde t es la variación del parámetro en el tiempo y el punto indica valor constante. Model QAICc

Delta QAICc

QAICc weights

Model likelihood

{Phi(.) p(.)}

138.0020

0.0000

0.9618

1.0000

63.5148

{Phi(t) p(.)}

144.8118

6.8098

0.0319

0.0332

50.4384

Num. Par

QDeviance

{Phi(.) p(t)}

148.0921

10.0901

0.0062

0.0064

53.7187

{Phi(t) p(t)}

158.2188

20.2168

0.0000

41.0956

27.0

Wing display Flight Grooming Fighting Temperature

9 8 7

26.5 26.0 25.5

23.0

22.5

22.0 16:00–15:59

23.5

15:00–15:59

14:00–14:59

24.0

13:00–13:59

12:00–12:59

24.5

11:00–11:59

10:00–10:59

25.0

9:00–9:59

Time of day Fig. 1. Variation in activity of male P. mirabilis and air temperature over the day (average from 13–23 June 2014). Data represent the frequency of each behaviour per 10 min of observation (± SE). This plot is based on 162 focal observations of 10 min each, involving 28 marked males and eight unmarked males. No male was observed more than once on each day. Fig. 1. Variación en la actividad de los machos de P. mirabilis y en la temperatura del aire a lo largo del día (promedio del 13 al 23 de junio de 2014). Los datos representan la frecuencia de cada comportamiento en 10 minutos de observación (± EE). La figura está basada en 162 observaciones focales de 10 minutos cada una, de 28 machos marcados y ocho no marcados. Ningún macho fue observado más de una vez al día.

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Relative frequency

0.6 0.5 0.4 0.3 0.2 0.1 0

20 25 30 35 40 Fight duration (min)

Fig. 2. Histogram of duration of 47 agonistic encounters between males of P. mirabilis, in classes of one minute. Note that the distribution is bimodal, with very short or very long fights. Fig. 2. Histograma de la duración de 47 encuentros agonísticos entre machos de P. mirabilis, en clases de un minuto. Obsérvese que la distribución es bimodal, con peleas muy cortas o muy largas.

Territorial behaviour Males defended a territory of about 0.5–1 m in radius, which included one or more damp logs, used by females to lay eggs. The same area was defended on consecutive days, showing high site fidelity. A few non–territorial males were observed perching on the shore vegetation; they were not attacked by residents unless they flew into a territory. Males were highly aggressive against conspecific males. We timed 47 interactions involving 15 marked males and 16 unmarked males. These data indicate that fight duration is bimodal (fig. 2). Most fights (83 %) lasted less than 2 minutes and were face–to–face encounters where males exhibited their blue face while mimicking behaviour of the other (fig. 3B). Neighbour territorial males were frequently observed engaging in these interactions. A few cases (13 %) lasted up to 7 min, and two (4 %) lasted 50 and 60 min. These were escalated fights that started with very close face–to–face displays (at about one body length of distance) and with the abdomen of both males pointing upwards (fig. 3C, 3D and video 2 in supplementary material), intercalated with hovering periods (fig. 3E) of both males flying only with their forewings (video 3 in supplementary material). From high–speed video, forewing beat frequency can be estimated as 11.2–12.2 hz. Both males approached and retreated alternatively, maintaining a distance of 20–40 cm, while moving to treetops. The two escalated fights observed may have lasted more than one hour because it was difficult to follow the pair into the forest canopy. Escalated fights were sometimes interrupted by short periods (2–3 minutes) of perching in the shoreline vegetation.

Males also interacted aggressively with wasps (Eustenogaster nigra Saito and Nguyen, 2006), with short face–to–face confrontations similar to those observed with conspecific males (fig. 3F, 3G). These interactions were common (N = 29 cases) but never lasted more than a few seconds. Sometimes males were attacked by the wasps while they were perched. It was unclear whether the wasps acted as territorial or as predators, but male P. mirabilis were apparently acting as territorial when they confronted the wasps. Reproductive behaviour When a male detected a female perched near his territory, there was a brief courtship flight, of a few seconds, in which the male exhibited his coloured hindwings. Females were observed approaching the stream and started to lay eggs on dead logs if no males were around. Males readily tried to grasp ovipositing females in tandem, with no or very brief courtship. Immediately, the male grasped the female prothorax, a precopulatory tandem was formed (5–6 sec), and the pair copulated. Two females rejected copulation after having been grasped in tandem. There was no intra–male sperm translocation before copulation. Copulation (fig. 3H) occurred at low frequency, at any time between 10:50 and 15:14 h (fig. 4), with a maximum of three matings observed per day and a total of 14 matings over the study period. Copulation lasted on average 7.10 ± 0.98 (7) min, with a range from 2.6 to 10.0 min (fig. 4). A generalized linear model with copulation duration as the response variate and time of day and air temperature as predictors suggested that copulation duration was not

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A B

I J

Fig. 3. Agonistic and reproductive behaviour of P. mirabilis: A, a territorial male (marked on the right forewing); B, detail of the blue face of the male; C, D, agonistic behaviour of high intensity, where the males show coloured hind wings, and upward–oriented abdomen; E, a male in the static agonistic flight, with the abdomen in horizontal position, and situated about 20–40 cm from its opponent; males fly only with their forewings; F, G, agonistic encounter with the wasp Eutenogaster nigra, which patrolled the same area of the stream; H, copulation; I, intra–male sperm translocation, which takes place after copulation; J, one ovipositing female, inserting eggs in a damp log on the stream. (Photos: ACR.) Fig. 3. Comportamiento agonístico y reproductor de P. mirabilis: A, un macho territorial (marcado en el ala delantera derecha); B, detalle de la cara azul del macho; C, D, comportamiento agonístico de alta intensidad, donde los machos exhiben las alas posteriores coloreadas, y el abdomen orientado hacia arriba; E, un macho en vuelo agonístico estático, con el abdomen en posición horizontal, y situado a unos 20–40 cm de su oponente; los machos solo vuelan con sus alas anteriores; F, G, encuentro agonístico con la avispa Eutenogaster nigra, que patrullaba la misma zona del arroyo; H, cópula; I, la transferencia de esperma hacia la vesícula seminal del propio macho, que tiene lugar después de la cópula; J, una hembra en puesta, insertando huevos en un tronco húmedo en el arroyo. (Fotos: ACR.)

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12 Copulation Stage I

Duration (min)

Stage II

8 6 4 2 0 0:00 10:00 11:00 12:00 13:00 14:00 15:00 16:00 17:00 Time of day

Fig. 4. The relationship between time of day and copulation duration (and its stages) in P. mirabilis. Note that stage II shows very little variation in duration. Fig. 4. Relación entre la hora del día y la duración de la cópula (y sus etapas) en P. mirabilis. Obsérvese que la etapa II muestra muy poca variación en su duración.

affected by these variables (F3, 3 = 2.134, p = 0.275). Copulation consisted of two phases (video 4 in supplementary material), similar to those described for other Zygoptera (Miller and Miller, 1981). Stage I was the most variable phase (average: 6.21 ± 0.57 (10); range: 3.0–8.8 min) and occupied 87% of copulation (fig. 4). Stage II was short, and almost identical in duration among different pairs (average: 1.14 ± 0.05 (7); range: 0.9–1.3 min; fig. 4). Females had a large spherical bursa copulatrix and no spermatheca (fig. 5A–5C). Male genital ligula had a flexible head culminated by two processes, covered by backward oriented spines (fig. 5D–5E). Movements of the male abdomen during stage I of copulation were suggestive of sperm removal by male genitalia (video 4 in supplementary material). To test this idea, we preserved females found alone, mated females after copulation, and interrupted females at the end of stage I, presumably before insemination. Upon dissection, we found that females had 64 % less sperm stored in their bursa when copulation was interrupted (mean volume: 0.0233 ± 0.0076 (5) mm 3 ), compared to pre–copula females (0.0644 ± 0.0133 (4)), which in turn had a sperm volume very similar to females captured after copulation (0.0678 ± 0.0125 (5)). There were significant differences between groups (ANOVA, F2, 11 = 5.095, p = 0.027). One of the females whose copulation was interrupted had the bursa empty, which could indicate that she had never mated previously. If this female were excluded, the sperm volume in interrupted fe-

males would be 0.0292 ± 0.0063 (4). A comparison between this value and the mean sperm volume of precopula females indicated marginal differences (t = –2.388, p = 0.054). After copulation, most females expelled a drop of sperm (six out of eight females closely observed) by active movements of their genital valves (video 3 in supplementary material). One such drop was examined under microscope and its volume was similar to the volume of sperm stored in postcopula females. After copulation, the male remained close to the female, and then flew to a perch in the territory. We concentrated on female postcopulatory behaviour, until we discovered that the intra–male sperm translocation took place within 2 min of the end of copulation (fig. 3I). The last three matings observed, when we specifically followed the male to check for this behaviour, all completed intra–male sperm translocation (lasting about 20 sec) after copulation (video 3 in supplementary material). Oviposition was observed on dead wood inside the stream (fig. 3J; video 5 in supplementary material), with males remaining not far from their mate, although close guarding could not be confirmed. Discussion In agreement with our initial hypothesis, we found that P. mirabilis possess several unique behaviours when compared to other damselflies. Given their

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A 0.5 mm 0.5 mm

C 0.5 mm

E 10 μm

200 μm

Fig. 5. Female and male genitalia of P. mirabilis: Bursa copulatrix and genital plates of a female collected before copulation (A), compared to a female whose copulation was interrupted at the end of stage I, before insemination (B), and a female collected immediately after copulation (C). The sperm volume of interrupted females was clearly reduced. Male genital ligula at SEM (D), with a detail (E) of the spines in the genital process. This anatomical evidence points to sperm removal by male genital processes. Images edited to remove background dust. Fig. 5. Genitales masculinos y femeninos de P. mirabilis. Bursa copulatrix y placas genitales de una hembra capturada antes de la cópula (A), en comparación con una hembra cuya cópula fue interrumpida al final de la fase I, antes de la inseminación (B), y una hembra capturada inmediatamente después de la cópula (C). El volumen de esperma encontrado en hembras al interrumpir la cópula se redujo claramente. Lígula genital masculina observada al SEM (D), con un detalle (E) de las espinas en el proceso genital. Este rasgo anatómico apunta a la eliminación del esperma por los procesos genitales masculinos. Imágenes editadas para eliminar el polvo de fondo.

distinctiveness, limited distribution and threats to their specialised habitat (Zhu et al., 2015), it is a priority taxon for conservation (Xue et al., 2017). Furthermore it is phylogenetically isolated, to the extent that it has been considered a new (sub)family since its original description (Kirby, 1900), and there currently seems to be a consensus to consider it a monotypic family (Dijkstra et al., 2014; Reels and Zhang, 2015; Yu and Bu, 2011). Our results confirm that P. mirabilis is highly territorial (Reels, 2008). Male density during study period was about one male per meter in favourable zones, with a mean of 20 individuals in a section of 200 m. Our population had low density, but the species was

dominant in the studied stream. Our fieldwork was probably completed at the end of the flight season so the density might be higher earlier in the season. Females were rarely seen on the stream. Males are long–lived, showing very little activity for hours, but sometimes engaging in conspicuous and elaborate aggressive behaviour. The estimated survival rate of 0.8989 per day is similar to that of Calopterygidae (Cordero–Rivera and Stoks, 2008), a family that also shows high site tenacity, territoriality, and in some species, escalated agonistic encounters (Córdoba– Aguilar and Cordero–Rivera, 2005). However, in many respects, the morphology and behaviour of P. mirabilis recalls that of the genus Chalcopteryx, an endemism of

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the Amazon forest streams (Guillermo–Ferreira et al., 2014; Resende and De Marco, 2010), which belongs to a distant family (Polythoridae). Both Pseudolestes and Chalcopteryx have long hyaline forewings, short hindwings with conspicuous coloration. Both also show high territoriality and engage in aerial contests between neighbouring males, sometimes involving three or more males. There are therefore surprising ethological and morphological convergences between these two distantly related taxa. Similar ecological pressures might be behind their convergent behaviour and morphology, which is also found at the larval stages, with both genera having abdominal gills, albeit completely different in structure (dos Santos and Costa, 1987; Yu and Bu, 2011). The most conspicuous behaviour of P. mirabilis is the agonistic male–male encounters (Reels, 2008). For most of the time, activity of territorial males was reduced, and some even remained motionless for more than one hour. Wing displays broke this immobility, again recalling behaviour of males of Chalcopteryx, which remain motionless for long periods and exhibit their iridescent hindwings in dark spots of the forests (Resende and De Marco, 2010). Fights in P. mirabilis were highly ritualised, consisting of short chases of a few seconds or minutes, with males confronting each other and displaying their blue face and coloured hind–wings (fig. 3). It is noteworthy that some fights escalated and lasted for almost one hour (fig. 2). Other damselflies apparently use their face ornamentation during territorial displays as honest signals (Vilela et al., 2017), but it is not known whether this applies to P. mirabilis. During these displays, we observed that the hindwings remained motionless (video 2 and video 3 in supplementary material), and the males beat their forewings at only 11–12 Hz, which is considerably slower than other territorial damselflies (e.g. 29.5 Hz in Chlorocypha cancellata; Günther, 2015). This behaviour is likely to be very costly in terms of energy. In one of the long fights observed, the resident male lost the territory after this interaction, as sometimes occurs with the escalated fights in Calopteryx (Marden and Waage, 1990). Future studies of the phenotypic variables associated with flight endurance and the tendency to escalate agonistic encounters are needed to fully understand the territorial behaviour of P. mirabilis. An unexpected result from our fieldwork was the interspecific agonism between damselflies and wasps (fig. 3F; see also Garrison et al., 2015). Wilson and Reels (2001) described P. mirabilis as giving 'a strong bee–like impression'. It seems that its hymenopteran mimicry is enough to produce mistaken identity between the two species. Interspecific territoriality is not uncommon among similar species, particularly congeneric ones that might be unable to distinguish between conspecific and heterospecific females (Anderson and Grether, 2010). This is not the case for P. mirabilis and E. nigra. Both species might be paying the costs of sharing the habitat and having similar aggressive behaviour. It would be interesting to compare sympatric and allopatric populations of the two species, to detect possible character displacement.

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Reproductive behaviour of P. mirabilis also shows several unique characters. First, copulation was not preceded by the intra–male sperm translocation, which typically occurs in Zygoptera before each mating (Cordero–Rivera and Córdoba–Aguilar, 2010). This behaviour is needed because odonates do not have internal connection between the testis and the intromittent organ (Cordero–Rivera and Córdoba–Aguilar, 2010). Sperm translocation after mating has previously been recorded only in the small coenagrionid Mortonagrion hirosei (Naraoka, 2014), and in the genus Cora (Polythoridae; Fraser and Herman, 1993), and we observed the same behaviour in Chinese members of the family Euphaeidae (Cordero–Rivera, unpublished). Given that males were never observed translocating sperm to their seminal vesicle before mating, we assume that they were inseminating the sperm already stored in that organ, translocated after the last mating. This is counterintuitive, because mating frequency was found to be very low: most males would not find a female to mate with for over several days. Therefore, we hypothesize that the seminal vesicle of P. mirabilis has physiological mechanisms to maintain sperm alive for longer periods than most odonates. Alternatively, males could translocate sperm each morning, discarding previous unused sperm, although we did not observe this behaviour. In any case, this unusual behaviour needs further study. We are currently reviewing the diversity of intra–male sperm translocation in the entire order Odonata, as an example of ethological diversity (Cordero–Rivera, 2017). A second unusual fact about P. mirabilis reproductive behaviour is the almost complete absence of precopulatory courtship typically found in odonates with coloured wings (Outomuro et al., 2013; Svensson and Waller, 2013). Another taxonomically isolated damselfly, Hemiphlebia mirabilis, considered by some authors as the oldest living zygopteran (e.g. Fraser, 1955), was found to have a highly specialised courtship (Cordero–Rivera, 2016), suggesting that courtship behaviour might have evolved early in the damselflies. In agreement with this, recent fossil evidence also points to the development of specialised structures likely used for courtship in Cretacic damselflies (Zheng et al., 2017). Nevertheless, in P. mirabilis, the coloured wings have apparently evolved in the context of intra–sexual selection. Therefore, wing pigmentation might be an honest signal of male quality, and a study of inter–individual and inter–population variation in wing pigmentation would be of interest. Copulatory behaviour of P. mirabilis was typical, however, in the sense that it follows the copulatory stages already described for Coenagrionidae (Miller and Miller, 1981) and found to be widespread in Zygoptera (Córdoba–Aguilar and Cordero–Rivera, 2008). This points to a very old origin of sperm removal in the Odonata (Cordero–Rivera, 2016). The genital ligula of this species has two distal appendages 'each of which terminates in a white pad covered with microscopic recurved prickles' (Needham, 1931) (see fig. 5D–5E). This morphology is in agreement with the ability to remove sperm from the bursa copulatrix. We provide evidence for sperm removal in this spe-

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cies, and this information is of relevance to track the evolution of sperm displacement in the Odonata (Cordero–Rivera and Córdoba–Aguilar, 2010). It is noteworthy that females do not have spermathecae, which may indicate that males are controlling fertilization, because females usually retain control over the spermathecal sperm (Córdoba–Aguilar and Cordero–Rivera, 2008). Yet we found that after mating, most females expelled a drop of sperm in such a large quantity that it was clearly visible with the naked eye (video 4 in supplementary material). There are two possible explanations for this behaviour. Females might simply be expelling the sperm that the male removed during copulation (Lindeboom, 1998), or they might be cryptically choosing between ejaculates and expelling part of the sperm from the last male (Córdoba–Aguilar, 2006). These two possibilities are not mutually exclusive. A genetic analysis is needed to understand this behaviour. In conclusion, we found that a taxonomically isolated species of damselfly shows remarkable ethological features (see also Cordero–Rivera, 2016) and a striking resemblance to an ecologically similar group of damselflies from a different family (Polythoridae). The agonistic behaviour of P. mirabilis is unique and has no equivalents in odonates. This diversity is therefore not explained by the usual three levels of biodiversity assessments, which include genes, species and ecosystems but neglect behaviour (Cordero–Rivera, 2017). Acknowledgements We are very grateful to Ishizawa Naoya for estimating wing beat frequency from our videos, and to Vincent Kalkman for providing the SEM image of Figure 5e. We thank Professor Qinghua Cai, Mr Hongdao Wu and his wife Cuizhen Yan, and Rosser Garrison for help before or during fieldwork. This research was funded by grants from the Spanish Ministry with Competences in Science (CGL2011–22629 and CGL2014–53140–P), including FEDER funds. SEM images were obtained at the facilities of the CACTI (University of Vigo). References Anderson, C. N., Grether, G. F., 2010. Interspecific aggression and character displacement of competitor recognition in Hetaerina damselflies. Proceedings of the Royal Society of London Series B, Biological Sciences, 277: 549–555, http://doi.org/10.1098/ rspb.2009.1371 Arnason, N. A., Schwarz, C. J., Boyer, G., 1998. POPAN–5. A data maintenance and analysis system for mark–recapture data. Manitoba, Canada: Department of Computer Science, The University of Manitoba. Caro, T., Sherman, P. W., 2012. Vanishing behaviors. Conservation Letters, 5(3): 159–166, http://doi. org/10.1111/j.1755–263X.2012.00224.x

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larva of Pseudolestes mirabilis Kirby (Odonata: Pseudolestidae). International Journal of Odonatology, 14(2): 105–110, http://doi.org/10.1080/138 87890.2011.592486 Yule, C. M., 1996. Trophic relationships and food webs of the benthic invertebrate fauna of two aseasonal tropical streams on Bougainville Island, Papua New Guinea. Journal of Tropical Ecology, 12: 517–534. Zheng, D., Nel, A., Jarzembowski, E. A., Chang, S.–C.,

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Supplementary material

Video 1. Feeding and wing displays of male P. mirabilis. Video 1. Alimentación y exhibición de las alas del macho de P. mirabilis. https://youtu.be/UFAUZdSsuWE

Video 2. Agonistic interactions between territorial males of P. mirabilis. Note the different position of the abdomens at the start of interactions, and the increasing distance between males as the interaction continues. Males use only the forewings to confront each other, and maintain the abdomen in a horizontal position. Vídeo 2. Interacciones agonísticas entre machos territoriales de P. mirabilis. Obsérvese la diferente posición del abdomen al inicio de las interacciones y la creciente distancia entre los machos a medida que la interacción continúa. Los machos solo utilizan las alas posteriores en los enfrentamientos y mantienen el abdomen en posición horizontal. https://youtu.be/oBgFNACxPng

Video 3. High speed video (300 frames/second) of a male P. mirabilis in static confrontation with another male. Note that hindwings are not used in these agonistic displays. Vídeo 3. Vídeo a alta velocidad (300 fotogramas por segundo) de un macho de P. mirabilis en confrontación estática con otro macho. Obsérvese que las alas posteriores no se utilizan en estas exhibiciones agonísticas. https://youtu.be/UY9Z4AVVwx8

Video 4. Copulatory behaviour of P. mirabilis. Mating lasts about seven minutes, and has a first phase (stage I), which is used by males to remove sperm from the spermatheca, the only organ for sperm storing of females of this species. Stage II lasts about one minute and constitutes the phase of insemination. After copulation, most females expel a drop of sperm. Males do not translocate sperm before mating, but they do perform this behaviour a few minutes after copulation. Vídeo 4. Comportamiento de cópula de P. mirabilis. La cópula dura unos siete minutos y tiene una primera fase (etapa I), que los machos utilizan para extraer el esperma de la espermateca, que es el único órgano de almacenamiento de esperma que tienen las hembras de esta especie. La etapa II dura aproximadamente un minuto y constituye la fase de inseminación. Tras la cópula, la mayor parte de las hembras expulsa una gota de esperma. Los machos no transfieren el esperma antes de la cópula, pero sí lo hacen pocos minutos después. https://youtu.be/HEh203Pt57k

Video 5. Oviposition behaviour by female P. mirabilis on damp logs in forested streams. Females were sometimes guarded by males. Vídeo 5. Comportamiento de ovoposición de la hembra de P. mirabilis en troncos húmedos de arroyos de zonas boscosas. En ocasiones, los machos vigilaban a las hembras. https://youtu.be/PphV–HGz–vg

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Reconsidering mammal extinctions in the Pernambuco Endemism Center of the Brazilian Atlantic Forest G. S. T. Garbino, G. C. Rezende, H. Fernandes–Ferreira, A. Feijó

Garbino, G. S. T., Rezende, G. C., Fernandes–Ferreira, H., Feijó, A., 2018. Reconsidering mammal extinctions in the Pernambuco Endemism Center of the Brazilian Atlantic Forest. Animal Biodiversity and Conservation, 41.1: 175–184. Abstract Reconsidering mammal extinctions in the Pernambuco Endemism Center of the Brazilian Atlantic Forest. In the last 500 years, there have been an estimated 21 mammal extinctions in the Pernambuco Endemism Center. We critically reviewed the published historical and recent literature records and concluded that the actual number of mammal species extinction was seven, indicating that the previous figure of 21 species lost is an overestimation of approximately 30 %. Our checklist differs from previous publications by including species that are still extant (n = 5), and removing species that have never been recorded in the Pernambuco Endemism Center (n = 8). We point out that a more rigorous approach towards historical and recent records is needed when producing lists of regionally extinct fauna, given that the implications of misidentifications and false assumptions can potentially lead to loss of credibility by stakeholders and ultimately have a negative effect on species conservation. Key words: Conservation, Defaunation, Rewilding, Extirpation Resumen Reconsiderando la extinción de mamíferos en el Centro de Endemismo Pernambuco perteneciente al bosque atlántico brasileño. Se ha calculado que, en los últimos 500 años, se han extinguido 21 mamíferos en el Centro de Endemismo Pernambuco. En el presente estudio, realizamos un examen crítico de los datos aportados en las publicaciones científicas históricas y recientes, y concluimos que el número real de mamíferos extintos es de siete, lo que indica que la cifra anterior de 21 especies extintas es una sobrestimación de aproximadamente el 30 %. Nuestra lista difiere de las publicaciones previas en que incluye especies aún existentes (n = 5) y excluye otras que nunca habían sido registradas en el Centro de Endemismo Pernambuco (n = 8). Asimismo, señalamos que, al elaborar listas de fauna extinta a escala regional, es necesario adoptar un planteamiento más riguroso en relación con los registros históricos y recientes, dado que las identificaciones erróneas y las suposiciones falsas podrían conducir a la pérdida de credibilidad ante las partes interesadas y, en última instancia, ser negativas para la conservación de especies. Palabras claves: Conservación, Defaunación, Asilvestramento, Erradicación Received: 10 V 17; Conditional acceptance: 7 VIII 17; Final acceptance: 11 VIII 17 G. S. T. Garbino, Pós–graduação em Zoologia, Depto. de Zoologia, Inst. de Ciências Biológicas, Univ. Federal de Minas Gerais, Pampulha, 31270–901 Belo Horizonte, Brazil.– G. C. Rezende, IPÊ–Inst. de Pesquisas Ecológicas, C. Postal 47, 12960–000 Nazaré Paulista, Brazil.– H. Fernandes–Ferreira, Fac. de Educação, Ciências e Letras do Sertão Central, Univ. Estadual do Ceará, 639000–000 Quixadá, Brazil.– A. Feijó, Lab. de mamíferos, Depto. de Sistemática e Ecologia, CCEN, Univ. Federal da Paraíba, Campus I, 58051–900 João Pessoa, Brazil. Corresponding author: G. S. T. Garbino. E–mail: gstgarbino@hotmail.com

ISSN: 1578–665 X eISSN: 2014–928 X

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Motivation and data gathering Species are the basic unit for conservation actions, being a logical target and the linchpin for conservation assessments and management (Mace, 2004; Dunning et al., 2006). Evolutionary, biogeographic and ecological processes and patterns are inferred through knowledge of species diversity, summarized in the form of taxonomic lists (Diniz–Filho et al., 2013). Inaccurate lists of species (including, for example, false presences, unreliable records, misuse of taxonomic names, and statements of local extinction) affect all subsequent studies (from ecological to evolutionary), possibly misleading hypotheses to explain diversity patterns (Hortal et al., 2015) and precluding proper conservation actions. This may in turn lead to misuse of the limited conservation resources (Mace, 2004; Hortal et al., 2015). The negative effects of misconstrued lists are more pronounced in highly–degraded ecosystems with a strong ongoing defaunation process (Bini et al., 2006), where effectiveness and precision are key to preserving the remaining native species. The Atlantic Forest is a South American rainforest that originally encompassed around 150 million hectares from approximately 5 °S to 30 °S, spanning parts of Argentina, Brazil and Paraguay (Ribeiro et al., 2009; fig. 1). Today, only 11.7 % of the original forest cover of the Brazilian Atlantic Forest remains (Ribeiro et al., 2009). One of the most devastated areas of the Atlantic Forest is located at its northernmost portion, in the region known as the Pernambuco Endemism Center (Ribeiro et al., 2009; Bernard et al., 2011; fig. 1), hereafter PEC. Mainly due to sugarcane plantation over the last 500 years (Coimbra–Filho and Câmara, 1996), approximately 12 % of the original vegetation currently remains in PEC (Ribeiro et al., 2009). As a consequence of habitat reduction, many remaining fragments face defaunation, exemplified by local extinction of medium and large mammals (Silva Júnior and Mendes Pontes, 2008; Canale et al., 2012). Prior to 2016, no studies attempted to quantify how many species of medium and large–sized mammals were lost due to the historical fragmentation of the Atlantic Forest in PEC. In a recent article, Mendes Pontes et al. (2016) surveyed medium and large mammals from 21 Atlantic Forest fragments in PEC. The authors examined the relationship between species richness and fragment size, and compared the number of extant species with the number of historically present mammals, arguing that, from a total of 42 medium and large mammals present in historical times, 21 (50 %) had been extirpated. While the paper provides a panorama of the impoverishment of mammal fauna in the Atlantic Forest of northeastern Brazil, and undoubtedly will be an important tool for use in conservation actions in the region, we argue throughout this paper that this figure may be overestimated by as much as 30 %, significantly changing the 'mass extinction' scenario proposed by Mendes Pontes et al. (2016). However, it is important to point out that there is a time lag in biological responses to fragmentation and more species might be lost before long. Thus, urgent

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conservation actions, preferably based on scientific data, are demanded in the region in order to avoid future extinctions. With all this in mind, we reviewed the cases presented by Mendes Pontes et al. (2016) as well as the historical and current records of medium and large mammals in the PEC, and provided a checklist for the area based on reliable sources where the identification is unambiguous and/or preserved specimens from museums are mentioned. Our first divergence concerns the number of extinct species in the table presented in Mendes Pontes et al. (2016): while in the text the authors mention 21 extinct species, the table in the article (table 1 in Mendes Pontes et al. 2016) compiles 20. Besides this point, our checklist is significantly distinct from that in Mendes Pontes et al. (2016) for two other reasons: (1) we included species considered extirpated by the authors; and (2) we removed species that have never been confirmed in PEC. Below we explain the divergences in detail and discuss the potential consequences of misidentifications on conservation. False absences: species considered extinct from PEC that are still extant We found that five of the 21 species considered locally extinct by Mendes Pontes et al. (2016) still occur in the PEC (fig. 2). Feijó and Langguth (2013) mention the margay (Leopardus wiedii) in two localities in PEC: Roteiro (Alagoas), and Alhandra (Paraiba). Several recent records of the pygmy anteater (Cyclopes didactylus) are known for the region (Gardner, 2008; Miranda and Superina, 2010; Feijó and Langguth, 2013). The red–handed howler monkey (Alouatta belzebul), which is considered Vulnerable by the IUCN Red List, is still widespread at PEC (Fialho et al., 2014). The naked– tailed armadillo, whose species occurring in PEC is Cabassous tatouay, not C. unicinctus as reported by Mendes Pontes et al. (2016: see Feijó and Langguth, 2013), was recorded, based on voucher specimens, in two localities in PEC (Feijó and Langguth, 2013). Mendes Pontes et al. (2016) classified the Neotropical otter (Lontra longicaudis) as 'extinct before species confirmed' in the article’s first table, but elsewhere in the text the otter is classified among the 'living dead' species (p. 18). Nevertheless, L. longicaudis was recently recorded in nine localities in PEC by Astúa et al. (2010), Feijó and Langguth (2013) and Toledo et al. (2014). The tapir (Tapirus terrestris), the collared peccary (Pecari tajacu) and the white–lipped peccary (Tayassu pecari) are present in PEC, at Usina Serra Grande, Alagoas (Bachand et al., 2009; Lazure et al., 2010), but the population of these species there may have been reintroduced. On this basis, we agree with Mendes Pontes et al. (2016) in that these three species have been historically extirpated in the region. False presences: species that have not been confirmed for Pernambuco Endemism Center Eight of the 21 species considered extinct by Mendes Pontes et al. (2016) have never been recorded for the PEC (fig. 2). Two of the eight are open–area dwellers;

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38 ºW

36 ºW

34 ºW N

Brazil

6 ºS

Pernambuco Endemism Center 8 ºS

River Original Atlantic forest cover Atlantic forest remmant

Sã

Fr a

isc

10 ºS

0 50 100 150 200 km

Fig. 1. Location of Pernambuco Endemism Center (PEC) in northeastern Brazil. Fig. 1. Localización del Centro de Endemismo Pernambuco (PEC) en el nordeste de Brasil.

two are rarely sampled throughout their entire range, one is a supposedly undescribed species of deer that went extinct before description, and three are primates that are typical of the Amazon. The two open–area dwellers, the three–banded armadillo (Tolypeutes tricinctus) and the hog–nosed skunk (Conepatus amazonicus), do not occur naturally in the PEC. The distribution of genus Tolypeutes was recently revised based on interviews, direct observations, fossil, historical and recent records up to 2013 (Feijó et al., 2015), and all 168 records of Tolypeutes tricinctus were restricted to the Caatinga scrubland (Brazilian ecosystem adjacent to PEC) and Cerrado savanna of northeastern Brazil. Conepatus amazonicus is also typical of open areas (Kasper et al., 2009; Feijó and Langguth, 2013). From the 17 records of the species from northeastern Brazil, none came from the Atlantic Forest (Feijó and Langguth, 2013). Therefore, when all available evidence is considered, it is improbable that either T. tricinctus or C. amazonicus occurred in the Atlantic Forest, as suggested by Mendes Pontes et al. (2016). The other two species with no records for PEC are the bush dog (Speothos venaticus) and the lesser long–nosed armadillo (Dasypus septemcinctus).

Despite the authors' assertion that the bush dog was mentioned in historical documents of the 16th and 17th centuries, there are no accurate references to this citation. In fact, there is no known historical or extant record of S. venaticus for PEC (see Feijó and Langguth, 2013; Fernandes–Ferreira, 2014). Occurrence of the Bush dog in PEC has only been suggested by a study that inferred the habitat suitability for the species through ecological niche modeling (DeMatteo and Loiselle, 2008). Occurrence records of the lesser long–nosed armadillo (Dasypus septemcinctus) are very scarce for northeastern Brazil, and it has only been recorded for the Caatinga of Pernambuco, among the states that comprise the PEC (Feijó and Langguth, 2013). Mendes Pontes et al. (2016) reported two supposedly undescribed species of Mazama for PEC. However, there is neither historical nor current evidence that there existed another species of Mazama in PEC besides Mazama guazoubira (Feijó and Langguth, 2013), although it is currently extinct there. Three monkeys considered locally extinct by Mendes Pontes et al. (2016) based on 'historical records', are restricted to the Amazonian basin, in north–western Brazil, and could not have occurred in PEC. The referred

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Table 1. Checklist of medium and large–sized mammals present in the Pernambuco Endemism Center (PEC) in the past 500 years. The recently introduced Saimiri sciureus in Pernambuco was not included. Conepatus semistriatus (here named Conepatus amazonicus), Dasypus septemcinctus, Tolypeutes tricinctus and Speothos venaticus, cited by Mendes Pontes et al. (2016) were removed due to the lack of evidence: 1 previously treated as Leopardus tigrinus (see Nascimento and Feijó 2017). Sources: 1, Feijó and Langguth (2013); 2, Guerra (1981); 3, Miranda and Superina (2010); 4, Marcgrave (1648); 5, Feijó et al. (2016); 6, Percequillo et al. (2007); 7, Fialho et al. (2014); 8, Oliveira and Langguth (2006); 9, Langguth et al. (1987); 10, Silva Júnior and Mendes Pontes (2008); 11, Astúa et al. (2010); 12, Toledo et al. (2014); 13, Vieira (1952); 14, Vieira (1953); 15, Mayer and Wetzel (1987); 16, Mendes Pontes et al. (2013). Tabla 1. Lista de los mamíferos de talla mediana y grande presentes en el Centro de Endemismo Pernambuco (PEC) en los últimos 500 años. No se incluyó la especie recientemente introducida en Pernambuco, Saimiri sciureus. Se eliminaron las especies Conepatus semistriatus (aquí conocida como Conepatus amazonicus), Dasypus septemcinctus, Tolypeutes tricinctus y Speothos venaticus, citados por Mendes Pontes et al. (2016), debido a la falta de datos: 1 anteriormente tratada como Leopardus tigrinus (véase Nascimento y Feijó, 2017). (Para las abreviaturas de las fuentes, véase arriba). Taxon

Present in Extirpated from PEC PEC

Source

Order Cingulata, Family Dasypodidae Euphractus sexcinctus (Linnaeus, 1758)

Cabassous tatouay (Desmarest, 1804)

1, 2

Dasypus novemcinctus Linnaeus, 1758

Order Pilosa, Family Cyclopedidae Cyclopes didactylus (Linnaeus, 1758)

1, 3

Order Pilosa, Family Myrmecophagidae Myrmecophaga tridactyla Linnaeus, 1758 Tamandua tetradactyla (Linnaeus, 1758)

4 1, 5

Order Pilosa, Family Bradypodidae Bradypus variegatus Schinz, 1825

1, 6

Order Primates, Family Cebidae Sapajus flavius (Schreber, 1774)

1, 7, 8

Callithrix jacchus (Linnaeus, 1758)

Order Primates, Family Atelidae Alouatta belzebul (Linnaeus, 1766)

1, 7, 9

Order Lagomorpha, Family Leporidae Sylvilagus brasiliensis (Linnaeus, 1758)

1, 6

Order Carnivora, Family Felidae Leopardus emiliae (Thomas, 1914) 1

1, 6

Leopardus pardalis (Linnaeus, 1758)

Leopardus wiedii (Schinz, 1821)

Puma yagouaroundi (É. Geoffroy, 1803)

Puma concolor (Linnaeus, 1771)

Panthera onca (Linnaeus, 1758)

Order Carnivora, Family Canidae Cerdocyon thous (Linnaeus, 1766)

1, 6

Order Carnivora, Family Mustelidae Eira barbara (Linnaeus, 1758)

Galictis cuja (Molina, 1782)

Lontra longicaudis (Olfers, 1818)

1, 11, 12

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Table 1. (Cont.) Taxon

Present in Extirpated from PEC PEC

Source

Order Carnivora, Family Procyonidae Nasua nasua (Linnaeus, 1758)

Procyon cancrivorus (G. Cuvier, 1798)

1, 5

Potos flavus (Schreber, 1774)

1,13

Order Perissodactyla, Family Tapiridae Tapirus terrestris (Linnaeus, 1758)

Order Artiodactyla, Family Tayassuidae Pecari tajacu (Linnaeus, 1758)

4, 14

Tayassu pecari (Link, 1795)

Order Artiodactyla, Family Cervidae Mazama gouazoubira (Fischer, 1814)

Order Rodentia, Family Caviidae Hydrochoerus hydrochaeris (Linnaeus, 1766)

Order Rodentia, Family Cuniculidae Cuniculus paca (Linnaeus, 1766)

Order Rodentia, Family Dasyproctidae Dasyprocta iacki Feijó and Langguth, 2013

Order Rodentia, Family Erethizontidae Coendou prehensilis (Linnaeus, 1758)

1, 5, 16

Coendou speratus Mendes Pontes et al., 2013

1, 16

Order Rodentia, Family Sciuridae Guerlinguetus brasiliensis (Gmelin, 1788) Total

authors based the presence of a spider monkey (Ateles sp.) in northeastern Brazil in a Portuguese translation of the work of Caspar van Baarle, (latinized as Caspar Barlaeus in publications), about the Dutch possessions in Brazil, made by Claudio Brandão (Barlaeus, 1940). Brandão stated in a translation note that the name Cajatayae, used by Barlaeus to describe a long–tailed reddish monkey, was similar to the word Coatá, the name commonly used for the spider monkeys, genus Ateles. Marcgrave (1648), however, described a monkey called Caitaia, a name that resembles Barleus’s Cajatayae. The animal described by Marcgrave as Caitaia has been considered as Sapajus flavius, a capuchin monkey still extant in PEC, especially due to the reference to the yellowish color of its pelage (Oliveira and Langguth, 2006). Moreover, cay or cai, as in caitaia, is the name of capuchin monkeys in the indigenous language Tupi–Guarani. Recent re–discovery of animal drawings made by the artist Frans Post in the Dutch Brazil area revealed an Ateles–like monkey among the depicted fauna (De Bruin, 2016). It is very improbable that naturalists such

X 27

5, 6

as Georg Marcgrave or Wilhelm Piso would have failed to detect a population of large–bodied spider monkeys in northeastern Brazil (Marcgrave, 1648). A more probable explanation is that the animal illustrated was obtained elsewhere, as transportation of primates was common in the colonial Americas (Browne, 1789; Teixeira and Papavero, 2010). According to Mendes Pontes et al. (2016), Pero de Magalhães Gandavo (Gandavo, 1924) and the Franciscan friar Vicente do Salvador (do Salvador, 1889) mentioned the presence of squirrel monkeys (Saimiri sp.) in the PEC. In both works, there is no reference to animals morphologically similar to squirrel monkeys, and, more importantly, according to Capistrano de Abreu (in Gandavo, 1924), Gandavo never visited Pernambuco. We believe that Mendes Pontes et al. (2016) assigned some monkeys described in Gandavo (1924) and do Salvador (1889) to Saimiri due to the characteristic odor, mentioned by the two Portuguese authors. Probably, Mendes Pontes et al. (2016) associated the presence of odoriferous glands with the common name of Saimiri in Portuguese, 'mico–

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180

False presences

Confirmed extirpations

Ateles sp. C. amazonicus D. apella D. septemcinctus M. guazoubira Mazama sp. Saimiri sp. M. tridactyla S. venaticus P. concolor T. tricinctus P. onca P. tajacu T. pecari T. terrestris

A. belzebul

Taxa considered extinct

C. tatouay C. didactylus L. longicaudis L. wiedii

False absences

Fig. 2. Diagram showing the seven extinctions in Pernambuco Endemism Center (PEC) according to the present study (yellow circle) compared with the 20 extinctions reported by Mendes Pontes et al. (2016) (green circle). False presences (blue circle) refer to taxa considered present in PEC by Mendes Pontes et al. (2016) but that were not recorded in the area. False absences (red circle) refer to species considered extinct in PEC by Mendes Pontes et al. (2016) that are still extant in the area. (For full species names see table 1). Fig. 2. Diagrama en el que se indican las siete extinciones en el Centro de Endemismo Pernambuco (PEC) según el presente estudio (círculo amarillo) en comparación con las 20 extinciones registradas por Mendes Pontes et al. (2016) (círculo verde). Las presencias falsas (círculo azul) hacen referencia a los taxones considerados presentes en el Centro por Mendes Pontes et al. (2016), pero que no se observaron en la zona. Las falsas ausencias (círculo rojo) hacen referencia a las especies consideradas extintas en el Centro por Mendes Pontes et al. (2016) que siguen presentes en la zona. (Véase la tabla 1 para consultar los nombres completos de las especies).

de–cheiro', which means 'monkey with odor'. However, the presence of scent glands is widely distributed in New World monkeys (Perkins, 1975; Heymann, 2006). Marcgrave (1648, p. 227), for example mentions a 'musky odor' for Sapajus flavius. The genus Saimiri is endemic from the Amazon basin and Central America (Groves, 2001), and therefore squirrel monkeys were not historically present in the PEC. The current records of the genus in the region came from introduced animals, as pointed out by Mendes Pontes et al. (2016). The equivocal inclusion of 'Cebus apella' (a species that is now classified in the genus Sapajus) in PEC fauna by Mendes Pontes et al. apparently has a simpler solution. Using a now–outdated taxonomy of capuchin monkeys, Hershkovitz (1987, p. 23) mentions 'Cebus apella libidinosus' among the mammals described by Marcgrave in PEC. After the comprehensive taxonomic

review of Silva Júnior (2001), Cebus apella (nowadays Sapajus apella), then considered a widespread polytypic species, was restricted to the Amazonia of northern Brazil. As noted earlier, the capuchin monkey described by Marcgrave that occurs in PEC is Sapajus flavius (Oliveira and Langguth, 2006). Consequences for conservation and species management Our ability to understand species extinctions and plan conservation actions are dependent on correctly identifying the components of the ecosystem (Mace, 2004). Taxonomic errors (such as identifying the Dasypus armadillos depicted in Nassau’s handbook as Cabassous; Mendes Pontes et al., 2016, p. 11), false assumptions (as reporting Saimiri and Ateles species as native to

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the Atlantic Forest), or unreliable records (as considering Tolypeutes tricinctus, Conepatus amazonicus and Dasypus septemcinctus present in PEC), could potentially lead to negative impacts on conservation planning and actions, especially in an already fragile and endangered area such as the Pernambuco Endemism Center. When the ambiguous historical records are considered, a total of 42 'original' species (excluding the exotic Saimiri sciureus) of medium and large–sized mammals were present in Pernambuco Endemism Center in the yer 1500, 21 (50 %) of which remain (Mendes Pontes et al., 2016, p. 7). However, after excluding the animals that occur in Amazonia and Caatinga (but not in the Atlantic Forest), and the misidentified historical records, we are left with 34 species that truly occurred in PEC originally, of which there are reliable recent records for 27 (79.4 %) (table 1). The loss of seven (20.6%) medium and large mammal species is, of course, highly relevant, especially if we consider the extremely small populations and rarity of the remaining mammals in the Brazilian northeast, which make them very vulnerable to extinction, and the ecological role of medium and large mammals (Canale et al., 2012). Although the situation is critical, and many species such as Alouatta belzebul, Leopardus pardalis and Sapajus flavius face serious threats and have an extremely reduced occupancy there, the previously proposed 'mass extinction' scenario is not backed by empirical evidence or systematic data. In a more optimistic point of view, this extinction scenario can still be reverted with immediate conservation actions. Stating that a species is locally extinct in an ecosystem may lead to erroneous decisions when planning conservation actions. For example, if conservationists base their decisions on the fact that Tolypeutes and Ateles are no longer present in the area and should be 'reintroduced', this action could cause an unprecedented impact on the local fauna and flora. This has happened in the case of the Amazonian squirrel monkeys (Saimiri sciureus) that were intentionally released at Reserva Biológica Saltinho, in Pernambuco in 1987. The introduction of this non–native species has negatively affected the other syntopic primate in the area, Callithrix jacchus (Camarotti et al., 2015). Moreover, the presence of Saimiri sciureus in the forest reserve has hindered the efforts to reintroduce the extremely endangered Sapajus flavius in Saltinho (Camarotti et al., 2015). For these reasons, the eradication of Saimiri sciureus is a priority in the National Action Plan for primates in northeastern Brazil (Brasil, 2012). Conversely, if the proposed notion of Mendes Pontes et al. (2016) is accepted, and Saimiri is considered as a native genus of the Atlantic Forest, based on misinterpreted historical records, this invasive species would be tolerated in the area. The raccoons (Procyon spp.) from the Caribbean islands of New Providence, Barbados and Guadaloupe are an emblematic case of a poorly understood taxonomy affecting conservation. They were traditionally treated as three endemic and threatened species of these islands, until it was demonstrated through genetic and morphological evidence that they are the result of recent human introduction (Helgen and Wilson, 2003).

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Understanding the primary composition of a community, and the role of each species, is crucial for planning conservation actions that can reverse the extinction process and maintain the balance of the ecosystem. In this context, this historical information is useful for reintroducing key species in 'empty landscapes', a process known as rewilding (Navarro and Pereira, 2012). Although in some cases a key species may be extinct and ecological interactions must be restored using phylogenetically close taxa (e.g. the giant Cylindraspis tortoises extinct from the Mascarene Islands; Griffiths et al., 2010, 2011), the use of native species is generally preferred over the use of non–natives (Lerdau and Wickham, 2011). Besides rewilding with native fauna, the environmental benefits of reproducing the 'original' biological community of an area when conducting conservation management has been highlighted in studies dealing with forestry management around the world. Fernandes et al. (2016) pointed out that planting trees in some Cerrado areas in Central Brazil that have historically never been forested (a process called 'afforestation'), may be harmful for the environment and also for the human settlements that depend on open habitats. In another case, Szabó et al. (2017) suggested that during most of the Holocene, the flora of Central Europe was dominated by conifers rather than broadleaves tree species, the latter being the species promoted by nature conservation and forestry policies. The declaration that a species is locally extinct must be based on solid evidence, such as failure to record it after several decades of sampling (Collar, 1998) or paleontological data (e.g. Carleton and Olson, 1999). In this context, analysis of museum specimens and correct interpretation of historical documents (such as naturalist travel accounts and field notes) have a crucial role in species conservation, by providing unknown records of endangered species, confirming known ones, elucidating taxonomic problems and even unveiling species unknown to science (Schlick– Steiner et al., 2003; Mace, 2004; Helgen et al., 2013). Ultimately, museum collections help in constructing a more realistic scenario of the past and current diversity, and should be consulted when conducting biodiversity conservation management. If this conservative approach towards species extinction is not followed, constant change in conservation status may result in loss of credibility from stakeholders, including key decision makers such as governmental agencies and non–governmental organizations that act for the species conservation. A premature declaration of extinction can affect the way conservation science is treated by the general public, impairing their support (Monte–Luna et al., 2007). Final considerations A sound taxonomic basis and a comprehensive historical review can lead to a better understanding of species extinction as they provide records of historical changes in the biological communities. This approach is decisive for species conservation planning since it

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provides a framework for management actions such as reintroductions for the restoration of species assemblage in the community and ecosystem functions. Therefore, as Rylands (2007) states, taxonomy is not a trivial pursuit, and a hasty taxonomic practice can have disastrous consequences for conservation. Acknowledgements We are grateful to Jean Paul Metzger and two anonymous referees whose commentaries and suggestions greatly improved a previous version of this manuscript, to Deleece McLaren for reviewing the text, and Jose Eduardo Serrano–Villavicencio for the Spanish translation of the abstract. This work was supported by CAPES (Coordenação de Aperfeiçoamento de Pessoal de Nível Superior) (AF, GCR, GSTG) and CNPq (Conselho Nacional de Desenvolvimento Científico e Tecnológico) (AF). References Astúa, D., Asfora, P. H., Aléssio, F. M., Langguth, A., 2010. On the occurrence of the Neotropical Otter (Lontra longicaudis) (Mammalia, Mustelidae) in Northeastern Brazil. Mammalia, 74: 213–217, doi: 10.1111/mam.12098 Bachand, M., Trudel, O. C., Ansseau, C., Cortez, J. A., 2009. Dieta de Tapirus terrestris Linnaeus em um fragmento de Mata Atlântica do Nordeste do Brasil. Revista Brasileira de Biociências, 7(2): 188–194. Barlaeus, G., 1940. História dos feitos recentemente praticados no Brasil, durante oito anos, sob o governo do Ilustrissimo Conde João Maurício de Nassau. Serviço Gráfico do Ministério da Educação, Rio de Janeiro. (Original work published in 1647). Bernard, E., Melo, F. P. L., Pinto, S. R. R., 2011. Challenges and opportunities for biodiversity conservation in the Atlantic Forest in face of bioethanol expansion. Tropical Conservation Science, 4: 267–275, doi: 10.1177/194008291100400305 Bini, L. M., Diniz–Filho, J. A. F., Rangel, T. F. L. V. B., Bastos, R. P., Pinto, M. P., 2006. Challenging Wallacean and Linnean shortfalls: knowledge gradients and conservation planning in a biodiversity hotspot. Diversity and. Distribution, 12: 475–482, doi: 10.1111/j.1366–9516.2006.00286.x Brasil, 2012. Instituto Chico Mendes de Conservação da Biodiversidade. Aprova o Plano de Ação Nacional para a Conservação dos Primatas do Nordeste – PAN Primatas do Nordeste, contemplando cinco espécies ameaçadas de extinção, estabelecendo seu objetivo geral, objetivos específicos, ações, prazo de execução, abrangência e formas de implementação e supervisão. Portaria nº 37 de 23 Março 2012, http://www.icmbio.gov.br/portal/ images/stories/docs–plano–de–acao/pan–primatas–nordeste/portaria–panprimatasnordeste.pdf Browne, P., 1789. The civil and natural history of Jamaica. London, UK. Camarotti, F. L. M., Silva, V. L., Oliveira, M. A. B.,

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Multi–criteria evaluation and simulated annealing for delimiting high priority habitats of Alectoris chukar and Phasianus colchicus in Iran I. Momeni Dehaghi, A. Salmanmahiny, S. Karimi, A. A. Shabani Momeni Dehaghi, I., Salmanmahiny, A., Karimi, S., Shabani, A. A., 2018. Multi–criteria evaluation and simulated annealing for delimiting high priority habitats of Alectoris chukar and Phasianus colchicus in Iran. Animal Biodiversity and Conservation, 41.1: 185–193. Abstract Multi–criteria evaluation and simulated annealing for delimiting high priority habitats of Alectoris chukar and Phasianus colchicus in Iran. Habitat degradation and hunting are among the most important causes of population decline for Alectoris chukar and Phasianus colchicus, two of the most threatened game species in the Golestan Province of Iran. Limited data on distribution and location of high–quality habitats for the two species make conservation efforts more difficult in the province. We used multi–criteria evaluation (MCE) as a coarse–filter approach to refine the general distribution areas into habitat suitability maps for the species. We then used these maps as input to simulated annealing as a heuristic algorithm through Marxan in order to prioritize areas for conservation of the two species. To find the optimal solution, we tested various boundary length modifier (BLM) values in the simulated annealing process. Our results showed that the MCE approach was useful to refine general habitat maps. Assessment of the selected reserves confirmed the suitability of the selected areas (mainly neighboring the current reserves) making their management easier and more feasible. The total area of the selected reserves was about 476 km2. As current reserves of the Golestan Province represent only 23 % of the optimal area, further protected areas should be considered to efficiently conserve these two species. Key words: Common pheasant, Chukar partridge, Conservation planning, Habitat suitability, Multi–criteria evaluation, Marxan, Simulated annealing Resumen Evaluación de múltiples criterios y recocido simulado para delimitar los hábitats de alta prioridad de Alectoris chukar y Phasianus colchicus en Irán. La degradación del hábitat y la caza son algunas de las causas más importantes del descenso demográfico de Alectoris chukar y Phasianus colchicus, que son dos de las especies cinegéticas más amenazadas de la provincia de Golestán del Irán. La escasez de datos relativos a la distribución y localización de hábitats de alta calidad para las dos especies dificulta las iniciativas de conservación en la provincia. Utilizamos la evaluación de múltiples criterios para hacer una primera selección de las zonas de distribución general y elaborar mapas de idoneidad de los hábitats para las especies. A continuación, utilizamos estos mapas en forma de algoritmo heurístico en el recocido simulado por medio del programa informático Marxan, a fin de establecer un orden de prioridad entre las zonas para la conservación de ambas especies. Para hallar la solución óptima, probamos varios valores del modificador de longitud de frontera en el proceso de recocido simulado. Nuestros resultados pusieron de manifiesto que la evaluación de múltiples criterios resultó útil para refinar los mapas de hábitat general. La evaluación de las reservas seleccionadas confirmó la idoneidad de las zonas seleccionadas (que principalmente son contiguas a las reservas actuales) lo que facilita su gestión y la hace más viable. La superficie total de las reservas seleccionadas fue de unos 476 km2. Como las reservas actuales de la provincia de Golestán solo representan el 23 % de la superficie óptima, deberá estudiarse la posibilidad de añadir otras áreas protegidas a efectos de conservar de forma eficiente estas dos especies.

ISSN: 1578–665 X eISSN: 2014–928 X

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Palabras clave: Faisán común, Perdiz chucar, Planificación de la conservación, Sostenibilidad del hábitat, Evaluación de múltiples criterios, Marxan, Recocido simulado Received: 8 XII 14; Conditional acceptance: 10 II 15; Final acceptance: 9 IX 17 Iman Momeni Dehaghi, Sahebeh Karimi & Afshin Alizadeh Shabani, Dept of Environmental Science, Fac. of Natural Resources, Univ. of Tehran, Tehran, Iran.– Abdolrassoul Salmanmahiny, College of Fisheries and Environmental Sciences, Gorgan Univ. of Agricultural Sciences and Natural Resources, Golestan, Iran. Corresponding Author: Iman Momeni Dehaghi. E–mail address: momeni.iman@gmail.com

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Introduction Based on the International Union for Conservation of Nature and Natural Resources (IUCN) Red List of Threatened Species, around 13 % of all known bird species worldwide are under threat (IUCN, 2014). In addition, some species are exposed to threats and are undergoing rapid population declines locally, but their global status is ranked as Least Concern (LC) in the IUCN Red List (Mansoori, 2008). This is especially true for game species in developing countries. The common pheasant (Phasianus colchicus) and chukar partridge (Alectoris chukar) are considered galliform species of special concern according to their ecological and economic importance in the Golestan Province of Iran (Salmanmahiny, 2008). While the common pheasant prefers forests and shrublands in humid climates, the chukar partridge prefers mountainous, rocky habitats and avoids dense–forests (Mansoori, 2008). Inability to fly long distances makes these species vulnerable to hunting. Habitat degradation caused by land–use/land–cover change (Minaei and Kainz, 2016) and hunting (Panayides et al., 2011) are among the most important causes of population decline for these species and they require considerable conservation efforts in Iran (Mansoori, 2008). Knowledge about species' habitats and distribution is fundamental for conservation efforts (Lawler et al., 2011; Underwood et al., 2009). However, in general, there is a lack of high–quality data about species distributions (Store and Kangas, 2001; Store and Jokimäki, 2003), especially in developing countries such as Iran (Momeni et al., 2013). Habitat suitability modeling as a surrogate of distribution data can be used to fill this gap in conservation studies (Store and Kangas, 2001; Store and Jokimäki, 2003; Lawler et al., 2011). MAXENT (Phillips et al., 2006), ENFA (Hirzel et al., 2002) and GAMs (Hastie and Tibshirani, 1986) are among the most popular models used to provide habitat suitability maps. However, these methods need accurate occurrence data of species and very often such data are not available. Multi–criteria evaluation (MCE) is a good, first step towards achieving refined information about species suitability maps (Store and Kangas, 2001; Store and Jokimäki, 2003). With the MCE method, the environmental variables deemed relevant and important are combined according to their relative weight for the species under study (Momeni, 2011). Habitats with the highest suitability are then singled out, showing the most probable points of a species' occurrence. This information can be fed into the methods requiring accurate presence data. Establishing reserve networks is an important and effective tool for conserving high–quality habitats and biodiversity (Possingham et al., 2006; Ceballos, 2007; Lawler et al., 2011). Systematic approaches to reserve selection are preferred to ad hoc approaches because the former are data driven, goal directed, efficient, explicit, transparent, repeatable and flexible (Pressey, 1999). A wide array of systematic conservation tools and algorithms has been developed to assist the reserve selection process (Lawler et al., 2011). Simu-

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lated annealing (SA) as a heuristic algorithm inspired from annealing in metallurgy (Kirkpatrick et al., 1983) has been developed for purposes of optimization and spatial configuration (Aerts and Heuvelink 2002; Pressey, 2002) considering a ‘minimum set problem’ (McDonnell et al., 2002, Game and Grantham, 2008). SA is subject to iterative improvement, although it accepts bad moves randomly to prevent getting trapped in local minimum solutions (Ardron et al., 2010). The SA algorithm has been successfully used in conservation planning and reserve selection around the world (Airame et al., 2003; Andelman and Willig, 2002; Hermoso et al., 2010; Leslie et al., 2003). Its applicability was recently also tested in Iran (Mehri et al., 2014; Momeni et al., 2013). The main goal of this study was to prioritize habitats of two galliform species (Alectoris chukar and Phasianus colchicus) in Golestan Province and to introduce these areas as possible, new reserves to protect the declining populations of these species. We loosely linked the MCE approach and the SA algorithm within Marxan software (Ball and Possingham, 2000) to find the optimum network of habitats for conservation. The optimum solution in our research was defined as the selection of a conservation network with a minimum area that could meet the conservation targets of the species. Finding the optimum network is important in developing countries because of chronic shortages in funds for conservation efforts. Material and methods Study area Golestan Province is located in north east Iran (fig. 1). It has a total area of 20,430 km2 and diverse climatic and ecological conditions. The province is one of the richest areas in Iran in terms of biodiversity. Golestan National Park, Khoshyeylagh Wildlife Sanctuary, Jahannama, Loveh, and Zav protected areas constitute the reserve networks of the Golestan Province. However, all these reserves were selected on an ad hoc basis and do not necessarily encompass the most important areas for the protection of biodiversity (Momeni et al., 2013). The Golestan Province has three different climates: plain moderate, mountainous, and semi–arid. The average annual temperature is around 18º Celsius and the annual rainfall is around 550 mm (Malekinezhad and Zare–Garizi, 2014). The altitude in the study area varies between –15 m (in the vicinity of the Caspian Sea) and 3,363 m (Alborz mountains). The Hyrcanian forests and urban areas cover 17.5 and one percent of the study area, respectively. Habitat suitability We selected high priority habitats of Alectoris chukar and Phasianus colchicus in Golestan Province. As a first step, we developed habitat suitability maps for the species under study using the MCE method (Eq. 1) and the general species distribution data in the form of

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320000000000

360000000000

400000000000

440000000000

408000000000

412000000000

Golestan

416000000000

2 3

408000000000

412000000000

416000000000

420000000000

Golestan Reserves

420000000000

Iran

280000000000

404000000000

4 0 20 40 240000

000000

280000

000000

320000

000000

360000

000000

80 km

400000

000000

404000000000

240000000000

440000000000

Fig. 1. Location of Golestan Province (study area) in Iran: 1, Golestan National Park; 2, Zav Protected Area; 3, Loveh Protected Area; 4, Jahannama Protected Area; 5, KhoshYeylagh Wildlife Sanctuary. Fig. 1. Localización de la provincia de Golestán (zona de estudio) en Irán: 1, Parque Nacional de Golestán; 2, área protegida de Zav; 3, área protegida de Loveh; 4, área protegida de Jahannama; 5, refugio de vida silvestre de KhoshYeylagh.

polygons. Store and Kangas (2001) suggested using this method for areas with a lack of suitable data. These polygons had been drawn by wildlife wardens on general maps and used as guides to select the most appropriate environmental variables for habitat mapping. The species habitat requirements were specified using nominal habitat suitability models provided by Salmanmahiny (2008) (table 1). Altitude (METI and NASA, 2011), vegetation cover (DiMiceli et al., 2011), edge and interior diversity (neighborhood analysis of land–use), climatic variables (Hijmans et al., 2005) and distance to roads, rivers, and settlements (NCC, 2005) were used in this connection. Pairwise–correlation tests were applied to single out the uncorrelated parameters (r < 0.8). These information layers were studied inside and outside the general distribution polygons delimited by wildlife wardens, along with resorting to scientific documents on the species' habitat requirements (Mansoori, 2008; Salmanmahiny, 2008). This helped to define ranges for the selected factors reported in table 1. As habitat factors do not have the same importance for each species, weighting is necessary. We used the Analytic Hierarchy Process (AHP) for factor weights and applied the combination using equation 1 S = Σi = 1 to n WiXi

(Eq. 1)

where S is suitability, Wi, weight of the environmental variable i and Xi, value of environmental variable i.

Habitat suitability maps provided in step 1 were standardized in the range of 0 to 255 (figs. 2A, 2C), a common range in fuzzy calculations. We next used experts’ opinions to define a threshold limit for re– classification of the produced habitat suitability layers. We chose areas with values higher than 150 and a minimum size of 0.1 km2 (10 hectares) as suitable habitats (value = 1), and areas with values less than 150 as unsuitable habitats (value = 0) (figs. 2B, 2D). We defined a minimum area of 0.1 km2 because management of the small habitat patches was not feasible. Finally, to assess the accuracy of the habitat suitability models of the species, we calculated partial AUC (pAUC) models using few occurrences data (seven points for the Chukar Partridge and five points for the Common Pheasant) gathered by wildlife wardens. Calculations were made in Niche Analysts 3.0 software (Qiao et al., 2016). Reserve selection Binary habitat layers were fed into the Marxan software. Marxan is the most widely used and global leader of conservation planning software and has been used in approximately 184 countries for the selection of nature reserves (http://marxan.net). This software allows users to find the minimum number of sites needed to represent all conservation targets

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Table 1. Habitat requirements of Alectoris chukar (Ac) and Phasianus colchicus (Pc) and their relative weights (W) in the Golstan Province (Salmanmahiny, 2008): temp, temperature. Tabla 1. Requisitos del hábitat de Alectoris chukar y Phasianus colchicus y su peso relativo en la provincia de Golestán (Salmanmahiny, 2008): temp, temperatura.

Environmental variable Elevation (m)

Ac 0–3,500

W 0.02

Pc 0–2,500

Weights source

0.02 DEM (METI, NASA, 2011)

Tree cover (%)

0–20

0.10

0–39

0.16 MODIS (DiMiceli et al. 2011)

Herbal cover (%)

3–99

0.07

34–100

0.06 MODIS (DiMiceli et al. 2011)

Bare land (%)

0–80

0.10

0–50

0.09 MODIS (DiMiceli et al. 2011)

Edge diversity (Unitless)

1–42

0.13

1–42

0.17 Land–use

Interior diversity (Unitless)

1 to 6

0.13

2–6

0.17 Land–use

Minimum temp in coldest month (ºC) –15–1.2

0.02 –11.1–6.3 0.02 WorldClim (Hijmans et al., 2005)

Annual temp range (ºC)

34–42

0.02

–

Annual mean temp (ºC)

–

8.3–19.4

0.05 WorldClim (Hijmans et al., 2005)

5.7–15.9

0.01 WorldClim (Hijmans et al., 2005)

Mean temp in wettest season (ºC)

–

Mean temp in driest season (ºC)

–

Mean temp in coldest season (ºC) Mean temp in warmest season (ºC)

–5–7

–

WorldClim (Hijmans et al., 2005)

19 to 28.2 0.01 WorldClim (Hijmans et al., 2005)

0.02 –4.3–11.6 0.01 WorldClim (Hijmans et al., 2005)

–

20–30

0.01 WorldClim (Hijmans et al., 2005)

Annual precipitation (mm)

167–264

0.01

–

WorldClim (Hijmans et al., 2005)

Precipitation in wettest season (mm)

84–125

0.01

–

WorldClim (Hijmans et al., 2005)

Precipitation in driest season (mm)

10–21

0.01

–

WorldClim (Hijmans et al., 2005)

Precipitation in warmest season (mm) 10–27

0.01

–

WorldClim (Hijmans et al., 2005)

Precipitation in coldest season (mm)

0.01

–

WorldClim (Hijmans et al., 2005)

Distance to roads (m)

47–99

150–27,000 0.05 89–12,000 0.05 1: 25,000 topography map

Distance to rivers (m)

> 16,000

0.09 < 12,000

Distance to human settlement (m)

> 1,000

0.20

> 500

(NCC, 2005)

0.08 1: 25,000 topography map

(Ardron et al., 2010). The simulated annealing in the Marxan requires definition and selection of planning units, conservation targets, boundary length, and objective function. Planning units (PUs) are parts or parcels of land that Marxan works on during selection of the desired reserves. PUs must cover all parts of the study area and their size should be appropriate in terms of the species being considered for conservation and the size of the final reserves (Game and Grantham, 2008). Hexagons, watersheds, and grids are common PUs (Game and Grantham, 2008). The study area was partitioned into 20,430 hexagon planning units with a minimum area of 1 km2 (100 hectares), relatively the same number and area as that used in Heller et al. (2015). Targets are the quantitative values of each conservation feature to be achieved in the final reserve

(NCC, 2005)

0.09 1: 25,000 topography map

(NCC, 2005)

solution (Game and Grantham, 2008). We used a target of 20 %, meaning that the final reserves selected should contain at least 20 % of the total area of each species' habitats. We used 50 repeat runs and 10,000,000 iterations for Marxan. Boundary length modifier (BLM) is a multiplier that determines the importance of boundary length relative to the cost of the reserve. Normally, a trial and error approach is used to find an appropriate BLM value (Game and Grantham, 2008). We examined a range of BLM values (0 to 60) to select the optimum configuration of the selected habitats. The objective function used in Marxan (Eq. 2) is designed to minimize the total cost of the selected areas (Ball and Possingham, 2000). ∑ Cost + BLM ∑ Boundary + + ∑ SPF × Penalty + Cost Threshold Penalty (Eq. 2)

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High: 255 Low: 0

W C

0 1 E

0 1

High: 255 Low: 0 0 20 40

120 160 km

Fig. 2. Habitat suitability and Boolean maps for Alectoris chukar (A and B) and Phasianus colchicus (C and D). Fig. 2. Idoneidad del hábitat y mapas booleanos para Alectoris chukar (A y B) y Phasianus colchicus (C y D).

BLM = 0

BLM = 1

BLM = 10

BLM = 15

BLM = 20

BLM = 25

BLM = 30

BLM = 35

BLM = 40

BLM = 50

BLM = 60

Fig. 3. Selected reserves using different BLM values in Marxan to conserve at least 20 % of habitat for Alectoris chukar and Phasianus colchicus in the Golestan Province. Red lines indicate boundaries of current reserves. Fig. 3. Reservas seleccionadas utilizando distintos valores del modificador de longitud de frontera en Marxan para conservar al menos el 20 % del hábitat para Alectoris chukar y Phasianus colchicus en la provincia de Golestán. Las líneas rojas indican las fronteras de las reservas actuales.

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Total boundary length (km)

1,800 1,600

BLM = 0

1,400

BLM = 1

1,200 1,000 800

BLM = 10

600

BLM = 25BLM = 35 BLM = 15 BLM = 20 BLM = 30

400 200 0

465

470

475

480

BLM = 40

BLM = 60

BLM = 50

485 490 495 500 Total area (km2)

505

510

515

520

Fig. 4. Boundary/area curve used to find the optimum BLM value. Fig. 4. Curva frontera/superficie utilizada para encontrar el valor óptimo del modificador de longitud de frontera.

In our study, the cost was considered to be equal to the total area of the selected reserves, the boundary was the total length of the boundary surrounding the selected habitats, and both species received the same SPF (species penalty factor: weighting factor for the conservation feature). The penalty term is a penalty associated with each under–represented conservation feature, and the cost threshold penalty is a penalty applied to the objective function if the target cost is exceeded. Results Habitat suitability Figure 2 shows habitat suitability maps for Alectoris chukar and Phasianus colchicus. Accuracy assessment of models showed results were in the acceptable range as indicated by the calculated pAUCs, which were 0.773 and 0.719 for the chukar partridge and the common pheasant, respectively. Assessment of habitat suitability values at occurrence points showed that the selected threshold (150) was relatively close to the minimum suitability value of the occurrences (183.6 for the chukar partridge and 165 for the common pheasant). The areas of suitable habitat were mainly located in the southern parts of the province and represented 1,475 km2 for Alectoris chukar and 2,150 km2 for Phasianus colchicus. These areas are higher and have more vegetation cover than central and northern parts of the province.

Reserve selection Visual inspection of the selected habitats showed that applying higher BLM values caused selection of a more compact reserve system (fig. 3). The total area of the selected reserves ranged between 513 km2 for BLM = 60, and 469 km2 for BLM = 0 (fig. 4). Boundary/area comparison To select an optimal BLM value, the total area and total boundary length of the reserve system were important factors. Stewart and Possingham (2005) suggested that boundary/area curve is a good tool to select an optimum BLM value. Figure 4 compares boundary/area in the various BLM values used in this study. According to the boundary/area ratio, when the BLM value is 15, the selected reserve system is the optimum solution. Discussion In recent years, Iran’s Department of Environment (DOE) has endeavored to conserve biodiversity but because of the shortage of funds and experts, these efforts have not fully achieved their set goals (Makhdoum, 2008). In most cases, protected areas in Iran have been selected on an ad hoc basis and consequently they do not necessarily fit conservation objectives and goals as shown by Momeni et al. (2013). Hence, it is time to use and apply up–to–date methods for reserve selection and species conservation.

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A comparison of the optimum reserves selected by Marxan (BLM = 15) and the current reserve system in Golestan Province showed little overlap between the optimum reserves and the available reserves (from seven newly identified patched, only two patches are located in the current reserves). The current reserve network contained only 23 % (112 km2) of the total area of reserve selected by Marxan (476 km2). Hence, we found it necessary to introduce new zones as protected areas to conserve 20 % of the suitable habitats of Alectoris chukar and Phasianus colchicus in the Golestan Province. The key note about the newly selected areas is that these areas are mainly near the current reserves and so it is relatively straightforward to complement current reserves through corridors. Because this study is based only on two bird species, we suggest hunting restrictions in selected areas (no–hunting area). If future studies indicate that these patches are also important habitats of other species, then promoting conservation level of the selected patches to protected areas can be considered. It is also notable that primary polygons of suitable habitats of species and occurrence data are gathered from experienced wildlife wardens without any sampling design and hence there might be some bias in the results, which is worthy of further research. Furthermore, the results of this study showed the possibility to approach conservation of the target species even in regions with limited data on the occurrence of the species and lack of habitat suitability maps. This lack was partially tackled using the coarse filter multi–criteria evaluation approach that refined the general and large areas of occurrences defined by field experts and wildlife wardens. In Iran, like many other developing countries, the data for some species distribution is limited and of low–quality, making conservation planning more difficult in these instances. As suggested by Store and Kangas (2001), we showed that the MCE approach coupled with the systematic reserve selection, namely Marxan, can help researchers form a good general picture of suitable habitats for a species to be conserved. Acknowledgements We express our gratitude to the anonymous reviewers of the earlier version of the manuscript. References Aerts, J., Heuvelink, G., 2002. Using Simulated Annealing for Resource Allocation. International Journal of Geographical Information Science, 16 (6): 571–587. Airame, S., Dugan, J. E., Lafferty, K. D., Leslie, H., McArdle, D. A., Warner, R., 2003. Applying Ecological Criteria to Marine Reserve Design: A Case Study from The California Channel Islands. Ecological applications, 13(1): 170–184. Andelman, S. J., Willig. M. R., 2002. Alternative

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Animal Biodiversity and Conservation 41.1 (2018)

Animal Biodiversity and Conservation

Manuscrits

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

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

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

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

ISSN: 1578–665X eISSN: 2014–928X

Introducción. S'hi donarà una idea dels antecedents del tema tractat, així com dels objectius del treball. Material y métodos. Inclourà la informació perti nent de les espècies estudiades, aparells emprats, mètodes d’estudi i d’anàlisi de les dades i zona d’estudi. Resultados. En aquesta secció es presentaran úni cament les dades obtingudes que no hagin estat publicades prèviament. Discusión. Es discutiran els resultats i es compa raran amb treballs relacionats. Els suggeriments de recerques futures es podran incloure al final d’aquest apartat. Agradecimientos (optatiu). Referencias. Cada treball haurà d’anar acom panyat de les referències bibliogràfiques citades en el text. Les referències han de presentar–se segons els models següents (mètode Harvard): * Articles de revista: Conroy, M. J., Noon, B. R., 1996. Mapping of species richness for conservation of biological diversity: conceptual and methodological issues. Ecological Applications, 6: 763–773. * Llibres o altres publicacions no periòdiques: Seber, G. A. F., 1982. The estimation of animal abundance. C. Griffin & Company, London. * Treballs de contribució en llibres: Macdonald, D. W., Johnson, D. P., 2001. Dispersal in theory and practice: consequences for conservation biology. In: Dispersal: 358–372 (T. J. Clober, E. Danchin, A. A. Dhondt, J. D. Nichols, Eds.). Oxford University Press, Oxford. * Tesis doctorals: Merilä, J., 1996. Genetic and quantitative trait vari ation in natural bird populations. Tesis doctoral, Uppsala University. * Els treballs en premsa només han d’ésser citats si han estat acceptats per a la publicació: Ripoll, M. (in press). The relevance of population studies to conservation biology: a review. Animal Biodiversity and Conservation. La relació de referències bibliogràfiques d’un tre ball serà establerta i s’ordenarà alfabèticament per autors i cronològicament per a un mateix autor, afegint les lletres a, b, c,... als treballs del mateix any. En el text, s’indic aran en la forma usual: "... segons Wemmer (1998)...", "...ha estat definit per

Robinson i Redford (1991)...", "...les prospeccions realitzades (Begon et al., 1999)...". Taules. Es numeraran 1, 2, 3, etc. i han de ser sempre ressenyades en el text. Les taules grans seran més estretes i llargues que amples i curtes ja que s'han d'encaixar en l'amplada de la caixa de la revista. Figures. Tota classe d’il·lustracions (gràfics, figures o fotografies) entraran amb el nom de figura i es numeraran 1, 2, 3, etc. i han de ser sempre 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 41.1 (2018)

III

Animal Biodiversity and Conservation

Manuscritos

Animal Biodiversity and Conservation es una revista inter disciplinar, publicada desde 1958 por el Museo Ciencias Naturales de Barcelona. Incluye artículos de investigación empírica y teórica en todas las áreas de la zoología (sistemática, taxo nomía, morfología, biogeografía, ecología, etología, fisiología y genética) procedentes de todas las regiones del mundo, con especial énfasis en los estudios que permitan comprender, desde un punto de vista pluri disciplinar e integrado, los patrones de evolución de la biodiversidad en su sentido más amplio. La revista no publica compilaciones bibliográficas, catálogos, listas de especies sin más o citas puntuales. Los estudios realizados con especies raras o protegidas pueden no ser aceptados a no ser que los autores dispongan de los permisos correspondientes. Cada volumen anual consta de dos fascículos. Animal Biodiversity and Conservation está 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.

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. Los caracteres en cursiva se utilizarán para los nombres científicos de géneros y especies y para los neologismos que no tengan traducción; las citas textuales, independientemente de la lengua en que estén, irán en letra redonda y entre comillas; el nombre del autor que sigue a un taxón se escribirá también en redonda. Se eviatá el uso de términos extranjeros (latín, aleman,...) Al citar por primera vez una especie en el trabajo, deberá especificarse siempre que sea posible su nombre común. Los topónimos se escribirán bien en su forma original o bien en la lengua en que esté redactado el trabajo, siguiendo el mismo criterio a lo largo de todo el artículo. Los números del uno al nueve se escribirán con letras, a excepción de cuando precedan una unidad de medida. Los números mayores de nueve se escribirán con cifras excepto al empezar una frase. Las fechas se indicarán de la siguiente forma: 28 VI 99 (un único día); 28, 30 VI 99 (días 28 y 30); 28–30 VI 99 (días 28 al 30). Se evitarán siempre las notas a pie de página.

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. Las pruebas de imprenta enviadas a los autores de berán remitirse corregidas al Consejo Editor en el plazo máximo de 10 días. Los gastos debidos a modificaciones sustanciales en las pruebas de imprenta, introducidas por los autores, irán a cargo de los mismos. El primer autor recibirá una copia electrónica del trabajo en formato PDF. ISSN: 1578–665X eISSN: 2014–928X

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.) © 2018 Museu de Ciències Naturals de Barcelona

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 species richness for conservation of biological diversity: conceptual and methodological issues. Ecological Applications, 6: 763–773. * Libros y otras publicaciones no periódicas: Seber, G. A. F., 1982. The estimation of animal abundance. C. Griffin & Company, London. * Trabajos de contribución en libros: Macdonald, D. W., Johnson, D. P., 2001. Dispersal in theory and practice: consequences for conservation biology. In: Dispersal: 358–372 (T. J. Clober, E. Danchin, A. A. Dhondt, J. D. Nichols, Eds.). Oxford University Press, Oxford. * Tesis doctorales: Merilä, J., 1996. Genetic and quantitative trait 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 y Redford (1991)...", "...las prospecciones realizadas (Begon et al., 1999)...". Tablas. Se numerarán 1, 2, 3, etc. y se reseñarán todas en el texto. Las tablas grandes deben ser más estrechas y largas que anchas y cortas ya que deben ajustarse a la caja de la revista. Figuras. Toda clase de ilustraciones (gráficas, figuras o fotografías) se considerarán figuras, se numerarán 1, 2, 3, etc. y se citarán todas en el texto. Pueden incluirse fotografías si son imprescindibles. Si las fotografías son en color, el coste de su publicación irá a cargo de los autores. El tamaño máximo de las figuras es de 15,5 cm de ancho y 24 cm de alto. Deben evitarse las figuras tridimensionales. Tanto los mapas como los dibujos deben incluir la escala. Los sombreados preferibles son blanco, negro o trama. Deben evitarse los punteados ya que no se reproducen bien. Pies de figura y cabeceras de tabla. Serán claros, concisos y bilingües en castellano e inglés. Los títulos de los apartados generales del artículo (Introducción, Material y métodos, Resultados, Discusión, Agradecimientos y Referencias) no se numerarán. No utilizar más de tres niveles de títulos. Los autores procurarán que sus trabajos originales no excedan las 20 páginas incluidas figuras y tablas. Si en el artículo se describen nuevos taxones, es imprescindible que los tipos estén depositados en alguna institución pública. Se recomienda a los autores la consulta de 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 41.1 (2018)

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 untranslatable neologisms must be in italics. Quota tions 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. Avoid the use of foreing terms (Latin, German,...) 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. 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.

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

ISSN: 1578–665X eISSN: 2014–928X

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 species richness for conservation of biological diversity: conceptual and methodological issues. Ecological Applications, 6: 763–773. * Books or other non–periodical publications: Seber, G. A. F., 1982. The estimation of animal abundance. C. Griffin & Company, London. * Contributions or chapters of books: Macdonald, D. W., Johnson, D. P., 2001. Dispersal in theory and practice: consequences for conservation biology. In: Dispersal: 358–372 (T. J. Clober, E. Danchin, A. A. Dhondt, J. D. Nichols, Eds.). Oxford University Press, Oxford. * PhD thesis: Merilä, J., 1996. Genetic and quantitative trait variation in natural bird populations. PhD thesis, Uppsala University. * Works in press should only be cited if they have been accepted for publication: Ripoll, M. (in press). The relevance of population studies to conservation biology: a review. Animal Biodiversity and Conservation. References must be set out in alphabetical and 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 and Redford (1991)...", "...the prospections that have been carried out (Begon et al., 1999)..." Tables. Must be numbered in Arabic numerals with reference in the text. Large tables should be narrow (across the page) and long (down the page) rather than wide and short, so that they can be fitted into the column width of the journal. Figures. All illustrations (graphs, drawings, 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 41.1 (2018)

VII

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161–174 Cordero–Rivera, A., Zhang, H. Ethological uniqueness of a damselfly with no near relatives: the relevance of behaviour as part of biodiversity 175–184 Garbino, G. S. T., Rezende, G. C., Fernandes– Ferreira, H., Feijó, A. Reconsidering mammal extinctions in the Pernambuco Endemism Center of the Brazilian Atlantic Forest

185–193 Momeni Dehaghi, I., Salmanmahiny, A., Karimi, S., Shabani, A. A. Multi–criteria evaluation and simulated annealing for delimiting high priority habitats of Alectoris chukar and Phasianus colchicus in Iran

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

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

Índex / Índice / Contents Animal Biodiversity and Conservation 41.1 (2018) ISSN 1578–665 X eISSN 2014–928 X 1–8 Obregón, R., Jordano, D., Cuadrado, M., Moreno– Benítez, J. M., Fernández Haeger, J. Dispersal of the monarch butterfly (Danaus plexippus) over southern Spain from its breeding grounds 9–17 Shekhovtsov, S. V., Sundukov, Yu. N., Blakemore, R. J., Gongalsky, K. B., Peltek, S. E. Identifying earthworms (Oligochaeta, Megadrili) of the Southern Kuril Islands using DNA barcodes 19–32 Pétillon. J., François, A., Lafage, D. Short–term effects of horse grazing on spider assemblages of a dry meadow (Western France) 33–60 Gée, A., Sarasa, M., Pays, O. Long–term variation of demographic parameters in four small game species in Europe: opportunities and limits to test for a global pattern 61–73 Telahigue, K., Hajji, T., El Cafsi, M., Saavedra, C. Genetic structure and demographic history of the endemic Mediterranean scallop Pecten jacobaeus inferred from mitochondrial 16s DNA sequence analysis 75–95 Charra, M., Sarasa, M. Applying IUCN Red List criteria to birds at different geographical scales: similarities and differences

97–108 Torma, A., Bozsó, M., Gallé, R. Secondary habitats are important in biodiversity conservation: a case study on orthopterans along ditch banks 109–115 Książkiewicz–Parulska, Z. The light–dark cycle of Desmoulin’s whorl snail Vertigo moulinsiana Dupuy, 1849 (Gastropoda, Pulmonata, Vertiginidae) and its activity patterns at different temperatures 117–120 Rodrigues, V. B., Jesus, F. M., Campos, R. I. Local habitat disturbance increases bird nest predation in the Brazilian Atlantic rainforest 121–130 Díaz–Páez, H., Canales–Arévalo, C. Effect of temperature and type of diet on the metamorphosis of Pleurodema thaul (Lesson, 1826) in a population of south–central Chile 131–140 Moreno, Á. C., Carrascal, L. M., Delgado, A., Suárez, V., Seoane, J. Striking resilience of an island endemic bird to a severe perturbation: the case of the Gran Canaria blue chaffinch 141–159 Arriaga–Flores, J. C., Rodríguez–Moreno, A., Correa– Sandoval, A., Horta–Vega, J. V., Castro–Arellano, I., Vázquez–Reyes, C. J., Venegas–Barrera, C. S. Spatial and environmental variation in phyllostomid bat (Chiroptera, Phyllostomidae) distribution in Mexico

Amb el suport de / Con el apoyo de / With the support of: