ABC 40.1 (2017) by Museu Ciències Naturals de Barcelona MCNB

Ben J. Hatchwell, Univ. of Sheffield, UK Editor en cap / Editor responsable / Editor in Chief

2Joan Carles Senar

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

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

Assessorament lingüístic / Asesoramiento lingüístico / Linguistic advisers Carolyn Newey Pilar Nuñez Animal Biodiversity and Conservation 40.1, 2017 © 2017 Museu de Ciències Naturals de Barcelona, Consorci format per l'Ajuntament de Barcelona i la Generalitat de Catalunya Autoedició: Montserrat Ferrer Fotomecànica i impressió: Inspyrame Printing ISSN: 1578–665 X eISSN: 2014–928 X Dipòsit legal: B. 5357–2013 Animal Biodiversity and Conservation es publica amb el suport de: l'Asociación Española de Ecología Terrestre, la Sociedad Española de Etología y Ecología Evolutiva i la Sociedad Española de Biología Evolutiva The journal is freely available online at: www.abc.museucienciesjournals.cat Dibuix de la coberta: Aphanius iberus, fartet, fartet, Spanish toothcarp (Jordi Domènech)

Animal Biodiversity and Conservation 40.1 (2017)

Hunting passerines with non–selective trapping methods was a source of conflict in Spain as far back as 1933 J. J. Ferrero–García

Ferrero–García, J. J., 2017. Hunting passerines with non–selective trapping methods was a source of conflict in Spain as far back as 1933. Animal Biodiversity and Conservation, 40.1: 1–6. Abstract Hunting passerines with non–selective trapping methods was a source of conflict in Spain as far back as 1933.— We here show unpublished documentation regarding a complaint presented to the Spanish Government by the Iberian Federation of Societies for the Protection of Animals and Plants in 1933. This complaint concerned apparent non–compliance with the International Convention for the Protection of Birds (1902). The reason was hunting with non–selective trapping methods (nets and birdlime) that were prohibited by the convention but authorized in certain cases by the Spanish Government in 1929. Such hunting could have contributed to the elimination of large num� bers of passerines, some protected by law. According to the documentation studied, the complaint from this Iberian Federation was triggered by a letter sent by Léon Pittet, president of the Comité National Suisse pour la Protection des Oiseaux. This event emphasizes the relationships between European organizations whose purpose was the conservation of birds, and certain Spanish associations whose objectives included the defense of passerines in the years before the Spanish Ornithological Society was created. In addition, it indicates that the 1902 Convention had some positive practical consequences, although these later decreased due to pressure from important hunting sec� tors in Spain. The case presented here shows that the current conflict in Spain between the use of certain hunting methods and legislation for the conservation of birds dates back at least to the first half of the twentieth century. Key words: History of conservation, International agreements, Hunting passerines, Parany, Protected species Resumen La caza de paseriformes con métodos de captura no selectivos ha sido una fuente de conflictos en España desde 1933.— En este estudio se muestra documentación inédita referente a una queja presentada al Gobierno de España en 1933 por la Federación Ibérica de Sociedades Protectoras de Animales y Plantas, por el aparente incumplimiento de la Convención Internacional para la Protección de las Aves (1902). El motivo era la caza con métodos de captura no selectivos (redes y liga), que estaban prohibidos por dicho tratado, pero que el Gobierno de España, en 1929, había autorizado en determinados casos. Este tipo de caza pudo contribuir a la eliminación de grandes cantidades de paseriformes, algunos protegidos por ley. Según la documentación estudiada, la queja de esta federación fue impulsada por una carta remitida por Léon Pittet, presidente del Comité National Suisse pour la Protection des Oiseaux. Estos hechos ponen de relieve las relaciones existentes entre las organizaciones europeas, cuya finalidad era la conservación de las aves, y ciertas asociaciones españolas, entre cuyos objetivos figuraba también la defensa de los paseriformes, en una época en la que aún no se había constituido la Sociedad Española de Ornitología. Además, indican que la Convención de 1902 tuvo algunas consecuencias prácticas po� sitivas, que posteriormente disminuyeron debido a la presión ejercida desde importantes sectores cinegéticos de España. El caso que se presenta aquí pone de manifiesto que el conflicto que existe actualmente en el país entre determinados métodos de caza y las leyes para la conservación de las aves se remonta al menos a la primera mitad del siglo XX. Palabras clave: Historia de la conservación, Acuerdos internacionales, Caza de paseriformes, Parany, Especies protegidas Received: 4 V 16; Conditional acceptance: 30 V 16; Final acceptance: 17 VI 16 J. J. Ferrero–García, Consejería de Medio Ambiente y Rural, Políticas Agrarias y Territorio, Avda. Luis Ramallo, s/n., 06800 Mérida, Badajoz, Spain. E–mail: juanjo.ferrero.g@gmail.com ISSN: 1578–665 X eISSN: 2014–928 X

Ferrero–García

Introduction In 1902, twelve European States signed the Interna� tional Convention for the Protection of Birds Useful to Agriculture (henceforth the Paris Convention) that remained fully valid until 1950 when it was replaced by the International Convention for the Protection of Birds (Van Heijnsbergen, 1997; Boardman, 2006; Bowman et al., 2010). The 1902 Paris Convention can be considered the first legally binding international re� gulation for the defence of wildlife (Van Heijnsbergen, 1997; Kiss & Shelton, 2007). Over the years, it has been criticized for its anthropocentric vision based on the division of animals into useful and harmful (Fe� rrero–García, 2013, 2014), though it has also been praised, among other things, for prohibiting massive and non–selective trapping methods (Gillespie, 2011; Sands & Peel, 2012). Currently in the European Union, such methods can only be used for bird hunting in exceptional cases (EPC, 2010). But despite such legislation, one region in Spain (the Comunidad Valenciana) has attempted to legalize the use of some these methods (specifically the parany, a method of capture based on the use of birdlime, which catches the birds when they alight on trees pruned especially for this purpose). Nevertheless, such attempts have all failed in the courts of justice and have been opposed by the Spanish Government because of the lack of selectivity with such strategies. It has been shown in this region that the parany method has been responsible for the elimination of many bird species, including 1.5–2 million thrushes Turdus spp. per year (Bort, 2005, 2006; González & Vega, 2009; Giménez, 2010, 2013; Murgui, 2014). The recent ruling of the Constitutional Tribunal of Spain (TC, 2013) is of particular interest, but despite such legislation, parany remains a source of conflict in Spain (CC SEO/BirdLife, 2013; Díaz et al., 2016). Some authors consider that the Paris Convention had few positive practical consequences (Boardman, 2006; Bowman, 2014). However, recent studies sug� gest that these pioneer regulations —or the current of thought that promoted them in the 19th century— could have benefited some birds, particularly passerines (Torres–Vila et al., 2015; Ferrero–García et al., in press). Spain signed the 1902 Paris Convention thanks to the work of Mariano de la Paz Graells (Ferrero–García, 2012b), who in 1896 drafted the first Spanish catalogue of protected birds (Royal Order of 25 November 1896). The birds in this catalogue were included on the basis of their status as insec� tivorous birds (Ferrero–García, 2011, 2012a; Casado, 2013). Seventy percent of the birds were passerines (Ferrero–García, 2011). Graells, a prestigious 19th century Spanish zoologist (Cervantes, 2009), was thus the precursor of the first legal steps towards conservation of birds in Spain. Several protective measures were approved at local levels around the same period (Ferrero–García et al., 2014), and they sometimes included a ban on hunting birds with nets (Torres–Vila et al., 2009). A few years later, the content of the 1896 catalogue was incorporated into the Regulation of the 1902 Hunting Law. For more

than half a century, these were the main hunting and fauna protection regulations in Spain (Martínez, 1998; Ferrero–García, 2010). This study presents previously unpublished docu� mentation that testifies how, in 1933, certain asso� ciations filed a complaint to the Spanish Government concerning the use of massive and non–selective methods of capture (nets and birdlime) that contributed to the possible elimination of large numbers of pas� serines, some of which were protected. By examining these archival sources, this paper aims to assess the effective application in practical terms of the 1902 Paris Convention in Spain. Early efforts to promote the pro� tection of birds in Spain, prior to the foundation of the Spanish Ornithological Society (Sociedad Española de Ornitología, SEO; now SEO/BirdLife), are also docu� mented. SEO was founded in 1954 (De Juana, 2004; Fernández, 2004). Furthermore, the paper links these events to the current conflict between certain types of hunting and bird conservation. Material and methods This research is based on documents from the General Archive of the Ministry of Foreign Affairs and Cooperation in Madrid (Spain). We reviewed documents from the archive concerning the Paris Convention (catalogue number: AGMAE, R65558, exp. 5; henceforth the AGMAE). More specifically, we reviewed a three–page letter signed on 25 Febru� ary 1933 by Joaquín Juliá, on behalf of the Madrid section of the Iberian Federation of Societies for the Protection of Animals and Plants (Federación Ibérica de Sociedades Protectoras de Animales y Plantas; henceforth the FSPAP) and addressed to the Minis� try of State (former name of the Ministry of Foreign Affairs in Spain), where it was received on 3 March. We also studied the reply from the Ministry of State and the reply from the Ministry of Agriculture (the Ministry of State requests a report to the Ministry of Agriculture). This report was signed on 6 Janu� ary 1934 by Miguel Pastor, General Director of the Forests, Fishing and Hunting sector of the Ministry of Agriculture and received at the Ministry of State on 10 January. To support the discussion, we reviewed publications of the time such as texts about the in� ternational agreements for the protection of birds and the conservationist law, the Spanish official bulletin of the time (Gaceta de Madrid), the daily press (Span� ish newspapers as La Vanguardia and the Swiss Le Confédéré) and some ornithological journals (Der Ornithologischer Beobachter). Results In the writing of 25 February 1933, Juliá began by alluding to another letter received in the FSPAP, in December 1932, in which Dr. Léon Pittet, president of the Comité National Suisse pour la Protection des Oiseaux (henceforth the CNSPO), requested informa� tion from the conservationist legislation in Spain to

Animal Biodiversity and Conservation 40.1 (2017)

substantiate a complaint to the Secretariat of the Paris Convention and the League of Nations against the Spanish Government for the breach of its international commitments (Pittet´s letter does not appear in the file consulted and therefore it is not possible to know when it was written). Juliá claimed that, according to Pittet, the number of migrant birds in Switzerland had decreased in the two previous years, attributing this to their massive elimination in Spain where they were hunted with nets and birdlime, both prohibited by Arti� cle 3 of the Paris Convention. Next, it stated what, in the FSPAP’s opinion, was the source of the problem: the Royal Order of 6 September 1929 (henceforth the Royal Order of 1929), which declared it legal in Spain to hunt non–insectivorous birds with nets or birdlime from 1 September to 31 January. Juliá argued that any type of bird, including Hirundinidae —he mentioned swallows 'golondrinas'—, were hunted throughout the year under that excuse. He also detailed a case that occurred in October 1931 in Valencia, when the FSPAP complained to the authorities about the use of nets for catching large quantities of swallows. He also complained about the daily use of nets and birdlime in Madrid for the indiscriminate hunting of birds for their sale and consumption in catering establishments. (See this letter in supplementary material). The Ministry of State forwarded the letter to the Ministry of Agriculture in March 1933, warning about possible international consequences of the complaint and asking whether to repeal the Royal Order of 1929 (AGMAE). Almost a year later, on 6 January 1934, the Ministry of Agriculture presented a report to the Ministry of State (AGMAE). This report stated that no evidence had been provided to link the decline of birds in Switzerland with the alleged massacres carried out in Spain. It then focused on the possible infraction of the Paris Convention, affirming that there was no breach at all. To substantiate this, the Ministry of Agriculture analyzed both the referred treaty and Spanish regulations. According to its conclusions, Article 3 of the Paris Convention (banning the use of nets and birdlime) was invalidated by Article 9 (excep� tions), due to the fact that Spain had a list of legally protected species (insectivorous birds) since 1896 —Royal Order of 25 November 1896 and Regulation of the 1902 Hunting Law—, making the use of these methods of capture possible for the remaining birds. The Ministry of State wrote to the FSPAP in January 1934, sending them a copy of the report and closing the matter (AGMAE). We have been unable to locate any documentation in the AGMAE that would allow us to know the opinion of the FSPAP in relation to this response, or whether this entity had communicated the reply to the CNSPO. Neither have we found any information about complaints —if there were any— from the CNSPO against the Spanish Government. Discussion The first efforts to prevent animal cruelty emerged in the early 19th century (Nash, 1989; Baker, 2015). In the second half of the 19th century, various animal

protection associations at a local level were consti� tuted in Spain (Ferrero–García, 2010). In 1925 the Madrid FSPAP section was created, as had occurred earlier in Barcelona and other places in Spain (see La Vanguardia, 1925a; 1925b). Also in 1925, the Span� ish Government declared these associations to be of public utility, and in 1928 its members were granted the status of agents of authority (Pérez, 2015). Its statutes were approved in 1933 (MG, 1933). These protection associations applied humanitarian prin� ciples and moral concerns to a variety of topics related to animal wellbeing, prevention of cruelty, proper care of pets and domestic animals and the like. But in some cases —as is shown in this paper— they also participated in conservation debates and proposals. In this sense, agronomist engineers such as Zacarías Salazar, doctors such as Eduardo Alfonso, and also scientists such as Ángel Cabrera and Cándido Bolívar were members of the FSPAP (see La Asociación, 1925); Cabrera and Bolívar were prestigious Spanish zoologists (Gomis, 1998; Merino, 2002b). The FSPAP sometimes had, among its objectives, the protection of at least some groups of wild birds —most probably above all passerines— as we can see in Juliá’s writing (where a vernacular name 'pájaro' was used instead of the word 'ave', bird). 'Pájaro' has several meanings (Bernis, 1995), but it seems most probable that Juliá referred primarily to passerines, because the use of nets and birdlime had aimed at hunting passerines such as starlings Sturnus spp. and thrushes Turdus spp. (Parsons, 1960; Giménez, 2010; Murgui, 2014). Another example appeared a few years earlier when Joaquina Casablancas, represent� ing the FSPAP, denounced bird hunting with paranys in Arenys del Mar (Barcelona), so the authorities took part confiscating and destroying them (see La Vanguardia, 1924a). The FSPAP also took a stand against the illegal sale of passerines for consumption (see La Vanguardia, 1924b). Overall, although massive and non–selective meth� ods were used in Spain during the first third of the 20th century (Parsons, 1960), opposition began to increase. There are examples, although they are not related to the FSPAP, where the massive losses provoked in insectivorous species —which were also protected by the Paris Convention (Herman, 1907; ME, 1907)— were denounced (AO, 1909). Some public institutions requested the complete disappearance of hunting with nets due to the damage caused to useful avifauna and agriculture (CPFT, 1911). Even in the hunting field, some voices were raised in criticism of the use of methods such as the 'paranys' (Bernat, 1924). Nevertheless, even today, the illegal use of some methods of catching birds remains a major problem today in several Mediter� ranean countries —including Spain— and affects many passerines (Brochet et al., 2016). L. Pittet (1866–1939) was a physician and ornitholo� gist who presided Ala, the Swiss Society for the Study and Conservation of Birds, between 1928 and 1932 (Bruderer & Marti, 2009). As president of the CNSPO and also as Swiss delegate at the International Com� mittee for Bird Protection (ICBP), Pittet sometimes complained about the ineffectiveness of the Paris

Ferrero–García

Convention regarding the abuse committed by some countries during the migration of birds (see Pittet, 1929; Le Confédéré, 1932). Pittet had already drawn attention to the indiscriminate killings of birds in Italy, France, Spain and Belgium, and had welcomed the constitution of the ICBP (Anonymous, 1925). In 1922, the ICBP was designed to co–ordinate the activities of national NGOs concerned with avian conservation, not only within Europe but also across the Atlantic (Bow� man, 2014). The ICBP, which later became BirdLife (Bowman, 2014), was the first international organization concerned with the preservation of wildlife (Campbell & Lack, 1985). Regarding Juliá, his link to the FSPAP is known since the association was first created (see La Asociación, 1925). He was the General Secretary of the Third International Congress of the Societies for Protection of Animals (ABC, 1927; Juliá, 1927), and also President of the International Office in Paris of the Societies for the Protection of Animals (Wöbse, 2012). Ultimately, evidence suggests that important members of the FSPAP (Juliá) maintained contacts with prominent members of the ICBP, the CNSPO and the Ala (Pittet), with the aim of improving the protection of birds in Spain throughout is over twenty years before the creation of the SEO. Thanks to the SEO, in 1963 the Spanish section of the ICBP was established (Anónimo, 1963; Fernández, 2004). But why did the Spanish Government approve the Royal Order of 1929? The answer is found in the text of this norm adopted by the Minister Rafael Benjumea (MF, 1929), which states that its purpose was to satisfy the demands presented by the Royal Association of Hunters and Fishers of Spain (Real Asociación de Cazadores y Pescadores de España), through their president, Fernando Luca de Tena. Hunters complained about the pressure put on them by some authorities that understood that the use of nets, birdlime and claims was an illegality (Article 3 of the Paris Convention and Regulation of the 1902 Hunting Law). As previously explained, such pressure was partly the result of the actions of the FSPAP. It is of note, on the other hand, that large scale hunting of some passerines was increasing at that time because of the high economic benefits obtained through the commercializing the captured birds (González, 1993). Also of note was the contribution of Benjumea, an engineer who developed intense activity as Minister of the Spanish Government (Merino, 2002a). It therefore seems reasonable to consider that, for nearly three decades, some Spanish authorities tried to strictly comply with Article 3 of the Paris Convention. At the same time some measures were promoted to protect natural spaces, including the declaration of the first Spanish National Parks (Casado, 2002, 2010). It seems that no fully satisfactory results were obtained, however, despite the Spanish Government passing various provisions over these years recalling the need to respect the laws protecting birds (Ferrero–García, 2010). In fact, at the beginning of the 20th century, it was said that the main problem facing bird conserva� tion in Spain was the insufficient compliance with laws (Macpherson, 1909). Nevertheless, in the 1920s, the

Spanish Government tried in several ways to appease the existing discontent among hunters, who consid� ered that the executive had not acted adequately to resolve their problems (González, 1993). In 1929, the Spanish Government wanted to close the mass and non–selective methods issue, explaining, between September 1 and January 31, that the use of nets and birdlime for hunting non– insectivorous birds was completely legal. The attempt of the FSPAP to reverse this situation, in 1933, was fruitless. Finally, we can deduce some issues in relation to the report of the Ministry of Agriculture: 1) as is well known, legal con� servationist actions should be based, both today and a century ago, on scientific rationale (Bertouille, 2012; Casado, 2013); and 2) as suggested in other studies (Bowman et al., 2010; Bowman, 2014), the regime of exceptions in the Paris Convention has hindered the achievement of the objectives of the treaty. In conclusion, in Spain, and partly thanks to the involvement of animal protection associations, for nearly 30 years the Paris Convention seems to have had some positive practical consequences for the conservation of birds, at least for passerines. However, these advances decreased after 1929 due to pressure from important sectors linked to the hunting activities. A similar situation reoccurred, but with more dramatic consequences, when in 1950, Spain signed the new Convention of Paris, while simultaneously adopting a standard —supported by the hunting sectors— for the extermination of, among other animals, most diurnal birds of prey (Ferrero–García, 2015). Moreover, even today, some hunting practices are causing major conservation problems in Spain and other countries of the Mediterranean region. Acknowledgements I sincerely thank Santos Casado for useful comments on a previous version of the manuscript, María del Carmen de San José Moreno for her help, and staff at the Archives of the Ministry of Foreign Affairs and Cooperation (Madrid). I am also grateful to Blanca Moreno Fontela for advice on the English. Mario Díaz and José Jiménez García–Herrera made valuable comments that helped to improve the manuscript. References ABC, 1927. El Congreso Internacional de las So� ciedades Protectoras de Animales. ABC, 7744 [25 June 1927]: 25–26. Anónimo, 1963. El Consejo Internacional para la Conservación de las Aves: la Sección española. Ardeola, 8: 297–301. Anonymous, 1925. Une Association mondiale pour la protection des Oiseaux. Der Ornithologischer Beobachter, 22 (supplement to 8th): 135: 1–16. AO, 1909. Notas bibliográficas. La Vanguardia, 13301 [25 December 1909]: 2. Baker, L. W., 2015. Animal Rights and Welfare: A Documentary and Reference Guide. ABC–CLIO,

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Animal Biodiversity and Conservation 40.1 (2017)

Underlying factors promoting nestedness of bird assemblages in cays of the Jardines de la Reina archipelago, Cuba A. García–Quintas & A. Parada Isada

García–Quintas, A. & Parada Isada, A., 2017. Underlying factors promoting nestedness of bird assemblages in cays of the Jardines de la Reina archipelago, Cuba. Animal Biodiversity and Conservation, 40.1: 7–16. Abstract Underlying factors promoting nestedness of bird assemblages in cays of the Jardines de la Reina archipelago, Cuba.— Assessing the factors associated with nestedness patterns is a crucial aspect in studies of community structure. Bird assemblages in the Jardines de la Reina archipelago have a stable nested structure but the underlying influences have not been evaluated. We constructed a presence–absence data matrix based on a bird inventory obtained from 43 cays of this archipelago. We calculated nestedness using the NODF metric based on the overlap and decreasing fill and evaluated its significance by running 1,000 iterations of four null models. The matrix columns were rearranged to evaluate seven factors possibly related to the nestedness of bird communities. Bird assemblages exhibited a significant nested pattern (67.93) and all factors contributed (p < 0.01) to the nestedness patterns of bird communities. Habitat diversity and cay area and perimeter were the factors that contributed most to the nested structure. The nestedness pattern in the bird assemblages of the Jardines de la Reina archipelago was potentially caused by the interaction of selective extinction and differential colonization of species, with the former having a more remarkable effect. Key words: Nested structure, Fragmented biota, Selective extinction, Differential colonization Resumen Factores subyacentes que promueven el anidamiento de ensamblajes de aves en cayos del archipiélago de los Jardines de la Reina, Cuba.— La evaluación de los factores asociados a los modelos de anidamiento se ha convertido en un aspecto esencial de los estudios sobre estructuración de comunidades. Los ensambla� jes de aves del archipiélago de los Jardines de la Reina presentan una estructura anidada estable, aunque sus causas permanecen sin evaluar. Se elaboró una matriz de datos de presencia y ausencia a partir de un inventario de aves obtenido en 43 cayos de este archipiélago. Se calculó el anidamiento mediante el índice NODF basado en el relleno superpuesto y decreciente. La significación del anidamiento se evaluó mediante 1.000 iteraciones de cuatro modelos nulos. Las columnas de la matriz se reordenaron para evaluar siete fac� tores que podrían estar relacionados con el anidamiento en las comunidades de aves. Los ensamblajes de aves presentaron un modelo de anidamiento significativo (67,93) y todos los factores contribuyeron (p < 0,01) a los modelos de anidamiento de las comunidades de aves. La diversidad de hábitats y el área y el perímetro de los cayos fueron los factores que más contribuyeron a la estructura anidada. El modelo de anidamiento de los ensamblajes de aves en los Jardines de la Reina podría estar causado por la interacción de la extinción selectiva y, en menor medida, por la colonización diferencial de especies. Palabras clave: Estructura anidada, Biota fragmentada, Extinción selectiva, Colonización diferencial Received: 23 I 15; Conditional acceptance: 25 V 15; Final acceptance: 18 VI 16 Antonio García–Quintas & Alain Parada Isada, Centro de Investigaciones de Ecosistemas Costeros (CIEC), Cayo Coco, Ciego de Ávila, 69400 Cuba. Corresponding author: Antonio García Quintas. E–mail: antonio@ciec.cu ISSN: 1578–665 X eISSN: 2014–928 X

Introduction Nestedness is a characteristic pattern in insular bio� tas (Rodríguez–Gironés & Santamaría, 2006; Moore & Swihart, 2007) although it may be uncommon in some oceanic islands (e.g., Florencio et al., 2015; Matthews et al., 2015). This model structure can be observed in fragmented scenarios, where the species of depauperated fragments constitute 'proper' subsets of those in richer fragments (Patterson & Atmar, 1986; Patterson, 1987). Nestedness could be generated by one or many factors, depending on the taxonomic entity and the biogeographical features of the study site in question. Wright et al. (1998) noted that any factor that favors or affects the assembly of species communities from a common pool, in a consistent order, may produce nested patterns. Elucidating the factor(s) promoting nested distributions can be a complicated process due to the confluence of many environmental variables, such as 'nuisance factors' (Méndez, 2004; González–Oreja et al., 2012). Two mechanisms that play an essential role in the unfolding of nested structures in natural assemblages of species are selective extinction (species loss within fragments in a predictable sequence based on their adaptability) and species' differential colonization (occupancy of the fragments by species based on their dispersal capabilities) (Patterson & Atmar, 1986; González–Oreja et al., 2012). Selective extinction gen� erates non–random losses because species requiring large minimum areas or forming small populations tend to be more prone to extinction events. In the differential colonization approach, stronger dispersers will take up more fragments than the rest, and this produces ordered differences of species between fragments. Other factors promoting nestedness are the area and isolation among the fragments (Feeley, 2003), disturbances (Bloch et al., 2007), interspecific com� petition, habitat fragmentation and quality, behavioral responses and species´ environmental tolerance, and landscape continuity (Méndez, 2004; González–Oreja et al., 2012). The effect of area upon nestedness rests on the fact that species requiring large home ranges/territories will occupy only large fragments, while species �� with more phenotypic flexibility re� garding habitat exploitation strategies will occur in every fragment (Wright et al., 1998; Feeley, 2003). Isolation is related to the distance effect, in which �� spe� cies with greater dispersal capabilities can be found in all fragments (including the most isolated ones), while more sedentary species could occupy the frag� �� ments closest to the dispersion source (Wright et al., 1998; Longo–Sánchez & Blanco, 2009). The shape of fragments is another potential factor contributing to nestedness because islands with more complex shapes may exhibit greater topographic complexity and habitat heterogeneity, enabling the coexistence of a higher number of species (Hu et al., 2011). Bloch et al. (2007) stated that intense disturbances would alter nestedness patterns at small spatial scales because these phenomena can lead to local extinc� tion. Conversely, periodic disturbance can facilitate the coexistence of species that, in absence of disturbance,

García–Quintas & Parada Isada

might be mutually exclusive. In general, a disturbance can destroy nested structure if rare species are elimi� nated or if local extinctions are density–independent (Bloch et al., 2007). Nevertheless, González–Oreja et al. (2012) based their assertion on the differences of species´ tolerance to consider disturbance as a promoter of nestedness. The influence of interspecific competition on nested structures may be ambiguous due to the existence of divergent criteria about this factor (Méndez, 2004). Bloch et al. (2007) consider that competitive exclusion reduces nestedness by preventing the co–occurrence of species that could otherwise share the same ecological niches. However, competition also affects nestedness by shifting the species composition or the checkerboard structure of the assemblages (Feeley, 2003; Almeida–Neto et al., 2007). Nevertheless, McLain & Pratt (1999) considered that competition along with habitat heterogeneity can reinforce nested� ness because competitive exclusion becomes stronger as the fragments’ habitat complexity decreases. Hierarchical habitat distributions (habitat nested� ness) also play an important role in species nestedness (Higgins et al., 2006; Ulrich et al., 2009; Watson et al., 2009). According to this factor, species nestedness is a direct outcome of non–random distribution of habitats. Despite the different ways that nestedness models can be generated, Ulrich et al. (2009) state that all underlying factors are usually defined by environmental or biological gradients leading to orderly changes of colonization and extinction events in fragmented areas. Nestedness is a typical feature of a wide variety of insular biotas that include plants, arthropods, reptiles, birds and mammals (Calmé & Desrochers, 1999; Al� meida–Neto et al., 2007). It characterizes the structure of meta–communities and describes species´ spatial distributions in less discrete habitats (Bloch et al., 2007; Moore & Swihart, 2007). Nestedness has been used to estimate minimum viable population sizes, to evaluate fragment connectivity, to characterize the resilience of disturbed communities (Bloch et al., 2007), and to predict species´ extinction rates (Azeria & Kolasa, 2008). It has also been widely used in decision–making in conflicting conservation scenarios to help determine whether the protection of small fragments should be prioritized over larger areas (Bloch et al., 2007). In the Jardines de la Reina archipelago (JRA), south of Cuba, bird assemblages exhibit a stable nestedness pattern which becomes more remarkable during the spring season (García–Quintas & Parada, 2014). Owing to the relatively little geographic isolation (for birds) and the poor landscape complexity (with low habitat diversity) of the JRA, we proposed that factors related to physical characteristics of cays (e.g., area, shape) may play a crucial role in bird nestedness unfolding and its stability over time. We would thus expect that the effect of these factors on nestedness would be greater than any other factor studied herein. The objective of this work was to assess the effects of possible driving factors on the nested structures of avian assemblages in the JRA. Such data may be relevant to detect the ecological components that sta� bilize the nested pattern of birds in this insular region.

Animal Biodiversity and Conservation 40.1 (2017)

Sancti Spiritus

Cuba

Ciego de Ávila Camagüey

80 km

Jardines de la Reina archipelago

1:190,000 4

16 km

78º 50' 0'' W

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Fig. 1. Geographic location of the three main insular sub–groups of the Jardines de la Reina archipelago, Cuba: A. Cays of Ana María; B. Central cays of the gulf of Ana Maria; and C. Cays of Doce Leguas. Fig. 1. Ubicación geográfica de los tres subgrupos insulares principales del archipiélago de los Jardines de la Reina, Cuba: A. Cayos de Ana María; B. Cayos del centro del golfo de Ana María; C. Cayos de las Doce Leguas.

Material and methods Study area The JRA stretches along the southern coast of the Isle of Cuba and is made up of numerous cays, flats, sandbanks and coral reefs. The present study en� compasses 43 cays from the three main sub–groups: cays of Ana María, cays of the central part of the Gulf of Ana María, and cays of Doce Leguas (fig. 1). In general, the study site is characterized by fragile ecosystems with low species richness, and the most representative vegetation type is mangrove forests. Xeromorphic scrub and complexes of sandy and rocky shoreline vegetation can also be found. Zúñiga (2000) noted that the origin of the cays of Doce Leguas experienced a gradual growth along the E–W axis during the Holocene through various geological and climatic processes. This is evident

in the different geological development of the cays (more complex cays eastward), and the diversity and structural complexity of its vegetation. The emergence of these cays acted as a barrier, restricting water exchange between the Gulf of Ana María and the Caribbean Sea. This favored the accumulation of muddy sediments in the gulf, which in turn led to the origin of many other cays (e.g., cays of Ana María and the central keys of the Gulf of Ana Maria) (Zúñiga, 2000). Those cays and islets are characterized by an oceanic origin with a recent geological history. Sampling effort Current knowledge of species occurrence in the study site is spatially biased towards the larger cays with higher vegetation diversity. Most surveys to date have been conducted in such cays and more comprehen� sive sets of census techniques have been applied to

describe their avifauna. Differences in the inventories between cays have thus been minimized because the sampling effort was in accordance with the cays’ characteristics. To test the effects of the sampling effort, we developed an indicator of sampling effort resulting from multiplying the number of surveys by the number of census techniques used in each cay, divided by the cay´s area. High values of this indicator correspond to the best sampling efforts conducted. Nevertheless, due �� to the standardization of the sam� pling intensity with respect to the cays’ area, we did not consider large differences in sampling efforts in our study (table 1s in supplementary material). Nestedness analysis We obtained the presence–absence data of bird spe� cies occurring at each surveyed cay from a recent work by García–Quintas & Parada (2014). This study comprised an up–to–date list of the published papers on species inventories carried out in the study area until 2012. We excluded species relying strongly on marine–oceanic habitats, as well as those with no explicit reference to their locality name when first re� ported. These steps aimed to increase the reliability of checklists and to avoid biases during the nestedness analysis. García–Quintas & Parada (2014) classified the migratory status of bird species at the JRA ac� cording to Garrido & Kirkconnell’s (2011) criteria and based on their own knowledge on the local avifauna as follows: permanent resident, winter resident, summer resident or transient. Transients were excluded from the analyses because they �� usually occur in low num� bers and exploit resources during limited timeframes while migrating through the area. We thus expected that these species could exert only a negligible influ� ence on the communities' structure and functions. We created a presence–absence (1–0) matrix who� se rows and columns represented species and cays, respectively. Entries were arranged in the matrix in increasing order, starting from the number of cays oc� cupied per species and the number of species per cay in the rows and columns, respectively. We calculated the degree of matrix nestedness using the nestedness metric based on the overlap and decreasing fill (NODF) (Almeida–Neto et al., 2008) by running the software ANINHADO 3.0.3 (Guimarães & Guimarães, 2006). NODF values ranged from 0 (no nestedness) to 100 (perfect nestedness) and this score was compared with those obtained from the simulations (1,000) of four null models (Er, Ce, Co and Li, Bascompte et al., 2003; Almeida–Neto et al., 2007) in order to estimate the significance level of the degree of overall nestedness. The null models followed these rules: Er. Presences are randomly assigned to any cell within the matrix (equal probability of presence of each species in each cay); Ce. The probability of a cell aij showing a presence is the average of the probabilities of occupancy of its row and column (probability of presence/occupancy is proportional to the degree of generalization in the ecological characteristics of the species and cays); Co. Presences are randomly assigned within the columns (species frequencies fixed with equiprobable site fre�

García–Quintas & Parada Isada

quencies); Li. Presences are randomly assigned within the rows (site frequencies fixed with equiprobable frequencies of species). Underlying factors Seven factors were used to assess their potential effects on the avian nested structures in the JRA. These factors included cay area, perimeter and shape, minimum distance between cays and from the Isle of Cuba (isolation measures), avian habitat diversity in the cays, and habitat nestedness. We measured the perimeter and area of each island using an unsupervised classification (k–means method, change threshold set at 5%, and three iterations) of a mosaic made up by two Landsat 5 TM satellite images, sceneries 13–45 and 13–46, from 11 July and 22 April 2001, respectively. We used ENVI 4.7 software to per� form the classification, which yielded seven classes. These were then grouped to help delimit deep waters from the remaining cover types (land, terrestrial vegeta� tion and shallow waters). Thus, a mask file was created to discard information on deep waters, by assigning it a zero value, and to delineate the emerged land of cays and other critical habitats for waterbirds, such as sha� llow waters. The resulting image (once classified and filtered through the mask) was converted from raster to vector format to calculate the area and perimeter of the 43 cays. For the cays formed by many smaller fragments, the values of area and perimeter of each fragment were added up to obtain the total score of these two variables. We used the shape index for islands (Hu et al., 2011) to characterize the shape complexity of the chosen cays. In this index, value one means circle–shaped fragment, and the score increases as shape tends to differ from the perfect circular symmetry. The isolation degree was calculated by measuring the minimum dis� tance between each cay and the isle of Cuba (potential source–sink system) and the distance between each cay and its nearest neighbor (potential species flux). Distances were measured from the classified image by using ENVI 4.7 software. Seven habitat classes of critical importance to the avifauna were identified: sandy and rocky coasts, close inland lagoons, open inland lagoons, mangroves, sandy vegetation, rocky vegetation and secondary vegetation. The number of habitats per cay (habitat diversity) was recorded from field observations, information provided by specialists and available cartographic materials. Furthermore, these bird habitat classes were used to evaluate the existence of habitat nestedness through a nested analysis in a new matrix where the rows represented habitat classes instead of species (Wang et al., 2013). The last analysis was performed by using the NODF index and the same null models (Er, Ce, Co and Li). We followed Lomolino´s (1996) method to assess the influence of each factor on nestedness, although we used the NODF index to quantify the degree of nestedness. This method is regarded as one of the most suitable and broadly applied to infer the rela� ted factors of nestedness (Fernández–Juricic, 2000; Ulrich et al., 2009; Valencia–Pacheco et al., 2011).

Animal Biodiversity and Conservation 40.1 (2017)

Essentially, it is a way to evaluate the possible drivers of nestedness where it is assumed that immigration probability decreases whereas isolation increases, and the extinction probability decreases insofar as the area increases. The main advantage of this method lies in its flexibility to assess the effect of each factor from the principle of fragment rearrangement within the data matrix (Fernández–Juricic, 2000; Méndez, 2004). It also allows, though indirectly, the linking of the nestedness patterns with the processes of selec� tive extinction (e.g., matrix with fragments ordered by area) and differential colonization (e.g., matrix with fragments ordered by isolation). Thus, seven presence–absence matrices (one per factor) were assembled in the same way that the matrix was built for the general analysis of the nestedness, but columns (cays) within the matrices were rearranged according to the factor to be eva� luated. Ordination criteria of the matrices´ columns were arranged in a decreasing order from left to right for area, perimeter, shape and habitat diversity. Cay–to–cay and cay–isle of Cuba distances were arranged in descending order. If analysis of habitat nestedness was positive, then ordination of the co� lumns to evaluate this factor would be equal to the matrix of habitat nestedness. We evaluated the underlying factors of nestedness in the seven matrices through the NODF index and their level of significance was estimated by running 1,000 simulations of the null model Li per matrix. The selection of this null model is based on its unique capability to randomize presence within the matrices, but to keep the initial order of the columns unaltered, according to the Lomolino (1996) approach. The remainder of the null models (Er, Co and Ce) can not be used for this test because their randomiza� tion algorithms provoke changes in the order of the columns, limiting the evaluation of the nestedness factors. Factors whose matrices yielded significant scores on the NODF index were considered as factors related to the nestedness of the bird assemblages, and the index value also indicated how important the influence of each factor was on this pattern. Factors assessed were related, although indirectly, to selective extinction (area, perimeter, shape, bird habitat diversity and habitat nestedness) and species differential colonization (isolation). To calculate the descriptive statistics of the scores from null models, we used the software Statistica 8.0 (StatSoft, 2007) and evaluated statistical significance at three p values: < 0.1, < 0.05 and < 0.01.

in the central part of the Gulf of Ana María; Palomo and Santa María cays showed distances over 10 km. The westernmost cays of the Doce Leguas group (Bretón, Alcatracito, Alcatraz and Cinco Balas) were the farthest from Cuba´s southern coast (over 80 km offshore). Algodón Grande, Cachiboca, Caguama and Anclitas were the richest cays in terms of bird habitat diversity (table 1); with the exception of Anclitas, all these cays presented secondary vegetation, a fairly uncommon feature in the landscape of the JRA.

Results

Discussion

Main physical characteristics of the cays

Bird assemblages of JRA were nested, showing a cohesive and non–random structure at the level of the regional meta–community. Formally, the implications of the nestedness patterns for biological studies focus on the associated factor(s) leading to such structuring (Cutler, 1994). Nestedness of bird assemblages in this archipelago was influenced by several factors, but bird habitat diversity played a fundamental role. Factors

Cays such as Cinco Balas and Grande were among the three localities with the highest scores of area and perimeter. Cayuelo, Obispito and Quitasol showed the lowest values and have an almost circular sha� pe (table 1). Highest isolation levels with respect to the nearest cays were reported in 75% of the cays

Nestedness of the avifauna of the JRA A total of 77 bird species were found in the sampled cays (transients excluded), but Fregata magnificens (Magnificent Frigatebird) was discarded prior to the nestedness analyses due to its strong dependen� ce on marine resources rather than on terrestrial ecosystems (table 2s in supplementary material). Bird assemblages in the JRA showed a significantly nested structure (NODF = 67.93; p < 0.01) with respect to those randomly generated by the four null models. The NODF simulated scores were 39.19 ± 1.25 [35.14–43.48], 43.41 ± 0.86 [40.57–46.04], 29.87 ± 0.96 [26.94– 32.68] and 40.79 ± 0.53 [39.18–42.60] for Ce, Co, Er and Li, respectively. Underlying nestedness factors Overall distribution of bird habitats among cays was significantly nested (NODFobs = 69.67; p < 0.01) according to the scores (mean ± SD [min–max]) of three null models (NODFCe = 60.71 ± 3.46 [50.10– 70.12], NODFCo = 53.60 ± 1.32 [50.48–60.36], NO� DFEr = 54.59 ± 3.34 [44.79–65.86]). Only Li showed low probabilities (for the three levels of p significance) of occurrence of this pattern (NODFLi = 65.54 ± 3.77 [49.26–75.63]; p = 0.13). We found that all factors contributed to the nested� ness of bird assemblages in the JRA (table 2). Hie� rarchically, bird habitat diversity reached the highest value of the NODF index, which means that it was the main driving factor of nested structures and the existence of strong habitat–bird associations in the study site. The latter statement was further supported by the habitat nestedness' contribution. The area, perimeter and shape of the cays were also important factors related with the observed nestedness degree of bird assemblages (table 2). However, the isolation of the cays (minimum distances) also contributed, although to a lesser extent than other factors.

García–Quintas & Parada Isada

Table 1. General features of 43 cays of the Jardines de la Reina archipelago, Cuba. The shape index (SI) describes the contour of the cays based on a perfect circular shape (SI = 1): C–C. Cay–to–cay; C–I. Cay–to–isle of Cuba; NBh. Number of bird habitats; NBsp. Number of bird species. Tabla 1. Características generales de 43 cayos del archipiélago de los Jardines de la Reina, Cuba. El índice de forma (SI) describe el contorno de los cayos a partir de la forma de un círculo perfecto (SI = 1). (Para consultar las abreviaturas, véase arriba.) Cay

Area (km2)

Perimeter (km)

Cayuelo Obispito Quitasol La Loma Obispo Guinea Cargado La Tea Bergantines Caoba Palomo Santa María Boca Rica Largo Juan Grin Algodoncito Camposanto Flamenco Cana Arenas Tío Joaquín Providencia Alcatracito Boca de la Piedra de Piloto Piedra Grande Guásimas Balandras Boca Seca Alcatraz Manuel Gómez Punta de Los Machos Cuervo Cachiboca Boca Piedra Chiquita Algodón Grande Las Cruces Cabeza del Este Bretón Caguama Anclitas Grande Caballones Cinco Balas

0.02 0.03 0.05 0.06 0.09 0.13 0.15 0.17 0.22 0.26 0.28 0.29 0.36 0.48 0.63 0.77 0.82 0.84 0.91 0.97 1.21 1.29 1.34 1.52 1.53 1.59 1.62 1.76 1.84 2.11 2.14 2.16 2.44 2.88 3.64 3.64 6.82 7.51 7.66 9.06 24.29 33.52 43.56

0.72 0.84 1.02 1.50 2.76 1.92 3.48 2.46 4.20 3.84 6.84 3.12 6.96 7.32 16.74 5.82 6.42 7.50 11.34 9.84 11.07 13.38 11.04 19.14 16.51 8.70 15.36 30.24 16.38 34.80 26.61 35.28 57.00 11.28 32.70 55.37 94.44 71.46 87.42 158.64 193.17 73.68 151.20

SI 1.44 1.38 1.34 1.72 2.66 1.50 2.50 1.71 2.54 2.14 3.63 1.63 3.27 2.98 5.93 1.88 2.00 2.31 3.35 2.81 2.84 3.32 2.69 8.38 3.76 1.95 3.41 6.44 3.40 6.76 5.13 6.78 10.30 1.88 4.84 8.19 10.20 7.36 8.92 14.87 11.06 3.59 6.46

Minimum distance (km) C–C C–I 0.52 1.08 0.48 0.55 1.08 1.55 8.21 1.06 7.14 1.52 14.72 10.85 0.71 5.20 0.05 3.66 0.12 0.29 0.20 1.18 0.00 1.52 0.48 0.64 0.09 1.00 0.54 0.05 0.08 3.66 0.00 8.21 1.83 0.42 5.42 0.09 0.07 0.77 0.68 0.42 0.00 0.00 0.08

11.50 6.21 11.95 17.64 4.32 7.30 46.16 8.00 35.66 11.85 40.20 13.17 37.75 37.88 39.25 28.07 38.50 11.86 18.69 20.05 13.64 8.86 85.64 45.05 46.28 12.63 12.36 37.49 81.93 33.03 14.88 46.85 42.37 46.47 21.10 43.42 31.44 91.00 34.38 46.65 58.23 52.65 81.43

NBh NBsp 1 4 2 2 4 4 4 2 4 3 4 4 2 3 3 4 4 4 5 4 5 5 4 4 5 4 2 2 4 4 3 5 6 5 6 4 5 4 6 6 5 5 3

3 8 5 6 14 11 20 8 18 17 23 22 12 5 16 29 11 35 16 18 18 5 19 25 20 6 3 25 14 11 9 27 26 24 50 29 34 37 60 68 61 47 27

Animal Biodiversity and Conservation 40.1 (2017)

Table 2. Evaluation of the effects of seven factors potentially related to the nestedness of bird assemblages in 43 cays of the Jardines de la Reina archipelago, Cuba. The simulated scores (N = 1,000) of the null model Li are shown as mean ± SD (min–max): NODF. Nestedness metric based on overlap and decreasing fill; P. Probability. Tabla 2. Evaluación del efecto de siete factores que podrían estar relacionados con el anidamiento de los ensamblajes de aves en 43 cayos del archipiélago de los Jardines de la Reina, Cuba. Los valores simulados (N = 1.000) del modelo nulo Li se muestran como media ± DE (mín–máx): NODF. Índice de anidamiento basado en el relleno superpuesto y decreciente; P. Probabilidad. Factor

NODFobs

NODFLi

Area

64.90

35.52 ± 0.73 (33.51 - 37.88)

< 0.01

Perimeter

64.82

35.53 ± 0.73 (33.57 - 38.21)

< 0.01

Shape

64.26

35.58 ± 0.76 (32.94 - 38.39)

< 0.01

Minimum distance (cay–to–cay)

61.42

35.51 ± 0.73 (33.34 - 38.05)

< 0.01

Minimum distance (cay–isle of Cuba)

57.16

35.56 ± 0.72 (33.41 - 37.93)

< 0.01

Bird habitat diversity

65.85

35.52 ± 0.70 (33.09 - 37.91)

< 0.01

Bird habitat nestedness

64.11

35.52 ± 0.74 (33.27 - 38.34)

< 0.01

associated with the physical characteristics of the cays (area, perimeter and shape) also contributed to the nested structure in a significant way. We indirectly inferred that selective extinction could be the princi� pal historical mechanism that triggers and stabilizes the observed nested patterns, without excluding the contribution of differential colonization events. Diversity of avian habitats was the most influential ecological factor regarding nestedness in the JRA. Although this factor may not overshadow the effects of other factors, it could diminish their effects. In this regard, Calmé & Desrochers (2000) consider that if species richness is correlated with habitat diversity, the area per se constitutes a secondary factor. The latter could be related to marked preferences of cer� tain groups of species over some specific habitats (Calmé & Desrochers, 1999). While shorebirds, gulls, herons and other waterbirds exploited common habi� tats (e.g., mangroves, coasts, lagoons) in most cays, several species of warblers, cuckoos and thrushes were restricted to sandy vegetation, a bird habitat present in fewer cays. Seemingly, the habitat factor plays a key role in the nestedness patterns of the avian communities inhabiting the JRA. This may well reinforce the intricate species–habitat relationships even further as a pivotal ecological factor determining species distribution patterns, mainly because birds are a highly–mobile group with great dispersal capability. The contributions of habitat nestedness reinforce the effect of habitat diversity, but represent a superior level of organization. The relationship between this pattern and the birds' nested structures reflects the strong dependence of bird species upon their habitats, since such habitats constitute their sources of forag� ing, reproduction and refuge. Habitat nestedness is

thought to be among those processes that explain the nested structures found in bird communities that have eluded much criticism since it rests mostly on the links between birds and their habitats, and dis� regards species' population dynamics and natural history (Calmé & Desrochers, 1999; Wang et al., 2013). Thereby, species exploiting common habitats should be widespread whereas species depending on uncommon habitats should be confined to a few sites (Wright et al., 1998). Hu et al. (2011) summarize that area, isolation and shape of the fragments are among the main factors shaping species richness patterns and meta� community assemblages. Shape may be related to the physical complexity of fragments, and thus to their potential capacity for supporting more or fewer numbers of species. Perimeter might also influence the degree of nestedness of the fragments (e.g., cays made up by several fragments) or the habitat avail� ability in transition zones. For instance, the water/land interface offers habitats and food webs exploited by terrestrial, marine and those organisms confined to this ecotone (Pizarro et al., 2012), which in turn might generate nested patterns. The strong species–area relationships, long consid� ered a cornerstone of the MacArthur & Wilson (1967) Theory of Biogeography of Isles, can reflect features of habitat spatial distribution, growth dynamics and popu� lation extinction as well as the dispersion and habitat selection statistics (Coleman et al., 1982). In the JRA, the area of the cays was one of the factors generating nestedness of bird communities, as illustrated by the finding that larger cays (e.g., Grande, Caguama) sup� ported higher species richness than smaller ones, and bird richness in small cays is a subset of the big cays.

This coincided with the results of Ambuel & Temple (1983) and Fernández–Juricic (2000), who focused on the avian assemblages associated with the fragmented deciduous forests of eastern North America and the urban parks of Madrid, respectively. Isolation metrics (cay–to–cay and cay–isle of Cuba) contributed less than other factors to the avian com� munities' nestedness detected in the JRA. Of these metrics, minimum distance between cays was the most important factor, probably due to the different degrees of isolation between the three insular sub� groups. In this regard, the central cays of the gulf of Ana María were the most isolated ones in the study site, whereas many cays of Doce Leguas are closer to one another. These isolation differences along with the varying dispersal capability shown by bird species may favor the development of the nestedness structures in the JRA's avian assemblages. Longo–Sánchez & Blanco (2009) mentioned that the effect of distance or isolation could be accounted for by the geographical isolation and the species dispersal skills. The curved shape of the southern coast of central Cuba, along with the relative position of the cays in the study site (fig. 1), clearly illustrates their proximity to the mainland. This condition may favor the flux of species between these two areas, especially for birds, which in turn, would account for the minor contribution (among the factors evaluated) of the distance among cays and between the cays and the Isle of Cuba to the development of nested structures. The species dispersal movements may also be favored by the low isolation of the cays of each insular subgroup and because birds are among the vertebrates with the greatest dispersal skills over the water (Cook & Quinn, 1995). Lees & Peres (2006) assert that the distance and isolation metrics become relevant at predicting species richness when habitat fragments are scattered in ranges from 100 to 10,000 m. Most of the minimum distances reported in the JRA fall within this range. Moreover, Higgins et al. (2006) considered that population dynamics of the birds inhabiting the Greater Antilles are largely determined by natality and mortality processes rather than by the species migra� tory behavior, given the long distances separating the islands. This view should not be generalized, however, because an important assortment of the Cuban avi� fauna is made up of migratory species coming from various regions of the American continent. Area, perimeter and shape of the cays in the JRA were the factors related to the selective extinction of species. This finding relies on the potential effects of these mechanisms regarding the capability of the cays for harboring bird species as well as on the species' adaptability. The isolation metrics were indirectly cor� related with the species' dispersal capacity, a key ele� ment for analyzing the differential colonization process. Differences in the diversity of bird habitat among the sampled cays favored extinction over the colonization process, as the degree of isolation does not appear to pose an effective barrier for preventing most of the species from wandering across this insular region. Nevertheless, the isolation of the cays could limit the distribution of some species of birds with rather

García–Quintas & Parada Isada

low dispersion in the JRA. It is of interest that endemic taxa such as Xiphidiopicus percussus, with common and broadly distributed year–round populations in many cays of Doce Leguas, are not known to oc� cur in the central and northern cays of the Gulf of Ana María, although they are closer to the southern coast of central Cuba, where the species’ preferred habitats are highly represented (mangrove forests) (Parada & Garcia–Quintas, 2012). Further evidence regarding the limited connectivity between the JRA and the mainland bird populations comes from the possible relictual populations of Quiscalus niger, with the subspecies Q. n. caribaeus persisting in the northern and southern archipelagos, including the Isle of Pines and the westernmost region of the mainland, having apparently been replaced elsewhere on the mainland by Q. n. gundlachi (Buden & Olson, 1989). Patterson & Atmar (1986), Wright et al. (1998) and Feeley (2003) state that selective extinction is, in natural archipelagos, a more frequent phe� nomenon rendering higher nestedness scores over the species' differential colonization. However, the presumed prevalence of extinction over colonization as a mechanism that promotes nested structures is not generalized. In the JRA, as in the study of Va� lencia–Pacheco et al. (2011), both colonization and extinction played essential roles in the development of nestedness patterns, coinciding with Murgui (2010), who considers these two processes as not mutually exclusive. Nonetheless, the effects of the other fac� tors suggest that species' selective extinction could contribute to the avian assemblages´ nestedness in the JRA. This insular region possesses peculiar fea� tures such as an oceanic origin and location not far from larger landmasses (e.g., the Isle of Cuba, the North American continent). Such location may have facilitated connectivity between the avifauna of the JRA and the neighboring emerged lands as the flux of species should have not been highly restricted. Despite the differences in the sampling effort across the study site, the probable absence of true nestedness patterns was ruled out. As shown by the sampling effort indicator, survey intensity corresponded to the cays' area. Furthermore, cays such as Caguama, Anclitas, Grande and Algodon Grande have been surveyed more frequently with the aid of several census techniques, as it was long assumed that they may harbour higher levels of avian biodiversity. The least surveyed cays are generally characterized by more homogeneous vegetation and fewer habitat types. For example, Cayuelo, Quitasol and La Tea are small cays that have a low diversity of bird habitat and predominance of mangrove forests. In these cays, bird species can be detected in few sampling sections. However, it will be important to improve the sampling design to reduce or eliminate the associated biases and use other metrics to quantify the sampling intensity. We believe that the nested structures detected in the JRA avifauna may have initially been generated by differential colonization (due to the oceanic origin of the JRA), and later reinforced by the selective ex� tinction of species. Such extinction events could have taken place through demographic processes such as

Animal Biodiversity and Conservation 40.1 (2017)

mortality and emigration, as well as by displacement of certain species via competitive exclusion. Differential colonization may contribute to nestedness stability but does not seem determinant in this insular region, as reflected by the lack of influence of the annual migrations on the degree of assemblages' nested� ness, as shown by García–Quintas & Parada Isada (2014). Therefore, cays with greater habitat diversity, larger area and higher bird species richness will act as source patches within the JRA and thus preserve the nestedness of avian assemblages. In this case, Anclitas, Grande, Caguama, Caballones and Algodón Grande cays are the most important sites for avian conservation. The latter is not currently included within the Cuban system of protected areas and its future inclusion might well bes the next step towards the efficient design of regional reserve networks in which the avifauna´s functional connectivity and nestedness are pivotal theoretical frameworks. Acknowledgements We thank Vicente Osmel Rodríguez Cárdenas for his valuable comments and corrections, which greatly im� proved the work, and the researchers and technicians at CIEC who contributed to this research. References Almeida–Neto, M., Guimarães, P., Guimarães, P. R. Jr., Loyola, R. D. & Ulrich, W., 2008. A �� consis� tent metric for nestedness analysis in ecological systems: reconciling concept and measurement. Oikos, 117: 1227–1239. Almeida–Neto, M., Guimarães, P. R. Jr. & Lewinsohn, T. M., 2007. On nestedness analyses: rethinking matrix temperatura and anti–nestedness. Oikos, 116: 716–722. Ambuel, B. & Temple, S. A., 1983. Area–dependent changes in the bird communities and vegeta� tion of southern Wisconsin forests. Ecology, 64: 1057–1068. Azeria, E. T. & Kolasa, J., 2008. Nestedness, niche metrics and temporal dynamics of a metacommu� nity in a dynamic natural model system. Oikos, 117: 1006–1019. Bascompte, J., Jordano, P., Melián, C. J. & Olesen, J. M., 2003. The nested assembly of plant–animal mutualistic networks. Proceedings of the National Academy of Sciences, USA, 100: 9383–9387. Bloch, C. P., Higgins, C. L. & Willing, M. R., 2007. Effects of large–scale disturbance on metacom� munity structure of terrestrial gastropods: temporal trends in nestedness. Oikos, 116: 395–406. Buden, D. W. & Olson, S. L., 1989. The avifauna of the cayerias of southern Cuba, with the ornithological results of the Paul Bartsch expedition of 1930. Smithsonian Contributions to Zoology, 477: 1–34. Calmé, S. & Desrochers, A., 1999. Nested bird and micro–habitat assemblages in a peatland archipe� lago. Oecologia, 118: 361–370.

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Horn Biosphere Reserve (54–55˚S), Chile. Polar Biology, 35: 39–51. Rodríguez–Gironés, M. A. & Santamaría, L., 2006. A new algorithm to calculate the nestedness tempe� rature of presence–absence matrices. Journal of Biogeography, 33: 924–935. StatSoft Inc., 2007. STATISTICA (data analysis software system), version 8.0. Statsoft, Inc., Tulsa, OK. Ulrich, W., Almeida–Neto, M. & Gotelli, N. J., 2009. A consumer’s guide to nestedness analysis. Oikos, 118: 3–17. Valencia–Pacheco, E., Avaria–Llautureo, J., Muñoz– Escobar, C., Boric–Bargetto, D. & Hernández, C. E., 2011. Patrones de distribución geográfica de la riqueza de especies de roedores de la tribu Oryzomyini (Rodentia: Sigmodontinae) en Suda� mérica: Evaluando la importancia de los procesos de colonización y extinción. Revista Chilena de Historia Natural, 84: 365–377. Wang, Y., Ding, P., Chen, S. & Zheng, G., 2013. Nestedness of bird assemblages on urban wo� odlots: Implications for conservation. Landscape and Urban Planning, 111: 59–67. Watson, J. E. M., Watson, A. W. T., Fischer, J., Ingram, J. C. & Whittaker, R. J., 2009. Using nestedness and species–accumulation analyses to streng� then a conservation plan for littoral forest birds in south–eastern Madagascar. International Journal of Biodiversity and Conservation, 1: 67–80. Wright, D. H., Patterson, B. D., Mikkelson, G. M., Cut� ler, A. & Atmar, W., 1998. A comparative analysis of nested subset patterns of species composition. Oecologia, 113: 1–20. Zúñiga, A., 2000. Caracterización básica de la geología de los cayos de la porción centro occidental del subarchipiélago Jardines de la Reina. Cayos Algodón Grande, Anclitas y Caballones. Enlace, 6: 1–5.

Animal Biodiversity and Conservation 40.1 (2017)

Supplementary material

Table 1s. Evaluation of the sampling effort in 43 cays of the Jardines de la Reina archipelago, Cuba, for a study about the nested structure of the bird assemblages in these cays: Ns. Number of surveys; Nct. Number of census techniques; Ise. Indicator of sampling effort. (Values of areas are given in table 1.) Tabla 1s. Evaluación del esfuerzo de muestreo en 43 cayos del archipiélago de los Jardines de la Reina, Cuba, para un estudio sobre la estructura anidada de los ensamblajes de aves de estos cayos: Ns. Número de muestreos; Nct. Número de técnicas de censo; Ise. Indicador del esfuerzo de muestreo. (Los valores de las áreas se presentan en la tabla 1.)

Cays

Nct

Ise

Cinco Balas

0.05

Caballones

Cabeza del Este

Bretón

Punta de Los Machos

Cays

Nct

Ise

Algodoncito

1.30

0.06

Piedra Grande

1.31

0.15

Boca de la Piedra de Piloto 2

1.32

0.27

Juan Grin

1.59

0.47

Largo

2.08

Manuel Gómez

0.47

Anclitas

2.32

Alcatraz

0.54

Caguama

2.35

Las Cruces

0.55

Camposanto

2.44

Balandras

0.62

Caoba

3.85

Guásimas

0.63

Boca Rica

5.56

Boca Piedra Chiquita

0.69

La Tea

5.88

Alcatracito

0.75

Santa María

6.90

Providencia

0.78

Palomo

7.14

Cachiboca

0.82

Guinea

7.69

Algodón Grande

0.82

Bergantines

9.09

Tío Joaquín

0.83

Obispo

11.11

Cuervo

0.93

Cargado

13.33

Grande

0.99

La Loma

16.67

Arenas

1.03

Quitasol

20.00

Cana

1.10

Obispito

33.33

Boca Seca

1.14

Cayuelo

50.00

Flamenco

1.19

García–Quintas & Parada Isada

Table 2s. Incidence matrix of non–transient bird species occurring in 43 cays in the Jardines de la Reina archipelago, Cuba: A. Anclitas; B. Grande; C. Caguama; D. Algodón Grande; E. Caballones; F. Bretón; G. Flamenco; H. Cabeza del Este; I. Las Cruces; J. Algodoncito; K. Cinco Balas; L. Cuervo; M. Cachiboca; N. Boca de la Piedra de Piloto; O. Boca Seca; P. Boca Piedra Chiquita; Q. Palomo; R. Santa María; S. Piedra Grande; T. Cargado; U. Alcatracito; V. Arenas; W. Tío Joaquín; X. Bergantines; Y. Caoba; Z. Juan Grin; a. Cana; b. Alcatraz; c. Obispo; d. Boca Rica; e. Camposanto; f. Guinea; g. Manuel Gómez; h. Punta de Los Machos; i. La Tea; j. Obispito; k. La Loma; l. Guásimas; m. Providencia; n. Quitasol; o. Largo; p. Cayuelo; q. Balandras; * Presence; – Absence.

Cay Species

A B C D E F G H

J K L M N O P Q R S

Setophaga petechia

* *

* * *

* *

* * *

* *

Ardea herodias

* *

* * *

* *

* * *

* *

Quiscalus niger

* *

* * *

* *

* * *

* * * *

– *

Phalacrocorax auritus

* *

* * *

– *

* * *

* *

Charadrius wilsonia

* *

* * –

* *

* * * – * –

Thalasseus maximus

* *

* * –

* *

* * *

* *

Patagioenas leucocephala

* * *

* *

– * – * *

* –

–

Pandion haliaetus

* * *

– *

* * *

* *

– –

Zenaida asiatica

* * *

– *

– * – * *

–

Pelecanus occidentalis

* *

* * *

– *

* * *

* *

– –

Tyrannus dominicensis

* * *

* *

– * – * *

* –

– –

Tyrannus caudifasciatus

* *

– * *

* *

* – – * * –

– –

Setophaga discolor

* *

* * *

* *

–

* – *

–

Butorides virescens

* * *

* *

–

* * * – *

– *

–

Vireo altiloquus

* * *

* –

– * *

* – *

– –

–

* *

* – *

Ardea alba

* * *

* *

– * – * *

* –

–

Egretta rufescens

* *

* * *

– –

* * *

* *

– –

Arenaria interpres

* *

* * – – – –

– * * – *

* *

Contopus caribaeus

* *

* * *

* – – – * –

– –

* *

Parkesia noveboracensis

* *

* * *

* *

–

* – * – – –

–

Zenaida macroura

–

– *

* – –

* *

– * – – *

– –

–

Cathartes aura

* *

* * *

* *

– * – * – *

– –

Chordeiles gundlachii

* *

* – –

* –

– – – * * –

* –

–

Chlorostilbon ricordii

* *

* * *

* –

* * – – *

Egretta tricolor

* *

* * *

* *

– * *

Setophaga ruticilla

* * *

* *

–

* – * – * –

Thalasseus sandvicensis

* *

* – – – *

–

* – * – – –

* –

–

* – *

– –

–

– –

Setophaga palmarum

* *

* * –

* – –

* * * – – –

– *

–

Eudocimus albus

* *

– * –

* –

– – – * *

– –

–

Buteogallus gundlachii

* * *

– *

– * *

* – *

– –

Myiarchus sagrae

* * *

* *

– – – * * –

– –

–

Agelaius humeralis

* *

– * – – *

– – – * *

* –

–

Egretta caerulea

* – *

– *

–

* * * – – –

– –

–

Rallus longirostris

* *

* – *

– *

–

* – * – – –

– *

– –

Leucophaeus atricilla

* *

* – –

* – * – * –

* –

–

Anhinga anhinga

* *

– * *

– *

– * – – – –

* –

– –

Megaceryle alcyon

* *

* * *

* – –

* – * – – –

– *

–

Animal Biodiversity and Conservation 40.1 (2017)

iii

Tabla 2s. Matriz de incidencia de las especies de aves no transeúntes registradas en 43 cayos del archipiélago de los Jardines de la Reina, Cuba: A. Anclitas; B. Grande; C. Caguama; D. Algodón Grande; E. Caballones; F. Bretón; G. Flamenco; H. Cabeza del Este; I. Las Cruces; J. Algodoncito; K. Cinco Balas; L. Cuervo; M. Cachiboca; N. Boca de la Piedra de Piloto; O. Boca Seca; P. Boca Piedra Chiquita; Q. Palomo; R. Santa María; S. Piedra Grande; T. Cargado; U. Alcatracito; V. Arenas; W. Tío Joaquín; X. Bergantines; Y. Caoba; Z. Juan Grin; a. Cana; b. Alcatraz; c. Obispo; d. Boca Rica; e. Camposanto; f. Guinea; g. Manuel Gómez; h. Punta de Los Machos; i. La Tea; j. Obispito; k. La Loma; l. Guásimas; m. Providencia; n. Quitasol; o. Largo; p. Cayuelo; q. Balandras; * Presencia; – Ausencia.

Cay

T U V

W X

Y Z

c d

l m n

o p q

* *

–

– *

–

– –

–

– – –

* *

–

– – –

* –

–

* –

–

– –

–

* –

–

– – –

* –

–

– *

–

– – –

–

– –

– – –

* –

–

– –

–

– – –

–

* –

–

– *

–

– – –

–

– – – – –

–

* –

–

– *

–

– – –

–

* –

–

– – –

– –

–

– –

–

– –

– – –

–

– –

–

– – –

* –

–

– –

–

– –

– – –

–

* *

–

– –

–

– – –

–

– *

–

– – –

– *

–

– –

–

– – –

– –

–

* –

–

– – –

–

* *

–

* –

–

– – –

– –

–

– –

–

– – –

– – *

–

* –

–

– –

– – –

–

* –

–

– – –

–

* *

–

* –

–

– – –

–

* –

–

* *

–

– – –

– – *

–

– –

–

– – –

– –

–

– –

–

– – –

– –

–

– –

–

– – –

– –

–

– –

–

– – –

–

– *

–

– – –

–

* –

–

– –

–

– – –

–

* *

–

– –

–

– – –

–

– *

–

– – –

– – *

–

– –

–

– – –

–

* –

–

– –

–

– – –

– – *

–

* –

–

– – –

–

– *

–

– – –

–

– –

–

– – –

García–Quintas & Parada Isada

Table 2s. (Cont.)

Cay Species

A B C D E F G H

J K L M N O P Q R S

Actitis macularius

* * –

* *

–

– – – – – –

– –

Geothlypis trichas

– – *

* – –

* – * – – –

– –

Platalea ajaja

– – – – –

* – *

* – –

– –

Pluvialis squatarola

* * – – – –

– – – – – –

– –

–

Calidris minutilla

* *

* * – – – –

– * – – – *

– *

– –

Setophaga americana

* *

* * *

* – – – – –

– –

Setophaga caerulescens

* *

* * *

* – –

* – – – – –

– *

– –

Mniotilta varia

– * *

* *

– – – – – –

– *

– * – – *

– – – – – *

* –

– –

* – *

–

– – * – – –

– –

Xiphidiopicus percussus

Setophaga dominica

–

* *

* – –

* *

–

Egretta thula

* *

* * – – –

– – – * – –

– –

Sternula antillarum

* –

* – – – *

–

– – – * – –

– –

Seiurus aurocapilla

* – *

* – –

– – – – – –

– –

Progne cryptoleuca

– * – – – –

– – – – – –

– –

Nyctanassa violacea

– –

* * –

–

– – – – – –

– –

Himantopus mexicanus

– *

* * – – – –

* – – – – –

– –

Tringa semipalmata

* –

* * – – – –

– * – – – –

– –

Coccyzus americanus

* –

– – *

– – –

– – – * – –

– *

– –

* *

–

Petrochelidon fulva

* –

* * – – – –

– * – – – –

– –

Dumetella carolinensis

– – *

* – –

– – – – – –

– –

Charadrius vociferus

–

– *

* – –

– – – – – –

– –

Calidris alba

–

– *

* – – – – –

– –

Hydroprogne caspia

– –

* – – – – –

– – – – – –

– –

–

Crotophaga ani

* –

– * – – – –

– – – – – –

– –

–

Setophaga tigrina

– – – – – –

– –

Columbina passerina

–

– *

* * – – – –

* – – – – –

– –

Charadrius semipalmatus

* *

– – – – – –

– –

Tringa melanoleuca

– *

– – – – – –

– –

Patagioenas squamosa

– –

– – *

– – –

– – – – – –

– –

Turdus plumbeus

* –

– * – – – –

– – – – – –

– –

Mimus polyglottos

– – – – – –

– –

Zenaida aurita

– *

– – – – – –

– –

Vireo griseus

* –

– – – – – –

– –

Helmitheros vermivorum

* –

– – – – – –

– –

Coccyzus minor

– –

– – – – – –

– –

Tiaris olivaceus

– –

– – – – – –

– –

Sula leucogaster

–

– –

– – *

– – –

– – – – – –

– –

Tringa flavipes

–

– –

– – – – – –

– – * – – –

– –

Haematopus palliatus

–

– –

– – – – – –

– –

Animal Biodiversity and Conservation 40.1 (2017)

Tabla 2s. (Cont.)

T U V

W X

Y Z

Cay b

c d

l m n

o p q

– – –

–

– –

–

– – –

–

– –

–

– – –

–

– *

–

– – –

– –

–

– –

–

– – –

– –

–

– –

–

– – –

–

– –

–

– – –

–

– –

–

– –

– – –

–

– –

–

– – –

–

– –

–

– – –

– – *

–

– –

–

– – –

–

– –

–

– – –

–

– –

–

– – –

–

– –

–

– – –

– – *

–

– –

–

– – –

–

– –

–

– – –

–

– –

–

– – –

–

– –

–

– – –

–

– –

–

– – –

–

– –

–

– – –

–

– –

–

– – –

–

* –

–

– –

– – –

– –

–

– –

–

– – –

– –

–

– –

–

– – –

–

– –

–

– – –

–

– –

–

– –

– – –

–

– –

–

– – –

–

– –

–

– – –

– –

–

– –

–

– – –

–

– –

–

– – –

–

– – –

–

– –

–

– – –

–

– – –

–

– –

–

– – –

–

– –

–

– – –

–

– –

–

– – –

–

– –

–

– – –

–

– –

–

– – –

–

– –

–

– – –

–

– –

–

– – –

–

– –

–

– – –

–

– –

–

– – –

– –

110

Morelle et al.

Animal Biodiversity and Conservation 40.1 (2017)

Panel of informative SNP markers for two genetic lines of European bison: Lowland and Lowland–Caucasian M. Wojciechowska, Z. Nowak, A. Gurgul, W. Olech, W. Drobik & T. Szmatoła

Wojciechowska, M., Nowak, Z., Gurgul, A., Olech, W., Drobik, W. & Szmatoła, T., 2017. Panel of infor� mative SNP markers for two genetic lines of European bison: Lowland and Lowland–Caucasian. Animal Biodiversity and Conservation, 40.1: 17–25. Abstract Panel of informative SNP markers for two genetic lines of European bison: Lowland and Lowland–Caucasian.— As the result of a population bottleneck, the present population of European bison shows a high level of inbreeding, and a significant loss of genetic variability. In studies on such specific species there is a need to apply methods that obtain as much information about the genome as possible in a short time. The aim of the study was to define a panel of SNP (single nucleotide polymorphism) markers that could serve in genetic diversity analysis of European bison from two lines: Lowland (LB) and Lowland–Caucasian (LC). The study used 57 individuals from the LB line and 72 from the LC line. To identify well–performing SNPs in European bison, we used two microarrays with different markers densities: BovineSNP50 v2 BeadChip and BovineHD BeadChip. As a result of the adopted criteria, 1,421 and 22,122 markers, respectively, were selected. On the basis of statistical analysis (allele frequencies, Fisher’s exact test, and the Z–test), a panel of 1,536 informative SNP markers was ultimately selected for further study; 26 of these with private alleles for the LB line and 611 with private alleles for the LC line. The data obtained in this study could further enrich and support breeding programs in the context of relatedness between particular specimens and herds from captive breeding centres. Key words: Bison bonasus, European bison, Genetic marker, Microarray, Single nucleotide polymorphism Resumen Conjunto de marcadores de tipo PSN para dos líneas genéticas de bisonte europeo: Lowland y Lowland– Caucasiana.— Debido al cuello de botella demográfico, la población actual de bisonte europeo muestra un elevado grado de endogamia y una pérdida significativa de variabilidad genética. Es necesario que en los estudios realizados con estas especies específicas se apliquen métodos que permitan obtener tanta informa� ción genómica como sea posible en un tiempo breve. El objetivo de este estudio era definir un conjunto de marcadores de tipo PSN (polimorfismo de un solo nucleótido) que pudiera servir para analizar la diversidad genética de las dos líneas de bisontes europeos: Lowland (LB) y Lowland–Caucasiana (LC). En el estudio se analizaron 57 individuos de la línea LB y 72 de la línea LC. Para caracterizar bien el rendimiento de los PSN en el bisonte europeo, se usaron dos micromatrices multigénicas (genochip) con densidades diferentes de marcadores: BovineSNP50 v2 BeadChip y BovineHD BeadChip. Como consecuencia de los criterios adop� tados, se seleccionaron 1.421 y 22.122 marcadores, respectivamente. Sobre la base del análisis estadístico (frecuencias alélicas, prueba exacta de Fisher y prueba Z), en última instancia se seleccionó un conjunto de 1.536 marcadores informativos de PSN para los estudios adicionales, 26 de los cuales tienen alelos privados para LB y 611, para la línea LC. La información obtenida en este estudio podría enriquecer aún más y apoyar a los programas de reproducción en un contexto de parentesco entre especímenes particulares y manadas que viven en cautividad en centros reproductivos. Palabras claves: Bison bonasus, Bisonte europeo, Marcador genético, Micromatriz multigénica (genochip), Polimorfismo de un solo nucleótido

ISSN: 1578–665 X eISSN: 2014–928 X

Wojciechowska et al.

Received: 4 VI 15; Conditional acceptance: 28 X 15; Final acceptance: 18 VII 16 Marlena Wojciechowska, Zuzanna Nowak, Wanda Olech, Wioleta Drobik, Dept. of Genetics and Animal Breeding, Warsaw Univ. of Life Sciences, Ciszewskiego 8, 02–786 Warsaw, Poland.– Artur Gurgul & Tomasz Szmatoła, Lab. of Genomics, National Research Institute of Animal Production, Krakowska 1, 32–083 Balice, Poland. Corresponding author: Marlena Wojciechowska. Email: marlena_wojciechowska@wp.pl

Animal Biodiversity and Conservation 40.1 (2017)

Introduction By the 1920s, Bison bonasus were extinct in the wild. The only remaining European bison were kept in ma� naged enclosures and amounted to just 54 individuals (29 males and 25 females). The current population of the species is derived from 12 founders: 11 indi� viduals of the subspecies Bison bonasus bonasus and the last representative of the subspecies Bison bonasus caucasicus. After successful reintroductions, there are now two genetic lines: Lowland (LB) and Lowland–Caucasian (LC). The Lowland line, also called the Bialowieza line, is derived from seven B. b. bonasus individuals and is a closed line, meaning that only offspring of Lowland European bison may be classified as belonging to it. The Lowland–Cau� casian line includes European bison whose pedigree includes the last and only male representative of B. b caucasicus (Pucek, 1991; Olech, 1999) As the result of past bottlenecks, the present po� pulation (5,249 specimens registered in 2013 in the European Bison Pedigree Book) shows a high level of inbreeding, and a significant loss of genetic variability (Olech, 2010). The European Bison Pedigree Book was created in 1924 and is published to this day. Pedigree data now provide a basis for carrying out breeding, and make it possible, among other things, to estimate the coefficient of inbreeding and kinship of the animals in the European bison breeding centres. However, in such a specific population, the pedigree, though being extremely valuable, cannot constitute the only source of information concerning the genetic value of the animals. In studies on European bison there is a need to apply methods that ensure that as much information about the genome as possible is obtained in a short time. Over the years, a considerable number of studies have been conducted on European bison. The analy� ses included, among others, allozymes (Hartl & Pucek, 1994), blood groups (Sipko et al., 1995), the genes from the group of the MHC (major histocompatibility complex) (Udina & Shaikhaiev, 1998; Łopieńska et al., 2003, 2011) and microsatellites (Gralak et al., 2004; Roth et al., 2006; Nowak & Olech, 2008a). One of the more recent techniques used to estimategenetic variation involves microarrays, an approach used for several years to determine SNP (single nucleotide polymorphism) genotypes in various sites of the ge� nome. This method allows the analysis of hundreds of thousands of markers at the same time, significantly reducing the time required to achieve a huge amount of data (Illumina®). Tokarska et al. (2009) compared effectiveness of 17 microsatellite and 960 SNP mar� kers for paternity and identity analysis in the Lowland line of the European bison. Oleński et al. (2015) used for the first time the BovineHD microarray to find SNP markers associated with posthitis in the same genetic line. The first SNP analysis using the BovineSNP50 microarray which included both genetic lines (five individuals from LB and five individuals from LC) was performed by Kamiński et al. in 2012. The aim of the present study was to identify a panel of SNP markers (among those assayed on the

Illumina BovineSNP50 and BovineHD arrays) that could be used to analyse genetic structure, identify individuals and control the origin and relatedness of the European bison, as well as identify alleles specific to the two genetic lines: Lowland (LB) and Lowland–Caucasian (LC). This is the first study using BovineHD microarray to compare genetic structure of two genetic lines of European bison. The data obtained in this study will further supple� ment and confirm analysis carried out on the basis of pedigree data in the context of relatedness between particular specimens and herds from captive breeding centres, making it possible, among other things, to estimate inbreeding based on multiple sources of infor� mation. Proper management of the breeding program is important for protection of the species against in� creasing inbreeding and its negative impact. Currently, this program is being conducted in both the European bison breeding centres and in the wild. Most of the animals from captive breeding have a known pedigree, which contributes to the control of their origins and aids population management. However, with such a low genetic variability as occurs in European bison, pedigree information may be insufficient. In addition, in the case of animals from free roaming herds, the information on their relatedness is incomplete or im� possible to determine. For this reason, the assignment of SNP markers characteristic for particular genetic lines, as well as the populations within them, is clearly a great advantage, not only in future research but also to enrich and support breeding programs. Material and methods The animals The study used 144 samples of European bison DNA (Bison bonasus), including eight samples that were analysed on two types of microarrays (BovineSNP50 v2 BeadChip–54,609 SNPs and BovineHD Bead� Chip–777,962 SNPs) to control the repeatability of the results and to increase the initial pool of markers. Additional control samples constituted the DNA of do� mestic cattle Bos taurus. A positive genotyping result was obtained from 129 individuals (57 individuals from LB and 72 from LC). The Lowland line included 32 ma� les and 25 females, while the Lowland–Caucasian line included 36 animals of each sex. The biological material was collected from European bison from Polish and other breeding centres, as well as from free roaming herds. To select the most representative samples of European bison species, we were guided by inter alia, the genetic line, parental lines and the participation of ancestors. In addition, to exclude P–C (parent–child) errors and P–P–C (parent–parent–child) errors, the research included related animals —one full family: mother (sample ID K233), father (sample ID K238) and offspring (sample ID K294), as well as eight pairs of father–offspring and 14 pairs of mother– offspring. These animals were selected on the basis of the pedigree book (see table 1s in supplementary material).

Wojciechowska et al.

DNA isolation

Genotyping BovineSNP50 v2 BeadChip and BovineHD BeadChip microarrays (Illumina, Inc. San Diego, CA) were used for genotyping 54K and 777K SNPs. Analyses were performed according to the Infinium Ultra and Infinium Super protocols (Illumina), and the microarrays were scanned using HiScanSQ (Illumina). The resulting intensity reading was analysed in the GenomeStu� dio (Illumina) software. Using the BovineSNP50 v2 BeadChip microarray we tested 91 samples. We then genotyped 46 samples using the BovineHD BeadChip microarray, including eight samples that we repeated to verify reproducibility of the results (see table 1s in supplementary material). Criteria of markers selection In the selection of markers for further analysis, we took into account: call rate ≥ 90%, only those markers that were genotyped in at least 90% of individuals were included; polymorphic markers whose frequency of minor allele amounted to ≥ 0.01 —adopting such a low MAF value as a criterion ensues from the speci� fics of the species, whose genetic variation is extre� mely low; no deviations from HW (Hardy–Weinberg) equilibrium at a significance level of 0.01, no P–C errors or P–P–C errors. Markers meeting the above criteria were individually tested in the GenomeStudio (Illumina) software and re–verified for proper cluster assignment, by the analysis of the GenCall Score value in SNP Graphs (figs. 1, 2). GenCall Score is a quality metric that indicates the reliability of each genotype call (GenomeStudioTM, Genotyping Module v1.0 User Guide, Illumina). Automatic verification was insufficient because the measure of reliability was developed for cattle. We then made a listing and comparison of co� rrectly clustered markers, selected after verification, obtained from both microarrays: BovineSNP50 v2 BeadChip and BovineHD BeadChip. For the analysis of allele frequencies, we used the number of private

1.20 1 0.80 Norm R

The test material consisted of whole blood samples and soft tissues collected by the European bison Gene Bank in the Animal Genetics and Breeding Department (Warsaw University of Life Sciences) according to decision (WPN–I.6401.90.2014.EB.1) of Regional Directorate of Environmental Protection in Warsaw. DNA from blood was isolated by the mag� netic method using a MagMAXTM Express (Applied Biosystems) and a MagMAXTM Total Nucleic Acid Isolation Kit (Ambion), as well as with use of QIAamp DNA Mini Kit (QIAGEN). DNA from soft tissues was isolated using a QIAamp DNA Mini Kit (QIAGEN). The quality and concentration of the isolates was checked using NanoDrop2000 (Thermo Scientific). The DNA was then normalized to a 50ng/μl concen� tration. Samples with a low concentration of DNA were subjected to concentration in a Concentrator 5301 (Eppendorf AG).

0.60 0.40 0.20 0

–0.20

48 0

0.20 0.40 0.60 0.80 Norm Theta

Fig. 1. SNP Graph showing the division of individuals into clusters corresponding to the genotypes at a given locus. The X–axis represents normalized theta (the angle deviation from a pure A signal, where 0 represents a pure A signal and 1 represents a pure B signal), and the Y–axis represents the distance of the point to the origin. Samples are divided according to their genotype. Samples lying within the left region are called AA; samples within the middle region are called AB and samples lying within the right region are called BB. Fig. 1. Gráfico relativo a los PSN en el que se muestra la división de individuos en aglomerados correspondientes a los genotipos de un locus determinado. El eje de abscisas representa theta normalizada (la desviación del ángulo desde una señal A pura, donde 0 representa una señal A pura y 1 representa una señal B pura). El eje de ordenas representa la distancia del punto al origen. Las muestras se han dividido en función de su genotipo. Las muestras que quedan en la región izquierda se denominan AA; las que quedan en la región central se denominan AB; y las que quedan en la región derecha se llaman BB.

alleles and PE (probability of exclusion) GenAlEx 6.5 (Peakall & Smouse, 2012). Using the R envi� ronment version 2.15.3, we carried out the Fisher exact test and the Z–test on two unrelated propor� tions for large samples to determine the statistical significance of the differences in allele frequency between the Lowland and Lowland–Caucasian lines. The final choice was made from SNP markers for which allele frequencies were significantly different, according to both tests, between the genetic lines of the European bison.

Animal Biodiversity and Conservation 40.1 (2017)

1.40 1

0.80

0.60

Norm R

1.20

0.60 0.40

0.20

0.20 0 –0.20

0.40

85 0

0.20 0.40 0.60 0.80 1 Norm Theta

Fig. 2. SNP Graphs showing an abnormal division into clusters. The X–axis represents normalized theta (the angle deviation from pure A signal, where 0 represents pure A signal and 1 represents pure B signal), and the Y–axis represents the distance of the point to the origin. Samples are coloured according to their genotype. Samples marked in black are classified as 'no calls'. Any ambiguous division into clusters excluded a marker from further analysis. Fig. 2. En los gráficos relativos a los PSN se muestra una división anómala en conglomerados. El eje de abscisas representa theta normalizada (la desviación del ángulo desde una señal A pura, donde 0 representa una señal A pura y 1 representa una señal B pura). El eje de ordenas representa la distancia de punto al origen. Las muestras se han coloreado en función de su genotipo. Las marcadas en color negro se clasifican como ''sin determinar''. Las divisiones ambiguas en conglomerados excluyeron un marcador de los análisis posteriores.

The population structure of all tested European bison was evaluated on SNPs common to both mi� croarrays using Bayesian clustering analysis in the software STRUCTURE 2.3.4 (Pritchard et al., 2000; Falush et al., 2003). Analysis was performed under the Correlated Allele Frequencies Model and Admix� ture Model with 30,000 burn–in steps and 100,000 Marcov–chain Monte Carlo (MCMC) replicates for K = 1–6. Tests were conducted five times for each value of K. To determine the most likely value of K, we used the ΔK statistic (Evanno et al., 2005) Structure Harvester software (Earl & vonHoldt, 2012). Results In rounds of arrays preparation, the analysed Bos taurus samples performed well and showed call rates close to 99%. This assured us that the assay per� formance was essentially optimal and no flaw in the laboratory procedure would affect the results for bison. From the 54,609 probes included on the BovineS� NP50 v2 BeadChip, in correctly genotyped individuals there were 51,609 markers with a call rate equal to or greater than 90%, of which 5,997 were polymorphic in European bison (MAF ≥ 0.01). Only 1,421 SNP markers met all the aforementioned criteria.

The BovineHD BeadChip has 777,962 types of probes on its surface. A total of 735,667 SNPs showed a call rate equal to or greater than 90%, and 22,122 of these markers fulfilled all the conditions set. After manual verification of SNP graphs in the GenomeStudio (Illumina) software for all the markers obtained after automatic analysis from both micro� arrays, we selected 806 SNPs and 15,062 SNPs respectively, of which 505 markers were present on both platforms. For these 505 markers, the genotypes of all eight samples analysed on both microarrays were identical. We found highly significant differences in allele frequency between two European bison lines in the case of 1,904 SNPs from both arrays. Finally 1,536 markers were selected for the design of a microarray specific to bison and further analyses: 47 from BovineSNP50 v2, 1,463 from BovineHD, and 26 common to both microarrays; 1,505 selected markers were distributed on autosomes and 31 SNPs on chromosome X. The number of markers on each chromosome ranged from 8 to 136 (table 1). As� suming a similar distribution of the studied SNPs in Bison bonasus and Bos taurus genomes, based on UMD3.1 cattle genome assembly, we found that the mean genomic distance for the selected SNPs was 1,443 kbp and the median distance was estimated for 211 kbp. We found that the highest median distance

Wojciechowska et al.

Table 1. The number of markers per chromosome, genomic distances (according to the UMD3.1 cattle genome build): Chr. Chromosome; N. Number of SNPs per chromosome; MD. Mean distance; SD. Standard deviation. Tabla 1. Número de marcadores por cromosoma, distancias genómicas (según la versión UMD3.1 del genoma de vacuno): Chr. Cromosoma; N. Número de PSN por cromosoma.; MD. Distancia media; SD. Desviación estándar.

Chr

MD (kbp)

SD (kbp)

MD (kbp)

Chr

MD (kbp)

SD (kbp)

MD (kbp)

108

1,456

3,170

253

2,210

4,859

243

2,479

5,584

328

8,524

10,295

1,092

110

952

1,718

229

1,844

3,785

539

1,356

3,715

258

1,297

3,147

174

1,733

7,736

115

1,987

3,579

374

136

822

2,175

116

3,147

6,224

242

118

955

1,797

166

1,126

2,866

179

1,691

4,483

388

2,906

7,550

502

1,238

1,719

579

1,271

2,648

272

2,442

5,582

529

1,079

2,591

164

1,470

4,271

153

1,373

2,986

250

932

1,421

220

1,109

1,681

216

2,206

6,160

248

1,148

1,360

582

2,452

6,432

398

295

813

2,017

3,714

424

All 1,536

1,443

3,852

211

1,037

1,901

189

between SNPs was for chromosome 18 (1,092 kbp) and the lowest was for chromosome X (31 kbp). The number of private alleles in the Lowland–Cau� casian line was considerably higher than in Lowland line (611 and 26 respectively). Selected 1,536 SNPs were plotted against cattle chromosomes in figure 1s in supplementary material. We calculated the prob� abilities of exclusion coefficients (PE, both parents known; PE1, only one parent known; PE2, exclude both parents) were calculated for combined loci from each microarray, for pooled genetic lines, and separately. For 47 SNPs from Bovine SNP50, all coefficients obtained for the LB line were significantly lower than in the LC line and in the pooled samples. The statistical difference between the value of this rate for LC and the whole population was found for PE1 only (table 2). In contrast, all analyses of PE for 1,489 markers from Bovine HD gave a result of 1,000. The minor allele frequency (MAF) was calculated separately for both lines to pre–estimate the genetic variability among the European bison analysed. Fig� ure 3 shows the distribution of minor allele frequency within Lowland and Lowland–Caucasian lines. In order to minimize miscalculation arising from the different

Table 2. Probability of exclusion combined for all loci from Bovine SNP50 microarray (47 SNPs): PE. Probability of exclusion (both parents known); PE1. Probability of exclusion (only one parent known); PE2. Probability of exclusion (neither parents known); * P ≤ 0.05; ** P ≤ 0.01 Tabla 2. Probabilidad de exclusión combinada para todos los "loci" de la micromatriz multigénica Bovine SNP50 (47 PSN): PE. Probabilidad de exclusión (ambos progenitores conocidos); PE1. Probabilidad de exclusión (solo un progenitor conocido); PE2. Probabilidad de exclusión (ningún progenitor conocido); * P ≤ 0,05; ** P ≤ 0,01.

PE1

PE2

0.741**

0.441**

0.897**

0.997**

0.875** *

1.000**

LB + LC

0.986**

0.720* **

0.999**

Animal Biodiversity and Conservation 40.1 (2017)

50% 40% LB

30%

20% 10%

D E F G H Minor allele frequency

Fig. 3. Distribution of minor allele frequency (MAF) within the Lowland line (LB) and the Lowland–Caucasian line (LC). The 0 value of MAF indicates that loci were polymorphic overall, but monomorphic within one of the genetic lines: A. 0.00–0.05; B. 0.05–0.10; C. 0.10–0.15; D. 0.15–0.20; E. 0.20–0.25; F. 0.25–0.30; G. 0.30–0.35; H. 0.35–0.40; I. 0.40–0.45; J. 0.45–0.50. Fig. 3. Distribución de la frecuencia alélica mínima dentro de la línea Lowland (LB) y la línea Lowland– Caucasiana (LC). El valor 0 de la frecuencia alélica mínima indica que los "loci" eran polimorfos en general, pero monomorfos dentro de una de las líneas genéticas. (Para las abreviaturas, véase arriba.)

number of polymorphic loci in both genetic lines and to firmly demonstrate the difference between them, we also included loci polymorphic in one line but monomorphic in the other. In the Lowland line, we found almost 50% of mo� nomorphic SNPs, indicating a high level of inbreeding, which is unavoidable in a closed group. In contrast to the Lowland population, in the Lowland–Caucasian line, more than 80% of the markers had an MAF > 0.2; of these, about 50% were characterized by MAF as greater than 0.3, indicating a far greater variation bet� ween individuals in the Lowland–Caucasian line than in the Lowland line. Analysis of the genetic structure of the population carried out in STRUCTURE 2.3.4 on all tested samples also showed clear differences between genetic lines. The highest value DK pointed to the division of the population into two clusters (K = 2), dividing individuals from both lines to clearly separate groups. The results of this analysis are given in figure 4. Discussion The BovineSNP50BeadChip, which was designed for domestic cattle, has been successfully used to analyse the genetic structure of several wild species. In the case of goats, two subspecies: the Tatra chamois (Rupicapra rupicapra tatrica) and the Alpine chamois (Rupicapra rupicapra rupicapra), were genotyped using the above mentioned microarray (Demontis et al., 2011). In this study, 505 of among 54,000 markers were found to be polymorphic, although only 151 were correctly clus� tered after manual verification. Such a low number of correctly clustered markers could indicate low

genetic variability, but could also be the result of species differences. In turn, a study by Haynes & Latch (2012) for the deer species Odocoileus hemionus columbianus, O. h. hemionus and O. Virginianus, obtained 21,131 genotyped markers in at least 90% of the animals tested, of which 1,068 were polymorphic. Although Odocoileus spp. are genetically more distant from domestic cattle than bison, use of the same microarray allowed to obtain a relatively high number of polymorphic markers. The evolution of the genus Bison shows that the European bison as a wild species is genetically more similar to the domestic cattle than the Ame� rican bison. The reason for this is the hybridization of the aurochs, which —like the introgression of yak in Bison bison— influenced the distance of these two subspecies of Bison (Nowak & Olech, 2008b). For comparison, in studies by Tokarska et al. (2009) the 50 Lowland European bison (LB) tested gave a reading of 52,978 SNPs, of which 960 markers were polymorphic. In contrast, Kamiński et al. (2012), despite testing only 10 European bison (five LB and five LC specimens), obtained 1,337 polymorphic SNPs. The number of markers was higher due to inclusion of both genetic lines in the studies and LC is intrinsically more diverse than LB, which is also noticeable in the present study. The participation of ancestors is different in each of the genetic lines, therefore the testing of only one of them is insufficient and cannot be used to estimate the genetic structure of the entire species. Tokarska et al. (2009) presented results of paternity analysis carried out on two marker sets: the most heterozygous SNPs, and a randomly selected set of markers. They concluded that in the case of the first set, 50–60 SNPs would be needed to assign

Wojciechowska et al.

100% 90% 80% 70% 60% 50% 40% 30% 20% 10% 0%

Lowland–Caucasian line

Lowland line

Fig. 4. Bayesian clustering analysis performed by STRUCTURE. Each individual is represented by one bar divided into segments, illustrating the proportion of estimated membership in each cluster. The vertical black lines separate a group of European bison from the Lowland line. Fig. 4. Análisis bayesiano de conglomerados llevado a cabo por STRUCTURE. Cada individuo está representado por una barra dividida en segmentos que ilustra la proporción estimada de individuos en cada conglomerado. Las líneas negras verticales separan un grupo de bisonte europeo de la línea Lowland.

paternity with 95% confidence and 80–90 loci for the random set. Our study for the probability of ex� clusion (PE) partly confirms these results. Analyses were carried out for pooled genetic lines (LB + LC), and each line separately. We checked how many loci would be needed for a 99% of confidence. For the PE (both parents known) in the Lowland line —the same genetic line as in Tokarska et al. (2009)— 57 markers would be sufficient, but in the case of PE1 (only one parent known) 160 would be necessary. Other results were obtained for the Lowland–Caucasian line; only 27 SNPs in the case of PE, and 53 markers for PE1. For combined lines we estimated that 50 markers would be needed for PE, and 59 SNPs in the case of PE1. In 2015, Oleński et al. used the BovineHD BeadChip for an association study in the Lowland line. Besides reporting SNP markers significantly associated with postitis disease, the authors concluded that infor� mation from the subsets of SNPs could be a useful tool for the European bison breeding program, from a conservation and epizootic point of view. When anticipate that our design of a specific SNP panel for European bison with characteristic markers for particular genetic lines (Lowland and Lowland– Caucasian) and parental lines will provide a key tool for the future analyses of the genetic structure of the species, specimen identification, and control of the origin and relatedness of European bison, both in captive breeding centres and in free roaming populations. Such knowledge is crucial for optimal management of breeding programs for these highly valued animals, and will contribute to their direct protection in the future.

Acknowledgements We thank the veterinarians and breeders at breeding centers in Poland and other regions for providing bio� logical material from European bison. The study was funded by the Project ‘Ex situ conservation of Euro� pean bison Bison bonasus in Poland’ under a measure 5.1 priority V of Operational Programme Infrastructure and Environment 2007–2013, in accordance with the grant agreement no. POIS.05.01.00–00–155/09–02 and the agreement on co–financing of eligible expen� ditures of the Project implemented under priority V of Operational Programme Infrastructure and Environ� ment no. 615/2010/Wn–50/OP–WK–PS/D and from a grant from Warsaw University of Life Sciences no. 505–10–070200–K00253–99. The authors declare that they have no conflict of interest. References Demontis, D., Czarnomska, S. D., Hajkova, P., Zema� nova, B., Bryja, J., Loeschcke, V. & Pertoldi, C., 2011. Characterization of 151SNPs for population structure analysis of the endangered Tatra chamois (Rupicapra rupicapra tatrica) and its relative, the Alpine chamois (R. r. rupicapra). Mammalian Biology, 76: 644–645. Earl, D. A. & vonHold, B. M., 2012. STRUCTURE HARVESTER: a website and program for visu� alizing STRUCTURE output and implementing the Evanno method. Conservation Genetics Resources, 4(2): 359–361. Evanno, G., Regnaut, S. & Goudet, J., 2005. Detect� ��

Animal Biodiversity and Conservation 40.1 (2017)

ing the number of clusters of individuals using the software STRUCTURE: a simulation study. Molecular Ecology, 14: 2611–2620. Falush, D., Stephens, M. & Pritchard, J. K., 2003. Inference of population structure using multilocus genotype data: Linked loci and correlated allele frequencies. Genetics, 164: 1567–1587. Gralak, B., Krasińska, M., Niemczewski, C., Krasiński, Z. A. & Żurkowski, M., 2004. �� Polymorphism of bo� vine microsatellite DNA sequences in the lowland European bison. Acta Theriologica, 49(4): 449–456. Hartl, G. B. & Pucek, Z., 1994. Genetic depletion in the European bison (Bison bonasus) and the significance of electrophoretic heterozygosity for conservation. Conservation Biology, 8: 167–174. Haynes, G. D. & Latch, E. K., 2012. Identification of novel single nucleotide polymorphisms (SNPs) in Deer (Odocoileus spp.) using the BovineSNP50 BeadChip. PLoS ONE, 7(5): e36536. Kamiński, S., Olech, W., Oleński, K., Nowak, Z. & Ruść, A., 2012. Single nucleotide polymorphisms between two lines of European bison (Bison bonasus) detected by the use of Illumina Bovine 50 K BeadChip. Conservation Genet Resour, 4: 311–314. Łopieńska, M., Nowak, Z., Charon, K. M. & Olech, W., 2003. Estimation of polymorphism in chosen region of MHC in two genetic lines of European bison (Bison bonasus L.). Applied Science Reports: Cattle production and breeding, 68(1): 17–24. – 2011. A comparison of polymorphism of DQA genes in European bison belonging to two genetic lines: Lowland and Lowland–Caucasian. Ann. Warsaw Univ. of Life Sci.–SGGW, Anim. Sci., 49: 93–102. Nowak, Z. & Olech, W., 2008a. Microsatellite vari� ability within the sex chromosomes in European bison. European Bison Conservation Newsletter, 1: 72–78. [In Polish] – 2008b. Verification of phylogenetic hypothesis concerning the evolution of genus Bison. Ann. Warsaw Univ. of Life Sci.–SGGW, Anim. Sci., 45: 65–72. Olech, W., 1999. The number of ancestors and their contribution to European bison (Bison bonasus L.)

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Animal Biodiversity and Conservation 40.1 (2017)

Supplementary material

Table 1s. Detailed information about individuals used in the study: SID. Sample ID; GL. Genetic line (LB. Lowland line, LC. Lowland–Caucasian line); PL. Paternal line (15. Begründer, 45. Plebejer, 100. Kaukasus); ML. Maternal line (16. Plavia, 35. Plewna, 42. Planta, 89. Bilma); FID. Father ID; MID. other ID; UM. Used microarray (50. BovineSNP50 v2 BeadChip, HD. BovineHD BeadChip). Tabla 1s. Información detallada sobre los individuos empleados en el estudio: SID. Identificador de cada muestra; GL. Línea genética (LB. Línea Lowland, LC. Línea Lowland–Caucasiana); PL. Línea paterna (15. Begründer, 45. Plebejer, 100. Kaukasus); ML. Línea materna (16. Plavia, 35. Plewna, 42. Planta, 89. Bilma); FID. Identificador paaterno; MID. Otro identificador; UM. Micromatriz multigénica utilizada (50. BovineSNP50 v2 BeadChip, HD. BovineHD BeadChip). SID

623

Białowieża

624

Białowieża

625

Białowieża

626

Białowieża

631

Białowieża

632

Białowieża

637

Białowieża

672

Białowieża

701

Białowieża

637

712

Białowieża

903

716

Białowieża

742

Białowieża (free–living herd)

745

Białowieża

758

Białowieża (free–living herd)

767

Białowieża (free–living herd)

773

Białowieża

781

Białowieża

782

Białowieża

785

Białowieża (free–living herd)

806

Białowieża

L110

834

Białowieża (free–living herd)

838

Białowieża (free–living herd)

841

Białowieża

868

Białowieża (free living herd)

871

Białowieża (free living herd)

877

Białowieża

878

Białowieża

888

Białowieża (free–living herd)

900

Białowieża

901

Białowieża

903

Białowieża

915

Białowieża

L002

Międzyzdroje

Sex

Breeding centre

FID

MID

L110 632

631

878

Wojciechowska et al.

Table 1s. (Cont.)

SID

Sex

L024

Borki

L034

Ebeltoft

L110

Białowieża

L111

Borki

L143

Białowieża

L147

Borki

L149

Borki

L201

Pszczyna

L209

Białowieża

903

L227

Białowieża

781

L304

Wrocław

L331

Borki

L342

Niepołomice

L410

Bydgoszcz

L457

Białowieża (free–living herd)

L460

Białowieża (free–living herd)

L498

Panevėžys

50&HD

L540

Białowieża

L541

Białowieża

L570

Białowieża (free–living herd)

L584

Pszczyna

L585

Pszczyna

L619

Gołuchów

Breeding centre

FID

MID

781

L149

L147

L640

Smardzewice

K015

Amsterdam

K024

100

Vanatori Neamt

K026

100

Vanatori Neamt

50&HD

K027

Vanatori Neamt

K032

100

Bussolengo

K033

100

Bussolengo

K052

Avesta

K106

Bayerischer Wald

K107

Praga

K109

100

Karlsruhe

K110

100

Karlsruhe

K111

100

Karlsruhe

K153

Damerower Werder

K173

Gera

K174

100

Karlsruhe

K026

50&HD

Animal Biodiversity and Conservation 40.1 (2017)

iii

Table 1s. (Cont.)

SID

Sex

K175

100

Breeding centre

Karlsruhe

FID

K176

100

Karlsruhe

K189

Damerower Werder

K194

Goldau

K198

Borås

K213

Damerower Werder

K214

Eriksberg

K219

Gera

K220

Gera

K225

Damerower Werder

K233

Hardehausen

K234

Hardehausen

K235

Hardehausen

K236

Hardehausen

K237

Hardehausen

K238

100

Hardehausen

K239

100

Hardehausen

K244

K240

Hardehausen

K242

50&HD

K241

100

Hardehausen

K235

K242

Hardehausen

K244

Hardehausen

K245

Hardehausen

K248

50&HD

K247

Hardehausen

K244

50&HD

K248

Hardehausen

K250

Damerower Werder

K274

Damerower Werder

K275

Damerower Werder

K282

Bern

K284

Goldau

K286

Thoiry

K289

Damerower Werder

K290

Damerower Werder

K291

Damerower Werder

K292

Damerower Werder

K294

100

Hardehausen

K302

Damerower Werder

K358

Neumünster

K373

100

Karlsruhe

K377

100

München, Hellabrunn

K238

MID

K173

K233

K244

K233

50&HD

Wojciechowska et al.

Table 1s. (Cont.)

SID

Sex

K378

München, Hellabrunn

Breeding centre

FID

MID

UM HD

K380

München, Hellabrunn

K461

Neumünster

K358

K502

100

Weilburg

K533

100

Bad Orb

K534

Eulbach

K545

100

Hanau

K559

100

Sababurg

K545

K560

100

Sababurg

K545

K563

100

Edertal–Hemfurth

K586

100

Hardehausen

K238

K591

100

Hardehausen

K238

KB317

Bieszczady (free–living herd)

KB342

Bieszczady (free–living herd)

KB344

Bieszczady (free–living herd)

KB394

Bieszczady (free–living herd)

KB396

Bieszczady (free–living herd)

KB397

Bieszczady (free–living herd)

Animal Biodiversity and Conservation 40.1 (2017)

Chromosome length (Mb)

1 2

3 4

5 6

Chromosome 7 8 9 10 11 12 13 14 15 16 17 18 19 20 21 22 23 24 25 26 27 28 29 X

100

150

Freq LB LC Chr

Fig. 1s. Localization of SNPs with significant differences in allele frequency between LB and LC populations and SNPs with private alleles in each population. Black line indicates chromosome length: Freq. SNPs differing in frequency between LB and LC population; LB. SNPs with private allele in LB population; LC. SNPs with private allele in LC population; Chr. Chromosome. Fig. 1s. Localización de los PSN que presentan diferencias significativas en las frecuencias alélicas entre las poblaciones LB y LC y de los PSN con alelos privados en cada una de las poblaciones. La línea negra indica la longitud del cromosoma: Freq. PSN cuya frecuencia difiere entre las poblaciones LB y LC; LB. PSN con alelos privados en la población LB; LC. PSN con alelos privados en la población LC; Chr. Cromosoma.

Animal Biodiversity and Conservation 40.1 (2017)

Eco–geographical characterization of aquatic microhabitats used by amphibians in the Mediterranean Basin M. Benítez, D. Romero, M. Chirosa & R. Real

Benítez, M., Romero, D., Chirosa, M. & Real, R., 2017. Eco–geographical characterization of aquatic microhabitats used by amphibians in the Mediterranean Basin. Animal Biodiversity and Conservation, 40.1: 27–40. Abstract Eco–geographical characterization of aquatic microhabitats used by amphibians in the Mediterranean Basin.— Small freshwater ecosystems, whether of natural or artificial origin, are aquatic microhabitats for many species and are particularly important in the Mediterranean region. This study characterizes the aquatic microhabitats suitable for amphibian reproduction in the Andalusian Mediterranean Basin and identifies the environmental and geographical features that determine the presence of different amphibian species in these water bodies. Geographical and environmental favourability models were performed to determine the relationship between characteristics of the microhabitats and species presence. The characteristics analysed were geographical location, external environment (climate and topography), surrounding conditions (connectivity and conserva� tion), type of water body, water conditions, and water dimensions. Microhabitats located in the western and central part of the study area were geographically favourable for most species. In descending order, the most common environmental factors characterizing the microhabitats were typology, surrounding conditions, water condition, external environment and size of the water body. The most common variables in the models were the connectivity between water bodies and old wells, a frequent type of microhabitat in areas of traditional cultures. Management plans should take these results into account in efforts to preserve these habitats for wildlife and especially amphibians. Key words: Water bodies, Environmental favourability, Freshwater ecosystem, Iberian peninsula, Conservation Resumen Caracterización ecogeográfica de los microhábitats acuáticos utilizados por los anfibios en la cuenca mediterránea.— Los ecosistemas de agua dulce de pequeño tamaño, independientemente de su origen natural o artificial, constituyen microhábitats acuáticos de gran valor para muchas especies, especialmente en la región mediterránea. En este estudio se caracterizan los microhábitats acuáticos disponibles para la reproducción de los anfibios en la cuenca mediterránea andaluza y se identifican las características ambientales y geográficas que determinan la presencia de las distintas especies de anfibios en ellos. Se utilizaron modelos de favorabilidad geográfica y ambiental para determinar la relación entre las características de los microhábitats y la presencia de especies. Las características analizadas fueron la ubicación geográfica, el ambiente externo (clima y topografía), las condiciones del entorno (co� nectividad y conservación), el tipo de masa de agua, las condiciones del agua y las dimensiones de la masa de agua. Los microhábitats ubicados en la parte occidental y central de la zona de estudio fueron geográficamente favorables para la mayoría de las especies. En orden decreciente, los factores ambientales más comunes que caracterizaron los microhábitats fueron la tipología, las condiciones del entorno, las condiciones del agua, el ambiente externo y el tamaño de la masa de agua. Las variables más comunes en los modelos fueron la conectividad entre las masas de agua y un tipo de microhábitat frecuente en zonas de cultivos tradicionales: los pozos antiguos. Los planes de gestión deberían tener en cuenta estos resultados en las iniciativas encaminadas a conservar estos hábitats para la fauna y especialmente para los anfibios. Palabras clave: Masas de agua, Favorabilidad ambiental, Ecosistemas de agua dulce, Península ibérica, Conservación

ISSN: 1578–665 X eISSN: 2014–928 X

Benítez et al.

Received: 21 IV 15; Conditional acceptance: 27 V 16; Final acceptance: 1 VIII 16 Maribel Benítez, Grupo de Biología y Conservación de Vertebrados Mediterráneos, Depto. de Zoología, Fac. de Ciencias, Univ. de Granada, Campus Fuentenueva s/n., 18071 Granada, Spain.– David Romero & Raimundo Real, Grupo de Biogeografía, Diversidad y Conservación, Depto. de Biología Animal, Fac. de Ciencias, Univ. de Málaga, E–29071 Málaga, Spain.– Manuel Chirosa, Inst. de Investigación y Formación Agraria y Pesquera (IFAPA), Consejería de Agricultura y Pesca, Junta de Andalucía, Camino de Purchil, Apdo. 2027, 18080 Granada, Spain. Corresponding author: M. Benítez. E–mail: mbenitez@ugr.es

Animal Biodiversity and Conservation 40.1 (2017)

Introduction Water is an essential element for wildlife but human exploitation of this resource can lead to changes in freshwater ecosystems and even to their disappearan� ce (Blondel et al., 2010). This conflict between humans and other species is especially critical in areas such as the Mediterranean region due to summer xericity, high biodiversity values (Grillas et al., 2004; Blondel et al., 2010) and high levels of species endemism (Zacharias & Zamparas, 2010). Maintenance of such aquatic habitats in Mediterranean climatic regions is thus essential for life and biodiversity (Grillas et al., 2004; Zacharias & Zamparas, 2010). Many of these freshwater ecosystems are small and they are often characterized as microhabitats. Humans have traditionally built a vast number of artificial water bodies such as pools, ponds, basins, fountains, and water troughs. In Andalusia, the first of these water bodies were constructed in the Bronze Age (about 3,700 BP) with the rise of the metallurgical practices of the Argar culture (García–Alix et al., 2013). In the Andalusian Mediterranean Basin, in particular, artificial water bodies have been used extensively by wildlife, especially amphibians, to supply their physi� ological requirements and have constituted important breeding habitats for many amphibian species (Beja & Alcázar, 2003; Casas et al., 2012). Thus, both natural and artificial water bodies must be preserved to maintain the breeding habitat of amphibian species (Beja & Alcázar, 2003; García–Muñoz et al., 2010). This is particularly important given that amphibians are the most threatened vertebrates on the planet (Beebee & Griffiths, 2005). In recent decades, the conflict between humans and amphibians over water has intensified due to changes in human water use linked to economic ac� tivity (Brühl et al., 2013), especially due to intensive agricultural practices. These practices have led to the replacement of outdoor, open water infrastructures with pipelines or underground water bodies, with the aim of increasing efficiency. These new infrastructures do not provide the conditions necessary for amphibian breeding. Traditional water infrastructures have been consequently abandoned, and amphibian conserva� tion status has become a concern, as much as the conservation of natural microhabitats. Amphibians are good indicators of the state of conservation of aquatic habitats, because part of their life cycle is linked to water bodies (Duelman & Trueb, 1986). Different species take advantage of different characteristics of the water bodies for their breeding success (Egea–Serrano et al., 2006; Richter–Boix et al., 2006). For this reason, characterization of water bodies is important to determine the role that their features have on the maintenance of specific amphib� ian populations (Calhoun & Hunter, 2003). The relationship between amphibian presence and aquatic habitat characteristics is widely recognized, but remains vaguely defined. A key issue in amphibian conservation planning is to identify the conditions that determine the presence of each amphibian species at each water body. In this context, habitat models can be

used to identify the environmental features that explain species distributions (Fielding & Haworth, 1995). These models use mathematical algorithms to reveal the char� acteristics of the habitats that are most relevant for the physiological and biological requirements of the species (Barbosa et al., 2001; Muñoz et al., 2005; Romero et al., 2013). However, the geographical context used to build the models affects the relationship between environmental variables and species distribution that are retained in the models (Acevedo et al., 2012). Previous studies have analysed amphibian distribution and habitat characteristics separately. Some authors have studied reproductive habitats of species without taking into account the spatial structure of the popu� lations (Egea–Serrano et al., 2006), whereas others have analysed species distribution without taking into account the specific characteristics of the microhabi� tats (Guerrero et al., 1999). Few authors have used local variables measured in situ as predictors in local distribution models (Gómez–Rodríguez et al., 2012; Ferreira & Beja, 2013). The aim of this study was to characterize the ty� pes of aquatic microhabitats available for amphibian reproduction in the Andalusian Mediterranean Basin and to identify the environmental and geographical features that determine the presence of amphibian species in these water bodies. We also discuss the role of the characteristics of the microhabitats in the conservation of amphibian species. Material and methods The study area This study was carried out in the Andalusian Medi� terranean Basin, located in the south of the Iberian peninsula, and comprising the hydrographical basins within the Autonomous Community of Andalusia that flow into the Mediterranean Sea (Fig. 1). This territory has a surface area of 18,193 km2 roughly contained in a 50 km by 350 km strip between the Strait of Gibraltar (Cadiz) and the Almanzora river basin (Almeria), and includes 652 km of coastline (MMARM, 2008). This geographical area includes large altitudinal differences and rugged mountains. The climate is Mediterranean with geographically–driven subclimates: subtropical, subdesert, continental, and mountain. The rainfall gradient ranges from 2000 mm in the west, due to the Atlantic influence, to 200 mm in the east in the Tabernas Desert (CAPMA, 2012). The extremes of the ombroclimatic belt are hyper–humid in the west and arid in the east (López et al., 2008). The temperature gradient ranges from an annual average of 18ºC on the coast to 2.5ºC in the mountains of the Sierra Nevada (Ninyerola et al., 2005). Sampling of microhabitats The classification of aquatic ecosystems is difficult due to their great variety in size, typology, hydroperiod, and vegetation. The aquatic microhabitats analysed include Williams’s mesohabitats (Williams, 2006),

Benítez et al.

which were defined as temporary streams and ponds, snow–melt pools, monsoon rain pools, floodplain pools, dewponds, and wetland pools. Additionally, the aquatic microhabitats investigated in this work differ according to whether they are artificial, natural, or mixed (Beja & Alcázar, 2003; Grillas et al., 2004) and range from temporary aquatic ecosystems to water storage infrastructures of different sizes that are scattered in agricultural landscapes (Zacharias & Zamparas, 2010). We defined aquatic microhabitats as systems linked to epicontinental, non–lotic, temporary or permanent waters, linked to upwelling, drainages, or natural or artificial ponds, with an approximate maximum volume of 200 m3 and surfaces of up to 500 m2 and with an associated biological community (Benítez et al., 2011). The following literature was consulted to identify potential sites for amphibian breeding within the study area: (i) official cartographic sources (Junta de An� dalucía, 2004; Instituto Geográfico Nacional, 2009); (ii) databases and scientific collections of amphibians (Asociación Herpetológica Española; colecciones del Museo Nacional de Ciencias Naturales; Estación Bio� lógica de Doñana; Estación Experimental de Zonas Áridas; colección del departamento de Zoología de la Universidad de Granada); and (iii) theses and unpu� blished reports on amphibians (Reques et al., 2006). We identified a total of 13,650 water bodies. All these points were small and 43% of them had the confirmed presenceof amphibians, or suitability for them. The location of these points was processed using ESRI ArcMap 9.2 software. The itineraries for sampling were designed to include the maximum number of representative and accessible water bodies in all the river basins. Between 2009 and 2011, we sampled 568 water bodies over 64 days, travelling a total of 14,500 km (see fig. 1). Two biologists obser� ved and sampled each water point for an average of 15 minutes. Each sampler recorded information on the presence of species, spatial situation, and environmental variables at each water body as well as the external environment (table 1). Geographic co� ordinates and altitude wereobtained by GPS (Garming X12). The following sampling method was used: (1) searching for and counting the number of individual amphibians present in the vicinity of the water body; (2) searching for and counting the eggs in the water body; and (3) examining the entire water body to detect larvae (Heyer et al., 1994). Given that amphibians are difficult to detect be� cause their reproductive cycles are closely linked to weather (Mazerolle et al., 2007), some authors propose different methods in order to increase the probability of amphibian detection (Mazerolle et al., 2007; Gómez–Rodríguez et al., 2012). However, in our study these methods were not feasible due to the large extension of the territory and the great difficulty of access to each water body. We considered instead the phenology of each species in each geographical location to visit the microhabitats when detectability was highest. Sampling efforts were concentrated in winter and spring in most locations, but in summer for sites above 1,500 meters of elevation. To locate

amphibians, we looked for adults, larvae or breeding calls in both the water body and the surroundings in a radius of 10 m. We detected larvae by dip–net� ting (Heyer et al., 1994). When low detectability of a particular species was an issue, the water bodies were visited several times; approximately 14.4% of the water bodies were revisited. The number of visits was included in the analysis to determine whether the different sampling effort had an effect on each species distribution. Variables and explanatory factors The presence of each amphibian species around each water body and its surroundings was recorded when eggs, larvae, juveniles, adults, calls, or identifiable remains (skin, bones, dry larvae, etc) were detected at the water bodies or in their vicinity. We obtained in situ data on 38 variables related to seven explanatory factors: five environmental factors, one geographical factor and one related to sampling effort (table 1). These environmental and geographical factors were used to investigate the relationship between the characteristics of the microhabitats and the presence of the species at each water body. We opted for an analytical approach that evaluated the role of each factor individually, because the investigation included the whole range of microhabitat characteristics that are relevant for amphibians. This is particularly important in the context of the current decrease of water bodies in the Mediterranean region, as all the critical water bodies relevant to amphibian species must be pre� served and this cannot be achieved without knowing their critical characteristics separately. In a synthetic model that combined the effect of all factors (Romero et al., 2015) it would be more difficult to determine the role of each individual factor, because the most relevant factors might overshadow the effect of those of lesser, yet relevant, importance. The location of each microhabitat, identified by longitude (X) and latitude (Y), was considered in order to evaluate the effect of the geographical fac� tor. In this way, latitude and longitude were used to build nine spatial variables that were useful only to perform a trend surface analysis (Legendre, 1993). Thus, we included a series of polynomial expansions —X, Y, X2, Y2, X3, Y3, X × Y, X2 × Y, Y2 × X— in a logistic regression to detect the spatial structure of the water bodies used by each amphibian species. The climatic variables (air temperature and wind) and the topographic variable (altitude) were included in the external environmental factor. The variables related to connectivity between water bodies, according to their proximity to each other, and the degree of conserva� tion of the microhabitat, according to the number of identified threats, were included in the surrounding factor. Connectivity was estimated by considering a 2 km buffer for each point, given that the species with the greatest displacement is the Natterjack toad (Bufo calamita), whose maximum detected displacement is 2.6 km during the reproductive period (Sinsch, 1992). The other environmental factors describe the cha� racteristics of the water bodies, such as the type of

Animal Biodiversity and Conservation 40.1 (2017)

Iberian peninsula

(b)

(a) (c) Altitude

100 km

0–300

300–600

600–1,000

1,000–1,500

1,500–2,000

2,000–3,480

Number of species 0

1–2 3–7

Fig. 1. Study area and distribution data showing number of species found in each water point. Altitudinal levels are indicated. The dotted ovals (a) and (b) show the most favourable zones for most of species. The dotted circle (c) shows the most favourable zone for two species (P. perezi and B. calamita). Fig. 1. Zona de estudio y distribución de los datos, indicando el número de especies encontradas en cada punto de agua. Se indican los niveles altitudinales. Las líneas de puntos ovales (a) y (b) indican las zonas favorables para la mayoría de las especies. La linea de puntos circular (c), indica la zona más favorable para dos de ellas (P. perezi y B. calamita).

water body (Morrell, 2008), the biological and physical conditions of the water, and the size of the water body. Nine natural and nine artificial types of water bodies were identified in the study area (table 1), including, for the first time in this kind of study, old shallow wells with stone or brick walls, whose width allowed sufficient light to enter for aquatic flora and fauna to proliferate (Lanz & Greenpeace España, 1997). Water temperature, pH, and conductivity were measured using a thermometer (Eutech, ECScan, accuracy: ± 0.5ºC), a pH meter (Eutech, pHScan 2, accuracy: ± 0.1 pH), and a conductimeter (Eutech, ECScan, SE: ± 0.01 mS), respectively. Each instrument was immer� sed to a depth of 2 cm. The vegetation at water bodies was assessed according to the macrophyte index of Suárez et al. (2005) and greater value was placed on the dominant taxa to assess water quality. Similarly, the macroinvertebrates were assessed according to the index of Alba–Tercedor et al. (2002). Water colour was measured on a gradient from transparent to opa� que (table 1). The movement of the water mass was characterized as a function of water velocity in cm/s in an ascending gradient. The small microhabitats were measured on site and those associated to large ponds, lagoons, or reservoirs were measured using orthophotos (aerial photos corrected to represent

an orthogonal projection without perspective effects, published by Junta de Andalucía, 2004). Wind, connectivity, conservation, vegetation, ma� croinvertebrates, colour, and movement are presented as semi–quantitative categorical variables (table 1). We derived wind and connectivity semi–quantitatively from speed in km/h and distance in meters from the closest water point, respectively, by considering only substantial dissimilarities in them that make a real difference for amphibians, and which go unnoticed in a continuous gradient. The typology of water bodies is presented as 18 binary variables and provides information on the presence (1) or absence (0) of each type at each body. Geographic and environmental favourability models A trend surface analysis was performed to obtain information on the relationship between the presence/ absence of each amphibian species at each water body and geographical location (Legendre, 1993). We used backward stepwise logistic regression of the analysed water bodies on the nine spatial variables to identify the geographical probability trend and to remove the components of longitude and latitude that were redundant. Finally, the geographical pro�

Benítez et al.

Table 1. Description of variables used to build the environmental models: P/A. Presence/absence. Tabla 1. Descripción de las variables utilizadas para elaborar los modelos: P/A. Presencia/ausencia.

Factors Variables

Units

Explanations

Geographical location Longitude (X)

The Universal Transverse Mercator (UTM) coordinate system

Latitude (Y)

projected onto the zone 30 S

External environment Air temperature

ºC

Altitude

Wind

Range of

values 1–6 0–2 km/h, smoke rises vertically; 2. Light air, 2–6 km/h, wind

Values according to speed in km/h at the following intervals: 1. Windless,

direction is defined by the smoke; 3. Breeze, 7–11 km/h, wind is

noticeable on face, tree leaves move; 4. Light breeze, 12–19 km/h,

tree leaves move continuously; 5. Moderate breeze, 20–29 km/h,

small branches move, dust rises; 6. Strong wind, > 30 km/h, small

trees move, waves form in pools

Surroundings Connectivity

Range of Degrees of distance in metres of each buffer zone around the values 0–3 microhabitat, with the following values: 0. Isolated point, the closest

point, the closest points are 2,000 m away; 1. The closest points

are 1,000 m away; 2. The closest points are between 500 and 1,000 m;

3. The closest points are between 0 and 500 m

Conservation

Range of Values according to degree of threat: 1. Points with more than 3 types values 1–3 of threats; 2: Points with 1, 2 or 3 types of threats; 3. Highest level

of conservation, void of threats. Types of threats: chemical pollution

organic pollution, construction excavation, residual waste water,

alien species, wild boar impact, neglect, floods, drought,

water harvesting, cleaning or emptying, excessive livestock

Typology River

P/A

Permanent drainage system. In this study, river also refers to small

ponds formed on the banks of the riverbed and the pools in the

headwaters

Stream

P/A

Short water flow, almost continuous

Spring–fed river

P/A

Small wetland

P/A

Bank–side spring that gives rise to or adds water in significant quantities to a river or stream An area of land that is permanently wet due to shallow superficial

or subterranean sources of water including roadside water ditches

Spring

P/A

Natural upwelling of groundwater

Mine water

P/A

Artificial underground gallery that collects groundwater by gravity. If

this gallery is not fully enclosed the it is called a ditch

Seepage

Small quantities of groundwater which flow from a non–permanent

P/A

sources that may produce small wetlands

Temporary pond

Accumulation of non–permanent water, resulting from periods of

P/A

heavy rain or other overflow water

Animal Biodiversity and Conservation 40.1 (2017)

Table 1. (Cont.)

Factors Variables

Units

Natural pond

P/A

Explanations Accumulation of permanent water, including high mountain lakes

and other naturally occurring small bodies of water

Fountain

Simple man–made construction to raise water from a spring for

P/A

daily use

Drinking trough

A receptacle built at a water source to provide livestock with

P/A

drinking water or to wash clothes

Plastic–lined pond P/A

An artificial hole lined with plastic for water storage

Earth–lined pond

P/A

An artificial hole lined with soil that can be filled naturally or

artificially for water storage

Concrete pool

An artificial water tank or pool with masonry walls to store water

P/A

for utilitarian purposes, such as irrigation and fish breeding or for

ornamental purposes

Cistern

P/A

Underground tank

Irrigation ditch

P/A

Stone, concrete, or earth–lined ditch or channel for irrigation and

Well

P/A

other purposes Hole excavated to locate a usable vein of water. In this study, well

refers to old shallow holes that are suitable microhabitats for

wildlife (as opposed to modern, covered wells)

Dyke or Levee

Low transverse barriers built across ravines or streams to stop

P/A

sedimentation and erosion during periods of heavy rain, in contrast

to dykes built for controlling water flow with dams

Water conditions Vegetation

Range of Values according to presence of plant species that are indicators values 0–3 of water quality: 0. No vegetation, clear or turbid due to sediment

and/or water phytoplankton; 1. Presence of helophytes and/or

benthic filamentous algae; 2. Characeae algae; 3. Presence of other

macrophytes

Macroinvertebrates Range of

Values based on the presence of species that are indicators

values 0–3 of water quality: 0. No macroinvertebrates; 1. Presence of Diptera

and/or Hemiptera and/or Annelids; 2. Molluscs and/or Odonata

and/or Ephemeroptera and/or Coleoptera and/or Platyhelminthes;

3. Plecoptera and/or Trichoptera

Colour

Range of Values based on opacity: 1. Clear; 2. Semi–transparent green; values 1–7 3. Semi–transparent brown; 4. Semi–transparent black or gray;

5. Opaque green (unicellular algae); 6. Opaque brown (due to

ground solutes); 7. Opaque gray (pollution)

Movement

Range of Values according to velocity in cm/s: 1. Stagnant without renewal; values 1–5 2. Stagnant with renewal (laminar flow without moving the entire water

mass); 3. Slow movement < 20 cm/5 s; 4. Fast moving >20 cm/5 s;

5. Fast moving > 20 cm/1 s

Benítez et al.

Table 1. (Cont.)

Factors Variables

Units

Explanations

pH Conductivity

Water temperature ºC Dimensions Length

Width

Depth

Water level

Surface

Pond volume

Water volume

bability values were transformed into geographical favourability values using the equation provided by Real et al. (2006): F = [p/(1–p)] / [(n1/n0) + (p/[1–p])], where n1 is the number of water bodies where the species was found, n0 is the number of water bodies where the species was not found, and p is the geo� graphic probability. The favourability function reflects the degree (between 0 and 1) to which the probability values obtained in each model differ from that expected according to the species’ prevalence, where 0.5 indi� cates no difference between both probability values. Probability depends both on the response of the spe� cies to the predictors and on the overall prevalence of the species, whereas favourability values reflect only the response of the species to the predictors. Geographic favourability values range from 0 indicating a completely unfavourable location for the waterIn a similar manner, an environmental favourability model was obtained for each environmental factor and spe� cies. In this case, conditional forward stepwise logistic regression (p < 0.05 to include a variable, p > 0.1 to exclude a previously included variable) was used to comply with the parsimony principle by not including unnecessary explanatory variables in the models. Wald’s test (1943) was used to assess the relative importance of each variable in the models. Wald’s test relates the coefficient β of each variable with the coefficient of variation of that coefficient β in order to know the importance of each variable in the model. Akaike’s information criterion (Akaike, 1973) was used to test if the model selected in the last step was more parsimonious than models in previous steps.

When performing our models, both type I and type II errors were possible. The risks of these two errors are inversely related and a researcher should determine which error has more severe consequences for the analysed situation. In our case, a type II error entails more severe consequences, as we are trying to identify the characteristics of the water bodies that are relevant for amphibian conservation. The accepta� ble probability of making a type I error is α, which is the level of significance established for a hypothesis test (0.05 in our case). The issue here is that when making multiple tests the type I error increases. To lower this risk, a lower value for α must be used, as the Bonferroni correction, for example, does. However, Bonferroni correction increases enormously the type II error (Pearce & Ferrier, 2000; Nakagawa, 2004). Our approach to deal with both types of error was to increase the power of the test by visiting as many water bodies as feasible, in order to reduce type II error, and to identify a low number of critical factors to test, to reduce type I error. We used 568 sampled water bodies to obtain each model, so that the power of the tests would be high, and grouped the variables into seven factors to reduce the number of models tested per species. Within each model, the stepwise procedure ensures that there is no increase in type I errors at individual steps with the number of variables tested, because at each step only the most significant variable is allowed to enter the model, whereas the significance of the rest of variables is re–analysed in the following step. We avoided excessive multicollinearity by chec� king the variable inflation factor (VIF) and pair–wise variable correlations. Therefore, in variables included in any model the VIF was considered acceptable up

Animal Biodiversity and Conservation 40.1 (2017)

to 10 (Montgomery & Peck, 1992) and Spearman correlation coefficient up to 0.7. The amount of spatial autocorrelation in the variables was assessed with Moran’s I coefficient (Moran, 1950). We mapped the locations of favourable microhabi� tats for each species according to the characteristics related to each factor, including sampling effort. These favourable microhabitats were those with favourabili� ty > 0.5 in each favourability model. Results We found three urodela species (Pleurodeles waltl, Salamandra salamandra longirostris and Triturus pygmaeus) and eight anuran species (Aytes dickhilleni, Discoglossus galganoi jeanneae, Pelobates cultripes, Pelodytes ibericus, Bufo spinosus, Bufo calamita, Hyla meriodionalis and Pelophylax perezi). All correspond to the amphibian species previously recorded in the study area. About 45% of these species are endemic to mainland Spain and the remainder are distributed throughout southwestern Europe and/or northern Africa (Pleguezuelos et al., 2002). Most of the sampled water bodies (78%) contained at least one species and 4% contained three or more (3–7). Microhabitats with three or more species were located in mountain areas, except for one that was located in a coastal natural reserve (fig. 1). Spatial autocorrelation of all variables was very low, with Moran’s I values below 0.1 except for altitude, which had a Moran’s I value of 0.5. Environmental characteristics of the microhabitats significantly favoured the presence of all species except the spadefoot toad (P. cultripes) (table 2; figs. 1s–10s in supplementary material). We thus obtained significant models about fa� vourable geographic locations of the microhabitats for every species. Favourable geographic locations for amphibians were concentrated in two areas, one westward and one central (supplementary material). The western area was favourable to D. galganoi jeanneae, P. ibericus, H. meridionalis, P. perezi, P. walt, S. salamandra longirostris, and T. pygmaeus whereas the central zone was favourable to all spe� cies except S. salamandra longirostris. The western zone includes the medium–altitude mountains of the Serranía de Ronda and those of the Grazalema and Alcornocales Natural Parks; and the central zone includes the mountain ranges of the Sierra Tejeda, Alhama, Almijara, Lújar, and the southern slopes of the Sierra Nevada range. About 85% of water bodies were visited once, but a significant correlation was found between the number of recorded species and the number of visits using Spearman’s correlation (rs= 0.241; P < 0.05). Sampling effort significantly favoured recording the presence of 70% of species (table 2). The location of favourable microhabitats (favourability > 0.5) for each species according to sampling effort are represented in figures 1s–10s in supplementary material. We did not obtain a significant model for each envi� ronmental factor and species. Only for H. meridionalis,

P. perezi and P. waltl we obtained significant char� acterizations according to every environmental factor (table 2). As regards the variables used to measure each factor, the models for H. meridionalis included more variables than those obtained for other species (36.8% of all variables). The fewest characterizations (three models) were obtained for S. salamandra longirostris and these models included the fewest variables (7.9% of all variables). In descending order, the most common environmen� tal factors in the models were typology, surrounding conditions, water condition, external environment, and size of the water point (table 2). Typology entered the models for all the species, and at least one type of natural and one type of artificial microhabitat were relevant for each of them. Surrounding conditions (in terms of connectivity with other microhabitats and conservation status) were relevant for 90% of species (the exception was B. calamita), whereas dimensions of the water point were relevant for only 50% of spe� cies (table 2). Regarding the external environment, higher air temperature was unfavourable for 50% of amphibian species (table 2). Regarding the number of variables, the models for H. meridionalis included more variables than those obtained for other species (36.8% of all variables). The fewest characterizations (three models) were obtained for S. salamandra longirostris and these models included the fewest variables (7.9% of all variables) (table 2). Discussion The richness of amphibian species in the study area follows a longitudinal gradient, with the highest number of species in the wetter west (CAPMA, 2012). Howe� ver, we found that favourable geographic locations for amphibians were concentrated in two mountainous areas, one western and one central (fig. 1). This may be because the rest of the study area has been subject to more change and degradation from human activities, agricultural activities and urban development, while the eastern zone contains few natural water points due to its more arid climate. In the arid eastward zone, water bodies were only favourable to B. calamita and P. perezi (table 2). It could be because B. calamita reproduces in any ponds with a short hydroperiod (Reques & Tejedo, 2002), while P. perezi is able to use even abundant, agricultural plastic–lined man–made ponds (Llorente et al., 2002). The central area of the Guadalhorce River and the coastline of Malaga and Cadiz used to host numerous streams and wetlands with optimal characteristics for amphibians (Real et al., 1993), but many of them have disappeared due to urban, agricultural, and industrial development. Hyla meridionalis and P. cultripes are the species most affected by this process, although the geographical favourability for H. meridionalis still suggests a coastal distribution. However, Pelobates cultripes is in decline in the study area (Benítez et al., 2012), because the sandy and loose soils needed by the species to burrow in have become increasingly scarce.

Benítez et al.

Table 2. Variables included in each environmental favourability model with the signs of the coefficients (β), + or –: Ad. Alytes dickhilleni; Dgj. Discoglossus galganoi jeanneae; Pi. Pelodytes ibericus; Bs. Bufo spinosus; Bc. Bufo calamita; Hm. Hyla meridionalis; Pp. Pelophylax perezi; Pw. Pleurodeles waltl; Ssl. Salamandra salamandra longirostris; Tp. Triturus pygmaeus.

Air temperature Altitude Wind Connectivity Conservation River Stream Small wetland Spring–fed river Spring Mine water Seepage Temporary pond Natural pond Fountain Drinking trough Plastic–lined pond Earth–lined pond Concrete pool Cistern Irrigation ditch Well Dike or Levee Vegetation Macroinvertebrates Colour Movement pH Conductivity Water temperature Length Width Depth Water level Surface Pond volume Water volume

Tabla 2. Variables incluidas en cada modelo de favorabilidad ambiental con el signo del coeficiente (β), + o –. (Para consultar las abreviaturas de las especies, véase arriba.)

Ad – + + – – – Dgj + + + + + + + – Pi + + + + Bs + + + + + – – Bc

– + + + +

Hm – – – + + – + + + + + + + Pp – + + + + + + – + + + Pw – + + + + + Ssl + + + Tp

– + + + + –

The study confirmed the importance of the type of aquatic habitat for amphibian species (Calhoun & Hunter, 2003; Zacharias & Zamparas, 2010). Wells entered into 60% of the models and were the most frequent typology. These structures were essential to anuran and urodele reproduction and provided shelter during dry periods (Lanz & Greenpeace España, 1997). The second most frequent aquatic microhabitat in the models was the earth–lined pond (table 2), commonly used by fauna and specifically by amphibians in the south–eastern Iberian peninsula (García–Muñoz et al., 2010). Regarding the variables that characterized water bodies, temperature was a frequent variable in the models (appearing in 50% of them), with low values being more favourable for most of the species (1–10ºC). This could be due to their nocturnal behaviour and to their location in mountain zones (Fig. 1) where there is less evaporation and therefore water bodies last longer. Our results also support the importance of connectiv� ity between microhabitats (Tabla 2). The connection between water bodies should be maintained to allow gene flow between the different populations (Stevens et al., 2006), increasing the diversity of species (Seml�

itsch & Bobie, 1998). Connectivity was more important for species that were more closely associated with water, such as D. galganoi jeanneae, P. ibericus, H. meridionalis, P. perezi, P. waltl and T. pygmaeus. Water bodies wererather isolated in some eastern watersheds (Almanzora River, Tabernas Desert, and Campo de Níjar). We suggest that management measures in these areas should be undertaken to ensure the long–term conservation of amphibians. Some authors consider the water body size of microhabitats to be an important factor for the coloni� zation of new water bodies by amphibians (Semlitsch & Bobie, 1998). However, in this study, water body size was only significant for 50% of the species, i.e., D. galganoi jeanneae, H. meridionalis, P. perezi, and P. walt, probably because they are the most aquatic species and select large water bodies that are more favourable to their biological activity. Regarding water conditions, vegetation and colour were the variables most frequently included in the models. This could be because vegetation is essential for P. ibericus, T. pygmaeus, and B. spinosus to lay their eggs (García–París, 2004; Montori & Herrero, 2004), and because H. meridionalis needs helophytic vegetation

Animal Biodiversity and Conservation 40.1 (2017)

for its biological activity (Díaz–Paniagua, 1986). The intensity of colour was important for P. walt and T. pygmaeus, possibly because turbid water is a refuge from predators for these species. Detectability of individuals differs among species, since some of them are conspicuous due to habits such as jumping into the water or singing, whereas others are more cryptic (De Solla et al., 2005). Some authors propose different methodologies in order to increase the detection probability of amphibians (Mazerolle et al., 2007; Gomez–Rodriguez et al., 2012). However, these methods were not feasible in our study due to the heterogeneity of the territory and water bodies. For this reason, we scheduled our visits according to the phenology of each species in each area to maximize detectability and 85.6% of water bodies were visited only once. However, when the microhabitat was visited in the appropriate phenologi� cal period for one species but not for others, the water body was revisited, thereby increasing the likelihood of finding more species. These multiple visits revealed additional species in some microhabitats, but not in others (supplementary material). Regarding the species individually, microhabitats favourable for A. dickhilleni were characterized by their good conservation status, low temperatures in air and water, and higher altitudes, indicating its ad� aptation to mountain areas where habitat protection is also greater. Discoglossus galganoi jeanneae has been reported to reproduce in any shallow water with helophytic vegetation where females can lay eggs (García–París, 2004). Natural microhabitats such as small streams or temporary ponds or springs were more favourable to this species than artificial ones, though the species was also found in some artificial microhabitats, such as old wells. It was the only spe� cies favoured by wind, possibly because selecting windy places may avoid competition with species such as P. perezi and H. meridionales, which make greater use of calls and are therefore more affected by frequent strong winds (table 2). The most important characteristics of favourable microhabitats for B. spinosus were the presence of vegetation, transparent water, and low conductivity. In contrast, sites with a large surface area but little volume were selected by the other congeneric spe� cies, B. calamita, which preferred temporary ponds. This typology is associated with the more frequent climate fluctuations to which this species is better adapted (Romero & Real, 1996). Hyla meridionalis and P. perezi are closely as� sociated with large masses of water and selected water bodies with large dimensions and with high connectivity (table 2). The water bodies that were most favourable to H. meridionalis were temporary ponds and natural ponds with surrounding vegetation (Sillero, 2009), whereas artificial ponds were more favourable to P. perezi (García–Muñoz et al., 2010). Although P. perezi is considered a generalist species (Llorente et al., 2002) that can adapt to conditions unfavourable to the other amphibian species and that can be present at almost any water body, our results show that this species is favoured by certain types

of microhabitats, particularly rivers, streams, plastic– lined pools and concrete pools. In addition, over 50% of characteristics of the P. perezi microhabitats, such as high water temperature, high pH, and width of the site, were different from those selected by other species. Thus, the models for this species indicated their preference for sites at a medium or low altitude with warm temperatures. This could be due to their preference for basking on aquatic vegetation or banks during the day, or to the fact that the reproduction of this species correlates positively to ambient tempera� ture (Richter–Boix et al., 2006). The selection of sites of a certain width could be due to their behaviour of jumping away from threats (García–París, 2004) or to the territorial disputes between males for places to call (Díaz–Paniagua et al., 2005). Microhabitats favourable to P. waltl and T. pygmaeus shared certain characteristics such as low temperature and connectivity and the same type of habitat: natural ponds and wells. These species are nocturnal, repro� duce in natural ponds, and hide in damp places during the terrestrial phase (Montori & Herrero, 2004). Regarding S. salamandra longirostris, fewer varia� bles entered the models and only shared the conser� vation status of the microhabitat with other species. This was the only species that selected the presence of macroinvertebrates and fountains as a habitat; both these variables are related to the high–quality water needed by salamanders. Their larvae prefer permanent bodies of water (Baumgartner et al., 1999), but this species occupies fountains or ponds in the south of the Iberian peninsula (table 2), where water availability is lower (Egea–Serrano et al., 2006). Conclusion This study identified some of the geographical and ecological characteristics to take into account to maintain the conservation value of small water bodies for one of the most endangered vertebrate groups: amphibians (Beebee & Griffiths, 2005). The combination of high sampling field effort, the on–site characterization of water bodies, and modelling tools is a useful and applicable methodology. The results suggest that typology and surrounding conditions of the water bodies are critical for constituting a breeding habitat for the amphibian species that inhabit the study area. The main results of this study emphasize the importance of the typology of aquatic microhabitats and the need for connectivity between them. These results should be used to develop management tools to regulate land use, making it more compatible with the conservation of amphibians (Scoccianti, 2001), especially in the most isolated habitats in medium and high mountain areas. In these areas, it is necessary to maintain existing water bodies and even to create new artificial bodies that would maintain and improve the situation of amphibians in the southern basin. Finally, coastal wetlands should be protected by encouraging moderate urban development that is more compatible with the environment and that respects the breeding sites of amphibians and their shelters.

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Supplementary material Distribution of each species (A) and location of their favourable microhabitats (favourability > 0.5) according to the predictive factors (B to H). Distribución de cada especie (A) y ubicación de sus microhábitats favorables (favorabilidad > 0,5) en función de los factores de predicción (de B a H).

Alytes dickhilleni

A 0 25 50

100 km

Fig. 1s. Presence of Alytes dickhilleni in the sampled microhabitats (A), and favourable microhabitats (F > 0.5) according to the factors that characterize them significantly: B. Geographical location; C. Number of visits; D. External environment; E. Surrounding conditions; F. Typology; G. Water conditions. Fig. 1s. Presencia de Alytes dickhilleni en los microhábitats muestreados (A) y microhábitats favorables (F > 0,5) en función de los factores que los caracterizan de forma significativa: B. Localización geográfica; C. Número de visitas; D. Entorno externo; E. Condiciones del entorno; F. Tipología; G. Condiciones de agua.

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Supplementary material. (Cont.)

Discoglossus galganoi jeanneae

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100 km

Fig. 2s. Presence of Discoglossus galganoi jeanneae in the sampled microhabitats (A), and favourable microhabitats (F > 0.5) according to the factors that characterize them significantly: B. Geographical location; C. Number of visits; D. External environment; E. Surrounding conditions; F. Typology; H. Dimensions of microhabitat. Fig. 2s. Presencia de Discoglossus galganoi jeanneae en los microhábitats muestreados (A) y microhábitats favorables (F > 0,5) en función de los factores que los caracterizan de forma significativa: B. Localización geográfica; C. Número de visitas; D. Entorno externo; E. Condiciones circundantes; F. Tipología; H. Dimensiones de microhabitat..

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iii

Supplementary material. (Cont.)

Pelodytes ibericus

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100 km

Fig. 3s. Presence of Pelodytes ibericus in the sampled microhabitats (A), and favourable microhabitats (F > 0.5) according to the factors that characterize them significantly: B. Geographical location; C. Number of visits; E. Surrounding conditions; F. Typology; G. Water conditions. Fig. 3s. Presencia de Pelodytes ibericus en los microhábitats muestreados (A) y microhábitats favorables (F > 0,5) en función de los factores que los caracterizan de forma significativa: B. Localización geográfica; C. Número de visitas; E. Condiciones del entorno; F. Tipología; G. Condiciones de agua.

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Supplementary material. (Cont.)

Bufo spinosus

0 25 50

100 km

Fig. 4s. Presence of Bufo spinosus in the sampled microhabitats (A), and favourable microhabitats (F > 0.5) according to the factors that characterize them significantly: B. Geographical location; C. Number of visits; D. External environment; E. Surrounding conditions; F. Typology; G. Water conditions. Fig. 4s. Presencia de Bufo spinosus en los microhábitats muestreados (A) y microhábitats favorables (F > 0,5) en función de los factores que los caracterizan de forma significativa: B. Localización geográfica; C. Número de visitas; D. Entorno externo; E. Condiciones del entorno; F. Tipología; G. Condiciones de agua.

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Supplementary material. (Cont.)

Bufo calamita

0 25 50

100 km

Fig. 5s. Presence of Bufo calamita in the sampled microhabitats (A), and favourable microhabitats (F > 0.5) accorading to the factors that characterize them significantly: B. Geographical location; C. Number of visits; D. External environment; F. Typology; H. Dimensions of microhabitat. Fig. 5s. Presencia de Bufo calamita en los microhábitats muestreados (A) y microhábitats favorables (F > 0,5) en función de los factores que los caracterizan de forma significativa: B. Localización geográfica; C. Número de visitas; D. Entorno externo; F. Tipología; H. Dimensiones de microhabitat.

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Supplementary material. (Cont.)

Hyla meridionalis

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100 km

Fig. 6s. Presence of Hyla meridionalis in the sampled microhabitats (A), and favourable microhabitats (F > 0.5) according to the factors that characterize them significantly: B. Geographical location; C. Number of visits; D. External environment; E. Surrounding conditions; F. Typology; G. Water conditions; H. Dimensions of microhabitat. Fig. 6s. Presencia de Hyla meridionalis en los microhábitats muestreados (A) y microhábitats favorables (F > 0,5) en función de los factores que los caracterizan de forma significativa: B. Localización geográfica; C. Número de visitas; D. Entorno externo; E. Condiciones del entorno; F. Tipología; G. Condiciones de agua; H. Dimensiones de microhabitat.

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Supplementary material. (Cont.)

Pelophylax perezi

0 25 50

100 km

Fig. 7s. Presence of Pelophylax perezi in the sampled microhabitats (A), and favourable microhabitats (F > 0.5) according to the factors that characterize them significantly: B. Geographical location; D. External environment; E. Surrounding conditions; F. Typology; G. Water conditions; H. Dimensions of microhabitat. Fig. 7s. Presencia de Pelophylax perezi en los microhábitats muestreados (A) y microhábitats favorables (F > 0,5) en función de los factores que los caracterizan de forma significativa: B. Localización geográfica; D. Entorno externo; E. Condiciones del entorno; F. Tipología; G. Condiciones de agua; H. Dimensiones de microhabitat.

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Supplementary material. (Cont.)

Pleurodeles waltl

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100 km

Fig. 8s. Presence of Pleurodeles waltl in the sampled microhabitats (A), and favourable microhabitats (F > 0.5) according to the factors that characterize them significantly: B. Geographical location; C. Number of visits; D. External environment; E. Surrounding conditions; F. Typology; G. Water conditions; H. Dimensions of microhabitat. Fig. 8s. Presencia de Pleurodeles waltl en los microhábitats muestreados (A) y microhábitats favorables (F > 0,5) en función de los factores que los caracterizan de forma significativa: B. Localización geográfica; C. Número de visitas; D. Entorno externo; E. Condiciones del entorno; F. Tipología; G. Condiciones de agua; H. Dimensiones de microhabitat.

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Supplementary material. (Cont.)

Salamandra salamandra longirostris

0 25 50

100 km

Fig. 9s. Presence of Salamandra salamandra longirostris in the sampled microhabitats (A), and favourable microhabitats (F > 0.5) according to the factors that characterize them significantly: B. Geographical location; E. Surrounding conditions; F. Typology; G. Water conditions. Fig. 9s. Presencia de Salamandra salamandra longirostris en los microhábitats muestreados (A) y microhábitats favorables (F > 0,5) en función de los factores que los caracterizan de forma significativa: B. Localización geográfica; E. Condiciones del entorno; F. Tipología; G. Condiciones de agua.

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Supplementary material. (Cont.)

Triturus pygmaeus

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100 km

Fig. 10s. Presence of Triturus pygmaeus in the sampled microhabitats (A), and favourable microhabitats (F > 0.5) according to the factors that characterize them significantly: B. Geographical location; D. External environment; E. Surrounding conditions; F. Typology; G. Water conditions. Fig. 10s. Presencia de Triturus pygmaeus en los microhábitats muestreados (A) y microhábitats favorables (F > 0,5) en función de los factores que los caracterizan de forma significativa: B. Localización geográfica; D. Entorno externo; E. Condiciones del entorno; F. Tipología; G. Condiciones de agua.

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Description of Silvinichthys pedernalensis n. sp. (Teleostei, Siluriformes) from the Andean Cordillera of southern South America L. Fernández, E. A. Sanabria & L. B. Quiroga

Fernández, L., Sanabria, E. A. & Quiroga, L. B., 2017. Description of Silvinichthys pedernalensis n. sp. (Teleostei, Siluriformes) from the Andean Cordillera of southern South America. Animal Biodiversity and Conservation, 40.1: 41–47. Abstract Description of Silvinichthys pedernalensis n. sp. (Teleostei, Siluriformes) from the Andean Cordillera of southern South America.— Silvinichthys pedernalensis, a new species, is described from an Andean stream in Provincia San Juan, Argentina, based on its coloration pattern, and its meristic, morphometric and osteological characters. S. pedernalensis differs markedly from all other known members of the genus Silvinichthys as a result of the combination of the absence of pelvic girdle and fin, the wide supraorbital bone, the number of interopercle odontodes 14–18, the number of opercular odontodes 6–8, the branched pectoral–fin rays 7, the dorsal–fin rays 11, the number of dorsal pterygiophore 7, the branchiostegal rays 6, the dorsal procurrent caudal–fin rays 14 and ventral 15, the ribs 16, the vertebrae 39, the dark marmorated pigmentation on the body and head, the head depth 9.9–12.2% SL, the interorbital wide 28.3–36.1% HL, the nasal barbel length 27.3–39.0% SL, the maxillary barbel length 39.5–61.7% SL, the submaxillary barbel length 24.7–41.9% SL, the snout length 40.6–44.4% HL, the body depth 10.1–12.6% SL, the anal base fin 10.2–11.7% SL, and the caudal peduncle length 19.3–21.5% SL. Key words: Neotropical, Catfish, Trichomycteridae, Silvinichthys pedernalensis, New species Resumen Descripción de Silvinichthys pedernalensis sp. n. (Teleostei, Siluriformes) de la cordillera de los Andes en la parte meridional de Sudamérica.— Se describe una nueva especie, Silvinichthys pedernalensis, en un arroyo andino de la provincia de San Juan, en Argentina, a partir del patrón de coloración y caracteres merísticos, morfométricos y osteológicos. S. pedernalensis difiere notablemente de todos los demás miembros conocidos del género Silvinichthys debido a la combinación de los siguientes rasgos: ausencia de cintura y aleta pélvica, hueso supraorbital ancho, 14–18 odontoides interoperculares, 6–8 odontoides operculares, 7 radios ramificados de la aleta pectoral, 11 radios de la aleta dorsal, 7 pterigióforos de la aleta dorsal, 6 radios branquióstegos, 14 radios dorsales procurrentes de la aleta caudal y 15 ventrales, 16 costillas, 39 vértebras, pigmentación marmórea oscura de la cabeza y el cuerpo, altura de la cabeza (9,9–12,2% de la longitud estándar [LE]), ancho interorbital (28,3–36,1% de la longitud de la cabeza [LC]), longitud de la barbilla nasal (27,3–39,0% LE), longitud de la barbilla maxilar (39,5–61,7% LE), longitud de la barbilla submaxilar (24,7–41,9% LE), longitud del hocico (40,6–44,4% LC), altura del cuerpo (10,1–12,6% LE), ancho de la aleta anal (10,2–11,7% LE) y longitud del pedúnculo caudal (19,3–21,5% LE). Palabras claves: Neotropical, Bagre, Trichomycteridae, Silvinichthys pedernalensis, Nueva especie Received: 5 II 16; Conditional acceptance: 26 V 16; Final acceptance: 6 IX 16 L. Fernández, CONICET–IBN Fundación Miguel Lillo, CP 4000 Tucumán; FACEN, Univ. Nacional de Catamarca, Belgrano 300, 4700 Catamarca, Argentine.– E. A. Sanabria & L. B. Quiroga, CONICET, Inst. de Ciencias Básicas, Fac. de Filosofía Humanidades y Artes, Univ. Nacional de San Juan, Av. José Ignacio de la Roza 230, CP 5400 San Juan, Argentine.– E. A. Sanabria, Fac. de Ciencias Exactas y Naturales, Univ. Nacional de Cuyo, Padre Contreras 1300, CP 5500 Mendoza, Argentine. Corresponding author: L. Fernández. E–mail: luis1813@yahoo.com ISSN: 1578–665 X eISSN: 2014–928 X

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Introduction Silvinichthys is the second most speciose genus of the Trichomycterinae (exclusive of Ituglanis and Scleronema, which are currently unassigned to subfamily (Costa & Bockmann, 1993; de Pinna 1998; Fernández & de Pinna, 2005) with five species recognized to date and two undescribed species. The genus Silvinichthys was erected by Arratia (1998) for a species previously placed in Trichomycterus (T. mendozensis Arratia et al., 1978) mainly based on the reduction of the cephalic laterosensory canal system to the nasal portion of the supraorbital canal and the postotic canal and the entire skin surface perforated by pores of the ampullary organs. In later years, four new species of Silvinichthys have been described (Fernández et al., 2011, 2013, 2014) and other known new species await formal description. The genus inhabits headwaters and temporary endorrheic streams, characterized by cold waters and rocky bottom, and included one species from phreatic waters. Silvinichthys shows a restricted distribution between 24°S to 32°S latitude in the western of Argentina and it is endemic to this arid region (Fernández et al., 2014). We describe here a sixth species of Silvinichthys, the fifth to lack the pelvic girdle from a mid–elevation location in western Argentina. Material and methods Measurements were taken from the left side of the specimens using digital calipers under a binocular microscope following the methods outlined by Tchernavin (1944). Cleared and counterstained specimens were prepared following the procedure of Taylor & Van Dyke (1985) and osteological nomenclature follows de Pinna (1989). Counts of dorsal and anal fin rays follow the methods proposed by de Pinna (1992) and taken from radiographs and cleared and stained specimens. Meristic values are followed by the number of specimens with that count in brackets; meristic values for the holotype in the text are indicated (*). Following de Pinna (1992), the vertebral counts exclude the vertebrae in the Weberian apparatus, with the compound caudal centrum counted as one element. Counts of caudal vertebrae follow Fernández & Schaefer (2003) with counts of vertebrae and ribs taken from one cleared and stained specimen. The numbering system and terminology for laterosensory pores of the head follow Northcutt (1989). Institutional abbreviations are as listed at Sabaj Perez (2014), with the addition of FACEN, Facultad Ciencias Exactas y Naturales, Universidad Nacional de Catamarca, Catamarca; FLBS, Flathead Lake Biological Station, Poulson; and MPSZI, Museo de Ciencias Naturales 'P. Antonio Scasso', San Nicolás de los Arroyos, Buenos Aires. Abbreviations are head length (HL) and standard length (SL). Comparative material examined Additional material is that cited in Fernández & Vari (2009) and Schaefer & Fernández (2009). The number

of specimens indicated refers to those examined for this study, not necessarily to the total number in lot. Abbreviations are: number specimens (ex), holotype (h), paratypes (p), cleared and stained specimens (cs), radiographed specimen (r): Hatcheria macraei: FACEN 0012 , 1 ex; FACEN 0014, 1 ex; MCNI 1521, 1 ex, 1 cs; MCNI 1561, 2 ex. Silvinichthys bortayro: AMNH 233621, 1 p; FACEN 0040, 1 ex. Silvinichthys gualcamayo: MCNI 1531, 5 p; MCNI 1532, 1 cs p. Silvinichthys huachi: MCNI 1515, 2 cs p; MCNI 1517, 4 p, 3 r. Silvinichthys leoncitensis: MCNI 1511, 1 h; MCNI 1512, 1 p, 1 cs; ILPLA 2171, 1 p. Silvinichthys mendozensis: FACEN 0078, 2 ex, 1 cs. Silvinichthys sp. A: MPSZI 1381, 1 ex; MPSZI 1382, 1 ex; ILPLA 1807, 1 ex. Trichomycterus alterus: FACEN 35, 8 ex; FML 2085, 9 ex, 1 cs. Trichomycterus areolatus: MCNI 1370, 1 ex. Trichomycterus barbouri: MCNI 0048, 3 ex; MCNI 1163, 6 ex; MCNI 1178, 3 ex; MCNI 1376, 1 ex. Trichomycterus belensis: FML 2531, 4 p, 1 cs; FACEN 0068, 1 ex; FACEN 0082, 3 ex. Trichomycterus boylei: MCNI 0795, 2 ex. Trichomycterus catamarcensis: FACEN 0069, 1 ex; FACEN 0083, 3 ex; FML 2510, 4 ex, 1 cs. Trichomycterus corduvensis: MCNI 1530, 4 ex; MCNI 1372, 4 ex; MCNI 1375, 1 ex; UNCa 66, 4 ex. Trichomycterus hualco: FML 2602, 1 p, 1 cs; USNM 383794, 4 p. Trichomycterus minus: MCNI 1529, 1 p, 1 cs. Trichomycterus pseudosilvinichthys: FML 2589, 4 p, 1 cs. Trichomycterus ramosus: FML 2071, 4 p, 1 cs. Trichomycterus roigi: MCNI 0757 2 ex; MCNI 0994, 5 ex. Trichomycterus spegaziinii: FACEN 0067, 1 cs; MCNI 0321, 3 ex; MCNI 0356, 5 ex; MCNI 0815, 1 ex Trichomycterus yuska: FML 1132, 4 p, 2 cs. Results Silvinichthys pedernalensis n. sp. (fig. 1, table 1) Holotype: FACEN 0071, 45.1 mm SL; Argentina, Provincia de San Juan, Departamento Sarmiento, Río Pedernal (31° 59’ S, 68° 44’ W), 1.092 m elevation, collected by L. Fernández, E. Sanabria, and L. Quiroga, 20 VII 2013. Paratypes: three specimens, 37.5–43.2 mm SL, collected with holotype: FACEN 0072, 2 specimens, 37.5–43.2 mm SL; FACEN 0073, 1 specimen, 42.7 mm SL CS. Diagnosis Silvinichthys pedernalensis is distinguished from S. mendozensis by the absence of pelvic girdle and fin (versus presence), the number of interopercle odontodes 14–18 (vs. 30–42), the marmorated pigmentation

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Fig. 1. Silvinichthys pedernalensis n. sp., holotype, FACEN 0071, 45.1 mm SL. Fig. 1. Silvinichthys pedernalensis sp. n., holotipo, FACEN 0071, 45,1 mm de LE.

on the body and head (vs. uniformly brown); from S. bortayro by the number of branched pectoral–fin rays 7 (vs. 5), the number of total dorsal–fin rays 11 (vs. 9), the anal base fin 10.2–11.7% SL (vs. 8.8–10.2), the supraorbital tendon bone wide (vs. slender), the dark marmorated pigmentation on the body and head (vs. the lack of dark pigmentation in larger individuals), the nasal barbel length 27.3–39.0% SL (vs. 47.1–74.4), the maxillary barbel length 39.5–61.7% SL (vs. 60.5–105.9), the submaxillary barbel length 24.7–41.9% SL (vs. 41.2–57.1), the snout length 40.6–44.4% HL (vs. 35.8–40.5), the number of opercular odontodes 6–8 (vs. 2–4), the number of interopercle odontodes 14–18 (vs. 9–12); from S. huachi by the body depth 10.1–12.6% SL (vs. 12.6–16.5), the supraorbital tendon bone wide (vs. slender), the number of interopercle odontodes 14–18 (vs. 21–28), the number of branchiostegal rays 6 (vs. 7–8), the number of dorsal procurrent caudal–fin rays 14 (vs. 11), the number of ventral procurrent caudal–fin rays 15 (vs. 10); from S. gualcamayo by the caudal peduncle length 19.3–21.5% SL (vs. 21.1–23.6), the head depth 9.9–12.2% SL (vs. 9.1–9.8), the interorbital wide 28.3–36.1% HL (vs. 26.1–27.8), the number of branchiostegal rays 6 (vs. 7), the number of dorsal procurrent caudal–fin rays 14 (vs. 11), the number of ventral procurrent caudal–fin rays 15 (vs. 9), and

the number of total vertebrae 39 (vs. 38); from S. leoncitensis by the total number of dorsal–fin rays 11 (vs. 13), the number of dorsal pterygiophore 7 (vs. 8), the number of ribs 16 (vs. 20), the number of total vertebrae 39 (vs. 40), and the number of interopercle odontodes 14–18 (vs. 18–28). Description Table 1 shows the morphometrics for holotype and paratypes of Silvinichthys pedernalensis. Body elongate, approximately cylindrical overall and gradually becoming more compressed transversely across the entire vertical extent of the body towards the caudal fin. Dorsal and ventral profiles of trunk region are nearly straight. Caudal peduncle smoothly continuous with dorsal and ventral profiles of trunk. Papillae–like structures absent on body. Urogenital and anal openings vertical through base of first or second branched dorsal–fin rays. Head profile nearly triangular from dorsal view, slightly longer than broad. Eye circular located on dorsal surface of head but visible from lateral view. Skin covering eye thin, transparent and separate from surface of eyeball. Anterior nostril slightly smaller than posterior nostril and bordered medially by fleshy flap and laterally by base of nasal barbel. Flap and base

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Table I. Morphometric data for Silvinichthys pedernalensis. Measurements are based on the holotype (FACEN 0071) and three paratypes (FACEN 0072, FACEN 0073). Tabla 1. Datos morfométricos de Silvinichthys pedernalensis. Mediciones basadas en el holotipo (FACEN 0071) y en tres paratipos (FACEN 0072 y FACEN 0073). Range LS (mm)

Mean

37.5–45.1 42.1

Percentage of LS Body depth

10.1–12.6 11.4

Caudal peduncle length 19.3–21.5 20.6 Caudal peduncle depth 7.8–9.0

8.5

Predorsal length

64.2–67.4 65.9

Preanal length

67.4–70.0 68.5

Dorsal–fin base length 10.3–13.4 11.9 Anal–fin base length

10.2– 11.7 10.7

Head length

16.5–18.1 17.5

Head width

15.0–17.6 16.3

Head depth

9.1–9.8

9.1

Percentage of HL Interorbital width

26.1–27.8 27.2

Snout length

40.6–44.4 42.0

Nasal barbel length

27.3–39.0 33.2

Maxillary barbel length 39.5–61.7 47.9 Rictal barbel length

24.7–41.9 34.5

Mouth width

33.9–42.5 39.0

of barbel continuous and forming short tube. Posterior nostril located approximately midway between anterior nostril and anterior orbital rim. Anterior margin of posterior nostril bordered by flap of thin skin. Infraorbital canal absent. Supraorbital canal incomplete, with segment between pores s1 and s2 present. Preoperculomandibular sensory canal absent. Postotic canal with two pores, with pterotic branch present at junction of pterotic and posttemporo–supracleithrum. Laterosensory canal along midlateral portion of trunk reduced, with three pores on anterior most portion of lateral line, with single terminal pore opening situated slightly posterior to posterior tip of opercular patch of odontodes. Mouth subterminal, with rictus directed posteriorly. Mesethmoid T–shaped, elongate, with anterior margin straight and shaft slightly smaller than lateral cornua, its posterior process extending between anterior portions of frontals, lateral ethmoids, and vomer.

Premaxilla rectangular and approximately equal in size to maxilla from ventral view. Premaxilla bearing 3 or 4 rows of teeth. Outer premaxillary tooth row with 8–9 teeth and total of 20 to 25 teeth. Maxilla enlarged, L–shaped, with pair of condyles, projecting between premaxilla and anterior border of autopalatine. Supraorbital tendon bone (= frontolachrymal or sesamoid supraorbital) wide. Anterior portion of sphenotic laterally directed in dorsal view. Autopalatine rectangular, broad anteriorly with short posterior process dorsally placed to broad metapterygoid. Medially, autopalatine articulates with both vomer and lateral ethmoid. Dentary with 3 rows of teeth, with 9 teeth in outer row. Lower lip fleshy anteriorly with anterior, and to a lesser degree, anteroventral surfaces covered with papillae. Lower lip with prominent lobes along lateral limits. Upper lip fleshy and bearing numerous papillae. Barbels relatively short and tapering distally, but not thread–like or with distal branching. Tip of maxillary barbel falling short of vertical through anterior limit of patch of opercular odontodes in some specimens but extending somewhat posterior of that point in other individuals. Submaxillary barbel shorter than maxillary barbel and falling short of vertical through anterior limit of opercular patch of odontodes. Nasal barbel extending posteriorly distinctly beyond posterior margin of eye. Branchiostegal membrane narrowly attached to isthmus anteriorly at midline, with wide and almost free branchial opening. Branchiostegal rays 6 in one cs specimen. Interopercular odontode patch elongate, straight and bearing 14 to 16 odontodes and 18 odontodes present in one cs specimen. Interopercular odontodes patch with maximum of 3 irregular rows. Opercular odontode patch small and rounded; odontodes straight overall. Opercular odontode patch bearing up to 6 odontodes arranged in up to 2 irregular rows and 8 odontodes present in one cs specimen; odontodes not imbedded in fleshy tissue covering of opercle. Dorsal–fin rays obvious in whole specimens 11 [4], with 4 unbranched rays and 7 branched rays, including one cs specimen. Pterygiophores 7. Dorsal–fin fleshy basally. Distal margin of dorsal fin semicircular in expanded fin. Dorsal–fin origin located distinctly anterior to vertical through anterior limit of vent. First proximal dorsal–fin pterygiophore inserting posterior to neural spine of vertebra 23. Anal–fin rays 10 [2] or 11* [1] with 4 [2] or 5* [1] unbranched rays and 6 [3] branched rays. Total of 11 rays in one cs specimen, with 5 unbranched rays and 6 branched rays. Pterygiophores 6. Anal–fin fleshy basally. Anal–fin relatively elongate; equal in size to, or slightly smaller than, dorsal fin with distal margin slightly rounded. Anal–fin origin located approximately at vertical through posterior portion of dorsal–fin base. First proximal anal–fin pterygiophore inserting posterior to haemal spine of vertebra 24. Dorsal–fin base either terminating at vertical through insertion of anal fin or overlapping anal–fin base for distance of up to 2 vertebrae. Pectoral–fin rays 8 [4], with lateral–most ray unbranched, including one cs specimen. Distal margin of pectoral–fin straight to slightly convex. First pectoral–fin ray terminating at

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–

n¢

•

400

800 km

Fig. 2. Map of Argentina showing distribution of known Silvinichthys species. 1, S. gualcamayo (l); 2. S. leoncitensis (n); 3. S. mendozensis (•); 4. S. bortayro («); 5. S. huachi (–); 6. S. pedernalensis (¢). Fig. 2. Mapa de Argentina en el que se muestra la distribución de las especies conocidas de Silvinichthys: 1. S. gualcamayo (l); 2. S. leoncitensis (n); 3. S. mendozensis (•); 4. S. bortayro («); 5. S. huachi (–); 6. S. pedernalensis (¢).

fin margin without forming distal filament. Pelvic–fin, girdle and muscles absent. Distal margin of caudal fin nearly straight or slightly convex. Principal caudal–fin rays 6 + 7 [3]. Three principal dorsal caudal–fin rays attaching to fused fourth plus fifth hypurals and 3 rays attaching to third hypural. Seven principal ventral caudal–fin rays attaching to fused hypurals 1–2 and separate parahypurals. Dorsal procurrent caudal–fin rays 14 and ventral procurrent caudal–fin rays 15. Total vertebrae 39, with 7 precaudal and 32 caudal vertebrae. Ribs on each side of body 16. No externally obvious sexual dimorphism present in examined population samples. All specimens with cysts on head, body, and fins (fig. 1). Color in alcohol Head and body with distinct marmoration formed by patches of small, dark chromatophores. Ventral surface of head from hyaline to slightly darkly pigmented. All barbels except by submaxillary barbel, with diffuse pattern of scattered dark chromatophores. Dorsal, anal, and pectoral fins with irregular, dark pigmentation on rays and membranes usually more intense along rays. Variation in intensity of dark dorsal–fin pigmentation sometimes in form of indistinct transverse bar. Caudal–fin membranes irregularly darker than those

of dorsal and anal fins. Pectoral–fin hyaline to slightly dusky ventrally, with irregular dark pigmentation on dorsal surface that becomes less intense distally. Opercular and interopercular odontodes and oral dentition unpigmented. Opercular, but not interopercular, patch of odontodes with web–like pattern of dark pigmentation around base of odontodes. Color in life Head and body with dark pigmentation as described for preserved specimens in alcohol, but with marmoration pattern more pronounced. Dorsal surface of head and body with slight yellowish coloration. Ventral surface of body silvery from isthmus to anus. Distribution and habitat Silvinichthys pedernalensis is known only from Río Pedernal (31° 59' S, 68° 44' W) in San Juan, Argentina (fig. 2). The type locality is a small creek, approximately 0.50 m deep and 1 to 3 m wide with silt in suspension, rock bottom without aquatic vegetation (fig. 3) at an elevation of 1,092 m a.s.l. The drainage lies within an endorheic system that experiences torrential hydrological conditions associated with scarce but intense summer rains. All captured specimens were hiding under

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Fig. 3. Type locality of Silvinichthys pedernalensis. Río Pedernal, Sarmiento, Provincia San Juan, Argentina. Fig. 3. Localidad tipo de Silvinichthys pedernalensis. Río Pedernal, Sarmiento, Provincia de San Juan, Argentina.

rocks; the usual habit for fishes in other streams in the Andes. The one other species of fish collected at that site was Hatcheria macraei (Siluriformes, Trichomycteridae). The Río Pedernal is impacted by limestone mining operations. Many Andean drainage systems are altered by mining activities, including mountain mining/ valley fill practices primarily for extraction of minerals. It is difficult to provide reliable conservation recommendations for Andean catfishes, mainly because data are deficient as geographic distributions are still poorly known; in many cases data are restricted to streams and access is difficult. Etymology The specific name, pedernalensis, is in reference to the type locality of the species, the Río Pedernal. A noun in apposition. Discussion The new species is a member of Silvinichthys, diagnosed by five synapomorphies: the perforation of the entire skin surface by the pores of the ampullary or-

gans; the reduction of the laterosensory canal system, with the posterior region of that system on the head reduced to the postotic portion (pores p1–p2) and the nasal portion of the supraorbital canal (pores s1–s2); the narrow and elongate opercle; the unossified gill rakers, and a urohyal with two foramina (Arratia, 1998; Fernández & de Pinna, 2005). Additional evidence for a sister–group relationship between the new species and four species of Silvinichthys is found in various other anatomical traits. Silvinichthys pedernalensis shares with S. bortayro, S. gualcamayo, S. huachi, and S. leoncitensis (Fernandez et al., 2014) the absence of pelvic girdle. Several trichomycterids species show reductive trends in their pelvic fins and girdle, such as Trichomycterus anhanga, T. candidus, T. catamarcensis, T. tropeiro, Eremophilus mutisi, the Tridentinae Miuroglanis platycephalus, the Glanapteryginae (except for some specimens of Glanapteryx), Ituglanis apteryx, and some specimens of Ituglanis parahybae (Fernández & Vari, 2000; Ferrer & Malabarba, 2011; Dutra et al., 2012; Datovo, 2014). S. pedernalensis also shares the three derived characters mentioned by Fernandez et al. (2013) along with S. bortayro, S. gualcamayo, S. huachi, and S. leoncitensis: the reduced numbers of odontodes on the opercular (2–9) and interopercular (9–28), and the absence of the orbitosphenoid bone. Possession of these characters may indicate sister species, but confirmation of such a hypothesis requires a broader comparative analysis incorporating information from multiple character systems. Acknowledgements Research associated with this project was supported by PIP (Proyecto Investigación Plurianual, CONICET) project nº 11420090100321. For loans and other assistance we thank S. Schaefer, B. Brown and R. Arrindell (AMNH), M. Sabaj Pérez (ANSP), D. Catania (CAS), J. Maclaine (BMNH), J. Andreoli–Bize (FACEN), M. Rogers (FMNH), M. Retzer (INHS), A. Bentley (KU), G. Chiaramonte (MACN), F. Lobo and V. Martínez (IBIGEO), K. Hartel (MCZ), J. Lima de Figueiredo and O. Oyakawa (MZUSP), L. Malabarba and R. Reis (PUCRS), H. López and L. Protogino (MLP), M. Arraya, F. Carvajal, and M. Maldonado (UMSS), P. Buckup (MNRJ), H. Ortega and M. Velasquez (MUSM), V. Jerez (MZUC), R. Robins (UF), D. Nelson (UMMZ), S. Raredon and R. Vari (USNM). This paper benefited from comments and suggestions from M. Hilal (UNT), the Editor of ABC, and three anonymous reviewers. References Arratia, G., 1998. Silvinichthys, a new genus of trichomycterid catfishes from the Argentinean Andes, with redescription of Trichomycterus nigricans. Ichthyological Exploration of Freshwaters, 9: 347–370. Arratia, G., Chang, A., Menu–Marque, S. & Rojas, R., 1978. About Bullockia gen. nov., Trichomycterus

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mendozensis n. sp. and revision of the family Trichomycteridae (Pisces: Siluriformes). Studies on Neotropical Fauna and Environment, 13: 157–194. Costa, W. W. E. M. & Bockmann, F. A., 1993. Un nouveau genre Néotropical de la famille des Trichomycteridae (Siluriformes: Loricarioidei). Revue Francaise d’Aquariologie, 20: 43–46. Datovo, A., 2014. A new species of Ituglanis from the Rio Xingu basin, Brazil, and the evolution of pelvic fin loss in trichomycterid catfishes (Teleostei: Siluriformes: Trichomycteridae). Zootaxa, 3790: 466–476. Dutra, G. M., Wosiacki, W. B. & de Pinna, M. C. C., 2012. Trichomycterus anhanga, a new species of miniatura catfish related to T. hasemani and T. johnsoni (Siluriformes: Trichomycteridae) from the Amazon basin, Brazil. Neotropical Ichthyology, 10: 225–231. Fernández, L. & de Pinna, M. C. C., 2005. A phreatic catfish of the genus Silvinichthys from southern South America (Teleostei, Siluriformes, Trichomycteridae). Copeia, 2005: 100–108. Fernández, L., Dominino, J., Brancolini, F. & Baigu, C., 2011. A new catfish species of the genus Silvinichthys (Teleostei: Trichomycteridae) from Leoncito National Park, Argentina. Ichthyological Exploration of Freshwaters, 22: 227–232. Fernández, L., Sanabria, E. & Quiroga, L., 2013. Silvinichthys gualcamayo, a new species of catfish from the central Andes of Argentina (Siluriformes: Trichomycteridae). Ichthyological Exploration of Freshwaters, 23: 367–373. Fernández, L., Sanabria, E. A., Quiroga, L. B. & Vari, R. P., 2014. A new species of Silvinichthys (Siluriformes, Trichomycteridae) lacking pelvic fins from mid–elevation localities of the southern Andes, with comment on the genus. Journal Fish Biology, 84: 372–382. Fernández, L. & Schaefer, S. A., 2003. Trichomycterus yuska, a new species from high elevations of Argentina (Siluriformes: Trichomycteridae). Ichthyological Exploration of Freshwaters, 14: 353–360. Fernández, L. & Vari, R. P., 2000. A new species of Trichomycterus (Teleostei: Siluriformes: Trichomycteridae) lacking a pelvic girdle from the Andes of Argentina. Copeia, 2000: 990–996. – 2009. New species of Trichomycterus from the

Andean Cordillera of Argentina (Siluriformes: Trichomycteridae). Copeia, 2009: 195–202. Ferrer, J. & Malabarba, L. R., 2011. A new Trichomycterus lacking pelvic fins and pelvic girdle with a very restricted range in southern Brazil (Siluriformes: Trichomycteridae). Zootaxa, 2912: 59–67. Northcutt, G., 1989. The phylogenetic distribution and innervation of craniate mechanoreceptive lateral lines. In: The mechanosensory lateral line: 17–18 (S. P. Coombs, S. P., P. Gorner & H. Munz, Eds.). Springer, New York. de Pinna, M. C. C., 1989. A new sarcoglanidine catfish, phylogeny of its subfamily, and an appraisal of the phyletic status of the Trichomycterinae (Teleostei, Trichomycteridae). American Museum Novitates, 2950: 1–39. – 1992. A new subfamily of Trichomycteridae (Teleostei, Siluriformes), lower loricarioid relationships and a discussion on the impact of additional taxa for phylogenetic analysis. Zoological Journal of the Linnean Society, 106: 175–229. – 1998. Phylogenetic relationships of Neotropical Siluriformes (Teleostei: Ostariophysi); historical overview and synthesis of hypotheses. In: Phylogeny and classification of Neotropical fishes: 279–330 (L. R. Malabarba, R. E. Reis, R. P. Vari, Z. M. S. Lucena & C. A. S. Lucena, Eds.). EDIPUCRS, Porto Alegre, Rio Grande do Sul, Brazil. Sabaj Pérez, M. H., 2014. Standard symbolic codes for institutional resource collections in herpetology and ichthyology: an online reference. Version 5.0 (22 September 2014). Electronically accessible at http://www.asih.org/, American Society of Ichthyology and Herpetologists, Washington, DC. Schaefer, S. A. & Fernández, L., 2009. Redescription of the Pez Graso, Rhizosomichthys totae (Trichomycteridae), of Lago de Tota, Colombia, and aspects of cranial osteology revealed by microtomography. Copeia, 2009: 510–522. Taylor, W. R. & Van Dyke, G. C., 1985. Revised procedures for staining and clearing small fishes and other vertebrates for bone and cartilage study. Cybium, 9: 107–119. Tchernavin, V., 1944. A revision of some Trichomyc– terinae based on material preserved in the British Museum (Nat. Hist.). Proceedings of the Zoological Society of London, 114: 234–275.

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Worldwide distribution of non–native Amazon parrots and temporal trends of their global trade E. Mori, G. Grandi, M. Menchetti, J. L. Tella, H. A. Jackson, L. Reino, A. van Kleunen, R. Figueira & L. Ancillotto

Mori, E., Grandi, G., Menchetti, M., Tella, J. L., Jackson, H. A., Reino, L., van Kleunen, A., Figueira, R. & Ancillotto, L., 2017. Worldwide distribution of non–native Amazon parrots and temporal trends of their global trade. Animal Biodiversity and Conservation, 40.1: 49–62. Abstract Worldwide distribution of non–native Amazon parrots and temporal trends of their global trade.— Alien species are the second leading cause of the global biodiversity crisis, after habitat loss and fragmentation. Popular pet species, such as parrots and parakeets (Aves, Psittaciformes), are often introduced outside their native range as a result of the pet trade. On escape from captivity, some such species, such as the ring–necked parakeet and the monk parakeet, are highly invasive and successfully compete with native species. Popula� tions of Amazon parrots (Amazona spp.) can be found throughout the world, but data on their status, distribu� tion and impact are incomplete. We gathered and reviewed the available information concerning global trade, distribution, abundance and ecology of Amazon parrots outside their native range. Our review shows that at least nine species of Amazon parrots have established populations outside their original range of occurrence throughout the world (in Europe, South Africa, the Caribbean islands, Hawaii, and North and South America). Their elusive behaviour and small population size suggest that the number of alien nuclei could be underes� timated or at undetected. Despite international trade bans, the large trade of wild–caught Amazon parrots in past decades appears to have contributed to the establishment of alien populations worldwide. Establishment success seems to differ geographically. While European populations are still small and growing slowly, USA populations are large and expanding geographically. This difference is not related to large propagule pressure (trade) but possibly to a better niche match between native and introduced ranges. Amazona aestiva is the most frequently encountered Amazona parrot, with at least eight alien populations reported to date. All these populations, with the exception of those in the USA where the climate is more suitable for their establishment, are composed of a low number of individuals even though they have been established for a long period of time. Further research is required as little information is available on the ecology and potential impact of these alien populations. Key words: Alien species, Amazona, Distribution range assessment, Establishment success, Impacts Resumen Distribución en el mundo de los loros introducidos del género Amazona y tendencias temporales de su comercio a escala mundial.— Las especies exóticas son la segunda causa de la crisis de biodiversidad mundial, precedida por la pérdida y la fragmentación del hábitat. Algunas especies populares como mascotas, como los loros y las cotorras (Aves, Psittaciformes) suelen introducirse fuera de su área de distribución nativa a consecuencia del comercio de animales de compañía. Si escapan de su cautiverio, algunas de estas especies, como la cotorra de Kramer y la cotorra argentina, son sumamente invasivas y compiten con las especies autóctonas. Las poblaciones de loros del género Amazona pueden encontrarse en todo el mundo, pero los datos relativos a su estado, distribución y efectos son incompletos. Recopilamos y examinamos la información disponible relativa a la ecología, la abundancia, la distribución y el comercio en el mundo de los loros del género Amazona fuera de su área de distribución nativa. Nuestro examen revela que al menos nueve especies de loros de este género han establecido poblaciones fuera de su área de distribución original en todo el mundo (en Europa, Sudáfrica, las islas del Caribe, Hawaii y América del Norte y del Sur). Su comportamiento esquivo y el reducido tamaño de la población sugieren que se podría haber infravalorado el número de núcleos intro� ISSN: 1578–665 X eISSN: 2014–928 X

Mori et al.

ducidos o que podrían no haberse detectado todos. Pese a las prohibiciones impuestas al comercio interna� cional, parece que el gran volumen de loros Amazona capturados en libertad que se ha comerciado en los últimos decenios ha contribuido al establecimiento de poblaciones foráneas en todo el mundo. Parece que el éxito del establecimiento varía en función de la zona geográfica. Mientras que las poblaciones europeas siguen siendo de pequeño tamaño y de crecimiento lento, las de los Estados Unidos son numerosas y están en expansión. Esta diferencia no guarda relación con una elevada presión del propágulo (comercio), pero sí lo haga posiblemente con una mejor correspondencia de nichos entre las áreas de distribución originales y las de introducción. Amazona aestiva es la especie del género que se observa con mayor frecuencia y hasta la fecha se han notificado al menos ocho poblaciones foráneas. Todas estas poblaciones, salvo aquellas que se encuentran en zonas de los Estados Unidos donde el clima les es más propicio, están formadas por unos pocos individuos, a pesar de que lleven establecidas un largo período de tiempo. Es necesario seguir estudiando sobre la ecología de estas especies exóticas y sus posibles repercusiones debido a la escasa información disponible al respecto. Palabras clave: Especies exóticas, Amazona, Evaluación del área de distribución, Éxito del establecimiento, Repercusiones Received: 29 IV 16; Conditional acceptance: 16 VI 16; Final acceptance: 20 IX 16 Emiliano Mori, Dept. of Life Sciences, Univ. of Siena, Via P. A. Mattioli 4, 53100 Siena (SI), Italy.– Gioele Grandi, Dept. of Earth and Environmental Sciences, Univ. of Pavia, Via A. Ferrata, 9, I–27100 Pavia, Italy.– Mattia Menchetti, Dept. of Biology, Univ. of Florence, Via Madonna del Piano 6, 50019 Sesto Fiorentino (Fi), Italy.– José L. Tella, Dept. of Conservation Biology, Estación Biológica de Doñana (CSIC), Sevilla, Spain.– Hazel Jackson, Durrell Inst. of Conservation and Ecology, School of Anthropology and Conservation, Univ. of Kent, Canterbury, Kent, CT53AU, U.K.– Luís Reino, CIBIO/InBIO–Centro de Investigação em Biodiversidade e Recursos Genéticos, Univ. do Porto, Campus Agrário de Vairão, Rua Padre Armando Quintas, 7, 4485–661 Vairão, Portugal; CIBIO/InBIO–Centro de Investigação em Biodiversidade e Recursos Genéticos, Univ. de Évora, 7004–516 Évora, Portugal.– André van Kleunen, Sovon Dutch Centre for Field Ornithology, Toernooiveld 1, 6525 ED Nijmegen, Netherlands.– Rui Figueira, CIBIO/InBio, Centro de Investigação em Biodiversidade e Recursos Genéticos, Univ. do Porto. Campus Agrário de Vairão, Vairão, Portugal; CEABN/InBIO, Centro de Ecologia Aplicada 'Professor Baeta Neves', Inst. Superior de Agronomia, Univ. de Lisboa, Tapada da Ajuda, 1349–017 Lisboa, Portugal.– Leonardo Ancillotto, Wildlife Research Unit, Lab. di Ecologia Applicata, Sezione di Biologia e Protezione dei Sistemi Agrari e Forestali, Dipto. di Agraria, Univ. degli Studi di Napoli Federico II, via Università 100, 80055 Portici (NA), Italy. Corresponding author: E. Mori. E–mail: moriemiliano@tiscali.it

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Introduction Human–assisted transport of live animals has occu� rred since ancient times (Meyerson & Mooney, 2007; Tella, 2011). Recent globalization trends, however, have facilitated the international wildlife trade and the consequent introduction and spread of alien species (Hulme, 2009). Throughout the world, introduced spe� cies have led to a large number of local and global extinctions, and the population decline of native spe� cies (Wohnam, 2006). Introduced species may also damage human activities (e.g., agriculture), resulting in economic damage and loss of wellbeing (Vitousek et al., 1996; Mack et al., 2000). In spite of this, the impact of many introduced species remains poorly known or hard to assess, especially that concerning birds (Kumschick & Nentwig, 2010). Thus, it is impor� tant to determine the extent of species distribution in non–native environments in order to observe trends in population growth and spread and to predict and manage the impact of introduced species. Many species kept as pets or attractions in urban parks, in zoos and in private homes may escape from captivity, sometimes establishing self–sustainable po� pulations (Reino & Silva, 1996; Duncan et al., 2003; Abellán et al., 2016). Parrots (Aves, Psittaciformes) are prominent among internationally traded birds because of their worldwide popularity as pets (Tella & Hiraldo, 2014), likely leading to the establishment of a number of non–native populations (Duncan et al., 2003; Cassey et al., 2004; Mori et al., 2013a; Abellán et al., 2016). Currently, approximately 60 out of 355 known parrot species have established at least one breeding population outside their native ranges (Menchetti & Mori, 2014). Although these species may be widely distributed and have easily–detectable populations (Mori et al., 2013a), the impact of intro� duced parrots on native biodiversity/environment has been largely overlooked and is still poorly understood (Juniper & Parr, 1988; Menchetti & Mori, 2014). To date, the impact of such invasion has mainly been competition with native hole–nester species (Strub� be et al., 2010; Mori et al., 2013b; Menchetti et al., 2014; Hernández–Brito et al., 2014), and damage to crops and infrastructures (Avery et al., 2002; Stafford, 2003; Menchetti & Mori, 2014), but it should be kept in mind that parrots and parakeets are also potential reservoirs of a variety of diseases transmittable to humans, domestic animals and wildlife (Fletcher & Askew, 2007; Runde et al., 2007), thus emphasizing the need for early detection and assessment of intro� duced populations in order to reduce risks of damage to local wildlife and society. The genus Amazona includes 32 species of medium–sized parrots, native to Central and South America (cf. Menchetti & Mori, 2014). Hybridization between species is known to occur both in nature and in captivity (McCarty, 2006). Amazon parrots are very popular in the pet trade due to their sociability and ability to imitate human voices (Tella & Hiraldo, 2014). Global population trends of Amazon parrots in their native distribution ranges have not been assessed for all the species, but several population

declines have been related to legal and illegal capture of wild individuals (Tella & Hiraldo, 2014). According to CITES (www.cites.org), over 31,660 wild–caught individuals were recorded in the international trade database between 1981 and 2005. Although anecdotal and fragmented, some in� formation is available on the presence of alien po� pulations of Amazon parrots throughout Europe (A. aestiva, A. oratrix and A. ochrocephala) and USA (A. viridigenalis, A. aestiva, A. autumnalis, A. albifrons, A. finschi, A. oratrix, A. ochrocephala). Menchetti & Mori (2014) analysed the known, certified effects of introduced parrots on native biodiversity but the status and impact of these populations and their worldwide ranges has not been systematically assessed. Given the importance of assessing the distribution of alien species (Genovesi & Shine, 2004), we aimed to fill this gap by reviewing the occurrences of alien populations of Amazon parrots worldwide and by assessing the status of these populations from the available literature, local experts and web–portals for bird observations. Trade data were also obtained for each country to explore temporal trends in trade and relationships with the establishment of non–native populations. Material and methods Occurrences were first searched for through online databases (i.e., ISI Web of Science, Scopus, Go� ogle Scholar). Search terms included all possible combinations of these words, in several languages (English, French, Italian, Portuguese, German, Dutch and Spanish): Amazon, Amazona aestiva, Amazona ochrocephala, Amazona oratrix, Amazona amazonica, Amazona autumnalis, Amazona viridigenalis, Amazona, alien population, introduction. Information on detected introduced populations was also obtained by contac� ting 64 local ornithologists and birdwatchers, including the authors of ornithological bulletins and the mailing list of the COST funded project named 'ParrotNet' (Action ES1304), i.e. a network of researchers, practitioners and policy–makers in Europe studying distribution and impacts of free–ranging parrots. Addi� tional occurrences were searched on citizen science based databases, i.e. iNaturalist (www.inaturalist.org) and eBird (www.ebird.org). Owners of data uploaded on these databases were also contacted for further information on their observations. We also checked National European databases of birds and non–native species and we reviewed the Christmas Bird Count (CBC), the American citizen–science, peer–reviewed database of the National Audubon Society, to assess the status of Amazon parrots populations introduced in North America (www.audubon.org; www.christmas� birdcount.org). A GLMM with binomial distribution (response varia¬ble: established or not) and logistic link function was used to assess the relation between the number of individuals of each Amazona species per country (i.e., a proxy of propagule pressure) and the esta� blishment success. The model fitted the number of

individuals of each species imported per country as an explanatory covariate. Species and country identities were used as random effects. All records of worldwide trade on Amazona spp. between 1980 and 2013 were obtained from the CITES trade database, to detect temporal trends in the international trade. Some discrepancies were identified between reported exports and imports; in these cases, trade data were filtered to obtain records of gross imports for wild birds. Data were taken from the CITES Trade Database of the United Nations Envi� ronment Programme (World Conservation Monitoring Centre: www.trade.cites.org/cites_trade_guidelines/ en–CITES_Trade_Database_Guide.pdf [Accessed on 22nd July 2016]). The CITES gross trade output compares the quantities reported by the exporter and importer, providing an estimate of the total number of individuals recorded in international trade. In other words, gross imports were used to take into account records of imports and re–exports. Results Records of wild populations of Amazona spp. A total of 22 papers, books and book chapters mentioning the genus Amazona outside its native range were identified through our literature screening. Publications were written in five languages: English (N = 11), German (N = 5), Italian (N = 3), Portuguese (N = 2) and Dutch (N = 1). Another six reviews sum� marizing the distribution of alien species in France, the Arab Peninsula, the Far East and North America were checked, although no data on Amazon parrots were found. Furthermore, 64 ornithologists or local experts were contacted from all countries reported in the 'Results' paragraph; of these, only 36 provi� ded us with feedback and 16 sent us unpublished data (see 'Acknowledgements'), or other published works we missed in our research (N = 7 papers in English, on North American populations). None of the others (N = 20 experts) added any relevant data on the population of Amazona spp. Furthermore, three papers from local newspapers provided us with data on Amazon parrots in Italy and Germany. Figure 1 shows the distribution of introduced breeding popu� lations of Amazona parrots. A total 44 records from 24 geographical areas of 9 countries were obtained from citizen–science platforms, as well as from social networks (e.g., Facebook) and online forums (e.g., Natura Mediterraneo: www.naturamediterraneo.com). Detailed data on breeding population trends were only available for three European populations, two from Italy and one from Germany (see data in the paragraphs below: fig. 2). These showed a linear increase in population size, though the oldest one (A. oratrix in Stuttgart) best fitted an exponential growth curve (fig. 2). Establishment success of each species was not related to the number of individuals imported by each country (GLMM: Estimate ± SE: 88 ± 2.31, df = 1, P = 0.26).

Mori et al.

Italy Two reproductive populations of Amazon parrots are currently present in Italy, one in Genoa (Liguria, North–Western Italy) and one in Milan (Lombardy, Northern Italy). In Genoa, the earliest presence of A. aestiva dates back to 1991, with the first breeding event documented in 1993 (Maranini & Galuppo, 1993, 1998). In recent years, mixed flocks of A. aestiva × A. ochrocephala, together with individuals with intermediate phenotypes, suggested that hybridization has occurred (Andreotti & Piacentino, 2009). McCarthy (2006) showed that hybridization among these species is possible in captivity, possibly because of their genetic similarity (Ribas et al., 2007). Recorded dietary preferences of the wild populations in Genoa comprised tree seeds and fruits, but no evidence of damage to plants has been documented (Andreotti & Piacentino, 2009). In 2009, 5–6 breeding pairs were present within two city districts of Genoa (i.e., Castelletto and Albaro districts). The number of breeding pairs might have been underestimated because of the elusive habits of these parrots during the breeding season (Seixas & de Miranda Mourão, 2002; Andreotti & Piacentino, 2009). About 20–30 individuals of Amazona are cu� rrently present in Genoa (fig. 2). No chicks produced by hybrid pairs have been observed, suggesting a probable low fitness of the hybrids/mixed pairs (An� dreotti & Piacentino, 2009). Andreotti & Piacentino (2009) reported rats and jackdaws Corvus monedula, as possible predators of chicks, although aggressive interactions have only been observed among jack� daws. A single A. amazonica was also repeatedly observed in 2008 and 2009 (Andreotti & Piacentino, 2009). In Milan (Northern Italy), free–ranging A. aestiva were first documented in 1994 (2 individuals, N. Ferrari and A. Peruz, pers. comm., 2015). A group of 8–10 A. aestiva h as been observed at a roost within the Indro Montanelli Gardens. The roost is shared with several individuals of Psittacula krameri. Two A. ochrocephala have also been observed at the same roost since 2014 (A. Peruz, pers. comm., 2015). These parrots feed in the Botanical Garden of Milan (N. Ferrari, pers. comm., 2015), and in Parco Lambro (4.5 km North–East to the roost: E. Mori, pers. obs., 2015), and roost mainly on the canopies of Platanus orientalis and Gingko biloba. Although nests of this species are often located very high on the tree trunks and are hard to detect, the long–term reported presence of this population in Milan, as well as the observation of young individuals (< 1 year), suggests that they are successfully reproducing (Andreotti & Piacentino, 2009). Only one breeding occurrence has been recorded, with a nest and two chicks observed in a hole of a P. orientalis in Piazza della Republica, Milan (April 2011: A. Marangoni, pers. comm., 2015). Furthermore, new releases and escapes may have maintained the population of A. aestiva in Milan (at least two individuals have escaped in the last 5 years: cf. fig. 1s in supplementary material). In addition to the populations in Genoa and Milan, two A. aestiva were documented nesting in a tree hole in a private garden from January to May 2007 in Giaveno (Province of Turin,

Animal Biodiversity and Conservation 40.1 (2017)

A. autumnalis USA

A. ochrocephala USA

A. oratrix USA Germany

A. ventralis Caribbean

A. finschi USA Florida

A. viridigenalis USA Caribbean

A. albifrons USA Caribbean

A. aestiva A. amazonica USA Caribbean Spain (Tenerife)

USA Italy Brazil Argentina South Africa

Fig. 1. Global distribution of introduced, breeding Amazona populations. Photos by: iNaturalist (jmoralesrbpc) and Wikimedia Commons (Brian Gratwicke, Leonhard F. Charlesjsharp, DickDaniels, Joe Quick, Smartneddy and Joseph C. Boone). Map from Wikimedia Commons (Crates). Fig. 1. Distribución mundial de las poblaciones reproductoras de Amazona introducidas. Fotografías por cortesía de iNaturalist (jmoralesrbpc) y Wikimedia Commons (Brian Gratwicke, Leonhard F. Charlesjsharp, DickDaniels, Joe Quick, Smartneddy y Joseph C. Boone). Mapa de Wikimedia Commons (Crates).

North–Western Italy: M. Colonna, pers. comm., 2015). The parrots nested within the hole of a P. orientalis, at a height of 5 m. The female Amazon showed aggres� sive behaviour towards Corvus monedula that tried to enter the nest. Two fertilized eggs were laid, but both adults and eggs were recaptured and caged before hatching. A number of incidental observations and escapes from captivity were also recorded throughout Italy; an average of 2.66 ± 2.42 escapes were reported per year, with a total of 34 Amazon parrots recorded to have escaped between 2004 and 2012 in Italy (see map in fig. 1s in supplementary material).

Germany According to the recent review by Nehring & Rabitsch (2015), the current status of A. aestiva in Germany is unclear, as many breeding events were observed in the past, but no established population of this species seems to occur currently. Bauer & Woog (2008), referring to Herkenrath (1995), mentioned that a breeding pair of A. aestiva was observed in Nordrhein–Westfalen in 1883. However, in Herken� rath (1995), there is no reference to this and it may represent a confusion with Niethammer (1963), who mentioned a breeding pair of A. aestiva in 1893 in

Switzerland, where the species has never established (Mori et al., 2013a). Four individuals of A. aestiva, erroneously recorded as A. ochrocephala, were ob� served between 1984 and 1998 in the Schlosspark Von Wiesbaden–Biebrich (Hessen) (Zingel, 1990). Cross–breeding events between A. aestiva and A. amazonica were observed between 2000 and 2003 (Stübing et al., 2010). This small population does not seem to increase, as it never exceeded four parrots, individually identified by observers (D. Franz, pers. comm., 2015). Between 1991 and 1993, a bree� ding pair of A. aestiva was observed close to Köln (Kretzschmar, 1999), and in 1999 two individuals were observed in Rosensteinpark, Stuttgart (Hoppe, 1999). Since 1984, a breeding population of yellow–hea� ded Amazon parrots (A. oratrix) has established in Stuttgart (Martens et al., 2013), starting with a single pair which bred for the first time in 1985; reproduction occurred every year, bringing the population to nearly 50 individuals in 2015 (D. Franz, pers. comm., 2015). A few individuals of A. aestiva and hybrids of A. aestiva × A. oratrix (i.e. individuals with intermediate phenotype) are also regularly observed (Martens et al., 2013). Amazon parrots in Stuttgart feed on a variety of cultivated and wild plant species, with a preference for Rosaceae and Betulaceae; eaten parts include unripe fruits, seeds and blossoms (Martens et al., 2013). Native and non–native plants do not seem to be affected by the feeding behaviour of Amazon parrots in Stuttgart, possibly because of the small population size and the wide foraging area (Martens et al., 2013). In Stuttgart, A. oratrix have been observed while mobbing C. monedula coming close to a nesting hole on a plane tree (D. Franz, pers. comm., 2015). Spain Abellán et al. (2016) reported the observations of free–ranging exotic birds recorded in Spain between 1912 and 2012, including 94 records (165 individuals) belonging to seven Amazona species. Amazona aestiva was the most commonly recor� ded species (46 records, 79 individuals) following detection for the first time in Santa Cruz de Tenerife (Canary Islands) in 1992: this population decreased and became extinct before 2000. Further isolated in� dividuals were recorded in Mallorca (Balearic Islands) and in the continental provinces of Barcelona, Burgos, Girona, Málaga, Toledo and Valencia. In Valencia, the reproduction of A. aestiva was recorded in 2009 (one breeding pair). There appear to be no established populations currently, and most sightings throughout the country appear to involve escaped birds. Amazona ochrocephala was the second most re� corded species (23 records, 53 individuals), closely followed by A. amazonica (18 records, 25 indivi� duals), observed for the first time in 1995 and 2001 respectively. Amazona ochrocephala was recorded as a breeding species in Tenerife in 2003, 2004 and 2005, while single escaped individuals were obser� ved in the provinces of Alicante, Barcelona, Seville and Valencia. Reproduction of A. amazonica was recorded in Barcelona only once, in 2004 (Abellàn et al., 2016), and in Tenerife in 2014 (grupodeave�

Mori et al.

sexoticas.blogspot.com.es [Accessed on 23rd March 2016]. A possible successful hybridization between A. amazonica and A. ochrocephala was recorded in Tenerife in 2015 (D. Hernández–Brito & G. Blanco, pers. comm., 2015). A group of four individuals of A. amazonica was observed in Tenerife in 2011. Single individuals were observed in the provinces of Alicante, Barcelona, Madrid, Málaga, Seville and Valencia, and two individuals were observed in Mallorca (Balearic Islands). Additional records include three A. amazonica individuals (provinces of Tenerife, Seville and Vizcaya, in 2013–2015) and three A. ochrocephala (Tenerife, in 2013 and 2014). A pair of A. leucocephala was also recorded in 1997 in Tenerife, successfully breeding in 1998. These birds were then captured (R. Zamora, pers. comm., 2015) and only two additional records of escaped individuals in mainland Spain are known. Single records were obtained from A. albifrons, A. farinosa and A. festiva in mainland Spain. One individual of A. oratrix was observed in Córdoba in 2014 (grupodeavesexoticas.blogspot.com.es [Ac� cessed on 23rd March 2016]). To conclude, despite the relatively high number of observations and a few reproduction events of Amazon parrots in Spain, a breeding population, small and mixed (A. amazonica/ A. ochrocephala), is known only in Tenerife. Portugal Single records of A. ocrocephala, A. amazonica and other unidentified Amazon parrots have been reported for Lisbon. Matias (2011) reported a sighting of A. ochrocephala (one individual in 2007) and a group of four individuals of A amazonica in a small city park in 2009 (Matias, 2011). New records have since descri� bed a group of three individuals of A. amazonica in 2012 and one individual in 2014 (Gomes, 2014). In 2014, a single individual was observed in an urban park in Póvoa do Varzim (NW Portugal) over several months (Gomes, 2014). Netherlands In the Netherlands, 13 individuals belonging to four Amazona species (A. aestiva, A. ochrocephala, A. viridigenalis and A. amazonica) were observed bet� ween 1984 and 2012, mostly in the urban centres surrounding The Hague (waarneming.nl/; van Kleunen et al., 2014). Among these, A. aestiva was observed in Bunnik (August 1994) and in Voorburg (December 2006) in a roosting flock mixed with Psittacula krameri. Amazona amazonica was observed in Brabantse Biesbosch (June 2011) and Losser (March 2012). No evidence of reproduction was reported. Since 2012, no Amazona species have been reported for the Netherlands. USA–Florida At least 12 Amazona species have been reported for Florida, mainly concentrated within the greater metropolitan Miami area (Florida Fish and Wildlife Conservation Commission, 2003). Amazona aestiva was recorded as breeding in Miami–Dade County, where it appeared to take hold in the late 1980s (Kale et al., 1992; Florida Fish and Wildlife Conservation

Animal Biodiversity and Conservation 40.1 (2017)

R2 = 0.8368

Number of individuals

60 50

•

40 30

•

R2 = 0.88128

R2 = 0.97278

10 0 1980

• •

1985

• n

1990

• n

1995 2000 Year

2005

2010

2015

Fig. 2. Trend in the number of breeding adults in Amazona populations in Europe: A. aestiva, Genoa, Italy (u) and Milan (n); A. oratrix, Stuttgart, Germany (•). Data sources were provided by local ornithologists or taken from publications regarding these populations (see text). Fig. 2. Evolución del número de adultos reproductores en las poblaciones de Amazona en Europa. A. aestiva, Génova, Italia (u) y Milán (n); A. oratrix, Stuttgart, Alemania (•). Las fuentes de los datos fueron proporcionadas por ornitólogos locales o se tomaron de publicaciones referentes a estas poblaciones (véase el texto).

Commission, 2003). Recent assessments indicate a positive population trend for this species, but reliable quantitative data are not available (Runde et al., 2007). Amazona viridigenalis was released in Florida between the late 1960s and the early 1970s, with at least 11 individuals. Owre (1973) reported this species as the most abundant Amazon parrot established in Florida, counting a flock of 32 individuals in 1972 (Robertson & Woolfenden, 1992). The population experienced a rapid growth since the 1980s, although a negative trend occurred since 2005 (Runde et al., 2007), with only a few scattered individuals observed in Broward, Miami–Dade, Fort Lauderdale, Palm Beach and in the Florida Keys, where hybrids with A. ochrocephala were also observed (cf. National Audo� bon Society, 2016). Epps & Karalus (2007) suggested that competition with A. amazonica, locally much more abundant, may have occurred for food resources. Amazona finschi was first reported in the 1970s (Robertson & Woolfenden, 1992) in Broward County and Southern Miami (N = 4). In 2006, a population was still present (Epps & Karalus, 2007) and a positive trend in population size was recorded (Runde et al., 2007). In 2016, about 15–20 individuals have been observed (D. Marty, pers. comm., 2016). A single population of A. amazonica is present in Southern Florida (Miami– Dade and Broward Counties). No population estimate is available, but the species in currently considered to be the most abundant parrot in Southern Florida (Epps & Karalus, 2007). A. ochrocephala was considered as

established in Florida in 1986 (Troops & Dilley, 1986), although, apart from isolated records of a single or few individuals within flocks of other species in Miami, no observation has been reported since 2007 (Epps & Karalus, 2007). A few individuals of A. auropalliata have been recorded in Florida (Broward County), with successful breeding by one pair documented (i.e., observation of fledged chicks) in 2000 and 2001 (Epps & Karalus, 2007). Groups of A. oratrix have bred in Broward County since 1985, most likely in small numbers, and hybridization with A. viridigenalis has also been observed (Epps & Karalus, 2007); A. ochrocephala has been also recorded as a breeding species in Florida (Toft & Wright, 2015). Escapes of other Amazona species are often reported in Florida, mainly in the Miami area (A. albifrons, A. autumnalis, A. petrei, A. ventralis: Robertson & Woolfenden, 1992), with groups of up to 30 A. albifrons reported in 2015 by the eBird portal. No data on the initial propagule pressure are available. USA–California Six species of Amazon parrots have been reported in California. Amazona aestiva has been introduced, with a small number of individuals (N = 2) reported in the Los Angeles basin, the San Gabriel Valley, and urban Orange County, possibly sustained by repeated escapes, and often detected in mixed flocks with other parrot species. Breeding has been reported for one pair in the San Gabriel Valley (Mabb, 2002) and in

Orange County, where 8–20 A. aestiva are currently individually monitored (www.californiaparrotproject. org; National Audobon Society, 2016). Amazona viridigenalis is present in California with a population founded in 1963 by two pairs, released in near Pasadena. This geographical area hosts the largest population of this species in California, with about 750 individuals in 1996 (Mabb, 1997): in 2015, a total of 263 individuals was counted. Other groups are present in the north–eastern area of Los Angeles, in Malibu, Mill Creek, San Diego and Orange County (National Audobon Society, 2016). Garrett (1997) conservatively estimated a total population count of 1,080 individuals in California, subsequently finding a significant population increase over time, reaching about 2,500 individuals in 2016 (www.forbes.com/ sites/grrlscientist/2016/04/07/are–there–more–free– living–mexican–red–headed–parrots–in–us–cities– than–in–all–of–mexico/#12412ac4675a [Accessed on 25th April 2016]; National Audobon Society, 2016). Mixed pairs A. viridigenalis × A. finschi were also observed in Pasadena in the late 1990s, with no recent confirmations (Mabb, 1997). Amazona finschi was recorded in California in 1976 for the first time, and has been considered to be established in Los Angeles since 1987. Garrett (1997) estimated 100 individuals in this population, although the cu� rrent count is no more than 55 individuals (National Audobon Society, 2016). Mabb (1997) observed a breeding pair, nesting in a utility pole, aggressively chasing Sturnus vulgaris and Corvus brachyrynchos. Two individuals of A. autumnalis were recorded in San Bernardino in 1972 by Hardy (1973) and in 1997 by Mabb (1997); 4–6 individuals were obser� ved in 2002 in the San Gabriel Valley (Mabb, 2002), with evidence of breeding. This species exhibited an evident increase in population size (Runde et al., 2007), and a total of 32 individuals were counted in 2015 in Orange County (National Audobon So� ciety, 2016). Amazona ochrocephala is present in California, with the first 10 breeding pairs in 1963; around 30 individuals were counted in 1973 (Hardy, 1973), but no recent population count is available, and only 1–2 individuals have been observed since 2010 (National Audobon Society, 2016). Although possibly confused with A. ochrocephala, A. oratrix was once widespread in southern California (Los Angeles, San Diego, Pasadena) but its population seems to have declined in recent years (Lever, 2005), with 5 individuals observed in 2015 in San Diego and 5 in Pasadena (National Audobon Society, 2016). The total population for California was estimated at about 60 individuals in late 1990s (Garrett, 1997). Toft & Wright (2015) also reported A. albifrons as an established species in Los Angeles County; ob� servations of fewer than 10 individuals occurred in 2015, also in Orange County and Pasadena (National Audobon Society, 2016). USA–other States A small population of A. viridigenalis persisted in Southern Texas (La Feria) between the 1920s and 1930s (Le� ver, 1987, 2005). Two groups of A. viridigenalis were

Mori et al.

present in 1973–75 in Texas (total N = 12, in Rio Grande), with an estimated population of about 400 individuals in 1995 (Butler, 2005) and about 700–1000 individuals (Brownsville, Harlingen, Wes� laco, Anzalduas–Bentsen) in 2016 (www.forbes.com/ sites/grrlscientist/2016/04/07/are–there–more–free– living–mexican–red–headed–parrots–in–us–cities– than–in–all–of–mexico/#12412ac4675a. Accessed on 25th April 2016; National Audobon Society, 2016). This species was reported since 1970 on the island of Oahu, Hawaii (Lever, 2005), where it reproduced until 1980s and then, possibly, disappeared (cf. Run� de et al., 2007). A single individual of A. finschi was observed at El Paso, in Texas, in 2015 (National Au� dobon Society, 2016). Haphazard observations were reported of single individuals of A. ochrocephala in New York and small numbers (2–4) in Texas in the 1970s (Lever, 1987). Puerto Rico and other Caribbean islands Probably introduced in the late 1960s (Lever, 2005), A. viridigenalis was reported in Puerto Rico (T. Silva, pers. comm., 1985) and later confirmed by Raffaele et al. (1998), who recorded as many as 40 individuals, indicating an established population. Forshaw (1980) reported the presence of hundreds of A. ventralis, including hybrids with A. aestiva, breeding in Puerto Rico after releasing a shipment of traded birds. The Puerto Rican population is growing, unlike the native population in Hispaniola. Other established popula� tions are reported from St. Croix and St. Thomas (Virgin Islands) (Lever, 2005). Amazona amazonica is also present with an established population in Puerto Rico since the late 1960s (Owre, 1973; currently about 130 individuals: T. White, pers. comm., 2015) and in Martinique (Raffaele et al., 1998). A. oratrix was probably introduced in Puerto Rico in the early 1970s, but data on its breeding success are lacking (Lever, 2005). Eleven records of A. albifrons, possibly breeding, in groups of 3–11 individuals, are reported from 2001 to 2013 by eBird. South America Alien populations of A. aestiva are recorded in some South American cities outside the native range of this species. For example, flocks of 6–10 individuals have been observed in São Paulo and Porto Alegre, Brazil (J. L. Tella, pers. obs.). In Argentina, an alien population occurs in Buenos Aires, where individuals have been observed since the late 1990s. In 1999, 16 individuals were present; in 2015, 40 animals were counted (T. Calatoso, pers. comm., 2015). A group of 5 A. aestiva was recorded in 2002 in Río Cuarto (Argentina), but no recent observations are available (T. Calatoso, pers. comm., 2015). South Africa Symes (2014) compiled information on A. aestiva (up to 6 individuals) observed in Pinetown since 1989, where two pairs seem to breed sporadically but most chicks are poached from their nests. Apart from this small population, only one other Amazona sp. indi� vidual is recorded in Johannesburg.

Animal Biodiversity and Conservation 40.1 (2017)

A Total USA Germany Spain Japan Portugal Italy South Africa France UK Singapore

Per country

20,000 15,000 10,000 5,000 0

Total

30,000 20,000 10,000 0 1980

1985

1990

1995 Year

2000

2005

2010

Per country

8,000

Total USA Netherlands Singapore UK Japan Spain Germany Portugal South Africa Italy Belgium Mexico Hong Kong U. Arab Emirates

6,000 4,000 2,000 0

Total

8,000 6,000 4,000 2,000 0 1985

1990

1995 2000 Year

2005

2010

Fig. 3. Total of Amazona specimens imported globally per year, highlighting the two most traded species, A. aestiva (A) and A. ochrocephala/A. oratrix (B). Fig. 3. Total de individuos de Amazona importados en todo el mundo por año, se destacan las dos especies más comercializadas: A. aestiva (A) y A. ochrocephala/A. oratrix (B).

Temporal trends in global trade of Amazona spp. Between 1980 and 2013, a total of 372,988 traded wild Amazon birds were reported by CITES. A. aestiva was the most commonly traded species (288,112 in� dividuals), followed by A. ochrocephala (68,401 indivi� duals) (fig. 3). After the 1992 ban on wild–bird trade in CITES–listed species in USA, most (66%) of this trade was redirected to the European Union. A rapid increase in the number of globally traded A. aestiva individuals occurred from 37 birds in 1995 to over 5,000 individuals in 2004, at which point the number of recorded traded birds declined sharply to 374 in 2006. This sharp reduc� tion coincided with the first European ban on trade in wild birds in 2005, which became permanent in 2007. Between 1981 and 2007, before the EU ban on wild bird trade was implemented (Commission Regulation (EC) No. 318/2007), the predominant importers of A. aestiva were Portugal (7,991), Spain (5,551) and Italy (3,681). Only small numbers (74) were recorded as imported into the USA (fig. 3). As to A. ochrocephala, approximately 1,500 imports per year were recorded between 1996 and 2004. Subsequently, yearly recor� ded imports decreased to approximately 500 birds, apart from in 2012 when over 1,000 importations were recorded, possibly after the release of an animated feature movie with parrots as main characters ('Rio', from 20th Century Fox). The predominant importers of A. ochrocephala between 1981 and 2007 were the Ne� therlands (3,717), Singapore (3,216) and Spain (2,066), with small numbers (789) being imported into the USA (fig. 3). Amazon parrots are listed within the CITES Appendices (several species in Appendix I, which includes species whose trade should be controlled to avoid an unsustainable withdrawal from the wild). The earliest countries to record the trade of Amazon parrots by subscribing CITES were the USA and South Africa (1975), followed by UK (1976), France (1978), Portugal (1981), Belgium and the Netherlands (1984), Spain (1986), Singapore (1987) and Mexico (1991). Trade of CITES–listed wild birds was banned in 1992 in the USA, after which the EU remained responsible for about 87% of worldwide trade. In the EU, the first ban of wild bird trade occurred in October 2005 and become permanent in 2007. Discussion Our review showed that at least 14 species of Amazon parrots have been reported to be free–living outside their native ranges, with nine species having estab� lished alien populations in Europe (A. aestiva, A. oratrix and A. amazonica), Africa (A. aestiva), South (A. aestiva) and North America (A. aestiva, A. albifrons, A. amazonica, A. autumnalis, A. finschi, A. ochrocephala, A. oratrix, A. viridigenalis), and the Caribbean islands (A. ventralis, A. viridigenalis, A. amazonica and A. aestiva). The most widespread of these is A. aestiva, with at least 8 known alien populations. Our work showed that although Amazon parrots were widely traded as pets, a small number of introduced popula� tions occurs worldwide.

Mori et al.

A species is defined as 'invasive' if, once intro� duced, it spreads and exerts negative ecological im� pacts on native biodiversity (Genovesi & Shine, 2004). Prior to the trade bans imposed by US and by the European Union in Europe, most of the traded Amazon parrots were wild–caught, a factor which may have favoured the establishment of non–native populations (Carrete & Tella, 2008, 2015, 2016; Cabezas et al., 2013). The European Union has banned the trade of wild–caught individuals since 2005, allowing only the sale of captive–born parrots, which usually show lower invasiveness potential than their wild–caught counterparts (Gismondi, 1991; Carrete & Tella, 2015). Some illegal trade still occurs across the Mexi� co–USA boundary, although no information on the numbers of traded birds is available (Tella & Hiraldo, 2014). The illegal trade might have contributed to a much larger introduction and escape of birds and a higher establishment success and population growth in the most populated southern USA states (e.g., California, Florida and Texas). The establishment of non–native populations may be due to patterns of climate–matching between the native and intro� duced ranges (Ancillotto et al., 2015; Jackson et al., 2015; Cardador et al., 2016) and ecological niche expansion into colder climates (Strubbe et al., 2015). Our analysis showed that establishment success of Amazon parrots was not related to initial propagule pressure, although one cannot rule out the possibility that further releases/escapes after the first observa� tions would have helped alien populations to establish. Therefore, niche suitability may be more important for establishment success than propagule pressure (Cardador et al., 2016) for Amazon parrots. Accord� ingly, the most widespread Amazon species outside their native range are not only those most traded (A. aestiva, A. ochrocephala/oratrix and A. Viridigenalis, in this order), but also those showing the widest natural extent of occurrence (Forshaw, 1980). Although living mainly in densely forested areas, species with large extent of occurrence have evolved adaptations to cope with climatic conditions in their distribution ranges (Ancillotto et al., 2015; Menchetti et al., 2016). This may represent an adaptive feature in establishing alien populations outside the native range, i.e. where climatic conditions are different from those occurring within the core area of the extent of occurrence of the species (Duncan et al., 2003; Ancillotto et al., 2015). Main European introduced nuclei and isolated breeding instances occurred in warmest countries (e.g., Italy and Spain), while the only German popula� tion was first human–assisted (Bauer & Woog, 2008; Martens et al., 2013). In contrast, large populations of Amazon parrots are flourishing in southern USA and Puerto Rico, where climate is more similar to that of their native distributions (Hijmans et al., 2005; Toft & Wright, 2015). From a general perspective, the probability of establishing new populations is also related to propagule pressure, i.e. the number of individuals introduced, which is probably correlated to the number of traded animals, though this information is often lacking. As Amazon parrots are popular and expensive pets (Tella & Hiraldo, 2014), their presence

Animal Biodiversity and Conservation 40.1 (2017)

in natural environments outside the natural range is mainly due to unintentional escapes (Abellán et al., 2016). In Italy, an average of 3.4 Amazon parrots per year were recorded as lost or escaped over the last 10 years, with the largest numbers in the largest cities (fig. 1s in supplementary material). Although new non–documented releases may play a pivotal role in determining local population increases even without reproduction (fig. 1s in supplementary material), the observation of fledglings or juvenile individuals suggests that breeding may have occurred also where observation of nesting sites lacks. Alien populations of Amazon parrots grow up at very low rates, being long–lived, slow–reproducing species, suggesting that timely and successful control of these population is still feasible at the start of their establishment process (Edelaar & Tella, 2012). The population curve for A. oratrix in Germany showed a steeper trend than that of Italian populations, pos� sibly because this population is still fed by humans in urban parks (Martens et al. 2013). Impact exerted by European populations seems to be negligible or nearly absent, possibly because these nuclei are composed of a few individuals (Andreotti & Piacentino, 2009; Martens et al., 2013). Nevertheless, even for the largest populations in the USA, studies on the impact are still lacking. Further investigations should be carried out on other, often overlooked, typologies of impact, e.g., on parasites and potential diseases carried by introduced Amazon parrots (Menchetti & Mori, 2014; Mori et al., 2015). Despite these considerations, small and localized populations together with limited expansion rates pre� vent us from identifying the impact of Amazon parrots in Europe. Neither can we rule out the possibility that the impact of these parrots might be limited. Stud� ies on feeding ecology in Genoa (Italy: Andreotti & Piacentino, 2009) and Stuttgart (Germany: Martens et al., 2013) show a wide trophic spectrum for these par� rots, without any detectable impact on plants. Some food items containing poisonous compounds are only used by Amazon parrots, thus reducing competition for food resources with native birds (Martens et al., 2013). These alkaloid–rich, poisonous species (e.g., Taxaceae, Cupressaceae and Robinia pseudoacacia) may reach the 60% of the diet of A. oratrix in Stuttgart (Martens et al., 2013). In Europe, aggressive behav� iour towards jackdaws and rats has been observed in the vicinity of the nests, when chicks were present (Andreotti & Piacentino, 2009). Similarly, harassment of starlings and American crows by Amazon parrots was observed in California (Mabb, 1997, 2002). As to potential impact, Amazon parrots are con� sidered agricultural pests. For instance, in its native range, Amazona aestiva may damage up to 100% in� dividual fruit crop size (e.g., citrus orchards: Navarro et al., 1991). Other impacts by Amazon parrots included fungal and microbial infections in captive individuals, transmittable to humans and other animal species (De Freitas Raso et al., 2004; Romanov et al., 2006; Hannon et al., 2012). Observed harassment toward jackdaws and starlings in invaded regions seems to be the only certified impact of these parrots, although

no study has measured whether they affected the reproductive success of native species. Apart from any possible concerns due to invasion potential, in� troduced populations may have a conservation value (e.g., genetic pool) as reservoirs that could be used to rescue endangered populations in their native ranges (Bauer & Woog, 2008), e.g., A. oratrix in Stuttgart (Germany) and A. ventralis in Puerto Rico. It is im� portant to note that due to the frequent hybridization found between species co–occurring in the invaded regions, care should be taken before considering these populations valuable for conservation (e.g., for captive breeding or translocations). A growing body of global evidence recognizes biological invasions as one of the main drivers of the current biodiversity crisis (Wonham, 2006; Vilà et al., 2010; Scalera et al., 2012; Mazza et al., 2014). For instance, over 12,000 introduced species currently occur in Europe (DAISIE; www.europe–aliens.org/ aboutDAISIE.do [Accessed on 21st March 2016}. A total of 12 billion euros per year is required for damage caused by only 15% of introduced species in Europe (Kettunen et al., 2008). Genovesi & Shine (2004) proposed a 3–stage hierarchical approach to reduce the risks posed by introduced species, which includes: i) prevention of new introductions, ii) early detection of new establishments and iii) mitigation of impact through eradication or numerical control of populations. In contrast with other parrot species (e.g., Myiopsitta monachus and Psittacula krameri: Men� chetti & Mori, 2014; Menchetti et al., 2016), Amazon parrots are alien non–invasive species as their spread and impact on native environments seem to be low even after more than 30 years after the first release. Only a few species, i.e. mainly those with wide native ranges, have thrived outside their native range, even if their population growth seems to be mainly helped by new releases or escapes from captivity, rather than by breeding success. The reduction of propagules entering invasive Amazon parrot populations, after trade bans and CITES agreement, has further redu� ced the survival of alien populations. Nevertheless, with a precautionary principle approach, a continuous trend–monitoring would be recommended for all the established populations in order to follow the recom� mendations for the reduction of impact by alien parrots postulated by Menchetti & Mori (2014). Acknowledgements The authors are grateful to W. Rabitsch, who kindly translated the information on free–ranging Amazon parrots from the German literature. CBC data were provided by the National Audubon Society and through the generous efforts of Bird Studies Canada and countless volunteers across the western hemisphere. Thanks are due to A. Andreotti, N. Baccetti, G. Blanco, T. Calatoso, N. Fattorini, N. Ferrari, D. Franz, C. Gotti, D. Hernandez–Brito, A. Marangoni, D. Marty, A. Peruz, T. Silva, T. White and R. Zamora for the information and help provided. We acknowledge the support provided by European Cooperation in Science and Technology

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Animal Biodiversity and Conservation 40.1 (2017)

Supplementary material

Escaped A. aestiva A. albifrons A. amazonica

45.0

A. autumnalis A. barbadensis A. chrocephala

Latitude

Populations A. aestiva

42.5

40.0

37.5

Longitude

Fig. 1s. Distribution of reported escapes and free ranging populations of Amazon parrots in Italy. Fig. 1s. DistribuciĂłn de las fugas reportadas y de las poblaciones libres de loros del gĂŠnero Amazona en Italia.

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Animal Biodiversity and Conservation 40.1 (2017)

Nesting and feeding habits of the Indian giant squirrel (Ratufa indica) in Karlapat wildlife sanctuary, India A. K. Pradhan, S. Shrotriya, S. D. Rout & P. K. Dash

Pradhan, A. K., Shrotriya, S., Rout, S. D. & Dash, P. K., 2017. Nesting and feeding habits of the Indian giant squirrel (Ratufa indica) in Karlapat wildlife sanctuary, India. Animal Biodiversity and Conservation, 40.1: 63–69. Abstract Nesting and feeding habits of the Indian giant squirrel (Ratufa indica) in Karlapat wildlife sanctuary, India.— The Indian giant squirrel (Ratufa indica) is one of four species of giant squirrels in the world. It is endemic to India and its populations are severely fragmented. The ecology of squirrels in Asia has been little studied, hindering conservation and management efforts. We studied the Indian giant squirrel’s nesting and feeding habits during spring in the Karlapat Wildlife Sanctuary, India. We surveyed 122.5 km of natural trails for direct observation of these squirrels, their nests and feeding evidence, and we sampled plot–based quadrats to assess the availability of resources. We used Manly’s resource selection function and log–likelihood test ratios to analyse the data for preference. The mean encounter rate of the Indian giant squirrel was 0.57 (± 0.18 SD) individuals/km. Haldinia cordifolia (Wi = 4.899, p < 0.001) and Mangifera indica (Wi = 4.322, p = 0.001) were the preferred tree for nesting, whereas Xylia xylocarpa (31.30%) and Bauhinia vahlii (28.24%) were the most commonly eaten plants. Nest site preference was for taller tree species. As current management practices directly damage the preferred nesting sites and food resources, our findings aim to promote effective conservation of the Indian giant squirrel. Key words: Giant squirrel ecology, Forest management, Nest tree selection, Nest structure, Food preference, NTFP (non–timber forest products) collection Resumen Hábitos de nidificación y alimentación de la ardilla gigante hindú (Ratufa indica) en el refugio de vida silvestre de Karlapat, India.— La ardilla gigante hindú (Ratufa indica) es una de las cuatro especies de ardilla gigante del mundo. Se trata de una especie endémica de la India, pero sus poblaciones se encuentran muy fragmentadas. La ecología de las ardillas en Asia se ha estudiado poco, lo que ha entorpecido los esfuerzos de conservación y de gestión. Se estudiaron los hábitos de nidificación y alimentación de la ardilla gigante hindú durante la primavera en el refugio de vida silvestre de Karlapat, en la India. Se inspeccionaron 122,5 km de senderos naturales para la observación directa de estas ardillas, sus nidos e indicios de alimentación, y se muestrearon varios cuadrantes para evaluar la disponibilidad de recursos. Para analizar los datos sobre preferencias, se emplearon los índices de la función de selección de recursos de Manly y del logaritmo de la prueba de razón de verosimilitud. El promedio del índice de encuentros de la ardilla gigante hindú fue de 0,57 (± 0,18 DE) individuos/km. Haldinia cordifolia (Wi = 4,899; p < 0,001) y Mangifera indica (Wi = 4,322; p = 0,001) fueron los árboles preferidos para la nidificación, mientras que Xylia xylocarpa (31,30%) y Bauhinia vahlii (28,24%) fueron las plantas de las que se alimentaron en mayor medida. La preferencia del sitio de nido fue por especies arbóreas más altas. Como las actuales prácticas de gestión dañan directamente los sitios preferidos de nidificación y los recursos alimentarios, con nuestras conclusiones tratamos de promover la conservación eficaz de la ardilla gigante hindú. Palabras clave: Ecología de la ardilla gigante, Gestión forestal, Selección de árboles para la nidificación, Estructura de los nidos, Preferencia alimentaria, Recogida de productos forestales no madereros Received: 15 II 15; Conditional acceptance: 17 IV 15; Final acceptance: 4 X 16 Anup K. Pradhan & Shivam Shrotriya, Wildlife Inst. of India, P. O. Box 18, Chandrabani, Dehradun 248001 (India).– S. D. Rout, Dept. of Wildlife and Conservation Biology, North Orissa Univ., Mayurbhanj, Odisha (India).– Prasad Kumar Dash, Dept. of Forest and Environment, Odisha Biodiversity Board, Nayapalli, Bhubaneswar, Odisha (India). Corresponding author: S. Shrotriya. E–mail: shivam@wii.gov.in ISSN: 1578–665 X eISSN: 2014–928 X

Pradhan et al.

Introduction

Data collection

The Indian or Malabar giant squirrel (Ratufa indica Erxleben, 1777) is endemic to the Peninsular India (South India) (Corbet & Hill, 1992). Although it is widely distributed within its range, it occurs in severely fragmented populations (Molur et al., 2005). It has faced local extinction and range restriction in several areas due to hunting and habitat loss and suitable habitat is limited in the areas where it occurs (Molur, 2016). The Indian giant squirrel is currently listed in the ‘Least Concern’ category of IUCN Red List, Appendix II of CITES and Schedule II of the Wildlife (Protection) Act, 1972 of India (Favre, 1989; Molur, 2016). The Indian giant squirrel occurs in the elevation range of 180 to 2,300 m and inhabits deciduous, mixed deciduous and moist evergreen forests (Prater, 1980). It is a large–bodied (90 to 100 cm), diurnal and arboreal squirrel (Hayssen, 2008). A solitary living species, it is seen in pairs only during the breeding season. It usually constructs more than one nest, or drey, within a single breeding season. The nests, which are made of leaves and twigs, are built in tall, profusely branched trees, in the higher canopy (Borges, 1989; Ramachandran, 1992). The species is omnivorous and feeds on fruits, flowers, nuts, bark, bird eggs and insects (Payne, 1979; Ramachandran, 1992). The ecology of squirrels from Asian countries has been little studied and published information is scarce (Pradhan et al., 2012). We studied the nesting and feeding preferences of the Indian giant squirrel in the Karlapat wildlife sanctuary, India, the easternmost limit of their distribution. We discuss threats to the species in our study area and envision that the information herein will contribute to the conservation of the Indian Giant Squirrel.

Data were collected from February 2009 to April 2009 (spring) when the squirrels are more active and easily seen. We searched for animals and their nests along the natural trails in the forest. The trails were selected using geographic information system (GIS) attempting to cover the entire sanctuary and keeping the trails spatially equidistant. A total distance of 122.5 km was walked over 55 trails of about 2 km each. The trails were walked in morning (06:00–10:00 hr) or evening (16:00–18:00 hr) hours when activity of the species is high. Feeding behaviour was recorded by direct observation and from food found in nests. Nest location was recorded using GPS. We also collected data on nesting tree species, height of the tree and height of the nest from the ground. Data on the nest building material and dimensions of the nest were collected only when nests were empty. Plot based random quadrat sampling was conducted to assess the availability of nesting tree species (Mishra, 1968). At least 20 plots (15 x 15 m2) were sampled for each of the six administrative blocks. Plants above 10 cm in girth at breast height were considered trees and categorised as dry deciduous, moist deciduous and semi–evergreen types.

Material and methods Study area Karlapat wildlife sanctuary (19° 36' 50'' to 19° 50' 51'' North and 82° 56' 18'' to 83° 19' 35'' East) is located in Kalahandi district of Odisha, India (fig. 1). The sanctuary covers an area of 175.50 km2 and altitude ranges from 400 to 996 m. According to the biogeographic classification of India, the sanctuary falls within the Deccan Peninsula Eastern highlands province close to the eastern coast (Rodgers et al., 2002). Geology of the region features patches of granite and bauxite rocks. The vegetation of the sanctuary is a mosaic of moist peninsular sal (Shorea robusta), moist mixed deciduous, dry deciduous and riparian semi–evergreen forests (Champion & Seth, 1968). The major fauna of the sanctuary are Leopard (Panthera pardus), Sloth bear (Melursus ursinus), Indian elephant (Elephas maximus), Indian bison (Bos gaurus), Sambar deer (Rusa unicolor), Barking deer (Muntiacus muntjak), Mouse deer (Tragulus meminna), Smooth–coated otter (Lutra persipicillata) and Indian Pangolin (Manis crassicaudata). The sanctuary is divided into six administrative blocks, and contains 17 villages.

Data analysis Preference for nesting trees was analysed using Manly’s resource selection function (Manly et al., 2002). The data on nesting tree availability and use were obtained and treated in design I study framework of habitat use, where availability and use of the resource by individuals are not identified separately (Thomas & Taylor, 1990). Evidence for the non–random selection of food plants was tested using log–likelihood test ratios (Karanth & Sunquist, 2000). Summary statistics were calculated wherever appropriate. All statistical analysis were performed in the program 'R', version 3.3.0 and package 'adehabitatHS', version 0.3.12 (Calenge, 2006; R Core Team, 2016). Results We sighted a total of 70 Indian giant squirrels during the trail surveys, corresponding to a mean encounter rate of 0.57 (± 0.18 SD) individuals/km. We recorded 277 nests built in 37 tree species. The Indian giant squirrel built significantly more nests in dry deciduous trees (61.07%) than in semi–evergreen (30.15%) and moist deciduous trees (8.83%) (Χ2 = 12.584, df = 2, p = 0.002; table 1). Resource selection function revealed a preferential selection for nesting tree species (Khi2L = 199.288, df = 36, p < 0.001). Analysis showed that Haldinia cordifolia (Wi = 4.899, p < 0.001) and Mangifera indica (Wi = 4.322, p = 0.001) were the most preferred nesting tree species whereas Cassia fistula (Wi = 0.096, p < 0.001) and Mallotus philippensis (Wi = 0.137, p < 0.001) were the least preferred species (fig. 2). Most nests were built on Terminalia alata (11.03%) and Anogeissus latifolia

Animal Biodiversity and Conservation 40.1 (2017)

83º 0' 0'' E

83º 10' 0'' E Kalahandi North Division

Junagarh Range

19º 40' 0'' N

Karlapat Wildlife Sanctuary

TH Rampur Range

2.5

10 km

83º 0' 0'' E

83º 10' 0'' E

Fig. 1. Location and map of Karlapat wildlife sanctuary, Odisha, India. Black triangles represent villages inside the sanctuary. Fig. 1. Ubicación y mapa del refugio de vida silvestre de Karlapat, en Odisha, la India. Los triángulos negros representan las aldeas dentro del refugio.

(8.82%) trees, which were also among the preferred nesting trees (Wi = 1.925 and 2.073, p = 0.004 and 0.008, respectively). The mean height of nesting trees was 11.08 (± 2.11 SD) m, and an average height of the nests from the ground was 9.64 (± 2.04 SD) m.

We observed that about 80% of the nesting trees had an association with climber plants species such as Bauhinia vahlii, Combretum decandrum, Entada rheedii and Calycopteris floribunda. The squirrels picked most of the nesting materials from the nes-

Table 1. Availability and use of nesting trees and height of nesting trees and nests in different tree types of Karlapat wildlife sanctuary. Tabla 1. Disponibilidad y uso de árboles para la nidificación y altura de los mismos y de los nidos en distintos tipos de árbol del refugio de vida silvestre. Tree type No. of species

Dry deciduous

Moist deciduous

Semi–evergreen

Availability of trees (%)

54.07

16.46

29.47

Use for nesting (%)

60.65

8.66

30.69

168

Tree height (m ± SD)

11.26 ± 2.26

10.66 ± 1.75

10.81 ± 1.86

Nest height (m ± SD)

9.8 ± 2.18

9.29 ± 1.70

9.42 ± 1.81

No. of nests

Pradhan et al.

Table 2. Preferred food plants of the Indian giant squirrel in Karlapat wildlife sanctuary: Fr. Feeding records (%); * Used for nesting. Tabla 2. Alimentos vegetales preferidos de la ardilla gigante hindú en el refugio de vida silvestre de Karlapat: Fr. Registros de alimentación; * Utilizado para anidar.

Food species

Fr (%)

Xylia xylocarpa

31.30

Bauhinia vahlii

28.24

Terminalia alata

9.92

Shorea robusta Mangifera indica

Food species *

Fr (%)

Cleistanthus collinus

0.76

Madhuca indica

0.76

Mitragyna parvifolia

0.76

9.16

Zanthoxylum armatum

0.76

6.87

Schleichera oleosa

0.76

Albizia lebbeck

3.05

Stereospermum suaveolens

0.76

Butea monosperma

1.53

Syzigium cuminii

0.76

Garuga pinnata

1.53

Tamarindus indica

0.76

Terminalia arjuna

1.53

Terminalia belerica

0.76

ting tree and associated climber species. Stem and leaves of B. vahlii were used in 86% of the nests. We measured the dimensions of 14 intact and abandoned nests. The mean length, width and depth of the nests were 27.35 (± 4.87 SD), 17.24 (± 3.40 SD) and 17.44 (± 3.33 SD) cm, respectively. Circular opening of the nests had a diameter of 5.09 (± 0.48 SD) cm. We collected data on feeding through 131 direct and indirect records. Log–likelihood test ratio showed that evidence for non–random selection of food plants was significant (G2 = 236.98, df = 17, p < 0.001). Xylia xylocarpa (31.30%) and B. vahlii (28.24%) were the most frequently used feeding plants, implying their importance (table 2). We observed the squirrels eating new leaves of Tamarindus indica, ripe fruits of X. xylocarpa, Ficus recimosa, B. vahlii, Schleichera oleosa, M. indica and T. alata and flowers of Butea monosperma. We also found other food materials nibbled upon by the squirrels inside the nests (fig. 3). Discussion The encounter rate or direct observations of the Indian giant squirrel varied little across the sanctuary area. However, the encounter rate was high in the pristine forest patches with dense canopy cover. Preference for nesting trees could depend on factors such as access to nesting material and food, nest safety and the branching pattern of the tree species. Our data showed that tree height was a major factor governing nesting tree selection. The Indian giant squirrel consistently selected tall trees with a mean height of 11.08 (± 2.11 SD) m. A nesting preference for dry deciduous trees could also be attributed to tree height (table 1). The nesting trees with high selection ratios were comparatively taller species with a long main trunk topped by a large crown

with many branches. Tree squirrels are known to build nests on trees with interlinking crowns, allowing easy access and movement in the canopy (Patton, 1975; Hall, 1981; Ramachandran, 1992). Datta & Goyal (1996) also observed that the Indian giant squirrel preferred large trees as nesting sites, probably to avoid predators. Kumara & Singh (2006) sighted the Indian giant squirrel mostly at a height of 16 to 20 m in moist forests and 11 to 15 m in dry forests. We observed the Indian giant squirrel nesting on a large variety of the tree species (n = 37) in Karlapat wildlife sanctuary. Kanoje (2008) also reported the use of a large variety of tree species (n = 30) for nesting in Sitanadi wildlife sanctuary, India. Some of the preferred nesting trees, such as M. indica, Albizia lebbeck, Syzygium cumini and T. alata, also provided food material (table 2). An association with climber plants, which provide nest building resources, could influence the selection of nesting trees. B. vahlii was a crucial climber species as it was the most frequently used nesting material and also a preferred plant food (table 2). The squirrels constructed their nests on a forked branch by interweaving climber plant stems and twigs and padding it with the leaves. Entrance to the globular (or oval) shaped nest was horizontal. Size and shape of the nests varied from nest to nest, but the diameter of the entry was consistent at 5.09 (± 0.48 SD) cm. The height of the nest from the ground depended on the height of the tree, which is evident from the similar standard deviation for both (table 1). However, it is interesting to note that the distance of the nest from the top of the tree was averaged at 1.185 (± 0.48 SD) m. The nests were not built on the highest possible branch, as the squirrels sought cover above the nest. Such cover might help avoid direct heat from the sun and serve as a hiding–place from birds of prey (Pradhan et al., 2012).

Zanthoxylum armatum

Xylia xylocarpa

Terminalia belerica

Terminalia arjuna

Terminalia alata

Tamarindus indica

Syzigium cuminii

Stereospernum suaveolens

Shorea robusta

Schleichera oleosa

Mitragyna parvifolia

Mangifera indica

Madhuca indica

Garuga pinnata

Cleistanthus collinus

Butea monosperma

Bauhinia vahlii

80%

70%

60%

50%

40%

30%

20%

10%

Fig. 3. Plant species and their parts used as food by the Indian giant squirrel.

Fig. 3. Especies vegetales y sus partes utilizadas como alimento por la ardilla gigante hindú.

Bark

Cassia fistula

Mallotus philippensis

Dalbergia sissoo

Terminalia arjuna

Buchanania lanzan

Cleistanthus collinus

Diospyros melanoxylon

Gmelina arborea

Grewia tiliiofolia

Bridelia retusa

Diospyros montana

Radermachera xylocarpa

Aegle marmelos

Mitragyna parvifolia

Madhuca indica

Shorea robusta

Tamarindus indica

Lagerstroemia parviflora

Pterocarpus marsupium

Ficus recimosa

Stereospermum suaveolens

Casearia elliptica

Xylia xylocarpa

Terminalia belerica

Sterculia villosa

Garuga pinnata

Terminalia chebula

Albizia chinensis

Desmodium oojeinense

Terminalia alata

Schleichera oleosa

Anogeissus latifolia

Syzygium cumini

Diospyros malabarica

Albizia lebbeck

Mangifera indica

Haldina cordifolia

Albizia lebbeck

Selection ratios (± SE)

Animal Biodiversity and Conservation 40.1 (2017) 67

Fig. 2. Manly’s resource selection function for nesting tree preference of the Indian giant squirrel. Plant species above selection ratio one are preferred for nest building.

Fig. 2. Función de selección de recursos de Manly para las preferencias de árboles para la nidificación de la ardilla gigante hindú. Las especies vegetales cuyo cociente de selección es superior a uno son las preferidas para la construcción del nido.

100%

90%

Fruits

Flowers

Leaves

Seeds

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We observed that the Indian giant squirrel depended on X. xylocarpa (31.30%) and B. vahlii (28.24%) for the bulk of its diet (table 2). This finding could be the result of the higher availability of these plants in the forest and the fact that the survey season matched with the fruiting of both species. Baskaran et al. (2011) found that the principal food source of the Indian giant squirrel was Tectona grandis (41%) in Mudumalai wildlife sanctuary, southern India. Nevertheless, it could be inferred that X. xylocarpa and B. vahlii are valuable food sources of the Indian giant squirrel in spring. B. vahlii also provided a greater variety of food items than other species (fig. 3). The Indian giant squirrel depends on a variety of plant species and their different parts to meet its nutritional needs. Similar dietary patterns of consuming different plant parts are reported from other parts of its range as well (Borges et al., 1992; Kanoje, 2008; Baskaran et al., 2011). Karlapat wildlife sanctuary faces severe pressure from the collection of non–timber forest products (NTFP) collection and demands for wood in Kalahandi district. Fruits of M. indica and S. cumini and leaves of Diospyros melanoxylon are among the top NTFP collections, which are also the preferred nesting trees of the Indian giant squirrel (fig. 2). Leaves of B. vahlii and S. robusta, seeds and latex of S. robusta, and flowers and fruits of Madhuca indica, which are also extracted heavily, are all central to the diet of the Indian giant squirrel (table 2). Each month, 500 to 1,000 large, old trees are cut down to sell as part of crop management by the forest department in the surrounding areas (Kalahandi Forest Department, pers. comm.). Heavy grazing also hinders forest regeneration within the sanctuary. The wood industry requires quality timber from tall, mature trees, trees that the Indian giant squirrel specifically need for nesting. Datta & Goyal (1996) also found that the Indian giant squirrel depends mainly on large, mature trees for feeding. Another issue of concern is the increased hunting of the Indian giant squirrel for meat consumption by locals in Karlapat wildlife sanctuary (Pradhan et al., 2012). The threats to this squirrel population in Karlapat wildlife sanctuary are immediate and visible. The results of this study support the need to implement the following conservation measures for the Indian giant squirrel: prevention of lopping and felling of mature, tall trees of the preferred nesting species; prevention of forest fires and mitigation of heavy grazing to allow regeneration of trees; regular monitoring of NTFPs; conservation of food plants such as X. xylocarpa and B. vahlii, and strict implementation of the law to minimise hunting of the squirrel. Karlapat wildlife sanctuary holds one of the easternmost populations of this endemic mammalian species in India. The sanctuary is a natural mosaic of different forest types and effective conservation management could have a positive, long–lasting impact on the population of the Indian giant squirrel. Acknowledgements We are thankful to the Principal Chief Conservator of Forest, PCCF (wildlife), Odisha for granting research

permission. Dr. Susil Kumar Dutta and Dr. H. K. Sahoo, Department of Zoology, North Orissa University, and Dr. L. A. K. Singh, Senior Research Officer, Forest Department, Odisha are acknowledged for their input during field surveys. Dr. Prautush Mahapatra and Rupa Majhi, forest department officials, were a great support for logistics. We also thank the anonymous reviewers for their suggestions and improvement of the article. References Baskaran, N., Venkatesan, S., Mani, J., Srivastava, K. S. & Desai, A. A., 2011. Some aspects of the ecology of the Indian giant squirrel Ratufa indica (Erxleben, 1777) in the tropical forests of Mudumalai Wildlife Sanctuary, southern India and their conservation implications. Journal of Threatened Taxa, 3(7): 1899–1908. Borges, R. M., 1989. Resource heterogeneity and the foraging ecology of the Malabar giant squirrel (Ratufa indica). Ph D Thesis, University of Miami, Florida. Borges, R. M., Mali, S. & Ranganathan, S., 1992. The status, ecology and conservation of the Indian giant squirrel (Ratufa indica). Technical Report No. 1. Wildlife Institute of India, Dehradun, India. Calenge, C., 2006. The package adehabitat for the R software: a tool for the analysis of space and habitat use by animals. Ecological Modelling, 197: 516–519. Champion, H. G. & Seth, S. K., 1968. A Revised Survey of the Forest Types of India, Govt. of India Press, New Delhi. Corbet, G. B. & Hill, J. E., 1992. The mammals of the Indo–Malayan region. Natural History Museum Publications, Oxford University Press, Oxford. Datta, A. & Goyal, S. P., 1996. Comparison of forest structure and use by the Indian giant squirrel (Ratufa indica) in two riverine forests of Central India. Biotropica, 28: 394–399. Favre, D. S., 1989. International trade in endangered species guide to CITES. Martines Nijhoff Publishers, London. Hall, J. G., 1981. A field study of the Kaibab squirrel in Grand Canyon National Park. Wildlife Monographs, 75: 1–154. Hayssen, V., 2008. Patterns of body and tail length and body mass in sciuridae, Journal of Mammalogy, 89(4): 852–873. Kanoje, R. S., 2008. Nesting sites of Indian giant squirrels in Sitanadi Wildlife Sanctuary, India. Current Science (Bangalore), 7: 882–884. Karanth, K. U. & Sunquist, M. E., 2000. Behavioural correlates of predation by tiger (Panthera tigris), leopard (Panthera pardus) and dhole (Cuon alpines) in Nagarahole, India. Journal of Zoology (London), 250: 255–265. Kumara, H. N. & Singh, M., 2006. Distribution and relative abundance of giant squirrels and flying squirrels in Karnatak, India. Mammalia, 70: 40–47. Manly, B. F., McDonald, L. L., Thomas, D. L., McDonald, T. L. & Erickson, W., 2002. Resource Selection

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by Animals: Statistical Design and Analysis for Field Studies. Kluwer Academic Publishers, Dordrecht. Mishra, R., 1968. Ecology Work Book. Oxford and IBH Publishing Company, Calcutta. Molur, S., 2016. Ratufa indica. The IUCN Red List of Threatened Species 2016: 6.T19378A22262028 [Accessed on 05 September 2016]. Molur, S., Srinivasalu, B., Srinivasalu, C., Walker, S., Nameer, P. O. & Ravi Kumar, L., 2005. Status of South Asian Non– Volant small mammals: Conservation Assessment and Management Plan (C. A. M. P.) workshop report. Zoo Outreach Organisation/ CBSG– South Asia, Coimbatore, India. Patton, D. R., 1975. Abert squirrel cover requirements in south–western ponderosa pine. USDA Forest Service Research Paper RM–145, Fort Collins, Colorado. Payne, J. B., 1979. Abundance of diurnal squirrels at the Kuala Lompat post of the Krau Game Reserve, Peninsular Malaysia. In: Abundance of animals in Malaysian rain Forests: 37–51 (A. G. Marshall, Ed.). Department of Geography, University of Hull.

Pradhan, A. K., Shrotriya, S. & Rout, S. D., 2012. Observation on Nest–Site selection by Indian giant squirrel in Karlapat Wildlife Sanctuary, Odisha. Small Mammal Mail, 4(2): 12–13. Prater, S. H., 1980. The Book of Indian Animals. Bombay Natural History Society and Oxford University Press, Mumbai, India. R Core Team, 2016. R: A language and environment for statistical computing. R Foundation for Statistical Computing, Vienna, Austria. URL https:// www.R–project.org/ Ramachandran, K. K., 1992. Certain aspects of ecology and behaviour of Malabar giant squirrel Ratufa indica (Schreber). Ph. D. Thesis, University of Kerala. Rodgers, W. A., Panwar, H. S. & Mathur, V. B., 2002. Wildlife Protected Area Network in India: A Review (Executive summary). Wildlife Institute of India, Dehradun. Thomas, D. & Taylor, E., 1990. Study designs and tests for comparing resource use and availability. Journal of Wildlife Management, 54: 322–330.

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Living on the edge: regional distribution and retracting range of the jaguar (Panthera onca) G. A. E. Cuyckens, P. G. Perovic & M. Herrán

Cuyckens, G. A. E., Perovic, P. G. & Herrán, M., 2017. Living on the edge: regional distribution and retracting range of the jaguar (Panthera onca). Animal Biodiversity and Conservation, 40.1: 71–86. Abstract Living on the edge: regional distribution and retracting range of the jaguar (Panthera onca).— To preserve biodi� versity we need to understand how species are distributed and which aspects of the environment determine these distributions. Human–induced changes in land–cover and loss of habitat threaten many species, particularly large carnivores, in many parts of the world. Differentiating the influence of climate and human land use on the distribu� tion of the jaguar (Panthera onca) is important for the species’ conservation. Historically distributed from the United States to southern Argentina, the jaguar has seen its distribution range decreased at regional and local scales. Here we predict the species’ distribution range using historical records of its presence, climate variables, and MaxEnt predictive algorithms. We focus especially on its southernmost limit in Argentina to indicate the historical limits of this species, and describe its present niche in these edge populations. To estimate the effect of human activity we used a raster of land cover to restrict the jaguar’s distribution. We collected a large amount of presence records through the species’ historical range, and estimated a historical regional distribution ranging from Patagonia up to latitude –50°S. Our findings show the range of the jaguar is decreasing severely in its southern limit and also in its northern limit, and that changes in land cover/use are threats to the species. After subtracting non–suitable land–cover from the studied niche, we found the environmentally suitable area for the jaguar in the study area has decreased to 5.2% of its original size. We thus warn of the high extinction risk of the jaguar in Argentina. Key words: Habitat suitability, Land–cover, MaxEnt, Species distribution models (SDM), Ecological Niche Factor Analysis (ENFA) Resumen Vivir al límite: distribución regional y superficie ocupada por el jaguar en retroceso (Panthera onca).— Para conservar la biodiversidad, es necesario entender cómo se distribuyen las especies y qué variables ambientales determinan dicha distribución. Los cambios inducidos por el hombre en la ocupación del suelo y la pérdida de hábitat ponen en peligro a numerosas especies de todo el mundo, especialmente grandes carnívoros. Diferenciar la influencia del clima y la de los usos del suelo en la distribución de jaguar (Panthera onca) es importante para su conservación. Esta especie, que tradicionalmente se distribuía desde los Estados Unidos hasta el sur de Argentina, ha visto reducida su distribución a escala regional y local. En este trabajo prede� cimos el rango de distribución de la especie utilizando registros de presencia histórica, variables climáticas y algoritmos predictivos obtenidos con MaxEnt. Nos centramos especialmente en su límite más austral en Ar� gentina para indicar los límites históricos de esta especie y describir el nicho que ocupa actualmente en estas poblaciones marginales. Para estimar el efecto de las acciones antrópicas, utilizamos una capa de ocupación del suelo para limitar la distribución del jaguar. Recopilamos una buena cantidad de registros de presencia en todo el área de distribución histórica de la especie y estimamos una distribución regional histórica desde la Patagonia hasta los –50° de latitud. Nuestros resultados ponen de manifiesto que el área de distribución del jaguar se está contrayendo de forma alarmante en el límite meridional y también en el septentrional, y que los cambios de ocupación y de uso del suelo son una amenaza para la especie. Tras restar del nicho estudiado la ocupación del suelo que no es adecuada, descubrimos que la superficie idónea para el jaguar desde el punto de vista ambiental en la zona del estudio se ha reducido hasta el 5,2% de su tamaño original. Por consiguiente, advertimos del elevado riesgo de extinción que acecha al jaguar en Argentina. ISSN: 1578–665 X eISSN: 2014–928 X

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Palabras clave: Idoneidad del hábitat, Ocupación del suelo, MaxEnt, Modelos de distribución de especies, Análisis factorial de nicho Received: 2 III 15; Conditional acceptance: 15 IV 15; Final acceptance: 5 X 16 Griet A. E. Cuyckens, Consejo Nacional de Investigaciones Científicas y Tecnológicas (CONICET) y Centro de Estudios Territoriales Ambientales y Sociales (CETAS).– Pablo G. Perovic, Administración de Parques Nacionales, Delegación Noroeste.– M. Herrán, Universidad Nacional de Salta (UNSa). Corresponding author: G. A. E. Cuyckens. E–mail: grietcuyckens@yahoo.com

Animal Biodiversity and Conservation 40.1 (2017)

Introduction Habitat loss and degradation are the main threats to mammals and can cause considerable range contrac� tions (Schipper et al., 2008). Large carnivores, in par� ticular, are highly affected by such threats (Rabinowitz & Zeller, 2010) because they need extensive surfaces and have low population densities. Some carnivore species can inhabit the anthropogenic habitats that have replaced natural landscapes, but in such areas they are especially vulnerable to killing associated with livestock depredation (McLellan, 1990) and direct hunt� ing. Edge populations (groups of individuals near the boundaries of their geographic ranges) are particularly vulnerable to extinction due to stochastic factors (Sodhi et al., 2009) because abundances in the periphery of the distribution of species tend to decrease and unusual (and often random and detrimental) events become prominent when population sizes are small. Habitat fragmentation causes gaps in distribution ranges, reducing the area of occupancy �� (IUCN Stan� dards and Petitions Subcommittee, 2011), depending on the species’ ability to use the surrounding matrix of habitats. In the management and conservation of species it is fundamental to determine the species’ niche in order to describe habitat requirements and estimate the amount and arrangement of suitable habitats in a landscape (Araújo �� & Guisan, 2006; Mill� spaugh & Thompson, 2009). Predictive modelling of species distributions is a useful tool for answering practical questions in applied ecology and conservation biology (Guisan & Thuiller, 2005). Species distribution modelling (SDM) is the most widely used approach to answer questions about the present and historical distributions of a species. Combined with geographic information systems (GIS) and land cover data, SDM enable geographical analysis of species distribution ranges (Guisan & Thuiller, 2005; Kearney & Porter, 2009). It is important to understand the underlying principles and assumptions of SDM. A convenient and practical postulation is to assume that the modelled species is in pseudo–equilibrium with its environment (Guisan & Theurillat, 2000). In practice, few species are in equilibrium with their environment and the retraction or expansion of species can violate this principle. Furthermore, SDM does not solve the conservation problems related to the Wallacean shortfall (Whittaker et al., 2005); i.e. insufficient knowledge of distributions could have consequences on conservation prioritisation, and local or smaller scale conservation assessments are still required. It is also important to bear in mind the reliance of these techniques on the niche concept (Guisan & Zimmermann, 2000). The niche concept can help understand geograph– ical distributions of species. Nevertheless, this concept is not only one of the most confusing terms in ecology but also one of the most extensively discussed and rethought (Morrison & Hall, 2002; Mitchell, 2005). Its definition is a subject of debate (Milesi & Lopez de Casenave, 2005) and it is difficult to determine which niche theory is applied to species distribution modelling (Shenbrot, 2009). Soberón & Peterson (2005) clarified this confusion by introducing the BAM–diagram which

indicates that the occupied distributional area of a species is the intersection of the area where biotic variables are suitable (B), the area where scenopoetic variables are suitable (A) and the area accessible for this species (M); (for more details see Soberón & Peterson, 2005). Jackson & Overpeck (2000), working with scenopoetic environmental spaces and following the main ideas of Hutchinson, defined the 'fundamental niche' as the subset of the environmental space defined by the n dimensions that describe the suite of combinations of variables that permit survival and reproduction of individuals. The 'realized niche' of Hutchinson (1957) represents the part of the existing fundamental niche that remains habitable after reductions caused by competitors and other negative interactors (Soberón, 2007). The portion of the fundamental niche that actually exists somewhere in the study region at the time of analysis is 'the existing fundamental niche' (Soberón 2007), also called the potential niche (Jackson & Overpeck, 2000). Species distribution models estimate niche–related objects along a continuum between the existing fundamental niche and the realized niche (Jiménez–Valverde et al., 2008). The most relevant ecological factors shaping habitat suitability can be identified by ecological–niche factor analysis (ENFA) (Hirzel et al., 2002; Basille et al., 2008; Calenge & Basille, 2008). This analysis identifies �� the response of a species to the main envi� ronmental variations in a study area (Rotenberry et al., 2006), thereby reflecting its realized niche (Braunisch et al., 2008). This niche can be geographically projected (Guisan & Zimmermann, 2000) and the resulting map summarises environmental suitability across the landscape (i.e. an estimate of the abiotically suitable area). Additional refinements to and analyses of model outputs can be used to estimate the occupied distributional area (Peterson, 2011). The distributional areas of a species are subsets of geographic space in which the presence of individuals or populations of a species can be detected (Soberón, 2007). Some other areas, lacking observable populations or individuals, but otherwise suitable, can also be defined (Peterson, 2011). The jaguar (Panthera onca) is a large and widely distributed felid, which, according to the IUCN, is considered to be Near–Threatened due to habitat loss and persecution (Caso et al., 2008). Historically, it was distributed from the Southwestern United States to southern Argentina. In its northern limit in the USA, the species underwent range contractions and it has almost disappeared there since the 1990s (Van Pelt & Johnson, 2002). In the southern limit, its range has been decreasing since the 19th century (Arra, 1974) and it is currently extinct in Uruguay (González & Martínez Lanfranco, 2010) and El Salvador (Sanderson et al., 2002b). Changes in land cover have caused local extinctions and led to further fragmentation of its distribution (Swank & Teer, 1989; Koford, 1991; Sanderson et al., 2002b). The jaguar is more sensitive to habitat transformation than other big felids such as the puma (Puma concolor) (De Angelo, 2011). The historical southernmost distribution limit is unknown,

but the species could have inhabited landscapes as far south as the Negro River at latitude 40ºS (Carman, 1984) or the Colorado River, at 39ºS (Lehmann– Nitsche, 1907) (fig. 1s in Supplementary material). The ancestor of the current species, P. onca augusta, was present during the Pleistocene and Rancholabreano in Chilean and Argentine Patagonia (Borrero, 2001), but it is not clear if today’s species reached these latitudes even though it was reported in historical documents (see Diaz, 2010). The lack of information on species' distribution (dubbed the Wallacean Shortfall) can have consequences on conservation prioritisation (Whittaker et al., 2005). It may be particularly important for fringe populations that are extremely vulnerable to extinction and of utmost importance for the focus of conservation actions. For this reason, we assessed the geographical aspects of the conservation status of the jaguar in Argentina, the country with the southernmost occurrence of this species. The jaguar today is only present in the northern portion of the country (as far as latitude 55° south in the Misiones province), with rapid local extinctions reducing its local distribution (Perovic & Herrán, 1998; Altrichter et al., 2006; Quiroga et al., 2014). Argentina is an agricultural–oriented country with a strong soybean production (Gudynas, 2008; Izquierdo & Grau, 2009). This involves transforming land to use that is incompatible with the jaguar’s ecological requirements. To protect the most adequate areas for conservation and management of the species and its habitat, several issues must be addressed: the decreasing regional distribution, the doubts regarding the southernmost limit of the species, and the loss of suitable areas for jaguar in Argentina due to changes in land cover/use. The main aim of this study was to model the historical niche of the jaguar in its entire range using SDM and records from the bibliography. We aimed to model how its distribution has changed over the last centuries. To do so, we modelled the present niche in Argentina using only current presence records and compared this with the historical niche. We estimated the impact of changes in human land use by subtracting non suitable areas from the distribution map. We also attempted to clarify the southern–most historical distribution limit in Argentina. Methods We gathered presence records of the jaguar throughout the species’ historical range from published literature, grey literature, publicly–available databases, museum collections, and previous works by some of the authors, and unpublished field data kindly shared by colleagues (see table 1s in supplementary material for a complete list). Distributions recorded across dif� ferent time periods can be used to test predictions of range shifts over time (Araújo et al., 2005; Martínez– Meyer & Peterson, 2006). To do this we separated the records arbitrarily into historical records (from 1741 to 2011) and present records for Argentina (from 1994 to 2011). In total we obtained 1,447 presence records for the entire range of the jaguar (from 1741 to 2011).

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To avoid errors in identifying the species we carefully checked all records (Newbold, 2010) using different methods. For published historical records, we read the description of the original authors; we excluded records that were not clear (for example, if only 'big felid' was mentioned) or when records were based on general place names (for example, 'El Tigre'). Pres� ence indicated by Diaz (2010) in Chilean and Argen� tine Patagonia based on place names and historical stories were revised carefully, but we used only three as the others possibly implied observer bias or mis� identification of species. For databases we carefully checked the metadata and we did not use records flagged as dubious. For museum records we used records previously checked by colleagues for correct species identification, and the museum records of the Argentinean museums were checked by ourselves. Finally, we filtered historical records using a shapefile of the historical species' distribution (kindly provided by the Panthera Foundation, and hereafter called the Historical Known Range); no historical records outside of these limits were used for modelling. We used all the records provided by felid �� specialists. To georefer� ence records, we used the coordinates provided with each record; when not provided, we digitalized the record based on maps given in the publication (i.e. Carman, 1984). Museum records were georeferenced with gazeteers, but we used only those records that could be georeferenced with an error of < 1 km. As our presence records were spatially biased because they were not gathered following a standardised method, we used a bias file in MaxEnt (see below) to upweight records with few neighbors in geographic space (Elith et al., 2010, 2011; Yackulic et al., 2013). We dealt with contingent, environmental and methodological absences (sensu Lobo et al., 2010) by creating background points to use in MaxEnt; these were not real absences although we assumed they corresponded to absences. Lobo et al. (2010) defined contingent absences as 'environmentally favourable places from where a species is absent due to restrictive forces such as dispersal limitations or local extinctions'. To determine contingent absences we used the Historical Known Range, assuming that the area outside of this range implies historical ab� sence and locating records along the spatial gradient under consideration (Lobo et al., 2010). To deal with environmental absences, we first created a minimum convex polygon (i.e. the smallest polygon in which no internal angle exceeds 180 degrees and contains all presence sites) using records from the present day. We used areas outside this polygon to randomly cre� ate environmental absences (Yackulic et al., 2013). To interpret methodological absences we calculated the Kernel density to identify places where representation is uncertain and places which are underrepresented. The Ecological–Niche Factor Analysis (ENFA) provides an exploratory analysis of the distribu� tion of a species (Hirzel et al., 2002; Basille et al., 2008; Calenge & Basille, 2008). It is used to identify the species' response to the main environ� mental variations in the study area (Rotenberry et al., 2006). Climate and topography are considered

Animal Biodiversity and Conservation 40.1 (2017)

the main determinants of the distribution of species at continental and sub–continental scales such as that studied here (Caughley et al., 1987; Wiens et al., 1987; Pearson et al., 2002, 2004; Soberón & Peterson, 2005). Hence, as input for the ecogeogra� phical variables of the ENFA, we used the nineteen bioclimatic variables derived from monthly tempera� ture and rainfall values, provided by the WorldClim ver. 1.4 interpolated map database (Hijmans et al., 2005; www.worldclim.org/). The ENFA was performed using Biomapper (Hirzel et al., 2002). The first factor extracted by the ENFA maximizes the marginality of the species, i.e. the ecological distance between the optimum for the species and the average condition in the study area. The other factors generated by the ENFA maximize specialization, defined as the ratio between the average overall variance for the study area and the variance observed for the species. The first five factors from eigenvalues were selected for mapping habitat suitability (HS). Biomapper provides the Boyce index (B) to indicate the spatial robustness of the model (Boyce et al., 2002). Species distribution models were generated with MaxEnt (Phillips et al., 2006). Although a number of different modeling approaches are available, Max� Ent performs relatively well compared to alternative approaches for modeling species that are widely distributed (Hernandez et al., 2008; Norris, 2014), such as the jaguar. One study claimed it has a good general performance when only presence records are available (Elith et al., 2006), but this work has been criticized for its design and findings (Lobo, 2008; Hij� mans, 2012), especially because the Area Under the Receiving Operator Curve (AUC) was used to evaluate performance. When the potential distribution is the goal of the research, the AUC is not an appropriate performance measure because the weight of commis� sion errors is much lower than that of omission errors (Jiménez–Valverde, 2012) and the AUC value can be artificially inflated because of distances between train� ing and testing points (Hijmans, 2012). We therefore used specificity to evaluate the models as it is suitable for the distribution performed (Tessarolo et al., 2014). Specificity is the probability that the model will correctly classify an absence (Allouche et al., 2006). MaxEnt uses the principle of maximum entropy and presence–background data to estimate a set of functions that relate environmental variables with habitat suitability to approximate the species' potential geographic distribution (Phillips et al., 2006). We set the program so that it performed both linear and quadratic features, as these generally perform better than the models which consider just linear features (Anderson & Gonzalez, 2011). We used the logistic output of MaxEnt which approximates better to the probability of occurrence; however, due to its effectiveness in describing the geographic distribution of the species' record, Maxent should be interpreted as a measure of the realized distribution of the species, rather than of the potential distribution that is typically characterised by habitat suitability (Jiménez–Valverde et al., 2008). This makes this tool adequate to characterise the changes in the current geographic distribution of the jaguar.

We developed two species distribution models. The first one was for the entire distribution range of the species, from USA to Argentina, using historical presence records (1741 to 2011), hereafter called the Historical Regional Model. As we had a lot of dubi� ous records (see stars in fig. 1) we trained the model using only presence records of the species that fell inside their Known Historical Range and, to the same extent, environmental variables. We then projected the model for the entire extension of America to check whether those records that were not used for model training were predicted as suitable. As we had more presence records (see Kernel density map; fig. 2) from Argentina, due to personal work in this area, and because we wanted to assess distributions of fringe populations in the said area, a smaller–scale analysis was more appropriate to meet our objectives. We thus developed a second model for Argentina alone that included current presence records f (from 1994 to 2011), hereafter called the Present Argentine Model. In this way we modeled the realized niche of the past and present in Argentina. We used 19 bioclimatic variables (WorldClim) that consist of interpolated data, derived from monthly temperature and rainfall (Hijmans et al., 2005). We used variables at a resolution scale of 2.5 arc–minutes (~21 km2) that captures local to regional climate variability and avoids reflecting internal dynamics of the communities (competition, predation, mutualisms, etc) of more refined resolutions. To define which environmental variables to use, we chose the variables of the ENFA analysis that were not intercorrelated (Pearson R < 0.7). We selected temperature seasonality, minimum temperature of the coldest month, and precipitation of the warmest quarter. MaxEnt offers a range suitability map as a logistic output. A threshold should be applied to transform this probability map into a binary presence/absence map. Applying a threshold is one of the most controversial issues in MaxEnt (Papeş & Gaubert, 2007; Rebelo & Jones, 2010). Therefore, we used a different threshold in each model because we had different objectives. For the Historical Regional Model we applied the fixed cumulative value 1 logistic threshold, which was 0.0878. When mapping, we divided the areas into white (non suitable) and grey (suitable) areas; the suitable areas were divided into 'low' (probability values of 0.25 or lower), intermediate (0.25–0.50), high (0.5–0.75) and maximum (> 0.75) suitability. To represent a niche more closely related to the niche for the Present Argentine Model we used a stricter threshold; the 10 percentile training presence logistic threshold commonly used for conservation objectives which was 0.4212. This threshold excludes 10% of presence records (according to a statistically calculated error), ensures that species presence is not exaggerated and enables us to focus better on conservation actions on site. To estimate losses due to human land use, we extracted all pixels with land covers incompatible with jaguar presence, and obtained in this way the occupied distributional area using a raster from the Global Land Cover database (ESA & UCLouvain, 2010). We considered 'croplands' (rain–fed

Cuyckens et al.

Records used for modelling Records outside historic range KHR (Panthera Foundation) 0

1,000

2,000

3,000 km

Fig. 1. Presence records (l) of jaguar Panthera onca used for modelling and outside the known distribution range («); the black line indicates known historical range (KHR, based on the Panthera Foundation shapefile). Fig. 1. Registros de presencia del jaguar Panthera onca empleados para la elaboración del modelo (l) y fuera del rango de distribución conocida («); la línea negra indica el rango histórico conocido (KHR, basado en el archivo Shape de la Fundación Panthera).

and mosaic) and 'artificial areas' as not suitable for the maintenance of jaguar populations. We compared the historical and present distribution of the species in Argentina by superimposing the Regional Model (projected in Argentina) with the Present Argentine Model. We also compared the occupied distributional area with the present fundamental niche to indicate how much of the jaguar’s potential area was already lost by changes in human land use. These analyses provide calculations of the overall surface area occupied by each aspect of the distribution of the jaguar. We also indicate historical and current presence of the species in the ecoregions and provinces of Argentina.

Results From a total of 1,447 presence records obtained for the species (fig. 1), only 718 records were used for the Historical Regional Model after considering only those that were inside the Known Historical Range and filtering them on a scale of 21 km2. This model had a specificity of 0.996. Despite the limitations of this model, it shows that the jaguar was historically present in 19,921,440 km2, from the south–eastern United States, including Mexico and Central America, to Argentina and Chile (fig. 2). This range includes part of Argentinean Patagonia,

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KHR (Panthera Foundation) Habitat suitability Absence Low Intermediate High Maximum 0

1,000

2,000

3,000 km

Fig. 2. Geographic projection of the historical (1741 to 2011) regional distribution of Panthera onca, based on environmental variables (temperature seasonality, minimum temperature of coldest month, and precipitation of warmest quarter). White areas are not suitable for the species and the scale of grey indicates increasing suitability. Fig. 2. Proyección geográfica de la distribución regional histórica (entre 1741 y 2011) de Panthera onca, basada en variables ambientales (estacionalidad de la temperaturas, temperatura mínima del mes más frío y precipitación del trimestre más cálido). Las zonas blancas no son adecuadas para la especie y la escala de grises indica el aumento de idoneidad.

especially the coastal area, the coast of the Chilean Valdivian Forests and part of the Chilean Matorral. The kernel density map (fig. 3) indicates that Brazil was under–sampled and northern Argentina and parts of Mexico were over–sampled. The map of presence records indicates a historical range contraction from latitudes 39.9º to 26.3º South (fig. 4). According to the Historical Regional Model, 77% of the Argentinean territory was environmentally suitable for the jaguar in the past (see light grey areas in fig. 5). Since 1994 this extent has been reduced to 5.5% of the territory, according to the model for the present day (see

dark grey areas in fig. 6), which had a good specificity of 0.904. We found 25% of this environmentally suitable area has already been lost due to changes in human land use, leaving a suitability of 4.2% of the Argentine territory or 5.2% of the area originally (historically) suitable for the jaguar (fig. 6). The species survives in only 6 of the original 23 Argentinean provinces (fig. 5). While it was originally present in almost all 12 Argentine ecoregions, today it survives only in the Dry Chaco (DCH), Humid Chaco (HCH), Yungas (YU) (including High pastures), a small part of Campos and Malezales, and Esteros de Iberá (IB) (fig. 6).

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Under–sampled Relatively well–sampled Well–sampled Somewhat over–sampled Over–sampled 0

1,000

2,000

3,000 km

Fig. 3. Kernel density of presence records used for modelling, as proxy for sample density. Fig. 3. Densidad del kernel de los registros de presencia utilizados para la elaboración del modelo, como variable sustitutiva de la densidad de la muestra.

The ENFA identifies that the most important environmental variables in determining jaguar distribution are temperature seasonality and minimum temperature of the coldest month, contributing to 14 and 12% of the variability, respectively. The ENFA indicates a high marginality for the jaguar. The first five factors were selected for mapping habitat suitability, and their ac� cumulative contribution reached 75.1%, accounting for 100% of marginality and 75.1% of the total specialization. The habitat suitability model showed a high predictive power (B = 0.64; s = 0.32). The limit value to differentiate between marginal habitat and suitable habitat was 25. Discussion We collected a large number of presence records for the jaguar through the species’ historical range, allowing

us to assess changes in the distribution of a species of high conservation value that is undergoing a retracting range process. Thereby, we provided a broad coverage of the environmental and geographical variability shown in current records (Lobo et al., 2007). According to our models, the jaguar occupies intermediate values of temperature seasonality, indicating the optimality of relatively stable climates for the species. The importance of precipitation in the warmest quarter, a variable related to plant growth, may indicate the importance of vegetation cover for the jaguar to hunt (Hopcraft et al., 2005). Minimum temperature of the coldest month was less important but could indicate the coldest limits to jaguar distribution. The regional distribution map presented here is the first to represent the historical distribution of the jaguar in South America. Our model adequately represents jaguar absence from the high altitudes of the Puna or

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Past

Recent

N Jaguar records Extent of occurrence 0 210 420

840

1,260

1,680 km

Fig. 4. Historical (1741 to 2011) and present (from 1994 to 2011) presence records and distribution range (Convex hull; black line) of jaguar (Panthera onca) in Argentina. Fig. 4. Registros de presencia histórica (entre 1741 y 2011) y actual (entre 1994 y 2011) y rango de distribución (envolvente convexa; línea negra) del jaguar (Panthera onca) en Argentina.

High Andes (Perovic & Herrán, 1998; Sanderson et al., 2002a). In this area, the main environmental factor restricting jaguar distribution could be extreme aridity (Clarke, 2006). This result contrasts with the prediction of presence in Chilean Valdivian Forests and Matorral, which never hosted jaguar populations. SDM results are independent of the presence of geographic barriers and therefore may extrapolate the species distribution towards areas that are not truly reachable for the species, such as the many islands deemed suitable by our model, i.e. the Caribbean Islands or Tierra del Fuego. It is therefore clear that dispersal limitations have played a role (Svenning & Skov, 2005) in shaping the jaguar’s distribution, and can be as important as or even more important than environmental suitability. The absence from Central Patagonia predicted by our model could be explained by a lack of vegetation cover (Burkart et al., 1999). The jaguar needs a certain level of cover to apply its hunting technique of ambushing (Brown & López–González, 2001; Hopcraft et al., 2005). Besides the lack of records, the model indicates a wide cover in Brazil, although with low suitability in the Brazilian

basin, as mentioned by Tôrres et al. (2007). Although the Historical Regional Model shows a wide distribution of areas of high suitability, the ENFA indicates that the jaguar has high marginality, which, in turn, indicates that it inhabits landscapes different from the majority present in the study area. Out of the dubious presence records (outside the Known Historical Range), only one group in Florida, United States and two records in the Argentine Patagonia were predicted by our model. Therefore, we deduce that a higher proportion of northern records were erroneous and that the model predicts well in Argentine Patagonia and is therefore useful to indicate the historical southern distribution limit. Interestingly, the southern limit of the distribution of the jaguar was neither the Colorado nor the Negro River, but an irregular limit further southwards. This makes sense as rivers do not imply determinative restrictions or limitations to jaguar dispersal, as they are excellent swimmers. Two records observed by historians were neither used nor predicted by the model (stars in fig. 1). The difference between the predicted presence by our Historical Regional Model and the Known

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N Present Fundamental Niche Areas with incompatible land use Historic Fundamental Niche 0 250 500

1,000 km

Fig. 5. Superimposition of the geographic projection of the Historical Regional Model (light grey area) and present Argentina (dark grey area) of jaguar (Panthera onca). (Present distribution is totally included in the historical distribution). Areas not suitable for jaguar presence due to land–use inside current range (black areas). Rectangle shows zoom in view in figure 6. Fig. 5. Superposición de la proyección geográfica del modelo regional histórico (área de color gris claro) y presente en Argentina (área de color gris oscuro) del jaguar (Panthera onca). (La distribución presente queda totalmente dentro de la distribución histórica). Zonas que no son idóneas para la presencia del jaguar debido al uso de la tierra dentro del rango actual (zonas negras). El rectángulo ampliado se encuentra en la figura 6.

Historical Range in the extreme north of its distribution in the southern USA (a small white strip in fig. 2) can be due to recent retraction of the species in those areas, suggesting southwards retractions of the species. The Kernel density map indicates over–sampling in the north of Argentina and Mexico. This is encouraging as these are fringe populations and conservation actions are more urgently needed. It further indicates many areas in Brazil are under–sampled, probably due to their difficult access. According to our estimates, the Historical Regional Distribution of the jaguar encompassed more than 19 million km2, more than twice the 8.75 million km² estimated previously by Sanderson et al. (2002a). There are some key differences between the two maps. In the southern portion of the potential distribution, our model predicts presence in the Valdivian forest and Chilean Matorral, and the area of predicted presence in Patagonia is larger than formerly accepted. Also,

our model predicts a strip of suitable territory on the southern–eastern coast of the USA (Texas, Louisiana, Mississippi, Alabama and Florida), not indicated on the Sanderson map. Furthermore, our model correctly classifies several historical presence records from Florida that were not used for its calibration, so it is possible that the jaguar once reached this region. The distribution of the jaguar could be depicted by a current records map and the Argentinean convex hull (fig. 4). However, using only atlas records is limited as it does not provide information on species presence between known records. Furthermore, the convex hull may generalise too much, failing to delineate the distribution of species populations at finer scales (for example, at a local scale) (Hurlbert & Jetz, 2007; Hortal, 2008), thereby overestimating the area of occupancy (Hurlbert & Jetz, 2007). Therefore, we believe that our SDM results are a better approximation of the actual distribution.

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Fundamental Niche Areas not compatible with jaguar presence

DCH

HCH

PA N

0 50 100 200

300

400 km

Fig. 6. Fundamental Niche and areas not suitable for jaguar presence due to land–use. Black circle indicates imminent loss of connection between Yungas and Chaco ecoregions: YU. Yungas; DCH. Dry Chaco; HCH. Humid Chaco; IB. Ibera Wetlands; PA. Paranaen Forests. Fig. 6. Nicho fundamental y zonas que no son adecuadas para la presencia del jaguar debido al uso del suelo. El círculo negro indica la pérdida inminente de conexión entre las ecorregiones de Yungas y Chaco: YU. Yungas; DCH. Chaco seco; HCH. Chaco húmedo; IB. Humedales Iberá; PA. Bosques paranaenses.

No records further south than latitude –26.5º of were observed after 1994, suggesting a contraction of the southern range limit towards the north. Both the map with presence records of Argentina and the difference between the Historical Fundamental Niche and the Present Niche (fig. 5) indicate that the jaguar has contracted its range in the past 200 years. This contraction and local habitat loss are indicated by the black areas in its current species range, suggesting that the present known distribution is smaller than its potential distribution in Argentina. Areas suitable for the species, but where the species is not present or was not detected, need special attention and could probably be recovered provided that conservation strategies are implemented. The Present Argentinean Model (figs. 5, 6) represents the current known distribution of the species well, with presence in only six, and marginally in seven, political provinces of the original 23 provinces of Argentina with environmentally suitable areas (fig. 5). The species is currently present in only five ecoregions: Dry Chaco, Humid Chaco, Yungas, Paranaen forest and Iberá Wetlands (fig. 6). The lowest elevation of the Argentinean Yungas forest, piedmont forest, has been highly transformed to other land uses (Brown & Malizia, 2004). Therefore, the jaguar is locally extinct in most of this ecoregion (black areas in fig. 6B), except

for the higher elevations. In the Yungas, jaguars have currently disappeared from sites where they were still present two decades ago (Perovic & Herrán, 1998), indicating the speed at which the distribution of this species is diminishing. The highest probabilities of occurrence in Argentina occur in just a few pixels, in the Upper Bermejo River, on the border with Bolivia. These areas are important to ensure connectivity with the Bolivian jaguar populations (Cuyckens et al., 2014). Also, the Upper Bermejo River Basin has already been signalled as important for Yungas conservation, due to its high biodiversity and presence of continuous forest (Brown et al., 2006). Although the models indicate a high probability of presence, and there are several presence records from the Dry and Humid Chaco (fig. 4), jaguar populations are rapidly declining in this ecoregion (Altrichter et al., 2006; Quiroga et al., 2014). This indicates that, despite environmental suitability, jaguar populations are threatened, probably by direct hunting and prey losses or land uses changes not detected by the scale and time of our land cover data. Furthermore, the connectivity between Chaco and Yungas ecoregions is disappearing (black circle in fig. 6), jeopardizing the connection. The connection is also lost between Chaco and Paranaen Forests, as no jaguar individuals were recorded in Iberá Wetlands, despite its suitability for

Cuyckens et al.

the species and historical presence of the species. In the southeast of Argentina, in the Paranaen forest, the highest probabilities occur on the limit with Brazil. This ecoregion is isolated from others in Argentina, but still has connections with Brazil and Paraguay. Therefore, tri–national conservation actions for jaguars would be crucial here (Paviolo et al., 2006). Jaguars have been seriously affected by habitat destruction and are more vulnerable to changes in land–cover than, for example, the puma; their habitat was already reduced by more than 90% in this ecoregion (De Angelo et al., 2011). Argentina represents the edge of distribution of the species and thus has a key role in jaguar conservation, with possible consequences for the species in neighbouring countries (Brazil, Bolivia, and Paraguay). The species is considered as a 'Natural Monument', the highest conservation category in the country (National Law 25.463), and hunting is totally prohibited. Nevertheless, there seems to be a lack of implementation of protection measures where it would be most important due to the small numbers of individuals present and the threat of direct hunting. On the other hand, Argentina is a country oriented towards crop production, promoting the increase of soy bean crops and other cultivars which directly threaten jaguars through habitat loss. The challenge is to combine production with jaguar conservation, while simultaneously promoting the peaceful coexistence between rural communities and this large carnivore. Jaguar distribution has been used as an example of the problems of biodiversity data and how it is mapped (McInerny et al., 2014), so we applied several methods to improve the models and the scientific rigor of the MaxEnt analyses. We gathered larger numbers of more widely distributed presence records and used adequate absence records as a bias file in MaxEnt (Elith et al., 2010; Anderson, 2012), worked at a finer scale (21 km2 vs. 100 km2), and selected uncorrelated environmental variables (Rocchini et al., 2011). We also had knowl� edge about the species to interpret the map and we focused on the area we know best. We generated a Kernel density map as a proxy for sample density in accordance with 'ignorance maps' (sensu Rocchini et al., 2011). We therefore believe that our results are robust and provide a fair representation of the distribution of the jaguar. However, our model presents several limita� tions, as the jaguar is not a species in equilibrium with its environment due to its ever–decreasing distribution. Nevertheless, we believe that the method we applied here was adequate to accomplish the proposed aims. This work contributes significantly to understanding the distribution of the jaguar and its conservation status, including the interpretation of the factors that could be influencing the conservation of this endangered species. We found evidence of range retraction in both the southern and the northern distribution limits. Due to our personal experience, we were able to assess the retraction in its southern limit in more detail. Our results support the idea that there is an imminent, high extinction risk of the jaguar in Argentina, as evidenced by its northward range contraction over time and the velocity of this decline, its range contraction in the Chaco ecoregion (Quiroga et al., 2014), and the lack

of areas of growth or recovery of populations. We encourage changes in policies as records of killed animals are still common, and there is a general lack of assessment of which actions may be worth taking. Acknowledgements One of the authors (G. A. E. C.) was financially supported by the Argentine Scientific Council (CONICET). We thank our colleagues C. de Angelo, A. Paviolo, V. Quiroga and M. Castro for providing field data, M. Tognelli and C. de Angelo for revisions of previous versions of this manu� script, J. Osorio and M. Zietsman for revising the English, and the anonymous reviewers and the editors of the journal. We also especially thank J. Hortal for reviewing and substantially improving the manuscript. References Allouche, O., Tsoar, A. & Kadmon, R., 2006. Asses� sing the accuracy of species distribution models: prevalence, kappa and the true skill statistic (TSS): Assessing the accuracy of distribution models. Journal of Applied Ecology, 43: 1223–1232. Doi: 10.1111/j.1365–2664.2006.01214.x Altrichter, M., Boaglio, G. & Perovic, P., 2006. The decline of jaguars Panthera onca in the Argentine Chaco. Oryx, 40: 1–8. Anderson, R. P., 2012. Harnessing the world’s biodi� versity data: promise and peril in ecological niche modeling of species distributions: Niche modeling to harness biodiversity data. Annals of the New York Academy of Sciences, 1260: 66–80. Doi: 10.1111/j.1749–6632.2011.06440.x Anderson, R. P. & Gonzalez, I., 2011. Species–spe� cific tuning increases robustness to sampling bias in models of species distributions: An implemen� tation with Maxent. Ecological Modelling, 222: 2796–2811. Doi: 10.1016/j.ecolmodel.2011.04.011 Araújo, M. B. & Guisan, A., 2006. Five (or so) challen� ges for species distribution modelling. Journal of Biogeography, 33: 1677–1688. Doi: 10.1111/j.1365– 2699.2006.01584.x Araújo, M. B., Pearson, R. G., Thuiller, W. & Erhard, M., 2005. Validation of species–climate impact models under climate change. Global Change Biology, 11: 1504–1513. Arra, M. A., 1974. Distribución de Leo onca (L) en Argentina. Neotropica 20: 156–158. Basille, M., Calenge, C., Marboutin, É., Andersen, R. & Gaillard, J.–M., 2008. Assessing habitat selection using multivariate statistics: Some refinements of the ecological–niche factor analysis. Ecological Modelling, 211: 233–240. Doi: 10.1016/j.ecolmo� del.2007.09.006 Borrero, L. A., 2001. Regional Taphonomy: Back� ground Noise and the Integrity of the Archaeological Record. In: Ethnoarchaeology of Andean South America. International Monographs in Prehistory, Ann Arbor: 243–254. Boyce, M. S., Vernier, P. R., Nielsen, S. E. & Schmie�

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Supplementary material

Tabla 1s. Sources and types of sources of presence records gathered. Tabla 1s. Fuentes y el tipo de fuentes de registros de presencia recopilados.

Source

Type of source

French, 1839

Published literature

Chauvet, 1857 de Moussy, 1864 Dobson, 1900 Río & Achával, 1904 Cann, 1939 Sánchez, 1946 Saenz, 1970 Carman, 1984 Zeballos, 1994 Barquez, 1997 Perovic & Herrán, 1998 Sierra Iglesias, 1998 Anderson, Peterson & Gómez–Laverde, 2002 Altrichter et al., 2006 Pautasso, 2008 Diaz, 2010 Estrada Hernández & Juárez Sánchez, 2003

Grey literature

De Angelo, 2009 CONABIO, 2010

Publicly available databases

Centro de Referência em Informação Ambiental CRIA & Fundação de Amparo à Pesquisa do Estado de São Paulo, 2014 Global Biodiversity Information Facility (GBIF) Administration of National Parks of Argentina (APN) Jaguar Network (Red Yaguareté) American Museum of Natural History (AMNH)

Museum collections

New York, USA Argentine Museum of Natural Sciences

'Bernardino Rivadavia' (MACN), Buenos Aires, Argentine Perovic, 2002

Previous works by some of the authors

Cuyckens, 2013 C. de Angelo pers. comm.

Unpublished field data kindly

A. Paviolo pers. comm.

shared by colleagues

V. Quiroga pers. comm. M. Castro, pers. comm

Cuyckens et al.

Supplementary material (Cont.)

–70ºW

–60ºW

Bolivia

–20ºS

Brazil Paraguay

Chile

Argentine

–30ºS

Uruguay

Col

ora

–10ºS

gro

Riv er

170

241

850 km

JHR (Panthera Foundation) Land cover (GlobCover) Irrigated croplands Rainfed croplands Mosaic croplands / vegetation Mosaic vegetation / croplands Artificial areas

Fig. 1s. Land cover in the southern cone of South America, indicating uses not compatible with jaguar presence (scale of greys). The black line indicates jaguar’s historical distribution (JHR according to Panthera Foundation) and we indicated the Negro and Colorado Rivers which were signalized as possible southernmost limits of jaguar’s distribution (Carman, 1984; Lehmann–Nitsche, 1907). Fig. 1s. Ocupación en el suelo del cono meridional de América del Sur que indica los usos incompatibles con la presencia del jaguar (escala de grises). La línea negra indica la distribución histórica del jaguar (según la Fundación Panthera) y nosotros indicamos los ríos Negro y Colorado que se señalizaron como los posibles límites más meridionales de la distribución del jaguar (Carman, 1984; Lehmann–Nitsche, 1907).

Animal Biodiversity and Conservation 40.1 (2017)

Eucalypt plantations reduce the diversity of macroinvertebrates in small forested streams A. Cordero–Rivera, A. Martínez Álvarez & M. Álvarez

Cordero–Rivera, A., Martínez Álvarez, A. & Álvarez, M., 2017. Eucalypt plantations reduce the diversity of macroinvertebrates in small forested streams. Animal Biodiversity and Conservation, 40.1: 87–97. Abstract Eucalypt plantations reduce the diversity of macroinvertebrates in small forested streams.— Land use patterns of a river basin have a significant effect on the structure and function of river ecosystems. Changes in the composition of riparian plant communities modify the quantity, quality and seasonality of leaf–litter inputs, determining changes in macroinvertebrate colonization and activity. The main goal of this study was to test the effect of land–use modifications, and particularly the impact of eucalypt plantations, on the macroinvertebrate communities of sixteen headwater streams. Macroinvertebrates were counted and identified to family level. Land uses were classified in five categories using aerial photography: native forest, eucalypt plantations, agricultural land, shrubland, and urban areas. We found that macroinvertebrate diversity increased with basin size and with the proportion of ba� sin covered by native forest. This variable correlated negatively with the land occupied by eucalypt plantations. Macroinvertebrate richness diminished with the increase of land surface covered by eucalypt plantations, and a similar tendency was observed with diversity. Furthermore, streams whose drainage basin was mainly covered by Eucalyptus were more likely to dry up in summer. This observation adds to evidence from previous studies that concluded that fast–growing tree plantations affect hydric resources, an important ecosystem service in the context of global warming. To minimize the impact of industrial sylviculture, we suggest that maintaining and/or restoring riparian forests could mitigate the effects of intensive eucalypt monocultures. Key words: Eucalyptus globulus, Biodiversity, River ecosystems, Land uses, Forest, Tree plantations Resumen Las plantaciones de eucaliptos reducen la diversidad de macroinvertebrados en pequeños arroyos forestales.— Los usos del suelo de una cuenca hidrológica ejercen un efecto importante en la estructura y el funcionamiento de sus ecosistemas fluviales. Los cambios en la composición de las comunidades de plantas ripícolas modifican la cantidad, calidad y estacionalidad de las entradas de materia y energía a los ríos, lo que afecta a la colonización y actividad de sus comunidades de macroinvertebrados. El principal objetivo de este estudio es analizar los efectos de los cambios en el uso del suelo y, en particular, de las plantaciones de eucalipto, en las comunidades de macroinvertebrados de 16 arroyos de cabecera. Se contaron macroinvertebrados y se identificaron hasta el nivel de familia. Los usos del suelo se clasificaron en cinco categorías utilizando fotografías aéreas: bosque autóctono, eucaliptal, zona agrícola, matorral y zona urbana. Observamos que la diversidad de macroinvertebrados aumentó con el tamaño de la cuenca y con la proporción de superficie de la cuenca cubierta por bosque autóctono, lo cual resultó estar inversamente correlacionado con la superficie ocupada por eucaliptales. La riqueza de macroinvertebrados disminuyó a medida que aumentaba el suelo ocupado por eucaliptales y se produjo una tendencia similar con la diversidad. Además, nuestras observaciones indican que los arroyos cuyas áreas de captación están cubiertas principalmente por eucal� iptales presentan una mayor probabilidad de secarse completamente en verano. Esta observación añade un nuevo indicio concordante con otros estudios que concluyen que las plantaciones de árboles de rápido crecimiento afectan a los recursos hídricos, que constituyen un servicio ecosistémico importante en el contexto del calentamiento de la Tierra. Con vistas a minimizar los efectos de la silvicultura industrial, se sugiere que mantener o recuperar bosques ribereños podría mitigar las repercusiones de los monocultivos intensivos de eucaliptos. Palabras clave: Eucalyptus globulus, Biodiversidad, Ecosistemas fluviales, Usos de suelo, Bosque, Planta� ciones de árboles Received: 16 III 16; Conditional acceptance: 1 VI 16; Final acceptance: 5 X 16 Adolfo Cordero–Rivera, Alba Martínez Álvarez & Maruxa Álvarez, ECOEVO Lab., Univ. de Vigo, EUE Forestal, Campus Universitario, 36005 Pontevedra, Spain. Corresponding author: Adolfo Cordero–Rivera. E–mail: adolfo.cordero@uvigo.es ISSN: 1578–665 X eISSN: 2014–928 X

Introduction Rivers are closely related to the terrestrial ecosystems of their drainage basins, to the point that modifications in land use patterns determine changes in the physico– chemical properties of rivers, and as a consequence, in their biotic structure (Margalef, 1983; Allan, 2004a; Sandin, 2009). In effect, land use changes are an integrator of many human activities that have a nega� tive impact on stream ecosystems (Allan & Castillo, 2007). Degradation of stream ecosystems derived from modifications in land use patterns is likely manifested in changes to flows and water temperature, bank ero� sion and silt deposition, thus affecting benthic habitat conditions (Hickey & Doran, 2004). In regions that naturally have riparian vegetation, trees strongly influ� ence energy pathways by changing the availability of light and input of organic matter (Gregory et al., 1991). The impact of invasive exotic species on native communities is widely known, and is considered an important component of global change (Sakai et al., 2001), particularly in relation to the loss of biodiversity (Calviño–Cancela et al., 2012, 2013; Calviño–Cancela, 2013)�� . Furthermore, the management of invasive spe� cies is one of the greatest challenges facing conserva� tion in Europe in this millennium (Genovesi & Shine, 2003). In the Iberian peninsula, vast areas of potentially deciduous forest have been colonised by plantations of the Australian tree Eucalyptus globulus Labill, which is clearly invasive in the North of Iberia (Dana et al., 2003; Calviño–Cancela & Rubido–Bará, 2013). Other species, such as Australian beetles (Cordero Rivera et al., 1999) and fungi (Díez, 2005), become invasive through their association with eucalypts. The spread of Eucalyptus is partially explained by the economic benefits of fast–growing tree plantations whose wood is used by the paper industry (see review in Canhoto et al., 2004), but also because, as a pyrophytic species, eucalypt are favoured by wildfires (Guitián Rivera & Cordero–Rivera, 2007). In small forested streams in which food–webs are based on detrital inputs from surrounding forest (Wal� �� lace et al., 1997)�� , modifications in watershed vegeta� tion also alter the quantity, quality and seasonality of leaf–litter inputs (Abelho & Graça, 1996; Molinero & Pozo, 2004). This variation in detritus quality may influence the colonisation and activity of decomposers (Kearns & Bärlocher, 2008)�� . In particular, the replace� ment of native mixed deciduous forests by evergreen monospecific plantations of eucalypts is known to produce structural and functional modifications on river ecosystems (Graça et al., 2002). Eucalyptus leaves are a resource of lower quality for aquatic organisms than those of native species, such as Alnus glutinosa (Canhoto & Graça, 1995; Santiago et al., 2011), with a lower amount of nitrogenous and phosphorous and a higher quantity of compounds of difficult degradation (lignin, oils, tannins and other phenolic compounds) (Canhoto & Graça, 1996; Molinero & Pozo, 2004). The substitution of native riparian vegetation by eu� calypt plantations also determines changes in light and temperature regimes, as well as alterations of the substrate and habitats, due to the frequent deposition

Cordero–Rivera et al.

of large particulate organic matter and soil particles (Graça et al., 2002). These changes also increase soil hydrophobicity. The presence of eucalypt plantations in riverine habitats thus modifies hydrologic regimes, particularly when clear cutting operations take place in the basin (Fernández et al., 2006). The resultant hydrophobicity affects the infiltration rate of water in the soil, increasing surface run–off, promoting erosion, and diminishing subterranean water reservoirs. Another side effect of this hydrophobicity is the possible summer dry out of streams whose basins are covered by eucalypt plantations (Graça et al., 2002; Jackson et al., 2005). Given the above considerations, it is surprising that empirical data on the effects of eucalypt plantations on native biota are scarce (but see Calviño–Cancela & Neumann, 2015), particularly in small streams, which are particularly relevant to maintain biodiversity (Finn et al., 2011). The main goal of this study was to test whether the observed effect of Eucalyptus planta� tions on stream macroinvertebrate communities are generalizable to the situation in NW Spain, where the numbers of eucalypt plantations have increased dramatically in the last decades (Cordero–Rivera, 2012), and where the eucalypts clearly show invasive behaviour after fire (Guitián Rivera & Cordero–Rivera, 2007). We analysed the effect of land use patterns, and particularly the effect of eucalypt plantations, on the macroinvertebrate communities of headwater streams. We hypothesized that streams whose basins were mainly covered by eucalypt plantations would have less richness and diversity and would more likely dry out in summer than comparable streams running through native riparian forests. Material and methods Study area For this study, we sampled 16 stream tributaries of the Lérez River in Pontevedra province (NW Spain) (fig. 1) in 2011. Streams were selected for their accessibility, size and land use patterns, covering the variability of land uses and vegetation types observed in the drainage (table 1). The average stream basin size is 2.66 ± 0.86 km2, ranging between 0.05 km2 in A Ceira and 12.93 km2 in Os Calvos (table 1). In agreement with the granitic geology of the region, water is slightly acid (mean ± SE pH = 6.01 ± 0.01) and has a low ionic content (mean conductivity = 66.58 ± 7.46 µS/cm; table 1; see also Membiela et al., 1991, for a review of typical values of rivers of the region). Most of the streams are on a southern slope (fig. 1), and thus have less flow during summer months, with some remaining as isolated pools for several weeks. To avoid the dry season, we performed the sampling at the beginning of spring (March–April) and the end of spring (May). Sampling methodology Macroinvertebrates were collected using a Surber net with a mesh size of 250 µm and a sampling area of 0.1 m2. In one of the streams, As Pozas (table 1),

Animal Biodiversity and Conservation 40.1 (2017)

Suapica 838 m

Outeiro Grande 777 m

Mourente 240 m Serra da Fracha 542 m

rio no Te

Ri ve rL ér ez

Gr a Gr nde an de 1 Rego F 2S 1 eix Qu iña ire za

i fre no l A

Serra Castrada 698 m

e nt Po Castro

Monte do Seixo

Cabeiro

o eir eix Fr

Candán 1,014 m

Sei xo

Pe niz as

rez Lé r e Riv

ira Ce

ez r Lér Rive

ó Lix B. o d ba m a C

iros Lace

Moído

as lv Ca

Cabale iros

Outeiro Branco 382 m

do Acebe Poz os

Acibal 594 m

Ladróns Manese s

Cregos 789 m

Serra do Condo 993 m

10 km

O Serrón 951 m N

Basin of the River Lérez

Fig. 1. The River Lérez basin (based on Río Barja & Rodríguez Lestegás, 1992) with the approximate location of sampling sites (dark circles). Fig. 1. La cuenca del río Lérez (basado en Río Barja y Rodríguez Lestegás, 1992) con la ubicación aproximada de los puntos de muestreo (círculos oscuros).

this net was too large for use so we used a hand net, following the standard norm UNE–EN 27828:1994. The two methods are comparable, although Surber sampling is usually more efficient (Torralba Burrial & Ocharan, 2007). In each stream, we collected two samples in fast flowing areas (where Surber sampling is efficient) and measured pH, water temperature and conductivity. Samples were preserved in ethanol. In the laboratory, macroinvertebrates were counted and identified to family level. Land uses and land cover To estimate basin area we used topographic maps from the Instituto Geográfico Nacional (Spanish Geographic Institute) (scale 1:25,000). From this in� formation, we estimated the limits of the watersheds over digitized maps, using Adobe Photoshop Exten� ded CS5 software (www.adobe.com). To calculate the area of the different land uses, we worked with aerial photographs. Images were obtained from the

SIGPAC webpage (http://sigpac.mapa.es/fega/visor/) and Google Earth software, and analysed with Com� peGPS Land version 7.3 (www.compegps.com) and Adobe Photoshop Extended CS5. Land uses and sizes were classified in five categories (table 1): (1) native forest, areas dominated by native trees, which in the region are mainly Alnus glutinosa, Betula alba, Quercus robur, Salix sp., Castanea sativa, Frangula alnus, Corylus avellana and Fraxinus excelsior; (2) eucalypt plantations, areas covered by Eucalyptus globulus with an undergrowth dominated by several species of Ulex and Erica; (3) agricultural land, areas where the main land use is agricultural crops, with or without irrigation; (4) shrubland, areas where the dominant vegetation is grasses, shrubs and pastures, with isolated trees; this category also includes rock and areas with no vegetation; and (5) urban, areas with small villages and buildings. We estimated the total basin area and the different land uses, and calculated the proportion of each area occupied by each land use/land cover.

Cordero–Rivera et al.

Table 1. Land uses and physical–chemical characteristics of the 16 streams included in this study. See figure 1 for their location: WT. Water temperature (ºC); WC. Water conductivity (µS/cm). Ba. Basin area (km2). Land uses: F. Native forest; S. Shrubland; E. Eucalypt plantations; Ag. Agricultural; U. Urban. Tabla 1. Usos del suelo y características físicoquímicas de los 16 arroyos incluidos en este estudio. Véase la figura 1 para conocer su ubicación: WT. Temperatura del agua (ºC); WC. Conductividad del agua (µS/cm); Ba. Superficie de la cuenca (km2). Usos del suelo: F. Bosque autóctono; S. Matorral; E. Eucaliptal; Ag. Zona agrícola; U. Zona urbana. WT

Land use cover (%) BA

A Ceira

5.6 13.2 34.9 0.05 33.99 66.01

As Pozas

4.9 14.6 106.9 0.22

0.00 25.75 74.25 0.00 0.00

Eucalypt–shrub

Barranq. de Lixó 6.3 14.0 35.8 0.28

0.00 44.26 55.74 0.00 0.00

Eucalypt–shrub

Rego 1

6.3 17.2 78.8 0.38

1.29 53.80 44.90 0.00 0.00

Eucalypt–shrub

As Penizas

6.4 17.4 95.4 0.39

0.00 54.71 45.29 0.00 0.00

Eucalypt–shrub

As Laceiras

5.9 13.8 28.9 0.76 88.28 11.72

Grande 1

5.7 15.8 55.3 1.26 12.18 29.02 50.10 5.93 2.77

Eucalypt

Fonte Seixiña

5.7 14.3 32.8 1.55 18.15 67.64

0.00 12.85 1.36

Shrubs

O Cambado

6.3 15.1 57.2 1.56 13.04 70.99 10.48 5.49 0.00

Shrubs

Os Ladróns

6.5 14.5 66.1 1.69 45.09

Native forest–Agricultural

O Moído

5.9 14.0 65.9 2.12 25.81 26.95 20.54 23.38 3.32

Mixed

Acevedo

5.7 14.6 116.7 2.91

4.13 45.06

Agricultural–shrubs

Grande 2

6.1 17.0 27.7 2.96

7.00 22.44 60.57 8.77 1.22

Os Cabaleiros

6.2 13.9 97.5 6.02 49.19 50.81

0.00 0.00 0.00

Native forest–shrubs

Os Maneses

6.4 13.7 98.8 7.44 57.52 20.64

9.87 9.40 2.57

Native forest

Os Calvos

6.3 14.1 66.6 12.93 30.96 40.19 15.75 12.10 1.00

9.36

Predominant U land use category

Stream

0.00 0.00 0.00

3.54 33.33 8.68 0.00 50.82 0.00

Native forest–shrubs

Native forest

Eucalypt

Native forest–shrubs

Data analysis In each sampling period (early and late spring), we studied the structure of communities in each stream by calculating the number of Families (a surrogate for richness), the number of individuals (Abundance), and diversity (Shannon index). In the analyses, we used the average values of both sampling periods and the number of families per stream. This allowed inclusion of the stream Barranqueira de Lixó, which was completely dry during the second sampling. The diversity index was calculated using a Box–Cox transformation to meet the assumptions of normality. This study was a sample survey, and strictly speak� ing, there were no treatments (Shaffer & Johnson, 2008), because we could not manipulate basin cover or size of our streams. We nevertheless calculated statistical relationships to test a priori ideas derived from ecological theory, which is a powerful way to identify possible cause–effect relationships. To explore the relationship between the variables describing macroinvertebrate communities and the variables describing land uses and river characteristics (i.e.

proportion of catchment area covered by each land use, basin size, pH and water temperature and con� ductivity), first, we used standard Pearson correlation analysis. Proportions (p) were transformed before the analysis (arcsin √p). The response variables were analysed by GLM assuming normal errors and identity link (diversity) or Poisson with log link (number of families). We used the Akaike Information Criterion to control over–fitting in statistical modelling and thus avoid the use of frequentist methods and their associated statistical tests, which may be misleading in obser� vational studies (Burnham et al., 2011). Agricultural and urban land cover was low in all streams, and no correlation between these variables and any measure of community structure was detected in exploratory analyses (see table 2). Land cover types made up 100%; therefore, to avoid problems of multicollinearity we included only forest and eucalypt land cover, and l basin size in the models, and no interactions were fitted because we had only 16 streams. Analyses were done using xlStat2013 (www.xlstat.com) and Genstat 18th edition (GenStat, 2015).

Animal Biodiversity and Conservation 40.1 (2017)

Table 2. Pearson correlations between variables (below the diagonal) and the p–values (above the diagonal). Diversity was Box–Cox transformed and the angular transformation was applied to all land cover variables to meet the assumptions of normality: T. Temperature; C. Conductivity; B. Basin size; Ab. Abundance; F. Families; D. Diversity; F. Forest; S. Shrubland; E. Eucalypts; Ag. Agricultural; U. Urban. (Values in bold are different from 0 with a significance level alpha = 0.05.) Tabla 2. Correlaciones de Pearson entre las variables (debajo de la diagonal) y valores de probabilidad asociados (encima de la diagonal). Para cumplir los supuestos de normalidad, la diversidad se transformó mediante el método Box–Cox y se aplicó la transformación angular a todas las variables de cobertura del territorio: T. Temperatura; C. Conductividad; B. Tamaño de la Cuenca; Ab. Abundancia; F. Familias; D. Diversidad; F. Bosque; S. Matorral; E. Eucaliptal; Ag. Agrícola; U. Urbano. (Los valores en negrita son distintos de cero con un nivel de significación alfa = 0,05.) Variables pH pH

0.422 0.825 0.206 0.015 0.823 0.278 0.497 0.971 0.743 0.779 0.386

0.216 0.738 0.317 0.306 0.120 0.278 0.018 0.732 0.014 0.639 0.752

–0.060 0.091 0.385 0.164 0.626 0.428 0.395 0.950 0.835 0.510 0.647

0.334 –0.267 0.233 0.724 0.049 0.019 0.183 0.697 0.463 0.306 0.405

–0.597 –0.273 0.366 0.096 0.083 0.962 0.742 0.901 0.806 0.332 0.935

0.061 –0.405 –0.132 0.499 0.447 0.000 0.041 0.768 0.018 0.065 0.141

0.289 –0.289 –0.213 0.579 –0.013 0.777 0.005 0.648 0.093 0.236 0.096

0.183 –0.581 –0.229 0.351 –0.089 0.515 0.664 0.133 0.003 0.728 0.246

–0.010 0.093 –0.017 –0.106 –0.034 0.080 –0.124 –0.393 0.536 0.375 0.026

–0.089 0.598 0.057 –0.198 –0.067 –0.583 –0.435 –0.697 –0.167 0.244 0.917

0.076 –0.127 0.178 0.273 0.259 0.472 0.314 0.095 –0.238 –0.309 0.013

0.233 –0.086 –0.124 0.223 0.022 0.385 0.430 0.308 –0.552 –0.028 0.603

Finally, to test whether predominant land uses were related to changes in the structure of ma� croinvertebrate communities, we calculated two or� dinations with non–metric multidimensional scaling (MDS). The first ordination was calculated using the average density of macroinvertebrates over the two sampling periods, and the second ordina� tion was calculated using presence/absence data. These ordinations, obtained using the similarity matrix based on the Bray–Cutis and Jaccard index, respectively. They allow a visual representation of the relationship between land use categories and macroinvertebrate communities of each stream, which were compared using an index of similarity (SIMPER, similarity percentages). This procedure examines the contribution of each family of macroin� vertebrates and identifies the average similarity and dissimilarity between two groups of samples (land use categories). This analysis is therefore restricted to land use categories with at least two streams (i.e. forest, shrubland and eucalypt plantations). MDS and SIMPER analyses were performed using PRIMER v.6 (Clarke & Warwick, 2001). Means are presented with their standard errors and sample size (mean ± SE (N)).

Results We found a total of 56 families of macroinvertebra� tes in the 16 streams. Table 2 shows the pair–wise correlations between variables. In agreement with ecological theory, the diversity of macroinvertebra� tes was positively correlated with catchment area (R = 0.58, P = 0.019), but up to a limit, thus describing a power function (fig. 2A). Diversity also increased with the proportion of the basin covered by native forest (R = 0.66, P = 0.005), but in this case in the form of a hump–shaped curve (fig. 2B). As expected given that native forest and eucalypt plantations are the main land uses, these two categories were negatively correlated (r = – 0.70; P = 0.003). There� fore, as the eucalypt plantation cover increased, both macroinvertebrate diversity (R = –0.43, P = 0.093) and richness (R = –0.58, P = 0.018) were negatively affected, although only the latter value was significant, and both seem to follow a non–linear trend (fig. 3). The proportion of shrubland, agricultural and urban land on the catchment did not show any relationship with macroinvertebrate diversity or richness (table 1). The abundance of macroinvertebrates showed a negative correlation with pH (R = –0.60; P = 0.015;

Cordero–Rivera et al.

3.0

3.5 3.0 2.5

2.0

Diversity

2.5

1.5 1.0

2.0 1.5 1.0 0.5 0.0

0.5 0.0

y = 0.3116 ln(x) + 1.9219 R2 = 0.3993

5 10 Basin area (km2)

–0.5 15 B

–1.0 0.0 0.2 0.4 0.6 0.8 1.0 1.2 Eucalypt cover (asin Öproportion)

3.5

2.5 2.0 1.5 1.0 0.5 0.0 –0.5

y = –3.9478x2 + 6.1296x – 0.0571 R2 = 0.7165

–1.0 0.0 0.5 1.0 1.5 Forest cover (asin Öproportion)

Number of Families

3.0

Diversity

y = –4.4594x2 + 2.8268x + 1.6104 R2 = 0.3651

y = –16.113x2 + 1.3694x + 20,973 R2 = 0.3745

25 20 15 10 5 0 0.0 0.2 0.4 0.6 0.8 1.0 1.2 Eucalypt cover (asin Öproportion)

Fig. 2. The relationship between basin size (A) and the proportion of native forest in the basin (B) and the average macroinvertebrate diversity (Shannon index, Box–Cox transformed) found in 16 streams of the River Lérez basin. Equations refer to the adjusted curves.

Fig. 3. The relationship between the proportion of eucalypt plantations in the basin and the average macroinvertebrate diversity (Shannon index, Box– Cox transformed) (A) and richness (number of Families) (B) of macroinvertebrate communities found on 16 streams of the River Lérez basin.

Fig. 2. Relación entre el tamaño de la cuenca (A) y la proporción de bosque nativo en la cuenca (B) y la diversidad media (índice de Shannon, con la transformación Box–Cox) de los macroinvertebrados hallados en 16 arroyos de la cuenca del río Lérez. Las ecuaciones se refieren a las curvas ajustadas.

Fig. 3. Relación entre la proporción de la cuenca ocupada por plantaciones de eucalipto y la diversidad media (índice de Shannon, con la transformación Box–Cox) (A) y la riqueza (número de Familias) (B) de las comunidades de macroinvertebrados encontradas en 16 arroyos de la cuenca del río Lérez.

table 2). Nevertheless, this relationship was due to the extraordinary density of Asellidae in the source of the most acid stream, As Pozas. Excluding this datapoint, a clear outlier, the relationship was not significant (R = –0.23, P = 0.402). No other significant correlation was found between abundance and the explanatory environmental variables.

Water temperature was positively correlated with the proportion of catchment area covered by eu� calypt plantations (R = 0.60; P = 0.014) (fig. 4A). We observed an opposite pattern for the native forests (R = –0.58; P = 0.018) (fig. 4B). Mean water tempe� rature for the 6 streams included under the eucalypt and eucalypt–shrub categories was approximately 2ºC

Animal Biodiversity and Conservation 40.1 (2017)

Discussion Our results show that the structure of macroinvertebrate communities follows the expected trends derived from ecological theory: higher complexity relates positively to basin size and native forest cover and negatively to eucalypt cover (figs. 2, 3). In a general sense, running waters are one of the most impacted ecosystems on the planet as they have been the focus for human set� tlement and are heavily exploited for water supplies, irrigation, electricity generation, and waste disposal (Malmqvist & Rundle, 2002). But besides these direct anthropogenic effects, streams are also beginning to show the pressure of global climate change through

Water temperature (ºC)

18 16 14 12 10 8

y = 2.9751x + 14.098 R2 = 0.3444 P = 0.0014

6 0.0 0.2 0.4 0.6 0.8 Eucalypt cover (asin Öproportion) B Water temperature (ºC)

higher than the temperature measured in the 6 streams with highest forest cover (16.0 ± 0.6ºC and 13.8 ± 0.2ºC, respectively) (table 1). In agreement with this observation, the streams with a large eucalypt cover showed dramatic changes in water flow between the two sampling periods. One of them, Barranqueira de Lixó, completely dried out on the second sampling date, and another, As Pozas, had water remaining only at the source, forming a single pool. In contrast, the smallest stream, A Ceira, which had mainly native forest in its catchment area, had a similar flow on sampling days in both periods. The saturated model (including all three explanatory variables; i.e. basin size, eucalypt and forest cover) explained 57.8% of variance, and was therefore a good starting model (table 3). The most supported model to explain variability in macroinver� tebrate diversity included basin size and forest cover, both with a positive effect, but with a much larger effect size for forest cover (table 3). Two other models were within 2 units of AIC, including the saturated model, suggesting that all three variables are of relevance. The saturated model explains 45.6% of variance in richness (number of Families; table 3). The most supported model to describe variability in richness includes basin size and eucalypt cover, the former having a positive effect and the latter having a negative effect. The next models, again close to the first, included only eucalypt cover (explaining alone 31.5% of variance) and all three variables (table 3), suggesting that all variables are of relevance. Multi� variate ordinations also supported the observed effect of land use and catchment size on macroinvertebrate communities. Basins with eucalypt plantations (and a percentage of the stream drainage covered by na� tive forest below 20%), were the most differentiated communities (fig. 5). Based on SIMPER analysis, the proportion of similarity between the macroinvertebrate communities of catchments dominated by native forests and by eucalypt plantations was only 33.3%. Further� more, we observed that the macroinvertebrates that are typical of streams with large areas of native forest belong to several families of mayflies (e.g., Heptagenii� dae), stoneflies (e.g., Chloroperlidae) and caddisflies (Sericostomatidae, Hydropsychidae, Brachycentridae), taxa that are rare on the streams dominated by eucalypt plantations, where filter feeders simulids and limonids were the most characteristic taxa.

18 16 14 12 10 8 6 0.0

y = -2.1945x + 15.798 R2 = 0.3376 P = 0.0018

0.5 1.0 1.5 Forest cover (asin Öproportion)

Fig. 4. The relationship between the proportions of basin covered by eucalypts (A) and native forest (B) and water temperature in 16 streams of the River Lérez basin. Fig. 4. Relación entre la proporción de la cuenca ocupada por plantaciones de eucalipto (A) y por bosque nativo (B) y la temperatura del agua en 16 arroyos de la cuenca del río Lérez.

alterations in hydrology, thermal regimes and riparian vegetation (Meyer & Pulliam, 1992), which will directly affect the quantity and quality of the leaf litter that is the source of the detritus–based food webs. Within this context, the most relevant result of this study is the positive association between the proportion of the basin covered by native forests and macroinvertebrate diversity. This relation seems non–linear (fig. 2B), and a close scrutiny suggests that even low values of 15–20% of forest cover in the basin might provide enough resources for stream macroinvertebrate communities. The take–home message is clear: if forests cover more than one third of the basin, stream communities are expected to retain high diversity, particularly if riparian vegetation remains intact.

Cordero–Rivera et al.

Table 3. Summary of model selection, with the estimated effect, SE and Akaike Information Criterion (AIC). The explanatory variables were the area of the catchment (BASIN), the percentage of the catchment covered by eucalypts (EUCALYPT) and native forests (FOREST). Models are ordered by increasing AIC. R2 refers to the proportion of variance explained. Tabla 3. Resumen de la selección de modelos, con el efecto estimado, la desviación estándar (SE) y el criterio de información de Akaike (AIC). Las variables explicativas fueron el área de la cuenca (BASIN) y el porcentaje de la cuenca cubierto por eucaliptos (EUCALYPT) y por bosques autóctonos (FOREST). Los modelos están ordenados de menor a mayor AIC. R2 indica la proporción de varianza explicada. Response variable Model

Estimate

AIC

ΔAIC

BASIN + FOREST

0.120

0.059

18.006

0.000

57.79

1.574

0.576

FOREST

1.988

0.598

19.892

1.886

44.13

BASIN + EUCALYPT + FOREST 0.120

0.061

20.000

1.994

57.81

0.057

0.739

1.617

0.822

EUCALYPT + FOREST

0.158

0.814

2.105

0.864

BASIN + EUCALYPT

0.156

0.064

Diversity

BASIN EUCALYPT

–0.938

21.846 21.874

3.840 3.868

44.29 44.20

0.595

0.176

0.066

22.907

4.901

33.53

–1.224

0.678

27.075

9.069

18.88

0.053

0.028

18.056

0.000

45.35

Number of families (richness) BASIN + EUCALYPT

–0.793

0.326

EUCALYPT

–0.839

0.034

19.110

1.054

31.50

BASIN + EUCALYPT + FOREST 0.051

0.030

20.000

1.944

45.60

–0.731

0.431

0.104

0.443

EUCALYPT + FOREST

–0.659

0.453

20.648

2.592

33.60

0.283

0.442

BASIN + FOREST

0.045

0.031

0.592

0.360

FOREST

0.713

0.034

20.963

2.907

23.10

BASIN

0.061

0.031

21.736

3.680

19.60

20.795

2.739

32.93

Eucalypt plantations are the dominant vegetation in large areas of the river Lérez basin, and they are the main land use near many small rivers. As previously shown (see for example Larrañaga et al., 2009a), we observed that streams flowing through eucalypt plan� tations had lower taxon richness and less diversity of macroinvertebrates than those flowing through native forests (fig. 3). In fact, the percentage of the stream drainage covered by forest was negatively correlated

with eucalypt plantations cover. Moreover, as observed in our study, previous studies in Spain and Portugal on the effects of Eucalyptus plantations in streams have shown changes in macroinvertebrate communi� ties (Abelho & Graça, 1996; Larrañaga et al., 2009b). More precisely, although we did not analyze changes on feeding groups, we observed that the trophic structure of streams under native vegetation versus altered sites pointed to a change from shredder

Animal Biodiversity and Conservation 40.1 (2017)

Resemblance S7 Jaccard 2D Stress: 0

Eucalypt plantation > 10% Native forest < 20%

Stream A Ceira Acevedo As Laceiras As Penizas As Pozas O Cambado Fonte Seixiña Grande_1 Grande_2 O Moído Os Cabaleiros Os Calvos Os Ladróns Os Maneses Rego_1

Fig. 5. Bidimensional ordination (MDS) of all the sampled streams using presence/absence data of the macroinvertebrate community. Fig. 5. Ordenación bidimensional (MDS) de todos los arroyos muestreados usando datos de presencia y ausencia de la comunidad de macroinvertebrados.

dominance (relying on native litter–fall inputs, such as the caddisflies of the families Sericostomatidae and Brachycentridae) to collector dominance (detritivores of fine organic matter, such as simulids or limonids) at sites under Eucalyptus plantations. Álvarez et al. (2001) also observed comparable tendencies in temporary Mediterranean streams. Similarly, in a litter exclusion experimental study assessing the impact of eucalypt plantations on benthic macroinvertebrate communities, Larrañaga et al. (2006) found that shredders were less abundant at sites with eucalypt leaves than at sites with leaves collected from deciduous forests. Besides the potential effect of changes in vegeta� tion inputs on stream biota, land use changes have a direct effect on hydrology through their links with the evapotranspiration regime. Temperature is also of major importance for poikilothermic aquatic organisms due to its effects on physiology and behaviour. As shown elsewhere (see for example Allan, 2004b), the results of our study suggest that land use patterns also affect stream water temperature. In effect, the streams with a larger proportion of their drainage basin covered by forests were colder than those dominated by eucalypt plantations, probably as a result of the reduced flow and higher evaporation rates of the latter (Ferreira et al., 2006)�� , and the decrease in shading due to the verti� cal leaf orientation of eucalypt (James & Bell, 2000). In agreement with the general hydrological consequences of fast–growing tree plantations (see for example Jack� �� son et al., 2005; Oyarzún et al., 2005), we observed that in the second sampling period, water flow was reduced or had ceased altogether in some streams run� ning through basins with a high proportion of eucalypt

trees. However, given the small number of streams surveyed in this study, our data are not conclusive. Nevertheless, they agree with numerous examples in other regions. For instance, in the pampas grasslands of Argentina, Engel et al. (2005) observed that plantations of Eucalyptus camaldulensis (a phreatophytic species commonly grown in the region) often caused localized drawdowns of water tables, as a consequence of the species transpiration demand during dry periods. In China, in r Eucalyptus and Pinus plantations, it has been shown that erosion increased and water storage diminished in comparison with native vegetation (Hou et al., 2010). Similarly, a global analysis of 504 annual catchment observations (Jackson et al., 2005) showed that afforestation of grasslands, shrublands, or croplands with eucalypt and Pinus plantations decreased stream flow by 52% per year (227 mm) and that 13% of streams dried up for at least one year. In NW Spain, large areas have been deforested for centuries (Guitián Rivera & Cordero–Rivera, 2007) and even if eucalypt plantations do not usually substitute native forests that disappeared long time ago (but see Teixido et al., 2010), eucalypts have been introduced over shrub vegetation (mainly Ulex sp. and Erica sp.) and have spread invasively (Calviño–Cancela & Rubido–Bará, 2013), particularly after wildfires (Guitián Rivera & Cordero–Rivera, 2007). Therefore, in a climate change scenario, eucalypt plantations can be detrimental to water availability and rapidly accelerate drying of wet soils (Montoya, 1995). Another finding of interest is that in the same region of NW Spain, when eucalypts are defoliated by beetles, annual stream water can increase by 22% (Fernández et al., 2006).

Besides the effect of changes in basin land uses, our results also indicate that macroinvertebrate communi� ties were affected by basin size and land use patterns. Large areas are known to harbour more species due to the increase in resources and diversity of habitats (Shafer, 1990). A similar relationship usually occurs for river systems (Margalef, 1983). Accordingly, we observed that macroinvertebrate diversity increased with basin size, probably as the result of a combination of higher physical diversity (microhabitats) and more resources (more species of riverine trees, greater complexity of shoreline vegetation, and so on). Nevertheless, as shown in figure 2A, even streams draining very small basins, around 2–3 km2, were able to harbour most of the families of macroinvertebrates of the regional fauna. This is in agreement with recent studies showing the relevance of headwaters for riverine biodiversity con� servation (Finn et al., 2011). Our findings add evidence to previous studies that concluded that fast–growing tree plantations affect hydric resources and support the need to maintain and/or restore riparian forests to minimize the impacts of intensive industrial sylviculture on aquatic commu� nities. In conclusion, tree plantations cannot be used as substitutes of all ecosystem properties of forests, especially when plantations are established with exotic species (Cordero–Rivera, 2011, 2012). References Abelho, M. & Graça, M. A. S., 1996. Effects of Eucalyptus afforestation on leaf litter dynamics and macroinvertebrate community structure of streams in Central Portugal. Hydrobiologia, 324: 195–204. Allan, J. D., 2004a. Influence of land use and lands� cape setting on the ecological status of rivers. Limnética, 23: 187–198. – 2004b. Landscapes and riverscapes: the influence of land use on stream ecosystems. Annual Review of Ecology, Evolution, and Systematics, 35: 257–284. Allan, J. D. & Castillo, M. M., 2007. Stream ecology: structure and function of running waters. 2nd edi� tion. Chapman and Hall, New York. Álvarez, M., Pardo, I., Moyá, G., Ramón, G. & Martí� nez–Taberner, A., 2001. Invertebrate communities in temporary streams of the island of Majorca: A comparison of catchments with different land use. Limnética, 20: 255–266. Burnham, K. P., Anderson, D. R. & Huyvaert, K. P., 2011. AIC model selection and multimodel inferen� ce in behavioral ecology: some background, ob� servations, and comparisons. Behavioral Ecology and Sociobiology, 65: 23–35. Calviño–Cancela, M., 2013. Effectiveness of eucalypt plantations as a surrogate habitat for birds. Forest Ecology and Management, 310: 692–699. Calviño–Cancela, M., López de Silanes, M.E., Rubido– Bará, M. & Uribarri, J., 2013. The potential role of tree plantations in providing habitat for lichen epiphytes. Forest Ecology and Management, 291: 386–395. Calviño–Cancela, M. & Neumann, M., 2015. Ecologi� cal integration of eucalypts in Europe: Interactions

Cordero–Rivera et al.

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Brief communication 99

Animal Biodiversity and Conservation 40.1 (2017)

Feeding habits of garfish, Belone belone euxini Günther, 1866 in autumn and winter in Turkey´s south–east coast of the Black Sea Ş. Kaya & H. Saglam

Kaya, Ş. & Saglam, H., 2017. Feeding habits of garfish, Belone belone euxini Günther, 1866 in autumn and winter in Turkey’s south–east coast of the Black Sea. Animal Biodiversity and Conservation, 40.1: 99–102. Abstract Feeding habits of garfish, Belone belone euxini Günther, 1866 in autumn and winter in Turkey´s south–east coast of the Black Sea.— We studied the stomach content of Belone belone in the south–east Black Sea during autumn and winter months in 2010–2011. The most frequent feeding items in the diet were insects in autumn and fish in winter. Other items in the diet were mollusks, crustaceans and isopods. Flying ants were mostly consumed by male garfish, particularly the smaller fish, in autumn. Key words: Garfish, Belone belone euxini, Feeding habits, Black Sea, Flying ants Resumen Hábitos alimentarios de la aguja, Belone belone auxini Günther, 1866 en otoño e invierno en la costa suroriental del Mar Negro, en Turquía.— En el estudio se examinó el contenido estomacal de Belone belone en la costa suroriental del Mar Negro durante los meses de otoño e invierno de 2010 y 2011. Las hormigas voladoras constituyeron la presa más frecuente de la dieta en otoño, mientras que en invierno fueron los peces. Otras presas que aparecen en la dieta son moluscos, crustáceos e isópodos. Las hormigas voladoras son consumidas básicamente por los machos, en especial por los más pequeños, durante el otoño. Palabras clave: Aguja, Belone belone euxini, Hábitos alimentarios, Mar Negro, Hormigas voladoras Received: 31 V 16; Conditional acceptance: 9 IX 16; Final acceptance: 14 X 16 Şeyda Kaya, Agriculture and Rural Development Support Institution, Trabzon, Turkey.– Hacer Saglam, Fac. of Marine Science, Karadeniz Technical Univ., 61530 Camburnu, Trabzon, Turkey. Corresponding author: H. Saglam. E–mail: hacersaglam@yahoo.com

ISSN: 1578–665 X eISSN: 2014–928 X

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Introduction

Results

Garfish, Belone belone (Family Belonidae), are a commercially important pelagic species in Turkey, with landings amounting to 314.2 tones (t) in 2015. They are caught mostly along the Black Sea (139.1 t per year) and Marmara (134.7 t per year) (TUIK, 2015). Garfish is a pelagic species that moves to warmer coastal areas in summer months to spawn before returning to the deeper, open sea (Dorman, 1989; Zorica & Cikes Kec, 2012). The species matures to about 38 cm total length (TL) (Samsun et al., 2006). Young specimens are most common near the Cystoseira and Zostera belts (Radu et al., 1999). Garfish tend to leap and skitter at the surface (Collette, 2003). They are carnivorous, feeding primarily on small fishes which they catch sideways in their beaks (Collette, 2003). They can move for a considerable length of time at high speeds, exploring a large area in search of prey (Pavlov & Kasumyan, 2002). Studies on the dietary patterns of garfish (Belone belone) are scarce: Dorman, 1988 (Irish waters), Sever et al., 2009 (Aegean Sea) and Zorica & Cikes Kec 2012 (Adriatic Sea). This is the first study on the feeding habits of this species in the Black Sea region of Turkey. The objective of the paper was to study the dietary composition of Belone belone in autumn and winter by analyzing stomach contents.

We collected 533 specimens in autumn (October–November 2011) and 146 specimens in winter (December 2010–January 2011). The total sample size consisted of 679 specimens (405 females and 272 males). Total length (cm) ranged from 15 to 49.5 cm and total weight ranged from 70.0 and 120.0 g. The sample was composed of 39% males and 61% females. The sex ratio (F:M) was significantly different from 1:1 (P < 0.01; c2 = 26.173; SD = 1). Only 2% of the individuals analyzed had empty stomachs. The total number of prey was 2,785. The maximum number of prey per stomach was 200 and the mean number of prey per stomach was 21.28 ± 3.34. Table 1 shows the food composition of B. belone euxini in autumn and winter. We identified nine types of prey, most of them at family, genus or species level. The most frequent feeding items were insects in autumn and fish in winter. Less frequent items were mollusks, crustaceans, isopods, and unidentifiable matter. The index of relative importance showed that the diet of B. belone euxini in autumn was dominated by flying ants belonging to the Formicidae Family (Subfamily Myrmicinae; genus Pheidole Westwood) of insects (IRI% = 59.74). In autumn, the mean number of flying ants per stomach was 74.2. The maximum number of flying ants in a stomach was 200. Anchovy, Engraulis encrasicholus (IRI% = 28.25), was also an important prey in the diet in autumn. Garfish also fed on smaller garfish (table 1). In winter the preferred food item was fish (anchovy and unidentified fish) (P < 0.05). Shrimp larvae, isopods and copepods were found only in winter. Garfish also ate the young of their own species in winter (table 1). The small–size group (≤ 38 cm) consumed mainly flying ants while the large–size group (> 38 cm) consumed mainly fish (P < 0.05). Small garfish consumed copepods, but this was not an important prey. Female and male diets differed significantly (P < 0.05) (table 1). Females (35.6 ± 3.9 cm) were significantly larger than males (31.6 ± 10.8 cm) (P < 0.05). In both same size groups, males consumed more flying ants than females (table 1). The highest value of diet overlap was 0.71 between females and males, and the lowest was 0.28 between autumn and winter. This index was 0.55 between small– and large–sized groups.

Material and methods Garfish were obtained from commercial catches during autumn and winter months 2010–2011 in the Trabzon region of the south east Black Sea. The samples were collected from artisanal fishermen using garfish trammel nets. We collected specimens in autumn (October–November 2011) and winter (December 2010–January 2011). Total length (TL) measured to the nearest cm and total weight (TW) were recorded for each fish. Stomach contents were removed and frozen at –18˚C. We were unable to identify food items from stomach contents to the lowest possible taxa due to the high degree of digestion of prey. Data analysis was carried using the index of relative importance (IRI) (Pinkas et al., 1971): IRI = O% (N% + W%) where O% is frequency of occurrence of each item, N% is the percentage of total number of food items, and W% is the percentage of total weight of stomach contents (wet weight) calculated for each food category. IRIs were standardized to 100% by calculating the percentage of total IRI contributed by each prey type (IRI %) (Hyslop, 1980). The sample was split into two groups according to size: small fish (≤ 38 cm TL) and large fish (≥ 38 cm TL). We compared stomach contents of similar size classes between sexes (to avoid the effect of size on sex). Schoener’s diet overlap index (Schoener, 1970) was used to measure diet overlap between size, sex and season. Schoener’s index values above 0.6 are usually considered to indicate significant overlap.

Discussion The dietary pattern of the garfish, B. belone, varies depending on geographic location. The most important prey in the Middle Eastern Adriatic Sea are copepods (Sever et al. 2009 in Aegean Sea; Zorica & Cikes Kec., 2012) while in Irish waters the most common food items are crab larvae, and clupeids (mainly juveniles) (Dorman, 1988). In the Black Sea, garfish feed on small fishes, particularly clupeids and anchovy (Engraulis sp.) (Collette & Parin, 1986; Dorman, 1988; Sever et al., 2009). Although in the present study we found that garfish feed mainly on flying ants in autumn, fish (particularly clupeids and anchovy) constitute its

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Table 1. Diet composition of garfish, B. belone euxini, from southeastern Black Sea in autumn and winter: %N. Percentage by number; %W. Percentage by weight; %O. Percentage frequency of occurrence; %IRI. Percentage index of relative importance: W. Winter; A. Autumn; F. Female; M. Male. Tabla 1. Composición de la alimentación de la aguja, B. belone euxini, en la costa suroriental del Mar Negro en otoño e invierno: %N. Porcentaje por número; %W. Porcentaje por peso; %O. Porcentaje de la frecuencia de presencia; %IRI. Porcentaje del índice de importancia relativa: W. Invierno; A. Otoño; F. Hembra; M. Macho.

Composition

% IRI

Length (cm) ≤ 38

> 38

Flying ant

90.8

11.7

28.5 59.7

69.0 23.0

4.1 61.7

47.8 73.4

Anchovy

1.5

54.1

24.8 28.3

20.8 55.9

18.7 27.7

38.0 14.9

Red mullet

0.1

5.8

2.9

0.4

0.1

0.9

0.3

Sex

Prey

Young garfish

%IRI

Season

0.4

0.0

< 0.1

0.3

0.7

0.0

< 0.1

0.0

0.0 < 0.1

0.8

19.2

16.8

6.9

5.1

14.0

39.4 5.2

6.1

Isopoda

< 0.1

0.0

0.7 < 0.1

< 0.1

0.0

0.4

0.0

Copepod

0.1

< 0.1

0.7 < 0.1

< 0.1

0.0

0.8

0.0

Shrimp larvae

1.9

1.7

1.5

< 0.1

0.0

28.6 0.0

0.0

0.6

Gastropod larvae

1.2

0.8

0.7 < 0.1

0.2

0.0

8.1

0.0

0.1

0.0

Unidentified tissues

3.5

6.4

22.6

4.4

6.9

0.0

5.1

7.1

2.1

Unidentified fish

0.1 4.6

main food item in the other seasons —as indicated by our data and other works performed in the region (Yüce, 1975; Colette & Parin, 1986; Dorman, 1988). In Turkey’s south east coast of the Black Sea, 78% of total marine fish production is anchovy and sprat of small pelagic fish (TUIK, 2015). Therefore, although ants are important in autumn, two small pelagic fish (anchovy and sprat) are the basis of the garfish diet (if we consider the whole year) due to their high abundance in this region. Our study also confirms the observations of Sever et al. (2009) and Dorman (1988) that garfish display cannibalism of smaller specimens. Although Zorica & Cikes Kec (2012) reported that the composition of the garfish diet was not size related, we found differences in the present study related to size. We observed that fish prey (IRI%) made up 20.84% of the diet for garfish ≤ 38 cm TL but 55.88% for garfish > 38 cm TL. As Samsun et al. (2006) found that females reached 50% sexual maturity at 38.8 cm length, the differences in diet in these size groups (< 38 cm vs. > 38 cm) could be related to maturity. Similarly to our study, Dorman (1988) found size–related differences in the diet of garfish in Irish waters: garfish of < 70 cm preyed on insects more often than fish, although all sizes fed predominantly on crustaceans. Sever et al. (2009) also found that members of the Formicidae Family were an important prey. They

9.1

< 0.1 0.0 0.0 < 0.1

reported an index of relative importance of 308 (1.2%), 2,385 (17.2%) and 3,176 (32.5%) in March, April and August respectively (the %IRI was calculated by ourselves). However, we observed flying ants in the diet of garfish in October. This difference may be due to the availability of different insect species. Dorman (1988) stated that garfish in Irish waters fed on different orders of insects (Hemiptera, Diptera, Coleoptera and Hymenoptera), captured in summer (June–September), but the proportion of insects in the diet in these fish in Irish waters was lower than that in our study in the Black Sea. Ants are abundant and dominant members of almost every terrestrial ecosystem. Nuptial flight of ants is a common phenomenon, usually occurring in summer or autumn, depending on the species (Dunn et al., 2007). Different ant species have different requirements concerning the weather to begin the nuptial flight. Nene et al. (2016) found that nuptial flights occurred during the raining season and on days with high relative humidity and less sunshine. The climate of the Black Sea coastal region has high humidity and rain falls throughout the year, mostly in autumn and winter. In rainy conditions, the flying ants may reach the coast in rainy weather, with large clouds and wind, and they may drown at sea. Garfish likely eat any masses of dead ants floating on the surface of the sea. In other seas and oceans flying ants are often seen in summer and autumn during

102

the nuptial flights (Dunn et al., 2007) but information about distribution and the timing of nuptial flights of this ant species in the Black Sea is lacking. This is the first study to investigate the dietary pattern of garfish along the Turkish coast of the Black Sea but further research is needed to understand more about the trophic relationship between insects and garfish in this region. Acknowledgements The authors wish to thank Dr. Nihat Aktaç (Emeritus Prof.) for identifying the ant species and the Editor and anonymous reviewers for their constructive comments during the review process. The present study was carried out within the framework of the research project 'Feeding habits of garfish, Belone belone, in the southeastern Black Sea coast' funded by Karadeniz Technical University BAP Project no: 2009.117.01.4 References Collette, B. B., 2003. Family Belonidae Bonaparte 1832 needlefishes. Calif. Acad. Sci. Annotated Checklists of Fishes, 16: 22. Collette, B. B. & Parin, N. V., 1986. Belonidae. In: Fishes of the Northeastern Atlantic and Mediterranean: 604–609 (P. J. P. Whitehead, M.–L. Bauchot, J. C. Hureau, J. & Nielsen, E. Tortonese, Eds.), Unesco, Paris. Dorman, J. A., 1988. Diet of the garfish, Belone belone (L.), from Courtmacsherry Bay, Ireland. Journal of Fish Biology, 33(3): 339–346. – 1989. Some aspects of the biology of the garfish Belone belone (L.) from southern Ireland. Journal of Fish Biology, 35: 621–629. Dunn, R. R. Parker, C. R. Geraghty, M. & Sanders M., 2007. Reproductive phenologies in a diverse temperate ant fauna. Ecological Entomology, 32: 135–142.

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Trends in the extinction of carnivores in Madagascar B. Cartagena–Matos, I. Gregório, M. Morais & E. Ferreira

Cartagena–Matos, B., Gregório, I., Morais, M. & Ferreira, E., 2017. Trends in the extinction of carnivores in Madagascar. Animal Biodiversity and Conservation, 40.1: 103–114. Abstract Trends in the extinction of carnivores in Madagascar.— The extinction of top predators, such as mammalian carnivores can lead to dramatic changes in foodweb structure and ecosystem dynamics. Since all native Malagasy terrestrial mammalian carnivores are endemic, their extinction implies a significant loss of biodiversity in Madagascar. Here we review the literature on Madagascar’s mammalian carnivores, aiming to determine which species are most likely to become extinct in the near future in view of the factors threatening their survival. We scored each factor according to its impact on the species. According to our results, the giant–striped mongoose, Galidictis grandidieri, is the most likely species to next become extinct. This is no surprise because this species is considered one of the rarest carnivores in the world, inhabiting only a small, threatened forest ecosystem. Our results emphasize the need for robust data about each species to help and support decision–makers implement conservation measures. Key words: Eupleridae, Endemism, Biodiversity loss, Human impacts, Deforestation, Interspecific competition Resumen Tendencias de la extinción de carnívoros en Madagascar.— La extinción de los depredadores apicales, como los mamíferos carnívoros, puede conllevar cambios drásticos en la estructura de la red alimentaria y la dinámica de los ecosistemas. Dado que todos los mamíferos carnívoros terrestres autóctonos de Madagascar son endémicos, su extinción implica una pérdida notable de biodiversidad en este país. En el presente artículo examinamos las publicaciones sobre mamíferos carnívoros de Madagascar con el propósito de determinar cuáles son las especies que tienen mayor probabilidad de extinguirse en un futuro próximo, en vista de los factores que amenazan su supervivencia. Puntuamos cada factor en función de los efectos que ejerce en las especies. Según nuestros resultados, la especie que tiene más probabilidad de extinguirse es la mangosta rayada grande, Galidictis grandidieri, lo cual no es sorprendente porque esta especie se considera uno de los carnívoros más escasos del mundo, que habita solo en un ecosistema forestal pequeño y amenazado. Asimismo, nuestros resultados ponen de manifiesto la necesidad de disponer de datos sólidos sobre cada especie, a fin de ayudar y respaldar a las autoridades a poner en práctica medidas de conservación. Palabras clave: Eupleridae, Endemismo, Pérdida de biodiversidad, Efectos de los humanos, Deforestación, Competencia interespecífica Received: 25 II 16; Conditional acceptance: 25 VIII 16; Final acceptance: 14 X 16 Bárbara Cartagena–Matos, Inês Gregório & Marta Morais, �� Dept. of Biology, Univ. of Aveiro, Campus �� Universi� tário de Santiago, 3810–193 Aveiro, Portugal.– Eduardo Ferreira, Dept. of Biology & CESAM, Univ. of Aveiro, Campus Universitário de Santiago, 3810–193 Aveiro, Portugal. Corresponding author: B. Cartagena–Matos. E–mail: barbara.cartagena.matos@gmail.com

ISSN: 1578–665 X eISSN: 2014–928 X

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Introduction Mammalian carnivores are susceptible to local extinction through habitat loss and fragmentation, mainly due to their fairly large distribution ranges, small population size and conflicts with humans (Woodroffe & Ginsberg, 1998; Crooks, 2002; Logan et al., 2015). Moreover, they are top predators, so their extinction can seriously affect food webs and ecosystem dynamics (Crooks, 2002). The effects of anthropogenic activities in Madagascar and their impact on carnivore populations are poorly understood (Gerber et al., 2010; Logan et al., 2015). Madagascar is one of twenty–five global biodiversity hotspots, harbouring almost 3% of the world’s endemic vertebrates (Myers et al., 2000). Since all Malagasy native terrestrial mammalian carnivores are endemic (Yoder et al., 2003; Duckworth et al., 2014), their extinction would represent a significant biodiversity loss, both in Madagascar and globally. Carnivora is one of four terrestrial mammalian orders occurring in Madagascar, but it is represented by only one family, Eupleridae (Yoder et al., 2003). This family comprises twelve species and subspecies and includes an extinct species and a newly discovered species (Albignac, 1972; Durbin et al., 2010; Goodman & Helgen, 2010). The only carnivore known to be extinct in Madagascar is the giant fossa (Cryptoprocta spelea Grandidier, 1902), considered a larger relative of the extant fossa, Cryptoprocta ferox Bennet, 1833 (Hoffman & Hawkins, 2015). The reasons for its extinction are unclear, but likely due to the loss of their main prey (such as giant lemurs) and extensive habitat destruction (Yoder et al., 2003; Goodman et al., 2004). Although the Giant fossa is the only carnivore known to be extinct in Madagascar, others extinctions may have occurred. There are also three introduced carnivore species in Madagascar: the domestic dog (Canis familiaris Linnaeus, 1758), the feral cat (Felis sp. Linnaeus, 1758), and the small Indian civet (Viverricula indica É. Geoffroy Saint– Hilaire, 1803) (Gerber et al., 2010; Farris et al., 2015). Co–occurrence of native and exotic carnivores may alter ecological dynamics, such as predation, competition or resource use (Hunter & Caro, 2008; Vanak & Gompper, 2010). Farris et al. (2015) identified a strong temporal overlap between native and introduced carnivores in Madagascar, with the small Indian civet presenting the greatest overlap with native Malagasy carnivores. Here we review the ecological characteristics of Malagasy carnivores and their current threats and relevance. Our aim is to review and analyse current knowledge on the ecology, conservation and threats to Malagasy carnivores of the family Eupleridae. Based on this knowledge, we try to identify current conservation priorities and predict which species are currently more susceptible to extinction risk in the near future. Material and methods Literature search This is a theoretical study based on information compiled from available literature and IUCN (International

Union for Conservation of Nature and Natural Resources) data. We restricted our search to international peer–reviewed manuscripts and books. We searched for scientific papers on the Web of Science™ database using the following keywords: 'Madagascar', Malagasy species scientific and common names, 'carnivore', 'biodiversity loss', 'deforestation', 'conservation', and scientific and common names of introduced mammalian carnivores. We reviewed 81 scientific papers and eight books. Moreover, we sourced information on the IUCN about the mammalian carnivores studied, totalling 97 references (list provided on supplemental material, table 1sS1). Furthermore, the IUCN Red List kindly provided species distribution GIS shapefiles, which were useful to calculate distribution areas and generate maps for species’ distribution (fig. 1, table 1). Data analysis To evaluate the risk of extinction, we generated a table whereby we averaged the standardized scores for the following factors affecting Malagasy carnivores: distribution range, species information, dietary breadth, dietary overlap, habitat breadth, strata overlap, activity pattern overlap, forested areas within distribution range in 2015, and deforestation between 2007 and 2015. Species information was estimated based on the number of references indexed internationally in which the species is evaluated (see table 1s in supplementary material). To assess dietary breadth, we put food items together in categories, namely fruits, eggs, invertebrates, amphibians and reptiles, fish, birds, lemurs, and other small mammals, and then counted how many categories of food items each species uses (table 1). Habitat breadth for each species was obtained from PANTHERIA (Jones et al., 2009). To quantify the possible ecological overlap between species, we estimated dietary, strata and activity pattern overlaps using data on diet, compiled from the literature, and data on arboreality and activity patterns, obtained from PANTHERIA (equations provided in appendix 1s in supplementary material). We assumed that more ecologically similar species would be most affected by sharing scarce resources. The distribution range was calculated in QGIS (Brighton version 2.6.0), based on shapefiles provided by the IUCN, using the Albers equal–area conical projection (EPSG: 102 022). We assessed deforestation in each species' distribution range, using QGIS, by comparing the amount of forested and non–forested areas in 2007 and 2015 (within each species' distribution range), and using the online updated data on global forest/non W forest maps from ALOS PALSAR Data (Shimada et al., 2014; updated data available on http://www.eorc.jaxa. jp/ALOS/en/palsar_fnf/fnf_index.htm, June 2016). The percentage of forested areas in 2015 in each species distribution range was estimated using the same dataset. Although we are aware of the different weights that each factor may have in the ecology and probability of extinction of species, because of lack of information in the literature, we considered each factor to be of the same relevance.

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N Cryptoprocta ferox Eupleres goudotti Fossa fossana Galidia elegans Galidictis fasciata Galidictis grandidieri Mungotictis decemlineata Salanoia concolor 0 80 160

320 km

Fig. 1. Geographic distribution of the carnivore mammal species in Madagascar. Species distribution data were obtained from the IUCN Red List. Fig. 1. Distribución geográfica de las especies de mamíferos carnívoros en Madagascar. Los datos sobre la distribución de las especies se obtuvieron de la Lista Roja de la UICN.

The scores for each factor were represented with a positive sign if the factor was favourable to conservation and a negative sign if the factor threatened conservation. Scores for each factor were standardized, resulting in variables with 0 mean and unit standard deviation, and with the most threatened species presenting the most negative score values. Scores for each factor were averaged for each species, and 95% confidence intervals were estimated for the species average scores. Results Species accounts The fossa (Cryptoprocta ferox, Bennet, 1833) is found at low densities, in forested areas across the whole island except for the central plateau (Goodman et al., 2004; Hawkins & Racey, 2005; Hawkins & Dollar, 2008). Fossa prey on a variety of vertebrates (small mammals, lemurs, reptiles and amphibians) and occasionally feed on domestic animals such as pigs and poultry (Hawkins & Racey, 2005). Their

period of daily activity overlaps with that of Eupleres goudotii goudotii, Fossa fossana, Galidictis fasciata, Viverricula indica, and feral cats and dogs (Farris et al., 2015). Moreover, it exhibits a dietary overlap with Galidia elegans (Gerber et al., 2012a). With no natural predators, humans are the biggest threat for fossa as they are hunted out of fear and to protect livestock (Goodman et al., 2004; Hawkins & Racey, 2005; Hawkins & Dollar, 2008). Deforestation is destroying the formerly broad habitat of fossa (Goodman et al., 2004; Hawkins & Racey, 2005; Hawkins & Dollar, 2008). Fossa fossana (P. L. S. Müller, 1776), also known as the Malagasy civet, is the third largest carnivore in Madagascar (Gerber et al., 2012a). It is nocturnal and distributed across the eastern territory of the island in tropical low–land and mid–altitude forests (Kerridge et al., 2003; Hawkins, 2008a). It is a generalist predator, including crustaceans, reptiles, rodents and amphibians in its diet (Goodman et al., 2003). The species’ abundance has decreased due to deforestation (excessive logging or land use change for agriculture) and due to competition with other species (Hawkins, 2008a).

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Table 1. Diet, activity pattern, strata of occurrence, habitat breadth and distribution range for the eight endemic carnivore species in Madagascar. Data in this table were used to estimate the factors presented in table 2 and figure 2: Fr. Fruits; Eg. Eggs; In. Insects and other arthropods; He. Herps (amphibians and reptiles); Fi. Fish; Bi. Birds; Le. Lemurs; Sm. Small mammals.

Species

Diet Bi

Cryptoprocta ferox

Eupleres goudotii

Fossa fossana Galidia elegans

X X

Galidictis fasciata

Galidictis grandidieri

Mungotictis decemlineata

Salanoia concolor

The taxonomic organization of the genus Eupleres has been widely discussed over the years. Albignac (1973) considered Eupleres to be monospecific and represented by two subspecies: E. goudotii goudotii (Doyère, 1835) and E. goudotti major (Lavauden, 1929). However, recently, Goodman & Helgen (2010) proposed that these subspecies could be elevated to the rank of species based on subfossil evidence. Here we consider only Eupleres goudotii goudotii because it is the subspecies for which most relevant information is available. The small–toothed mongoose (Eupleres goudotii goudotii) is thought to be very uncommon across its range, which includes the east coast and the north of Madagascar (Albignac, 1972; Dollar, 2000). Its diet consists mainly of earthworms, but it occasionally feeds on amphibians and insects (Albignac, 1972; Macdonald, 1992; Garbutt, 1999). The main threat for this species is deforestation caused by slash–and–burn agriculture, logging and charcoal production (Schreiber et al., 1989; Nowak, 1999). It is also the only species selectively hunted for bushmeat (Dollar, 2000; Logan et al., 2015). E. g. goudotii and F. fossana have highly similar activity profiles with the introduced V. Indica, but there are considerable differences in niche requirements between these two native carnivores and V. indica (Farris et al., 2015). The giant–striped mongoose (Galidictis grandidieri, Wozencraft, 1986) is one of the rarest carnivores in the world (Andriatsimietry et al., 2009; Marquard et al., 2011). It inhabits a small,unique spiny forest ecosystem that is threatened by human activity, and it is also preyed on by introduced dogs (Hawkins, 2008b; Marquard et al., 2011). It preys on tortoise eggs, invertebrates and some vertebrates (Andriatsimietry et al., 2009; Currylow, 2014). G. grandidieri co–occurs

with C. ferox and with the introduced civet species V. indica, but has no dietary overlap with the latter. Competition for food with the latter is unlikely, as C. ferox preys more on vertebrates than the giant–striped mongoose (Andriatsimietry et al., 2009). The broad–striped mongoose (Galidictis fasciata Gmelin, 1788) is found only in the eastern rainforests of Madagascar (Garbutt, 1999; Goodman, 2003b). It is a generalist predator, feeding on rodents, small lemurs, reptiles, small amphibians and invertebrates (Garbutt, 1999; Goodman, 2003b). This species has not been extensively studied, perhaps due to their strictly nocturnal habits (Goodman & Pidgeon, 1998; Garbutt, 1999; Nowak, 1999; Goodman, 2003b). Like most Malagasy carnivores, Galidictis fasciata is threatened by deforestation and by direct competition with feral cats and dogs (Hawkins, 2008c; Farris et al., 2015). The Malagasy ring–tailed mongoose (Galidia elegans, I. Geoffroy Saint–Hilaire, 1837) is a well–studied diurnal carnivore that has become very common in disturbed habitats (Hawkins, 2008d; Bennett et al., 2009; Farris et al., 2014). It has even been seen following groups of tourists for food waste (Hawkins, 2008c; Bennett et al., 2009; Farris et al., 2014). Three subspecies are currently recognized: G. e. elegans found in the eastern rainforests; a western race, G. e. occidentalis, found in deciduous forests in the central western parts; and a northern race, G. e. dambrensis (Hawkins, 2008d; Schnoell, 2012). Despite its population decline of over 20 percent in the last ten years, probably due to habitat loss, G. elegans occurs in secondary habitats, at forest edges, and in exotic tree plantations near native forests (Hawkins, 2008d; Irwin et al., 2010). It preys on lemurs, like C. ferox, V. indica, and feral cats and dogs, but it also consumes

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Tabla 1. La alimentación, el régimen de actividad, los estratos de presencia, la amplitud del hábitat y el rango de distribución de las ocho especies endémicas de carnívoros en Madagascar. Los datos de esta tabla se emplearon para calcular los factores presentados en la tabla 2 y la figura 2: Fr. Frutos; Eg. Huevos; In. Insectos y otros artrópodos; He. Anfibios y reptiles; Fi. Peces; Bi. Aves; Le. Lemures; Sm. Pequeños mamíferos.

Activity pattern

Strata Arboreal

Terrestrial

Habitat Breadth

Distribution range (thousands of km2)

0.5

447.59

103.13

92.10

Diurnal

0.5

Nocturnal

93.50

0.5

68.30

0.5

1.01

0.5

13.43

18.75

other small mammals, invertebrates, reptiles, fish, birds, eggs, and fruit (Nowak, 1997; Garbutt, 1999; Hawkins, 2008d; Farris et al., 2014, 2015). Its daily activity period overlaps with that of the brown–tailed mongoose, Salanoia concolor (Farris et al., 2014, 2015). G. elegans is hunted in some areas, persecuted for raiding local poultry and killed by dogs (Hawkins, 2008d). The tail of the animal is known to be used for cultural purposes by some tribal groups (Goodman, 2003a; Hawkins, 2008d). The little–known brown–tailed mongoose, Salanoia concolor (I. Geoffroy Saint–Hilaire, 1837), is a diurnal species and is most frequently observed in relatively undisturbed rainforests (Hawkins et al., 2008). It feeds on small birds, mammals and coleopteran larvae (Albignac, 1972; Britt, 1999; Britt & Virkaitis, 2003). Like most Malagasy carnivores, S. concolor is believed to be threatened by deforestation (Hawkins et al., 2008; Farris & Kelly, 2011). This species presents high activity overlap with dogs and moderate overlap with feral cats (Farris et al., 2015). Salanoia durrelli is a newly–discovered species in Madagascar. It has been separated from S. concolor based on morphological and molecular traits (Durbin et al., 2010). Since this species is not listed in the IUCN Red List and almost no data have been published on it (at least in publicly accessible sources), we do not include it in our analyses. The Malagasy narrow–striped mongoose (Mun� gotictis decemlineata A. Grandidier, 1867) is relatively common within a small area of the deciduous forests of Menabe in the southwest of Madagascar (Schreiber et al., 1989; Hawkins, 2008e). This species is diurnal, terrestrial and mainly insectivorous, but complements its diet with small vertebrates (Rabeantoandro, 1997). Currently, the population of

M. decemlineata is threatened by predation by dogs (Hawkins, 2008e). In addition, habitat degradation caused by intensive logging and pasture conversion, and increased hunting by humans, contribute to its vulnerable status (Goodman & Raselimanana, 2003; Hawkins, 2008e). Summary of the factors affecting Malagasy carnivore species The scores for each factor affecting Malagasy terrestrial carnivore species were estimated and represented with either a positive or a negative sign depending on whether the factor was favourable (+) or unfavourable (–) to the conservation of the species (fig. 2). These scores were later standardized, resulting in variables with 0 mean and unit standard deviation, with the most threatened species presenting the most negative score values (table 2, fig. 3). Moreover, scores for the various factors were averaged for each species, and 95% confidence intervals were estimated for the species average scores (table 2, fig. 3). Except for Galidictis grandidieri, the average level of threat did not differ significantly (fig. 3). G. grandidieri appears to be the species most likely to become extinct first, with all factors negatively influencing its score (Av. Std. Score = –0.77). It is followed by M. decemlineata, G. fasciata, and S. concolor, which also had negative average standardized scores, though these were not significant (table 2, fig. 3). Galidia elegans appears to be the least threatened Malagasy carnivore species (Av. Std. Score = + 0.76), unlikely to disappear from Madagascar in the near future, followed by C. ferox, F. fossana, and E. g. goudotii, which also had positive average standardized scores, though not significant (table 2, fig. 3).

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Distribution range

Dietary breadth 8

5.00E+05 4.00E+05

3.00E+05

2.00E+05 2

1.00E+05 0.00E+00

Cf Egg Ff Ge Gf Gg Md Sc

Habitat breadth

Strata overlap

–2

Cf Egg Ff Ge Gf Gg Md Sc

–4

2 –6

1 0

Cf Egg Ff Ge Gf Gg Md Sc

–8

Dietary overlap 0.00 –0.10 –0.20 –0.30 –0.40 –0.50 –0.60 –0.70 –0.80 –0.90

Cf Egg Ff Ge Gf Gg Md Sc

Activity pattern overlap

–1

–3

Cf Egg Ff Ge Gf Gg Md Sc

–5

Cf Egg Ff Ge Gf Gg Md Sc

Fig. 2. Species scores (non–standardized) for factors affecting Malagasy terrestrial carnivores. Factors with an inferred negative impact are presented with negative scores while factors with a positive impact are shown with positive scores: Cf. Cryptoprocta ferox; Egg. Eupleres goudotti goudotti; Ff. Fossa fossana; Ge. Galidia elegans; Gf. Galidictis fasciata; Gg. Galidictis grandidieri; Md. Mungotictis decemlineata; Sc. Salanoia concolor. Fig. 2. Puntuaciones (no estandarizadas) de las especies con respecto a los factores que afectan a los carnívoros terrestres de Madagascar. Los factores que ejercen un efecto negativo se presentan con puntuaciones negativas, mientras que los que tienen un efecto positivo se muestran con puntuaciones positivas: Cf. Cryptoprocta ferox; Egg. Eupleres goudotti goudotti; Ff. Fossa fossana; Ge. Galidia elegans; Gf. Galidictis fasciata; Gg. Galidictis grandidieri; Md. Mungotictis decemlineata; Sc. Salanoia concolor.

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Fig. 2. (Cont.) Forested areas (2015)

Deforestation (2007–2015)

100%

80%

–10%

60%

–20%

40%

–30%

20%

–40%

Cf Egg Ff Ge Gf Gg Md Sc

–50%

Cf Egg Ff Ge Gf Gg Md Sc

Species information 50 40 30 20 10 0

Cf Egg Ff Ge Gf Gg Md Sc

Discussion We reviewed the literature on the terrestrial carnivore mammals of Madagascar, aiming to understand which species are more likely to go extinct in the near future, taking into account the factors that might be affecting their survival. We used resources such as available literature and data on different online datasets (PANTHERIA, IUCN, ALOS PALSAR Data) to assess the major factors affecting these species. According to our results, only Galidictis grandidieri was revealed to be significantly more threatened. However, final score values showed four species are likely to be more threatened than the other four, with the giant–striped mongoose Galidictis grandidieri most likely to become extinct first. All factors seem to have a negative impact on this species' risk of extinction. This came as no surprise because G. grandidieri inhabits a small, unique spiny forest ecosystem that is threatened by anthropogenic impacts (Hawkins, 2008b; Marquard et al., 2011). Moreover, it is considered one of the rarest carnivores in the world, with the smallest range of all Malagasy carnivores (Andriatsimietry et al., 2009; Marquard et al., 2011). Also, this species has a narrow

dietary breadth that overlaps with both diet and activity pattern of Cryptoprocta ferox, while there is also a generalized lack of information on its biology (Hawkins, 2008b; Andriatsimietry et al., 2009; Currylow, 2014). Our results are in agreement with the IUCN classification, which considers G. grandidieri one of the most endangered Malagasy carnivore species (Hawkins, 2008b). The Malagasy narrow–striped mongoose, Mungotictis decemlineata, is considered Vulnerable by the IUCN (Hawkins, 2008e). Our results support this. According to the available literature and online datasets, deforestation has a devastating effect on M. decemlineata, with the healthiest population found within the least disturbed forest in its range, and it has one of the smallest distribution ranges of the mammalian carnivores considered here (Woolaver et al., 2006; Hawkins, 2008e). The broad–striped mongoose, Galidictis fasciata, is considered Near Threatened by the IUCN (Hawkins 2008c), which is not unlike findings in our analysis. G. fasciata has a generalist diet, consuming mammals, amphibians, reptiles, and some invertebrates (Garbutt, 1999; Goodman, 2003b), and, in addition, it presents a large distribution range (Hawkins, 2008c; our analysis). However, this spe-

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Table 2. Standardized scores for each of the nine factors used to estimate the average threat score for each of the eight Malagasy carnivores. Average standardized scores (and 95% confidence intervals) are also provided for each species. Most negative average scores correspond to most threatened species, while most positive average scores correspond to least threatened species: DR. Distribution range; HB. Habitat breadth; SO. Strata overlap: APO. Activity pattern overlap; DB. Diet breadth; DO. Diet overlap; FA. Forested areas (2015); D. Deforestation (2007–2015); SI. Species information; ASS. Average standard score (± 95% CI). Tabla 2. Puntuaciones estandarizadas de cada uno de los nueve factores utilizados para calcular la puntuación media de la amenaza para cada uno de los ocho carnívoros de Madagascar. También se proporciona la media de las puntuaciones estandarizadas (y los intervalos de confianza del 95%) de cada especie. La mayoría de las puntuaciones medias negativas corresponden a las especies más amenazadas, mientras que la mayoría de las puntuaciones medias positivas corresponden a las especies menos amenazadas: DR. Rango de distribución; HB. Amplitud del hábitat; SO. Superposición de estratos: APO. Superposición de patrones de actividad; DB. Ancho de la dieta; DO. Superposición de la dieta; FA. Zonas forestales (2015); D. Deforestación (2007–2015); SI. Información sobre especies; ASS. Puntuacion estándar media (± 95% IC). Species

APO

ASS

1.00

1.62

–1.05

0.48

0.41

–1.39

0.32

2.19

0.66 (± 1.00)

–0.78 –0.54

1.35

–0.29 –0.26

0.30

0.58

0.40

0.08 (± 0.50)

Cryptoprocta ferox

2.38

Eupleres g. goudotii

–0.01

Fossa fossana

–0.09

0.11

–0.54

1.35

0.48

0.12

0.51

0.53 –0.11 0.26 (± 0.41)

1.89

–0.54

0.15

2.02

2.25

0.49

0.51

–0.78 –0.54

–1.05

–0.29 –0.63

0.67

0.45 –0.50 –0.32 (± 0.43)

–0.78 –0.54

–1.05

–1.06 –0.88 –0.41

–0.38 –1.13 –0.77 (± 0.22)

1.62

0.15

–0.29 –0.63 –1.46

–2.36 –0.50 –0.44 (± 0.85)

–0.78 –0.54

0.15

–1.06 –0.38

0.35 –0.50 –0.23 (± 0.55)

Galidia elegans

–0.08

0.14

0.76 (± 0.78)

Galidictis fasciata

–0.25

Galidictis grandidieri

–0.72

Mungotictis decemlineata

–0.63

0.11

Salanoia concolor

–0.60

cies' diet and activity pattern overlap with both native and introduced carnivores living in the same area as G. Fasciata that can have a negative impact on its risk of extinction (Hawkins, 2008c; Farris et al., 2015). The brown–tailed mongoose (Salanoia concolor) is considered Vulnerable by the IUCN (Hawkins et al., 2008), again supported by our analysis. This species is highly negatively affected by lack of information, small distribution range and habitat breadth, and restricted diet that overlaps with several other species (Farris et al., 2015; our analysis). Although there is currently no evidence of predation by humans or competition with exotic carnivores, S. concolor has been found to be absent from sites where feral cats occur (Albignac, 1972; Hawkins et al., 2008; Farris et al., 2012). As

1.31

there is little information on S. concolor, the IUCN recognizes the need to revaluate the extent of threats to the species, which might warrant reclassification from Vulnerable to Endangered (Hawkins et al., 2008). The ring–tailed mongoose, Galidia elegans, is considered Least Concern by the IUCN (Hawkins, 2008d), in agreement with our analysis. Its widespread distribution range occurs in a number of protected areas and forested fragments (Hawkins, 2008d). This species also has a large variety of prey n its diet, but it has a high dietary overlap with both native and exotic species (Hawkins, 2008d; Farris et al., 2014, 2015; our analysis). In contrast with our findings, the fossa (Cryptoprocta ferox) is considered Vulnerable by the IUCN (Hawkins & Dollar, 2008). Although it

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2.0

111

Average standardized score

1.5 1.0 0.5 0.0 –0.5 –1.0 –1.5

Egg

Fig. 3. Mean standardized scores per species, averaged across all scores for all factors (table 2), with 95% confidence intervals: Cf. Cryptoprocta ferox; Egg. Eupleres goudotti goudotti; Ff. Fossa fossana; Ge. Galidia elegans; Gf. Galidictis fasciata; Gg. Galidictis grandidieri; Md. Mungotictis decemlineata; Sc. Salanoia concolor. Fig. 3. Puntuaciones medias estandarizadas por especie, teniendo en cuenta las puntuaciones de to� dos los factores (tabla 2), con intervalos de confianza del 95 %): Cf. Cryptoprocta ferox; Egg. Eupleres goudotti goudotti; Ff. Fossa fossana; Ge. Galidia elegans; Gf. Galidictis fasciata; Gg. Galidictis grandidieri; Md. Mungotictis decemlineata; Sc. Salanoia concolor.

seems that C. ferox is one of the least threatened Malagasy carnivore species according to our results, we acknowledge its vulnerability, mainly caused by its ecological overlap with all native and introduced carnivores, and human–related conflicts (Hawkins & Dollar, 2008; Gerber et al., 2012a; Farris et al., 2015; Logan et al., 2015). It is of considerable concern that deforestation is destroying the habitat of the formerly widely distributed C. ferox (Goodman et al., 2004; Hawkins & Racey, 2005). Nevertheless, fossa is a well–studied, generalist predator, with a high distribution range (Albignac, 1972; Goodman et al., 1997; Hawkins & Racey, 2008; Gerber et al., 2012a; Farris et al., 2015). The Malagasy civet (Fossa fossana) is considered Near Threatened by the IUCN (Hawkins, 2008a). According to our results and online datasets, although its distribution range is one of the largest, it has a high dietary and activity overlap with other carnivore species (Farris et al., 2015). Moreover, this species appears to be particularly sensitive to forest disturbance, as it is absent from fragmented rainforests, does not occupy human–dominated landscapes, and is considered to be intolerant to degraded forests (Kerridge et al., 2003; Gerber et al., 2012b). Lastly, the small–toothed mongoose (Eupleres goudotii goudotii), is considered Endangered by the IUCN (Dollar, 2000), but not in our results. However, we acknowledge that it

is sensitive to habitat destruction due to deforestation, and that its narrow dietary niche overlaps with those of F. fossana and C. ferox (Albignac, 1972; Schreiber et al., 1989; Macdonald, 1992; Garbutt, 1999; Nowak, 1999; Dollar, 2000; Logan et al., 2015). One of the major concerns for Malagasy species is deforestation. Harper et al. (2007) reported that, since the 1950s, forest coverage on Madagascar had declined by 40%. Deforestation has altered carnivore assemblages, with contiguous rainforests harbouring the greatest number of native species (Gerber et al., 2012b). Despite conservation efforts, deforestation rates remain high, with its consequent negative impacts on biodiversity, soil compaction and erosion, water and carbon cycles (Erdmann, 2003; Raik, 2007). Economic interests and political lobbying, such as the state's forest concession policy (Jarosz, 1993; Klein, 2002), also contribute to the high rates of deforestation, Decisions by politicians in the 1920s led to the massive destruction of some of the most easily accessible forests on the island (Jarosz, 1993). Current knowledge on the majority of species is scarce and measures to improve this (such as ecological studies or monitoring programs) should be prioritized. For example, we found that the giant–striped mongoose G. grandidieri was most likely to be next to become extinct but current knowledge on this species

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(as for M. decemlineata, S. concolor, G. fasciata and F. fossana) is limited. This means that, although G. grandidieri was considered the most vulnerable in our analysis, there is some uncertainty associated with this outcome. Lack of knowledge on human–animal conflicts greatly inhibited our analysis, with no data available for F. fossana, G. fasciata and G. grandidieri. Additionally, information on the species’ prey status is also missing for G. elegans, so these two important factors were not taken into consideration in the final analysis for each species. It is clear that our current level of knowledge on species can influence our risk evaluations and the probability of a species going extinct. We strongly believe that if knowledge gaps are filled, better management actions can be taken to mitigate the most pressing threats. It should also be borne in mind that lack of knowledge is a threat in itself, as it is harder to preserve the unknown. In conclusion, Malagasy terrestrial mammalian carnivores are at risk of extinction. There is an immediate need to increase the size of protected areas across all Madagascar’s forested ecosystems (Gerber et al., 2012b). Conservation measures such as community– based actions are already contributing to the protection of C. ferox (Jones et al., 2008). A recent assessment across the tropics has shown thatwell–designed community–based conservation approaches, despite their possible flaws, often result in synergistic economic and ecological gains (Brooks et al., 2012). Emphasis should also be placed on the the need for conservation legislation to be strictly enforced. Finally, knowledge on each and every species is essential to correctly implement conservation measures and to support decision makers in formulating and enacting management plans. Acknowledgements We thank Professors Jorge Medina and Carlos Fonseca for their support. We also thank the IUCN Red List for data on species distributions. Co–author Eduardo Ferreira was supported by a post–doctoral grant from FCT (Program POPHQREN, ref: SFRH/ BPD/72895/2010). We would like to thank the University of Aveiro (Department of Biology) and FCT/ MEC for the financial support to CESAM RU (UID/ AMB/50017) through national funds and, where applicable, co–financed by the FEDER, within the PT2020 Partnership Agreement. References Albignac, R., 1972. The carnivora of Madagascar. Springer, Netherlands. – 1973. Mammifères Carnivores. ORSTOM/CNRS, Paris. Andriatsimietry, R., Goodman, S. M., Razafimahatratra, E., Jeglinski, J. W. E., Marquard, M. & Ganzhorn, J. U., 2009. Seasonal variation in the diet of Galidictis grandidieri Wozencraft, 1986 (Carnivora: Eupleridae) in a sub–arid zone of extreme south–western Madagascar. Journal of Zoology,

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279: 410–415. Bennett, C. E., Pastorini, J., Dollar, L. & Hahn, W. J., 2009. Phylogeography of the Malagasy ring–tailed mongoose, Galidia elegans, from mtDNA sequence analysis. Mitochondrial DNA, 20: 7–14. Britt, A., 1999. Observations on two sympatric, diurnal herpestids in the Betampona NR, eastern Madagascar. Small Carnivore Conservation, 20: 14. Britt, A. & Virkaitis, V., 2003. Brown–tailed mongoose Salanoia concolor in the Betampona Reserve, eastern Madagascar: photographs and an ecological comparison with ring–tailed mongoose Galidia elegans. Small Carnivore Conservation, 28: 1–3. Brooks, J. S., Waylen, K. A. & Mulder, M. B., 2012. How national context, project design, and local community characteristics influence success in community–based conservation projects. PNAS, 109: 21265–21270. Crooks, K. R. 2002. Relative sensitivities of mammalian carnivores to habitat fragmentation. Conserva� tion Biology, 16: 488–502. Currylow, A. F. T., 2014. Natural History Notes. Her� petological Review, 45: 116–117. Dollar, L., 2000. Eupleres goudotii. The IUCN Red List of Threatened Species. Version 2015.1. <www. iucnredlist.org> [Downloaded on 09 June 2015]. Duckworth, J. W., Hawkins, A. F. A., Randrianasolo, H., Andrianarimisa, A. & Goodman, S. M., 2014. Suggested English names for Madagascar’s species of Carnivora. Small Carnivore Conservation, 50: 54–60. Durbin, J., Funk, S. M., Hakins, F., Hill, D. M., Jenkins, P. D., Moncrieff, C. B. & Ralainasolo, F. B., 2010. Investigations into the status of a new taxon of Salanoia (Mammalia: Carnivora: Eupleridae) from the marshes of Lac Alaotra, Madagascar. System� atics and Biodiversity, 8: 341–355. Erdmann, T. K., 2003. The dilemma of reducing shifting cultivation. In: The Natural History of Madagas� car (S. M. Goodman & J. P. Benstead, Eds.). The University of Chicago Press, Chicago. Farris, Z. J., Gerber, B. D., Karpanty, S., Murphy, A., Andrianjakarivelo, V., Ratelolahy, F. & Kelly, M. J., 2015. When carnivores roam: temporal patterns and overlap among Madagascar’s native and exotic carnivores. Journal of Zoology, 296: 45–57. Farris, Z. J., Karpanty, S. M., Ratelolahy, F. & Kelly, M. J., 2014. Predator–primate distribution, activity, and co–occurrence in relation to habitat and human activity across fragmented and contiguous forests in northeastern Madagascar. International Journal of Primatology, 35: 859–880. Farris, Z. J. & Kelly, M. J., 2011. Assessing carnivore populations across the Makira Protected Area, Madagascar: WCS Pilot Camera Trapping Study. Submitted to the Wildlife Conservation Society Madagascar Program. Farris, Z. J., Kelly, M. J., Karpanty, S. M., Ratelolahy, F., Andrianjakarivelo, V. & Holmes, C., 2012. Brown– tailed vontsira Salanoia concolor (Eupleridae) documented in Makira Natural Park, Madagascar: new insights on distribution and camera–trap success. Small Carnivore Conservation, 47: 82–86.

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Supplementary material

Appendix 1s. Equations used for the estimation of dietary, strata and activity pattern overlap. Apéndice 1s. Ecuaciones utilizadas para la estimación de la dieta, los estratos y el patrón de actividad. Dietary overlap The dietary overlap of a species, relatively to other species, was estimated has the sum of the overlap (weighted by the maximum possible number of overlapping species) for each of the food items in its diet, weighted by the diet breadth (i.e. the number of different items on that species diet.

j [ n

Dietary_overlap =

1 Dietary_breadth

Overlapping_species Max_overlap

]

were dietary_breadth is the number of different food items, overlapping_species, the number of other species eating the same item; and max_overlap, the maximum number of species that could overlap (i.e. seven species) Strata overlap The strata overlap of species n was estimated as the sum of the products between species n scores (for arboreal and terrestrial strata) and all other i species scores (for arboreal and terrestrial strata).

j [(Arbn x Arbi) + (Tern x Teri)] n-1

Strata_overlapn =

were Arb is the score of arboreal strata for species n and i (0; 0.5 or 1); and Ter is the score of terrestrial strata for species n and i (0; 0.5 or 1) Activity pattern overlap The activity pattern overlap of species n was estimated as the sum of the products between species n scores (for diurnal and nocturnal activity) and all other i species scores (for diurnal and nocturnal activity). Cathameral species were coded as 50% diurnal and 50% nocturnal.

j [[(Din x Dii) + (Non x Noi)] n-1

Activity_pattern_overlapn =

were Di is the score of diurnal activity for species n and i (0; 0.5 or 1); and No is the score of nocturnal strata for species n and i (0; 0.5 or 1)

Cartagena–Matos et al.

Table 1s. List of references consulted in this study, with reference to species addressed and to the type of information available in each listed reference: D. Diet; C. Competition; M/O. Madagascar/other. Species codes: Cf. Cryptoprocta ferox; Eg. Eupleres goudotii; Ff. Fossa fossana; Ge. Galidia elegans; Gf. Galidictis fasciata; Gg. Galidictis grandidieri; Md. Mungotictis decemlineata; Sc. Salanoia concolor. Tabla 1s. Lista de las referencias consultadas en este estudio, con referencia a las especies que mencionan y el tipo de información disponible en cada una de ellas: D. Dieta; C. Competencia; M/O. Madagascar/otros. (Para los códigos de las especies, véase arriba).

References

Albignac (1972)

Gg Md X

Albignac (1973)

Ali & Huber (2010)

C M/O

Allnutt et al. (2008) Andriatsimietry et al. (2009)

Antona et al. (2002)

Antona et al. (2004)

Bennett et al. (2009) Borgerson (2013)

Britt (1999)

Britt & Virkaitis (2003)

Brooks et al. (2006)

Brooks et al. (2012)

Cincotta et al. (2000)

Corlett & Primack (2011)

Crooks (2002)

Crowley (2010)

Crutzen (2002)

Currylow (2014)

Davies et al. (2007)

Dirzo & Raven (2003) Dolch (2011)

Dollar (2000)

Duckworth et al. (2014)

Dunham (1998)

Durbin et al. (2010)

Farris & Kelly (2011)

Farris et al. (2012)

Farris et al. (2014)

Farris et al. (2015)

Garbutt (1999)

Gerber et al. (2010)

Gerber et al. (2012a)

Gerber et al. (2012b)

Gerber et al. (2012c)

Goillot (2009)

X X

Animal Biodiversity and Conservation 40.1 (2017)

iii

Table 1s. (Cont.)

References

Goodman (2003)

Goodman & Benstead (2005)

Goodman et al. (2003)

Goodman et al. (1997)

Goodman & Helgen (2010)

Goodman & Raselimanana (2003)

Gg Md

C M/O

Goodman et al. (2004)

Goodman & Pidgeon (1998)

Green & Sussman (1990)

Harper et al. (2007)

Hawkins & Dollar (2008)

Hawkins (2008a)

Hawkins (2008b)

Hawkins (2008c)

Hawkins (2008d)

Hawkins (2008e)

X X

X X X

Hawkins et al. (2008)

Hawkins et al. (2000)

Hawkins & Racey (2008)

Hawkins & Racey (2005)

Hector et al. (2001) Hunter & Caro (2008) Irwin et al. (2010)

Jarosz (1993)

Jones et al. (2008)

Jones et al. (2009)

Klein (2002) Köhncke & Leonhardt (1986)

Kremen et al. (2008)

Kremen et al. (1999)

Kremen et al. (1998)

Kull et al. (2014)

Logan et al. (2015)

Lürs & Dammhahn (2010)

Lürs & Kappeler (2013)

Macdonald (1992)

Marquard (2011)

Michalski & Peres (2005) Mitchell & Banks (2005)

Myers et al. (2000)

Cartagenaâ&#x20AC;&#x201C;Matos et al.

Table 1s. (Cont.) References

Gg Md

C M/O

Noss et al. (1996) Nowak (1999)

Paemelaere & Dobson (2011)

Rabeantoandro (1997)

Raik (2007) Ravoahangy et al. (2011)

Razafimahaimodison (2003)

X X

Razafindratsima (2014)

Scales (2012)

Schipper et al. (2008)

Schneider & Kappeler (2014)

Schnoell (2012) Schreiber et al. (1989)

Shimada (2014)

Tilman et al. (1994)

Vanak & Gompper (2010) Van Vuuren et al. (2012)

Watson et al. (2004)

Woodroffe & Ginsberg (1998)

Woolaver et al. (2006)

Yoder et al. (2003)

Totals

20 37

Animal Biodiversity and Conservation 40.1 (2017)

References listed in table 1s Referencias relacionadas en la tabla 1s.

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Wild felid species richness affected by a corridor in the Lacandona forest, Mexico M. Gil–Fernández, C. Muench, D. A. Gómez–Hoyos, A. Dueñas, S. Escobar–Lasso, G. Aguilar–Raya & E. Mendoza

Gil–Fernández, M., Muench, C., Gómez–Hoyos, D. A., Dueñas, A., Escobar–Lasso, S., Aguilar–Raya, G. & Mendoza, E., 2017. Wild felid species richness affected by a corridor in the Lacandona forest, Mexico. Animal Biodiversity and Conservation, 40.1: 115–120. Abstract Wild felid species richness affected by a corridor in the Lacandona forest, Mexico.— Wild felids are one of the most vulnerable species due to habitat loss caused by fragmentation of ecosystems. We analyzed the effect of a structural corridor, defined as a strip of vegetation connecting two habitat patches, on the richness and habitat occupancy of felids on three sites in Marqués de Comillas, Chiapas, one with two isolated forest patches, the second with a structural corridor, and the third inside the Montes Azules Biosphere Reserve. We found only two species (L. pardalis and H. yagouaroundi) in the isolated forest patches, five species in the structural corridor, and four species inside the Reserve. The corridor did not significantly affect occupancy, but due to the low detection rates, further investigation is needed to rule out differences. Our results highlight the need to manage habitat connectivity in the remaining forests in order to preserve the felid community of Marqués de Comillas, Chiapas, México. Key words: Habitat fragmentation, Connectivity, Neotropical felids, Corridor, Landscape ecology Resumen Los efectos de la presencia de un corredor en la selva Lacandona, en México, en la riqueza de especies de félidos silvestres.— Los félidos silvestres se encuentran entre las especies más vulnerables ante la pérdida de hábitat causada por la fragmentación de los ecosistemas. Se analizó el efecto de la presencia de un corredor estructural, definido como una franja de vegetación que conecta dos fragmentos de hábitat, en la riqueza y ocupación de félidos en tres sitios de Marqués de Comillas, en Chiapas: uno comprende dos fragmentos de bosque aislados, otro presenta un corredor estructural y el último se encuentra dentro de la reserva de la biosfera Montes Azules. Se encontraron cuatro especies en el interior de la Reserva, cinco en el corredor estructural y únicamente dos (L. pardalis y H. yagouaroundi) en los fragmentos de bosque aislados. La presencia del corredor no afectó de forma significativa a la ocupación, pero debido a la baja tasa de detección, se necesita seguir investigando para descartar diferencias. Nuestros resultados resaltan la necesidad de manejar la conectividad del hábitat en los bosques remanentes para lograr la conservación de la comunidad de félidos en Marqués de Comillas, en Chiapas, México. Palabras clave: Fragmentación del hábitat, Conectividad, Félidos neotropicales, Corredor, Ecología de los paisajes Received: 24 V 16; Conditional acceptance: 10 X 16; Final acceptance: 15 XI 16 Margarita Gil–Fernández, Anel Dueñas, Gustavo Aguilar–Raya & Eduardo Mendoza, Inst. de Investigaciones sobre los Recursos Naturales, Univ. Michoacana de San Nicolás de Hidalgo.– Margarita Gil–Fernández & Sergio Escobar–Lasso, Inst. Internacional en Conservación y Manejo de Vida Silvestre, Univ. Nacional, Costa Rica.– Carlos Muench, Inst. de Investigaciones en Ecosistemas y Sustentabilidad, Univ. Nacional Autónoma de México, Morelia, México.– Diego A. Gómez–Hoyos, ProCAT Internacional/The Sierra to Sea Inst., Las Alturas, Puntarenas, Costa Rica; Grupo de Estudio en Herpetología, Univ. del Quindío, Armenia, Colombia; Grupo de Investigación y Asesoría en Estadística, Red Mesoamericana y del Caribe para la Conservación de Anfibios y Reptiles (MesoHERP), Univ. del Quindío, Armenia, Colombia. Corresponding author: M. Gil–Fernández. E–mail: mgilfedz@gmail.com ISSN: 1578–665 X eISSN: 2014–928 X

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Introduction Habitat loss is a consequence of exponential growth of human populations around the world. Habitat destruction typically leads to fragmentation, the division of habitat into smaller and more isolated portions of land, separated by a matrix of human–transformed land cover (Haddad et al., 2015). Ecosystem fragmentation causes changes in landscape configuration. The implicit fragmentation also has well–documented consequences related to edge effects, such as increases in access for poaching, and dispersion of invasive species and decreases in the genetic flow of some species (Laurance & Useche, 2009). One of the most popular approaches to improve landscape connectivity is structural corridors, defined from a human perspective as habitat strips connecting habitat patches (Tellería, 2016). Structural corridors are crucial to connect populations that would otherwise be isolated, and they mitigate the effects of fragmentation (Bennett et al., 2004). Wild felids are particularly prone to the effects of fragmentation because of their intrinsic ecological traits, such as large food requirements and wide–ranging behavior (Ripple et al., 2014). They are therefore an interesting wildlife group to study, not only because of their susceptibility to such effects but also because of their key ecological roles within ecosystems (Zanin et al., 2014). This study was conducted in the Lacandona rainforest, where five species of wild felids occur: jaguar (Panthera onca), cougar (Puma concolor), ocelot (Leopardus pardalis), margay (L. wiedii) and jaguarundi (Herpailurus yagouaroundi) (Garmendia et al., 2013). The populations of these five species are currently declining right throughout their distribution range (IUCN, 2016). The aim of this research was to analyze the effect of a structural corridor on the richness and occupancy of wild felids in a tropical rainforest area. Material and methods The study was carried out at three sites (fig. 1) located in the Lacandona rainforest, one of the most biodiverse areas in Mexico (Lira–Torres et al., 2012). Two of the sites were located within the municipality of Marqués de Comillas, which supports 15.7% of the total area of the Lacandona rainforest (204,000 ha). This is a highly fragmented area, where agriculture and livestock activities occupy 52% of the territory (Bezaury–Creel & Gutíerrez–Carbonell, 2009). Habitat fragmentation has created spatially heterogeneous landscape patterns. Marqués de Comillas is divided into common lands called 'ejidos', and this study was developed at four ejidos: Reforma Agraria, Adolfo Lopez Mateos, Zamora Pico de Oro and La Corona (fig. 1). The third site was located at the southern part of the Montes Azules Biosphere Reserve (MABR), which protects 331,200.00 ha of the Lacandona rainforest. Contrary to Marqués de Comillas, the reserve has a continuum of tropical rainforest that has been protected since 1978 (Ortiz–Espejel & Toledo, 1998).

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We carried out a camera–trap survey in the Lacandona forest to record wild felid presence. We selected two sites with different landscape configuration, focusing on the presence–absence of a structural vegetation corridor. The first site consisted of two forest parches isolated by an anthropogenic matrix, without corridor connections (16° 22' 18''–16° 20' 33'' N, 90° 42' 22''–90° 40' 32'' W). Three landscape elements were identified in this site: (1) a forest patch in La Corona (17.27 km2), (2) a matrix in between, and (3) a forest patch in Zamora Pico de Oro (17.84 km2) (figs. 1A, 2A). There is a 1 km linear distance between patches. The mean distance between camera–trap stations and population settlements was 2.31 km (± 1.04). The camera–trap stations were on average 6.2 km (± 0.94) from the MABR. The second site is located within the ejidos of Reforma Agraria and Adolfo Lopez Mateos (16° 15' 95''– 16° 13' 41'' N y 90° 49' 25''–90° 48' 45'' W). It consists of two forest patches connected by a structural corridor. Here, we defined four landscape elements: (1) a forest patch of Reforma Agraria (23.85 km2), (2) a structural corridor (defined as corridor in Muench, 2012), (3) a matrix surrounding the corridor, and (4) a forest patch of Adolfo Lopez Mateos (75.77 km2) (figs. 1B, 2B). There was a minimum distance of 2.14 km between patches and the mean distance to population settlements was 2.86 km (± 1.08). The camera–trap stations were placed an average of 3.8 km (± 1.34) from the MABR. We set four camera–traps on each of the seven landscape elements. The cameras were located according to a systematic arrangement, spaced 1 km from each other (fig. 2). Only the corridor cameras were spaced 500 m from each other, due to spatial limitations. Both study sites in Marqués de Comillas (figs. 2A, 2B) were sampled from 28 I 2012 to 15 XII 2012. The minimum distance between both camera–trap arrays was 16.2 km, reducing the likelihood to capture the same individual felids in the different camera–trap arrays. The third site was used as a control. It was located inside the MABR Biosphere Reserve Montes Azules, where 15 camera–traps were located, in the circuits Miranda, La Granja and Sabana II, located to the south of the reserve. The distance between cameras was 500 m (fig. 2C). These camera–traps were active from 30 VIII 2013 to 20 V 2014. The average distance to population settlements was 2.36 km (± 0.58). This site is at least 15.6 km far from the other camera–trap systems, and there is also a fast–flowing river between the reserve and Marqués de Comillas, so we can consider this site independent from the other two. Owing to equipment restrictions, we used four different models of camera–traps (Bushnell trophy cam, Primos truth cam 35, Wildview extreme, and Stealth Cam STC–U838NXT). Nevertheless, each site had the same proportion of each model to standardize the sampling between sites. The sampling efforts at each site were: 842 camera–trap days for the isolated patches site, 1,664 camera–trap days for the corridor site, and 1,016 camera–trap days for the reserve.

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Fig. 1. Study area, Marqués de Comillas, Chiapas, and its location in Mexico. The study sites where the camera–traps were located are shown in the lower panel: A. Isolated ejidos Zamora Pico de Oro and La Corona, site of the two isolated forest patches; B. Connected, ejidos Reforma Agraria and Adolfo Lopez Mateos, site of forest patches connected by a structural corridor (in white); C. Reserve, southern part of the Montes Azules Biosphere Reserve, site of the continuum rainforest. Fig. 1. Zona de estudio, Marqués de Comillas, en Chiapas, y su ubicación dentro de México. En el panel inferior se muestran los sitios donde se colocaron las cámaras trampa: A. Aislado, ejidos Zamora Pico de Oro y La Corona, sitio de los dos fragmentos de bosque aislados; B. Conectado, ejidos Reforma Agraria y Adolfo López Mateos, sitio de los fragmentos de bosque conectados mediante un corredor estructural (en blanco); C. Reserva, parte meridional de la reserva de la biosfera Montes Azules, sitio con bosque húmedo continuo.

All of the picture records were considered sufficiently clear to avoid false positives. To be able to compare species richness at each site, regardless of the camera–trap effort, we used rarefied species accumulation curves and extrapolations (Gotelli & Coldwell, 2001; Magurran, 2004). The history of detection of felid species at each camera–trap was used to estimate occupancy at each site. To calculate these parameters we used the single–season occupancy models of software PRESENCE 11.5 (MacKenzie et al., 2002). For these models we assumed occupancy (Ψ) was alternately constant or dependent on the presence–absence of the structural corridor and presence of the reserve. Likewise, detection probability (p) was alternately considered constant, varible over time, or dependent on the presence of the structural corridor or reserve. We ran a total of 16 models for each felid species. The best model was chosen based on the Akaike Information Criterion (AIC), where the smaller values indicate a higher likelihood (Burnham & Anderson, 2002).

Results We recorded four species in the reserve site; L. pardalis, L. wiedii, P. concolor, and P. onca (fig. 1s in supplementary material). We recorded all five species in the connected site (fig. 3, fig. 2s in supplementary material). This latter site included the landscape elements with the most felid independent captures, the patch of Reforma Agraria and the corridor, each with seven records. There were no records of felids in the matrix surrounding the corridor. In the isolated site we recorded only two species, L. pardalis and H. yagouaroundi (fig. 3s in supplementary material). These records were confined to La Corona patch; no sightings were obtained in the matrix or in the Pico de Oro patch. The model with the lowest AIC was the null model, in which the occupancy rate and probability of detection were constant. The model with the second lowest AIC value had constant occupancy and probability of presence dependent on presence–absence of the

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corridor; nevertheless, the fitting was insufficient to accept this model (∆AIC > 3). There was a tendency to a higher occupancy rate for L. pardalis than for other species Ψ = 0.27 (95% CI: 0.10–0.56; fig. 4). Overall, there was high uncertainty associated with occupancy estimations ranging from 0.29 (95% CI: 0.04–0.78) for P. onca, to 0.59 (95% CI: 0.09–0.91) for P. concolor. Discussion

Structural corridor

Fig. 2. Camera–trap locations. Study sites: A. Isolated, two isolated forest patches; B. Connected, two forest patches connected by a structural corridor (in white dots); C. Reserve, sampling along the paths inside the Montes Azules Biosphere Reserve, Chiapas, Mexico. Four camera–traps were located on each landscape element (corridor, matrix, patch) of the sites outside the reserve. Fig. 2. Ubicación de las cámaras trampa. Sitios de estudio: A. Aislado, dos fragmentos de bosque aislados; B. Conectado, dos fragmentos de bosque conectados por un corredor estructural (en puntos blancos); C. Reserva, muestreo en senderos en el interior de la reserva de la biosfera Montes Azules, en Chiapas, México. Se colocaron cuatro cámaras trampa por cada elemento del paisaje (corredor, matriz y fragmento) en los sitios fuera de la reserva.

Our results suggest that loss of structural connectivity has a marked effect on wild felids species richness in our study site, as observed in other sites (Haddad et al., 2015). We should also emphasize that the only two species found in the isolated patches site, L. pardalis and H. yagouaroundi, are the most generalist of the five species in the study area. In contrast, it has been reported that presence of jaguar and cougar is favored in connected habitat (Grigione et al., 2009). The margay is considered the most vulnerable felid in this group, which could also explain its absence on the isolated patches site (Payan et al., 2008). We did not record H. yagouaroundi inside the reserve, possibly in view of its low density, which has been previously reported (Towns et al., 2013). When comparing occupancy between sites, there was acceptance of the null model. We attribute this result to the low detection rates, which might produce false negatives, specially for the isolated patches site. The low detection rates could also be attributed to the high mobility and low density of these felids (Dillon & Kelly, 2007; Silver et al., 2004). Other factors may affect our results. First, the area of patches is contrasting. In fact, the Adolfo Lopez Mateos patch is three times larger than the other patches, which could account for the higher richness of felids in the connected site (Scheiner, 2003). Still, in the connected site, there were more species and more records of felids in the Reforma Agraria patch, which is only slightly larger than the patches of the isolated site. Other factors that could also play a role are the distance to the reserve, and the distance to roads and population settlements, which we considered to have negligible effects on the probability of capture. It should also be noted that the camera–trap array was different in the reserve site; cameras were placed along the trails, possibly increasing the captures of felids, which are common users of trails (Harmsen et al., 2010). However, the detection rate did not vary between sites, so we could assume that the array did not affect our results. The low variation in detectability also suggests adequate distribution of the camera–trap models, which could otherwise be a source of error (Kelly & Holub, 2008). To conclude, our results indicate that habitat connectivity is an important landscape feature, which requires to be maintained or even increased to favor conservation of felid species richness in the Lacandon rainforest. If habitat connectivity continues to decrease due to deforestation, the species that could be the most affected, in the short term, might be P. concolor, P. onca and L. wiedii.

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Fig. 3. Rarefied species accumulation curves of wild felids for the study sites located in Marqués de Comillas (connected and isolated) and south of the Montes Azules Biosphere Reserve (reserve), Chiapas, Mexico. Fig. 3. Curvas de acumulación de especies escasas de félidos silvestres para los sitios del estudio en Marqués de Comillas (conectado y aislado) y al sur de la reserva de la biosfera Montes Azules (reserva), en Chiapas, México.

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Fig. 4. Estimated occupancy of felid species in the Lacandona forest, Chiapas, México. Ponc. Panthera onca; Pcon. Puma concolor; Lpar. Leopardus pardalis; Lwie. Leopardus wiedii; and Hyag. Herpailurus yagouaroundi. Fig. 4. Ocupación estimada de las especies de félidos de la selva Lacandona, en Chiapas, México: Ponc. Panthera onca; Pcon. Puma concolor; Lpar. Leopardus pardalis; Lwie. Leopardus wiedii; Hyag. Herpailurus yagouaroundi.

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Acknowledgements We would like to thank the Corredor Biológico Mesoamericano for financial support, and the communities who provided shelter and accepted the use of camera–traps on their territory. We also thank Angela A. Camargo–Sanabria and J. Manuel Mora for their valuable comments. Finally, we would like to thank Idea Wild, whose equipment donated to M. Gil–Fernández was used in the writing and analyses of this study. References Bennett, A. F., 2004. Enlazando el paisaje: el papel de los corredores y la conectividad en la conservación de la vida silvestre. IUCN–the World Conservation Union, San José, Costa Rica. Bezaury–Creel, J. & Gutiérrez–Carbonell, D., 2009. Áreas naturales protegidas y desarrollo social en México. In: Capital Natural de México, Vol. II: Estado de conservación y tendencias de cambio: 385–431 (R. Dirzo, R. González & I. J. March, Eds.). CONABIO, Mexico City. Burnham, K. P. & Anderson, D. R., 2002. Model Selection and Multimodel Inference: A Practical Information–Theoretic Approach, 2nd ed. Springer– Verlag, New York. Dillon, A. & Kelly, M. J., 2007. Ocelot Leopardus pardalis in Belize: the impact of trap spacing and distance moved on density estimates. Oryx, 41: 469–477. Garmendia, A., Arroyo–Rodríguez, V., Estrada, A., Naranjo, E. J. & Stoner, K. E., 2013. Landscape and patch attributes impacting medium and large–sized terrestrial mammals in a fragmented rain forest. Journal of Tropical Ecology, 29: 331–344. Gotelli, N. J. & Colwell, R. K., 2001. Quantifying biodiversity: procedures and pitfalls in the measurement and comparisons of species richness. Ecological Letters, 4: 379–391. Grigione, M. M., Menke, K., López–González, C., List, R., Banda, A., Carrera, J., Carrera, R., Giordano, A. J., Morrison, J., Sternberg, M., Thomas, R. & Van Pelt, B., 2009. Identifying potential conservation areas for felids in the USA and Mexico: integrating reliable knowledge across an international border. Oryx, 43: 78–86. Haddad, N. M., Brudvig, L. A., Clobert, J., Davies, K. F., Gonzalez, A., Holt, R. D., Lovejoy, T. E., Sexton, J. O., Austin, M. P., Collins, C. D., Cook, W. M., Damschen, E. I., Ewers, R. M., Foster, B. L., Jenkins, C. N., King, A. J., Laurance, W. F., Levey, D. J., Margules, C. R., Melbourne, B. A., Nicholls, A. O., Orrock, J. L., Song, D. & Townshend, J. R., 2015. Habitat fragmentation and its lasting impact on Earth’s ecosystems. Science Advances, 1: e1500052. Harmsen, B. J., Foster, R. J., Silver, S., Ostro, L. & Doncaster, C. P., 2010. Differential Use of Trails by Forest Mammals and the Implications for Camera– Trap Studies: A Case Study from Belize. Biotropica,

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42: 126–133. IUCN, 2016. Red List of Threatened Species. Version 2015–4. In: www.iucnredlist.org [Accessed on 14 May 2016]. Kelly, M. J. & Holub, E. L., 2008. Camera Trapping of Carnivores: Trap success among camera types and across species, and habitat selection by species, on Salt Pond Mountain, Giles County, Virginia. Northeastern Naturalist, 15: 249–262. Laurance, W. F. & Useche, D. C., 2009. Environmental synergisms and extinctions of tropical species. Conservation Biology, 23: 1427–1437. Lira–Torres, I., Galindo–Leal, C. & Briones–Salas, M., 2012. Mamíferos de la Selva Zoque, México: riqueza, uso y conservación. Revista de Biología Tropical, 60: 781–797. MacKenzie, D. I., Nichols, J. D., Lachman, G. B., Droege, S., Royle, A. & Langtimm, C. A., 2002. Estimating site occupancy rates when detection probabilities are less than one. Ecology, 83: 2248–2255. Magurran, A. E., 2004. Measuring Biological Biodiversity. Blackwell Publishing. Australia. Muench–Spitzer, C. E., 2012. Mapa: Áreas estratégicas para el mantenimiento de la conectividad al sureste de la selva Lacandona, Chiapas, México. Url: http://www.conabio.gob.mx/informacion/gis/ [Accessed on 14 October 2016]. Ortiz–Espejel, B. & Toledo, V. M., 1998. Tendencias en la deforestación de la Selva Lacandona (Chiapas, México): El caso de Las Cañadas. Interciencia, 23: 318–327. Payan, E., Eizirik, E., de Oliveira, T., Leite–Pitman, R., Kelly, M. & Valderrama, C., 2008. Leopardus wiedii. UICN Lista roja de especies amenazadas. Version 2011.2. Url: www.iucnredlist.org [Accessed 12 November 2014]. Ripple, W. J., Estes, J., Beschta, R. L., Wilmers, C. C., Ritchie, E. G., Hebblewhite, M., Berger, J., Elmhagen, B., Letnic, M., Nelson, M. P., Schmitz, O. J., Smith, D. W., Wallach, A. D. & Wirsing, A. J., 2014. Status and ecological effects of the world’s largest carnivores. Science, 343(1241484): 151–162. Scheiner, S. M., 2003. Six types of species–area curves. Global Ecology and Biogeography, 12: 441–447. Silver, S. C., Ostro, L. E. T., Marsh, L. K., Maffei, L., Noss, A. J., Kelly, M. J., Wallace, R. B., Gómez, H. & Ayala, G., 2004. The use of camera traps for estimating jaguar Panthera onca abundance and density using capture/recapture analysis. Oryx, 38: 1–7. Tellería, J. L., 2016. Wildlife Habitat Requirements: Concepts and Research Approaches. Current Trends in Wildlife Research, 1: 79–95. Towns, V., León, R., de la Maza, J. & Sánchez– Cordero, V., 2013. Aportaciones al listado de los mamíferos carnívoros del sur de la Reserva de la Biosfera Montes Azules, Chiapas. Therya, 4: 627–640. Zanin, M., Palomares, F. & Brito, D., 2014. What we (don’t) know about the effects of habitat loss and fragmentation on felids. Oryx, 49.01: 96–106.

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Supplementary material

Fig. 1s. Felid species recorded on the Montes Azules Biosphere Reserve, Chiapas, Mexico: A. Puma concolor; B. Panthera onca; C. Leopardus wiedii; and D. Leopardus pardalis. Fig. 1s. Especies de félidos registradas en la reserva de la biosfera Montes Azules, en Chiapas, Mexico: A. Puma concolor; B. Panthera onca; C. Leopardus wiedii; D. Leopardus pardalis.

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Fig. 2s. Felid species recorded at the connected site of Marqués de Comillas, Chiapas, Mexico: A. Puma concolor; B. Leopardus wiedii; C. Leopardus pardalis; D. Panthera onca; and E. Herpailurus yagouaroundi. Fig. 2s. Especies de félidos registradas en el sitio conectado en Marqués de Comillas, en Chiapas, México: A. Puma concolor; B. Leopardus wiedii; C. Leopardus pardalis; D. Panthera onca; E. Herpailurus yagouaroundi.

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Fig. 3s. Felid species recorded at the isolated site of Marqués de Comillas, Chiapas, Mexico: A. Herpailurus yagouaroundi; B. Leopardus pardalis. Fig. 3s. Especies de félidos registradas en el sitio aislado en Marqués de Comillas, en Chiapas, México: A. Herpailurus yagouaroundi; B. Leopardus pardalis.

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Peasant farmer–raptor conflicts around Chembe Bird Sanctuary, Zambia, Central Africa: poultry predation, ethno–biology, land use practices and conservation V. R. Nyirenda, F. Musonda, S. Kambole & S. Tembo Nyirenda, V. R., Musonda, F., Kambole, S. & Tembo, S., 2017. Peasant farmer–raptor conflicts around Chembe Bird Sanctuary, Zambia, Central Africa: poultry predation, ethno–biology, land use practices and conservation. Animal Biodiversity and Conservation, 40.1: 121–132. Abstract Peasant farmer–raptor conflicts around Chembe Bird Sanctuary, Zambia, Central Africa: poultry predation, ethno– biology, land use practices and conservation.— Raptors provide ecosystem services to African rural communities by: (1) preying on rodents, (2) regulating harmful snake populations, (3) shaping cultural beliefs, and (4) being part of tourist attractions. Peasant farmers, however, connect them with poultry depletion, telepathic omens, and traditional witchcraft. Consequently, raptors suffer human–induced persecution. Using a qualitative content analysis technique, we investigated the interaction between farmers and raptors in areas adjoining the Chembe Bird Sanctuary. Our results unravel negative perceptions, attitudes and practices that could threaten the extinction of five raptors in the study area. We propose the use of transformative cognitive measures (e.g., raising stakeholder awareness, ensuring stringent law enforcement for raptors and protecting their habitat, and strengthening relational social capital) and physical measures (e.g., providing appropriate fencing and poultry breeding of high resilient phenotypes) to improve the co–existence between farmers and raptors. Key words: Social–ecological system, Stakeholder participation, Technical ecological knowledge, On–farm counter–measures, Ecosystem services Resumen Los conflictos entre campesinos y rapaces alrededor del refugio de Chembe Bird, en Zambia, África central: depredación de aves de corral, etnobiología, prácticas de uso de la tierra y conservación.— Las rapaces prestan servicios ecosistémicos a las comunidades rurales de África: (1) depredando roedores, (2) regulando las poblaciones de serpientes dañinas, (3) configurando las creencias culturales y (4) formando parte de las atracciones turísticas. Sin embargo, los campesinos las relacionan con la disminución de las aves de corral, las profecías telepáticas y la brujería tradicional. En consecuencia, las rapaces son perseguidas por los humanos. Mediante una técnica de análisis cualitativo de contenido, analizamos la interacción entre los campesinos y las rapaces en zonas adyacentes al refugio de Chembe Bird. Nuestros resultados revelan las prácticas, actitudes y percepciones negativas que podrían poner en peligro de extinción a cinco rapaces en la zona de estudio. A fin de mejorar la coexistencia entre agricultores y rapaces, proponemos utilizar medidas transformadoras de carácter conceptual (por ejemplo, sensibilizar a las partes interesadas, garantizar el cumplimiento riguroso de la legislación relativa a las rapaces y proteger su hábitat, así como reforzar el capital social relacional) y medidas prácticas (como proporcionar cercados apropiados y aves de corral de fenotipos de alta resistencia). Palabras clave: Sistema socioecológico, Participación de partes interesadas, Conocimiento técnico–ecológico, Contramedidas en las explotaciones agrícolas, Servicios ecosistémicos Received: 23 I 16; Conditional acceptance: 10 VI 16; Final acceptance: 18 XI 16 Vincent R Nyirenda, Frederick Musonda & Saviour Kambole, Dept. of Zoology and Aquatic Sciences, School of Natural Resources, Copperbelt Univ., Jambo Drive, Riverside, P. O. Box 21692, Kitwe, Zambia.– Sydney Tembo, Conservation and Management Section, Dept. of National Parks and Wildlife, P. O. Box 260240, Kalulushi, Zambia. Corresponding author: V. R. Nyirenda. E–mail: vrnyirenda@hotmail.com, vincent.nyirenda@cbu.ac.zm ISSN: 1578–665 X eISSN: 2014–928 X

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Introduction Human–wildlife conflicts commonly arise when humans encroach into previously uninhabited areas and disrupt animal habitat (Lamarque et al., 2009; Treves et al., 2009) through competition for space and food resources (Balmford et al., 2001; Okello, 2005). Increasing populations of humans at the human–wildlife interfaces such as park boundaries exacerbate such conflict (Wittemyer et al., 2008). Social and human dimensions such as cultural values, beliefs, and experiences with wildlife contribute to negative local perceptions and attitudes against wildlife (Dickman, 2010). Due to their carnivorous feeding habits, raptors are usually resented for their opportunistic predation on poultry (Fowler et al., 2009) even though some raptors have beneficial social–ecological roles. For instance, owls help farmers control rodents that forage crops and stored food (Magige & Senzota, 2006; Kopij et al., 2014). The greater the poultry losses, the less local communities support wildlife conservation (Gadd, 2005). Such losses prompt farmers to use lethal, illegal control methods such as retaliatory killings of the raptors considered 'pests' (Etheridge et al., 1997; Lamarque et al., 2009). These persecutions translate to species extirpation as raptor breeding rates and density decrease within large geographical ranges (Graham et al., 2005; Thirgood et al., 2005; Peterson et al., 2010). Such decreases may have perceived and actual economic impacts associated with the incurred losses (Sarasola et al., 2010; Margalida et al., 2014). From a theoretical perspective, human–raptor conflicts are influenced by prevailing ecological traps in human–dominated landscapes. Ecological traps, which are attractive habitats of choice, yet poor quality habitats to fauna for their survival and population growth, are a management concern as they may result in species decline or local extinction (Battin, 2004). According to Schlaepfer et al. (2002), organisms can become ecologically trapped by their evolutionary responses to environmental cues by making decisions regarding behavioural and life–history habitat selection, thus decreasing survival or reproduction. According to Kokko & Sutherland (2001), human–modified landscapes may become ecological traps if species, using indirect cues, miscalculate the habitat suitability when making preferences. Consequently, raptors succumb to several environmental stressors such as destruction of their habitats through poisoning (Kramer & Redig, 1997; Henny & Elliott, 2007), electrocution on power–lines (Harness, 2007), disease (Rodríguez et al., 2010), trapping (Duncan et al., 2002; Romulo et al., 2009), road kills, and land conversions (Meunier et al., 2000; Jaarsma et al., 2006). These anthropogenic disturbances extend towards protected areas such as national parks and wildlife reserves (Gill & Sutherland, 2000). However, the ecological impact on specific groupings of birds is still poorly understood. Resolving human–raptor conflicts would contribute positively to human–raptor coexistence (Gibson et al., 2000; Riley et al., 2003; Fernandez–Juricic et al., 2004) and foster benefits from ecological services such as cultural and tourism development for local communities.

Dealing with peasant farmer–raptor conflicts requires the analysis of long–term empirical data on an array of socio–economic aspects concerning those affected and also on the ecological aspects of species considered pests. However, empirical data are rarely available, particularly in developing countries where research is limited due to technical and financial constraints and the conservation status of numerous species in such a setting is usually unknown (Costello et al., 2013). Therefore, the use of traditional ecological knowledge (TEK) can be helpful to develop and implement conservation plans when scientific information is scarce (Freeman, 1992; Usher, 2000). TEK refers to 'common' knowledge and resulting perceptions shaping attitudes and practices, locally passed on through generations (Berkes, 2004). In this study, we explored TEK, resultant perceptions, attitudes and practices by peasant farmers to determine anthropogenic threats to raptor conservation in an important bird area. The study addressed two broad questions: (1) what are the perceived factorial conditions for occurrence of poultry losses caused by steppe buzzards (Buteo vulpinus) and black kites (Milvus migrans) and (2) what is the perceived underlying proximate nature of peasant farmer–raptor conflicts in terms of poultry predation, ethno–biology (i.e. cultural treatment and usage of raptors by peasant farmers) and land use practices. We hypothesised that entrenched TEK shaped perceptions, attitudes and practices among peasant farmers, who either negated or supported avian raptor conservation. Steppe buzzards, black kites and owls are protected under Zambia wildlife legislation, which implements the Convention on International Trade in Endangered Species of Wild Fauna and Flora (CITES) regulations. All these birds are CITES Appendix II species and, as such, are protected internationally. Unlike steppe buzzards and black kites, none of the four owls (Barn owls, Tyto alba; spotted eagle owls, Bubo africanus; pearl spotted owls, Cilaucidium perlatum and giant eagle owls, Bubo lacteus) inhabiting the study area predate on poultry, yet they are persecuted due to local myths associated with them. Material and methods Study area The study was conducted within a 20 km radius area around Chembe Bird Sanctuary (539 ha in dimension; with central coordinates of 27.9955º E, 12.8302º S) in Kalulushi District, Zambia (fig. 1). Chembe Bird Sanctuary is renowned for its high diversity of resident and migratory birds, attracting numerous visitors and tourists. Raptors such as steppe buzzards, black kites, barn owls, spotted eagle owls, pearl spotted owls and giant eagle owls use the park and proximate areas for physiological purposes such as foraging, breeding and roosting. However, peasant farmers, originating from diverse Bantu ethnic and cultural groups occupy much of the original raptor habitat around Chembe Bird Sanctuary. Peasant farmers rear 'village' free–ranging

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Country boundary Main roads National Park Forest reserve Game management area

Fig. 1. Location of the study area in the peripherals of Chembe Bird Sanctuary, Zambia. Fig. 1. Localización geográfica de la zona de estudio en la periferia del refugio de Chembe Bird, en Zambia.

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chickens (Gallus gallus domesticus), crested guinea fowls (Guttera pucherani), helmeted guinea fowls (Numida meleagris), and common domestic ducks (Anas platyrhynchos domesticus). Other livestock reared are goats (Capra aegagrus hircus), sheep (Ovis aries) and cattle (Bos taurus).

topology of issues that culminated into underlying local perceptions, attitudes and practices associated with raptorial species.

Data collection protocols

Factorial conditions for occurrence of human–raptor conflicts: the multi–factorial theory

Data were collected from 15 IX 2015 to 23 XI 2015, coinciding with the peak breeding activities of the raptors. This period was the off–peak farming season, guaranteeing successful interviews with 138 peasant farmers, the survey participants. We used a semi– structured questionnaire (supplementary material) and adopted the purposive sampling approach described by Coetzee et al. (2014), but integrated within the interviews with only refereed, knowledgeable adult informants (> 35 years old, either males or females) with antecedents. We used the 'peer reference system' in the multi–stage sampling method as stipulated by Emery & Purser (1996), where the participant selection was iterative and based on knowledge, area of residence, longevity of residence, and involvement with poultry rearing. The purposive sampling strategy enabled the collection of the informants' own detailed viewpoints, serving as a social learning platform and enabling development of threads of grounded concepts as postulated by Strauss & Corbin (1998) and Muro & Jeffrey (2008). The participants, both males and females, were inductively identified and interviewed. The questions covered perceived TEK, attitudes, and practices, allowing us to gain an in–depth and broad understanding of the nature of peasant farmer–raptor conflicts and to obtain practical insights for the development of management strategies. The objective, contents, and applications of the research were explained to respondents before obtaining their individual, free and informed consent. The six potential participants that declined to participate on the grounds of suspicions that the researchers could be devil agents, working through raptors, were excused from further questioning. Vernacular language (i.e. Bemba) was used in place of English, the official language. Respondents were asked to prioritise and rank their responses based on their knowledge and experiences. The responses were probed until information given was clear and exhaustive. Participants were shown coloured pictures of prevalent raptors in the region to help identification. A WS–853 Olympus Voice Recorder device was used to capture participants’ views on local perceptions, attitudes and practices. These recordings were later transcribed for analyses. Interviews were discontinued once responses were complete and nonce no further new information was provided. Data analyses We used the content analysis approach to analyse the qualitative data as described by Coetzee et al. (2014). This approach allowed categorising and condensing of similar thematic issues, from which we derived the

Results

Peasant farmers perceived four factorial conditions relating to peasant farmer–raptor conflicts in areas adjacent to the Chembe Bird Sanctuary: (1) Motivation: most peasant farmers (94.20%, n = 130) indicated that the raptors were more likely to predate on poultry when energetic demands were highest, particularly when providing for newly hatched chicks, as the predation peaked during their breeding season in September–October. The raptors tended to prey on seemingly weak and highly nutritious small, free–range chickens of less than eight weeks that were easier to lift and more tender than older chickens; (2) Anti–internal inhibition: many peasant farmers (81.88%, n = 113) perceived that restraint measures would reduce prey frequency on poultry. They contended that the raptors would instead prey on substitutes such as small wild birds, snakes and snails; (3) Anti–external inhibition: most peasant farmers (97.83%, n = 135) perceived that raptors, however, find ways of circumventing interventions by humans against poultry predation, such as trappings and fences. All survey participants indicated that tall trees and vegetated anthills within reach of the raptors (tens of meters) on the premises facilitated predation by steppe buzzards. However, they perceived that the presence of people and pet domestic dogs (Canis lupus) at home combined with shouting, use of metals for noise, and stoning were deterrents to the marauding steppe buzzards and black kites in more than 75% of predation events (fig. 2). Furthermore, preponderance of peasant farmers (96.38%, n = 133) perceived that bare and cleared surroundings of chicken–holding panes increased exposure of chickens to raptors, amounting to a loss o of 20–45% more chickens than where tree cover and other escape features were available. However, all the respondents contended that raptors are a great challenge to poultry, as expressed in the following viewpoints: ''Despite adaptive changes in the chicken behaviour and implementation of multiple counter–measures, raptors remain a major threat to chickens' survival and cause considerable economic loss to us.'' Tina Mwamba, female, 63 years old. (4) Anti–victim resistance factor: majority of peasant farmers (85.51%, n = 118) perceived that the raptors must overcome chickens’ defence behaviour from motherly palliative care such as taking swift cover upon detecting danger. Raptor predation seasonality and their perceived underlying causes The preponderance of respondents (95.65%, n = 132) found most chicken predation by raptors occurred

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Fig. 2. Use of domestic dogs as deterrents against raptor attacks on free–ranging chickens in the peripherals of the Chembe Bird Sanctuary, Zambia. Fig. 2. Utilización de perros domésticos para disuadir a las rapaces de atacar pollos criados en libertad en la periferia del refugio de Chembe Bird, en Zambia.

in the hot dry season, coinciding with the breeding season for the raptors, though it may occur at any time of year. All respondents had had chicken losses, especially birds less than four months old. Though raptor–chicken predation occurred mainly in the mornings and evenings, it could take place any time

of the day (fig. 3). Many peasant farmers (92.03%, n = 127), however, perceived that this poultry loss was due to the type of counter–measures used and phenotype–based selective harvesting by size and colour. Furthermore, a large number of respondents (73.19%, n = 101) considered that distances from

Number of respondents

60 52

50 40

30 22

20 12

10 0

E M & E Time of chicken hunts

All

Fig. 3. Diurnal predation times of chickens by raptors in the peripherals of the Chembe Bird Sanctuary, Zambia: M. Morning; A. Afternoon; E. Evening; All. All the time. Fig. 3. Depredación de pollos por rapaces según el momento del día en la periferia del refugio de Chembe Bird, en Zambia: M. Mañana; A. Tarde; E. Anochecer; All. Todo el día.

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Mean number of chickens

27.31

25 20

17.65

15 10 5 0

Chickens reared

Chickens predated

Fig. 4. Chickens predated by raptors from the total reared chickens per month estimated by peasant farmers in areas surrounding Chembe Bird Sanctuary, Zambia. Fig. 4. Pollos depredados por rapaces del total de pollos criados por mes, según las estimaciones de los campesinos en zonas adyacentes al refugio de Chembe Bird, en Zambia.

Chembe Bird Sanctuary, the core raptor habitat, to their homesteads was inconsequential as raptors covered considerably large territorial areas. Life history traits Most respondents (81.16%, n = 112) had witnessed steppe buzzards predating small–medium sized chickens. Many respondents (72.46%, n = 100) had seen black kites predating chicks only. Raptors predated many free–ranging chickens (64.44%) before they were old enough for domestic consumption, for sale or as gifts to friends and relatives (fig. 4). Most respondents (94.20%, n = 130) reported a greater loss (> 90%) to raptors of white chickens than brown or black chickens that were camouflaged by the colour of the ground. Costs of raptor–poultry predation Raptor predation amounted to a loss for peasant farmers of K25.73 (USD 1.84) and K53.89 (USD 3.85) per chicken had they reached adult size. A single farmer would lose an average of 17.65 ± 1.35 chickens monthly to raptors (fig. 4). Peasant farmers (100%, n = 138) did not report any of the poultry predation incidents to authorities such as traditional leaders and wildlife agency, but they may have shared information socially among peers, the young and the elderly, relatives, and neighbours. Over half of the respondents (65.22%,

n = 90) used preventive or counter–measures such as fencing, provision of cover and housing, and traditional measures (e.g., shouting, stoning, metal banging and domestic dogs), whereas the rest allowed the chickens to roam around premises unattended during the day. Providing housing for chickens was prohibitively expensive for peasant farmers, especially if materials such as fencing wire had to be sourced from elsewhere. In addition, peasant farmers (87.68%, n = 121) incurred opportunity costs by forgoing other chores to keep vigil over their poultry. More than half of the 90 respondents (57.78%, n = 52) who acknowledged using preventive and counter–measures to mitigate raptor loss considered wire fencing was more effective than other methods to manage raptor–chicken predation on their premises (figs. 5, 6). Furthermore, most respondents (97.10%, n = 134) recognised that raptor shooting was illegal but that it was occasionally applied as a retaliatory counter–measure. However, the respondents (94.93%, n = 131) indicated that a combination of preventive and counter–measures was more effective than a single method. They further contended that the effectiveness of a given method also depended on how well the local farmers implemented the method. Traditional ecological knowledge and attitudes associated with steppe buzzards and black kites The peasant farmers (70.29%, n = 97) emphasized that they would occasionally find chicken carcasses on their premises and would blame raptors based on evidence from claw marks left on them, especially on the necks, or the heads having been cut off. They further indicated that the decision to throw away or eat the chicken remains left by raptors depended on the condition of the carcass; decaying carcasses and cases of suspected poisoning were discarded, while the size of the chicken and the damaged part were not considered. The peasant farmers had no direct anthropogenic uses for steppe buzzards and black kites. Locally, the mere presence of steppe buzzards and black kites signified emergent poultry depletion and perpetual poverty among the impoverished peasant farmers. Not surprisingly, 36.23% (n = 50) expressed resentment and considered these raptors should be eradicated due to considerable damage they caused. On the contrary, respondents (63.77%, n = 88) that supported raptor conservation based their response on conservation awareness and a religious belief of stewardship. Those in support of conservation perceived that enclosures and keeping raptors off their premises could reduce the economic loss resulting from raptor predation of their chickens. Traditional ecological knowledge of owls All the peasant farmers (100%, n = 138) who suffered poultry losses from steppe buzzards and black kites also regularly encountered four species of owls: barn owls, spotted eagle owls, pearl spotted owls and giant eagle owls. While their encountering circumstances varied across time and individuals, the owl

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Number of respondents

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Moderately effective

50 40

127

Lowly effective Not effective

30 20

17 8

10 0

2 1

9 3 2

IHr

1 2 1 1

Sc Pr Counter–measures

1 1

1 1 1 1

Fig. 5. Rating by respondents of effectiveness of counter–measures used in the peripherals of Chembe Bird Sanctuary, Zambia against chicken predation by raptors: Wf. Wire fencing; IHr. In–housing rearing; Sc. Stick caging: Pr. Prolonged release in mornings; Tm. Traditional methods; Rs. Raptor shooting. Fig. 5. Clasificación de la eficacia de las contramedidas empleadas en la periferia del refugio de Chembe Bird, en Zambia, contra la depredación de pollos por parte de rapaces, según los encuestados: Wf. Cercas de alambre; IHr. Cría en lugares cerrados; Cs. En jaula: Pr. Liberación prolongada por las mañanas; Tm. Métodos tradicionales; Rs. Disparo a las rapaces.

raptors were commonly encountered by respondents (77.61%, n = 104) perched on trees in proximity to their homesteads, within farmlands. However, most

of the respondents (92.75%, n = 128) perceived that the primary habitats for owls were the nearby thick forests and secondary habitats were the anthills within

Fig. 6. Wire fencing successfully implemented at one of the farms in the peripherals of the Chembe Bird Sanctuary, Zambia, where small chicks are confined to the fowl run and larger birds are free ranging within the enclosure. Fig. 6. Cercado de alambre instalado en una de las explotaciones situadas en la periferia del refugio de Chembe Bird, en Zambia, donde los polluelos quedan confinados en el corral y las aves más grandes viven en libertad dentro del cercado.

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their farmlands. They also saw them on the ground, on rooftops, along the roads, and in proximity with the water bodies like dams or rivers. Though peasant farmers generally perceived habitat losses and low productivity rates as threats to raptors, a number of the respondents (63.77%, n = 88) perceived that lethal counter–measures such as killing by use of shotguns, catapults and trappings against owls were the greatest conservation threats to raptors. Furthermore, respondents (97.10%, n = 134) reported that harvesting and destruction of raptor eggs as acts of aggression were common. Besides, 72.46% (n = 100) perceived short fallow periods (less than three years) which do not support tree recruitment, and that opening up of new farming areas, and charcoal production accounted for much of the habitat destruction for owls and other raptors. Many respondents (98.55%, n = 136) attributed the perceived reduction in owls to anthropogenic activities such as habitat conversions. Despite the antagonistic tendencies against raptors, all the respondents acknowledged the important ecological role owls play in regulating rodents which positively contributes to their food security. All the respondents (100%, n = 138) stated no they had no knowledge of owls attacking their poultry. Traditional uses of owls According to respondents (100%, n = 138), the use of owl parts such as feathers, feet and faces was secretive among local community members. They highlighted that owl parts were used to fortify persons from witchcraft attacks, and to protect crops, food and other property from magic. Respondents (86.21%, n = 100) held entrenched myths that encountering owls was a taboo, the debut of the demise of a family member. Local perception, attitudes and practices related to owls Perceptions were mixed as to what respondents felt they should do: whether they should take part in conservation of the owls or have owls removed by means such as killing them or translocation. While over half of the respondents (57.97%, n = 80) indicated willingness to participate in avian conservation, a considerable number of respondents (42.03%, n = 52) declared that they would rather have owls removed. The various local perceptions have led to strong intolerance and attitudes towards owls as expressed here: ''Owls are enemies to humankind and should be chased away or better still killed on encounter at all cost.'' Veronica Kabinda, female, 75 years old. ''We fear owls for the bad omens and so they should not be allowed near people....'' Florence Lufino, female, 46 years old. However, those (57.97%, n = 80) in support of owl conservation proffered their support based on belief that owls were God’s creation, worthy of good stewards’ care. Such a positive perception is reflected in the following quote: ''Owls are God’s creation, and humanity has the responsibility to preserve them for their own sake.'' Dawson Chinyama, male, 57 years old.

Most respondents (94.20%, n = 130) revered the owl hunters and shared information with the traditional leaders about the killing of owls by owl hunters but rarely reported the incidents to the wildlife agency for punitive consequences for infractions. Most respondents (95.65%, n = 132) confirmed they were not culturally attached to owl conservation, despite the benefits owls provide, such as their warning of imminent death, and their impact on food protection by preying on rodents. All respondents (100%, n = 138) indicated that there are local unpublished 'by–laws' governing owl conservation and prohibiting wanton destruction of wildlife in general, but that these are not specific and lack sanctions. If they find an owl carcass, most respondents (89.86%, n = 124) either leave it, or throw it away. If they find a live owl, however, most (97.83%, n = 135) chase it and often even kill it on the belief that owls are a bad omen and ugly. However, most respondents (88.40%, n = 122) envisioned that conservation measures such as sensitisation of local communities coupled with law enforcement by the wildlife agency increase benefits to local communities and that their involvement in conservation strategies could be critical for owl conservation. Discussion Poultry predation, ethno–biology and land use practices Traditional ecological knowledge (TEK) about raptor– driven poultry predation, ethno–biology and land use practices is generated and shared laterally among the clan and friends as well as in a hierarchical manner with affiliates and collaborators across local communities and other stakeholders. TEK of predator behaviour, and the determination of blood stains and claw marks on dead or injured chickens is based on social memory and social learning by societal members (Muro & Jeffrey, 2008) and prevails from challenges about a social–ecological landscape in social fabrics (Wenger, 1998). Consequently, based on the various constructs of realities by the local communities, social learning influences stakeholders' participation in natural resource management by shaping their identity and attitude (Glasser, 2010). The retention and dissemination of formal and informal channels of information can thus play a vital role in natural resource management (Berkes, 2004). For instance, peasant farmers know that steppe buzzards (being relatively large and conspicuous) often mimick poultry, mixing with free–ranging chickens and foraging together with them before attacking them. They also know black kites may predate in pairs, approaching the target chicken one behind the other to maximise the chances of catching their prey. Such innovative mechanisms clearly increase their effectiveness in predation. Relying on swiftness and strong sharp talons, the raptors likely harvest a chick in any predation event. Local knowledge also tells that peak predation coincides with the breeding sea-

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son in September–October, when steppe buzzards and black kites hatch their young and supplement their food with chicken protein. It is unlikely that peasant farmers yet realize that environmental stressors such as climate change could prompt raptors to increase poultry depletion. Several studies have already highlighted the possibility that climate change influences extinction of fauna and flora (Myers et al., 2000; Sinervo et al., 2010), and may consequently negatively impact on raptors. More local micro–data collection and analysis may be critical to determine the quality of extension services to peasant farmers on raptor ecology. In response to the actual and perceived human– raptor conflicts, peasant farmers have implemented several preventive and counter–measures, predominantly traditional methods. In addition, some farmers have adopted non–lethal methods, such as training dogs to protect their poultry. If well trained, dogs can effectively be used for poultry protection from predators (Gehring et al., 2010). There is a need to adopt a combination of effective methods to lessen the prevailing and underlying apprehensions between peasant farmers and wildlife agency staff due to wildlife based conflicts. According to Marshall et al. (2007) and Hill (2015), inter–human relationships are equally important in human–wildlife conflict resolution and management because of epitomised differences in management objectives among parties. As poultry has multiple uses even a few minor events of losses can be devastating to impoverished peasant farmers. As a result, some peasant farmers have resorted to clearing vegetation from their properties to prevent raptors from resting and perching before they attack their poultry, reinforcing their cultural belief that having trees within their yards signifies low sanitation. On the contrary, clearing vegetation around the premises increases the risk of chicken loss to raptors because limited cover enhances the visibility of chickens to predatory birds. Some peasant farmers have employed illegal lethal methods such as retaliatory shooting of raptors with shotguns and catapults. In some localities, education and Christian beliefs may help support conservation as local people have an entrenched understanding about wildlife stewardship, in contrast with areas where negative traditional beliefs and low conservation appreciation overshadow conservation efforts. Furthermore, the prohibitive costs of establishing and maintaining wire fencing counteracted the willingness by peasant farmers to adopt novel preventive interventions. Fencing material and labour cost about USD 15–USD 20 per meter, an expense that is out of reach of many peasant farmers. The prolonged dry season (May–November) probably puts the poultry at greater risk than in the rainy season, when raptor predation events decrease considerably. Preventive and counter–measures that take seasonality into consideration may help reduce human–raptor conflicts. Given that poultry predation by raptors takes place at any time of the day, the most highly recommend measures, such as wire fencing, that are effective day long are those most recommended in resolving of human–raptor conflicts.

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White phenotypic traits in chickens risk extirpation if not protected and only brown, black and other dull earthly coloured chickens —which express adaptation to the environment of the chickens— would persist in accordance with natural selection postulated by Sinclair et al. (2006). However, further detailed ecological empirical data are needed to understand the evolutional mechanisms that increase raptors’ ability to predate their prey in human–dominated, poor quality sink habitats while more innovative preventive counter–measures are being explored. Implications for local participation in avian conservation Local peasant farmers build their knowledge base about the environment around them, through long– term social learning, which is critical for avian conservation (Araya et al., 2009; Novotny et al., 2012). Though TEK has inherent limitations in some cases, such as a potentially significant degree of inaccuracy, it influences and shapes the local status of conservation (Becker & Ghimire, 2003; McGregor, 2004). Based on local knowledge, peasant farmers whose poultry has been preyed on develop negative perceptions and attitudes towards wildlife conservation and may retaliate by killing raptors and destroying their habitat, as described by Treves et al. (2009). Irrational emotions by peasant farmers and an often exaggerated magnitude of economic losses often account for illegal and destructive counter–measures such as animal poisoning, shooting and trappings (Nyirenda et al., 2013). Innovative techniques to identify illegal activities can be critical in generating relevant information for policy making (Cross et al., 2013), given that the magnitude of illegal killings of raptors is usually unknown. Local participation in conservation can be increased if conservationists and researchers integrate the understanding of ecological traps into conservation planning (Battin, 2004). For instance, in addition to land conversion and forest fragmentation, the inappropriate and indiscriminate use of fires for anthropogenic activities (such as traditional hunting of small animals for local consumption (Eriksen, 2007) and the use of unsustainable harvesting methods (such as cutting of trees to harvest edible caterpillars (Mbata et al., 2002) reduce habitat quality for raptors. Such unsustainable anthropogenic activities may affect physiological aspects of raptors by reducing forage and nesting resources. Furthermore, phenotype–based selective poultry predation may have evolutionary implications for poultry in the long term in favour of non–white chickens. Use of appropriate and exclusionary countermeasures such as fencing would minimise poultry predation and consequently, improve avian conservation. Effective interventions would negate the intolerance that would otherwise develop among the affected people through such means as poisoning, shooting and trappings (Treves et al., 2009). The lethal effects of such methods have great repercussions for target and even non–target species due to their non–selective nature (Berny et al., 1997; Anderson, 2000; Brakes & Smith, 2005).

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The development of effective interventions such as establishing raptor–proof fowl runs to address human–raptor conflicts therefore requires systematic participatory approaches. The responsible implementation of a combination of effective deterrents is recommended, such as: (1) well maintained and managed fence enclosures to keep raptors away from poultry; (2) healthily managed poultry which evade predation; (3) use of more aggressive guard dogs such as Anatolian shepherd guard dogs; and (4) consistent presence of people, especially during the peak raptor attack periods. Integration of multi–level analyses of empirical data, with the participation of multi–stakeholders, can be critical for human–raptor conflict resolution and management (Canavelli et al., 2014). And lastly, emerging tourism based on diverse avian populations is likely to stimulate economic benefits and local support for raptor conservation (Snyman, 2012). Acknowledgements We thank The Copperbelt University and Zambia Wildlife Authority (now the Department of National Parks and Wildlife) for authorizing and facilitating this research. Numerous respondents volunteered their time and information, making this research possible. References Anderson, M. D., 2000. Raptor conservation in the Northern Cape province, South Africa. Ostrich, 71: 25–32. Araya, Y. N., Schmiedel, U. & von Witt, C., 2009. Linking ‘citizen scientists’ to professionals in ecological research: examples from Namibia and South Africa. Conservation Evidence, 6: 11–17. Balmford, A., Moore, J. L., Brooks, T., Burgess, N., Hansen, L. A., Williams, P. & Rahbek, C., 2001. Conservation conflicts across Africa. Science, 291: 2616–2619. Battin, J., 2004. When good animals love bad habitats: ecological traps and the conservation of animal populations. Conservation Biology, 18(6): 1482–1491. Becker, C. D. & Ghimire, K., 2003. Synergy between traditional ecological knowledge and conservation science supports forest preservation in Ecuador. Conservation Ecology, 8(1): 1. http://www.consecol. org/vol8/iss1/art1/ Accessed 23 IX 15. Berkes, F., 2004. Rethinking community based conservation. Conservation Biology, 18: 621–630. Berny, P. J., Buronfosse, T., Buronfosse, F., Lamarque, F. & Lorgue, G., 1997. Field evidence of secondary poisoning of foxes (Vulpes vulpes) and buzzards (Buteo buteo) by Bromadiolone: a 4–year survey. Chemosphere, 35: 1817–1829. Brakes, C. R. & Smith, R. H., 2005. Exposure of non– target small mammals to rodenticides: short–term effects, recovery and implications for secondary poisoning. Journal of Applied Ecology, 42: 118–128. Canavelli, S. B., Branch, L. C., Cavallero, P., Gon-

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Supplementary material

Questionnaire with serial numbers per individual interviewed (Ethics letter detailing the study presented to the individual participants, followed by a brief explanation and consent from individual participants before conducting of interviews). 1. What is your name?........................................................................................................... (optional) 2. How old are you? ____ years 3. Gender? [noted] a. Male [ ]

b. Female [ ]

4. What is your marital status? a. Married [ ] c. Single [ ] b. Divorced [ ] d. Widowed [ ] 5. What is the highest level of education you attained? a. Primary [ ] c. Junior secondary [ ] b. Senior secondary [ ] d. College/university [ ] 6. What is your religion? (specify, if any) ............................................................................................... 7. What ethnic grouping do you belong to? specify .............................................................................. A. Conditions for occurrence of humanâ&#x20AC;&#x201C;raptor conflicts 1. What conditions would you attribute to the occurrence of peasant farmerâ&#x20AC;&#x201C;raptor conflicts in the peripherals of Chembe Bird Sanctuary? B. Diurnal times and seasonality 2. In which season is poultry predation most prevalent? a. Rainy season b. Cold dry season c. Hot dry season 3. Why do you think poultry predation is most prevalent in that season? 4. At what age in their growth were the chickens most vulnerable to raptors? 5. During which diurnal time was chicken predation most common? a. Morning b. Afternoon c. Evening d. Morning and evening e. All the time 6. Explain why the predation was most prevalent during the diurnal time you have indicated in Q5 above 7. Do you consider the distance from Chembe Bird Sanctuary to influence the quantum of chicken losses to raptors?

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C. Life history traits 8. What size of chicken is predated most by steppe buzzards? a. Small (< 8 weeks) b. Medium (8–12 weeks) c. Large (> 12 weeks) 9. What size of chicken is predated most by black kites? a. Small (< 8 weeks) b. Medium (8–12 weeks) c. Large (> 12 weeks) 10. State on average how many chickens are usually reared per month? 11. How many of those chickens reared are usually predated per month? 12. Indicate the colour of chickens lost to raptors? 13. Can you explain relationship that exists between colour and the chicken losses, if any? D. Costs of raptor–poultry predation 14. How much does each chicken cost at the point of disposal? 15. Do you report the poultry losses to any of the authorities? 16. If yes to Q15, where do you report to? 17. What actions are taken if reported? 18. What counter–measures do you deploy against raptors that prey on poultry? 19. How effective are the counter–measures? a. Effective (i.e. eliminate >75% of the risks) b. Moderate (i.e. eliminate 50–75% of the risks) c. Not effective (i.e. fail to eliminate half of the risks) 20. Do you incur any opportunity costs relating to protection of your poultry? a. Yes b. No c. Not sure d. Do not know at all 21. If yes to Q20, explain 22. Explain the mechanism of reducing the cost of losing poultry to raptors? E. Traditional ecological knowledge and attitudes associated with steppe buzzards and black kites 23. What evidence do you seek to indicate that raptors are responsible for the loss of the poultry at your farm? 24. What do you do to the dead chickens left behind by the raptors? 25. What determines your actions? 26. What are the uses for raptors that predate your chickens? 27. What is the significance of human co–existence with the raptors predating the chickens?

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28. Do you support conservation of the raptors? 29. If yes to Q28, explain F. Traditional ecological knowledge over owls 30. State which owls you encounter in your area? 31. What are the owls doing at the time of encounter? 32. In what habitats are the owls encountered? 33. State the major threats to owls in your area? 34. Do you view that the mentioned threats (in Q33) have negative impacts on owls? explain 35. Explain the ecological roles of owls, if any? 36. Is there any known owl that attack poultry in your area? a. Yes b. No G. Traditional uses of owls 37. What is the nature of use of owl parts? 38. How are parts of owls used, if any? 39. What do owls signify in your life? H. Local perception, attitudes and practices related to owls 40. Must owls be conserved or not? explain 41. Were any incidences of owl killings by local people reported to authorities? 42. How are your contemporary actions towards owls? 43. Are there local ‘by–laws’ for the protection of owls? 44. If yes to Q43, where do you report to? 45. What do you do to owls found dead? 46. How do you respond when you encounter owls? 47. What do you suggest should be done to ensure owls are conserved?

Thank you very much!

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Animal Biodiversity and Conservation

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

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ISSN: 1578–665X eISSN: 2014–928X

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III

Animal Biodiversity and Conservation

Manuscritos

Animal Biodiversity and Conservation es una revista interdisciplinar, publicada desde 1958 por el Museo Ciencias Naturales de Barcelona. Incluye artículos de investigación empírica y teórica en todas las áreas de la zoología (sistemática, taxonomía, morfología, biogeografía, ecología, etología, fisiología y 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. 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

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.

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, Nombre, Abstract, Key words, Resumen, Palabras clave y Dirección postal conformarán la primera página.)

Introducción. En ella se dará una idea de los ante� cedentes del tema tratado, así como de los objetivos del trabajo. Material y métodos. Incluirá la información referente a las especies estudiadas, aparatos utilizados, me� todología de estudio y análisis de los datos y zona de estudio. Resultados. En esta sección se presentarán úni� camente los datos obtenidos que no hayan sido publicados previamente. Discusión. Se discutirán los resultados y se compara� rán con otros trabajos relacionados. Las sugerencias sobre investigaciones futuras se podrán incluir al final de este apartado. Agradecimientos (optativo). Referencias. Cada trabajo irá acompañado de una bibliografía que incluirá únicamente las publicaciones citadas en el texto. Las referencias deben presentarse según los modelos siguientes (método Harvard): * Artículos de revista: Conroy, M. J. & Noon, B. R., 1996. Mapping of spe� cies richness for conservation of biological diversity: conceptual and methodological issues. Ecological Applications, 6: 763–773. * Libros y otras publicaciones no periódicas: Seber, G. A. F., 1982. The estimation of animal abundance. C. Griffin & Company, London. * Trabajos de contribución en libros: Macdonald, D. W. & Johnson, D. P., 2001. Dispersal in theory and practice: consequences for conserva� tion biology. In: Dispersal: 358–372 (T. J. Clober, E. Danchin, A. A. Dhondt & J. D. Nichols, Eds.). Oxford University Press, Oxford. * Tesis doctorales: Merilä, J., 1996. Genetic and quantitative trait vari� ation in natural bird populations. Tesis doctoral, Uppsala University. * Los trabajos en prensa sólo se citarán si han sido aceptados para su publicación: Ripoll, M. (in press). The relevance of population studies to conservation biology: a review. Animal Biodiversity and Conservation. Las referencias se ordenarán alfabéticamente por autores, cronológicamente para un mismo autor y con las letras a, b, c,... para los trabajos de un mismo autor y año. En el texto las referencias bibliográficas se indicarán en la forma usual: "...

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

Animal Biodiversity and Conservation 40.1 (2017)

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 untranslat� able neologisms must be in italics. Quotations in whatever language used must be typed in ordinary print between quotation marks. The name of the author following a taxon should also be written in lower case letters. When referring to a species for the first time in the text, both common and scientific names should be given when possible. Do not capitalize common names of species unless they are proper nouns (e.g. Iberian rock lizard). Place names may appear either in their original form or in the language of the manuscript, but care should be taken to use the same criteria throughout the text. Numbers one to nine should be written in full within the text except when preceding a measure. Higher numbers should be written in numerals except at the beginning of a sentence. Specify dates as follows: 28 VI 99 (for a single day); 28, 30 VI 99 (referring to two days, e.g. 28th and 30th), 28–30 VI 99 (for more than two consecu� tive days, e.g. 28th to 30th). Footnotes should not be used.

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

ISSN: 1578–665X eISSN: 2014–928X

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

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

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

Animal Biodiversity and Conservation 40.1 (2017)

VII

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

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Morelle et al.

103–114 Cartagena–Matos, B., Gregório, I., Morais, M. & Ferreira, E. Trends in the extinction of carnivores in Madagascar 115–120 Gil–Fernández, M., Muench, C., Gómez–Hoyos, D. A., Dueñas, A., Escobar–Lasso, S., Aguilar–Raya, G. & Mendoza, E. Wild felid species richness affected by a corridor in the Lacandona forest, Mexico

121–132 Nyirenda, V. R., Musonda, F., Kambole, S. & Tembo, S. Peasant farmer–raptor conflicts around Chembe Bird Sanctuary, Zambia, Central Africa: poultry predation, ethno–biology, land use practices and conservation

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 40.1 (2017) ISSN 1578–665 X eISSN 2014–928 X 1–6 Ferrero–García, J. J. Hunting passerines with non–selective trapping methods was a source of conflict in Spain as far back as 1933 7–16 García–Quintas, A. & Parada Isada, A. Underlying factors promoting nestedness of bird assemblages in cays of the Jardines de la Reina archipelago, Cuba 17–25 Wojciechowska, M., Nowak, Z., Gurgul, A., Olech, W., Drobik, W. & Szmatoła, T. Panel of informative SNP markers for two genetic lines of European bison: Lowland and Lowland–Caucasian 27–40 Benítez, M., Romero, D., Chirosa, M. & Real, R. Eco–geographical characterization of aquatic microhabitats used by amphibians in the Mediterranean Basin 41–47 Fernández, L., Sanabria, E. A. & Quiroga, L. B Description of Silvinichthys pedernalensis n. sp. (Teleostei, Siluriformes) from the Andean Cordillera of southern South America

49–62 Mori, E., Grandi, G., Menchetti, M., Tella, J. L., Jackson, H. A., Reino, L., van Kleunen, A., Figueira, R. & Ancillotto, L. Worldwide distribution of non–native Amazon parrots and temporal trends of their global trade 63–69 Pradhan, A. K., Shrotriya, S., Rout, S. D. & Dash, P. K. Nesting and feeding habits of the Indian giant squirrel (Ratufa indica) in Karlapat wildlife sanctuary, India 71–86 Cuyckens, G. A. E., Perovic, P. G. & Herrán, M. Living on the edge: regional distribution and retracting range of the jaguar (Panthera onca) 87–97 Cordero–Rivera, A., Martínez Álvarez, A. & Álvarez, M. Eucalypt plantations reduce the diversity of macroinvertebrates in small forested streams 99–102 Kaya, Ş. & Saglam, H. Feeding habits of garfish, Belone belone euxini Günther, 1866 in autumn and winter in Turkey’s south–east coast of the Black Sea

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