Animal Biodiversity and Conservation issue 36.1 (2013) by Museu Ciències Naturals de Barcelona MCNB

2013

and

Animal Biodiversity Conservation 36.1

Dibuix de la coberta: Salamandra salamandra, salamandra comuna, salamandra común, fire salamander (Jordi Domènech) Editor en Cap / Editor Responsable / Editor–in–Chief Joan Carles Senar Editors Temàtics / Editores Temáticos / Thematic Editors Ecologia / Ecología / Ecology: Mario Díaz (Asociación Española de Ecología Terrestre – AEET) Comportament / Comportamiento / Behaviour: Adolfo Cordero (Sociedad Española de Etología – SEE) Biologia Evolutiva / Biología Evolutiva / Evolutionary Biology: Santiago Merino (Sociedad Española de Biología Evolutiva – SESBE) Editors / Editores / Editors Pere Abelló Inst. de Ciències del Mar CMIMA–CSIC, Barcelona, Spain Javier Alba–Tercedor Univ. de Granada, Granada, Spain Russell Alpizar–Jara Univ. of Évora, Portugal Xavier Bellés Centre d' Investigació i Desenvolupament–CSIC, Barcelona, Spain Luís Mª Carrascal Museo Nacional de Ciencias Naturales–CSIC, Madrid, Spain Michael J. Conroy Univ. of Georgia, Athens, USA Adolfo Cordero Univ. de Vigo, Vigo, Spain Mario Díaz Univ. de Castilla–La Mancha, Toledo, Spain José A. Donazar Estación Biológica de Doñana–CSIC, Sevilla, Spain Jordi Figuerola Estación Biológica de Doñana–CSIC, Sevilla, Spain Gary D. Grossman Univ. of Georgia, Athens, USA Damià Jaume IMEDEA–CSIC, Univ. de les Illes Balears, Spain Jordi Lleonart Inst. de Ciències del Mar CMIMA–CSIC, Barcelona, Spain Jorge M. Lobo Museo Nacional de Ciencias Naturales–CSIC, Madrid, Spain Pablo J. López–González Univ. de Sevilla, Sevilla, Spain Santiago Merino Museo Nacional de Ciencias Naturales–CSIC, Madrid Juan J. Negro Estación Biológica de Doñana–CSIC, Sevilla, Spain Vicente M. Ortuño Univ. de Alcalá de Henares, Alcalá de Henares, Spain Miquel Palmer IMEDEA–CSIC, Univ. de les Illes Balears, Spain Javier Perez–Barberia The Macaulay Institute, Scotland, United Kingdom Oscar Ramírez Inst. de Biologia Evolutiva UPF–CSIC, Barcelona, Spain Montserrat Ramón Inst. de Ciències del Mar CMIMA–CSIC, Barcelona, Spain Ignacio Ribera Inst. de Biología Evolutiva CSIC–UPF, Barcelona, Spain Pedro Rincón Museo Nacional de Ciencias Naturales–CSIC, Madrid, Spain Alfredo Salvador Museo Nacional de Ciencias Naturales–CSIC, Madrid, Spain José L. Tellería Univ. Complutense de Madrid, Madrid, Spain Francesc Uribe Museu de Ciències Naturals de Barcelona, Barcelona, Spain Carles Vilà Estación Biológica de Doñana–CSIC, Sevilla, Spain Rafael Villafuerte Inst. de Investigación en Recursos Cinegéticos (IREC–CSIC–UCLM–JCCM), Ciudad Real, 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 Carolyne Newey Pilar Nuñez

Animal Biodiversity and Conservation 36.1, 2013 © 2013 Museu de Ciències Naturals de Barcelona, Consorci format per l'Ajuntament de Barcelona i la Generalitat de Catalunya Autoedició: Montserrat Ferrer Fotomecànica i impressió: ISSN paper: 1578–665 X ISSN digital: 2014–928 X Dipòsit legal: B. 5357–2013 Animal Biodiversity and Conservation es publica amb el suport de: l'Asociación Española de Ecología Terrestre, la Sociedad Española de Etología i la Sociedad Española de Biología Evolutiva The journal is freely available online at: www.abc.museucienciesjournals.cat

and

Animal Biodiversity Conservation 36.1

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

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

Arizaga et al.

Animal Biodiversity and Conservation 36.1 (2013)

Editorial / Editorial / From the Editor

L’any 2001, el Museu de Ciències Naturals de Barcelona va remodelar la seva publicació Miscel·lània Zoològica i la va convertir en Animal Biodiversity and Conservation. Un dels objectius d’aquell canvi va ser aconseguir l’índex d’impacte que caracteritza les revistes més competitives. Uns anys després, el 2012, ja l’hem assolit. Durant aquest temps hem publicat 291 articles i hem estat citats 1.626 vegades en revistes d’impacte. Dos dels articles publicats han arribat a acumular més de 100 cites. Ara, volem fer el pas següent, que és augmentar el nostre índex d’impacte fins a valors equiparables als d’altres revistes internacionals de prestigi. És per això que hem dotat la revista d’un equip d’editors de vàlua internacional, que amb el temps s’anirà ampliant, i hem buscat el suport de les societats científiques espanyoles més prestigioses. Tres societats han acceptat ja aquest repte: la Societat Espanyola d’Etologia, l’Associació Espanyola d’Ecologia Terrestre i la Societat Espanyola de Biologia Evolutiva, les quals han designat uns editors temàtics que ens ajudaran a potenciar cadascuna de les seves disciplines respectives. Això també vol dir que obrim les temàtiques de publicació de la nostra revista, de manera que prenem l’expressió animal biodiversity, que li dóna títol, en el sentit més ampli. Un altre front obert per potenciar Animal Biodiversity and Conservation ha estat la millora de la pàgina web que ha de servir de repositori dels diferents treballs publicats. Amb aquesta nova pàgina volem oferir un millor servei als lectors, però també continuar apostant per l’accés obert a la revista. Un open access en el seu veritable sentit, que ofereix la publicació als científics de forma gratuïta tant per als lectors com per als autors. Si volem que la ciència i el coneixement que aquesta genera —produït tot sovint amb diner públic— siguin patrimoni de tothom, no solament hem de permetre que tothom la pugui llegir sinó que també hem de facilitar que tothom hi pugui publicar, independentment del seu poder econòmic. Són moments difícils per a la ciència, especialment al nostre país. Esperem que iniciatives com la nostra, que aixopluguen l’esforç de diverses societats científiques i dels investigadors, i que obren el coneixement a tothom, hi puguin ajudar. Joan Carles Senar Editor en cap d’Animal Biodiversity and Conservation

ISSN: 1578–665X

Animal Biodiversity and Conservation 36.1 (2013)

En el año 2001, el Museo de Ciencias Naturales de Barcelona remodeló su publicación Miscel·lània Zoològica convirtiéndola en Animal Biodiversity and Conservation. Uno de los objetivos de aquel cambio fue alcanzar el índice de impacto que caracteriza a las revistas más competitivas. Unos años después, en 2012, ya lo hemos logrado. Durante este tiempo hemos publicado 291 artículos y hemos sido citados 1.626 veces en revistas de impacto. Dos de los artículos publicados han llegado a acumular más de 100 citas. Ahora queremos dar el paso siguiente, que es aumentar nuestro índice de impacto hasta valores equiparables a los de otras revistas internacionales de prestigio. Por ello hemos dotado a la revista de un equipo de editores de gran valía internacional, que con el tiempo se irá ampliando, y hemos buscado el apoyo de las sociedades científicas españolas más prestigiosas. Tres de ellas han aceptado ya este reto, la Sociedad Española de Etología, la Asociación Española de Ecología Terrestre y la Sociedad Española de Biología Evolutiva, que han designado a unos editores temáticos que nos ayudarán a potenciar cada una de sus respectivas disciplinas. Esto también significa que abrimos las temáticas de publicación de nuestra revista, por lo que tomamos la expresión animal biodiversity, que le da título, en su sentido más amplio. Otro frente abierto para potenciar Animal Biodiversity and Conservation ha sido la mejora de la página web que ha de servir de repositorio de los diferentes trabajos publicados. Con esta nueva página queremos ofrecer un mejor servicio a los lectores, pero también seguir apostando por el acceso abierto a nuestra revista. Un open access en su verdadero sentido, que ofrece la publicación a los científicos de forma gratuita tanto para los lectores como para los autores. Si queremos que la ciencia y el conocimiento que esta genera —producido muy a menudo con dinero público— sean patrimonio de todos, no sólo debemos permitir que todo el mundo pueda leerla sino que también hemos de facilitar que todo el mundo pueda publicar en ella, con independencia de su poder económico. Son momentos difíciles para la ciencia, especialmente en nuestro país. Esperamos que iniciativas como la nuestra, que acogen el esfuerzo de diferentes sociedades científicas y de los investigadores, y que abren el conocimiento a todo el mundo, puedan contribuir a ello. Joan Carles Senar Editor jefe de Animal Biodiversity and Conservation

In 2001, the Natural History Museum of Barcelona remodelled its journal Miscel·lània Zoològica as Animal Biodiversity and Conservation. One of the aims of this change was to obtain an impact factor, the index that characterises the international ranking of journals. This milestone was reached some years later, in 2012, having published 291 articles that have been cited in impact journals 1,626 times. Two of these papers have been cited more than 100 times. We are now ready to take the next step forward. Our new challenge is to increase our impact factor so that Animal Biodiversity and Conservation is comparable with prestigious international journals in the field. To meet this objective we have assembled a team of editors with international recognition, and we will continue to expand this team over time. We have sought support from the most prestigious Spanish scientific societies and three have accepted the task: the Spanish Society of Ethology, the Spanish Association of Terrestrial Ecology, and the Spanish Society of Evolutionary Biology. These societies have named thematic editors who will help us strengthen their respective disciplines. This means that we will also expand the range of fields covered in the journal in future, interpreting the expression animal biodiversity in the journal´s title in its widest sense. Another front that we have opened to boost Animal Biodiversity and Conservation is the new webpage, remodelled for use as a medium to post our published papers. With this move we not only hope to improve our services to readers but also look towards open access, open access in its broadest conception, offering the journal to scientists, both readers and authors, free of charge. If we want science and the knowledge it generates —produced with public money— to belong to us all, we not only have to enable everyone to read it but we also have to help everyone publish, independently of their economic state. These are difficult times for science, particularly in Spain. We hope that initiatives such as ours will accelerate the efforts of scientific societies and researchers to open knowledge to everyone. Joan Carles Senar Editor–in–Chief of Animal Biodiversity and Conservation

ISSN: 1578–665X

Animal Biodiversity and Conservation 36.1 (2013)

The impact of an invasive exotic bush on the stopover ecology of migrant passerines J. Arizaga, E. Unamuno, O. Clarabuch & A. Azkona

Arizaga, J., Unamuno, E., Clarabuch, O. & Azkona, A., 2013. The impact of an invasive exotic bush on the stopover ecology of migrant passerines. Animal Biodiversity and Conservation, 36.1: 1–11. Abstract The impact of an invasive exotic bush on the stopover ecology of migrant passerines.— Migration is highly energy–demanding and birds often need to accumulate large fuel loads during this period. However, original habitat at stopover sites could be affected by invasive exotic plants outcompeting native vegetation. The impact of exotic plants on the stopover behavior of migrant bird species is poorly understood. As a general hypothesis, it can be supposed that habitat change due to the presence of exotic plants will affect migrants, having a negative impact on bird abundance, on avian community assemblage, and/or on fuel deposition rate. To test these predictions, we used data obtained in August 2011 at a ringing station in a coastal wetland in northern Iberia which contained both unaltered reedbeds (Phragmites spp.) and areas where the reedbeds had been largely replaced by the invasive saltbush (Baccharis halimifolia). Passerines associated with reedbeds during the migration period were used as model species, with a particular focus on sedge warblers (Acrocephalus schoenobaenus). The saltbush promoted a noticeable change on bird assemblage, which became enriched by species typical of woodland habitats. Sedge warblers departed with a higher fuel load, showed a higher fuel deposition rate, and stayed for longer in the control zone than in the invaded zone. Invasive plants, such as saltbush, can impose radical changes on habitat, having a direct effect on the stopover strategies of migrants. The substitution of reedbeds by saltbushes in several coastal marshes in Atlantic Europe should be regarded as a problem with potential negative consequences for the conservation of migrant bird species associated with this habitat. Key words: Acrocephalus spp., Biological conservation, Biological invasion, Coastal marshes, Fuel deposition rate, Saltbush (Baccharis halimifolia). Resumen Impacto de un arbusto exótico invasor en la ecología de los puntos de parada de los paseriformes migradores.— La migración requiere un elevado gasto de energía y las aves suelen necesitar acumular grandes cantidades de grasa durante este período. Sin embargo, el hábitat original de los puntos de parada podría verse afectado por plantas exóticas invasoras que compiten con la vegetación autóctona. Se conocen poco los efectos de las plantas exóticas en el comportamiento de las especies de aves migradoras en cuanto a los puntos de parada. Como hipótesis general, puede suponerse que el cambio del hábitat debido a la presencia de plantas exóticas afectará a las aves migradoras e influirá negativamente en su abundancia, la composición de la comunidad de aves y el índice de deposición de grasa. Para comprobar estas predicciones, utilizamos los datos obtenidos en agosto de 2011 en una estación de anillamiento situada en los humedales costeros del norte de la península ibérica en los que había carrizos inalterados (Phragmites spp.) y en zonas en las que los carrizos habían sido sustituidos en gran parte por el bácaris invasor (Baccharis halimifolia). Se utilizaron como modelo a los paseriformes asociados a los carrizales durante el período de migración y se prestó especial atención al carricerín común (Acrocephalus schoenobaenus). El bácaris propició un cambio notable en la composición avícola, que se enriqueció con especies típicas de hábitats forestales. Los carricerines partieron con una cantidad de grasa superior, mostraron un índice de deposición de grasa más elevado y permanecieron más tiempo en la zona de control que en la zona invadida. Las plantas invasoras, como el bácaris, pueden forzar cambios radicales en el hábitat y tener un efecto directo en las estrategias de parada de las aves migradoras. La sustitución de los carrizales por bácaris en diversas marismas de la costa atlántica de Europa debería considerarse un problema con posibles consecuencias negativas para la conservación de las especies de aves migradoras asociadas a este hábitat. ISSN paper: 1578–665 X ISSN digital: 2014–928 X

Arizaga et al.

Palabras clave: Acrocephalus spp., Conservación de la diversidad biológica, Invasión biológica, Marismas costeras, Índice de deposición de grasa, Bácaris (Baccharis halimifolia). Received: 11 VI 12; Conditional acceptance: 25 IX 12; Final acceptance: 19 XI 12 Juan Arizaga, Edorta Unamuno & Ainara Azkona, Dept. of Ornithology, Aranzadi Sciences Society, Urdaibai Bird Center, Orueta 7, E–48314 Gautegiz–Arteaga, Bizkaia, España (Spain).– Oriol Clarabuch, Catalan Institute of Ornithology, Museu de Ciències Naturals de Barcelona, Passeig Picasso s/n., E–08003 Barcelona, Espanya (Spain). Corresponding author: J. Arizaga. E–mail: jarizaga@aranzadi–zientziak.org

Animal Biodiversity and Conservation 36.1 (2013)

Introduction Migration is considered to have carry–over effects on several parameters in the life cycle of avian migrants (Newton, 2004). Suboptimal stopover places may hamper an adequate fuel deposition rate and compromise not only survival, but also future life history aspects such as mating or breeding success (Sandberg & Moore, 1996; Smith & Moore, 2003). Generally, migrant birds cannot gain enough fuel at a single site to reach their destination areas in a single uninterrupted flight or in several flights without refuelling at en route stopovers. They therefore need to stop over periodically to accumulate sufficiently high energy stores to accomplish the next flight bout successfully. The rate of fuel accumulation at stopover sites influences speed of migration, and it has been considered to be an indicator of habitat quality at these stopover sites (Alerstam & Lindström, 1990). Accordingly, any factor that affects fuel deposition rate, such as food availability or predator disturbance, can be crucial for migrant bird species in terms of migration success or survival. Exotic plants can displace native vegetation, causing habitat changes which are often linked to changes in biodiversity (Vitousek et al., 1997). This phenomenon also affects bird migrants when they land to refuel in an altered habitat (e.g., Cerasale & Guglielmo, 2010). The impact of exotic plants on stopover behavior of migrant bird species is, however, poorly understood (Cerasale & Guglielmo, 2010). This issue is worth taking into consideration, especially if we consider that several bird species, both at population and individual levels, tend to use the same stopover sites year after year (Newton, 2008). Habitat changes in these areas can thus have negative consequences, even though alternative stopover sites may be available. As a global hypothesis, it can be stated that native habitat change (i.e., deterioration) due to the presence of exotic plants will have a negative impact, either in relation to bird assemblage or stopover behavior (fuel deposition rate, stopover duration, etc.), on individual migrants originally associated with native vegetation. The predictions tested in this study were: (1) A decrease in the area covered by native vegetation will have a negative impact on bird abundance at a local scale, because these species are adapted to use their particular native habitats. Consequently, these species would not use areas heavily invaded by exotic plants. From a structural standpoint, the community may change toward species better adapted to the new conditions (e.g., Sol et al., 2002). (2) If a zone affected by an invasive exotic plant species turns into a suboptimal area (e.g., offering worse fuelling opportunities), migrants should be expected to move from this to better nearby areas with native vegetation. Thus, we expect a higher number of within–season recaptures from the affected, a priori suboptimal site (affected by exotic plants), to the unaffected site, especially if migrants are able to look for better sites at a micro–scale level within the same stopover area (Delingat & Dierschke, 2000). (3) A decrease in the

rate of fuel deposition of migrants in a zone affected by exotic vegetation as compared to a zone with only native vegetation is also expected (Cerasale & Guglielmo, 2010), either because migrants associated with native vegetation are not good foragers in a foreign habitat (in suboptimal habitats, migrants should be expected to forage less efficiently; e.g., Jenni–Eiermann et al., 2011), or because local insects (i.e., birds’ food supply) cannot feed on exotic plants and, therefore, food availability is lower than that found in areas of native vegetation. Consequently, migrants departing from a stopover site affected by exotic plants should have lower fuel loads than migrants departing from an unaffected site. The aim of this study was to evaluate the impact of exotic plants on the community structure and stopover ecology of migrant birds, particularly their effect on fuel load and fuel deposition rate. We used data obtained at a coastal wetland in southwestern Europe that had both unaltered reedbeds (Phragmites spp.) and areas where the reedbeds had been largely replaced by the invasive saltbush (Baccharis halimifolia). The study was carried out at two levels: the first level focused on bird assemblages, while the second level focused on the sedge warblers (Acrocephalus schoenobaenus) as an avian model species typical of reedbeds to compare body mass, fuel deposition rate and stopover duration in invaded and control areas. Material and methods Study system Reedbeds in Europe play a relevant role as stopover sites for marsh–associated birds during migratory periods (Schaub et al., 2001; Arizaga et al., 2006). This habitat, however, has suffered a notable decline in several areas due to the saltbush (Sanz et al., 2004), a shrub originally found along the coast of eastern North America (Cronquist, 1980). Typical of plains within coastal marshes, it is currently widespread worldwide as an exotic, invasive plant that occupies wetlands with slight to moderate levels of brackish waters. The saltbush has become a problem of primary importance in many wetlands in Europe, Asia and Australia, and it has been the target of numerous (and usually costly) management plans (e.g., Palmer et al., 1993). Its eradication has thus been a priority in several projects (e.g., Life projects such as LIFE08NAT/E/000055) orientated to preserve habitats of interest in Europe. The reed–associated warblers (Acrocephalus spp.) are a group of closely–related species of small insectivorous birds that are normally adapted to exploit vertically–structured vegetation in Europe, Africa, Asia and some parts of Oceania (Cramp, 1992; Leisler & Schulze–Hagen, 2011). A paradigmatic case of this specialization is birds such as sedge warblers, which mainly forage on aphids (Hyalopterus spp.) found in reedbeds (Bibby & Green, 1981). Sedge warblers breed in humid habitats from the west of Europe to Central Siberia (90º E), between the July isotherms

of 12 and 30ºC (Cramp, 1992). They overwinter in tropical/southern Africa or southern Asia (Cramp, 1992). During migration (mainly from mid–July to September during the autumn migration period), they preferably occupy reedbeds, both along the coast and inland (Cramp, 1992). Sedge warblers have a direct dependence on certain aphid species on which they feed to accumulate large fuel loads before crossing the Sahara Desert in autumn (Bibby & Green, 1981). Stopover normally takes place in strategic areas with good foraging conditions (Bibby & Green, 1981; Bensch & Nielsen, 1999), so habitat changes at stopover areas could be particularly damaging to the conservation of this species. The sedge warbler is present in northern Iberia only during the migration period (Tellería et al., 1999), when they are common in many wetlands. During the autumn migration period, the species accounts for ca. 20% of the captures obtained at a constant–effort ringing station in Urdaibai, a coastal wetland in Northern Iberia (Unamuno & Arizaga, unpubl. data). The sedge warbler is a good avian model to test the effect of invasive plants on stopover behavior of migrants associated with reedbeds, since all birds captured in Urdaibai are non–breeding, true migrants. Study area and data collection Data were obtained at Urdaibai, a Ramsar coastal estuary in the southeastern Bay of Biscay, northern Iberia. The wetland spreads over an area of 945 ha and has a relatively high richness of habitats, determined by tide regimens and the degree of salinity. Such habitats range from beach and dunes (situated within the lowest part of Urdaibai) to tidal flats of limes and plant species adapted to tidal flooding in the lower marsh and reedbeds and freshwater–associated vegetation in the upper marsh and nearby polders. Birds were captured with mist nets of 16 mm mesh (144 linear m/zone) at two zones which had a different degree of invasion by saltbushes (table 1). Originally, both zones were occupied by reedbeds (as dominant vegetation) together with Juncus spp. and Aster spp., but currently the vegetation in one zone has been almost completely replaced by saltbushes. Hereafter, we will respectively call these zones 'control' and 'invaded'. Both zones are found in the upper marsh and are subject to daily tidal flooding. From July to September the area is used by warblers coming from the British Isles and Western–Central Europe (Cantos, 1998) en route to their wintering areas in tropical Africa (Cramp, 1992). Mist nets at both zones were open daily during August 2011 for a period of 4 h starting at dawn. Since the sampling was carried out in parallel with another project, focused on finding key stopovers for the aquatic warbler (A. paludicola) in the bay of Biscay, we also used tape lures with the song of a male aquatic warbler (one lure per 36 linear m of mist nets) (Julliard et al., 2006). Since the sampling effort with both mist nets and lures was constant in these two zones, using tape lures would not be expected to create any bias

Arizaga et al.

for the comparison of stopover behavior between the two trapping sites. Once captured, each bird was ringed and its age (first year or adult bird) was determined according to Svensson (1996). Additionally, we measured body mass (± 0.1 g) and the length of the P3 primary feather (± 0.5 mm; numbered from outermost to innermost). Following measurements, the birds were released. No bird was retained for longer than 1 h. Statistical analyses We compared community richness (number of birds), as well as a diversity index (H’) using a bootstrap procedure as calculated by the software PAST 2.1 (Hammer et al., 2001). H’ stands for the Shannon index, which ranges from 0 (communities with only a single taxon) to high values in communities with many taxa, each with few individuals. The bootstrap procedure consisted in generating a 95% confidence interval (CI) by taking 1,000 random sub–samples from the total pooled data. The number of captures per day at each zone was compared with a t–test. Number of captures was log–transformed to normalize the data. Contingency tables were used to compare the proportion of recaptures between zones, and to test whether cross–recaptures from the control zone into the invaded zone were more common than vice versa. As body mass in excess of structural mass in migrants is mainly stored as fat and, to a lesser extent, proteins (Jenni & Jenni–Eiermann, 1998), we used body mass controlled for body size (including P3 as a covariate) as a surrogate for fuel load (e.g., Arizaga et al., 2010, 2011a). Analyses of fuel load and fuel deposition rate were run only for sedge warblers. Captures from ordinary trapping sessions at a stopover site are subject to certain constraints. In particular, the first and last captures of each individual are not always obtained on the exact dates of arrival or departure (Schaub et al., 2001). It is thus impossible to estimate daily fuel load var iation for the entire stopover period of each individual. However, trapping sessions at ringing stations suffice to estimate fuel load of stopping–over migrants (Schaub & Jenni, 2000a, 2000b; Arizaga et al., 2010, 2011b). To obtain a closer estimation of fuel load both on arrival at the site and on departure, we selected the ten lightest and heaviest migrants in each zone (Ellegren & Fransson, 1992). Since migrants are expected to gain fuel whilst stopping–over at a site, the lightest birds are those most likely to have just arrived, whereas the heaviest birds are those most likely about to depart. We compared fuel loads between zones using an ANCOVA on body mass with zone and fuel category (lightest/heaviest) as factors, and P3 as a covariate that controlled for body size. Although body mass can differ between age classes (Grandío, 1999), we did not consider age as an additional factor due to the relatively low sample size, particularly in one of the zones (the invaded one). Body mass was normally distributed (K–S test: P > 0.05).

Animal Biodiversity and Conservation 36.1 (2013)

Table 1. Main (mean ± SE) vegetation characteristics at the two trapping stations. Statistics are calculated for a sample size of 12 nets at each trapping site. The habitat was studied within an area of 240 m2 from each side of each net (i.e., 480 m2 around each net). The distance (straight line) from one zone to another was 300 m: * We show here the first and second dominant herbaceous plants. Tabla 1. Características principales (media ± EE) de la vegetación en las dos estaciones de trampeo. Los estadísticos se calcularon para un tamaño de muestra de 12 redes en cada punto de trampeo. El hábitat se estudió en una superficie de 240 m2 a cada lado de las redes (es decir, 480 m2 alrededor de cada red). La distancia (en línea recta) entre las dos zonas era de 300 m: * Aquí mostramos la primera y la segunda plantas herbáceas dominantes.

Affected by saltbush (Invaded zone) 43º 20' 43.44'' N 02º 39' 44.30'' W

Non–affected by saltbush (Control zone) 43º 20' 52.90'' N 02º 39' 44.04'' W

Vegetation cover (%) No vegetation (mud flats and water) Tree

0.4 ± 0.4% –

4.6 ± 1.9% 3.3 ± 2.2%

Bush

88.8 ± 4.7%

5.6 ± 2.8%

Herbaceous

10.8 ± 4.5%

86.5 ± 3.8%

Dominant vegetation Tree Bush Herbaceus*

–

Tamarix

Baccharis

Juncus–Phragmites

Phragmites–Aster

Vegetation height (over 12 nets) < 1 m

–

2/12

1–2 m

–

3/12

2–3 m

6/12

7/12

> 3 m

6/12

–

Contingency tables were also used to see whether the proportions of the first–year birds and adults differed between the control and invaded zones. This analysis was done only for sedge warblers. The importance of a stopover site cannot be determined only in relation to how many birds a given site hosts. Consideration must also be given to sites which allow migrants to gain fuel (Alerstam & Lindström, 1990). Indeed, the fuel deposition rate is considered to reflect the habitat quality of a stopover site (Newton, 2008). To estimate fuel deposition rate, we considered the difference of body mass between the last and first capture event of each bird, divided by the number of days elapsed between these two events. To compare the fuel deposition rate between zones we used an ANCOVA on the fuel deposition rate with zone as a factor, and P3 and number of days between first and last capture as covariates. Since the estimation of the fuel deposition rate using recaptures of ringed migrants can be subject to bias due to a handling effect, especially when recaptures are obtained shortly after the first capture (Schwilch

& Jenni, 2001), we repeated the analysis removing the migrants recaptured the day after their first capture event. Intermediate recaptures (in migrants recaptured more than once) were not considered in any of the analyses. Finally, we also calculated the minimum stopover duration as: (ti–t0) + 1, where ti and t0 were respectively the date of the last and first capture event of each bird, respectively. The '+1' was added because sedge warblers are nocturnal migrants. We did not use a more accurate estimation of stopover duration (e.g. by means of the use Cormack–Jolly–Seber models, which allow separation of the survival and recapture probabilities) because our sample was too small to allow us to separately estimate survival (here, probability of a bird remaining at the site from one day to the next) and the recapture rate (Lebreton et al., 1992). We used a t–test assuming heteroscedasticity to test if the stopover duration differed between the control and invaded zones. For statistical procedures, we used the software PAST 2.1. (Hammer et al., 2001), and SPSS 18.0. All means are given ± SE.

Results Assemblage characteristics We captured a total of 30 bird species in the invaded zone and 28 of these were passerines. In the control zone we captured 28 species and 26 were passerines. The number of captures at each zone was 529 and 558, respectively (table 2). Bird assemblage did not differ in terms of richness, but it did differ in terms of diversity (table 3), which was lower in the control zone. Another notable aspect was that bird assemblage in the control zone was richer in species typical of reedbeds and wetlands, while in the invaded zone assemblage was richer in woodland–related species (table 3). In the control zone, six out of the ten most abundant species were clearly associated with reedbeds and wetlands (only classified as 'reed' in table 2), while conversely, in the invaded zone, six species were clearly associated with woodlands (classified as 'wood' in table 2; fig. 1). The aquatic warbler, the only species linked to wetlands and globally threatened, only appeared in the control zone (n = 7). Capture rates were similar in both zones (invaded zone: 18.0 ± 2.7 cap./day, n = 29; control zone: 18.2 ± 3.0 cap./day, n = 31; t58 = 0.25, P = 0.80). However, the difference was significant when only captures of species clearly related to reedbeds/wetlands were considered (invaded zone: 6.2 ± 0.7 cap./day, n = 29; control zone: 10.0 ± 1.0 cap./day, n = 31; t58 = 3.12, P = 0.003). Recaptures We found 26% of the birds captured in the control zone were recaptured (at any of the study zones), while only 15% of the birds captured in the invaded zone were recaptured (at any of the study zones). This difference was significant (x12 = 13.04, P < 0.001). Focusing on species related to reedbeds, the proportion was 38% and 25%, respectively (x12 = 4.71, P = 0.03). Moreover, 6% of the birds first captured in the control zone were thereafter recaptured in the invaded zone, while 24% of the birds first captured in the invaded zone were thereafter recaptured in the control zone. This difference being significant (x12 = 9.78, P = 0.002). Focusing on species typical of reedbeds, the difference was even higher (6% versus 32%; x12 = 12.97, P < 0.001), even for sedge warblers (3% versus 58%; x12 = 11.49, P = 0.005), indicating that species from reedbeds tended to move to the control zone much more often than the other way around. Body mass, fuel deposition rate, age ratios and stopover duration in sedge warblers Sedge warblers captured in the invaded and control zone did not differ in body mass when considering only the ten lightest migrants in each zone, but there was a difference at the other end of the scale (fig. 2; ANCOVA: zone, F1,39 = 20.73, P < 0.001; fuel category, F1,39 = 385.25, p < 0.001; zone × fuel category, F1,39 = 35.18, P < 0.001; P3, F1,39 = 0.001, P = 0.98).

Arizaga et al.

The body mass (controlled for body size) of the ten heaviest migrants in the control zone was higher than the body mass of the ten heaviest migrants in the invaded zone (fig. 2). Sedge warblers last recaptured in the control zone showed significantly higher fuel deposition rates than those recaptured in the invaded zone (table 4). The proportion of first–year birds and adults did not differ between the invaded (48.6%) and control zones (57.1%; x12 = 0.80, P = 0.37). Minimum stopover duration in the control zone (6.4 ± 0.9 days, n = 32) was longer than in the invaded zone (3.0 ± 0.6 days, n = 3; t17.228 = 3.20, P = 0.005). Discussion The saltbush is an exotic shrub that invades humid habitats in Europe and displaces the native vegetation (e.g., reedbeds). In this work, we show that the saltbush may have a negative impact on both the abundance and stopover characteristics (fuel management and habitat use) of migrants associated with reedbeds. Bird assemblage characteristics The diversity index was found to reach higher values in the invaded zone than in the control zone, mainly due to the fact that in the control zone there were three passerines accounting for 70% of the captures, whilst in the invaded zone only two species were clearly dominant, making up 50% of the captures. Reedbeds constitute particular habitats where vertical vegetation is dominant. Exploitation of this type of vegetation requires high adaptation/specialization, so reedbeds constitute an adequate habitat for relatively few species (Poulin et al., 2000, 2002; Leisler & Schulze–Hagen, 2011) compared to other habitats where horizontal vegetation is dominant (e.g., Arizaga et al., 2009). It is of note that in species that are typical of wooded areas and occupy reedbeds during the non–breeding period, such as robins (Erithacus rubecula), it is the juvenile fraction which is detected in reedbeds and wetlands (suboptimal habitats), whilst adults monopolize more suitable habitats such as forests (Figuerola et al., 2001). Thus, the captures in the control zone were mostly of species typical of reedbeds and wetlands, whereas the species detected in the invaded zone were, to a larger extent, typical of forested habitats. Moreover, reedbed–associated species were more abundant in the control zone than in the invaded zone. This result highlights that the saltbush not only had a negative impact on the abundance of reedbed–associated species, but also affected bird assemblage by facilitating the presence of woodland passerines. Stopover behavior The finding that the proportion of recaptures was higher for the control zone than for the invaded zone suggests that migrants tended to stay longer in the control zone. We cannot overlook the possibility that the recapture rate was site–dependent and this may have imposed some

Animal Biodiversity and Conservation 36.1 (2013)

Table 2. Number of captures / recaptures of species caught with mist nets in the invaded and control zones. Species were assigned to the specific habitats that they normally occupy during the non–breeding period, either when foraging or roosting (Cramp, 1988, 1992; Cramp & Perrins, 1994): Reed. Reedbeds and wetlands; Wood. Woodland, including shrubs, hedgerows, forested areas and parks with trees/ shrubs; Others. Urban areas, open habitats, crops. Recaptures refer to birds recaptured ≥ 1 days after the first capture event, to sites where birds were first captured (i.e., a bird captured in the control zone and subsequently recaptured in the invaded area is included in the column of recaptures for the control zone), and only one recapture per bird is considered: Cap. Captures; Rcap. Recaptures. Tabla 2. Número de capturas / recapturas de las especies atrapadas en redes japonesas en las zonas invadida y de control. Las especies se asignaron a los hábitats que suelen ocupar durante el período no reproductivo, bien mientras forrajeaban, bien mientras reposaban (Cramp, 1988, 1992; Cramp & Perrins, 1994): Reed. Carrizales y humedales; Wood. Tierras arboladas, incluidos arbustos, setos, áreas boscosas y parques con árboles y arbustos; Others. Áreas urbanas, hábitats abiertos y cultivos. Las recapturas hacen referencia a las aves que se volvieron a capturar uno o más días después de la primera captura y a los sitios en que se había capturado a las aves la primera vez (esto es, un ave capturada en la zona de control y posteriormente vuelta a capturar en la zona invadida figura en la columna de las recapturas de la zona de control); solo se tuvo en cuenta una recaptura por ave: Cap. Capturas; Rcap. Recapturas. Specific name

Invaded zone Code

Main habitats

Control zone

Cap.

Rcap.

Cap.

Rcap.

A. arundinaceus

ACRARU

Reed

A. paludicola

ACROLA

Reed

A. schoenobaenus

ACRSCH

Reed

106

A. scirpaceus

ACRSCI

Reed

106

145

A. caudatus

AEGCAU

Wood

A. atthis

ALCATT

Reed

C. brachydactyla

CERBRA

Wood

C. cetti

CETCET

Reed

C. juncidis

CISJUN

Others + Reed

E. rubecula

ERIRUB

Wood

F. hypoleuca

FICHYP

Wood

F. coellebs

FRICOE

Wood + Others

G. glandarius

GARGLA

Wood

H. pollyglotta

HIPPOL

Wood

J. torquilla

JYNTOR

Wood

L. collurio

LANCOL

Others + Wood

L. luscinioides

LOCLUS

Reed

L. naevia

LOCNAE

Wood + Reed

L. megarhynchos

LUSMEG

Wood

L. svecica

LUSSVE

Reed

P. caeruleus

PARCAE

Wood

P. cristatus

PARCRI

Wood

P. major

PARMAJ

Wood

P. domesticus

PASDOM

Others + Reed

P. montanus

PASMON

Wood + Reed

P. ibericus

PHYIBE

Wood

P. trochylus

PHYLUS

Wood + Reed

155

145

R. aquaticus

RALAQU

Reed

Arizaga et al.

Table 2. (Cont.)

Invaded zone

Specific name

Code

R. ignicapillus

REGIGN

S. torquata

Main habitats

Control zone

Cap.

Rcap.

Cap.

Rcap.

Wood

SAXTOR

Wood

S. vulgaris

STUVUL

Others + Reed

S. atricapilla

SYLATR

Wood

S. borin

SYLBOR

Wood

S. communis

SYLCOM

Wood

T. troglodytes

TROTRO

Wood

T. merula

TURMER

Wood

bias in relation to the estimation of stopover duration (Schaub et al., 2001). Unfortunately, our sample was too small to separately estimate survival (here meaning the probability of a bird remaining in the site from one day to the next) and the recapture rate (Lebreton et al., 1992). Mean fuel deposition rate in the control zone (+ 0.2 g/d) was even higher than the rate reported in another coastal reedbed in northern Iberia (+ 0.1 g/d; Grandío, 1998). In optimal stopover habitat, the species has been reported to reach mean rates of > + 0.3 g/d (Schaub & Jenni, 2000a). Thus, the rate of fuel accumulation at Urdaibai seemed high, and hence the control zone can well be considered an optimal habitat for the species. The fact that the proportion of birds that moved to the control zone from the invaded zone was higher than the opposite scenario supports the hypothesis that the saltbush area was suboptimal for migrants as compared to reedbeds. Also supporting the hypothesis that the saltbush did not provide a proper fuelling chance for migrants was the finding that the fuel deposition rate in the control zone was much higher than in the invaded zone. Together with the previous result, this is in accordance with the idea that migrants quickly depart from a site when experiencing a very low fuel accumulation rate (Alerstam & Lindström, 1990). From an evolutionary standpoint, this response allows migrants to look for better stopover places and thus have a second chance

to gain fuel during migration period. However, this flexible behavior depends on current fuel load, and birds with small fuel loads will therefore be hampered in finding a better stopover site if this is associated with long displacements. In this scenario, invasive plants imposing radical habitat changes, like the saltbush, constitute a severe problem from a fuelling standpoint. This is particularly applicable if native vegetation is replaced by exotic plants across very large areas. The substitution of reedbeds by saltbushes in several coastal marshes in the Bay of Biscay, including Urdaibai, must be regarded as a problem with unknown consequences for the conservation of migrant bird species associated with reedbeds. The body mass of the ten heaviest and the ten lightest sedge warblers from each zone was used as a surrogate of body mass at departure and arrival, respectively (Ellegren & Fransson, 1992). The analysis of these data indicate that sedge warblers arrived at both zones with a similar fuel load, but departed with more fuel from the control zone, a result that is in accordance with the control zone favoring fuel accumulation. This result cannot be considered to be caused by a possible bias between age classes (body mass can vary between age classes; Grandío, 1999) and zones, since the proportion of each age class was constant for the invaded and control zones. Our results also support the hypothesis that a number of

Table 3. Diversity–related statistics between the invaded and control zones. Tabla 3. Estadísticos relacionados con la diversidad entre las zonas invadida y de control. Taxa (richness) Shannon diversity index (H’)

Invaded zone

Control zone

Bootstrap (P–values)

0.556

2.485

2.124

< 0.001

Animal Biodiversity and Conservation 36.1 (2013)

Invaded zone Others TROTRO AEGCAU ERIRUB PHYIBE CETCET SYLATR TURMER ACRSCH ACRSCI PHYLUS

Control zone

Others ALCATT ACROLA PARCAE CETCET SYLCOM LUSSVE CISJUN ACRSCH PHYLUS ACRSCI

10 20 30 40 0 Captures (%)

10 20 30 Captures (%)

Fig. 1. Relative number of captures (first capture event) of the ten most frequent species at each sampling zone. Fig. 1. Número relativo de capturas (primera captura) de las diez especies más frecuentes en cada zona de muestreo.

warblers in the saltbush zone may have been able to compensate for their low fuel accumulation rate/ fuel load on departure by moving to a better site (the control zone) with higher fuelling chance.

Body mass (g)

16 14

Conclusion and perspectives The saltbush had a negative impact on several species closely associated with reedbeds, and promoted

Lightest Heaviest

12 10 8 6 4 2 0

Invaded

Control

Fig. 2. Mean (± SE) body mass of sedge warblers first captured at the invaded and control zones. For each zone, we considered the ten lightest and heaviest captures of sedge warblers in each zone (ages pooled). Also within each zone, average values bearing the same letters represent non–significant differences in body mass between fuel categories (lightest / heaviest). Differences in body size were controlled including P3 as a covariate. Fig. 2. Media (± EE) del peso corporal de los carricerines que se capturaron por primera vez en las zonas invadida y de control. Para cada zona, se tuvieron en cuenta las capturas de los diez carricerines de mayor y menor peso (agrupadas por edad). También para cada zona, los valores medios asociados a las mismas letras representan diferencias no significativas en cuanto al peso corporal entre las categorías de cantidad de grasa (más ligero / más pesado). Las diferencias de tamaño corporal se controlaron con la inclusión de la P3 como covariable.

Arizaga et al.

Table 4. Mean (± SE) fuel deposition rates for sedge warblers captured in the invaded and control zones. Sample size in brackets. Data on both zones were compared with an ANCOVA on fuel deposition rate with zone as factor and P3 and number of days elapsed between the first and last captures as covariates; we only show F–values used to test for the effect of zone. The number of recaptures differed from table 1 since not all birds had the P3 measured. Tabla 4. Media (± EE) de los índices de deposición de grasa para los carricerines capturados en las zonas invadida y de control. El tamaño de la muestra entre paréntesis. Los datos de ambas zonas se compararon con una ANCOVA del índice de deposición de grasa con la zona como factor y la P3 y el número de días transcurridos entre la primera captura y la última como covariables; mostramos únicamente los valores de F utilizados para comprobar los efectos de la zona. El número de recapturas difirió de las de la tabla 1 porque no se midió la P3 a todas las aves. Days All > 1

Invaded zone

Control zone

– 0.1 ± 0.1

+ 0.2 ± 0.1

0.062

0.812

(4)

(41)

– 0.2 ± 0.2

+ 0.1 ± 0.1

4.316

0.016

(3)

(31)

a noticeable change on bird assemblage, which was found to be enriched by species typical of woodland habitats. A migrant species strongly associated with reedbeds during migration period, sedge warblers captured at the control zone departed with a higher fuel load, showed a higher fuel deposition rate, and remained for longer than those that stayed at a nearby site occupied by saltbushes (an exotic invasive bush from America). Our results suggest that saltbushes reduce the habitat quality for migratory sedge warblers. Unfortunately, data comparing bird migration in the control and invaded areas before saltbushes invasion are not available. Consequently, although unlikely, it might be possible that both areas already differed in their quality before invasion. Future research with replicates will be necessary to confirm these results. Further studies are also needed to accurately quantify how the wetlands (coastal marshes) from southern Europe have been affected by saltbushes, and to better understand the ecology of migrants so as to to be able to properly evaluate the real impact of saltbushes on the stopover behavior and migratory performance of bird migrants associated with reedbeds. From a management standpoint, restoration of reedbeds must be regarded as a priority tool to preserve the optimal stopover habitat for passerines associated with this type of vegetation. Acknowledgements This research was funded by the Basque Government, the Bizkaia Council and BBK. Ringing activities were authorized by the Bizkaia Council. V. Salewski, D. Serrano and an anonymous editor provided very valuable comments that helped us to improve an earlier version of this work.

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Animal Biodiversity and Conservation 36.1 (2013)

Dalechampii oak (Quercus dalechampii Ten.), an important host plant for folivorous lepidoptera larvae M. Kulfan, M. Holecová & P. Beracko

Kulfan, M., Holecová, M. & Beracko, P., 2013. Dalechampii oak (Quercus dalechampii Ten.), an important host plant for folivorous lepidoptera larvae. Animal Biodiversity and Conservation, 36.1: 13–31. Abstract Dalechampii oak (Quercus dalechampii Ten.), an important host plant for folivorous lepidoptera larvae.— We conducted a structured analysis of lepidoptera larvae taxocenoses living in leaf bearing crowns of Dalechampii oak (Quercus dalechampii Ten.) in nine study plots in the Malé Karpaty Mountains (Central Europe). The differences between lepidoptera taxocenoses in individual oak stands were analyzed. A total of 96 species and 2,140 individuals were found. Species abundance peaked in May, while number of species and species diversity reached the highest values from April to May and from April to June, respectively. Abundance showed two notable peaks in flush feeders and in late summer feeders. Lepidoptera taxocenosis in the study plot Horný háj (isolated forest, high density of ants) differed significantly from all other taxocenoses according to Sörensen’s index of species similarity, species diversity, analysis of similarity on the basis of permutation and pairwise tests (ANOSIM), seasonal variability of species composition, and NMDS ordination. Key words: Moths, Caterpillars, Q. dalechampii, Malé Karpaty Mountains, SW Slovakia. Resumen El roble de dalechampii (Quercus dalechampii Ten.), una importante planta hospedadora de las larvas de lepidópteros filófagos.— Llevamos a cabo un análisis estructurado de las taxocenosis de larvas de lepidópteros que viven en las copas del roble de dalechampii (Quercus dalechampii Ten.) en nueve parcelas del estudio en los Pequeños Cárpatos (Europa central). Se analizaron las diferencias entre las taxocenosis de lepidópteros de cada roble. Se hallaron 96 especies y 2.140 individuos. La abundancia de especies alcanzó su valor más elevado en mayo, mientras que el número y la diversidad de especies fueron máximos desde abril hasta mayo y desde abril hasta junio, respectivamente. La abundancia mostró dos máximos notables en las larvas que se alimentan durante la brotación y las que se alimentan al final del verano. La taxocenosis de los lepidópteros en la parcela del estudio Horný háj (un bosque aislado con una elevada densidad de hormigas) difirió significativamente de las demás taxocenosis según el índice de Sörensen para la similitud de las especies, la diversidad de las especies, el análisis de la similitud sobre la base de las pruebas de permutación y las pruebas de pares (ANOSIM), la variabilidad estacional de la composición de especies y el escalamiento multidimensional no métrico (NMDS por sus siglas en inglés). Palabras clave: Polillas, Orugas, Q. dalechampii, Pequeños Cárpatos, Eslovaquia sudoccidental. Received: 11 VII 12; Conditional acceptance: 27 X 12; Final acceptance: 20 XII 12 Miroslav Kulfan & Pavel Beracko, Dept. of Ecology, Fac. of Natural Sciences, Comenius Univ., Mlynská dolina B–1, SK–84215 Bratislava, Slovakia.– Milada Holecová, Dept. of Zoology, Fac. of Natural Sciences, Comenius Univ., Mlynská dolina B–1, SK–84215 Bratislava, Slovakia. Corresponding author: M. Kulfan. E–mail: kulfan@fns.uniba.sk

ISSN: 1578–665X

Kulfan et al.

Introduction Oaks belong to the woody plants that host the richest insect assemblages in Central Europe (Patočka et al., 1999). Lepidoptera larvae have been shown to be the most important group of oak defoliators (Patočka et al., 1962, 1999). About 250 lepidoptera species are known to damage the assimilation tissue of oaks in Central Europe (Patočka et al., 1999; Reiprich, 2001). Lepidoptera fauna on some oak species in Central Europe have been relatively well studied (Patočka et al., 1962, 1999; Csóka, 1990–1991, 1998a, 1998b; Kulfan, 1990, 1997; Kulfan, 1992; Kulfan et al., 1997, 2006; Kulfan & Degma, 1999; Turčáni et al., 2009, 2010; Parák et al., 2012, etc.). Taxocenoses of lepidoptera caterpillars on three oak species from Slovakia and the Czech Republic (Quercus robur, Q. petraea and Q. cerris) have been used to explain why there are so many species of herbivorous insects in tropical rainforests (Novotny et al., 2006). However, the lepidoptera fauna related to Q. dalechampii growths has been poorly explored in Europe. A total of nine lepidoptera miner species from families Nepticulidae, Tischeriidae and Gracillariidae have been recorded on Q. dalechampii in southern Slovakia (Arboretum Čifáre) (Skuhravý et al., 1998). Kollár (2007) mentions the species Phyllonorycter roboris (lepidoptera miner) as a pest of Q. dalechampii in Slovakia. Stolnicu (2007) studied lepidoptera leaf– miners on Q. dalechampii in Romania. Kulfan (2012) partially studied economically most important pest species on Q. dalechampii in Central Europe. Dalechampii oak (Quercus dalechampii Ten.) is one of the most common oaks in Europe and is naturally distributed in Western Italy, Sicily, Greece, Albania, Montenegro, Macedonia, Bosnia & Herzegovenia, Serbia, Slovenia, Austria, Hungary, Slovakia, Romania, and Bulgaria. The main aims of the present study were: (i) to analyze the structure taxocenoses, alpha diversity and representation of trophic groups and seasonal guilds of lepidoptera en bloc on Dalechampii oak; (ii) to complete data concerning biodiversity of lepidoptera species feeding on oaks in Central Europe; and (iii) to highlight the differences among the individual study plots representing various types of oak forests, with emphasis on fragmentation, forest age and crown canopy. Material and methods Material was collected by the beating method into a tray of 1 m diameter (one quantitative sample = beating from 25 branches) on nine selected plots at regular 2–weekly intervals from April to October 2000–2002. Samples were taken from branches at a height of about 1–2.5 m above ground with varying exposure to cardinal points. Larvae were identified using the keys by Gerasimov (1952), Patočka (1954, 1980) and Patočka et al. (1999). Seasonal guilds of lepidoptera caterpillars were established according to Turčáni et al. (2009). The complete linkage clustering in combination with Sörensen’s index and Wishart’s similarity ratio

was used to classify the taxocenoses. Visualization of dendrograms was done by computer program Syn–tax, Version 5.0 (Podani, 1993). Diversity of taxocenoses was characterised using Pielou’s index of equitability, Shannon–Wiener’s index of total species diversity, and Simpson’s index of dominance (Poole, 1974; Ludwig & Reynolds, 1988). Shannon–Wiener diversity indices were compared using the t–test (Poole, 1974). Ordination was carried out with non–metric multidimensional scaling (NMDS) using the Bray–Curtis dissimilarity coefficient. One–way analysis of similarities (ANOSIM) was used to identify difference in species variability of the lepidoptera taxocenosis in the study plots during the year. Hierarchical (nested) ANOVA was used to examine spatial (locality) and temporal (sampling months) variation in the distribution of the total abundance, number of species, taxa and species diversity of lepidoptera. The model contained factors (terms) representing the effects of locality and sampling date nested in locality. Multiple sample comparisons were used to identify significant differences in the number of individuals, number of species and species diversity between localities and sampling months. The hypothesis that occurrences of three types of feeding specialization are randomly distributed throughout the vegetation season was tested according to Poole & Rathcke (1979). Differences of means and dispersion of species numbers in feedings groups were analyzed by Tukey’s pairwise comparison and Levene’s test in ANOVA, respectively. Analyses of variance and Tukey’s pairwise comparison were used to identify differences between the number of species and the number of individuals in seasonal gilds. The nomenclature and systematic classification of the lepidoptera species were used according to Laštůvka & Liška (2011). The trophic groups of lepidoptera larvae were established according to Brown & Hyman (1986). The map (fig. 1) and pedological and phytocoenological characteristics of the investigated area are given in detail by Zlinská et al. (2005). Voucher specimens (in ethanol) are deposited at the Faculty of Natural Sciences, Comenius University, Bratislava, Slovakia. Study area The lepidoptera larval stages on Quercus dalechampii were studied in the territories of the Protected Landscape Area of Malé Karpaty and Trnavská pahorkatina hills situated in the centre of Europe in the western part of Slovakia.The vast majority of the plots are located in the southern to northern part of the Malé Karpaty Mountains (Mts.) at altitudes of 240–350 m a.s.l. and an average annual temperature of 8–9°C. Study plots in Trnavská pahorkatina hilly land are situated near the Malé Karpaty Mts. at an altitude of 240 m. The annual precipitation in both territories is about 650–800 mm. Study plots (abbreviation of study plot in parentheses used in the text): Vinosady (VI), 48º 19' N, 17° 17' E, 280 m a.s.l.: a 60–80–year–old forest at the foot of the Kamenica

Animal Biodiversity and Conservation 36.1 (2013)

Study plot

N 72

Bradlom

500

Jablonica

300

GRN DFS

30 0

300

48º 36' N

Chtelnica

200 300

Rohožnik

500

Smolenice

48º 30' N

500 200

48º 24' N

48º 42' N

Vrbové

200

Čachtice

200

S Brezová pod

Settlement

48º 48' N 500

Contour line (m a.s.l.) Malé Karpaty Mts. border

300

200

Lozorno

CA 30

Stupava 77 300

200

48º 06' N

Bratislava 79 17º 00'E

17º 10' E

48º 00' N

Modra

Pezinok

Sv. Jur 48º 12' N

63 64 65 66 67 68 69 49º70 71 72 73 74 75 76 77 78 79 48º80 81 82 83

17º 20' E

25 km

48º 18' N 17º 30' E

17º 40' E

17º 50' E

17º 18º 19º 20º 21º 22º 63 67 68 69 70 71 72 73 74 75 76 77 78 79 80 81 82 83 84 85 86 87 88 89 90 91 92 93 94 94 96 97 98 99 00 01

GRN DFS: Grid references number of the Databank of Slovak fauna

Map design:

Fig. 1. Study area with location of the study plots. Fig. 1. Área del estudio con la ubicación de las parcelas del estudio.

hill, NW and W oriented, with drier subxerophilous meadows and shrub complexes. Besides Quercus dalechampii, the tree stratum consists of Q. cerris and Acer campestre. Cajla (CA), 48º 20' N, 17° 16' E, 260–280 m a.s.l.: an 80–100–year–old forest at the foot of the Malá cajlanská homola hill, S oriented and neighbouring meadows and vineyards on S and E, from N and W closed forest complexes. Quercus dalechampii and Carpinus betulus predominate in the tree layer.

Fúgelka (FU), 48º 22' N, 17° 19' E, 350 m a.s.l.: an 80–100–year–old forest near the Dubová village, S oriented. Besides Quercus dalechampii, the tree stratum consists of Acer pseudoplatanus. Lindava (LI) (Nature Reserve), 48º 22' N, 17° 22' E, 240 m a.s.l.: an 80–100 (120)–year–old forest near the village of Píla. Quercus dalechampii and Q. cerris predominate in the tree layer. Horný háj (HH), 48º 29' N, 17° 27' E, 240 m a.s.l.: a larger complex of an island forest 60–80–years old

Kulfan et al.

near the village of Horné Orešany, surrounded by fields and vineyards, W and SW oriented. Quercus cerris, Q. dalechampii, Carpinus betulus and Fraxinus excelsior predominate in the tree layer. Lošonec–lom quarry (LL), 48º 29' N, 17º 23' E, 340 m a. s. l.: an 80–100–year–old forest SW oriented, neighbouring with mesophilous meadows and pastures. The tree layer consists of Quercus dalechampii, Q. cerris and Carpinus betulus. The leaf litter, herbage undergrowth and trees are strongly covered with calcareous dust from a nearby quarry. Lošonský háj (LH) (Nature Reserve), 48º 28' N, 17º 24' E, 260 m a.s.l.: an 80–100–year–old oak–hornbeam forest NE oriented, surrounded by closed forest complexes. Quercus dalechampii, Q. cerris and Carpinus betulus predominate in the tree stratum. Naháč–Kukovačník (NA), 48º 32' N, 17º 31' E, 300 m a.s.l.: a small forest island, approximately 40–60– year–old surrounded by fields and pastures, NE oriented. Quercus dalechampii, Q. cerris and Carpinus betulus predominate in the tree layer. Naháč–Katarínka (NK) (Nature Reserve), 48° 33' N, 17º 33' E, 340 m a.s.l.: a 40–60–year–old forest NW oriented, surrounded by closed forest ecosystems. Quercus dalechampii and Carpinus betulus predominate in the canopy. Only abbreviations of the study plots are used in the following text. The study plots LI and HH are situated in Trnavská pahorkatina hills and the others are in the Malé Karpaty Mts.

Table 1. Family dominance (%) of lepidoptera larvae on Quercus dalechampii in the Malé Karpaty Mountains in 2000–2002 (based on total number of individuals). Tabla 1. Dominancia por familia (%) de las larvas de lepidópteros que se encontraron en Quercus dalechampii en los Pequeños Cárpatos entre los años 2000 y 2002 (con respecto al número total de individuos). Family / year

2000

2001

2002

Total

Psychidae

0.00

0.17

0.00

0.05

Bucculatricidae

0.00

0.43

0.19

Gracillariidae

0.17

0.00

0.05

Ypsolophidae

2.48

2.15

1.29

1.87

Chimabachidae

2.74

1.49

1.12

1.73

Peleopodidae

8.85

1.65

1.25

3.46

Coleophoridae

8.77

3.63

4.09

5.28

Gelechiidae

0.17

0.83

0.22

0.37

Tortricidae

6.79 19.64

9.68 11.68

Lycaenidae

0.33

0.43

0.37

Pyralidae

1.82

0.50

1.29

1.21

Drepanidae

0.33

0.11

0.23

Geometridae Results From 2000–2002, a total of 2,140 Lepidoptera larvae were collected in nine study plots with Quercus dalechampii. They represented 96 species from 17 families (appendix 1). The families Geometridae, Noctuidae and Tortricidae encompassed the highest number of species found (27, 23, and 13, respectively) (appendix 1). The lowest number of species (18 species) were found in HH (appendix 1). Six species (Coleophora siccifolia, Lomographa temerata, Peribatodes rhomboidaria, Acronicta auricoma, Orthosia opima and Amata phegea) were found on oaks for the first time in Slovakia (cf. Hrubý, 1964; Patočka et al., 1999). A. phegea is one of six species presenting first records of lepidoptera larvae feeding on oaks. This species probably entered the oak crown from the surrounding low vegetation because it has not been found previously on trees according to the literature (Reiprich, 2001). The most abundant families were Geometridae and Noctuidae (appendix 1, table 1). The families Tortricidae and Erebidae achieved relatively high dominance, mainly due to the species Aleimma loeflingiana (Tortricidae) and Lymantria dispar (Erebidae) (appendix 1, table 1). Species with dominance higher than 10% were Lymantria dispar in HH, Operophtera brumata in CA (calamitous oak pests), Cosmia trapezina in LI, Aleimma loeflingiana in FU (an important pest of oaks) and Cyclophora linearia in HH (cf. Patočka et al., 1999; appendix 1).

32.62 39.11 42.58 38.79

Notodontidae

5.46

0.17

Erebidae

7.95

5.94

9.04

7.85

Nolidae

3.48

1.98

1.72

2.29

Noctuidae No individuals

0.22

1.68

18.05 22.11 26.56 22.90 604

606

930 2,140

The species Lymantria dispar, Cyclophora linearia, Pseudoips prasinana and Carcina quercana reached the highest dominance on the species poorest study plot HH when compared with other plots (appendix 1). Characteristic species of the plot LL covered with calcareous dust are as follows: Tortrix viridana, Conobathra tumidana, Aleimma loeflingiana, Agriopis leucophaearia and Alsophila aceraria. Three lepidoptera species, Archips podana, Eudemis profundana and Apocheima hispidaria (appendix 1), were found only in this plot but abundance was low. Lepidoptera species Agriopis marginaria, Cosmia trapezina, Orthosia cruda and Lymantria dispar (apendix 1) were typical of the lighter, sparser and younger oak stands (study plots NK, LI, CA, VI). The vast majority of Lepidoptera belonged to the monovoltine species with main occurrence in spring. Further oligophagous species (Cyclophora linearia and Ennomos erosaria) and polyphagous species (Parectropis similaria and Colocasia coryli) belonged

Animal Biodiversity and Conservation 36.1 (2013)

to the bivoltine species. Watsonalla binaria proved to be trivoltine species (appendix 1). Most species found belonged to the trophic group of generalists (64 species). Narrow oligophages (18 species) feeding on oaks are considered to be typical oak species. Only six species belonged to wider oligophages. The value of Shannon–Wiener´s diversity index of the richest lepidoptera taxocenosis (NK, H' = 3.428) and the poorest taxocenosis (HH, H' = 2.505) was statistically significantly different from other taxocenoses (T–test, P < 0.05) (table 2). A detailed algorithm is given by Poole (1974). The richest taxocenosis NK includes 462 individuals representing 52 species; of these, seven species dominate at least 5%. The poorest taxocenosis HH includes only 44 individuals belonging to 18 species; 4 of these species dominate over 5% (appendix 1). Poor qualitative–quantitative taxocenosis of lepidoptera larvae on island forest HH is also expressed by Simpson’s index of dominance (c = 0.126) where dominance is concentrated in a small number of species (appendix 1). In other taxocenoses, dominance is spread to more co–dominant species (Simpson’s index of dominance values from 0.044 to 0.086). The value of equitability was highest at FU, NK and NA (table 2).

A dendrogram based on the qualitative representation (Sörensen’s index, complete linkage) separated the lepidoptera taxocenosis on the study plot HH (isolated forest, high density of ants, the lowest diversity of species) (fig. 2). Based on a qualitative–quantitative similarity (Wishart’s similarity ratio, complete linkage), the hierarchical classification divided the lepidoptera taxocenoses into two clusters connected on the relatively low level of similarity (fig. 3). The first cluster consisted of the taxocenoses HH and NA (island forests) with the lowest figures for abundance and individuals (44 and 133, respectively). The second cluster had two subclusters and included other taxocenoses. The first subcluster contained the taxocenoses from the denser and older plots (LL. Study plot affected by calcium dust deposition and with higher canopy cover of shrub story; LH. Lot with higher canopy cover of wood species crowns; and FU. Plot with higher canopy cover of both shrub story and wood species crowns). The second subcluster may be formed from the taxocenoses on lighter and younger plots (NK, LI, CA and VI) The NMDS showed plot HH was set apart from all the other study plots (fig. 4). The study plot NA was also separated (although less marked so) as confirmed by Wishart’s index.

Table 2. Species diversity test and basic characteristics of caterpillar taxocenoses at study plots in 2000–2002: H'. Shannon’s index of species diversity; e. Pielou’s index of equitability; c. Simpson’s index of dominance. (T–test values of H' are under the diagonal and degrees of freedom are above it; the testing process is detailed in Materials and methods; significance levels: *** P < 0.001; ** 0.001 < P < 0.01; * = 0.01 < P < 0.05; ns = 0.05 < P (non–significant); for abbreviations of the study plots see Material and methods). Tabla 2. Prueba de la diversidad de especies y características básicas de las taxocenosis de orugas en las parcelas del estudio entre los años 2000 y 2002. H'. Índice de Shannon para la diversidad de especies; e. Índice de Pielou para la equidad; c. Índice de Simpson para la dominancia. (Los valores de H' de la prueba t se encuentran debajo de la diagonal y los grados de libertad, encima; el proceso de la prueba se detalla en el apartado Material and methods; niveles de significación: *** P < 0,001; ** 0,001 < P < 0,01; * = 0,01 < P < 0,05; ns = 0,05 < P (no significativo); para consultar las abreviaturas de las parcelas del estudio, ver Material and methods).

0.851

0.801

0.872

0.809

0.867

0.838

0.849

0.867

0.868

0.066

0.086

0.063

0.126

0.067

0.063

0.066

0.044

3.097

3.065

3.101

3.212

2.505

3.091

3.194

3.197

3.428

3.1

541.083

349.481 636.538

55.534

445.853 356.73

228.02

670.81

3.07

0.343ns

421.247 582.174

65.467

495.143 431.18

294.16

477.82

3.1

0.048ns 0.346ns

396.331

64.992

372.919 350.76

265.84

289.7

3.21

1.356ns 1.495ns

1.127ns

59.702

491.438

402.8

259.76

603.69

2.51

3.549*** 3.218**

3.429** 4.162***

63.662

68.671

78.673

51.015

3.09

0.071ns 0.246ns

0.106ns 1.263ns

3.387**

386.06

272.78

381.14

3.19

0.99ns

0.842ns 0.182ns

3.901*** 0.957ns

284.73

299.12

192.36

2.201*

1.174ns

3.2

0.908ns 1.092ns

0.792ns 0.131ns

3.764*** 0.894ns 0.029ns

3.43

4.681*** 4.198***

3.784*** 2.773**

5.653**

4.013*** 2.567**

Kulfan et al.

0.8

Dissimilarity scale

0.7 0.6 0.5 0.4 0.3 0.2 0.1

0.0

Fig. 2. Classification of lepidoptera taxocenoses on individual study plots according to species presence/ absence (Sörensen’s index). Fig. 2. Clasificación de las taxocenosis de lepidópteros en cada una de las parcelas del estudio en función de la presencia o ausencia de las especies (índice de Sörensen).

Table 3 shows the overall result of the permutation test and pairwise ANOSIMs between all pairs of groups (provided as post–hoc test). Significant comparisons (at P < 0.05) are shown in bold. Analysis of similarity based on seasonal variability of species composition distinguished two significant

different lepidoptera taxocenoses. The lepidoptera taxocenosis of the HH had significantly lower abundance and number of species than the taxocenoses of the other eight study plots (table 4). Generally, lepidoptera larvae were weakly represented in HH because of the occurrence of numerous

1.0

Dissimilarity scale

0.9 0.8 0.7 0.6 0.5 0.4 0.3 0.2 0.1

0.0

Fig. 3. Classification of lepidoptera taxocenoses on individual plots according to qualitative–quantitative similarity (Wishart’s index). Fig. 3. Clasificación de las taxocenosis de lepidópteros en cada una de las parcelas del estudio en función de la similitud cualitativa y cuantitativa (índice de Wishart).

Animal Biodiversity and Conservation 36.1 (2013)

Stress: 0.073

0.16

0.12

NMDS_2

0.08

0.04 0

–0.04

–0.08 –0.12

–0.16 –0.2

–0.72 –0.6 –0.48 –0.36 –0.24 –0.12 0

0.12 0.24

NMDS_1 Fig. 4. Nonmetric multidimensional (NMDS) scaling plot based on Bray–Curtis similarities for species abundance data from nine study plots. (For abbreviations of the study plots see Material and methods.) Fig. 4. Gráfico del escalamiento multidimensional no métrico (NMDS) de las similitudes de Bray–Curtis a partir de los datos sobre la abundancia de las especies obtenidos en nueve parcelas del estudio (para consultar las abreviaturas de las parcelas del estudio, véase el apartado Material and methods).

colonies of ants as predators of lepidoptera larvae in this plot (appendix 1). The seasonal effect was reflected in all three examined parameters of the taxocenosis (table 4). The species abundance peaked in May, while number of species and species diversity reached the highest values from April to May and from April to June, respectively (fig. 5).

Figure 6 shows the number of species and the number of individuals in seasonal guilds. The number of species and the abundance showed two clear peaks in flush feeders and in late summer feeders. The number of flush feeder species was significantly higher than the number of species in other seasonal guilds (table 5).

Table 3. Results of analysis of similarity (ANOSIM): permutation number: 1,000; mean rank within: 5,047; mean rank between: 5,241; R = 0.037, P < 0.01. (For abbreviations of the study plots see Material and methods.) Tabla 3. Resultados del análisis de similitud (ANOSIM): número de permutaciones: 1.000; rango medio dentro: 5.047; rango medio entre: 5.241; R = 0,037; P < 0,01. (Para las abreviaturas de las parcelas del estudio, véase el apartado Material and methods).

–

0.16

0.39

0.73

0.15

0.23

0.02

0.06

0.18

0.58

0.67

0.77

0.46

0.01

0.68

0.21

–

0.68

0.15

0.32

0.01

0.73

0.17

0.38

0.62

0.02

0.83

0.33

–

0.48

0.01

0.18

0.22

–

0.3

–

0.09

0.11 0.3

–

Kulfan et al.

Table 4. Comparison of abundance, species number and species diversity in spatial and temporal scaling of studied lepidoptera taxocenoses in the hierarchical (nested) analysis of variance (ANOVA): * P < 0.05, ** P < 0.01; SSq. Sum of squares; MSq. Mean squares; SS. Sampling site; SD. Sampling date. Tabla 4. Comparación de la abundancia, el número de especies y la diversidad de especies en las escalas espacial y temporal de las taxocenosis estudiadas de lepidópteros en el análisis jerárquico (anidado) de la varianza (ANOVA): * P < 0,05, ** P < 0,01; SSq. Suma de los cuadrados; MSq. Media de los cuadrados; SS. Lugar de muestreo; SD. Fecha de muestreo. SSq

Abundance

MSq

F–statistic

13336.73

1667.09

4.29614 P < 0.01

61095.30

1131.39

3.51589 P < 0.01

Number of species taxa

Mann–Whitney pairwise comparison

P–value

HH < CA, VI, FU, LI, NK, NA, Ll* June–October < April–May*

268.798

33.600

3.0216

P < 0.01

HH < CA, VI, FU, LI, NK, NA, Ll*

4419.608

81.845

7.3602

P < 0.01

June–October < April–May*

Species diversity

6.50334

0.81292

1.2961 P = 0.2571

86.24383

1.59711

4.9800

P < 0.01

–– August–October < April–June**

The occurrence in time of lepidoptera species with two types of feeding specialization (generalists, narrow oligophagous) was non–randomly distributed throughout the season (table 6). The number of species in these feeding groups peaked in May. On the other hand, species in the wider oligophagous feeding groups exploited time in a random way. Discussion Taxocenoses of lepidoptera larvae observed on Quercus dalechampii in Malé Karpaty Mts. can be compared with taxocenoses on Q. cerris that were studied under similar conditions. A comparison shows similarities and differences (cf. Kulfan et al., 2006). Species richness was higher on Q. dalechampii (96 species on Q. dalechampii compared to 58 species on Q. cerris). The lowest number of species in both types of taxocenoses was found in the study plot HH. Lymantria dispar and Operophtera brumata belonged to the most abundant species both on Q. dalechampii and on Q. cerris. In the study plot HH, L. dispar reached a higher dominance on Q. cerris than on Q. dalechampii (cf. Kulfan et al., 2006). Regarding cumulative dominance, the families Ypsolophidae, Pyralidae and Drepanidae predominated on Q. cerris. On the contrary, the families Peleopodidae and Chimabachidae were noticeably more common on Q. dalechampii (cf. table 1, Kulfan et al., 2006). In general, when compared with other areas of Slovakia, the abundance of lepidoptera larvae found on Q. dalechampii corresponds to the latent phase of

the gradation cycle on oaks. No marked outbreaks of folivorous lepidoptera larvae have been observed on oaks in Slovakia since 1990 (cf. Kulfan, 1990, 1998, 2002; Kulfan, 1992; Kulfan et al., 1997, 2006). The values of species diversity of lepidoptera taxocenoses on Q. dalechampii are characterized by a greater variance than the values of diversity of taxocenoses on Q. petraea in the Malé Karpaty Mts. Diversity of lepidoptera taxocenoses on Q. petraea reached annual values ranging from 3.042 to 3.296 (Kulfan, 1990). The smallest diversity of lepidoptera taxocenoses on oaks in the Malé Karpaty Mts.was found on Q. cerris and it achieved a value of 2.230 (Kulfan et al., 2006) for the three–year period. As a rule, there is notable spring peak in abundance of lepidoptera caterpillars on oaks in Central Europe (Kulfan, 1992; Kulfan, 1983, 1990; Parák et al., 2012, etc.). This was also confirmed by the research on Q. dalechampii. Other peaks in caterpillar abundance on Q. dalechampii were not found throughout the growing season. Two peaks in the number of lepidoptera caterpillars during the season with prevalence in spring time were found on some oak species in Central Europe (Kulfan, 1992; Kulfan, 1983, 1990). The abundance of lepidoptera taxocenoses on Q. petraea throughout the growing season in the Malé Karpaty Mts in Slovakia has been found to have a very marked peak in spring (May–June) and a less noticeable peak in autumn (September). However, it is interesting that the autumn peak of total abundance of lepidoptera caterpillars on Q. petraea in the old oak stand in 1978 was more noticeable when compared to the spring peak (cf. Kulfan, 1983).

Animal Biodiversity and Conservation 36.1 (2013)

Diversity

0.8 0.6 0.4 0.2 0 iv

vii

viii

Number of individuals

200 160 120 80 40 0

Number of species

vii

viii

40 39 20 10 0 iv

vii

viii

Fig. 5. Box plots showing the monthly variation of diversity, abundance and number of species of lepidoptera taxocenosis: IV–X. Months of presence of lepidopteran caterpillars during the season. Fig. 5. Diagramas de caja en los que se muestra la variación mensual de la diversidad, la abundancia y el número de especies de la taxocenosis de los lepidópteros: IV–X. Meses de presencia de las orugas de lepidópteros durante la estación.

This was probably caused by unfavorable weather conditions in spring. Yoshida (1985) in northern Japan presented the highest abundance of lepidoptera caterpillars on oaks in summer. This difference compared to our results may be caused by different climate, because the frosts in May in northern Japan have a large impact on leaf phenology, which is associated with the development of the spring taxocenoses of caterpillars.

In oak forest on Mont Holomontas (Mediterranean area, Greece) even three peaks in insect abundance (consisting mainly of lepidoptera larvae) on six oak species were found (Kalapanida & Petrakis, 2012). Not only the species abundance but also the species richness and diversity of lepidoptera species on Q. dalechampii culminated in the vernal aspect. The marked increase of species diversity of lepidoptera taxocenoses on Q. dalechampii was in spring. A similar

Kulfan et al.

Number of species

50 40 30 20 10

Number of individuals

FIF

LSF

FIF

LSF

500 400 300 200 100 0

Fig. 6. Box plots showing the effects of seasonality on species number and abundance of lepidoptera taxocenosis: FIF. Flush feeders; SF. Summer feeders; LSF. Late summer feeders; AF. Autumn feeders. Fig. 6. Diagramas de caja en los que se muestra los efectos de la estacionalidad en el número de especies y en la abundancia de la taxocenosis de los lepidópteros: FIF. Se alimentan durante la brotación; S. Se alimentan en verano; LSF. Se alimentan al final del verano; AF. Se alimentan en otoño.

Table 5. Results of one–way ANOVA on differences between seasonal guilds in the number of species taxa and number of individuals. The post hoc multiple sample comparison test (Tukey's pairwise comparison) for differences in mean number of species taxa and number of individuals between seasonal guilds: * P < 0.05, ** P < 0.01; FIF. Flush feeders; LSF. Late spring feeders; SF. Summer feeders; AF. Autumn feeders. Tabla 5. Resultados de la ANOVA simple de las diferencias existentes entre los gremios estacionales en cuanto el número de taxones y el número de individuos. La prueba múltiple de comparación a posteriori de Tukey (comparación por pares de Tukey) de las diferencias en el número medio de taxones y el número de individuos entre gremios estacionales: * P < 0,05; ** P < 0,01; FIF. Se alimentan durante la brotación; LSF. Se alimentan al final de la primavera; SF. Se alimentan en verano; AF. Se alimentan en otoño.

F–statistic df P–value

Tukey's pairwise comparison

Number of species in seasonal guilds

85.49

P < 0.01

SF, LSF, AF < FIF**

Number of individuals in seasonal guilds

25.44

P < 0.01

SF, LSF, AF < FIF**

SF, AF < LSF**

Animal Biodiversity and Conservation 36.1 (2013)

Table 6. Results obtained from the Poole–Rathcke method used to segregate the moths in time. The null hypothesis (H1) states that the dispersion is not significantly different from random and the second null hypothesis (H2) that the two means and dispersions are not significantly different: N. Number of species; OV. Observed variance; EV. Expected variance; DR. Dispersion ratio; RD. Random dispersion (significance of H1); HM. Homogeneity of means; HD. Homogeneity of dispersion (significance of H2). (* P < 0.05, ** P < 0.01, ns. Non–significant) Tabla 6. Resultados obtenidos con el método de Poole–Rathcke empleado para segregar las polillas en el tiempo. La hipótesis nula (H1) afirma que la dispersión no es significativamente distinta de la aleatoria y la segunda hipótesis nula (H2), que las dos medias y las dos dispersiones no son significativamente diferentes: N. Número de especies; OV. Varianza observada; EV. Varianza esperada; DR. Razón de la dispersión; RD. Dispersión aleatoria (significación de H1); HM. Homogeneidad de las medias; HD. Homogeneidad de la dispersión (significación de H2). (* P < 0,05; ** P < 0,01; ns. No significativa).

Feeding specialization Generalists

9.87

0.19

51.94736842

Narrow oligophagous

18.7

0.27

69.25925926

Wider oligophagous

15.33

0.82

18.69512195

Feeding specialization compared Generalist / narrow oligophagous

Q = 8.54 P < 0.01

W = 2.14 P = 0.15

Generalist / wider oligophagous

Q = 8.74 P < 0.01

W = 13.55 P < 0.01

Narrow oligophagous / wider oligophagous

Q = 0.19 P = 0.98

W = 14.92 P < 0.01

course of diversity was observed on four oak species in the Borská nížina Lowland in Slovakia. (Kulfan & Degma, 1999). Southwood et al. (2005) found distinct seasonal patterns in species richness of the arthropod fauna on four oak species in the U.K. In terms of species richness, the values showed a general trend peaking in summer and early autumn, but biomass peaked in May on the native oak species, mainly due to lepidoptera larvae. A relatively steady decrease in the individuals from early spring to autumn is well known from the 'Quercus type' of host tree (Niemelä & Haukioja, 1982). These authors suggested that this effect was due to a decline in available resources. Feeny (1970) and Kamata & Igarashi (1996) stated that tougher leaves with a higher tannin concentration contributed to the lower richness of Lepidoptera later during the growing season. A negative correlation between some specialist oak feeders and condensed tannins in the canopy of Quercus alba and understorey of Q. velutina was found (Forkner et al., 2004). Their results generally indicated a negative response from both specialists and generalists to condensed tannins. A higher number of flush feeders in spring compared to the number of species in other seasonal guilds on Q. dalechampii was also found on three oak

species in Borská nížina lowland (western Slovakia, Central Europe), where the greatest proportion of flush feeders was found on Q. robur (cf. Turčáni et al., 2009). Acknowledgements This work was supported by the Slovak Grant Agency for Science (Grant Nos 1/0124/09, 1/0137/11, 2/0035/13 and 1/0066/13). We are grateful to Jaroslav Fajčík for technical assistance. References Brown, V. K. & Hyman, P. S., 1986. Successional communities of plants and phytophagous Coleoptera. Journal of Ecology, 7: 963–975. Csóka, G., 1990–1991. Oak hostplant records of Hungarian Macrolepidoptera. Erdészeti Kutatások, 82–83: 89–93. – 1998a. Oak Defoliating Insects in Hungary. In: Proceedings: Population Dynamics, Impacts, and Integrated Management of Forest Defoliating Insects: 334–335 (M. L. McManus & A. M. Liebhold, Eds.). USDA Forest Service General Technical

Report NE–247, Hungary. – 1998b. A Magyarországon honos tölgyek herbivir rovaregyüttese. Erdészeti Kutatások, 88: 311–318. Feeny, P., 1970. Seasonal changes in oak leaf tannins and nutrient as a cause of spring feeding by winter moth caterpillars. Ecology, 51: 565–581. Forkner, R. E., Marquis, R. J. & Lill, J. T., 2004. Feeny revisited: condensed tannins as anti–herbivore defences in leaf–chewing herbivore communities of Quercus. Ecological Entomology, 29: 174–187. Gerasimov, A. M., 1952. Insects. Lepidoptera 1(2). Caterpillars 1. Fauna of Soviet Union 56. Zoological Institute, Academy of Sciences USSR, Moscow, Leningrad. Hrubý, K., 1964. Prodromus lepidopter Slovenska. Vydavateľstvo SAV, Bratislava. Kalapanida, M. & Petrakis, V., 2012. Temporal partitioning in an assemblage of insect defoliators feeding on oak on a Mediterranean mountain. European Journal of Entomology, 109: 55–69. Kamata, N. & Igarashi, I., 1996. Seasonal and annual change of a folivorous insect guild in the Siebolds beech forests associated with the outbreaks of the beech caterpillar, Quadricalcarifera punctatella (Motschulsky) (Lep., Notodontidae). Journal of Applied Entomology, 120: 213–220. Kollár, J., 2007. The harmful entomofauna of woody plants in Slovakia. Acta entomologica serbica, 12: 67–79. Kulfan, J., 1992. Zur Struktur und Saisondynamik von Raupenzönosen (Lepidoptera) an Eichen. Biológia, Bratislava, 47: 653–661. Kulfan, M., 1983. The seasonal dynamics of lepidoptera caterpillar communities in the Malé Karpaty mountains. Biologia, Bratislava, 38: 1003–1009. – 1990. Communities of Lepidoptera caterpillars (Lepidoptera) on broadleaf tree species of Malé Karpaty. Veda, Vydavateľstvo SAV, Bratislava. – 1997. Lepidoptera living on oaks in Southwestern Slovakia lowlands. Folia faunistica Slovaca, 2: 85–92. – 1998. Lepidoptera as pests of oak crowns in Southwestern Slovakia lowlands. Folia faunistica Slovaca, 3: 119–124. – 2002. Lepidoptera larvae communities on Quercus robur and Cerasus mahaleb in NNR Devínska Kobyla (SW Slovakia). Folia Faunistica Slovaca, 7: 55–60. – 2012. Structure of lepidopterocenoses on oaks Quercus dalechampii and Q. cerris in cenral Europe and estimation of the most important species. Munis Entomology & Zoology, 7: 732–741. Kulfan, M. & Degma, P., 1999. Seasonal dynamics of lepiropteran larvae communities diversity and equitability on oaks in the Borská nížina lowland. Ekologia, 18: 100–105. Kulfan, M., Holecová, M. & Fajčík, J., 2006. Caterpillar (Lepidoptera) communities on European Turkey oak (Quercus cerris) in Malé Karpaty Mts (SW Slovakia). Biologia, Bratislava, 61: 573–578. Kulfan, M., Šepták, L. & Degma, P., 1997. Lepidoptera larvae communities on oaks in SW Slovakia.

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Biologia, Bratislava, 52: 247–252. Laštůvka, Z. & Liška, J., 2011. Annotated checklist of moths and butterflies of the Czech republic (Insecta: Lepidoptera). Biocont Laboratory, Brno. Ludwig, J. A. & Reynolds, J., 1988. Statistical ecology: a primer of methods and computing. Whiley – Interscience Public, New York. Niemelä, P. & Haukioja, E., 1982. Seasonal patterns in species richness of herbivores: Macrolepidoptera larvae on Finnish deciduous trees. Ecological Entomology, 7: 169–175. Novotny, V., Drozd, P., Miller, S. E., Kulfan, M., Janda, M., Basset, Y. & Weiblen, G. D., 2006. Why are there so many species of herbivorous insects in tropical rainforests? Science, 313: 1115–1118. Parák, M., Kulfan, J. & Svitok, M., 2012. Spoločenstvá húseníc (Lepidoptera) na troch druhoch dubov (Quercus spp.) z oblasti Čachtických Karpát (západné Slovensko). Folia faunistica Slovaca, 17: 247–256. Patočka, J., 1954. Caterpillars on oak trees in Czechoslovakia. Štátne pôdohospodárske nakladateľstvo, Bratislava. – 1980. Die Raupen und Puppen der Eichenschmetterlinge Mitteleuropas. Paul Parey Verlag, Germany. Patočka, J., Čapek, M. & Charvát, K., 1962. Contribution to the knowledge of the crown arthropod fauna on oaks of Slovakia, in particular with regard to the order Lepidoptera. Veda, Vydavateľstvo SAV, Bratislava. Patočka, J., Krištín, A., Kulfan, J. & Zach, P., 1999. Die Eichenschädlinge und ihre Feinde. Ústav ekológie lesa SAV, Zvolen. Podani, J., 1993. Syn–tax. Version 5.0. Computer programs for Multivariate Data Analysis in Ecology and Systematics. User’s guide. Scientia Publishing, Budapest. Poole, R. W., 1974. An Introduction to Quantitative Ecology. McGraw–Hill, New York. Poole, R. W. & Rathcke, B.–J., 1979. Regularity, randomness and aggregation in flowering phonologies. Science, 203: 470–471. Reiprich, A., 2001. Die Klassifikation der Schmetterlinge der Slowakei laut den Wirten (Nährpflanzen) ihrer Raupen. Správa Národného parku Slovenský raj vo vydavateľstve SZOPK, Spišská Nová Ves. Skuhravý, V., Hrubík, P., Skuhravá, M. & Požgaj, J., 1998. Occurrence of insect associated with nine Quercus species (Fagaceae) in cultured plantations in southern Slovakia during 1987–1992. Journal of Applied Entomology, 122: 149–155. Southwood, T. R. E., Wint, G. R. W., Kennedy, C. E. J. & Greenwood, S. R., 2005. The composition of the arthropod fauna of the canopies of some species of oak (Quercus). European Journal of Entomology, 102: 65–72. Stolnicu, A. M., 2007. Leaf–mining insects encountered in the forest reserve of Hârboanca, Vaslui county. Analele Ştiinţifice ale Universităţii, Al. I. Cuza” Iaşi, s. Biologie animală, 53: 109–114. Turčáni, M., Patočka, J. & Kulfan, M., 2009. How

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do lepidopteran seasonal guilds differ on some oaks? – A case study. Journal of Forest Science, 55: 578–590. – 2010. Which factors explain lepidopteran larvae variance in seasonal guilds on some oaks? Journal of Forest Science, 56: 68–76. Yoshida, K., 1985. Seasonal population trends of

macrolepidopterous on oak trees in Hokaido, northern Japan. The Entomological Society of Japan, 53: 125–133. Zlinská, J., Šomšák, L. & Holecová, M., 2005. Ecological characteristic of studied forest communities of an oak–hornbeam tier in SW Slovakia. Ekológia (Bratislava), 24 (Suppl. 2): 3–19.

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Appendix 1. The list of the lepidoptera species recorded in the nine study plots in the Malé Karpaty Mountains on Quercus dalechampii with dominance (%), months of occurrence of larvae (MO), trophic group (TG: S2. Narrow oligophagous; S3. Wider oligophagous species; G. Generalists; U. Unknown), and larval trophic specialization and seasonal guilds (SG: FIF. Flush feeders; LSF. Late spring feeders; SF. Summer feeders; AF. Autumn feeders). (For abbreviations of study plots see Material and methods.)

Families and species

Psychidae Sterrhopterix fusca (Haworth, 1809)

0.0

0.3

Bucculatricidae Bucculatrix ulmella Zeller, 1848

0.0

Gracillariidae Phyllonorycter sp.

0.0

0.6

0.0

Ypsolophidae Ypsolopha alpella (Denis et Schiffermüller, 1775)

1.8

0.0

0.9

Ypsolopha parenthesella (Linnaeus, 1761)

0.0

Ypsolopha ustella (Clerck, 1759)

0.3

1.1

0.0

0.3

Chimabachidae Diurnea fagella (Denis et Schiffermüller, 1775)

0.0

0.7

0.6

0.0

Diurnea lipsiella (Denis et Schiffermüller, 1775)

0.6

2.8

1.2

0.6

Peleopodidae Carcina quercana (Fabricius, 1775)

0.6

2.8

3.0

4.6

Coleophoridae Coleophora ibipennella Zeller, 1849

0.0

0.3

Coleophora kuehnella (Goeze, 1783)

0.0

Coleophora lutipennella (Zeller, 1838)

0.6

4.2

6.6

4.6

Coleophora siccifolia Stainton, 1856

0.0

0.3

Gelechiidae Anacampsis timidella (Wocke, 1887)

0.0

0.6

0.0

Carpatolechia decorella ( Haworth, 1812)

0.0

0.4

0.0

Psoricoptera gibbosella (Zeller, 1839)

0.0

0.3

Stenolechia gemmella (Linnaeus, 1758)

0.0

0.4

0.0

Tortricidae Aleimma loeflingiana (Linnaeus, 1758)

7.7

0.0

12.0

3.4

Archips crataegana (Hübner, 1799)

0.6

0.0

0.3

Archips podana (Scopoli, 1763)

0.0

Eudemis profundana (Denis et Schiffermüller, 1775)

0.0

Pammene albuginana (Guenée, 1845)

0.0

Pandemis cerasana (Hübner, 1786)

0.0

0.7

2.4

0.0

Pandemis corylana (Fabricius, 1794)

0.0

0.4

0.0

0.3

Pandemis heparana (Denis et Schiffermüller, 1775)

0.0

0.4

0.0

0.3

Ptycholoma lecheana (Linnaeus, 1758)

0.0

0.3

Spilonota ocellana (Denis et Schiffermüller, 1775)

0.0

2.1

0.6

0.0

Tortricodes alternella (Denis et Schiffermüller, 1775)

1.8

1.4

3.6

0.9

Tortrix viridana (Linnaeus, 1758)

4.9

0.7

1.2

0.3

Zeiraphera isertana (Fabricius, 1794)

1.8

0.4

1.8

0.3

Animal Biodiversity and Conservation 36.1 (2013)

Apéndice 1. Lista de las especies de lepidópteros registradas en Q. dalechampii en las nueve parcelas del estudio ubicadas en los Pequeños Cárpatos con la dominancia (%), los meses de presencia de las larvas (MO), el grupo trófico (TG: S2. Oligófagas estrictas; S3. Especies oligófagas más amplias; G. Generalistas; U. Desconocido) y la especialización trófica de las larvas y los gremios estacionales (SG: FIF. Se alimentan durante la brotación; LSF. Se alimentan al final de la primavera; SF. Se alimentan en verano; AF. Se alimentan en otoño). (Para las abreviaturas de las parcelas del estudio, véase el apartado Material and methods).

0.0

FIF

0.4

LSF

0.0

1.5

0.0

2.3

0.0

1.1

3.0

2.0

5–6

FIF

0.0

1.1

0.0

FIF

2.3

0.9

1.1

0.8

0.4

5–6

FIF

4.5

0.5

0.0

3.8

0.4

6–9

0.0

0.5

1.1

0.8

3.9

5–8

FIF

5–8

LSF

FIF

6.8

3.2

1.6

2.3

3.5

0.0

0.2

FIF

0.0

6.0

2.7

3.8

5.0

4–6

FIF

0.0

5.2

4–5

FIF

0.0

FIF

0.0

FIF

2.3

0.0

0.5

0.0

FIF

0.0

1.1

0.0

5–6

FIF

0.0

4–5

FIF

0.0

9.7

5.4

10.5

0.0

1.1

0.0

0.4

5–6

FIF

0.0

0.5

0.0

FIF

0.0

0.5

0.0

FIF

0.0

0.2

FIF

0.0

3.8

1.5

0.7

5–7

FIF

0.0

0.5

0.0

FIF

0.0

1.4

0.5

0.0

0.7

5–6

FIF

0.0

0.8

0.2

FIF

0.0

FIF

4.5

1.4

1.1

1.5

0.9

FIF

0.0

9.3

0.0

2.3

1.3

4–5

FIF

0.0

1.1

0.0

0.9

4–5

FIF

Kulfan et al.

Appendix 1. (Cont.)

Families and species

Lycaenidae Favonius quercus (Linnaeus, 1758)

0.3

0.0

1.2

0.3

Pyralidae Acrobasis repandana (Fabricius, 1798)

0.0

Acrobasis tumidana (Denis et Schiffermüller, 1775)

0.0

Phycita roborella (Denis et Schiffermüller, 1775)

0.9

0.4

0.0

0.9

Drepanidae Watsonalla binaria (Hufnagel, 1767)

0.0

0.4

0.6

0.0

Geometridae Agriopis aurantiaria (Hübner, 1799)

2.5

0.7

0.6

Agriopis leucophaearia (Denis et Schiffermüller, 1775)

3.1

1.4

1.8

6.8

Agriopis marginaria (Fabricius, 1776)

3.4

4.2

3.6

8.0

Alcis repandata (Linnaeus, 1758)

0.0

Alsophila aceraria (Denis et Schiffermüller, 1775)

0.3

0.4

0.0

1.5

Alsophila aescularia (Denis et Schiffermüller, 1775)

4.9

2.1

0.6

1.8

Apocheima hispidaria (Denis et Schiffermüller, 1775)

0.0

Biston betularia (Linnaeus, 1758)

0.0

0.4

1.2

0.0

Biston strataria (Hufnagel, 1767)

0.0

Campaea margaritaria (Linnaeus, 1761)

1.8

0.0

0.3

Colotois pennaria (Linnaeus, 1761)

1.5

0.4

0.0

1.8

Cyclophora linearia (Hübner, 1799)

0.3

3.2

8.4

2.5

Cyclophora punctaria (Linnaeus, 1758)

0.0

0.4

0.0

Ennomos autumnaria (Werneburg, 1859)

0.0

0.4

0.0

Ennomos erosaria (Denis et Schiffermüller, 1775)

0.9

0.0

0.6

0.3

Ennomos quercinaria (Hufnagel, 1767)

0.0

0.4

0.0

Epirrita dilutata (Denis et Schiffermüller, 1775)

1.2

7.7

1.2

0.9

Erannis defoliaria (Clerck, 1759)

1.8

2.1

0.0

0.6

Hypomecis punctinalis (Scopoli, 1763)

0.6

0.0

0.3

Lomographa temerata (Denis et Schiffermüller, 1775)

0.0

1.1

0.0

0.6

Lycia hirtaria (Clerck, 1759)

0.0

0.3

Operophtera brumata (Linnaeus, 1758)

11.4

22.9

10.2

8.0

Parectropis similaria (Hufnagel, 1767)

0.0

Peribatodes rhomboidaria (Denis et Schiffermüller, 1775)

0.0

Phigalia pilosaria (Denis et Schiffermüller, 1775)

0.0

Selenia lunularia (Hübner, 1788)

0.0

0.3

Selenia tetralunaria (Hufnagel, 1767)

0.0

Notodontidae Drymonia ruficornis (Hufnagel, 1766)

0.3

0.0

2.4

0.3

Phalera bucephala (Linnaeus, 1758)

0.0

Spatalia argentina (Denis et Schiffermüller, 1775)

0.0

Thaumetopoea processionea (Linnaeus, 1758)

0.0

0.6

Animal Biodiversity and Conservation 36.1 (2013)

0.0

0.5

0.0

0.6

FIF

0.0

0.8

0.0

FIF

0.0

2.3

1.6

0.0

0.7

FIF

0.0

0.5

1.1

0.0

0.9

4–5, 9

FIF

0.0

6, 8, 10

AF FIF

0.0

0.9

0.0

0.8

0.0

3.7

2.2

0.8

1.3

4–6

2.3

7.9

7.1

1.5

8.5

4–5

FIF

2.3

0.9

7.6

1.5

5.7

4–5

FIF

0.0

0.5

0.0

2.8

0.5

0.8

0.7

4–5

FIF

0.0

5.6

5.4

3.0

3.1

4–6

FIF

0.0

0.5

0.0

FIF

2.3

0.9

0.0

8–10

0.0

0.5

0.0

FIF

0.0

1.4

2.2

3.8

1.3

4–9, 11

LSF

0.0

1.6

1.5

3.3

4–5

FIF

11.4

3.2

3.3

8.3

2.8

6–10

LSF

0.0

0.5

0.0

6–7

LSF

0.0

0.9

0.0

1.5

0.0

5–7

LSF

0.0

5,9

FIF

0.0

0.5

0.0

5–6

FIF

2.3

0.5

2.3

1.5

4–5

FIF

0.0

0.5

0.0

0.7

4–5

FIF

0.0

1.3

6–9

0.0

0.4

7–8

0.0

0.4

FIF

4.5

15.7

15.8

5.3

9.2

4–5

FIF

0.0

0.9

0.0

7, 10

0.0

0.8

0.0

0.2

FIF

0.0

LSF

0.0

0.5

0.0

LSF

0.0

0.9

FIF

4.5

0.0

0.5

0.0

0.2

6–7

0.0

4.4

5–6

FIF

Kulfan et al.

Appendix 1. (Cont.)

Families and species

Erebidae Amata phegea (Linnaeus, 1758)

0.0

0.4

0.0

Calliteara pudibunda (Linnaeus, 1758)

1.2

0.0

0.9

Lymantria dispar (Linnaeus, 1758)

16.3

1.8

3.0

11.4

Orgyia antiqua (Linnaeus, 1758)

0.0

Nolidae Bena bicolorana (Fuessly, 1775)

0.3

0.4

0.0

0.3

Nycteola revayana (Scopoli, 1772)

0.0

0.6

0.0

Pseudoips prasinana (Linnaeus, 1758)

1.2

4.2

3.0

2.8

Noctuidae Acronicta auricoma (Denis et Schiffermüller, 1775)

0.0

0.7

0.0

0.3

Agrochola helvola (Linnaeus, 1758)

0.0

Amphipyra pyramidea (Linnaeus, 1758)

0.0

0.6

1.5

Colocasia coryli (Linnaeus, 1758)

0.0

0.4

0.0

Cosmia pyralina (Denis et Schiffermüller, 1775)

5.2

2.5

0.0

Cosmia trapezina (Linnaeus, 1758)

5.5

12.0

14.5

Dichonia convergens (Denis et Schiffermüller, 1775)

2.8

0.7

1.8

0.6

Dryobotodes eremita (Fabricius, 1775)

0.0

4.2

0.0

Eupsilia transversa (Hufnagel, 1766)

0.6

2.8

1.2

0.3

Lithophane ornitopus (Hufnagel 1766)

0.6

2.1

3.0

0.9

Moma alpium (Osbeck, 1778)

0.0

Noctuidae species 1

0.0

Noctuidae species 2

0.0

Noctuidae species 3

0.0

Noctuidae species 4

0.0

0.9

Noctuidae species 5

0.0

0.3

Noctuidae species 6

0.0

Noctuidae species 7

2.2

0.0

Orthosia cerasi (Fabricius, 1775)

5.5

2.1

0.0

3.4

Orthosia cruda (Denis et Schiffermüller, 1775)

2.2

1.1

1.8

2.5

Orthosia gothica (Linnaeus, 1758)

0.0

2.1

0.0

0.6

Orthosia incerta (Hufnagel, 1776)

0.0

0.4

0.0

Orthosia opima (Hübner, 1809)

0.0

2.4

3.4

No individuals

325

284

167

325

No species / taxons

Animal Biodiversity and Conservation 36.1 (2013)

4–5

2.3

0.0

FIF

0.0

0.8

0.0

6–8

29.5

0.0

2.2

18.0

3.5

4–7

FIF

0.0

0.2

LSF

0.0

0.8

0.0

4, 8

FIF

0.0

FIF

6.8

0.9

1.1

1.5

1.1

6–10

LSF

0.0

5–6

LSF

0.0

0.5

0.0

FIF

0.0

0.5

0.0

0.2

FIF

0.0

0.5

0.0

0.8

0.0

6, 8

LSF

4.5

0.5

0.0

1.5

4–5

FIF

0.0

5.6

9.8

0.0

5.9

4–5

FIF

0.0

0.5

0.0

3.0

1.3

4–5

FIF

0.0

0.5

0.0

0.8

0.0

FIF

4.5

2.3

0.5

3.0

0.7

4–6

FIF

0.0

0.9

0.5

2.3

0.2

4–6

FIF

0.0

0.5

0.0

0.8

0.4

7–8

0.0

0.8

0.0

FIF

0.0

0.8

0.0

FIF

0.0

1.5

0.0

FIF

0.0

FIF

0.0

FIF

0.0

1.6

0.0

FIF

0.0

3.2

6.5

0.0

1.1

4–7

FIF

0.0

1.4

0.0

7.0

4–6

FIF

0.0

0.8

0.4

5–6

FIF

0.0

FIF

4–6

FIF

0.0

0.5

0.0

2.8

216

184

133

462

Arizaga et al.

Animal Biodiversity and Conservation 36.1 (2013)

33 Forum

Why do ecologists aim to get positive results? Once again, negative results are necessary for better knowledge accumulation A. Martínez–Abraín

Martínez–Abraín, A., 2013. Why do ecologists aim to get positive results? Once again, negative results are necessary for better knowledge accumulation. Animal Biodiversity and Conservation, 36.1: 33–36. Abstract Why do ecologists aim to get positive results? Once again, negative results are necessary for better knowledge accumulation.— Hypothesis testing is commonly used in ecology and conservation biology as a tool to test statistical–population parameter properties against null hypotheses. This tool was first invented by lab biologists and statisticians to deal with experimental data for which the magnitude of biologically relevant effects was known beforehand. The latter often makes the use of this tool inadequate in ecology because we field ecologists usually deal with observational data and seldom know the magnitude of biologically relevant effects. This precludes us from using hypothesis testing in the correct way, which is posing informed null hypotheses and making use of a priori power tests to calculate necessary sample sizes, and it forces us to use null hypotheses of equality to zero effects which are of little use for field ecologists because we know beforehand that zero effects do not exist in nature. This is why only 'positive' (statistically significant) results are sought by ecologists, because negative results always derive from a lack of power to detect small (usually biologically irrelevant) effects. Despite this, 'negative' results should be published, as they are important within the context of meta–analysis (which accounts for uncertainty when weighting individual studies by sample size) to allow proper decision–making. The use of multiple hypothesis testing and Bayesian statistics puts an end to this black or white dichotomy and moves us towards a more realistic continuum of grey tones. Key words: Power test, Negative results, Effect size, Positive results, Observational data, Null hypothesis testing. Resumen ¿Por qué los ecólogos desean obtener resultados positivos? Una vez más, los resultados negativos son necesarios para mejorar la acumulación de conocimiento.— El contraste de hipótesis se emplea habitualmente en ecología y biología de la conservación como una herramienta para contrastar los valores de los parámetros de poblaciones estadísticas con las hipótesis nulas. Esta herramienta fue inventada por biólogos de laboratorio y estadísticos para tratar datos experimentales para los que se conocía de antemano la magnitud de los efectos biológicamente relevantes. Esto hace que a menudo en ecología no sea adecuado utilizar esta herramienta porque los ecólogos de campo generalmente trabajamos con datos observacionales y rara vez conocemos la magnitud de los efectos que son biológicamente relevantes. Ello nos impide usar el contraste de hipótesis adecuadamente, es decir, plantear hipótesis nulas que contengan información y emplear pruebas de potencia a priori para calcular los tamaños de muestra necesarios, y nos fuerza a emplear hipótesis nulas de efectos iguales a cero que son de poca utilidad para los ecólogos de campo porque sabemos por adelantado que en la naturaleza los efectos siempre son distintos de cero. Por esto los ecólogos siempre desean encontrar resultados positivos (estadísticamente significativos), porque los negativos siempre proceden de una falta de potencia para detectar efectos pequeños, que por lo general son biológicamente irrelevantes. A pesar de ello, los resultados negativos deberían publicarse porque son importantes en el contexto de los metanálisis (que analizan la incertidumbre al ponderar distintos estudios en función del tamaño de muestra) para permitir una adecuada toma de decisiones. El uso del contraste múltiple de hipótesis y la estadística bayesiana acaba con esta dicotomía, y nos sitúa en un contexto más realista en el que existe una escala de grises. Palabras claves: Pruebas de potencia, Resultados negativos, Magnitud del efecto, Resultados positivos, Datos observacionales, Contraste de hipótesis nulas. ISSN: 1578–665X

Martínez–Abraín

Received: 21 IX 12; Conditional acceptance: 13 XII 12; Final acceptance: 20 XII 12 Alejandro Martínez–Abraín, Dept. de Bioloxía Animal, Bioloxía Vexetal e Ecoloxía, Univ. da Coruña, Campus da Zapateira s/n., 15071 A Coruña, España (Spain); IMEDEA (CSIC–UIB), Population Ecology Group, Miquel Marquès 21, 07190 Esporles, Mallorca, España (Spain). E–mail: a.abrain@imedea.uib–csic.es

It is well known that citation rates of ecological papers are affected by the direction of the study outcome with respect to the hypothesis tested, with supportive papers being more frequently cited than unsupportive papers (Leimu & Koricheva, 2005). In part because of this, it is difficult to publish a result of our research which turns out to be negative (defining negative as ‘statistically non–significant’) (Dickersin et al., 1992), or at least to publish it in a journal with a high impact factor (Koricheva, 2003). But there is more than this behind our reluctance to publish and cite negative results. Why do we aim to get positive results? Frequentist inferential statistics is a tool developed by and for laboratory people (notably Fisher, Neyman and Pearson) in the 1930s. It was created as a forced alternative to already existing Bayesian statistics because at that time it was difficult to solve the integrals needed to estimate the denominator of Bayes’ formula (i.e. the probability of obtaining our data). Lab people —including all sorts of experimentalists in the life sciences and other sciences— have a huge advantage over field ecologists. Before starting an experiment, they often know which magnitude of effect is biologically relevant for them. On the contrary, we ecologists most often do not. I like to call this the 'Gordian knot' of ecological statistics. Hence we have borrowed an analytical framework that is not particularly appropriate for us, most often owners of observational rather than experimental data. When we know the magnitude of an effect that is of interest for our question we can use hypothesis–testing correctly, by making use of a priori power tests. These tests allow us to calculate the sample size required to obtain a statistically significant result only for a biologically relevant magnitude of the effect. For example, when comparing the length of the wings of two migratory moth populations to evaluate their potential as migrants we would only say that differences are statistically significant (that is, there is little or no overlap between the 95% confidence intervals of the point estimates of wing length of both populations) when they differ in at least 'x' millimetres if we knew beforehand that only beyond that difference level (i.e. effect size) a relevant biological phenomenon occurs.

In this case, our null hypothesis would not be equal to zero but equal to 'x'. But the problem is that we seldom (not to say 'never') know the magnitude of interest of our differences in ecology. And hence we end up testing uninformative null hypotheses (of equality to zero effects) and thus do not make use of informed a priori power tests. Hence it is not that ecologists necessarily make poor use of hypothesis testing; it is that we cannot do better with a tool that does not belong to us, as if we were trying to paint a wall with a brush designed to paint a canvas. The drawback of not being able to use a priori power tests to obtain the required size of samples to test for a biologically meaningful effect is that we use hypothesis testing blindly. We commonly collect data with a fairly small sample size (e.g. n = 30) and hope to reach conclusions. We reason, correctly, that if we are able to obtain positive ('statistically significant') results with such a small sample size, we can be quite confident that we have found a large effect, most likely a biologically relevant one. This is because small effects require a large sample size to be detected. Hence, it is also true that if we are using a large sample size to reach statistical significance we will certainly be able to do so even for tiny effects, which will often be biologically irrelevant. But this situation of getting into trouble for having 'too much' sample size is much less common in ecology (although it can also happen when pooling large data sets), except perhaps for theoretical ecologists, who can make use of huge sample sizes when using simulated data. If, on the contrary, without using a priori power tests, we find a negative result —that is, we have a P–value higher than the a priori agreed alpha risk of being wrong— we can only say that we have had a lack–of–power problem, something that both authors and journal editors dislike. This is so because by increasing the sample size we would always obtain a statistically significant result in the end, when our null hypotheses are of equality to zero effects, because there are always some effects or differences in nature due to natural variability between individuals and populations. Two populations may differ in a tiny amount only, but they do differ in some of their decimal points (Martínez–Abraín, 2007). If we have not found that difference it is just because of a low ability of our 'magnifying lens' (determined by our sample size) to do so. That seemingly makes our negative results not appealing for publication.

Animal Biodiversity and Conservation 36.1 (2013)

Negative results, however, are informative if we pay attention to sample size. A negative result obtained using a small sample size probably means that the effect we are studying is medium or small but not large. A negative result associated with a large sample size necessarily means that our biological effect is tiny. Negative results are indeed most informative when we have previously used a priori power tests but, I insist, it is not the ecologist’s fault not to be able to do so. It is something inherent to our observational science not to know beforehand in most cases the magnitude of an effect that is relevant for our questions. The use of alternative methods of data analysis, such as the simultaneous testing of multiple (sensical) hypotheses with selection of the most parsimonious models (representing hypotheses in mathematical format) by means of numerical criteria based on the loss of theoretical Kullback–Leibler information (Burnham & Anderson, 2002; Anderson, 2008), is a step forward that is increasingly gaining relevance in ecology. Obviously, this is a much better approach than testing a null hypothesis containing nil information against a unique alternative hypothesis which points by force in opposite direction of the null and for which we present no evidence whatsoever. Importantly, the use of Bayesian statistics may also help us to put an end to this old debate because we obtain the posterior distribution of the population parameter (given our data), thanks to combining the data–derived likelihood of the parameter with prior information on the parameter (or with a flat prior distribution if previous information is not available) by means of Bayes’ rule; the advantage, anyway, is that we can interrogate the posterior distribution not only about the probability of our parameter being equal to a small range of values including zero, but about the probability of our parameter being within any other interval of interest representing an area delimited by the posterior probability density function (Kéry, 2010; Kéry & Schaub, 2012). This way, the dichotomy between positive and negative results associated to classical hypothesis testing vanishes. Bayesian statistics, with all its particular drawbacks are, in this sense, more appropriate for the field ecologist and her/his battery of observational data. So, are negative results of any use? All of this does not imply that our negative results are not useful and that they should not be published. There is good information in each study that can be taken advantage of by means of meta–analysis (Borenstein et al., 2009). Means and standard deviations are valuable information, when duly weighted by sample size to obtain overall effect sizes (Gurevitch & Hedges, 2001). These overall effect sizes are in turn useful to break the undesirable situation of not having an a priori idea of the effect expected to be relevant in our studies, and hence they are useful to make use of statistical inference by hypothesis testing in the right way, which is using a priori power tests to calculate required sample sizes to couple biological

and statistical significance (Martínez–Abraín, 2008) or using them as prior information in Bayesian analyses. In addition, publishing negative results (abundant in the grey literature) helps prevent publication bias in meta–analysis, a common problem when trying to synthesize knowledge quantitatively in an unbiased way (Møller & Jennions, 2001), and basic for proper decision–making in applied ecology (Stewart, 2010). Publication bias arises because studies with larger effect sizes are more likely to be statistically significant (positive) for any given sample size, and hence larger effects are more likely to be published, leading to overestimation of the true overall effect (Borenstein et al., 2009). As Scargle (2000) stated, 'apparently significant, but actually spurious, results can arise from publication bias, with only a modest number of unpublished studies and hence, statistical combinations of studies from the literature can be trusted to be unbiased only if there is reason to believe that there are essentially no unpublished studies (almost never the case!)'. Preventing this, by publishing negative results, is especially relevant in times of ecological and economic crisis as currently prevail. Final remarks Developing journals that promote publication of negative results (such as the Journal of Negative Results of the University of Helsinki http://www.zoominfo.com/ company/Journal+of+Negative+Results–354476827 or the initiatives by the Centre for Evidence–Based Conservation at Bangor University http://www.cebc. bangor.ac.uk/ together with Cambridge University http://www.conservationevidence.com/) to publish and synthesize grey literature) is a fundamental step to improve knowledge accumulation in applied ecology. Additionally, the recent expansion among ecologists of model selection by means of information criteria and Bayesian statistics (see e.g. Halstead et al., 2012), as enabled by modern computers, will contribute to ending this conundrum, because results are not classified in a dichotomous way around an arbitrary a priori risk level of being wrong (alpha) but in a continuous way. Hopefully, we will see this old debate around positive and negative results die in the near future. This will translate into better decision–making in fields such as applied conservation biology. With a focus on simultaneous multiple hypothesis testing and effect sizes, black or white debates will be substituted by a scale of greys, better representing what really occurs out there in the complex world of Earth´s ecosystems. Meanwhile, let’s not discard negative results, because we need to extract as much information as possible from our costly data. Acknowledgements The author was funded by a Parga–Pondal postdoctoral contract from the Xunta de Galicia. Thanks also to two anonymous referees and Luis M. Carrascal for fruitful suggestions.

References Anderson, D. R., 2008. Model based inference in the life sciences: a primer of evidence. Springer, New York. Borenstein, M., Hedges, L. V., Higgins, J. P. T. & Rothstein, H. R., 2009. Introduction to meta–analysis. Wiley & Sons, Southern Gate, Chichester, West Sussex, UK. Burnham, K. P. & Anderson, D. R., 2002. Model selection and multimodel inference: A practical information–theoretic approach. Springer, New York. Dickersin, K., Min, Y. I. & Meinert, C. L., 1992. Factors influencing publication of research results. Follow– up of applications submitted to two institutional review boards. Journal of the American Medical Association, 267: 374–378. Gurevitch, J. & Hedges, L. V., 2001. Meta–analysis: Combining the results of independent experiments. In: Design and analysis of ecological experiments: 347–369 (S. M. Scheiner & J. Gurevitch, Ed.). Oxford Univ. Press, Oxford. Halstead, B. J., Wylie, G. D., Coates, P. S., Valcarcel, P. & Casazza, M. L., 2012. Exciting statistics: the rapid development and promising future of hierarchical models for population ecology. Animal Conservation, 15: 133–135. Kéry, M., 2010. Introduction to WinBUGS for ecologists:

Martínez–Abraín

A Bayesian approach to regression, anova, mixed models, and related analyses. Elsevier, Burlington (MA, USA), San Diego, London and Amsterdam. Kèry, M. & Schaub, M., 2012. Bayesian population analysis using WinBUGS: A hierarchical perspective. Elsevier, Burlington (MA, USA), San Diego, London and Amsterdam. Koricheva, J., 2003. Non–significant results in ecology: a burden or a blessing in disguise? Oikos, 102: 397–401. Leimu, R. & Koricheva, J., 2005. What determines the citation frequency of ecological papers? Trends in Ecology and Evolution, 20: 28–32. Martínez–Abraín, A., 2007. Are there any differences? A non–sensical question in ecology. Acta Oecologica, 32: 203–206. – 2008. Statistical significance and biological relevance: A call for a more cautious interpretation of results in ecology. Acta Oecologica, 34: 9–11. Møller, J. P. & Jennions, M. D., 2001. Testing and adjusting for publication bias. Trends in Ecology and Evolution, 16: 580–586. Scargle, J. D., 2000. Publication bias: The 'File Drawer' problem in scientific inference. Journal of Scientific Exploration, 14: 91–106. Stewart, G., 2010. Meta–analysis in applied ecology. Biology Letters, 6: 78–81.

Animal Biodiversity and Conservation 36.1 (2013)

The competitor release effect applied to carnivore species: how red foxes can increase in numbers when persecuted J. Lozano, J. G. Casanovas, E. Virgós & J. M. Zorrilla

Lozano, J., Casanovas, J. G., Virgós, E. & Zorrilla, J. M., 2013. The competitor release effect applied to carnivore species: how red foxes can increase in numbers when persecuted. Animal Biodiversity and Conservation, 36.1: 37–46. Abstract The competitor release effect applied to carnivore species: how red foxes can increase in numbers when persecuted.— The objective of our study was to numerically simulate the population dynamics of a hypothetical community of three species of small to medium–sized carnivores subjected to non–selective control within the context of the competitor release effect (CRE). We applied the CRE to three carnivore species, linking interspecific competition with predator control efforts. We predicted the population response of European badger, the red fox and the pine marten to this wildlife management tool by means of numerical simulations. The theoretical responses differed depending on the intrinsic rate of growth (r), although modulated by the competition coefficients. The red fox, showing the highest r value, can increase its populations despite predator control efforts if control intensity is moderate. Populations of the other two species, however, decreased with control efforts, even reaching extinction. Three additional theoretical predictions were obtained. The conclusions from the simulations were: 1) predator control can play a role in altering the carnivore communities; 2) red fox numbers can increase due to control; and 3) predator control programs should evaluate the potential of unintended effects on ecosystems. Key words: Predator control, Wildlife management, Competition, Generalist predator, Population dynamics, Population growth. Resumen El efecto liberador de competidores aplicado a las especies de carnívoros: cómo puede aumentar el número de zorros cuando son perseguidos.— El objetivo de nuestro estudio consistió en simular numéricamente la dinámica de poblaciones de una comunidad hipotética de tres especies de carnívoros de talla pequeña y mediana sometidas a un control no selectivo en el contexto del efecto liberador de competidores. Aplicamos el modelo del efecto liberador de competidores, que relaciona la competencia interespecífica con el control de predadores, a tres especies de carnívoro. Así, pudimos predecir la respuesta de las poblaciones de tejón, zorro y marta frente a este mecanismo de gestión de la fauna silvestre por medio de simulaciones numéricas. Las respuestas teóricas fueron distintas en función de la tasa intrínseca de crecimiento (r), si bien estuvieron reguladas por los coeficientes de competencia. El zorro, con el valor de r más elevado, puede aumentar sus poblaciones a pesar del control de predadores si este es moderado. Por el contrario, las poblaciones de las otras dos especies disminuyeron con el control hasta extinguirse. Obtuvimos también tres predicciones teóricas. Las conclusiones de las simulaciones fueron: 1) el control de predadores puede alterar las comunidades de carnívoros; 2) la población de zorros puede aumentar debido al control y 3) los programas de control de predadores deberían evaluar los efectos indeseados que podrían producirse en los ecosistemas. Palabras clave: Control de predadores, Gestión ambiental, Competencia, Predadores generalistas, Dinámica de poblaciones, Crecimiento poblacional. Received: 3 IV 12; Conditional acceptance: 9 V 12; Final acceptance: 8 I 13 Jorge Lozano, Dept. de Ecología, Univ. Autónoma de Madrid, c/ Darwin 2, Edificio de Biología, E–28049 Cantoblanco, Madrid, España (Spain).– Jorge Lozano, Jorge G. Casanovas & Juan M. Zorrilla, Dept. de Ecología, Univ. Complutense de Madrid, c/ José Antonio Novais 12, Ciudad Universitaria, E–28040 Madrid, España (Spain).– Emilio Virgós, Dept. de Biología y Geología, Univ. Rey Juan Carlos, c/ Tulipán s/n., E–28933 Móstoles, Madrid, España (Spain). Corresponding author: J. Lozano ISSN: 1578–665X

Introduction Populations of various taxonomic groups are declining sharply due to human activities in the environment (Groombridge, 1992). In particular, predator control has produced a negative and strong impact on populations of many species of large carnivores (e.g. Schaller, 1996; Breteinmoser, 1998; Rodríguez & Delibes, 2002). A lot of smaller carnivore species, such as the marten species (Martes spp.), the European wildcat (Felis silvestris Schreber, 1775) or the European badger (Meles meles L.), are also affected (Langley & Yalden, 1977; Lankester et al., 1991; Ruggiero et al., 1994; Caro & Stoner, 2003; Lozano et al., 2007). In contrast, adaptable species of carnivores, such as the red fox (Vulpes vulpes L.), can become more abundant in some places as the result of such activities (Baker & Harris, 2006; Beja et al., 2009). The main goal of predator control, in both natural reserves interested in protecting sensitive species and hunting lands with an interest in harvest management, is to reduce the incidence of predation (Tapper et al., 1991; Harris & Saunders, 1993; Reynolds & Tapper, 1996; Côte & Sutherland, 1997). Predator control techniques vary greatly both in their degree of selectivity and effectiveness with regard to the persecuted species (e.g. Calver et al., 1989; Windberg & Knowlton, 1990; Tapper et al., 1991; Hein & Andelt, 1994; Reynolds & Tapper, 1996; Harding et al., 2001; Rushton et al., 2006; Beja et al., 2009). Unfortunately, many of these methods are non–selective (e.g. snares, traps, poisoned baits), and negatively affect both the species considered a pest and others of conservation concern (Herranz, 2000; Duarte & Vargas, 2001; Whitfield et al., 2003; Rodríguez & Delibes, 2004; Virgós & Travaini, 2005; Beja et al., 2009; Cabezas–Díaz et al., 2009; Estes & Terborgh, 2010; Lozano et al., 2010). Many studies have been carried out around the world dealing with the effects of predator control on prey populations (Reynolds & Tapper, 1996; Côte & Sutherland, 1997; Valkama et al., 2005; Oro & Martínez–Abraín, 2007). Nevertheless, the effects of such control on the population parameters of the targeted predators, or on the structure of their natural communities, have received much less attention (e.g. Yoneda & Maekawa, 1982; Harris & Saunders, 1993; Reynolds et al., 1993; Estes & Terborgh, 2010). It is well established that predation, competition and their interaction are important factors shaping natural communities (Chase et al., 2002; Caro & Stoner, 2003). Interspecific interactions within the communities of carnivores can cause the extinction or exhaustion of specialist species or larger predators (e.g. lynxes, wolves, coyotes). For example, such interactions may be interguild predation or competition through exploitation (Erlinge & Sandell, 1988; Polis & Holt, 1992; Palomares & Caro, 1999; Müller & Brodeur, 2002). After the disappearance of these top predators, numbers of smaller species could increase, a pattern also observed in more generalist species such as the Iberian mongoose (Herpestes ichneumon L.) or the red fox (see Palomares et al., 1995; Creel & Creel, 1996; Palomares & Caro, 1999).

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These data imply that different predator management systems could have different effects on the communities of predators and indirectly affect the levels of predation on the prey species (Estes & Terborgh, 2010; Levi et al., 2012). This could even produce paradoxical effects, such as a reduction in the diversity and/or abundance of game species or those of conservation interest, as a result of the increase in the abundance of generalist carnivores (Soulé et al., 1988; Courchamp et al., 1999a; Crooks & Soulé, 2000; Caut et al., 2007). However, given the lack of field data in relation to this issue, and the difficulty involved in obtaining such information, an alternative to study the possible effects of the applied control techniques on the predator populations consists of developing simple mathematical models. With a minimum number of assumptions, it is possible analyse the population dynamics of these species, and the interspecific interactions on community composition (see the application of this type of procedure in the works by Shorrocks & Begon, 1975; Courchamp et al., 1999a, 1999b; Caut et al., 2007; Fenton & Brockhurst, 2007). Caut et al. (2007) used this theoretical approach to describe a new ecological mechanism named the Competitor Release Effect (hereafter CRE). According to this mechanism, an inferior competitor can increase in numbers if the superior competitor is controlled, due to the competitive interactions between them. This occurs even though the inferior competitor is also being killed. Moreover, at the same time, theoretical results show negative effects on the population of a shared prey. Shared prey can decline because while numbers of a superior competitor are decreasing, there may be an unwanted and unexpected increase in numbers of the inferior competitor. Caut et al used empirical data from an eradication program of rodents living on islands to test and support their competitor release hypothesis. They also suggested that the same effect could be found in communities of carnivore mammals, where the population of a competitor such as the red fox could increase if the community of predators is being managed using predator control. The objective of our study was to numerically simulate the population dynamics of a hypothetical community of three species of small to medium–sized carnivores subjected to non–selective control (i.e. where all the individuals are being eliminating with similar probability), within the context of the proposed CRE (Caut et al., 2007). The selected species for the simulations were the European badger, the red fox and the pine marten (Martes martes L.). Reasons for this choice were: (i) these species are sympatric in a wide range of Europe (Mitchell–Jones et al., 1999); (ii) there is evidence of competition among them (Lindström et al., 1995; Palomares & Caro, 1999; Trewby et al., 2008); (iii) all three species are often controlled (e.g. Côté & Sutherland, 1997; Virgós & Travaini, 2005; Trewby et al., 2008) and; (iv) information about their populational parameters can be found in the scientific literature (Bright, 1993). Our study differs from that of Caut et al. (2007) in that we evaluated a third species in the mathematical

Animal Biodiversity and Conservation 36.1 (2013)

model. Moreover, known (i.e. real) values for the population growth rates were used, so the model should be more realistic. We specifically wanted to know whether the CRE could increase the population of red fox when the three species are being controlled, and if so, under what conditions such an increase occurs. Furthermore, we used the results of numerical simulations from the theoretical model to obtain a set of predictions which could be tested with empirical data when available. Material and methods We modified the CRE model from Caut et al. (2007) by adding an equation to simulate the populational dynamics of a third competitor. This model was based on the classical Lotka–Volterra equations (Powell & Zielinski, 1983; Begon et al., 1996), which were modified to incorporate an additional factor of linear mortality (in a similar way that in Shorrocks & Begon, 1975; Fenton & Brockhurst, 2007). The slope of this new factor is independent from the density and represents the mortality caused by the predator control involved (degree of non–selectivity). Thus, this modification of the classical model, which only includes density–dependent mortality implicitly in the population growth rate, is analogous to those previously proposed by Gause (1934, 1935). In these studies, the author considered mortality independent from density, possibly caused by factors such as parasites, non–specific diseases, or other mortality factors (e.g. Caut et al., 2007; Fenton & Brockhurst, 2007). Our model is defined by three equations that govern the coupled dynamics of three species of competing carnivores: At+1= At + rA At (1 –

At + αAB Bt + αAC Ct KA

) – ω At

Bt + αBA At + αBC Ct Bt+1 = Bt + rB Bt (1 – ) – ω Bt KB Ct + αCA At + αCB Bt Ct+1 = Ct + rC Ct (1 – ) – ω Ct KC where A, B and C represent the number of individuals of each particular species at time t. The intrinsic growth rates of each population are rA, rB and rC. The effect of the interspecific competition of one species against another is represented by α (which is the competition coefficient), and the carrying capacity of the environment for each population is K. Finally, included in each equation is the parameter ω (control coefficient) which represents the extraction rate of each species as a result of the non–selective control applied. This can be interpreted as the proportion of the population of a given species which dies during a period t as a consequence of the control. Because the model attempts to determine the effect of non–selective control, this ω parameter was fixed with the same

value for all the species of the community, although more complex scenarios could be developed within the premise of non–selective control. In this scenario, the set of values of the parameters was chosen to take into consideration that each species of the model represented one for which the intrinsic growth rate (r) was available (Bright, 1993). These values are from British populations of each species, but we assumed that the growth rates were similar for other regions of Europe (Turchin, 2003). The three carnivore species were the European badger (species A), the pine marten (species B), and the red fox (species C). The carrying capacities (K) were set as constant and identical to facilitate the interpretation of results from the numerical analysis. The selected value is theoretical but also realistic in function of the considered spatial scale (K = 30, approximately equivalent to 25–30 km2 considering mean density values of the species in Europe; Wilson & Mittermeier, 2009), and it allows a sufficient range of variation to compare the populational dynamics among species. Likewise, the competitive interactions among the species were considered symmetrical (i.e. AB = AC = BA = BC = CA = CB). Although asymmetries can be expected in the wild (e.g. Palomares & Caro, 1999), unfortunately no quantitative data are available to create more realistic scenarios. Table 1 shows the demographic values used for the parameters in the model. A total of 72 deterministic numerical analyses were performed per species: one for each combination of parameters, varying in equal intervals α from 0.1 to 0.9, and ω from 0 to 0.7. The value reached by the population in the equilibrium for each combination of parameters was graphically represented. Furthermore, to test whether the predictions arising from the theoretical model were sufficiently robust to stand up against small variations in the values for the intrinsic growth rate (r), a sensitivity analysis was conducted. This consisted of creating simulations varying by a single parameter and leaving the other parameters constant. Thus, α = 0.5, ω = 0.3, rA = 0.46 and rC = 1.1 were fixed, whereas rB (the intermediate intrinsic growth rate regarding original values) varied in five intervals from 0.4 to 1.1. This also could be equivalent to simulating the population dynamics of a competiting species other than the pine marten (species B), showing different values of intrinsic growth rate. All these analyses were performed using the computer program STELLA v.9.1.4 for Windows (ISEE, 2010). The basic model used here and the Lotka–Volterra equations have been analytically studied elsewhere using the same assumptions (for more details see Gilpin, 1975; Shorrocks & Begon, 1975; Caut et al., 2007). Therefore, our work is an extension of these basic models, incorporating a level of predator community complexity more commonly found in natural systems. Results Simulation results From our numerical simulations, we found a qualitatively different behaviour among the carnivore popu-

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Table 1. Demographic values for parameters of the model: r is the intrinsic growth rate for each species, and K is the carrying capacity of the environment. Tabla 1. Valores demográficos utilizados para los parámetros del modelo: r es la tasa intrínseca de crecimiento para cada especie y K es la capacidad de carga del medio.

Letter in equation

M. meles (badger)

M. martes (pine marten)

V. vulpes (red fox)

0.46

0.57

1.1

Species

lations, considering the competitive interactions that can occur in the community as a result of a predator control program. The sensitivity analysis of the system with regard to the value of species B (r = 0.57, the pine marten) indicates that when there is intermediate competition (0.5) and moderate control (0.3), the variation in the intrinsic growth rate affects the equilibrium value of species C (r = 1.1, the red fox) but does not appear to affect the species A (r = 0.46, the European badger). There is a critical value for rB, at 0.9, above which the red fox population does not rise above K* (the maximum population value in the presence of the other two species, in this scenario K* = 15 individuals). At intermediate levels of competition, the model predicts a 'paradoxical effect' produced by the non–selective control. In other words, when species B shows an r > 0.9 (this being a threshold value), the population increases rather than decreases, despite the applied control. The simulations of the general scenario, where the intrinsic growth rates were fixed (0.46, 0.57 and 1.1), produced a system dynamic showing very clear patterns of population change depending on competition and predator control intensity. Thus, if the intrinsic growth rate (r) of one species was below the threshold value r' = 0.9, then the population would change in time, always showing a decrease in its numbers (see figs. 1, 2). The simulated population with the lowest intrinsic growth rate, corresponding to the European badger (r = 0.46, fig. 1), showed a linear decrease in its numbers when predator control was applied. This population became completely extinct with an intermediate degree of control intensity (0.5), in conditions of minimum competition (0.1). The increase in competition implies that the population could be destroyed under conditions of even less intense non–selective control. Likewise, the general pattern observed in the change of the population with a slightly higher growth rate (r = 0.57), belonging in this case to the pine marten, was practically the same (see fig. 2). The difference was that the higher growth rate implied that the population of this species needed a slightly higher level of control intensity than the badger to disappear: a value of 0.6, under conditions of minimum competition (0.1).

In strong contrast, the population of the red fox behaved in a very different manner under conditions of non–selective control, and depending on competition (see fig. 3). The pattern presented a very marked non– linearity which could be attributed to their high intrinsic growth rate (r = 1.1). Without competition, or with very low levels of competition (up to 0.2), control efforts reduced the population in a linear sense, such as in the previous species, but maintained a large number of individuals even under conditions of very intense control (0.7). The red fox population could reach extinction only with a very high degree of control intensity, the value of control coefficient being near the maximum. At a medium level of control intensity, the population maintained approximately half the individuals of the population maximum (K), regardless of competition. Furthermore, with a low level of competition (0.3) this population was not affected by the number of individuals when low intensities of non–selective control were applied (< 0.4). Surprisingly, when starting from this low level of competition (0.3), the low control intensities (< 0.4) led to a sharper population increase (the paradoxical effect mentioned above) if there was a higher degree of competition with the other two species. The increase in red fox occurred precisely when there was a decrease in the number of their competing species, whose populations showed lower intrinsic growth rates. The final result of the application of predator control under these conditions could therefore be a red fox population with double the initial number of individuals. When the degree of competition was higher, the intensity of control needed to be lower to reach the maximum level of increase. The absolute population maximum was then also reached when competition was maximal. Predictions of the theoretical model Based on the results of the numerical analysis and due to the CRE, we made the following set of three predictions: i) populations of red fox showing maximal abundances (or those of other generalist predators showing a high intrinsic growth rate) will be present in areas subjected to predator control (usually areas devoted to small game hunting); ii) statistically signi-

Animal Biodiversity and Conservation 36.1 (2013)

Meles meles

Equilibrium population

30.00

Competition coefficients

25.00

0.1 0.2 0.3 0.4 –0.5 –0.6 0.7 0.8 –0.9

20.00 15.00 10.00 5.00 0

0.0

0.1 0.2 0.3 0.4 0.5 0.6 Non–selective control coefficient

0.7

Fig. 1. Functions for European badger (Meles meles) populations considering the competition coefficients (α) that relate the value for the population in dynamic equilibrium according to the intensity degree of non–selective control (ω). All populations become extinct at intermediate degrees of control intensity, following a linear pattern. Fig. 1. Funciones de las poblaciones de tejón (Meles meles) teniendo en cuenta los coeficientes de competencia (α) que relacionan el valor de la población en equilibrio dinámico con el grado de intensidad de control no selectivo (ω). Todas las poblaciones se extinguen con una intensidad de control intermedia siguiendo un patrón lineal.

ficant differences between controlled and uncontrolled areas will not be found in the abundance of red fox (or those of other generalist predators showing a high intrinsic growth rate). But according to prediction 1), if differences appear, the red fox will be more abundant in controlled areas; and iii) the most abundant populations of competing species showing low intrinsic growth rates will be found in areas where predator control programs are not implemented. Thus, statistically significant differences will be found in the abundance of predator populations with low intrinsic growth rate between controlled and uncontrolled areas, with the more abundant populations inhabiting the uncontrolled areas. Discussion Many managers of natural areas, gamekeepers and the hunting community in general have the perception that generalist predators (including several species of rodents, corvids, gulls, and carnivores) increase continuously and are so abundant that their populations should be controlled (e.g. Herranz, 2000; Garrido, 2008). One of the most persecuted species is the red fox, blamed for reducing populations of game species (Herranz, 2000; Virgós & Travaini, 2005; Rushton et al., 2006; Beja et al., 2009). Although the belief that red fox populations are increasing everywhere and continuously is probably an exageration (see the case of a large Spanish region in Sobrino et al., 2009), it seems true that under certain conditions this species

can increase above normal values (e.g. Beja et al., 2009; Trewby et al., 2008). Surprisingly, the paradox is that these real increases in abundance, as in the case of other generalist predators, occur even though their populations are subjected to permanent control campaigns (Herranz, 2000; Virgós & Travaini, 2005; Beja et al., 2009). Caut et al.´s competitor release effect (CRE) described an ecological mechanism that is applicable to mesocarnivore populations and could theoretically explain this paradox (2007). The CRE framework implies a scenario where the carnivores community is shaped by a number of species presenting negative interspecific interactions (–,–) and experiencing a non–selective predator control program. Under these conditions, our theoretical results predicted that certain changes will occur in the composition and structure of the community. Thus, predator control efforts might eliminate populations with a low intrinsic growth rate (r) and only the population with a high rate of increase might persist (in our case the red fox, whose populations showed a growth rate higher than 0.9), unless the control is extremely intense. Surprisingly, if the control level is moderate, then the populations of these species of generalist predators could increase in a paradoxical way, even surpassing the theoretical value for maximum population in equilibrium in the presence of the remaining species. This is due to both the disappearance of competitors and their higher reproductive capacity. Moreover, the ecological consequences of a mechanism such as the CRE might be similar to those produced by

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Martes martes

Equilibrium population

30.00

Competition coefficients

25.00

0.1 0.2 0.3 0.4 –0.5 –0.6 0.7 0.8 –0.9

20.00 15.00 10.00 5.00 0

0.0

0.1 0.2 0.3 0.4 0.5 0.6 Non–selective control coefficient

0.7

Fig. 2. Functions for pine marten populations (Martes martes) follow a similar pattern to those of European badger, although their higher growth rate requires more intense control to completely eliminate the populations. Fig. 2. Las funciones de las poblaciones de marta (Martes martes) siguen un patrón parecido a las de tejón, aunque su mayor tasa de crecimiento hace necesario intensificar el control para eliminar totalmente las poblaciones.

the mesopredator release effect (see Soulé et al., 1988; Courchamp et al., 1999a), if net predation on certain prey species were to increase as a result of the population increase of control–resistant predators (Caut et al., 2007). Our modelling therefore supports both the CRE hypothesis and the MRE hypothesis. We found three predictions from our model evaluation that should be specifically tested with empirical data obtained in the field. In general, these predictions are based on the fact that red fox populations (in our case scenario) will increase or maintain their numbers despite the implementation of predator control efforts (Predictions 1 and 2), while other species of carnivores will become less abundant (Prediction 3). It is expected that the most sensitive species disappear over time, so that species richness will decrease in controlled areas (see Estes & Terborgh, 2010). Interestingly, the few data available in the scientific literature seem to support our findings. For example, the results of a study carried out in Portugal showed that in hunting grounds where predator control was practised, the abundance of red fox was almost twice as high as in non–hunting areas (Beja et al., 2009), which appears to bear out Predictions 1 and 2. Moreover, other species of predators tended to be more abundant in uncontrolled areas, supporting Prediction 3 in our study. Similarly, predator control on badgers in the UK increased the red fox population, again appearing to meet Prediction 1 (Trewby et al., 2008). The results obtained by Virgós & Travaini (2005) in Spain also appear to generally support the CRE predictions. These authors detected the absence of some carnivores

in controlled areas, while the red fox frequency of occurrence in both controlled and uncontrolled areas was similar. However, field data collected following a carefully designed study protocol are needed to reliably test the CRE model predictions found in this study. In our model, no explicit consideration was given to the effects that spatial heterogeneity, landscape pattern, and structure of a territory could have on the behaviour of population and community dynamics. There is evidence that landscape composition and quality affect interactions among species (Erlinge & Sandell, 1988; Hanski, 1995), the efficacy of predator control programs (Schneider, 2001; Rushton et al., 2006), and therefore the persistence at a regional level of a given pool of species. Given that the model predicts the probabilities of differential extinction of the species in fragmented landscapes and complex environments, the long–term configuration of the communities will also depend on the different probabilities of recolonisation (Hanski, 1994; Schneider, 2001; Rushton et al., 2006). It is possible to speculate about the existence of deterministic processes within a community of carnivores subjected to non–selective control. These processes would occur at the local level, but predictable consequences would result at the landscape scale. These aspects should also be tested independently through further research with empirical data. The fundamental objective of predator control is an effective increase in the populations of prey species of interest to hunting or conservation (Trout & Tittensor, 1989; Reynolds & Tapper, 1995, 1996; Côte & Sutherland, 1997; Virgós & Travaini, 2005;

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Vulpes vulpes

Equilibrium population

30.00

Competition coefficients

25.00

0.1 0.2 0.3 0.4 –0.5 –0.6 0.7 0.8 –0.9

20.00 15.00 10.00 5.00 0

0.0

0.1 0.2 0.3 0.4 0.5 0.6 Non–selective control coefficient

0.7

Fig.3. Functions for red fox populations (Vulpes vulpes) are qualitatively different from those of the previous species due to an intrinsic growth rate higher than 0.9. Red fox populations thus show a non–linear response when persecuted depending on the level of competition: populations increase when competition coefficients are greater than 0.3 and control intensity is moderate. Furthermore, red fox populations do not become extinct despite intense predator control. Fig. 3. Las funciones de las poblaciones de zorro (Vulpes vulpes) son cualitativamente diferentes de las de las especies anteriores debido a una tasa intrínseca de crecimiento mayor que 0,9. Así, las poblaciones de zorro muestran una respuesta no lineal cuando son perseguidas dependiendo del grado de competencia: las poblaciones aumentan cuando los coeficientes de competencia son mayores que 0,3 y la intensidad del control es moderada. Además, las poblaciones de zorro no se extinguen aunque se aplique un control de predadores intenso.

Reynolds et al., 2010). This is based on studies that found direct effects of control or natural reduction of predators on the abundance and dynamics of prey populations (e.g. Marcström et al., 1988, 1989; Small & Keith, 1992; Lindström et al., 1994). However, the success of predator control campaigns is variable and, in general, very expensive (Reynolds & Tapper, 1996; Côte & Sutherland, 1997). Some studies have thus suggested that these practices are effective (regarding the above indicated objective) when applied at the local level, in conditions of very intense control, but only in the short–term (e.g. Reynolds et al., 1993; Harding et al., 2001; Keedwell et al., 2002). Other studies have shown that the predator control was ineffective in meeting management goals (Reynolds & Tapper, 1996; Côte & Sutherland, 1997; Banks, 1999; Herranz, 2000; Kauhala et al., 2000; Keedwell et al., 2002; Martínez–Abraín et al., 2004; Baker & Harris, 2006; King et al., 2009). Moreover, there are many predator species that do not affect game or threatened species. Thus, it has been argued that predator control can not be effective when focused on them (see for the cases of lizards, snakes and large gulls Herranz, 2000; Oro & Martínez–Abraín, 2007). Overall, it has been considered that the unique nature of predator–prey relationships within communities makes it difficult to make generalizations, and that evaluation of the effectiveness of conducting a predator control

program thus requires individual consideration (Sih et al., 1998; Abrams & Ginzburg, 2000; Turchin, 2003; Holt et al., 2008; Valkama et al., 2005). On the other hand, predator management could have ecological costs that depend on the relative importance of the different uses and intrinsic values of the territory. For example, the ecological consequences of controlling one or various species of predators might be appraised positively or negatively depending on the environmental perception, and on the type of local use of the natural resources (Langley & Yalden, 1977; Banks et al., 1998). Thus, perception might be different if predator control is used to enhance an endangered species rather than a game species (e.g. Côte & Sutherland, 1997; Keedwell et al., 2002). Furthermore, predator control also appears to affect different demographic parameters of the target predator species, including density, age structure, and inmigration patterns (see Yoneda & Maekawa, 1982; Rushton et al., 2006). However, the more notable and more harmful effects of non–selective predator control are related to the conservation of threatened species of predators and the unwanted consequences on ecosystems due to the alteration of natural communities, such as the increase of generalist predators (including target species of the control) and pest species (including rodents), the decline of shared prey species (including

those of game interest), and unforeseen effects on vegetation, ecosystem function, and similar owing to chain reactions (e.g. Herranz, 2000; Martínez–Abraín et al., 2004; Rodríguez & Delibes, 2004; Virgós & Travaini, 2005; Caut et al., 2007; Cabezas–Díaz et al., 2009; Estes & Terborgh, 2010). The theoretical results obtained in this study highlight the importance of competitive ecological interactions among predators in the design of an optimum management strategy for their communities (Courchamp et al., 1999a, 1999b; Trewby et al., 2008). Caut et al´s CRE has shown how the complex network of interactions (see also Polis & Holt, 1992; Chase et al., 2002; Caro & Stoner, 2003) among carnivore mammals can also lead to undesired effects, such as a population increase in the target species (the red fox or any predator with high reproductive capacity in our study, or the American mink Neovison vison Schreber 1777; see Bright, 1993; King et al., 2009), and the elimination of more sensitive species that might be of conservation interest. The obtained results support the idea that the design of programs to manage predator populations should consider potential consequences to communities and the ecosystem as a whole (Schneider, 2001; Zavaleta et al., 2001; Courchamp & Caut, 2005; Caut et al., 2007), as well as the biological traits of the involved species. To validate our model findings, empirical data should evaluate these responses and not just the individual species’ responses of the targeted predator and prey. The development of management strategies for species such as the red fox populations should take the ecological framework into account, and predator control programs should be thoroughly evaluated to determine the potential impact on the community and ecosystem. This is particularly important for predator control programs using non–selective methods (e.g. Herranz, 2000; Virgós & Travaini, 2005; Beja et al., 2009), where numbers of red foxes and other generalist predators can increase despite the efforts of managers. Acknowledgements We are grateful to Ana Rojas, Raúl Bonal, Sara Cabezas and Cormac de Brun for help in different steps of the study, including the translation to English. We also thank Daniel L. Huertas, Mario Díaz and two anonymous referees who provided useful suggestions and comments to different drafts of this manuscript. References Abrams, P. A. & Ginzburg, L. R., 2000. The nature of predation: prey dependent, ratio dependent or neither? TREE, 15: 337–341. Baker, P. J. & Harris, S., 2006. Does culling reduce fox (Vulpes vulpes) density in commercial forests in Wales? European Journal of Wildlife Research, 52: 99–108.

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Assessing the extent of occurrence, area of occupancy, territory size, and population size of marsh tapaculo (Scytalopus iraiensis) L. Klemann Jr. & J. S. Vieira

Klemann Jr., L. & Vieira, J. S., 2013. Assessing the extent of occurrence, area of occupancy, territory size, and population size of marsh tapaculo (Scytalopus iraiensis). Animal Biodiversity and Conservation, 36.1: 47–57. Abstract Assessing the extent of occurrence, area of occupancy, territory size, and population size of marsh tapaculo (Scytalopus iraiensis).— First described in 1998, the marsh tapaculo (Scytalopus iraiensis) is an endangered bird of the family Rhinocryptidae. It is endemic to Brazil and is restricted to the wet flood plains of rivers and streams. Due to its cryptic habits and environments of occurrence, information available on its biology, natural history and distribution is scarce. We compiled occurrence records (99 records), delimited the extent of occurrences (296,584 km2), calculated the area of occupancy (84 km2), estimated territory size (5,313 ± 1,201 m2 per pair), population density (3.76 ± 0.85 individuals per hectare), and population size (31,584 ± 7,140 mature individuals) of marsh tapaculo. The species was recorded in marshes associated to four types of vegetation and in four ecological zones. This new information is extremely important to support revaluation of the species’ threat category and to enhance knowledge about this endemic and little known bird from Brazil. Key words: Scytalopus iraiensis, Distribution, Population, Endangered species, Southeastern Brazil. Resumen Evaluación de la extensión de presencia, la superficie de ocupación, el tamaño del territorio y el tamaño de la población del churrín palustre (Scytalopus iraiensis).— Descrito por primera vez en 1998, el churrín palustre (Scytalopus iraiensis) es una ave en peligro de extinción de la familia Rhinocryptidae. Es endémica de Brasil y su presencia queda restringida a los planos aluviales de los ríos y los cursos de agua. Debido a sus hábitos crípticos y a los ambientes en los que se halla presente, la información disponible sobre su biología, su historia natural y su distribución es escasa. Compilamos varios registros de presencia (99 registros), delimitamos la extensión de las presencias (296.584 km2), calculamos la superficie de ocupación (84 km2) y estimamos el tamaño del territorio (5.313 ± 1.201 m2 por pareja), la densidad de la población (3,76 ± 0,85 individuos por hectárea) y el tamaño de la población (31.584 ± 7.140 individuos maduros) del churrín palustre. La especie se registró en zonas de marismas asociada a cuatro tipos de vegetación y en cuatro zonas ecológicas. Esta nueva información es fundamental para respaldar la reevaluación de la categoría de situación de peligro de la especie y potenciar el conocimiento de esta ave endémica y poco conocida de Brasil. Palabras clave: Scytalopus iraiensis, Distribución, Población, Especie en peligro de extinción, Sureste de Brasil. Received: 13 X 12; Conditional acceptance: 18 XII 12; Final acceptance: 8 I 13 Louri Klemann Júnior, Ecology and Conservation Post–Graduation Program, Univ. Federal do Paraná, P. O. Box 19020, Curitiba, Paraná, Brazil.– Juliana S. Vieira, Entomology Post–Graduation Program, Univ. Federal do Paraná, P. O. Box 19020, Curitiba, Paraná, Brazil. Corresponding author: L. Klemann Jr. E–mail: klemannjr@yahoo.com.br

ISSN: 1578–665X

Klemann Jr. & Vieira

Introduction

Material and methods

The marsh tapaculo (Scytalopus iraiensis) is a species of the bird family Rhinocryptidae. It was first described in 1998 from specimens taken from two municipalities in the metropolitan region of Curitiba (capital of Paraná state, southern Brazil) and its occurrence is known to be restricted to the wet flood plains of rivers and streams (Bornschein et al., 1998). Its habitat is described as upland marshes and also, at Rio Grande do Sul state (south of Brazil), coastal wetlands associated to grasslands and Atlantic Forest biomes (Machado et al., 2005). According to Bornschein et al. (1998, 2001) this bird occurs in dense humid floodplain watercourses, usually surrounded by alluvial forests, ranging from less than 1 ha to approximately 350 ha, with a vegetation height ranging from 60 to 180 cm. After the species’ description, new records were obtained in Paraná, Santa Catarina and Rio Grande do Sul states (south of Brazil) and in Minas Gerais state (southeast of Brazil) increasing the number of locations, municipalities and states with records of marsh tapaculo (Bornschein et al., 1998, 2001; Accordi et al., 2003; Maurício, 2005; Straube et al., 2005; Bencke et al., 2006; Raposo et al., 2006; Corrêa et al., 2007, 2008; Vasconcelos et al., 2008; Fontana et al., 2008; Rodrigues et al., 2008). Despite the new records, knowledge of this species’ real distribution is still very basic. New records are expected even in well–studied regions, an expectation reinforced by the species’ description fifteen years ago (Bornschein et al., 1998), based on individuals from a metropolitan region with 1.75 million inhabitants (IBGE, 2007), and from records obtained in well–studied locations of Minas Gerais state (e.g. Serra da Canastra and Serra do Cipó) as of 2003 (Vasconcelos et al., 2008). The cryptic habits, common to all representatives of this genus, and the flooded environment where the species occurs make research of this tapaculo difficult, accounting thereby for the little information produced and available. To date, only two papers have been published on the species’ biology and behavior (Hassdenteufel et al., 2006a, 2006b), and three on the species’ distribution (Bornschein et al., 2001; Corrêa & Woldan, 2007; Vasconcelos et al., 2008). Apart from the little information available at the time of the description, a suggestion was made to include marsh tapaculo in the Brazilian list of endangered species (Bornschein et al., 1998, 2001). Nowadays, this tapaculo is considered ‘Endangered’ at national and international levels (Machado et al., 2005; BirdLife International, 2012). This lack of information on biological data and distribution details of the marsh tapaculo can lead to erroneous classification of the threat category. It is therefore extremely important to compile records, delimit the extent of occurrence, measure the area of occupancy, and estimate the territory size, population density and population size in order to reevaluate the threat category and to expand knowledge of this endemic and little known bird from Brazil.

Study area The study area included the extent of occurrence of marsh tapaculo, from Rio Grande do Sul (RG), Santa Catarina (SC) and Paraná state (PR) in Southern Brazil and São Paulo (SP) and Minas Gerais (MG) state in Southeastern Brazil, between coordinates 19º 07'–32º 19' S and 43º 20'–52º 28' W. Records Records of the species were obtained from literature, photos and recordings available on the internet (http:// www.xeno–canto.org/ and http://www.wikiaves.com. br/), unpublished records from third persons (personal communication) and bird surveys carried out by the first author between 2002 and 2012. For records taken from the literature, internet information and personal communications, the coordinates used were those provided by the author of each record. When this information was not available and the location could not be specified, the municipality centroid was used. Due to the species’ cryptic habits (Bornschein et al., 2001), the records gathered by the first author were obtained by listening to audio communications stimulated by means of song playbacks. The coordinates for the points where the species was recorded were obtained through a global positioning system (GPS) device. Extent of occurrence, area of occupancy, territory size and population size The records were represented in a geographical information system (GIS). The extent of occurrence, defined as the smallest area which can be drawn to encompass all the known, inferred or projected sites of present occurrence of a taxon (IUCN, 2001), was delimited by using the minimum convex polygon method (Mohr, 1947; IUCN, 2001); this was defined as the smallest polygon in which none of the internal angles exceed 180 degrees, and containing all the species points of occurrence (IUCN, 2001). This method, usually used to calculate the home range size (Hayne, 1949; White & Garrott, 1990; Harris et al., 1990), was adapted to the delineation of the extent of occurrence (IUCN, 2001). By overlaying the records and thematic maps, we identified the vegetal formation (IBGE, 1993) and the global ecological zones (ecozones; FRA, 2000) where the species occurs. Considering the minimum convex polygon method, which incorporate large areas that are not used or occupied by the species (Ostro et al., 1999; Powell, 2000; Burgman & Fox, 2003), and aiming to create a realistic extent of occurrence, this polygon was adjusted using the adjusted polygon method (Mills & Gorman, 1987; Li & Rogers, 2005; Grueter et al., 2009) by excluding the vegetal formation without records of the species. We also excluded small separated areas, created by the polygon adjustment, and great water bodies.

Animal Biodiversity and Conservation 36.1 (2013)

The area of occupancy, defined as the area inside the extent of occurrence occupied by the species (IUCN, 2001), was obtained using the grid cell method. To accomplish this goal, the recorded points were overlaid on a grid, the occurrence cells were identified, and their area was added, resulting in a value in square kilometers. This method, which is usually used to calculate home range size (Adams & Davis, 1967; White & Garrott, 1990), was adapted to calculate the area of occupancy according to IUCN (2001). The size of the cells is a factor that can influence results, overestimating or, more commonly, underestimating the calculated area (Kool & Croft, 1992; IUCN, 2001; Lehmann & Boesch, 2003; Grueter et al., 2009). Although the choice of cell size was usually a decision with no biological basis or known objective procedures (White & Garrott, 1990), the size of cells used in this research (1 km2) was defined considering the disjunctive distribution of the habitat where the species occurs. We aimed at not producing overestimated values for the area of occupancy, as is preferable from a conservation point of view (IUCN, 2001). The size of the territory of the marsh tapaculo was estimated by counting individuals in eight different marshes where the species was recorded. The marshes sampled are in two of the four states where the species occur, Paraná and Santa Catarina. These two states contain 74.75% of the compiled records, justifying the implementation of sample sites in these regions. The marshes studied are distributed in the two main ecological zones, subtropical humid forest and tropical mountain system (94.95% of the records have been collected in these zones), with two main vegetation types, araucaria moist forest and grassland (85.86% of the records come from these vegetation types). Each marsh was sampled twice, in March 2011 and April 2012, using song playback to stimulate vocalization. Considering that the species respond well to the playback and could be heard easily (Bornschein et al., 1998), we followed the point count method (Ferry & Frochot, 1970; Hutto et al., 1986). This is one of the most widely used counting methods in bird population studies (Rosenstock et al., 2002) , with an unlimited radius (see Simons et al., 2007). We used an adapted version of the double–observer approach (Nichols et al., 2000; Thompson, 2002), with two observers in each marsh, positioned to visually cover the whole sample area. Due to the species’ cryptic habits (Bornschein et al., 1998), we were unable to use other methodologies (such as banding, or observation) to estimate the number of individuals. For each location sampled we counted the number of individuals vocalizing at the same time, heard in a period of 15 min after playback (done for 10 min). The marshes sampled were vectorized utilizing satellite images, taken in 2010 with 0.5 meter resolution, and the areas were then calculated. Considering the high resolution of the images used and the great facility in marsh identification and delimitation it was possible to vectorize these vegetal formations with great precision, resulting in highly accurate calculations of their areas.

To measure the territory size, the marsh area was divided by the maximum number of individuals vocalizing at the same time after playback, and the average value of this ratio was used. Considering territory as a defended area (Howard, 1920) that provides food, nesting sites and mates (Perrins & Birkhead, 1983), the value obtained would be an estimate of the mean size of the territory occupied by a pair of marsh tapaculo. Population density was then estimated as twice the inverse of territory size, and population size was estimated by multiplying the area of occupancy by the population density value. Results Records and occurrence A total of 99 occurrence records was compiled; 47 from the literature, 22 based on images and recordings from the Internet, six personal communications, and 24 from bird surveys carried out by the first author. The occurrence locations (70) were distributed over 42 municipalities in Paraná, Santa Catarina, Rio Grande do Sul and Minas Gerais states (table 1). According to the global ecological zones (FRA, 2000), the records were located in four regions: subtropical humid forest (58 records), tropical mountain system (36), tropical moist deciduous forest (3) and tropical rainforest (2). According to the Brazilian vegetation map (IBGE, 1993) and Brazilian vegetation classification (Veloso et al., 1991), the records were located in marshes within the domains of four types of vegetation: araucaria moist forest (60 records) in its alluvial, montane and cloudy formations (e.g. Rio Grande do Sul, Santa Catarina and Paraná); semi–deciduous seasonal forest (7) and its formations submontane and montane (e.g. Minas Gerais); grassland (25) (e.g. Rio Grande do Sul, Santa Catarina, Paraná and Minas Gerais); pioneer formation areas (5) (e.g. Rio Grande do Sul coastal line); and ecological tension areas (2) between grassland and semi–deciduous seasonal forest (fig. 1). Extent of occurrence, area of occupancy, territory size and population size The minimum convex polygon method gave an extent of occurrence of 424,064 km2. After excluding the vegetal formation without any record of the species, the extent of occurrence obtained was 296,584 km2 (fig. 1). We used a 1 km2 cell grid to calculate the species’ area of occupancy (a total of 299,536 cells to cover all of the extent of occurrence), resulting in 84 km2. This area is distributed over Rio Grande do Sul (6 km2), Santa Catarina (23 km2), Parana (44 km2) and Minas Gerais (11 km2) states. The territory size found for the species was 5,313 ± 1,201 m2 (ranging from 3,589 to 6,990 m2), resulting in a population density of 3.76 ± 0.85 individuals per hectare. One to five individuals were

Klemann Jr. & Vieira

Table 1. Records of marsh tapaculo between 1997 and 2013, showing the municipalities, states (MG. Minas Gerais; PR. Parana; RS. Rio Grande do Sul; SC. Santa Catarina), coordinates (* Coordinate of the centroid of the municipality) and source of the records (LKJ. Records collected by the first author; personal communications: PSN. Pedro Scherer Neto; AER. Adrian Eisen Rupp). Tabla 1. Registros del churrín palustre entre 1997 y 2013 en los que se muestran los municipios, los estados (MG. Minas Gerais; PR. Parana; RS. Rio Grande do Sul; SC. Santa Catarina), las coordenadas (* Coordenadas del centroide del municipio) y los autores de los registros (LKJ. Registros recopilados por el primer autor; comunicaciones personales: PSN. Pedro Scherer Neto; AER. Adrian Eisen Rupp). Municipality

State Coordinate

Source

Bambuí

20° 14' S, 45° 58' W

Vasconcelos et al., 2008

Catas Altas

20° 07' S, 43° 27' W

Vasconcelos et al., 2008

Itabira

19° 36' 04" S, 43° 18' 03” W*

Silva, 2012

Mariana

20° 19' 46" S, 43° 19' 55" W*

Silva, 2011

Moeda

20° 19' 50" S, 43° 59' 35" W*

Franco, 2013

Morro do Pilar

19° 11' 05" S, 43º 23' 38" W

Faria, 2011

Morro do Pilar

19° 15' S, 43° 31' W

Araújo, 2011

Morro do Pilar

19° 15' S, 43° 31' W

Araújo, 2011

Morro do Pilar

19° 15' S, 43° 31' W

Araújo, 2011

Morro do Pilar

19° 15' S, 43° 31' W

Rodrigues et al., 2008

Ouro Preto

20° 23' 29" S, 43° 36' 39" W*

Silva, 2012

Santa Bárbara

20° 08' S, 43° 31' W

Vasconcelos et al., 2008

Santa Bárbara

20° 00' S, 43° 28' W

Vasconcelos et al., 2008

Santana do Riacho

19° 15' S, 43° 31' W

Vasconcelos et al., 2008

São João Batista do Glória

20° 28' S, 46° 26' W

Vasconcelos et al., 2008

Bituruna

26° 18' 03" S, 51° 30' 12" W

LKJ

Bituruna

26° 18' 09" S, 51° 30' 09" W

LKJ

Bituruna

26° 18' 20" S, 51° 29' 31" W

LKJ

Castro

24° 47' 58" S, 49° 50' 30" W*

IAP, 2009

Cruz Machado

25° 48' S, 51° 05' W

Straube et al., 2005

Cruz Machado

25° 47' S, 51° 07' W

Straube et al., 2005

Cruz Machado

25° 45' S, 51° 05' W

Straube et al., 2005

Curitiba

25° 37' 16" S, 49° 19' 47" W

LKJ

Curitiba

25° 28' 50" S, 49° 20' 02" W

PSN verbally 2011

Curitiba

25° 28' 59" S, 49° 11' 15" W

Straube et al., 2009

Curitiba

25° 35' 58" S, 49° 15' 59" W

Straube et al., 2009

General Carneiro

26° 18' 54" S, 51° 29' 26" W

LKJ

General Carneiro

26° 35' S, 51° 19' W

Straube et al., 2005

Guarapuava

25° 14' 13" S, 51° 17' 04" W

IAP, 2009

Guarapuava

25° 26' 00" S, 51° 13' 00" W

LKJ

Guarapuava

25° 25' 34'' S, 51° 13' 40" W

LKJ

Lapa

25° 48’ S, 50° 14' W

Bornschein et al., 2001

Palmas

26° 33' 25" S, 51° 34' 09" W

IAP, 2009

Palmeira

25° 26' S, 50° 12' W

Bornschein et al., 2001

Pinhais

25° 26' S, 49° 07' W

Bornschein et al., 2001

Piraí do Sul

24° 27' S, 49° 50' W

Bornschein et al., 2001

Piraquara

25° 27' S, 49° 07' W

Bornschein et al., 2001

Ponta Grossa

25° 14' 51" S, 50° 00' 35" W

PSN verbally 2011

Animal Biodiversity and Conservation 36.1 (2013)

Table 1. (Cont.)

Municipality

State

Coordinate

Source

Quatro Barras

25° 23' S, 49° 05' W

Bornschein et al., 2001

Quatro Barras

25° 23' 38" S, 49° 06' 06" W

PSN verbally 2011

Quatro Barras

25° 23' 38" S, 49° 06' 06" W

PSN verbally 2011

Quatro Barras

25° 23' 38" S, 49° 06' 06" W

PSN verbally 2011

São João do Triunfo

25° 47' S, 50° 13' W

Bornschein et al., 2001

São João do Triunfo

25° 50' S, 50° 14' W

Bornschein et al., 2001

São José dos Pinhais

25° 39' 52" S, 49° 05' 42" W*

Whittaker, 2011

São José dos Pinhais

25° 34' 15" S, 49° 03' 26" W

Athanas, 2008

São José dos Pinhais

25° 34' 12" S, 49° 03' 25" W

Luijendijk, 2010

São José dos Pinhais

25° 34' 59'' S, 49° 03' 37'' W

By, 1998

São José dos Pinhais

25° 36' S, 49° 09' W

Minns, 2002

São José dos Pinhais

25° 36' S, 49° 09' W

Bornschein et al., 2001

São José dos Pinhais

25° 36' S, 49° 06' W

Bornschein et al., 2001

São José dos Pinhais

25° 37' S, 49° 06' W

Bornschein et al., 2001

São José dos Pinhais

25° 30' S, 49° 09' W

Bornschein et al., 2001

São José dos Pinhais

25° 33' S, 49° 00' W

Bornschein et al., 2001

São José dos Pinhais

25° 34' S, 49° 03' W

Bornschein et al., 2001

São José dos Pinhais

25° 38' S, 49° 09' W

Bornschein et al., 2001

São José dos Pinhais

25° 36' S, 49° 10' W

Bornschein et al., 2001

Teixeira Soares

25° 21' S, 50° 25' W

Bornschein et al., 2001

Tijucas do Sul

25° 48' S, 49° 09' W

Bornschein et al., 2001

Tijucas do Sul

25° 47' S, 49° 09' W

Bornschein et al., 2001

Tijucas do Sul

25° 48' S, 49° 08' W

Bornschein et al., 2001

Tijucas do Sul

25° 51' S, 49° 06' W

Bornschein et al., 2001

Tijucas do Sul

25° 50' S, 49° 12' W

Bornschein et al., 2001

Tijucas do Sul

25° 47' S, 49° 08' W

Bornschein et al., 2001

Tijucas do Sul

25° 52' S, 49° 13' W

Bornschein et al., 2001

Bom Jesus

28° 28' 55" S, 50° 42' 48" W

Fontana et al., 2008

Cambará do Sul

29° 09' 34'' S, 50° 07' 35'' W

Patrial, 2006

Cambará do Sul

29° 09' 34'' S, 50° 07' 35'' W

Bencke et al., 2006

Rio Grande

32° 18' 16" S, 52° 25' 19" W

Jacobs, 2008a

Rio Grande

32° 18' 16" S, 52° 25' 19" W

Jacobs, 2008b

Rio Grande

32° 18' 16" S, 52° 25' 19" W

Mauricio, 2005

Rio Grande

32° 18' 57" S, 52° 25' 19" W

Andretti, 2010a

Rio Grande

32° 18' 57" S, 52° 25' 19" W

Andretti, 2010b

Viamão

30° 10' 02" S, 50° 52' 12" W*

Fenalti, 2011

Viamão

30° 06' S, 50° 52' W

Accordi et al., 2003

Água Doce

26° 45' 56" S, 51° 36' 46" W*

Corrêa et al., 2008

Campo Alegre

26° 05' 34" S, 49° 13' 38" W

LKJ

Campo Belo do Sul

27° 50' 51" S, 50° 46' 20" W*

Espínola, 2011

Campo Belo do Sul

27° 59' 45" S, 50° 52' 15" W

Rupp, 2010

Campo Belo do Sul

27° 58' 47" S, 50° 48' 29" W

Rupp, 2010

Jaraguá do Sul

26° 15' 37" S, 49° 13' 47" W

LKJ

Lages

28° 01' 09" S, 50° 20' 21" W*

Corrêa et al., 2008

Klemann Jr. & Vieira

Table 1. (Cont.)

Municipality

State

Coordinate

Source

Ponte Alta do Norte

27° 07' 39" S, 50° 22' 54" W

LKJ

Ponte Alta do Norte

27° 07' 06" S, 50° 24' 12" W

LKJ

Ponte Alta do Norte

27° 07' 41" S, 50° 22' 53" W

LKJ

Ponte Alta do Norte

27° 07' 15" S, 50° 25' 23" W

LKJ

Ponte Alta do Norte

27° 07' 33" S, 50° 23' 12" W

LKJ

Ponte Alta do Norte

27° 09' 19" S, 50° 22' 28" W

LKJ

Ponte Alta do Norte

27° 07' 30" S, 50° 25' 11" W

LKJ

Santa Cecília

27° 05' 30" S, 50° 23' 42" W

LKJ

Santa Cecília

26° 45' 08" S, 50° 21' 53" W

LKJ

Santa Cecília

26° 46' 00" S, 50° 22' 28" W

LKJ

São Cristóvão do Sul

27° 17' 08" S, 50° 18' 00" W

LKJ

São Cristóvão do Sul

27° 19' 21" S, 50° 16' 34" W

LKJ

São Cristóvão do Sul

27° 18' 38" S, 50° 17' 45" W

LKJ

São Cristóvão do Sul

27° 16' 41" S, 50° 19' 31" W

LKJ

São Cristóvão do Sul

27° 19' 42" S, 50° 16' 17" W

LKJ

Três Barras

26° 12' S, 50° 12' W

Corrêa et al., 2007

Urupema

27° 58' 28" S, 49° 58' 04" W

AER verbally 2011

found vocalizing in the marshes sampled, with an area from 3,589 to 30,422 m2, and this number remained constant in the two samples made in each marsh (table 2), reinforcing the efficiency of the method used. The estimated population size, based on the population density and in the area of occupancy, was 31,584 ± 7,140 mature individuals. Discussion Occurrence and environment The expected occurrence in the state of São Paulo, considered by Vasconcelos et al. (2008), was corroborated by the existence of the vegetal formations and ecological zones with known occurrences for the species. The inclusion of São Paulo state in the polygon of the extent of occurrence delimited herein is thus justified. The expected distribution for this state extends from the border with the state of Paraná to the Minas Gerais border, covering an area of approximately 150 km from east to west starting at the Serra do Mar (fig. 1). The presence of records in the tropical moist deciduous forest (3) and tropical rainforest (2), a few kilometers (11 km maximum) off the boundaries with the mountain tropical system, might be associated with the scale used for mapping the global ecological zones. It is thus convenient to consider

only the subtropical humid forest and the mountain tropical system as ecological zones for the species' occurrence. Occurrence of the marsh tapaculo in the vegetation types where the species were found (araucaria moist forest, semi–deciduous seasonal forest, grassland, pioneer formation zone and ecological tension areas between grassland and semi–deciduous seasonal forest) is directly associated with the presence of pioneer formation areas with fluvial influence, plant communities that occur throughout Brazil in floodplains and flooded depressions (Veloso et al., 1991). The vegetation structure in these communities is quite varied, although the species seems to be restricted to environments dominated by Cyperaceae and/or Poaceae (Bornschein et al., 1998, 2001; Vascocelos et al., 2008). These vegetal formations suffered greatly from the impact of the Brazilian Federal Government incentive program called Pro–Várzea (established in the 1970s to take advantage of the wetlands for agricultural production, financing the drainage of the wetlands), but are now protected by the Brazilian Forest Code and considered permanent preservation areas. However, the lack of specific data on the alteration of wetlands in Brazil makes it difficult to evaluate and measure the changes that have occurred herein. On the other hand, we can use the deforestation rates of the Atlantic Forest in the states where the

Animal Biodiversity and Conservation 36.1 (2013)

Marsh tapaculo records

Marsh tapaculo extent of occurrence

Vegetation types

0 40 80

160

240

320 km

Pioneer formation areas

Ecological tension areas

Deciduous seasonal forest

Semideciduous seasonal forest

Tropical moist forest

Araucaria moist forest

Grassland

Rivers and lakes

All other vegetation types

Fig. 1. Map of vegetation types with the location of marsh tapaculo records between 1997 and 2012 and the extent of occurrence delimited through the adjusted polygon method. Fig. 1. Mapa de los tipos de vegetación con la ubicación de los registros del churrín palustre entre 1997 y 2012 y la extensión de la presencia delimitada mediante el método del polígono ajustado.

marsh tapaculo occur to understand and measure the process of alteration in other vegetal formations associated with these forests. The deforestation rate was 1.23 from 2005–2008 and 0.45% from 2008–2010 in the state of Minas Gerais, 0.51 and 0.17% in Paraná, 0.31 and 0.18% in Rio Grande do Sul, 1.19 and 0.17% in Santa Catarina, and 0.11 and 0.02% in São Paulo (Fundação SOS Mata Atlântica & Instituto Nacional de Pesquisas Espaciais, 2009, 2011).

Considering these rates of deforestation, the reduction in the extent of suitable habitat and consequently the reduction in the species' population cannot be considered significant (≥ 30%). The species' sensitivity to environmental changes, specially caused by fire and by natural or anthropogenic changes in the vegetal structure, is still not fully understood, and therefore does not allow a more significant discussion about the impact of these changes on populations of the species.

Klemann Jr. & Vieira

Table 2. State of the marshes used to calculate the territory size of marsh tapaculo, with the number of vocalizing individuals at the same time after playback on two sample dates (March 2011 and April 2012), measures of marsh areas (Ma, in m2) and territory size (Ts, in m2): S1. Sample 1; S2. Sample 2. (For abbreviations of states, see table 1.) Tabla 2. Estado de las marismas utilizadas para calcular el tamaño del territorio del churrín palustre, número de individuos que emiten sonidos al mismo tiempo después de reproducir una grabación en las dos fechas del estudio (marzo de 2011 y abril de 2012), mediciones de las superficies de las marismas (Ma, en m2 y tamaño del territorio (Ts, en m2): S1. Muestra 1; S2. Muestra 2. (Para las abreviaturas de los estados, ver tabla 1.) Vocalizing indiv. S1

State

6,990

3,589

12,614

6,307

18,121

4,530

24,203

6,051

9,535

4,768

8,363

4,182

30,422

6,084

Mean

5,313

Extent of occurrence, area of occupancy, territory size and population size The increasing number of records in the 14 years since the species’ description shows a significant growth in the extent of occurrence, going from only one state with records until 2001 (Bornschein et al., 2001) to four states in 2008 (Vasconcelos et al., 2008). The extent of occurrence presented by BirdLife International (2012), 490 km2, is based on only 20 locations of occurrence, a much smaller number than that presented here: 70 locations and 42 municipalities. This difference in the amount of data used could explain the greater extent of occurrence, area of occupancy and population size presented here. The territory size found for marsh tapaculo (5,313 ± 1,201 m2), another factor that contributes to the increase in the estimated population size, is consistent with the value obtained for a species with similar body mass and environment (marsh antwren Stymphalornis acutirostris): an average of 2,500 m2 (Reinert et al., 2007) and 7,000 m2 (Reinert, 2008) in

tidal marshes and 32,000 m2 in saw grass marshes (Reinert et al., 2007). There is no available information about the size of territory in relation to other species with similar body size and environment. Averages for territory sizes of forest species with body mass of 10 to 15 g range from 6,000 to 150,000 m2 (Greenberg & Gradwohl, 1985; Silva, 1988; Terborgh et al., 1990; Skutch, 1996). An unpublished estimate, from Banhado do Maçarico in the state of Rio Grande do Sul (the southernmost locality known for the species), found a population density of 0.5 individuals of marsh tapaculo per hectare (40,000 m2 territory size) (BirdLife International, 2012). The difference between this value and that obtained here could be explained by the location of marshes sampled in relation to the extent of occurrence of species and/or by differences in vegetation structure of marshes sampled. Such variation can also be observed in the territory size of marsh antwren (2,500 and 32,000 m2), which varies with the vegetation structure of the marshes (Reinert et al., 2007). The area of occupancy calculated for marsh tapaculo (84 km2) represents only 11.24% of the environment considered suitable for this species in one state, Paraná (747 km2) (Bornschein et al., 2001). Furthermore, the area of occupancy obtained through the grid cell method is influenced by the sampling intensity (Grueter et al., 2009), generating underestimated values when the species is not recorded in points where it occurs. It is therefore expected that the area of occupancy will increase as the number of records increase. This projection is confirmed by the increasing number of new records in recent years and from unpublished information. The growing numbers of searches for the species inside and outside the known range of occurrence will also contribute to increasing the range of occurrence and area of occupancy. As it is expected that the area of occupancy will extend, the population size for marsh tapaculo is also expected to increase. Thus, given the significant growth of knowledge for the species’ distribution presented here, the population size obtained is much larger than the estimates presented previously, 250 to 999 mature individuals (BirdLife International, 2012), and this number tends to increase as new records are made in different localities. Threat category The species was classified as ‘Endangered’ based on the little information available covering aspects of distribution and population. The information compiled and the results presented here show that a revaluation of the threat category is needed. The urgency for this review and for a change in the threat category, due to the fact that the original category is considered misclassified, is recommended by IUCN (2001). Thus, considering the current extent of occurrence (> 20,000 km2), area of occupancy (< 100 km2), population size (> 10,000 individuals) and habitat conditions, the criteria needed to include the marsh tapaculo in the IUCN red list (BirdLife International, 2012) (criteria A3c+4c, B1ab (i, ii, iii, iv, v), C2a(i))

Animal Biodiversity and Conservation 36.1 (2013)

and in the Brazilian red list (Machado et al., 2005) (criteria B2ab) is not obtained for any of the threat categories ('Critically Endangered', 'Endangered' or 'Vulnerable'). Despite the small size of the area of occupancy, caused by the disconnected characteristics of the species' habitat, the number of locations where the marsh tapaculo was recorded was much greater than ten, and no fluctuations in the extent of occurrence, area of occupancy, number of locations or sub–populations, and number of mature individuals were observed. Revaluation of the species' threat category is thus strongly recommended. Acknowledgements Thanks to Adrian Eisen Rupp, Pedro Scherer Neto and Raphael Eduardo Fernandes Santos for the marsh tapaculo records kindly provided, and special thanks to Tiago Venâncio Monteiro, Liliane Klemann and Zora Morgenthaler for the translation of the text. References Accordi, I. A., Hartz, S. M. & Ohlweiler A., 2003. O sistema Banhado Grande como uma área úmida de importância internacional. II Simpósio de Áreas Protegidas – conservação no âmbito do Cone Sul. Pelotas, Brazil. Adams, L. & Davis, S. D., 1967. The internal anatomy of home range. Journal of Mammalogy, 48: 529–536. Andretti, C. B., 2010a. WA383164, Scytalopus iraiensis. Wiki Aves – The Encyclopedia of Brazilian Birds. www. wikiaves.com/383164 (Accessed 3 October 2011). – 2010b. XC82545, Scytalopus iraiensis. Xeno–Canto – Bird sounds around the world. www.xeno–canto. org/82545 (Accessed 3 October 2011). Araujo, F. M., 2011. WA318746, WA291931, WA300586, WA300597, WA300601, Scytalopus iraiensis. Wiki Aves – The Encyclopedia of Brazilian Birds. www.wikiaves.com/318746 (Accessed 3 October 2011). Athanas, N., 2008. XC22833, Scytalopus iraiensis. Xeno–Canto – Bird sounds around the world. www. xeno–canto.org/22833 (Accessed 3 October 2011). Bencke, G. A., Maurício, G. N., Develey, P. F. & Goerck, J. M., 2006. Áreas importantes para a conservação das aves no Brasil. Parte I – estados do domínio da Mata Atlântica. SAVE, São Paulo. BirdLife International, 2012. Species factsheet: Scytalopus iraiensis. www.birdlife.org/datazone/ speciesfactsheet.php?id=30032 (Accessed 9 October 2012). Bornschein, M. R., Reinert, B. L. & Pichorim, M., 1998. Descrição, ecologia e conservação de um novo Scytalopus (Rhinocryptidae) do sul do Brasil, com comentários sobre a morfologia da família. Ararajuba, 6: 3–36. – 2001. Novos registros de Scytalopus iraiensis. Nattereria, 2: 29–33.

Burgman, M. A. & Fox, J. C., 2003. Bias in species range estimates from minimum convex polygons: implications for conservation and options for improved planning. Animal Conservation, 6: 19–28. By, R. A. de, 1998. XC10687, Scytalopus iraiensis. Xeno–Canto – Bird sounds around the world. www.xeno–canto.org/10687 (Accessed 3 October 2011). Corrêa, L., Bazílio, S., Woldan, D. & Boesing, A. L., 2008. Avifauna da Floresta Nacional de Três Barras (Santa Catarina, Brasil). Atualidades Ornitológicas, 143: 38–41. Corrêa, L. & Woldan, D. R. H., 2007. Registro de Scytalopus iraiensis na Floresta Nacional de Três Barras, planalto norte do estado de Santa Catarina, Brasil.In: Livro de Resumos do XV Congresso Brasileiro de Ornitologia: 204 (C. S. Fontana, Ed.). EdiPUCRS, Porto Alegre. Corrêa, L., Woldan, D. R. H. & Bazílio, S., 2007. Registros de aves raras e ameaçadas no planalto norte de Santa Catarina, Brasil. In: Livro de Resumos do XV Congresso Brasileiro de Ornitologia: 203 (C. S. Fontana, Ed.). EdiPUCRS, Porto Alegre. Espínola, C., 2011. WA314543, Scytalopus iraiensis. Wiki Aves – The Encyclopedia of Brazilian Birds. www.wikiaves.com/314543 (Accessed 3 October 2011). Faria, L., 2011. XC76250, Scytalopus iraiensis. Xeno–Canto – Bird sounds around the world. www. xeno–canto.org/76250 (Accessed 3 October 2011). Fenalti, P. R., 2011. WA478969, Scytalopus iraiensis. Wiki Aves – The Encyclopedia of Brazilian Birds. www.wikiaves.com/478969 (Accessed 16 January 2012). Ferry, C. & Frochot, B., 1970. L’avifaune nidiﬁcatrice d’une foret de Chenes pedoncules en Bourgogne: Etude de deux successions ecologiques. Rev Ecol–Terre Vie, 24: 153–250. Fontana, C. S., Rovedder, C. E., Repenning, M. & Gonçalves, M. L., 2008. Estado atual do conhecimento e conservação da avifauna dos Campos de Cima da Serra do sul do Brasil, Rio Grande do Sul e Santa Catarina. Revista Brasileira de Ornitologia, 16(4): 281–307. FRA, 2000. Global Ecological Zones. Forest Resource Assessment. www.fao.org/geonetwork/srv/ en/main.home (Accessed 28 September 2011). Franco, E. S., 2013. WA844493, Scytalopus iraiensis. Wiki Aves – The Encyclopedia of Brazilian Birds. www.wikiaves.com/844493 (Accessed 8 January 2013). Fundação SOS Mata Atlântica & Instituto Nacional De Pesquisas Espaciais, 2009. Atlas dos remanescentes florestais da mata atlântica período 2005–2008. http://mapas.sosma.org.br/dados/ (Accessed 22 December 2011) – 2011. Atlas dos remanescentes florestais da mata atlântica período 2008–2010. http://mapas.sosma. org.br/dados/ (Accessed 22 December 2011) Greenberg, R. & Gradwohl, J., 1985. A comparative study of the social organization of Antwrens on Barro Colorado Island, Panama. Ornithological

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Biodiversitas, Belo Horizonte. Maurício, G. N., 2005. Taxonomy of southern populations in the Scytalopus speluncae group, with description of a new species and remarks on the systematics and biogeography of the complex (Passeriformes: Rhinocryptidae). Ararajuba, 13: 7–28. Mills, M. G. L. & Gorman, M. L., 1987. The scent– marking behaviour of the Spotted hyaena Crocuta crocuta in the southern Kalahari. Journal of Zoology, 212: 483–497. Minns, J., 2002. XC5976, Scytalopus iraiensis. Xeno–Canto – Bird sounds around the world. www. xeno–canto.org/5976 (accessed 3 October 2011). Mohr, C. O., 1947. Table of equivalent populations of North American small mammals. American Midland Naturalist, 37: 223–249. Nichols, J. D., Hines, J. E., Sauer, J. R., Fallon, F. W., Fallon, J. E. & Heglund, P. J., 2000. A double–observer approach for estimating detection probability and abundance from point counts. Auk, 117: 393–408. Ostro, L. E. T., Young, T. P., Silver, S. C. & Koontz, F. W., 1999. A geographic information system (GIS) method for estimating home range size. Journal of Wildlife Management, 63: 748–755. Patrial, E., 2006. XC18297, Scytalopus iraiensis. Xeno–Canto – Bird sounds around the world. www. xeno–canto.org/18297 (Accessed 3 October 2011). Perrins, C. M. & Birkhead, T. R., 1983. Avian Ecology. Glasgow & Blackie et Son, London. Powell, R., 2000. Animal home ranges and territories and home range estimators. In: Research techniques in animal ecology: controversies and consequences: 65–110 (L. Boitani, & T. Fuller, Eds.). Columbia Univ. Press, New York. Raposo, M. A., Stopiglia, R., Loskot, V. & Kirwan, G. M., 2006. The correct use of the name Scytalopus speluncae (Ménétriés, 1835), and the description of a new species of Brazilian tapaculo (Aves: Passeriformes: Rhinocryptidae). Zootaxa, 1271: 37–56. Reinert, B. L., 2008. Ecologia e comportamento do bicudinho–do–brejo (Stymphalornis acutirostris Bornschein, Reinert e Teixeira, 1995; Aves, Thamnophilidae). Ph. D. Thesis, Univ. Estadual Paulista 'Julio de Mesquita Filho'. Reinert, B. L., Bornschein, M. R. & Firkowski, C., 2007. Distribuição, tamanho populacional, hábitat e conservação do bicudinho–do–brejo Stymphalornis acutirostris Bornschein, Reinert e Teixeira 1995 (Thamnophilidae). Revista Brasileira de Ornitologia, 15: 493–519. Rodrigues, M., Mariana, L. M., Freitas, G. H. S. de, Dias, D. F., Mesquita, E. P., Cavalcanti, M., Rocha, R. P., Rodrigues, L. da C., Ferreira, J. D. & Diniz, F. C., 2008. Campos rupestres úmidos e secos da cadeia do espinhaço: uma abordagem ornitológica. In: Livro de Resumos do XVI Congresso Brasileiro de Ornitologia: 194 (T. Dornas & M. de O. Barbosa, Eds.). ECOAVES–UFT, Palmas. Rosenstock, S. S., Anderson, D. R., Giesen, K. M., Leukering, T. & Carter, M. F., 2002. Landbird counting techniques: Current practices and an alternative. Auk, 119: 46–53.

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Rupp, A. E., 2010. WA181812, Scytalopus iraiensis. Wiki Aves – The Encyclopedia of Brazilian Birds. www. wikiaves.com/181812 (Accessed 3 October 2011). – 2010. WA244703, Scytalopus iraiensis. Wiki Aves – The Encyclopedia of Brazilian Birds. www.wikiaves. com/244703 (Accessed 3 October 2011). Silva, J. C., 2011. WA487684, Scytalopus iraiensis. Wiki Aves – The Encyclopedia of Brazilian Birds. www. wikiaves.com/487684 (Accessed 16 January 2012). – 2012. WA601765, Scytalopus iraiensis. Wiki Aves – The Encyclopedia of Brazilian Birds. www.wikiaves. com/601765 (Accessed 16 May 2012). – 2012. WA706138, Scytalopus iraiensis. Wiki Aves – The Encyclopedia of Brazilian Birds. www.wikiaves. com/706138 (Accessed 8 January 2013). Silva, J. M. C., 1988. Aspectos da ecologia e comportamento de Formicivora g. grisea (Boddaert, 1789) (Aves: Formicariidae) em ambientes amazônicos. Revista Brasileira de Biologia, 48(4): 797–805. Simons, T. R., Alldredge, M. W., Pollock, K. H. & Wettroth, J. M., 2007. Experimental analysis of the auditory detection process on avian point counts. Auk, 124(3): 986–999. Skutch, A. F., 1996. Antbirds and ovenbirds. Univ. of Texas Press, Austin. Straube, F. C., Carrano, E., Santos, R. E. F., Scherer– Neto, P., Ribas, C. F., Meijer, A. A. R. de, Vallejos, M. A. V., Lanzer, M., Klemann–Júnior, L., Aurélio–Silva,

M., Urben–Filho, A., Arzua, M., Lima, A. M. X. de, Deconto, L. R., Bispo, A. Â., Jesus, S. de & Abilhôa, V., 2009. Aves de Curitiba–Coletânea de registros. Hori Consultoria, Curitiba. Straube, F. C., Krul, R. & Carrano, E., 2005. Coletânea da avifauna da região sul do estado do Paraná (Brasil). Atualidades Ornitológicas, 125: 10–71. Terborgh, J., Robinson, S. K., Parker Iii, T. A., Munn, C. A. & Pierpont, N., 1990. Structure and organization of an Amazonian forest bird community. Ecological Monographs, 60(2): 213–238. Thompson, W. L., 2002. Towards reliable bird surveys: accounting for Individuals present but not detected. Auk, 119(1): 18–25. Vasconcelos, M. F., Maurício, G. N., Kirwan, G. M. & Silveira, L. F., 2008. Range extension for marsh tapaculo Scytalopus iraiensis to the highlands of Minas Gerais, Brazil, with an overview of the species’ distribution. Bulletin BOC, 128(2): 101–106. Veloso, H. P., Rangel Filho, A. L. R. & Lima, J. C. A., 1991. Classificação da vegetação brasileira, adaptada a um sistema universal. IBGE, Departamento de Recursos Naturais e Estudos Ambientais, Rio de Janeiro. White, G. & Garrott, R., 1990. Analysis of wildlife radio tracking data. Academic, San Diego. Whittaker, A., 2011. WA511140, Scytalopus iraiensis. Wiki Aves – The Encyclopedia of Brazilian Birds. www. wikiaves.com/511140 (Accessed 16 January 2012).

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Drusia (Escutiella) alexantoni n. sp. (Gastropoda, Pulmonata, Parmacellidae), a new terrestrial slug from the Atlantic coast of Morocco A. Martínez–Ortí & V. Borredà

Martínez–Ortí, A. & Borredà, V., 2013. Drusia (Escutiella) alexantoni n. sp. (Gastropoda, Pulmonata, Parmacellidae), a new terrestrial slug from the Atlantic coast of Morocco. Animal Biodiversity and Conservation, 36.1: 59–67. Abstract Drusia (Escutiella) alexantoni n. sp. (Gastropoda, Pulmonata, Parmacellidae), a new terrestrial slug from the Atlantic coast of Morocco.— We describe a new parmacellid, Drusia (Escutiella) alexantoni n. sp. from the Moroccan Atlantic coast. The species most closely related to the new taxon are D. (E.) deshayesii and D. (D.) valenciennii. The new parmacellid differs from D. (E) deshayesii mainly by the presence of external spots and bands on both the back and the shield, a reproductive system with uneven atrial appendices of the horn–shaped organ, and a different reticulated pattern of the inner epiphallus. It differs from D. (D.) valenciennii mainly for the appearance of the shell and the pattern and disposition of the bumps inside the penis, the presence of an elbow–shape in this organ, and the reticulated appearance of the inner wall of the epiphallus. An updated dichotomous key of the family Parmacellidae is provided. Key words: Slug, Parmacellidae, Drusia (Escutiella) alexantoni, New species, Morocco, North Africa. Resumen Drusia (Escutiella) alexantoni sp. n. (Gastropoda, Pulmonata, Parmacellidae), una nueva babosa del litoral atlántico de Marruecos.— Se describe un nuevo parmacélido, Drusia (Escutiella) alexantoni sp. n., de la costa atlántica marroquí. Las especies más afines al nuevo taxon son D. (E.) deshayesii y D. (D.) valenciennii. De la primera se diferencia por presentar externamente manchas y bandas sobre el dorso y escudo, un aparato reproductor con apéndices atriales del órgano corniforme bastante desiguales, y por el distinto aspecto del reticulado del interior del epifalo. De D. (D.) valenciennii se diferencia principalmente por la forma de su concha, así como por el aspecto y la disposición de los mamelones del interior del pene, la presencia de un marcado acodamiento en este órgano, así como por el aspecto reticulado del interior del epifalo. Se proporciona además una clave dicotómica actualizada de la familia Parmacellidae. Palabras clave: Babosa, Parmacellidae, Drusia (Escutiella) alexantoni, Nueva especie, Marruecos, Norte de África. Received: 10 I 13; Conditional acceptance: 31 I 13; Final acceptance: 6 II 13 Alberto Martínez–Ortí & Vicent Borredà, Museu Valencià d’Història Natural i iVBiotaxa, L’Hort de Feliu–Alginet, P. O. Box 8460, 46018 València, Espanya (Spain) and Dept. de Zoologia, Fac. de Ciències Biològiques, Univ. de València, c/ Dr. Moliner 50, 46100 Burjassot, València, Espanya (Spain). Corresponding author: A. Martínez–Ortí. E–mail: amorti@uv.es

ISSN: 1578–665X

Introduction In a recent article (Martínez–Ortí & Borredà, 2012) we revised the systematics of the family Parmacellidae P. Fischer, 1856 and we proposed a new systematic scenario for this family, which would be formed by four genera: Candaharia Godwin–Austen, 1888 (2 subgen., 4 spp.), from Central Asia; Cryptella Webb et Berthelot, 1833 (7 spp.) from the Canary Islands; Parmacella Cuvier, 1804 (2 spp.) from Libya and Egypt; and Drusia Gray, 1855 (2 subgen., 4 spp.), with a wide distribution detailed below. In the previously mentioned paper, the genus Drusia Gray, 1855 was divided into two subgenera: D. (Escutiella) Martínez–Ortí & Borredà, 2012 and D. (Drusia) s. str. The subgenus D. (Drusia) includes three species: D. (D.) valenciennii (Webb et Van Beneden, 1836), from the South of the Iberian peninsula; D. (D.) tenerifensis (Alonso, Ibáñez & Díaz, 1985), from Tenerife and D. (D.) ibera (Eichwald, 1841) from the Caucasus–Caspian Sea area. Subgenus D. (Escutiella) was described to include only one species: D. (E.) deshayesii (Moquin–Tandon, 1848), from Algeria and Northern Morocco. In January 2011, we carried out a malacological prospection along the Moroccan Atlantic coast, collecting numerous specimens of a parmacelle which we propose as a new species to be included in the subgenus D. (Escutiella). Results After a detailed morpho–anatomical study of the collected specimens we observed that they corresponded to a parmacelle closely related to the species Drusia (Drusia) valenciennii and Drusia (Escutiella) deshayesii, particularly to the latter, but we believe it is a new species, and we propose naming it Drusia (Escutiella) alexantoni n. sp. Family Parmacellidae P. Fisher, 1856 Genus Drusia Gray, 1855 Subgenus D. (Escutiella) Martínez–Ortí & Borredà, 2012 Drusia (Escutiella) alexantoni n. sp. Typical locality Road from Marrakech to Essaouira, 12 km before Essaouira, Taftchet. Essaouira. Morocco (UTM = 29RMQ4388) (January 2, 2011). Collectors: A. Martínez–Ortí, A. López Alabau and A. Pérez Ferrer (MVHN–100111GH01). Other localities Road of Essaouira to Agadir–Smimov, Smimov. Essaouira. Morocco (UTM = 29RMQ3274) (January 3, 2011) (Collectors: A. Martínez–Ortí, A. López Alabau and A. Pérez Ferrer) (MVHN–100111GH02; five specimens); Agadir–Ida–Outanane, close to a lake with coots beside the road (February 6, 2009) (Martínez, 2009) (UTM = 29RMP46).

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Type material Formed by 29 specimens. The holotype is deposited at the Museu Valencià d’Història Natural (Valencia, Spain) with the code MVHN–100111GH01a. There are 13 paratypes (in ethanol 70%) with the code MVHN–100111GH01b and four paratypes (in ethanol 96%) with the code MVHN–100111GH01c, all at the same institution. In addition, three paratypes (in ethanol 70%) were deposited at the Museu de Ciències Naturals de Barcelona (Zoologia, MZB) with the code MZB 2012–0728; three paratypes (in ethanol 70%) at the Nationaal Natuurhistorisch Museum–Naturalis of Leiden (The Netherlands) with the code RMNH. MOL.323195; three paratypes (in ethanol 70%) at the Museo Nacional de Ciencias Naturales of Madrid (Spain) with the code MNCN–15.05/60078; and two paratypes (in ethanol 70%) at the Senckenberg Museum of Frankfurt am Main (Germany) with the code SMF 341354. Etymology Species dedicated to Alejandro Pérez Ferrer and Antonio López Alabau, co–collectors of the studied specimens and enthusiastic Valencian amateurs of the malacology. Common name Slug of Barbary; Babosa de Berbería; Limace de Berberie. Diagnosis Parmacelle of great size. Young specimens present an olive brown dorsum with black lines and spots, especially on the shield, while adult specimens are light orange–brown and with lighter lines and spots. Toward the back of the shield, multiple lines or black bands of different thicknesses converge on the protoconch showing individual pattern variation (character less patent in adults). This protoconch is bright greenish, covered in adults and protruding slightly on the body surface of young individuals. Inside the reproductive atrium there is a thick ligula that extends inside the largest of the horn–shaped accessorial appendices. It has a penis with a lateral bulge, giving it an elbow–like shape and it has two thick internal bumps. The interior of the epiphallus has a characteristic reticulated form with thick longitudinal folds that can spread out. Between these folds there are other less patent transverse folds that are almost perpendicular. External appearance (figs. 1–11): slug of the family Parmacellidae with external features characteristic of this family: large, rough skin, and large, granular shield with the pneumostome in its right posterior portion. Light orange dorsal keel on the caudal part of the animal. Orange dark keel clearly visible in the posterior part of the body, especially in well–developed adult specimens. Very acuminated tail. Foot is of aulacopod type and the sole is light in colour. Caudal gland absent. Adult individuals reach 15 cm in length. Young individuals present a dorsal olive brown background with black lines and spots, especially on the shield; dorsal black bands or lines converge

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1 cm 3 4 5 mm 5

1 cm 8

5 mm

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Figs. 1–11. Drusia (E.) alexantoni n. sp.: 1–2. Adult holotype; 3. Juvenile holotype; 4–5. Two paratypes; 6–7. Young paratypes showing the protoconch; 8–10. Shield pattern variability of three paratypes; 11. View of a group of several paratypes in the type locality. Figs. 1–11. Drusia (E.) alexantoni sp. n.: 1–2. Holotipo adulto; 3. Holotipo juvenil; 4–5. Dos paratipos; 6–7. Paratipos juveniles mostrando la protoconcha; 8–10. Variabilidad de los dibujos de los escudos de tres paratipos; 11. Vista de un grupo de varios paratipos en la localidad tipo.

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600 μm 17

1 mm 15

60 μm

1 mm 16

1 mm

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Figs. 12–18. Shell of Drusia (E.) alexantoni n. sp.: 12. Paratype (21.5 mm length) (digital photography); 13–14. Protoconch (scanning electron microscope); 14. Detail of the nucleus of the protoconch; 15–16. Posterior region of the shell showing the anchorage denticulation for the muscles; 17–18. Aspect of the surface of the protoconch; 18. Detail of the irregularly reticulated protoconch. Figs. 12–18. Concha de Drusia (E.) alexantoni sp. n.: 12. Paratipo (21,5 mm de longitud) (fotografía digital); 13–14. Protoconcha (microscopio electrónico de barrido); 14. Detalle del núcleo de la protoconcha; 15–16. Región posterior de la concha, con denticulación para el anclaje muscular; 17–18. Aspecto de la ornamentación de la superficie de la protoconcha; 18. Detalle del reticulado irregular de la protoconcha.

toward the shield end, having individual pattern variation. The greenish, bright protoconch is slightly protruded in young individuals and even in sub–adult specimens (figs. 6–7). In well–developed adults, the overall tone of the body is light orange brown, with more visible bands and spots found only on the edge of the shield, while the rest of the dorsum shows a uniform appearance. In general, adult coloration is lighter than in younger animals. Shell (figs. 6–7, 12–18): the shell is located under the mantle in the posterior part of the shield. It consists of a protoconch, from where a spiral begins, attached to a flat lamina, the limacella (or spatula). The protoconch is greenish, shiny, smooth, and relatively wide. The spiral is clearly visible. The limacella is white, slightly curved and paddle–shaped; it is slightly narrow in comparison

and not strictly flat, being more cupped than in other species of the family. The protoconch protrudes slightly from the posterior end of the mantle in young and sub– adult specimens; it is well–developed and presents a well–marked oval–circular opening (figs. 15–16). In the outer flange an arrowhead–shaped, anchoring tooth is appreciable (figs. 15–16). Although at a glance the protoconch looks smooth and glistening, high magnification reveals a characteristic form, consisting of longitudinal and transverse lines forming an irregular grid in some areas (figs. 17–18). The size of the shell from two of the adult paratypes varies from 12.0 to 14.0 mm in width and from 21.5 to 24.0 mm in length. Reproductive system (figs. 19–24): hermaphrodite gland partly covered by digestive organs is bilobed and formed by irregular acini. In young specimens it

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2 mm

2 mm 24

1 mm

2 mm

1 mm

Figs. 19–24. Reproductive system of Drusia (E.) alexantoni n. sp.: 19, 22–24. A paratype genitalia: 19. Complete genitalia; 22. Atrium containing the ligula; 23. Detail of the interior of the penis showing the bumps; 24. Form of the inner wall of the epiphallus. 20–21. Penis of another paratype. Figs. 19–24. Aparato reproductor de Drusia (E.) alexantoni sp. n.: 19, 22–24. Genitalia de un paratipo: 19. Genitalia completa; 22. Atrio con la ligula; 23. Detalle del interior del pene mostrando los mamelones. 24. Ornamentación de la pared interior del epifalo. 20–21. Pene de otro paratipo.

is lighter and in adults it is darker in colour, greyish, with the same colour as the hepatopancreas. Hermaphrodite duct long and winding. Very large, triangular, whitish and irregular albumen gland, larger than in D. (E.) desayesii and D. (D). valenciennii. Ovispermiduct relatively short, shorter than the albumen gland; distally it separates into feminine and masculine ducts. The masculine duct consists of vas deferens, epiphallus and penis, and together is longer than the ovispermiduct. The vas deferens is flared at its distal part,

turning into the epiphallus, which presents a series of very thick longitudinal folds that can spread out along with other transverse, perpendicular, some of them oblique, less patent folds which give it a reticular appearance interiorly (fig. 24). This reticular appearance is similar to that of D. (E.) deshayesii, although this species has both the transverse and longitudinal folds similarly well–marked. The retractor muscle is inserted in the distal part of the epiphallus and it enlarges markedly turning into the penis. The

Martínez–Ortí & Borredà

700 μm

200 μm

Figs. 25–26. Spermatophore of Drusia (E.) alexantoni n. sp.: 25. General view; 26. Anchoring disk detail. Figs. 25–26. Espermatóforo de Drusia (E.) alexantoni sp. n.: 25. Vista general; 26. Detalle del disco de anclaje.

penis has a lateral protrusion close to the retractor muscle, giving it an elbow–like shape. Interiorly, the penis is completely covered with tight papillae. Inside the penis, in its proximal part, there is a bump next to the area of insertion of the muscle retractor (figs. 21, 23). Another larger bump is present in a distal position inside the elbow area. No complete spermatophores have been recovered (figs. 25–26). Inside the bursa copulatrix of four adult paratypes occurred several spermatophores (up to four in one of them), partially digested but quite complete. The spermatophores have the characteristics of the parmacelle morphology, and they are formed by a spiral from which a long filament emerges ending in a star–shaped fixing disk. We did not find entire anchoring disks whose morphology is a character of possible taxonomic value among the partially digested spermatophores, but some of them fairly complete (fig. 26). The female duct begins with a short and cylindrical free oviduct which ends in a widened structure which also converges at the duct of the bursa copulatrix. This widened structure is smooth and ovoid, with a hemispherical bulge in front of the

end of the short bursa duct; the bursa is rather large and has very thin walls, although its size and shape vary greatly depending on the presence and degree of digestion of the spermatophores (fig. 19). The widened area increases its width becoming more glandular in aspect, having a bean–shape; it is the so–called perivaginal gland. The vagina is surrounded by this gland and ends in the atrium, which is rather short and has two conspicuous appendices attached, unequal in size and shape (figs. 19, 22). They are the atrial appendices; together they constitute the corniform organ, which has an irregular croissant shape. In the interior of the atrium, as is typical in the genus Drusia, there is a highly developed fleshy ligula that expands through the larger corniform organ appendix (Martínez–Ortí & Borredà, 2012) (fig. 22). Other characters (figs. 27–36): jaw of oxygnathous type and crescent–shaped (figs. 27–29), similar to that of D. (E.) deshayesii. In addition, it has a serrated edge, visible as tiny teeth at high magnification (fig. 29). The radulae of two examined paratypes

Animal Biodiversity and Conservation 36.1 (2013)

700 μm

300 μm

29 200 μm

1 mm

20 μm 34

20 μm

40 μm

30 μm

60 μm

Figs. 27–36. Jaw and radula of Drusia (E.) alexantoni n. sp.: 27–29. Jaw: 27. Paratype; 28–29. Other paratype; 29. Detail of the serrated edge. 30–35. Radula; 30. General view of the radula; 31. Central tooth and first lateral teeth; 32–33. Central tooth; 34. Lateral teeth next to the central tooth; 35. Transition from the lateral teeth toward the edge of the radula. 36. Last lateral teeth. Figs. 27–36. Mandíbula y rádula de Drusia (E.) alexantoni sp. n.: 27–29. Mandíbula: 27. Paratipo; 28–29. Otro paratipo; 29. Detalle del borde aserrado. 30–35. Rádula; 30. Vista general de la rádula; 31. Diente central y primeros laterales; 32–33. Diente central; 34. Dientes laterales próximos al diente central; 35. Transición de los dientes laterales hacia el borde de la rádula. 36. Últimos dientes laterales.

Martínez–Ortí & Borredà

New key for the determination of the family Parmacellidae P. Fischer, 1856. Nueva clave para la determinación de la familia Parmacellidae P. Fischer, 1856.

1 Vagina surrounded by a perivaginal gland not thickened and provided with a long finger–shape caecum Vagina with a swollen perivaginal gland, well–developed and bean–shaped. No caecum. 2 Genital atrium without appendices. Bursa copulatrix without thickening Genital atrium with two appendices, or at least one. Duct of the bursa with a thickening where the spermatophores are attached 3 Atrial appendices of similar size. Elongated and well–developed distal part of the atrium from the insertion of appendices to the genital pore. Without intraatrial stimulators, only fleshy folds, with small ridges on its wall Atrial appendices of different size. Short distal part of the atrium. One or more intraatrial large and fleshy stimulator folds 4 Ornamented protoconch with small parallel spiral grooves. Very long epiphallus with two bends Smooth protoconch. Epiphallus shorter and with a single curvature 5 Adults presenting dorsum with a shield that has dark stains and/or bands. Smooth penis without extrusion. Interior of the epiphallus not reticulated. Protoconch amber coloured and limacella in form of broad paddle Adults with dorsum and shield of uniform reddish–brown colour, or only with small lines at the end of the shield. Epiphallus internally reticulated. Penis with side extrusion, sometimes elbow–shaped. Greenish protoconch and a little wide limacella in the form of elongated paddle. Morocco and Algeria 6 Shell with a spatula (limacella) shaped shovel, very wide. Animals of large size (70–95 mm in ethanol). Anchoring disk of the spermatophore curved like an umbrella. Tenerife, Canary Islands Shell with spatula oval, much more narrow. Specimens of smaller size. Anchoring disk of the spermatophore almost flat 7 Wide spatula (limacella) of the shell (long/wide < 1.60). Stimulator fold in the Interior of the atrium thin and not very developed. Georgia, Kazakhstan, and other countries in the E of the Caspian Sea Spatula much narrower (long/wide > 1.85). Atrial appendices of very different sizes, sometimes only one. Stimulator fold of the atrium unique, pleated and very thick, occupying almost all of the intraatrial space. South of the Iberian peninsula, Spain and Portugal 8 Juvenile with dorsum and shield with black bands and spots, which tend to disappear in adults. Atrial appendices of the corniform organ quite unequal. Interior of the epiphallus with thick reticulate. Huge albumen gland. Atlantic coast of Morocco, Essaouira to Agadir Dorsum and shield, both juveniles and adults, of reddish–brown uniform colour, no bands or spots. Slender reticulate inside the epiphallus. Only slightly unequal atrial appendices. Northern Morocco and Algeria

Candaharia (Central Asia) 2 Cryptella (Canary Islands) 3

Parmacella (Libya, Egypt) Drusia

P. festae P. olivieri

D. (Drusia) s. str.

D. (Escutiella)

D. (D.) tenerifensis 7

D. (D.) ibera

D. (D.) valenciennii

D. (E.) alexantoni n. sp.

D. (E.) deshayesii

Animal Biodiversity and Conservation 36.1 (2013)

Mediterranean Sea

Essaouira

M or oc co

i nt

a ce

3 Agadir

Fig. 37. Map of geographical distribution of Drusia (E.) alexantoni n. sp. Fig. 37. Mapa de distribución geográfica de Drusia (E.) alexantoni sp. n.

consist of 100 and 116 rows and both measure 4.65 mm length and 3.0 mm wide. Its radular formula is: 51 + C + 51. Teeth are generally similar to D. (E.) deshayesii (figs. 30–36) (Martínez–Ortí & Borredà, 2012). The central tooth presents a deep cut in the shape of an isosceles triangle at the base of the mesocone and reaching the vertex and lower end of this triangle (figs. 31–33). The ectocones also present wing–shape expansions. Other teeth present at the base of the external ectocone with additional wing– shape expansions directed outwards (Martínez–Ortí & Borredà, 2012). Geographical distribution and habitat D. (E.) alexantoni n. sp. has been found on the Moroccan Atlantic coast, from Essaouira to Agadir (fig. 37), in crops of argan (Argania spinosa (L.) Skeels). One of the authors (Martínez–Ortí) collected all the specimens living in colonies underneath the stones and small walls between these crops along with the Papillionaceae plant Ononix natrix L. which is possibly part of their diet. It has also been cited in lacustrine riparian environments (Martínez, 2009). Discussion This new species undoubtedly belongs to the genus Drusia and we decided to include it in the subgenus D. (Escutiella) due to the appearance of its shell and other features. Besides, it is very similar to D. (E.) deshayesii

due to the following reproductive characters: i) penis with a lateral protrusion, ii) inside the penis there are two thick and solid bumps and iii) reticulated epiphallus inside with thick longitudinal folds. It differs from D. (E.) deshayesii by i) a reproductive system with uneven atrial appendices of the horn– shaped organ, ii) lateral protrusion that gives it an elbow–like shape that is not present in D. (E.) deshayesii, iii) the arrangement and number of bumps inside the penis, only two of them in D. (E.) alexantoni n. sp and up to four in D. (D.) desayeshii, iv) the reticulated appearance of the inner wall of the epiphallus is different, with the longitudinal folds being much larger in D. (E.) alexantoni n. sp. and v) the very large albumen gland of the new species. These reproductive characters are taxonomically more relevant than the external appearance which in juveniles, with spots and bands, could be confused with the subgenus D. (Drusia) s. str. and with the species Drusia (D.) valenciennii. Equally, the two appendices of the corniform organ are very unequal in the new species, which makes it more like D. (D.) valenciennii. However, due to the set of characters mentioned and described above, it seems much more related to D. (E.) deshayesii and we have included it in the subgenus D. (Escutiella). The radula maximum dimensions of D. (E.) alexantoni n. sp. are 4.65 x 3.00 mm, being slightly smaller than in D. deshayesii (6.75 x 3.95 mm) and D. (D.) valenciennii (7.00 x 4.00 mm). In addition, the radular formula of D. (E.) alexantoni n. sp. (51 + C + 51) is clearly different from D. deshayesii (70 + C + 70) and D. valenciennii (65 + C+ 65) (Martínez–Ortí & Borredà, 2012). Martínez–Ortí & Borredà (2012) provide a dichotomous key to identify the species in the family Parmacellidae but due to the discovery of D. (E.) alexantoni n. sp. and the new morpho–anatomical features provided it requires slight modifications (see above). Acknowledgements We thank Dr. Fouad Achemchem from l’ Université Ibn Zohr Agadir (Morocco) for his help during the sampling. We also thank the staff at the Sección de Microscopía Electrónica of the S. C. S. I. E., Universitat de València for their help using the SEM Hitachi S–4100. References Martínez, F., 2009. Parmacella sp. http://www.biodiversidadvirtual.org/insectarium/Parmacella+sp– img62285.html (accessed July 10, 2012) Martínez–Ortí, A. & Borredà, V., 2012. New systematics of Parmacellidae P. Fischer 1856 (Gastropoda, Pulmonata), with the recovery of the genus–name Drusia Gray 1855 and the description of Escutiella subgen. nov. Journal of Conchology, 41(1): 1–18.

Arizaga et al.

Animal Biodiversity and Conservation 36.1 (2013)

Connectivity patterns and key non–breeding areas of white–throated bluethroat (Luscinia svecica) European populations

J. Arizaga & I. Tamayo

Arizaga, J. & Tamayo, I., 2013. Connectivity patterns and key non–breeding areas of white–throated bluethroat (Luscinia svecica) European populations. Animal Biodiversity and Conservation, 36.1: 69–78. Abstract Connectivity patterns and key non–breeding areas of white–throated bluethroat (Luscinia svecica) European populations.— Using ring recovery data from the EURING databank, the aims of this study were: (1) to identify the chief migration and wintering areas of white–throated bluethroat European subspecies, L. s. namnetum, L. s. cyanecula and L. s. azuricollis, (2) to evaluate the degree of connectivity between breeding and non–breeding regions and determine the migration patterns of each subspecies, and (3) to evaluate whether recovery data are sufficient to answer the previous questions adequately. Most of the recoveries were obtained during the autumn migration period (n = 155, 68.9%), followed by winter (n = 49, 21.8%) and spring (n = 21, 9.3%). For L. s. azuricollis, we did not find any ring recoveries at more than 100 km in autumn or spring, and there were none at all in winter. All analyses thus relate to L. s. cyanecula and L. s. namnetum. Both subspecies move across a NE–SW axis from their breeding to their wintering areas within the circum–Mediterranean region, mainly in Iberia, following population–specific parallel migration routes. L. s. namnetum mainly uses the Atlantic coastal marshes from France to south–western Iberia, where the chief wintering areas are found. L. s. cyanecula, however, uses both Atlantic and Mediterranean wetlands in autumn, but only those in the Mediterranean in spring, thus giving rise to a loop–migration pattern. Telescopic migration was demonstrated for L. s. cyanecula. Recovery data were insufficient to identify in detail the entire wintering range for all white–throated bluethroat European populations. Technologies such as the use of geolocators will play a relevant role in this scenario. Key words: EURING databank, Mediterranean region, Migration and wintering, Recovery data, Wetlands. Resumen Patrones de conectividad y principales áreas de invernada de las poblaciones europeas del pechiazul (Luscinia svecica).— Utilizando los datos de recaptura recopilados en la base de datos de EURING, los objetivos del estudio fueron: (1) determinar las principales rutas migratorias y áreas de invernada de las subespecies de pechiazul L. s. namnetum, L. s. cyanecula y L. s. azuricollis; (2) evaluar el grado de conectividad entre las zonas de reproducción y las áreas de invernada, y determinar los patrones de migración de cada subespecie y (3) evaluar si los datos de recaptura son suficientes para responder adecuadamente a las preguntas anteriores. La mayor parte de las recapturas se obtuvieron durante el período de migración en otoño (n = 155; 68,9%), seguido del invierno (n = 49; 21,8%) y la primavera (n = 21; 9,3%). No se obtuvo ninguna recaptura de L. s. azuricollis en más de 100 km en otoño ni en primavera, ni tampoco en todo el invierno. Por consiguiente, todos los análisis hacen referencia a L. s. cyanecula y L. s. namnetum. Ambas subespecies de desplazan a lo largo de un eje NE–SO desde las zonas de reproducción hasta las zonas de invernada de la región circunmediterránea, principalmente en la península ibérica, y siguen rutas migratorias paralelas que son específicas de cada población. L. s. namnetum utiliza principalmente las marismas del Atlántico desde Francia hasta el suroeste de la península ibérica, donde se encuentran las principales zonas de invernada. Sin embargo, L. s. cyanecula utiliza los humedales tanto del Atlántico como del Mediterráneo en otoño, pero solo los del Mediterráneo en primavera; en consecuencia, se establece un patrón de migración en bucle. Se demostró que el patrón migratorio de L. s. cyanecula es de tipo telescópico. Los datos de recaptura fueron insuficientes para determinar con precisión la distribución invernal de todas las poblaciones europeas de pechiazul. Las tecnologías como la utilización de geolocalizadores desempeñarán una función fundamental en este contexto. ISSN: 1578–665X

Arizaga & Tamayo

Palabras clave: Base de datos de EURING, Región del Mediterráneo, Migración e invernada, Datos de recaptura, Humedales. Received: 28 XII 12; Conditional acceptance: 8 II 13; Final acceptance: 2 III 13 Juan Arizaga & Ibon Tamayo, Urdaibai Bird Center–Dept. of Ornithology, Aranzadi Sciences Society, Zorroagagaina 11, E20014 Donostia–S. Sebastián, España (Spain). Corresponding author: J. Arizaga. E–mail: jarizaga@aranzadi–zientziak.org

Animal Biodiversity and Conservation 36.1 (2013)

Introduction The term connectivity refers to the degree to which breeding and non–breeding areas used by a population of a migrant species are connected (Webster et al., 2002). High or strong connectivity occurs when all/most of the individuals from a population overwinter together in an area well–differentiated from that used by individuals from other populations (Swarth, 1920; Madsen et al., 1999; Chamberlain et al., 2000; Procházka et al., 2008). In contrast, connectivity is weak when individuals from several populations coexist in sympatry in the same wintering area (McGrady et al., 2002). Understanding connectivity patterns provides key clues about the spatio–temporal distribution range of birds during the non–breeding season and it has direct implications on population dynamics (Peach et al., 1991; Baillie & Peach, 1992; Szép, 1995; Newton, 2004) and conservation (Pain et al., 2004; Julliard et al., 2006). Studies on the connectivity patterns of European passerines have targeted mainly Afro–tropical winter quarters (e.g., Hill, 1997; Schäffer et al., 2006; Procházka et al., 2008; Zwarts et al., 2009), while less attention has been paid to the circum–Mediterranean region (although we have exceptions to this rule; e.g., see Newton, 1972; Bairlein, 2001) in spite of its importance as one of the chief wintering areas for several passerine species (Cramp, 1988, 1992; Cramp & Perrins, 1994b, 1994a). The bluethroat (Luscinia svecica) is a polytypic Holarctic passerine, breeding from Iberia in West Europe to Alaska and Canada. Two to three white–throated subspecies are recognized as breeding in Europe (Collar, 2005): L. s. namnetum breeds in the Atlantic wetlands of France and migrates along the coast of Northern Iberia (Arizaga et al., 2006a) to overwinter mainly in Portugal and Northern Africa (Zucca & Jiguet, 2002); L. s. cyanecula breeds from the east of France to Russia, and during the winter it is distributed over a wide geographic area ranging from Southern Europe to tropical Africa (Cramp, 1988); L. s. azuricollis breeds only in Iberia, and it is a subspecies well differentiated from the rest, as shown by genetic studies (Johnsen et al., 2006). The winter quarters of L. s. azuricollis remain to be discovered (Arizaga et al., 2006b, 2011b). Apart from this general distribution pattern, the degree of connectivity and the identification of key non–breeding areas for these subspecies have not been analysed, in part because the current data come from partial studies carried out in relatively small areas (Tellería et al., 1999; Zucca & Jiguet, 2002; Bønløkke et al., 2006; Spina & Volponi, 2009). Accordingly, there is still a need for a complete overview of how the breeding and non–breeding areas of white–throated European bluethroat populations are connected and where the chief stopover sites during autumn and spring migrations are placed considering a continental or even an inter–continental scale. The compilation of more ring recoveries in recent years could help to fill these gaps in knowledge. There are normally two main non–breeding longitudinal distribution patterns in migrants: parallel and non–parallel (centrifuge and fan) migrations. If bluethroat populations show parallel migration pat-

terns (indicating strong connectivity between breeding and non–breeding regions), longitudinal geographic location at breeding quarters should be positively correlated with longitudinal geographic location at non–breeding regions, either during the migration period or in winter (e.g., Zwarts et al., 2009). The other patterns would result in no relationship between geographic location at breeding and non–breeding quarters. Latitudinally, three chief patterns have been identified: leap–frog, chain, and telescopic migrations (Salomonsen, 1955; Newton, 2008). In the case of leap–frog migration, where northern breeders spend the winter in areas further south than those used by southern breeders, latitude of breeding origin should be negatively correlated with latitude of winter quarters. In this case, latitude of breeding origin should also be positively correlated with distance to winter location, as migrants breeding in the north must cover a much longer distance than those that breed further south. In contrast, in the case of chain migration where northern breeders also overwinter in areas further north than southern breeders, the latitude of breeding origin should correlate positively with latitude at winter quarters and not with distance to winter location, since populations occur in winter in the same latitudinal sequence as when occupying their breeding areas. Overall, both patterns will indicate strong connectivity, as each population occupies different wintering quarters. Alternatively, in the case of telescopic migrations, where populations occur in sympatry in the same wintering area, none of these correlations would be expected, but the latitude of breeding origin should be correlated with distance to winter quarters, as migrants coming from more distant regions will have to travel further to reach a sympatric wintering area shared with populations from closer places of origin. In this scenario, connectivity will be weak. Identifying main wintering areas is basic but insufficient when dealing with migrant species whose fitness could also depend on key sites for stopping– over (Fransson et al., 2005; Julliard et al., 2006). Such key sites can differ seasonally, such as in the case of loop migrations, when the migratory route to reach the winter quarters varies from that used to go to breeding quarters (Berthold, 2001). Within the Africa–European system of bird migrations, migrants in spring tend to displace themselves via routes which are further east than those used during autumn (Bairlein, 2001). This, however, has not been tested for all species and the causes underlying this strategy are still not fully understood. Using ring recovery data from the EURING databank, our aims here were to evaluate the degree of connectivity between breeding and non–breeding regions, and determine the migration patterns used by each bluethroat subspecies, including the detection of a possible seasonal change in the use of particular routes. For L. s. azuricollis, we did not find a single ring recovery at more than 100 km in autumn or spring, and there was norecovery at all in winter. All analyses henceforth will be relative to L. s. cyanecula and L. s. namnetum.

Arizaga & Tamayo

Material and methods Data collection The data used in this study were obtained from the EURING databank (ring recoveries of bluethroats ringed from 1926 to 2009). We only used data of bluethroats ringed in their breeding areas and recaptured during the autumn or spring migrations or during the winter, or vice versa. In the cases where the number of recoveries per bird and period were > 1 (e.g., a bird recaptured three times at the same stopover site during the autumn migration period), we only considered one of the recoveries. If such recoveries were obtained at different locations, then the most distant location was considered for the analysis. When a single bird was recovered in different periods (e.g., a bird is recaptured during the autumn migration period and also during the winter period), a recovery per period was considered (i.e. in the last case two recoveries were considered, one breeding–autumn, and the other breeding–winter). This resulted in a sample of 833 recoveries. The phenological periods were defined as follows (based on Cramp, 1988): breeding (May–Jul), autumn (Aug–Nov), winter (Dec– Feb), spring (Mar–Apr). From the 833 recoveries we considered the following for the analyses: (1) recovery data obtained at > 100 km from breeding site for the recoveries made in autumn and spring; (2) all the recoveries made during winter. We differentiated these periods because birds captured during the autumn or spring migration period could still be at (autumn) or have already arrived at (spring) their breeding sites, so these birds do not reflect the true geographic position during migration period. In contrast, bluethroats captured during the winter are truly wintering birds; in this case we cannot reject having resident bluethroat populations/individuals, which move a null or negligible distance. Some bluethroats captured in winter could be birds in active migration. Subspecies were identified using the areas where ring recoveries were obtained during the breeding period. Thus, bluethroats breeding in the Atlantic wetlands of France from Arcachon to Mont St. Michel were considered as L. s. namnetum (Zucca & Jiguet 2002), bluethroats breeding in Iberia as L. s. azuricollis, and the rest as L. s. cynaecula (Cramp, 1988). Statistical analyses Simple linear correlations were used to test whether the location (longitude and latitude) of bluethroats at their breeding areas was correlated with the location at their non–breeding areas during autumn, winter and spring. Similarly, correlations were also used to test a positive relationship of breeding origin location with distance to winter quarters. We used 95% kernel polygons to identify the main non–breeding quarters of L. s. cyanecula, using ArcGIS 9.2 ESRI. This analysis was carried out only on L. s. cyanecula since sample sizes for the other subspecies were too small to allow us to build these polygons.

Table 1. Number of recoveries of bluethroats of L. s. namnetum and L. s. cyanecula subspecies between breeding and non–breeding periods. (Source: EURING databank.) Tabla 1. Número de recapturas de las subespecies de pechiazul L. s. namnetum y L. s. cyanecula entre los períodos reproductivos y los no reproductivos. (Fuente: base de datos EURING.) Period

All recoveries Recoveries > 100 km

Autumn

470

155 (33.0%)

Winter

48 (98.0%)

Spring

184

21 (11.4%)

We used circular statistics to analyze the spatial distribution of recovery data. All angles were calculated from breeding to non–breeding (autumn, winter or spring) location. We used a Watson–Williams F–test to test between–period variations in mean migratory axis of recoveries plus a Mardia–Watson–Wheeler test to assess differences in spatial distribution pattern between periods (Arizaga et al., 2010b). Circular statistics were carried out using Oriana 3.0 software. SPSS software was used for the remaining statistical analyses. All means are given ± SE. Results Overall, we obtained a data set with most recoveries made during the autumn migration period (n = 155, 68.9%), followed by winter (n = 49, 21.8%) and spring (n = 21, 9.3%) (table 1). Most of the recoveries during the autumn and spring migrations were obtained at less than 100 km from breeding sites (table 1), thus indicating that many recoveries from this period were birds still remaining (autumn) or already present (spring) in their breeding areas. In contrast, during the winter, only one bird was recaptured at less than 100 km from its breeding location (table 1), indicating that most bluethroats left their breeding areas to spend this period in more distant regions (mean: 1,517.2 ± 86.3 km). The distribution (longitude and latitude) of bluethroats during the non–breeding period was positively correlated with their distribution range during the breeding period (table 2). Longitudinally, we observed that bluethroats from regions further east also appeared further east outside the breeding and vice versa, either in autumn, winter (with r coefficients > 0.6) or, especially, in spring (r > 0.8; table 2), thus supporting parallel migration patterns, more marked during the spring migration period. Furthermore, bluethroats from regions further east appeared in areas further north both in autumn and spring, although this might

Animal Biodiversity and Conservation 36.1 (2013)

Table 2. Linear correlations used to test the relationship between breeding and non–breeding geographic distribution of L. s. namnetum and L. s. cyanecula bluethroats. Recoveries obtained at less than 100 km during the autumn and spring migration periods were excluded to avoid considering birds still in their breeding areas. Tabla 2. Correlaciones lineales utilizadas para comprobar la relación entre la distribución geográfica de los períodos reproductivo y no reproductivo de las subespecies de pechiazul L. s. namnetum y L. s. cyanecula. Para no incluir las aves que aún permanecían en sus zonas reproductivas, se excluyeron las recapturas obtenidas a menos de 100 km durante los períodos de migración de otoño y primavera.

Subspecies

Period

Breeding latitude n

Breeding longitude r

All subspecies Non–breeding latitude +0.289

< 0.001

+0.250

0.002

Autumn

155

Winter

–0.258

0.074

+0.234

0.105

Spring

+0.337

0.135

+0.702

< 0.001

Non–breeding longitude +0.187

0.020

+0.686

< 0.001

–0.040

0.784

+0.659

< 0.001

+0.184

0.424

+0.877

< 0.001

Autumn

155

Winter Spring

L. s. namnetum Non–breeding latitude Autumn

+0.380

0.035

–0.249

0.177

Winter

–0.416

0.413

+0.405

0.426

Spring

–

Non–breeding longitude Autumn

+0.155

+0.406

–0.008

0.966

Winter

+0.432

+0.393

–0.419

0.409

Spring

–

L. s. cyanecula Non–breeding latitude Autumn

124

+0.315

< 0.001

+0.299

0.001

Winter

–0.269

0.081

+0.359

0.018

Spring

+0.239

0.325

+0.683

0.001

Non–breeding longitude Autumn

124

–0.045

0.617

+0.663

< 0.001

Winter

–0.340

0.026

+0.601

< 0.001

Spring

–0.155

0.526

+0.846

< 0.001

just be due to the fact that northern breeders also tended to breed further east (r = 0.261, p < 0.001, n = 224). Considering the two subspecies separately, such patterns were confirmed only for L. s. cyanecula (table 2). Latitudinally, bluethroats breeding further north were also located in areas further north, and east, in autumn, but not in winter or spring (table 2). This pattern was maintained when the two subspecies were considered separately.

The distance from breeding to winter quarters was positively correlated to breeding latitude (latitude: r = 0.612, p < 0.001; longitude: r = –0.049, p = 0.738, n = 49), indicating that bluethroats breeding in northern regions migrated a longer distance than those from areas further south. The breeding latitude was not correlated with the winter latitude (table 2), so bluethroats breeding in areas further north did not overwinter further to the south than those breeding

Arizaga & Tamayo

Autumn

L. s. cyanecula Winter

Spring

Autumn

L. s. namnetum Winter

Spring

Fig. 1. Recoveries of L. s. cyanecula and L. s. namnetum captured / recaptured during the breeding period and recaptured/captured during the non–breeding period in autumn, winter, and spring. Sample sizes are shown in table 2. In L. s. cyanecula, we have not shown the single winter relating to a bird ringed in The Netherlands and recaptured in Senegal. Fig. 1. Recapturas de L. s. cyanecula y L. s. namnetum capturadas / recapturadas durante el período reproductor y recapturadas/capturadas durante el período no reproductor en otoño, invierno y primavera. Los tamaños muestrales se indican en la tabla 2. En L. s. cyanecula no hemos mostrado la única recaptura de invierno de un ave que se anilló en los Países Bajos y se recapturó en Senegal.

in areas further south. Only a single recovery was found south of the Sahara and this was a bird ringed in The Netherlands. Most winter recoveries (n = 39, equivalent to 79.6%) appeared in Iberia, both in the case of L. s. namnetum (100%) and L. s. cyanecula (76.7%) (fig. 1). L. s. namnetum was concentrated in areas within south–western Iberia whereas L. s. cyanecula tended to overwinter in south–eastern and eastern Iberia. There were also additional recoveries of this subspecies in northern Africa (Morocco). In both subspecies bluethroats mostly overwintered in coastal regions, especially on the Mediterranean and southern Atlantic coasts of Iberia (fig. 1). The small sample sizes obtained for L. s. namnetum preclude us from making a firm conclusion with regard

to the possible use of different routes in autumn and spring (table 2; fig. 1), but this was not the case for L. s. cyanecula (table 2; fig. 1). In this subspecies, during autumn, we observed a high number of recoveries on both the Atlantic and the Mediterranean coasts, especially in France and in the east of Iberia (fig. 2). Sites with particularly high concentrations of recoveries were the Arcachon bay area and the Ebro basin and delta, together with the Mediterranean wetlands from La Camargue in France, and Doñana and its surroundings in Iberia (fig. 2). In spring, however, not a single recovery was obtained on the Atlantic side apart from the Doñana and surrounding areas. The remainder were all related to the Mediterranean. As in autumn, bluethroats in spring mainly used the wetlands existing from Doñana to La Camargue (fig. 2).

Animal Biodiversity and Conservation 36.1 (2013)

Autumn

270

Winter Winter

180

Spring

Fig. 2. 95% kernel polygons obtained from recovery data of L. s. cyanecula between breeding and non–breeding regions in autumn, winter and spring. Fig. 2. Funciones de kernel al 95% obtenidas a partir de los datos de recaptura de L. s. cyanecula entre las regiones reproductivas y no reproductivas en otoño, invierno y primavera.

Overall, bluethroats moved along a NE–SW axis between their breeding and non–breeding areas, without significant differences between periods (Watson Williams test: F2,221 = 0.698, p = 0.499; mean vector: 206.9 ± 2.7º, n = 224; data considered here: movements > 100 km). However, dispersion of data differed between periods (Mardia test: W = 11.936, p = 0.018). In particular, recovery data from winter were less dispersed than during autumn and almost

L. s. namnetum L. s. cyanecula

Fig. 3. Row data and mean migratory direction of bluethroats captured during the non–breeding period at a distance of more than 100 km from breeding localities. Angles established from breeding to non–breeding localities. Fig. 3. Datos no elaborados y dirección media de la migración de los pechiazules capturados durante el período no reproductor a una distancia de más de 100 km de las localidades de reproducción. Ángulos establecidos desde las localidades reproductivas y a las no reproductivas.

significantly less than during spring (pairwise comparisons: autumn–winter, W = 10.803, p = 0.005; autumn–spring, W = 0.416, p = 0.812; winter–spring: W = 5.087, p = 0.079; fig. 3). By subspecies, the migratory direction was observed to differ subspecifically in autumn (Watson Williams test: F1,153 = 19.224, p < 0.001; L. s. namnetum, n = 31; L. s. cyanecula, n = 124), but not in winter (F1,46 = 1.655, p = 0.205; L. s. namnetum, n = 6; L. s. cyanecula, n = 42) or spring (F1,19 = 0.002, p = 0.965; L. s. namnetum, n = 2; L. s. cyanecula, n = 19; fig. 3). The small sample sizes for L. s. namnetum in winter and spring should be noted, however. Discussion We used ring recovery data from the EURING databank to identify the main migration and wintering areas of white–throated bluethroat subspecies breeding in Europe, and also to understand the connectivity patterns between breeding and non–breeding grounds of each subspecies. Of the three white–throated bluethroat subspecies with recovery data available in the EURING databank, we only obtained valuable data for two of them: L. s. namnetum and L. s. cyanecula. Most L. s. namnetum recoveries were spread across the Atlantic coasts of France and Iberia, thus confirming that this subspecies uses a very specific, narrow corridor to move from its breeding area on the Atlantic side of France to its wintering area on the south–western coast of Iberia (Zucca & Jiguet, 2002). Outside this Atlantic scenario, L. s. namnetum should be regarded as a rarity (Arizaga et al., 2006a). If we consider L. s. cyanecula, it seems to move across broad fronts, although some degree of connectivity between breeding and non–breeding grounds exists. The clearest, most marked result was that L. s. cyanecula bluethroats breeding further east were also located in regions further east during the autumn migration period, winter and spring. This result highlights the occurrence of parallel routes of migration among L. s. cyanecula populations, so that migrants breeding in adjacent places follow the same (SW in this case) main direction but do not exhibit high overlap, as has already been shown for other species in Europe (Imboden, 1974; Bairlein, 2001). Particularly marked was the connectivity between breeding and non–breeding grounds in spring, with a correlation coefficient > 0.8, as compared to autumn and winter, with r values > 0.6. This result suggests that migrants use more direct, probably more population–specific routes in spring than in autumn, and that the overlap that they exhibit at their wintering areas is less marked in spring. In autumn, both subspecies were mainly concentrated in coastal marshes, both on the Atlantic (Arcachon, Doñana) and on the Mediterranean coasts (from the east of Iberia to La Camargue, and also around Venice). This suggests that these wetlands could play a key role for the conservation of both L. s. namnetum and L. s. cyanecula subspecies. A possible drawback in relation to this result is a possible geographic bias in sampling (ringing) effort. If this effort is higher at

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the large,main marshes with apparent good habitats, then this result must be considered cautiously, and we cannot fully reject the possibility that some inland humid zones could also be r elevant for bluethroats. The number of recoveries during the spring migration period was very low, so we should be cautious about making a firm conclusion with regard to this period. This is especially applicable to L. s. namnetum, for which there were only two recoveries in this period.. In the case of L. s. cyanecula, where autumn recoveries were obtained both on the Atlantic and the Mediterranean coasts, in spring the recoveries all had a Mediterranean–biased distribution, suggesting loop–migration. A reason commonly suggested is that using more direct routes allows migrants to reach their breeding areas faster, which is crucial in terms of fitness as it has breeding consequences (Kokko, 1999). Alternative hypotheses are the wind–assistance and the food provisioning hypotheses. The wind–assistance hypothesis states that migrants will gain advantage by migrating over areas that provide greater tail wind assistance or have weaker head wind. If the dominant wind varies seasonally and regionally, then migrants will use different routes depending on the season, and thus will show loop–migration. The food provisioning hypothesis states that migrants would use routes that provide a more abundant food supply. When this supply varies seasonally and regionally, loop–migration is expected. Most recoveries in winter (ca. 80%) appeared in Iberia, thus revealing the relevance of this area for the conservation of bluethroats (L. s. namnetum and L. s. cyanecula) in Europe. Other Mediterranean peninsulas, such as Italy (not Greece, where we found no recoveries), did not seem to be important wintering regions for the studied bluethroat populations. This conclusion agrees with previous research (Spina & Volponi, 2009). Especially relevant was the southern half of Iberia, with both the Atlantic and the Mediterranean coasts being utilized. Apart from the Mediterranean marshes in eastern Iberia (Peiró, 1997) and some Andalusian wetlands in the south–east which had been reported to host bluethroats in winter (Cortés et al., 2002), we discovered high concentrations within the Atlantic wetlands of Doñana and Bahía de Cádiz. It is of note that this was the area with the highest densities of wintering L. s. cyanecula in Europe, as revealed by ring recovery data. Again, as we have said previously, the identification of these key sites must be regarded with some caution because we have no data on a possible bias in sampling effort. The Ebro Valley, in northern Iberia, was pointed out in the past as being a chief corridor for the bluethroat during migration period (Hernández et al., 2003), but was not considered to be a main wintering area (Arizaga et al., 2010a). In our study the Ebro Valley (or more specifically part of it) appeared as a relatively important wintering area. Some Moroccan wetlands also appeared to host a relatively high number of bluethroats during the winter. Interestingly, although biometric analyses have shown that L. s. namnetum overwinter in Morocco (Zucca & Jiguet, 2002), no recoveries of this subspecies have

Animal Biodiversity and Conservation 36.1 (2013)

been obtained for the Maghreb. In part this is likely due to the low ringing effort carried out in Morocco as opposed to Iberia. Future campaigns in Morocco will be capital to clarify whether the main wintering area of the L. s. namnetum subspecies is Portugal or the Atlantic wetlands of Morocco. The distribution of recoveries at winter quarters did not correlate with the distribution at breeding quarters, thus not supporting the leap–frog and chain migration strategies. Conversely, telescopic migration was evidenced, with a high overlap between populations from different origin latitudes within their wintering areas. The distance from breeding to winter quarters correlated positively with breeding latitude, thus adding further support to this theory. This result, however, may be biased due to the lack of recoveries in tropical Africa. The single recovery obtained in this region was from a bird ringed in The Netherlands, one of the most northern breeding places for the L. s. cyanecula subspecies (Cramp, 1988). Future studies using other methods, such as stable isotopes analyses (Pain et al., 2004) and/or the use of geo–locators (Bächler et al., 2010) will clarify whether L. s. cyanecula populations demonstrate a telescopic migration strategy. Recovery data were insufficient to identify in detail the entire wintering range for all white–throated bluethroat European populations. Paradigmatic examples of this were the lack of recovery data from bluethroats breeding in Iberia (L. s. azuricollis), and also the lack of recoveries in the Sahelian non–breeding areas, where only one recovery (from an individual ringed in The Netherlands) was obtained. Morphological studies lend support to the hypothesis that bluethroats breeding in Iberia are long–distance migrants that probably overwinter in tropical Africa (Arizaga et al., 2006b). Additionally, the mean wing length of bluethroats captured in Senegal during the winter (Arizaga et al., 2011a) approached that reported for the Iberian population (Arizaga et al., 2011b). The difficulty to find/obtain recoveries in Africa may explain why the wintering area of this subspecies remains still unknown. Moreover, as the number of bluethroats ringed as breeders in Iberia is small (ca. less than 1,000 individuals from 2000 to 2010; J. García, per. com.), the chances of obtaining recoveries outside their breeding areas are low given the low recapture rates of ringed passerines. In conclusion, European populations of white– throated bluethroat (belonging to L. s. namnetum and L. s. cyanecula subspecies) move on a NE–SW axis from their breeding areas to their wintering areas within the circum–Mediterranean region, mainly in Iberia, following population–specific parallel migration routes . L. s. namnetum mainly uses the Atlantic coastal marshes from France to south–western Iberia, where the chief wintering areas are found, whilst L. s. cyanecula uses both the Atlantic and the Mediterranean wetlands in autumn but only the Mediterranean wetlands in spring, thus giving rise to a loop–migration pattern. Nothing is yet known, however, about the Afro–tropical winter grounds on a population scale, or about the L. s. azuricollis subspecies breeding in Iberia.

Acknowledgements This work was funded by the Basque Government and the Gipuzkoa Administration. The recovery data were provided by the EURING databank (www.euring.org). Thanks to I. Maggini and V. Salewski, who provided useful comments that helped us to improve an earlier version of this work. References Arizaga, J., Alonso, D., Campos, F., Unamuno, J. M., Monteagudo, A., Fernandez, G., Carregal, X. M. & Barba, E., 2006a. ¿Muestra el pechiazul Luscinia svecica en España una segregación geográfica en el paso posnupcial a nivel de subespecie? Ardeola, 53: 285–291. Arizaga, J., Alonso, D. & Fernández, E., 2010a. Presencia de pechiazules Luscinia svecica invernantes en Navarra. Rev. Cat. Ornitol., 26: 51–55. Arizaga, J., Alonso, D., Maggini, I., Romero, L., Vilches, A. & Belamendia, G., 2011a. Características de los paseriformes europeos que invernan en el Parque Nacional de las Aves del Djoudj (África occidental). Munibe, 59: 41–51. Arizaga, J., Campos, F. & Alonso, D., 2006b. Variations in wing morphology among subspecies might reflect different migration distances in Bluethroat. Ornis Fennica, 83: 162–169. Arizaga, J., García, J. & Suárez–Seoane, S., 2011b. Ruiseñor Pechiazul – Luscinia svecica. In: Enciclopedia virtual de los Vertebrados Españoles (A. Salvador, L. M. Bautista, Eds.). MNCN, Madrid: http://www.vertebradosibericos.org Arizaga, J., Herrero, A., Galarza, A., Hidalgo, J., Aldalur, A., Cuadrado, J. F. & Ocio, G., 2010b. First–year movements of Yellow–legged Gull (Larus michahellis lusitanius) from the southeastern Bay of Biscay. Waterbirds, 33: 444–450. Bächler, E., Hahn, S., Schaub, M., Arlettaz, R., Jenni, L., Fox, J. W., Afanasyev, V. & Liechti, F., 2010. Year–Round Tracking of Small Trans–Saharan Migrants Using Light–Level Geolocators. Plos One, 5. Baillie, S. R. & Peach, W. J., 1992. Population limitation in Palearctic–Africa migrant passerines. Ibis, 134: 120–132. Bairlein, F., 2001. Results of bird ringing in the study of migration routes. Ardea, 89: 7–19. Berthold, P., 2001. Bird migration – a general survey. Oxford Univ. Press, Oxford. Bønløkke, J., Madsen, J. J., Thorup, K., Pedersen, K. T., Bjerrum, M. & Rahbek, C., 2006. Dansk Traekfugleatlas. Rhodos, Humlebæk. Collar, N. J., 2005. Family Turdidae (Thrushes). In: Handbook of the Birds of the World, vol 10: 514–807 (J. del Hoyo, A. Elliot & D. A. Christie, Eds.). Lynx, Barcelona. Cortés, J. A., Cobos, V. & Vidoy, I., 2002. El plumaje de los pehiazules (Luscinia svecica) invernantes en la provincia de Málaga. Revista de Anillamiento, 9–10: 41–48. Chamberlain, C. P., Bensch, S., Feng, X., Akesson,

S. & Andersson, T., 2000. Stable isotopes examined across a migratory divide in Scandinavian willow warblers (Phylloscopus trochilus trochilus and Phylloscopus trochilus acredula) reflect their African winter quarters. Proceedings of the Royal Society of London Series B–Biological Sciences, 267: 43–48. Cramp, S., 1988. Handbook of the Birds of Europe, the Middle East and North Africa, vol. 5. Oxford Univ. Press, Oxford. – 1992. Handbook of the Birds of Europe, the Middle East and North Africa, vol. 6. Oxford Univ. Press, Oxford. Cramp, S. & Perrins, C. M., 1994a. Handbook of the Birds of Europe, the Middle East and North Africa, vol. 8. Oxford Univ. Press, Oxford. – 1994b. Handbook of the Birds of Europe, the Middle East and North Africa, vol. 9. Oxford Univ. Press, Oxford. Fransson, T., Jakobsson, S. & Kullberg, C., 2005. Non–random distribution of ring recoveries from trans–Saharan migrants indicates species–specific stopover areas. Journal of Avian Biology, 36: 6–11. Hernández, M., Campos, F., Arizaga, J. & Alonso, D., 2003. Migration of the bluethroat Luscinia svecica in the Iberian Peninsula. Ardeola, 50: 259–263. Hill, L. A., 1997. Trans–Saharan recoveries of House Martins Delichon urbica, with discussion on ringing, roosting and sightings in Africa. Safring News, 26: 7–12. Imboden, C., 1974. Zug, Fremdansiedlung und Brut– periode des Kiebitz Vanellus vanellus in Europa. Ornithol. Beob., 71: 5–134. Johnsen, A., Andersson, S., Garcia Fernandez, J., Kempenaers, B., Pavel, V., Questiau, S., Raess, M., Rindal, E. & Lifjeld, J. T., 2006. Molecular and phenotypic divergence in the bluethroat Luscinia svecica subspecies complex. Molecular Ecology, 15: 4033–4047. Julliard, R., Bargain, B., Dubos, A. & Jiguet, F., 2006. Identifying autumn migration routes for the globally threatened Aquatic Warbler Acrocephalus paludicola. Ibis, 148: 735–743. Kokko, H., 1999. Competition for early arrival in migratory birds. J. Anim. Ecol., 68: 940–950. Madsen, J. J., Cracknell, G. & Fox, T., 1999. Goose populations of the western Palaearctic. National Environment Research Institute, Rönde, Denmark. McGrady, M. J., Maechtle, T. L., Vargas, J. J., Seegar, W. S. & Pena, M. C. P., 2002. Migration and ranging

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of Peregrine Falcons wintering on the Gulf Of mexico coast, Tamaulipas, Mexico. Condor, 104: 39–48. Newton, I., 1972. Finches. Collins, London. – 2004. Population limitation in migrants. Ibis, 146: 197–226. – 2008. The migration ecology of birds. Academic Press, London. Pain, D. J., Green, R. E., Giessing, B., Kozulin, A., Poluda, A., Ottosson, U., Flade, M. & Hilton, G. M., 2004. Using stable isotopes to investigate migratory connectivity of the globally threatened aquatic warbler Acrocephalus paludicola. Oecologia, 138: 168–174. Peach, W., Baillie, S. & Underhill, L., 1991. Survival of British Sedge Warblers Acrocephalus schoenobaenus in relation to West African rainfall. Ibis, 133: 300–305. Peiró, I. G., 1997. A study of migrant and wintering bluethroats Luscinia svecica in south–eastern Spain. Ringing and Migration, 18: 18–24. Procházka, P., Hobson, K., Karcza, Z. & Kralj, J., 2008. Birds of a feather winter together: migratory connectivity in the Reed Warbler Acrocephalus scirpaceus. Journal of Ornithology, 149: 141–150. Salomonsen, F., 1955. The evolutionary significance of bird–migration. Dan. Biol. Medd., 22: 1062. Schäffer, N., Walther, B. A., Gutteridge, K. & Rahbek, C., 2006. The African migration and wintering grounds of the Aquatic Warbler Acrocephalus paludicola. Bird Conservation International, 16: 33–56. Spina, F. & Volponi, S., 2009. Atlante della migrazione degli uccelli in Italia. Vol. 2: Passeriformi. ISPRA–MATTM, Roma. Swarth, H. S., 1920. Revision of the avian genus Passerella, with special reference to the distribution and migration of the races in California. Univ California Publication in Zoology, 21: 75–224. Szép, T., 1995. Relationship between west African rainfall and the survival of central European San Martins Riparia riparia. Ibis, 137: 162–168. Tellería, J. L., Asensio, B. & Díaz, M., 1999. Aves Ibéricas. II. Paseriformes (J. M. Reyero, Ed.). Madrid. Webster, M. S., Marra, P. P., Haig, S. M., Bensch, S. & Holmes, R. T., 2002. Links between worlds: unraveling migratory connectivity. Trends in Ecology and Evolution, 17: 76–83. Zucca, M. & Jiguet, F., 2002. La Gorgebleue à miroir (Luscina svecica) en France: nidification, migration et hivernage. Ornithos, 9–6: 242–252.

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Intensive monitoring suggests population oscillations and migration in wild boar Sus scrofa in the Pyrenees M. Sarasa & J.–A. Sarasa

Sarasa, M. & Sarasa, J.–A., 2013. Intensive monitoring suggests population oscillations and migration in wild boar Sus scrofa in the Pyrenees. Animal Biodiversity and Conservation, 36.1: 79–88. Abstract Intensive monitoring suggests population oscillations and migration in wild boar Sus scrofa in the Pyrenees.— As few studies have analysed local variability in populations of wild boar Sus scrofa in Western Europe in recent years, our understanding of ecological processes currently affecting this species is limited. To analyse questions regarding local variability in wild boar abundance, we used information from 442 traditional drive hunts monitored throughout eight hunting periods in the Pyrenees mountain range (Urdués, N Spain). Results showed temporal oscillations in abundance, and a non–linear decrease of 23% in the number of wild boar seen per drive hunt between 2004 and 2011. Numbers of dogs and hunters per drive hunt also affected indexes of wild boar abundance. Inter–annual variations in bag size may cause overestimations of variations in boar abundance and may even deviate from the population dynamics inferred from the number of wild boars seen per drive hunt. The multimodal patterns of wild boar abundance during the hunting periods suggest migrations in the Pyrenees. Our findings highlight the limitations of hunting bag statistics in wild boar. Further studies are required to guarantee information–based sustainable management of wild boar populations. Key words: Wild boar, Sus scrofa, Animal migration, Big game traditional hunting, Population dynamics, Wildlife management. Resumen El seguimiento intensivo sugiere la existencia de oscilaciones demográficas y movimientos migratorios en las poblaciones de jabalí (Sus scrofa) en los Pirineos.— Muy pocos estudios recientes han analizado la variabilidad local de las poblaciones de jabalí (Sus scrofa) en Europa occidental, lo que limita nuestra comprensión de los procesos ecológicos que en la actualidad afectan a esta especie. Usando la información recopilada mediante el seguimiento de 442 batidas durante ocho temporadas de caza en los Pirineos (Urdués, norte de España), se analizaron cuestiones relacionadas con la variabilidad local de la abundancia de jabalí. Los resultados revelaron oscilaciones temporales de la abundancia y una disminución discontinua del 23% en el número de jabalíes avistados por batida entre 2004 y 2011. El número de perros y de cazadores por batida también afectó a los índices de abundancia de jabalí. Las variaciones interanuales de animales abatidos pueden provocar que se sobreestimen las variaciones de la abundancia de jabalí e incluso pueden desviarse de la dinámica de poblaciones inferida del número de jabalíes avistados por batida. En los Pirineos, el patrón multimodal de la abundancia de jabalí durante las temporadas de caza sugiere la existencia de movimientos migratorios. Los resultados obtenidos destacan las limitaciones de las estadísticas de abundancia realizadas sobre el número de jabalíes abatidos y ponen de manifiesto la necesidad de llevar a cabo nuevos estudios que permitan gestionar las poblaciones de jabalí de forma sostenible y fundamentada. Palabras clave: Jabalí, Sus scrofa, Migración, Caza mayor tradicional, Dinámica de poblaciones, Gestión de la fauna silvestre. Received: 13 I 13; Conditional acceptance: 3 III 13; Final acceptance: 22 III 13 Mathieu Sarasa, Grupo Biología de las Especies Cinegéticas y Plagas (RNM–118), España (Spain).– Juan– Antonio Sarasa, Grupo de Caza Mayor de Urdués, España (Spain). Corresponding author: M. Sarasa, Fédération Nationale des Chasseurs 13, Rue du Général Leclerc, F–92136 Issy les Moulineaux Cedex, France. E–mail: mathieusar@hotmail.com; msarasa@chasseurdefrance.com ISSN: 1578–665X

Introduction Over the last 30 years, most studies discussing or mentioning wild boar Sus scrofa abundance and densities in Western Europe have suggested that overall wild boar populations are increasing (Sáez–Royuela & Tellería, 1986; Melis et al., 2006; Marco et al., 2011). Wild boar populations have been associated with health problems in livestock and humans (Artois et al., 2002; Rossi et al., 2005; Meng et al., 2009), damage to crops and colonized ecosystems (Schley et al., 2008; Cuevas et al., 2010), and road accidents (Groot Bruinderink & Hazebroek, 1996; Lagos et al., 2012). It has also been suggested that wild boar might reduce the availability of summer grazing areas through soil disturbance (Bueno et al., 2010), although such issues have raised considerable controversy (Risch et al., 2010; Wirthner et al., 2011; Wirthner et al., 2012). In light of these concerns and of the predicted increase in wild boar populations as a response to global warming (Melis et al., 2006), management tools to control and reduce wild boar populations are of much interest (Massei et al., 2011). In recent years, however, few studies of wild boar population dynamics in Western Europe have been performed (Marco et al., 2011), limiting our understanding of current ecological processes in this species. Generalized increases in wild boar densities are thought to be responsible for increasing presence of wild boar in agricultural ecosystems and even in urban environments (Jansen et al., 2007). However, very few studies have addressed whether these increases are the result of a source–sink gradient, sustained by woodland environments with increasing numbers of wild boars, or whether wild boars are locally adapting to agricultural and urban environments in which effective and perceived hunting pressure is low and opportunist foraging is facilitated by city dwellers (Cahill & Llimona, 2004). Increasing interest is also being shown in the way in which wild boar use space, and a number of studies have revealed variable and complex movement patterns (Keuling et al., 2010). Whilst some authors suggest that wild boars are essentially sedentary animals (Saunders & Kay, 1996; Keuling et al., 2008; Mitchell et al., 2009), others indicate that wild boars might perform sex–specific habitat selection depending on their landscape of fear (Saïd et al., 2012) or even on local migrations (Andrzejewski & Jezierski, 1978; Singer et al., 1981; D’Andrea et al., 1995). The essential parameters regarding population dynamics and space use in wild boar are therefore unclear, thus hindering the establishment of appropriate management practices. In this study, we used hunting data collected over eight hunting periods in a locality of the Pyrenean mountain range (Urdués, N Spain) to analyse the population dynamics and space use in the wild boar. Historically, the Pyrenees have always been considered a propitious environment for this game species (Gortázar et al., 2000). Thus, if the increase in wild boar in agricultural and urban environments is the product of a source–sink gradient sustained by woodland environments, we would expect a positive population trend at our study

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site during the monitoring period (prediction 1). Given the overall increase in wild boar populations referred to in other studies (Sáez–Royuela & Tellería, 1986; Marco et al., 2011), we would also expect an increase in wild boar populations in the Pyrenees (prediction 1) Secondly, we investigated the spatial ecology of wild boar by testing two mutually exclusive hypotheses. If the wild boar is a sedentary species (hypothesis 1), we would expect a decrease in wild boar abundance in our study site during the hunting period due to the population reduction caused by hunting pressure (prediction 2). Nevertheless, if the wild boar is migratory in the Pyrenees (hypothesis 2), we would expect to observe a pattern of abundance that, rather than corresponding to a simple model of population decrease during the hunting period, exhibits a bell–shaped or a multi–modal pattern of abundance indexes during the hunting period (prediction 3). Thirdly, we also took into account previous studies of wild boar harvesting. Null or weak relationships were recorded between the numbers of dogs and hunters and bag size per drive hunt in Italy (Scillitani et al., 2010). Thus, we expected comparable results at our study site (prediction 4). Material and methods Study site We analysed local wild boar abundance on the southern side of the Pyrenees (Hecho valley, Aragón, northern Spain). This area is characterized by extensive woodlands (mainly Pinus sylvestris, Fagus sylvatica and Quercus sspp.) and few open habitats (Acevedo et al., 2006). Human population density is low and traditional agricultural practices are mostly focused on animal husbandry (cows, sheep, goats, and horses). In the Pyrenees, local agriculture has been changing in recent decades and the natural reforestation of former open areas has led to a loss of diversity in the landscape mosaic (Ortigosa et al., 1990; García–Ruiz et al., 1996; Roura–Pascual et al., 2005). In this area, traditional hunting drives for wild boar are conducted by one or more beaters on foot with dogs and with hunters on stands. Despite apparent intensive harvesting, the global hunting pressure on the species in the region might be low, due to the abundance of shelter areas (Acevedo et al., 2006; Herrero et al., 2008). In this study we monitored the hunting group from the village of Urdués, which harvests wild boar in their local area (25 km2). In this study, the moderate scale of the area and the detailed information of our data set allowed a more precise monitoring of the local abundance of wild boar than in previous studies on wild boar in Aragón (detailed thereafter). Data collection Close collaboration with hunters allowed us to generate a database that included details of the drive hunts that were not included in previous studies on wild boar. Although reliable estimates of ungulate abundance

Animal Biodiversity and Conservation 36.1 (2013)

can be made when the numbers of individuals seen on hunts are available (Ericsson & Wallin, 1999; Mysterud et al., 2007; Rönnegård et al., 2008), this information has rarely been available in previous studies on wild boars (Sáez–Royuela & Tellería, 1988; Tellería, 2004). Hunting statistics could also act as a good index of wild boar population abundance (Tellería, 2004; Imperio et al., 2010), even though official hunting statistics in Spain are incomplete and of questionable accuracy (Martínez–Jaúregui et al., 2011). We monitored the number of wild boars seen and culled during each drive hunt. The area in which the hunt took place and the number of dogs and hunters were also recorded. In total, data from 442 drive hunts were recorded during hunting periods (2.21 drive hunts per km2 per hunting period) from 2004 to 2012. For this study our monitoring allowed a fine–resolution that is close to forty times greater than those of previous studies on wild boar in this region (2,657 drive hunts per hunting period for 47,669 km2 in Aragón means close to 0.06 drive hunt per season per km2 [Acevedo et al., 2006]). This underlines the difference of resolution between the data used in previous studies on wild boar in Aragón and the data set used in this study. Analysis We considered two indexes of wild boar abundance: the number of wild boars seen (index 1) and the number of wild boars culled (index 2) per drive hunt. To analyse the determining factors in these indexes of abundance, we used General Additive Models (GAMs) (Wood, 2006; Zuur et al., 2007). The explanatory variables considered were: (Y) the year in which the hunting period started; (D) the day of the hunting period: for the first day of each hunting period, we used the day number according to the Gregorian calendar and then added the number of days up to the end of the hunting period; (Nd) the number of dogs; (Nh) the number of hunters. In the Pyrenees, the number of hunters per drive hunt in traditional hunting groups (mean ± SE = 7 ± 2.8) rarely or never allows coverage of all the potential escape routes of the hunted patch. Also, data concerning the exact surface hunted by beaters and dogs (mean number of dogs per drive hunt ± SE = 9.8 ± 3.6) is usually unavailable because the courses of the dogs are not systematically recorded with telemetric tools. The exact hunted area is thus usually unknown. In this study, the approximate area potentially hunted during each drive hunt was close to 2.5 km2. Furthermore, instead of using estimated surfaces characterized by overblown and unreliable accuracy, we tested the area in which the drive hunt took place as a potential co–factor to account for the potential effects of spatial heterogeneity on wild boar abundance. We used an information–theoretic approach based on the Akaike’s information criterion corrected for a small sample size (AICc; Burnham & Anderson, 2002). The analysis identified the most parsimonious model (lowest AICc) of possible subsets, ranging from the null model (MO, intercept only) to a model with all the considered explanatory variables. This analytical procedure selects the model that provides an accurate approximation to the structural information in the data at

hand, with the smallest possible number of parameters for adequate representation of the data (Burnham & Anderson, 2002). The Akaike weight of models (Wi) was presented —the weight of evidence in favour of the considered model being the best model for the situation at hand (Burnham & Anderson, 2002). The relative importance (RI) of the explanatory variables was estimated —by the sum of the Akaike weights over all models in which that variable appears– to highlight evidence for the importance of each variable within the set of models (Burnham & Anderson, 2002). Explained deviance values (Dev–expl), providing an estimate of the model fit (Wood, 2006), are also presented. All analyses were performed using the R statistical software (R Development Core Team, 2011). Results Variability in wild boar abundance indexes depended on temporal factors —the hunting period and the day in the hunting period— and on the characteristics of the drive hunt —the number of dogs and hunters and the area. Model selection suggests for both wild boars seen and wild boars culled that the best model for the data at hand includes as explanatory factors the year, the day of the season, the interaction between these two factors, the number of dogs and hunters, and the area (table 1 and 2). Over the considered period, the numbers of wild boars seen and culled per drive hunt showed non–linear trends (fig. 1). For the number of wild boars seen per drive hunt, the fitted model suggests an increase of 13% between 2004 and 2005, a decrease of 44% between 2005 and 2009, and an increase of 20% between 2009 and 2011. Between 2004 and 2011, this model suggests an overall reduction of 23% in the number of wild boars seen per drive hunt (fig. 1A, left). For the number of wild boars culled per drive hunt, the selected model suggests an oscillatory pattern with substantial increases (101% between 2004 and 2005; 57% between 2007 and 2008) and decreases (–46% between 2005 and 2007; –66% between 2008 and 2010). Between 2004 and 2011, this model suggests a 14% increase in the number of wild boars culled per drive hunt (fig. 1B, left). These inter–annual trends interact with a multimodal pattern that exhibits variations depending on the hunting period (fig. 2A). The number of wild boars seen per drive hunt was highest at the beginning of the hunting period (early October), in early January, and in February in 2004–2006. However, this pattern changed over the study period and the number of wild boars seen was highest in December and February in 2006–2009. Since 2009, however, the periods with greatest numbers of wild boars seen were the same as in previous years but with the difference that the peaks of abundances in boar seen decreased in comparison with the period 2004–2009 (fig. 2A, left). The number of wild boars seen per drive hunt was positively associated with the number of hunters (at least up to ten hunters) (fig. 2B, left) and also increased strongly in drive hunts with 10–18 dogs (fig. 2B, left). A decrease in the number of wild boar

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Table 1. Model selection for determining factors in the number of wild boar Sus scrofa seen per drive hunt: Y. Hunting period; D. Day of the hunting period; Nd. Number of dogs; Nh. Number of hunters; A. Area where the drive hunt took place; * interaction; K. Number of estimated parameters; AICc. Akaike’s Information Criterion corrected for small sample size, lower values indicate a most–parsimonious model for the observed data; ΔAICc. Difference of AICc between the model and the most parsimonious model; the larger the ΔAICc, the less plausible it is that the fitted model is the best model given the data set; L(gi/x). Probability of the model being the best model given the data set; Wi. Akaike weight of the model; Dev–expl. Explained deviance of the fitted model; RI. Relative Importance of factors. Only the ten best models are reported (Burnham & Anderson, 2002; Wood, 2006). Tabla 1. Selección de modelos para determinar los factores que condicionan el número de jabalíes (Sus scrofa) avistados por batida: Y. Temporada de caza; D. Día de la temporada de caza; Nd. Número de perros; Nh. Número de cazadores; A. Área en la que tuvo lugar la batida; * Interacción; K. Número de parámetros estimados; AICc. Criterio de información de Akaike corregido para un tamaño muestral pequeño, los valores bajos indican un modelo principalmente parsimonioso para los datos observados; ΔAICc. Diferencia de AICc entre el modelo y el modelo más parsimonioso, cuánto mayor sea ΔAICc, menos plausible será que el modelo ajustado sea el mejor para el conjunto de datos; L(gi/x). Probabilidad de que el modelo sea el mejor para el conjunto de datos; Wi. Peso de Akaike del modelo; Dev–expl. Variabilidad explicacada del modelo ajustado; RI. Importancia relativa de los factores. Solo se muestran los diez modelos mejores (Burnham & Anderson, 2002; Wood, 2006). Model

AICc

ΔAICc

L(gi/x)

Dev–expl

Y+D+Y*D+Nd+Nh+A

1490.79

0.00

1.00

0.87

0.30

1.00

Y+D+Y*D+Nd+A

1494.59

3.80

0.15

0.13

0.28

1.00

Y+D+Y*D+Nh+A

1509.94

19.15

0.00

0.28

Y*D 1.00

Y+D+Nd+Nh+A

1540.89

50.10

0.00

0.21

Nd 1.00

Y+D+Y*D+Nd+Nh

1548.35

57.56

0.00

0.25

Nh 0.87

D+Nd+A

1558.53

67.74

0.00

0.17

Y+D+Nh+A

1559.72

68.93

0.00

0.19

Y+D+Nd+A

1560.71

69.92

0.00

0.17

Y+Nd+A

1564.63

73.84

0.00

0.15

D+Nh+A

1567.45

76.65

0.00

0.17

seen was observed in drive hunts with more than 18 dogs, although this variation should be considered with caution due to its small sample size. The number of wild boar culled per drive hunt revealed three key periods in the hunting periods, above all in the periods 2004–2005, 2007–2009, and 2011–2012 (fig. 2A, right): the end of December–early January and February, both characterized by the greatest number of wild boar culled per drive hunt, and lastly, the beginning of the hunting period (although the number of wild boars culled per drive hunt in this period was lower than in the other two periods). The number of wild boars culled per drive hunt also increased with the numbers of hunters and dogs, above all in drive hunts with 12–18 dogs (fig. 2B, right). All the considered factors have very high relative importance (close to 1) in explaining the variability in the indexes of wild boar abundance (table 1 and 2). Yet, the explained deviance of the selected models was moderate (30% for wild boar seen and 23% for wild boar culled), which suggests that the considered factors only provide a partial understanding of the observed variability.

1.00

Discussion Multimodal patterns in wild boar abundance indexes during hunting periods suggest that wild boar conduct seasonal migrations in our study site. Migrations are a more likely explanation than nomadism (Mueller & Fagan, 2008) because the environment is highly seasonal in the Pyrenees and because pulsations in wild boar abundance during the hunting period occur over years and in certain predictable times of the hunting period. The boar mating season at the end of December–early January (Delcroix et al., 1990), for instance, is one of the periods when high abundances of wild boar are most predictable. Thus, the observed variations in wild boar abundance may be linked —at least in part— to the behavioural ecology of the species in the area. The observed evidence of wild boar migrations in our area differs from the sedentary patterns reported in Germany and Australia (Keuling et al., 2008; Mitchell et al., 2009) but agrees with results from Poland and mountainous environments in Italy and in Tennessee, USA (Andrzejewski & Jezierski, 1978; Singer et al., 1981; D’Andrea et al., 1995). Patterns in the use of

Animal Biodiversity and Conservation 36.1 (2013)

Table 2. Model selection for determining factors in the number of wild boar Sus scrofa culled per drive hunt. Only the ten best models are reported. (Burnham & Anderson, 2002; Wood, 2006). (For abbreviations see table 1.) Tabla 2. Selección de modelos para determinar los factores que condicionan el número de jabalíes Sus scrofa abatidos por batida. Solo se han mostrado los diez modelos mejores (Burnham & Anderson, 2002; Wood, 2006). (Para las abreviaturas, ver tabla 1.) Model

AICc

ΔAICc

L(gi/x)

Dev–expl

Y+D+Y*D+Nd+Nh+A

758.41

0.00

1.00

0.94

0.23

0.98

D+Nd+A

766.03

7.62

0.02

0.18

0.98

Y+Nd+A

767.11

8.70

0.01

0.16

Y*D 0.95

Y+D+Y*D+Nd+A

767.60

9.19

0.01

0.19

Nd 0.99 Nh 0.95

Y+D+Nd+A

767.84

9.43

0.01

0.18

Y+D+Nd+Nh+A

769.77

11.36

0.00

0.18

Y+D+Y*D+Nh+A

772.44

14.03

0.00

0.19

Y+D+Nh+A

772.94

14.53

0.00

0.19

Y+Nh+A

774.60

16.19

0.00

0.16

Y+D+Y*D+Nd+Nh

782.29

23.88

0.00

0.17

space in wild boar, therefore, appear to be highly dependent on the environment. As in other European ungulates (Albon & Langvatn, 1992; Mysterud, 1999; Ball et al., 2001) and as previously suggested for the wild boar (Andrzejewski & Jezierski, 1978), migrations may involve only part of the wild boar population (partial migration) and still require further study. The knowledge of wild boar migration in the Pyrenees may stimulate a reappraisal of significant variations in local populations. The sustainable management of migratory species requires an accurate understanding and familiarity with migratory routes (Thirgood et al., 2004; Bolger et al., 2008), and such knowledge would represent a substantial challenge for future management plans. Further studies should aim to characterize the life–history, the spatial scale, the phenology and the determining factors of migration (Ramenofsky & Wingfield, 2007) of wild boar in the Pyrenees. Previous studies on space use in wild boar suggested small home ranges at small time scales (< 1,000 ha in average; Massei et al., 1997; Keuling et al., 2008). The choice of temporal scale at which data are collected and the definition of home range can significantly influence biological inference (Börger et al., 2006). The size of our study site (2,500 ha) and our intense monitoring were key factors that allowed us to reach high–resolution analyses of variations in wild boar abundance. Further studies should use movement data at small temporal scale and take into account reproductive ecology and food availability, not just hunting period. Integrated and high–resolution monitoring is required to unravel the misunderstood complexity of space use in wild boar. The number of wild boar seen and culled per drive hunt varied substantially during the monitoring period

1.00

and the trends in these two abundance indexes differed. Inter–annual variations were greater for wild boars culled than for wild boars seen per drive hunt and on occasions the trends in abundance for each index were different. For instance, in the period 2004–2011, the results for the wild boars seen per drive hunt suggest a reduction of 23%, while the results for wild boars culled per drive hunt suggest an increase of 14%. The number of wild boars seen per drive hunt is probably a more reliable index of wild boar abundance than the number of wild boars culled because it is not dependent on shooting success and because indexes based on seen–individuals have previously been preferred in other ungulate species (Ericsson & Wallin, 1999; Mysterud et al., 2007). Variations in the migratory/resident ratio might also affect the relationship between the numbers of wild boar seen and culled through dilution effects on predation risk (Krause & Ruxton, 2002). Thus, strong inter–annual variations in the number of culled wild boar should be regarded with caution as this abundance index might overestimate or even deviate from true population dynamics. Further studies are required to unravel the relative importance of shooting success. As seen above, the number of wild boars seen per drive hunt suggests a non–linear decrease of 23% in 2004–2011 in our study site. Indirect evidence of wild boar migration were observed and, therefore, further studies should analyse the spatial scale of this decreasing population trend. The population dynamics of wild boar in other areas should also be examined using indexes other than hunting bag alone. The observed inter–annual pattern disagrees with the results reported by Marco et al. (2011) that suggest —on the basis of official hunting statistics and for an area that included

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A Mean number of wild boars

3.2

2.2 1.7 1.2 0.7

2004– 2005– 2006– 2007– 2008– 2009– 2010– 2011– 2005 2006 2007 2008 2009 2010 2011 2012 Hunting period

1.2 Mean number of wild boars

2.7

1.1 1 0.9 0.8 0.7 0.6 0.5 0.4 0.3 0.2

2004– 2005– 2006– 2007– 2008– 2009– 2010– 2011– 2005 2006 2007 2008 2009 2010 2011 2012 Hunting period

Fig. 1. Inter–annual variation in the indexes of wild boar abundance in Urdués (Pyrenees, northern Spain): A. Number of wild boars seen per drive hunt; B. Number of wild boars culled per drive hunt. Points and error bars represent mean values and related standard error. The solid lines represent the predicted patterns estimated by the best models and the dotted lines indicate the related standard error. Fig. 1. Variación interanual en los índices de abundancia de jabalí en Urdués (Pirineos, norte de España): A. Número de jabalíes avistados por batida; B. Número de jabalíes abatidos por batida. Los puntos y las barras de error representan los valores medios y el error estándar relacionado. Las líneas continuas representan los patrones previstos estimados con los mejores modelos y las discontinuas indican el error estándar relacionado.

our study site— that wild boar populations increased in Aragón during this period. This incongruence might be caused by differences in the spatial scale considered. Nevertheless, as highlighted by Martínez–Jauregui et al. (2011), in Spain official hunting statistics can be incomplete and thus this discrepancy may be due to differences in the accuracy and in the completeness

of the available information. Between 1997 and 2002, Acevedo et al. (2006) suggested a population increase in the Pyrenees (woodland habitat) and relative population stability or local decrease in central and south Aragón (which is characterized by a more developed agriculture than the Pyrenees). This might suggest that wild boar population dynamics might have changed in

Animal Biodiversity and Conservation 36.1 (2013)

0.7 of Number d rs culle wild boa

0.6

2.0

0.5

1.5

0.4

F J e h d D f t io N y o per a D ing nt hu

B 5.0 4.5 4.0 3.5 3.0 2.5 2.0 1.5 1.0

0.2

Ye a

F J he d D f t io N y o per Da ing nt hu

of Number rs seen wild boa

of Number d rs culle wild boa

2.0 1.5

r be

g do

0.5

20 18 16 14 12 10 8

1.0

20 18 16 14 12 10 8

0.3

0 42 0 40 0 38 0 36 0 34 0 32 0 30 0 28

Ye a

0 42 0 40 0 38 0 36 0 34 0 32 0 30 0 28

20 20 04 20 05 20 06 20 07 20 08 20 09 20 10 11

0.5

20 20 04 20 05 20 06 20 07 20 08 20 09 20 10 11

1.0

of Number rs seen wild boa

2.5

do of

Fig. 2. Determining factors in the number of wild boar seen (left) and number of wild boar culled (right) per drive hunt: A. Effect of year and day of the hunting period (O. October; N. November; D. December; J. January; F. February); B. Effects of the numbers of dogs and hunters. Fig. 2. Factores determinantes en el número de jabalíes avistados (izquierda) y número de jabalíes abatidos (derecha) por batida: A. Efecto del año y el día de la temporada de caza (O. Octubre; N. Noviembre; D. Diciembre; J. Enero; F. Febrero); B. Efectos del número de perros y de cazadores.

the Pyrenees, at least locally. However, this interpretation should be considered with caution because the results of Acevedo et al. (2006) were based on hunting bag size and their study had a much lower resolution than ours. We showed that inter–annual variations in bag size may deviate from the population dynamics inferred from the number of wild boars seen per drive hunt and, thus, results of previous studies should be considered with caution.

In Aragón, as in other European regions, wild boar can be hunted with no limit on bag size and the assumed population increase may have led authorities to advance the hunting period by two weeks since 2009 (Boletín Oficial de Aragón, 2008, 2009). Variations in wild boar abundance indexes observed during this study question the accuracy of the fine–tuning from one year to the next of the management of wild boar in Aragón. As in wild boar populations in Eastern Europe

(Danilov & Panchenko, 2012), wild boar populations in Western Europe oscillate —both increases and reductions occur— and this could be taken into account in further management plans. As for other game species, the sustainability of harvested wild boar populations —as a limited natural resource— will depend on the integration of results such as those we present here into future management plans to avoid population collapse (Fryxell et al., 2010). Past examples in Italy, Russia, Scandinavia and the United Kingdom already showed that severe population declines in wild boar or even a collapse were possible when the environmental conditions became too adverse (Apollonio et al., 1988; Leaper et al., 1999; Welander, 2000; Rosvold et al., 2010; Danilov & Panchenko, 2012). Thus, further studies are required to unravel the relative importance of down regulation (fructification, agricultural practices), characteristics of wild boar population (demographic structure, genetics) and of top regulation (pathogens, predators and hunting) in population dynamics of wild boar (e.g., Massolo & Mazzoni della Stella, 2006; Rosell et al., 2012). Wild boar population dynamics at our study site does not support the hypothesis of a generalized increase in wild boar densities in woodland areas as the origin of the increasing presence of wild boars in urban environments (Jansen et al., 2007). The hypothesis suggesting that the species might be locally adapting to agricultural and urban environments —where hunting pressure is low and where opportunist foraging might even be facilitated by city dwellers (Cahill & Llimona, 2004)— should be analyzed more closely. Such an adaptation has already been observed in wild boar by other authors (Cahill et al., 2012; Rosell et al., 2012) and the importance of areas where hunting was banned in explaining crop damage was also highlighted (Amici et al., 2012). In conclusion, as a result of the close collaboration with local hunters, our study was able to reveal that wild boar seen per drive hunt decreased by 23% in the period 2004–2011 in our study area in the Pyrenees. The observed patterns of wild boar abundance imply the existence of wild boar migrations and further studies should analyse population dynamics of wild boar in other areas using indexes other than hunting bag alone. Our study also highlighted the fact that the numbers of dogs and hunters affect the number of wild boars seen and culled per drive hunt, and that inter–annual variations in bag size might lead to overestimates and discrepancies with population dynamics inferred from the number of wild boars seen per drive hunt. Thus, there is still a need for further studies on the spatial ecology of wild boar and on the applied ecology of wild boar if we are to move towards sustainable and information–based management of wild boar populations. Acknowledgements We would like to thank Julia López, Agnès Sarasa, David Sarasa and Noelia Sánchez–López for logistic support during the study period. Thanks also to all the members of the Grupo de Caza Mayor de Urdués, and above all to César López, Jesús Jiménez, Francisco

Sarasa & Sarasa

López, Eduardo López, Roberto Jiménez, Pablo López and Gabriel López for their collaboration during the monitoring period. We are grateful to Michael Lockwood and Agnès Sarasa for the English revision and to Cédric Girard–Buttoz and Sylvain Losdat for valuable comments on an earlier draft of this paper. The research activities of MS are partially supported by the Plan Andaluz de Investigación Desarollo e Innovación de la Junta de Andalucía (RNM–118). This work was conducted without specific financial support and complies with current Spanish laws. References Acevedo, P., Escudero, M. A., Munoz, R. & Gortazar, C., 2006. Factors affecting wild boar abundance across an environmental gradient in Spain. Acta Theriologica, 51: 327–336. DOI: 10.1007/BF03192685. Albon, S. D. & Langvatn, R., 1992. Plant phenology and the benefits of migration in a temperate ungulate. Oikos, 65: 502–513. Amici, A., Serrani, F., Rossi, C. M. & Primi, R., 2012. Increase in crop damage caused by wild boar (Sus scrofa L.): the 'refuge effect'. Agronomy for Sustainable Development, 32: 683–692. Andrzejewski, R. & Jezierski, W., 1978. Management of a wild boar population and its effects on commercial land. Acta Theriologica, 23: 309–339. Apollonio, M., Randi, E. & Toso, S., 1988. The systematics of the wild boar (Sus scrofa L.) in Italy. Bolletino di Zoologia, 55: 213–221. Artois, M., Depner, K. R., Guberti, V., Hars, J., Rossi, S. & Rutili, D., 2002. Classical swine fever (hog cholera) in wild boar in Europe. Revue Scientifique et Technique de l’Office International des Epizooties, 21: 287–304. Ball, J. P., Nordengren, C. & Wallin, K., 2001. Partial migration by large ungulates: characteristics of seasonal moose Alces alces ranges in northern Sweden. Wildlife Biology, 7: 39–47. Boletín Oficial de Aragón, 2008. ORDEN de 10 de junio de 2008, del Departamento de Medio Ambiente, por al que se aprueba el Plan General de Caza para la temporada 2008–2009. 15277–15286. Boletín Oficial de Aragón, 2009. ORDEN de 11 de junio de 2009, del Departamento de Medio Ambiente, por al que se aprueba el Plan General de Caza para la temporada 2009–2010. 16253–16263. Bolger, D. T., Newmark, W. D., Morrison, T. A. & Doak, D. F., 2008. The need for integrative approaches to understand and conserve migratory ungulates. Ecology Letters, 11: 63–77. DOI: 10.1111/j.1461– 0248.2007.01109.x. Börger, L., Franconi, N., Ferretti, F., Meschi, F., De Michele, G., Gantz, A. & Coulson, T., 2006. An integrated approach to identify spatiotemporal and individual–level determinants of animal home range size. The American Naturalist, 168: 471–485. Bueno, C. G., Barrio, I. C., García–González, R., Alados, C. L. & Gómez–García, D., 2010. Does wild boar rooting affect livestock grazing areas in alpine grasslands? European Journal of Wildlife Research,

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How do we share food? Feeding of four amphibian species from an aquatic habitat in south–western Romania H. V. Bogdan, S.–D. Covaciu–Marcov, O. Gaceu, A.–S. Cicort–Lucaciu, S. Ferenţi & I. Sas–Kovács

Bogdan, H. V., Covaciu–Marcov, S.–D., Gaceu, O., Cicort–Lucaciu, A.–S., Ferenţi, S. & Sas–Kovács, I., 2013. How do we share food? Feeding of four amphibian species from an aquatic habitat in south–western Romania. Animal Biodiversity and Conservation, 36.1: 89–99. Abstract How do we share food? Feeding of four amphibian species from an aquatic habitat in south–western Romania.— The feeding of four amphibian species (Triturus cristatus, Lissotriton vulgaris, Bombina variegata and Pelophylax ridibundus) was studied in 2011, in south–western Romania. The diet of the newts was uniform and mostly composed of aquatic preys The diet of the anurans was more diversified, comprising more prey taxa, mostly terrestrial. The trophic niches of the two newt species overlapped highly but differed from those of the anurans. The trophic niches of the anurans differed from one another. The differences among the four species' diets were determined by the use of different trophic resources, originating from different environments, and by their different sizes. The newts’ diet was less diversified because the aquatic habitat was small and poor in trophic availability. The anurans used the aquatic habitat as a base from where they captured terrestrial preys in the surrounding terrestrial environment. Key words: Amphibian community, Diet, Food composition, Trophic niche, Diversity. Resumen ¿Cómo compartimos la comida? La alimentación de cuatro especies de anfibios de un hábitat acuático en el sureste de Rumania.— En 2011 se estudió la alimentación de cuatro especies de anfibios (Triturus cristatus, Lissotriton vulgaris, Bombina variegata y Pelophylax ridibundus) en Rumania. La dieta de los caudados era uniforme y se componía principalmente de presas acuáticas. La dieta de los anuros era más diversificada y comprendía más taxones de presas, en su mayor parte terrestres. Los nichos tróficos de las dos especies de caudado se solapaban en gran medida, pero eran distintos de los de los anuros. Los nichos tróficos de los anuros diferían entre sí. Las diferencias entre las dietas de las cuatro especies se debían a la utilización de recursos tróficos diferentes, procedentes de diversos ambientes, y a sus tallas distintas. La dieta de los caudados era menos diversificada porque el hábitat acuático era reducido y la disponibilidad de alimentos en el mismo, escasa. Los anuros utilizaban los hábitats acuáticos como base desde la que capturaban presas terrestres en el ambiente terrestre circundante. Palabras clave: Comunidad de anfibios, Dieta, Composición de los alimentos, Nicho trófico, Diversidad. Received: 10 XII 12; Conditional acceptance: 28 I 13; Final acceptance: 5 IV 13 Horia V. Bogdan, Severus–Daniel Covaciu Marcov, Alfred–Stefan Cicort–Lucaciu, Sára Ferenţi, István Sas– Kovács, Dept. of Biology, Fac. of Sciences, Univ. of Oradea, Universităţii str. 1, Oradea 410087, Romania.– Ovidiu Gaceu, Dept. of Geography, Fac. of Geography, Turism and Sport, Univ. of Oradea, Universităţii str. 1, Oradea 410087, Romania.– Sára Ferenţi, Dept. of Biology, Fac. of Biology and Geology, 'Babes–Bolyai' Univ., Gheorghe Bilascu (Republicii) str. 44, 400015 Cluj–Napoca, Romania and 'Iosif Vulcan' National College, Jean Calvin str. No. 5, Oradea, Romania. Corresponding author: S. D. Covaciu–Marcov. E–mail: severcovaciu1@gmail.com

ISSN: 1578–665X

Bogdan et al.

Introduction Small, isolated wetlands are extremely important ecologically but are among the most threatened ecosystems (see Pitt et al., 2012). For amphibians, a declining group due to multiple causes (e.g. Stuart et al., 2004; Collins, 2010), it is important to maintain a high diversity of aquatic habitats at a small spatial scale (e.g. Hartel, 2008; Hartel et al., 2011). However, not all amphibians are equally tied to the same aquatic habitats (Hartel et al., 2011). Thus, it is important to establish to what extent the aquatic habitats of amphibians meet their needs. Feeding can be considered a useful indicator for this purpose (see Kovacs et al., 2007). In recent years, several papers about the feeding of some amphibian species have been published from Romania (e.g. Aszalos et al., 2005; Balint et al., 2010; Cicort–Lucaciu et al., 2005, 2007a, 2007b, 2011; Cogalniceanu et al. 2000; Covaciu–Marcov et al., 2010a, 2010b, 2010c, 2011, 2012; Dobre et al., 2007; Kovacs et al., 2007; Sas et al., 2009; Iftime & Iftime, 2011), but most have been from the north–western part of the country. The few data available regarding the south west focus only on some species (e.g. Bogdan et al., 2011, 2012a). To date, there are few studies on the feeding of more amphibian species in relation to the habitat and with the way they use its resources (e.g. Fasola & Canova, 1992; Covaciu–Marcov et al., 2002, 2010c, 2012; Bisa et al., 2007; Guidali et al., 2000; Cicort–Lucaciu et al., 2011; Cogalniceanu et al., 2000; Junca & Eterovick, 2007), and even fewer studies that compare the diet of newts with that of anurans (e.g. Vignoli et al., 2009). To the best of our knowledge, such studies in Romania have only been made separately. This study was performed to determine the food composition of four amphibian species [Triturus cristatus (Laurenti, 1768), Lissotriton vulgaris (Linnaeus, 1758), Bombina variegata (Linnaeus 1758) and Pelophylax ridibundus

(Pallas 1771)] in an aquatic habitat in south–western Romania, to establish the differences between them, and to compare the trophic niches and the way the trophic resources are exploited. Material and methods The study area was located near Maru village, in the Ţarcu Mountains, in south–western Romania (45° 27' 26.21'' N, 22° 26' 42.13'' E). The habitat was a pond of approximately 8 m long and 4 m wide, situated at an altitude of 430 m a.s.l. In this habitat we found large newt populations (see Bogdan et al., 2012b, also see a detailed habitat description). The study took place between March and July, in 2011. We analyzed 599 amphibian stomach contents (407 newts and 192 anurans) during a total of six field trips (table 1). Field trips were made once every three weeks for the newts and one a month for the anurans, after the newts had left the habitat. The study was stopped in July because of drought. The amphibians were captured with nets. Stomach contents were sampled using the stomach flushing method (Solé et al., 2005), after which the amphibians were set free. The composition of the food was analyzed by percentage abundance (%A) and frequency of occurrence (%f), while also establishing the origin for each prey (aquatic or terrestrial). Food diversity was estimated using the Shannon–Wiever Index (H) (Shannon & Wiever, 1949). The trophic niche overlap, using the percentage abundance of the food items, was calculated using the Pianka Index (Q) (Pianka, 1973). The obtained pairwise values were used to create a correlation matrix to perform the tree clustering analysis (Statistica 6.0). The samples from different periods at the same species and between different species from the same period were compared by the values of the frequency of occurrence of the food items using the Mann–Whitney test (Zar, 1999).

Table 1. The number of studied amphibians, the frequency of occurrence of empty stomachs, the stomachs with animal content, plant fragments, shed–skin, amphibian eggs, and inorganic elements: Tc. Triturus cristatus; Lv. Lissotriton vulgaris; Pr. Pelophylax ridibundus; Bv. Bombina variegata.

25 III 2011

14 IV 2011

5 V 2011

% of empty stomachs

6.00

1.96

–

2.00

–

% with animal content

74.00

96.08

90.24

96.00

90.00 100.00

% with non–animal content

92.00

82.35

95.12

94.00

100.00 78.00

% with plant fragments

50.00

62.75

65.85

78.00

88.00

58.00

% with shed–skin

42.00

29.41

53.66

30.00

36.00

40.00

% with amphibian eggs

64.00

5.88

58.54

16.00

26.00

6.00

% with inorganic elements

6.00

7.84

2.44

20.00

–

No. of studied amphibians

Animal Biodiversity and Conservation 36.1 (2013)

Results Feeding differed for each species of amphibian and for each period. Unfed animals were more numerous in spring (table 1). We identified two types of stomach content: animal and non–animal. The first category consisted of 4,574 prey belonging to 62 prey taxa. Only quantitatively important taxa are shown in tables 2 and 3. Taxa found in low amounts or consumed only occasionally are included in the category 'Others' (Turbelaria–Planariidae, Nematomorpha, Hirudinea, Gastropoda, Arachnida–Acarina, Crustacea–Amphipoda, Crustacea–Ostracoda, Crustacea–Isopoda, Chilopoda, Plecoptera, Coleoptera–larve, Lepidoptera, Homoptera–Aphididae, Ortoptera, Dermaptera, Mecoptera, Trichoptera, Urodela–Larve, Urodela–L. vulgaris, Diptera–Brahicera–Tabanidae, Diptera–Nematocera–Tipulidae, Hymenoptera–undetermined and eight Coleopteran groups (Cantaridae, Carabidae, Coccinelidae, Curcullionidae, Chrysomelidae, Elateridae, Scarabeidae, Staphylinidae). The most important prey taxa for the newts were aquatic Nematocera larvae, Crustacea Cladocera and Copepoda. L. vulgaris consumed microcrustaceans more frequently, while T. cristatus fed more frequently on larger preys (earthworms or mayfly larvae). For the anurans, the most important prey were coleopterans and dipterans (tables 2, 3). L. vulgaris had the highest value for average number of preys/individual (12.42). The maximum number of preys/individual (37) was the same in the case of the two newts. Newts consumed the lowest number of preys at the beginning of spring (tables 2, 3). Most prey consumed by newts were aquatic, while most prey consumed by anurans were terrestrial (table 4). The non–animal stomach contents consisted of plant remains, shed skin fragments, amphibian spawn, and inorganic elements. The plant fragments and shed skin were consumed by all species, throughout all

the periods, while the inorganic elements were also consumed by all the species but not in each period. Spawn was consumed only by newts, varying in frequency during the seasons (table 1). Generally, the non–animal stomach contents appeared together with animal prey, but a few individuals from all species consumed them exclusively. P. ridibundus showed the highest food diversity (H = 3.161 in May). This species showed more diverse feeding than B. variegata during every sampling period (table 4). Newts had the lowest food diversity (table 4). During most sampling periods, L. vulgaris had a less diversified diet than T. cristatus, but the latter still had the smallest diversity value throughout the entire study (H = 0.645 in May). The trophic niches that overlapped most were those of the newts (fig. 1), both during the same period and in different time frames (between T. cristatus from May and July, Q = 0.99). The two anuran species used different trophic niches than newts (fig. 1). We also found differences in the trophic niches between B. variegata and P. ridibundus. Differences between the newts’ feeding were never significant, but those between the newts and P. ridibundus, were always significant. The greatest differences were found between P. ridibundus and L. vulgaris on the 18th of June (Mann–Whitney test Z = –4.022, df = 1, P = 0.000010). Significant differences were also recorded between the feeding of the newts and B. variegata. Significant differences occurred between B. variegata and P. ridibundus only one time (Mann–Whitney test Z = 2.247, df = 1, P = 0.0225). Discussion The food composition of the four amphibian species resembled the diet of other populations of the same species previously studied, both in Romania (e.g. Do-

Tabla 1. Número de anfibios estudiados, frecuencia de la presencia de estómagos vacíos, estómagos con contenido animal, fragmentos de plantas, mudas de piel, huevos de anfibios y elementos inórganicos: Tc. Triturus cristatus; Lv. Lissotriton vulgaris; Pr. Pelophylax ridibundus; Bv. Bombina variegata. 28 V 2011

18 VI 2011

19 VII 2011

–

1.69

–

3.33

–

95.56

97.96

88.14

90.91

98.04

96.00

96.67

93.33

95.24 90.00

95.56

87.76

71.19

95.45

96.08

90.00

70.00

76.67

76.19 86.67

51.11

65.31

54.24

95.45

74.51

62.00

53.33

63.33

61.90 76.67

37.78

10.20

16.95

36.36

13.73

26.00

16.67

13.33

23.81 23.33

60.00

40.82

–

62.75

20.00

–

15.56

14.29

10.17

–

17.65

8.00

10.00

3.33

–

Bogdan et al.

Table 2. Abundance of animal prey in percentages: l. Larvae; aq. Aquatic; ter. Terrestrial; * Mean values under 3%. (For abbreviations of species see table 1.)

25 III 2011

14 IV 2011

5 V 2011

Oligocheta – Lumbricidae

3.28

9.91

–

Gasteropoda [snails] [aq.]

–

4.20

–

Gasteropoda [snails] [ter.]

–

Bivalvia

–

8.40

–

Arachnida – Araneae

–

Crustacea – Cladocera

9.02

9.50

8.62

Crustacea – Copepoda

25.41

71.60

–

32.85

17.73

Crustacea – Ostracoda

–

Collembola

–

7.38

3.50

9.05

15.13

11.82

Ephemeroptera [l.] Ephemeroptera

–

Plecoptera [l.]

–

5.60

4.67

–

Odonata [l.]

–

5.04

–

Odonata

–

Homoptera – Cicadellidea

–

Heteroptera [aq.]

–

Heteroptera [ter.]

–

Coleoptera – Dytiscidae

–

Coleoptera – undet. [ter.]

–

Coleoptera – Carabidae

–

Trichoptera [l.]

6.56

3.45

3.78

Lepidoptera [l.]

–

45.08

16.87

61.21

46.38

44.12

56.16

Diptera – Nematocera – Culicidae

Diptera – Nematocera [l.]

–

Diptera – Brahicera [l.]

–

10.08

–

Diptera – Brahicera [l.]

–

Diptera – Brahicera – Muscidae

–

Hymenoptera – Formicidae

–

Hymenoptera – Apidae

–

Anura [l.]

–

3.02

3.27

8.03

7.76

6.6

9.25

5.67

Others

bre et al., 2007; Covaciu–Marcov et al., 2010a, 2011; Ferenţi et al., 2010; Cicort–Lucaciu et al., 2011; Bogdan et al., 2012a) and in other regions (e.g. Çiçek & Mermer, 2006; Paunović et al., 2010; Mollov, 2008; Mollov et al., 2010; Kuzmin, 1990). This is more obvious for newts, which more frequently consume micro–crustaceans and Nematocera larvae (e.g. Dobre et al., 2007; David et al., 2009; Covaciu–Marcov et al., 2010a). However,

the differences in feeding in different habitats seems to be greater for anurans than for newts (e.g. Tóth et al., 2007; Ferenţi et al., 2010; Çiçek, 2011). Although we found differences between the food composition for all four species, those between newts and anurans were almost always significant. The newts' diet was more uniform and focused on the aquatic habitat’s trophic resources. The fact that

Animal Biodiversity and Conservation 36.1 (2013)

Tabla 2. Abundancia de presas animales en porcentajes: l. Larvas; aq. Acuático; ter. Terrestre; * Valores medios inferiores al 3 %. (Para las abreviaturas de las especies ver tabla 1.)

28 V 2011

18 VI 2011

19 VII 2011

–

6.12

–

3.10

–

8.96

3.33

–

4.42

–

6.19

6.72

4.44

6.19

–

18.21

–

37.37

–

6.87

–

3.33

–

28.97

–

2.06

4.11

–

10.28

–

4.42

–

3.74

–

4.42

3.33

–

4.12

–

3.40

–

6.11

4.12

–

7.48

–

3.98

–

12.93

5.61

11.50

8.96

–

9.86

3.74

–

11.06

10.56 18.56 4.44

–

4.48

87.40

68.06

–

73.78

53.38

–

15.46

–

5.10

10.28

8.67

–

3.54

22.39

10.31

–

7.48

3.56

–

9.52

–

15.49

8.96

33.89 15.46

–

4.42

6.54

–

7.08

5.97

5.00

11.34

–

5.00

–

8.49

6.86

32.33

23.36

10.66

9.25

23.9

12.38

newts hunt in the aquatic habitat has been recorded previously (e.g. Denoël & Andreone, 2003; Kutrup et al., 2005; Vignoli et al., 2009; Covaciu–Marcov et al., 2010a, 2010b), and few exceptions have been documented (Cicort–Lucaciu et al., 2007a). The reduced food diversity of the newts is a consequence of the consumption of aquatic preys, which are probably less diversified in a small habitat. Quantitatively,

33.64 33.56

however, newts consume a greater number of prey than anurans, mainly due to the consumption of microcrustaceans. Microcrustaceans are the basis of the diet of L. vulgaris, a smaller species (Fuhn, 1960) that feeds on the smallest prey. Anuran feeding focuses on terrestrial preys. This brings them into contact with numerous insects, a group with a high number of species, mostly terrestrial (see Radu &

Bogdan et al.

Table 3. The frequency of occurrence of animal prey (are included the same taxa as for percentage abundance): l. Larvae; aq. Aquatic; ter. Terrestrial. (For abbreviations of species see table 1.)

25 III 2011

14 IV 2011

5 V 2011

Oligocheta – Lumbricidae

8.00

1.96

31.71

–

Gasteropoda [snails] [aq.]

–

7.32

–

8.00

–

Gasteropoda [snails] [ter.]

–

Bivalvia

–

12.00

2.00

–

Arachnida – Araneae

–

1.96

–

Crustacea – Cladocera

8.00

7.84

2.44

26.00

4.00

14.00

Crustacea – Copepoda

18.00

74.51

–

50.00

2.00

28.00

Crustacea – Ostracoda

–

1.96

–

2.00

Collembola

–

2.00

18.00

23.53

26.83

22.00

50.00

56.00

–

2.00

–

Ephemeroptera [l.] Ephemeroptera Plecoptera [l.]

–

9.76

16.00

–

8.00

4.00

–

8.00

22.00

–

Odonata

–

Homoptera – Cicadellidea

–

2.44

–

Odonata [l.]

Heteroptera [aq.]

–

2.44

–

2.00

Heteroptera [ter.]

–

2.00

–

2.00

–

4.00

–

2.44

2.00

8.00

–

Coleoptera – Dytiscidae Coleoptera – undet.

–

2.00

–

Trichoptera [l.]

Coleoptera – Carabidae

14.00

13.73

17.07

4.00

18.00

4.00

Lepidoptera [l.]

–

2.00

–

2.00

44.00

66.67

65.85

92.00

60.00

80.00

Diptera – Nematocera [l.] Diptera – Nematocera – Culicidae

–

7.32

–

6.00

Diptera – Brahicera [l.]

–

16.00

–

Diptera – Brahicera [l.]

–

2.00

–

Diptera – Brahicera – Muscidae

–

1.96

–

4.00

–

2.00

Hymenoptera – Formicidae

–

Hymenoptera – Apidae

–

Anura [l.]

–

12.20

2.00

Radu, 1967), so amphibians that feed from the terrestrial environment have a more diversified diet. Of the two anurans, P. ridibundus consumed a higher amount of terrestrial preys a fact also observed for other populations (e.g. Çiçek & Mermer, 2006; Mollov et al., 2010). Unlike P. ridibundus, B. variegata more often used the trophic resources from the aquatic habitat, a finding also noted previously (e.g. Ferenţi et al., 2010; Covaciu–Marcov et al., 2011).

The newts showed a uniform diet, as other populations from western Romania (e.g. Cicort–Lucaciu et al., 2005, 2007a, 2007b; Covaciu–Marcov et al., 2010a), probably because the habitats they used had approximately the same trophic offer. However, in other areas, the newts' diet may present differences, sometimes focusing on leeches (Griffiths & Mylotte, 1987). The overlap of the newts' trophic niches was also a consequence of the almost exclusive use of the

Animal Biodiversity and Conservation 36.1 (2013)

Tabla 3. Frecuencia de la presencia de presas animales (se incluyen los mismos taxones que para el porcentaje de abundancia): l. Larvas; aq. Acuático; ter. Terrestre. (Para las abreviaturas de las especies ver tabla 1.) 28 V 2011

18 VI 2011

19 VII 2011

4.44

–

13.56

–

16.67

3.33

4.76

–

1.96

–

6.78

9.09

–

3.33

23.33

9.52

3.33

–

2.04

–

1.96

–

20.34

9.09

7.84

–

36.67

23.33

33.33

20.00

–

30.61

–

1.96

54.00

–

12.24

–

4.00

–

1.96

–

1.69

9.09

–

3.33

–

3.33

26.67

4.08

1.69

–

9.80

10.00

–

20.00

–

10.17

18.18

11.76

–

16.67

10.00

9.52

3.33

–

2.00

3.33

–

8.89

4.08

–

1.96

6.00

10.00

–

14.29

–

18.64

9.09

–

10.00

–

9.52

3.33

–

3.39

13.64

–

30.00

3.33

23.81

–

4.44

2.04

3.39

4.55

–

2.00

10.00

–

10.00

–

2.04

15.25

9.09

–

16.67

10.00

47.62

10.00

4.44

–

6.78

31.82

–

20.00

6.67

9.52

–

4.44

4.08

32.20

18.18

7.84

2.00

63.33

30.00

47.62

43.33

–

28.81

13.64

–

43.33

6.67

28.57

–

2.00

3.33

–

4.76

6.67

–

5.08

4.55

5.88

2.00

6.67

20.00

14.29

6.67

93.33

91.84

1.69

–

76.47

80.00

3.33

–

9.52

30.00

–

20.34

22.73

17.65

–

23.33

56.67

9.52

16.67

–

9.09

–

10.00

–

1.69

9.09

17.65

–

9.52

3.33

–

4.08

32.20

13.64

–

2.00

50.00

26.67

71.43

36.67

4.44

–

18.64

22.73

1.96

–

33.33

16.67

33.33

30.00

–

1.69

–

6.67

–

23.81

–

2.22

–

1.69

–

1.96

–

3.33

–

trophic resources in the aquatic habitat. Using this habitat, the newts showed a similar food composition. The anurans, however, using the terrestrial environment, did not overlap their trophic niches, a finding also observed previously (e.g. Cogalniceanu et al., 2000). Thus, in both the newts and the anurans diet is as diversified as the habitat from where they hunt. For all four species, the differences between the feeding were due to the differences in size. L. vulgaris

and B. variegata are smaller than T. cristatus and P. ridibundus, respectively (see Fuhn, 1960), their diet being therefore less diversified and their number of prey taxa more limited. Trophic partitioning due to size of the predator has been indicated previously (e.g. Kuzmin, 1990; Cogalniceanu et al., 2000; Bisa et al., 2007; Ferenţi & Covaciu–Marcov, 2011). Aside from the fact that T. cristatus consumed larger preys, it also consumed spawn more frequently than L.

Bogdan et al.

Table 4. The number of animal prey, the maximum and average number of animal prey / individual, the percentage abundance of aquatic and terrestrial prey and the Shannon–Weaver diversity index. (For abbreviations of species see table 1.)

25 III 2011

14 IV 2011

5 V 2011

No. of animal preys

122

486

232

621

238

406

% of aquatic preys

96.72

99.18

87.07

98.87

96.22

98.28

% of terrestrial preys

3.28

0.82

12.93

1.13

3.78

1.72

2.44

9.53

5.66

12.42

4.76

8.12

Feeding diversity (Shannon–Wiever index) 1.55

0.98

1.47

1.39

1.91

1.37

Maximum no. of preys/individual Average no. of preys/individual

Single linkage 1–Pearson r T–III L–IV L–Va T–IV T–Vb T–VI L–Vb L–VI T–Va L–III P–Vb P–VI P–VII B–VIII B–VI B–Vb 0.0

0.1

0.2 0.3 Linkage distance

0.4

0.5

0.6

Fig. 1. Tree clustering based on Pianka’s pairwise values, representing the trophic niche overlap of the studied species over the study period: III. March; IV. April; Va. Start of May ; Vb. End of May; VI. June; VII. July; T. Triturus cristatus; L. Lissotriton vulgaris; P. Pelophylax ridibundus; B. Bombina variegata. Fig. 1. Tres agrupamientos basados en los valores de superposición de Pianka que representan la superposición de los nichos tróficos de las especies estudiadas a lo largo del período del estudio: III. Marzo; IV. Abril; Va. Inicio de mayo; Vb. Final de mayo; VI. Junio; VII. Julio; T. Triturus cristatus; L. Lissotriton vulgaris; P. Pelophylax ridibundus; B. Bombina variegata.

Animal Biodiversity and Conservation 36.1 (2013)

Tabla 4. Número de presas animales, número máximo y promedio de presas animales por individuo, porcentaje de la abundancia de presas acuáticas y terrestres e índice de diversidad de Shannon– Weaver. (Para las abreviaturas de las especies ver tabla 1.) 28 V 2011

18 VI 2011

19 VII 2011

365

335

294

107

450

281

226

134

180

95.07 97.91

5.78

10.28

82.00

98.58

9.29

3.73

5.56

21.65

4.93

2.09

94.22

89.72

18.00

1.42

90.71 96.27

94.44

78.35

8.11

6.84

4.98

4.86

8.82

5.62

7.53

4.47

8.57

3.23

0.64

1.06

3.16

2.52

1.17

1.12

3.05

2.85

2.60

2.42

vulgaris. In contrast, L. vulgaris consumed smaller preys from the water body, such as microcrustaceans. These results again show how these two species use the aquatic habitat (e.g. Dolmen & Koksvik, 1983; Covaciu–Marcov et al., 2010a). The feeding of the four species showed marked seasonal variations. The average number of consumed preys/individual (feeding intensity) increased in the warm season for the newts, whereas for the anurans, and especially for B. variegata, feeding intensity decreased in summer, probably due to the drought. In other cases the intensity of amphibian feeding was lower in summer, due to the higher temperature (e.g. Yu et al., 2009). Seasonal variation in the feeding of amphibians determined by environmental conditions has been frequently documented (see Fasola & Canova, 1992; Kovács et al., 2007; Sas et al., 2009; Bogdan et al., 2012a). Despite the fact that they occupied the same habitat, and that their trophic niches overlapped considerably, the two newt species coexisted, benefiting as in other cases by the generalist feeding and the high trophic offer (see Vignoli et al., 2009). Contact between the two species was low, as in other cases T. cristatus fed occasionally on common newts (e.g. Cicort–Lucaciu et al., 2005), while at Maru this was only recorded once. In the case of the anurans, competition is avoided by using different trophic niches in relation to their size, P. ridibundus hunting larger preys. Using different hunting grounds modifies the feeding of anurans even in habitats with an abundant yet uniform trophic offer (Covaciu–Marcov et al., 2010c). The habitat in Maru seemed to satisfy the amphibians’ trophic needs, as they consumed abundant and diversified food. Anurans and urodelas overlapped in the habitat only for a short period. However, even when they overlapped in space and time, they did not explore the same trophic resources and thus did not compete with each other. The newts consumed a lot of small aquatic preys, while anurans consumed many and diverse terrestrial preys, exploiting the aquatic habitat differently both in space

and time. For the newts, the aquatic habitat is crucial for feeding, and ultimately for survival of the population. As this is tied to the aquatic habitat, the newts are more vulnerable to its changes. This is of concern, as newts are very attached to the same aquatic habitat, returning to it for mating year after year (e.g. Joly & Miaud, 1989; Sinsch et al., 2006). For anurans, individuals migrate between habitats, as described for B. variegata (e.g. Hartel, 2008). The anurans probably used the habitat from Maru only for laying eggs, before moving to some of the streams in the area. Thus, the results from Maru confirm that for diverse amphibian communities comprising both newts and anurans, conservation strategies should look after both the aquatic habitat and the neighboring terrestrial areas (see Dodd & Cade, 1998; Semlitsch & Bodie, 2003; Denoël & Lehman, 2006). Acknowledgements This work was partially supported by the strategic grant POSDRU/88/1.5/S/53501, Project ID53501 (2009), co–financed by the European Social Fund–Investing in People, within the Sectorial Operational Programme Human Resources Development 2007–2013. References Aszalós, L., Bogdan, H., Kovács, É.–H. & Peter, V. I., 2005. Food composition of two Rana species on a forest habitat (Livada Plain, Romania). North– Western Journal of Zoology, 1: 25–30. Balint, N., Indrei, C., Ianc, R. & Ursut, A., 2010. On the diet of the Pelopylax ridibundus (Anura, Ranidae) in Ticleni, Romania. South Western Journal of Horticulture, Biology and Environment, 1(1): 57–66. Bisa, R., Sfenthourakis, S., Fraguedakis–Tsolis, S. & Chondropoulos, B., 2007. Population density and food analysis of Bombina variegata and Rana graeca in mountainous riverine ecosystems of

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A new species of Mysidopsis (Crustacea, Mysida, Mysidae) from coastal waters of Catalonia (north–western Mediterranean) C. San Vicente

San Vicente, C., 2013. A new species of Mysidopsis (Crustacea, Mysida, Mysidae) from coastal waters of Catalonia (north–western Mediterranean). Animal Biodiversity and Conservation, 36.1: 101–111. Abstract A new species of Mysidopsis (Crustacea, Mysida, Mysidae) from coastal waters of Catalonia (north–western Mediterranean).— A new species of the genus Mysidopsis (Crustacea, Mysida, Mysidae, Leptomysinae) is described based on specimens sampled with a suprabenthic sled in the littoral located near the coastal city of Mataró (north–western Mediterranean). The new species lives in the soft–bottom suprabenthic habitat near a Posidonia oceanica meadow, at depths between 17 and 21 m. The main distinguishing features of Mysidopsis iluroensis n. sp. are the small body size, a prominent rostrum, the absence of carapace dorsal nodules, and the armature of the antennule, telson and uropod. The morphology of the new species is compared with other species of Mysidopsis in the Mediterranean Sea. Key words: Mysida, Leptomysinae, New species, Mediterranean. Resumen Una nueva especie de Mysidopsis (Crustacea, Mysida, Mysidae) de las aguas costeras de Cataluña (Mediterráneo noroccidental).— Se describe una nueva especie del género Mysidopsis (Crustacea, Mysida, Mysidae, Leptomysinae) a partir de los ejemplares muestreados con un trineo suprabentónico en el litoral cercano a la ciudad de Mataró (Mediterráneo noroccidental). Esta nueva especie vive en el hábitat suprabentónico de los fondos blandos cercanos a una pradera de Posidonia oceanica, entre 17 y 21 m de profundidad. Las principales características que definen a Mysidopsis iluroensis sp. n. son su pequeña talla, un rostro prominente, la ausencia de nódulos dorsales en el caparazón y la armadura de la anténula, el telson y el urópodo. La morfología de la nueva especie se compara con las otras especies de Mysidopsis descritas en el mar Mediterráneo. Palabras clave: Mysida, Leptomysinae, Nueva especie, Mediterráneo. Received: 20 I 13; Conditional acceptance: 13 III 13; Final acceptance: 2 IV 13 Carlos San Vicente: c/ Nou 8, 43839 Creixell, Tarragona, Espanya (Spain). E–mail: csanvicente@gencat.cat

ISSN: 1578–665X

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Introduction

Material and methods

The Mediterranean Sea is one of the most well studied geographical areas in the world, but many regions and habitats remain insufficiently studied and little is known about several taxonomic groups (Coll et al., 2010). Additionally, the Mediterranean Sea is a complex region where ecological and human interaction has a strong impact on marine biodiversity. It is therefore of high priority to improve descriptions of species and to expand our knowledge of their geographical distribution. Mysids are an important component of the mobile fauna associated with seagrass meadows and algal communities (Ledoyer, 1962; Mazzella et al., 1989; Scipione et al., 1996; Sánchez–Jerez et al., 1999; Wittmann, 2001; Gan et al., 2010). Their diversity and abundance in the Mediterranean Posidonia oceanica meadows has been related to several environmental factors, including the influence of edge effects (Barberá–Cebrián et al., 2002). These edges could be considered a supplementary habitat for mobile fauna because they increase the spatial heterogeneity of seagrass ecosystems (Sánchez–Jerez et al., 1999). During a monitoring program of the Posidonia oceanica meadow off the Mataró coast (Barcelona, NE Iberian peninsula), many suprabenthic Mysida were sampled in the neighbouring soft bottoms in 2001 and 2011. The following species were collected: the Erythropinae Erythrops erythrophthalma (Goës, 1864), the Gastrosaccinae Anchialina agilis (Sars, 1877), the Leptomysinae Leptomysis gracilis (Sars, 1864) and Paraleptomysis banyulensis (Bacescu, 1966), as well as some individuals ascribed to a new species within the genus Mysidopsis (Leptomysinae). This paper deals with the morphological description of this new taxon and provides an updated identification key to the known European Mysidopsis. The genus Mysidopsis Sars, 1864 presently contains a heterogeneous group of 47 living species (Anderson, 2010; Mees & Meland, 2012), showing a wide variety of morphological characters (Bacescu, 1968; Brattegard, 1969; Tattersall, 1969; Bacescu & Gleye, 1979; Mauchline, 1980; Price et al., 1994; Bravo & Murano, 1996). The first three species described in genus Mysidopsis —M. didelphys (Norman, 1863), M. angusta Sars, 1864 and M. gibbosa Sars, 1864— are known from the north–eastern Atlantic and the Mediterranean. Within their respective distributional areas, all of them have been reported in numerous studies such as Sars (1872), Colosi (1929), Bacescu (1941), Tattersall & Tattersall (1951), Mauchline (1970), Mauchline & Murano (1977), Lagardère & Nouvel (1980), Sorbe (1982), Müller (1993), Cunha et al. (1997), and San Vicente (2004). Mysidopis cachuchoensis San Vicente et al., 2012 is the latest species to be described from bathyal soft–bottoms specimens of the Bay of Biscay. These four known European species are morphologically well described and easy to identify (Sars, 1872; Tattersall & Tattersall, 1951; Tattersall, 1969; Mauchline, 1971; San Vicente et al., 2012).

The specimens examined in the present study were recorded at Mataró II and Mataró III stations located at the edges of a Posidonia oceanica meadow at 20.6 and 17 m depth, respectively (fig. 1). This well conserved Posidonia oceanica meadow covers more than 730 hectares of substrate and the distribution of specimens has been studied in detail by Manzanera & Cardell (2002). In 2001 and 2011, the mysid fauna was qualitatively sampled using a suprabenthic sled with a rectangular opening of 50 cm wide and 25 cm high, equipped with a 0.5 mm mesh size net, designed to skim over the surface of the sediment to collect the fauna swimming within the near bottom water layer. The sled was towed by a scuba diver who opened and closed the net at the beginning and end of the bottom sampling. On board, samples were fixed with a solution of 4% formalin in sea water. At the laboratory, the Mysidopsis specimens were sorted and conserved separately in 70% ethanol for later examination. The total body length (TL) of individuals was measured from the apex of the rostrum to the posterior end of the telson, excluding setae. Specimens selected for the species description were dissected and temporarily mounted on slides. Dissected appendages were drawn with the aid of a camera lucida mounted on a Zeiss Axioscop 20 microscope. The terminology for cuticle projections (spines and setae) follows Watling (1989) and Garm (2004). Type material is deposited at the Institut de Ciències del Mar (CSIC) of Barcelona, Spain. The description below refers to both sexes, unless otherwise stated. Nomenclature of higher Mysida taxa follows Mees & Meland (2012). Results Taxonomy Order Mysida Haworth, 1825 Family Mysidae Haworth, 1825 Subfamily Leptomysinae Hansen, 1910 Mysidopsis G. O. Sars, 1864 Mysidopsis iluroensis n. sp. (figs. 2–5) Material examined Holotype: (catalogue number: ICMM12120401), 1 mature male, 4.0 mm TL, 14 VI 2001, station Mataró II, 41º 31' 37.9029'' – N, 2º 28' 17.861'' E, 20.6 m depth, 0–25 cm near–bottom layer; dissected, one vial. Paratypes: (ICMM12120402), 1 mature female, 4.9 mm TL, station Mataró II, data as for holotype, not dissected, one vial; (ICMM12120403), 1 mature male, 4.0 TL, 29 VI 2011, station Mataró III, 41º 30' – 2.574'' N, 2º 28' 56.901'' E, 17 m depth, 0–25 cm near–bottom layer, dissected, one vial. Other material: (ICMM12120404), 5 juveniles (1.9, 2.0, 2.0 2.1, 2.1 mm TL), station Mataró II, not dissected, one vial; (ICMM12120405), 3 immature males (2.7, 3.6 and 3.7 mm TL), station Mataró III,

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2º 25' E

2º 26' E

2º 27' E

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2º 28' E

2º 29' E

2º 30' E

2º 31' E 5 10 15

41º 33' N

20 25

41º 32' N

Mataró II 5 41º 31' N

Mataró III

20 15

Rocky substrate

41º 30' N 30

Rocky substrate with Posidonia Cymodocea nodosa Posidonia oceanica

Fig. 1. Location of the two stations sampled with a suprabenthic sled in the Posidonia oceanica meadow edge off the Mataró coast (NE Iberian peninsula) in 2001 (Mataró II) and 2011 (Mataró III). Posidonia and Cymodocea meadow distribution and isobaths (in metres) modified following Manzanera & Cardell (2002). Fig. 1. Localización de las dos estaciones muestreadas con un trineo suprabentónico en los márgenes de la pradera de Posidonia oceanica de la costa de Mataró (NE de la península ibérica) en 2001 (Mataró II) y 2011 (Mataró III). Distribución y curvas batimétricas (en metros) de las praderas de Posidonia y Cymodocea modificado en base a Manzanera & Cardell (2002).

not dissected, one vial; (ICMM12120406), 1 mature female (3.9, TL), station Mataró III, dissected, one vial; (ICMM12120407), 3 mature females (3.8, 4.1 and 4.1 mm TL) station Mataró III, not dissected, one vial; (ICMM12120408), 7 immature females (3.0, 3.1, 3.2, 3.3, 3.3, 3.4 and 3.5 mm TL), station Mataró III, not dissected, one vial; (ICMM12120409), 12 juveniles (1.1, 1.7, 1.7, 1.8, 1.9, 2.0, 2.1, 2.2, 2.2, 2.2, 2.4 and 2.4 mm TL), station Mataró III, not dissected, one vial. Etymology This species is dedicated to the archaeological remains of the ancient Roman city of Iluro preserved around the present village of Mataró. Diagnosis Small–size mysid characterized by a prominent triangular rostrum; carapace without dorsal nodules; eyestalk with a finger–like papilla on dorsal anterior margin; third article of antennular peduncle armed with one and three small cuspidate setae on the dorsal medium

and posterior margins, respectively; linguiform telson with a rounded apex armed with two median cuspidate setae; endopod of uropod armed on its inner medium margin with 6–8 cuspidate setae. Description General form robust and compact. Carapace with anterior margin produced in the middle line into a large triangular rostrum which extends to the middle of the first segment of antennular peduncle; without dorsal nodules; posterior margin emarginated dorsally, leaving the last two, and part of the sixth thoracic somites uncovered; posterolateral lobe not covering anterior abdominal somite (fig. 2A). Abdomen moderately robust, as wide as the middle portion of the carapace and not tapering posteriorly; first five somites subequal in length; last somite two times as long as the fifth. Eyes (fig. 2B) large, globular, broader than the eyestalk, laterally extending beyond carapace limits; eyestalk with dorsal finger–like papilla; ommatidial pigment orange (in preserved specimens).

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Antennular peduncle (fig. 2C) shorter than antennal scale. First article longer than wide; second article short, half as long as broad, third article broader than long, armed with one and three small cuspidate setae on the dorsal medium and posterior margins, respectively; male lobe large and hirsute. Antennal sympod (fig. 2D) with a pointed spine on the ventral outer posterior angle. Peduncle slightly shorter than the scale length; first article short as long as broad; second article three times as long as broad, inner posterior margin armed with two simple setae; third article half the length of the second article, posterior inner margin armed with three simple setae. Antennal scale lanceolate, four times as long as maximum width, extending beyond the antennular peduncle; outer margin straight, inner margin convex; a small suture present in posterior one–eleventh. Labrum (fig. 2E) rounded, asymmetrical, without frontal spiniform process, posterior margin with two clusters of short, irregularly distributed thin simple setae. Mandibles (fig. 2F–H) with a three–segmented palp, second article about twice as long as the third, with simple setae on both margins; third article armed on the posterior third of the inner margin with 10 ventral simple setae, five distally ventral serrate setae and one posterior large conspicuous pappose seta. Setal row and molar process reduced. Maxillule (fig. 2I) apex of outer lobe armed with eight strong cuspidate setae; inner lobe with two apical and one median simple seta. Maxilla (fig. 3A) with posterior article of endopod oval, longer than wide, inner posterior two–thirds margins armed with 18 pappose setae; exopod relatively narrow, extending to the third length of the posterior margin endopod article, with 10 pappose setae, the eight short simple setae on its anterior inner margin; inner margin of coxal and bilobulate basal endites armed with simple setae (fig. 2A). First thoracopod (fig. 3B–C) short and robust, with unarmed epipodite; inconspicuous coxa and articulation between basis and preischium; endopod with preischium and ischium fused, carpopropodus subequal in length to fused preischium and ischium; dactylus with one strong posterior spine and nine simple setae. Second thoracopod (fig. 3D) longer than the first; endopod with preischium and ischium not fused, merus shorter than carpopropodus; dactylus armed with one strong terminal curved setae; exopod subequal in length to the endopod, 10–segmented. Third to eighth thoracopods (fig. 3E–I) much longer than first and second, with endopod longer than exopod; isquium, merus and carpoprodus subequal in length; carpopropodus three–segmented; dactylus with a posterior simple curved seta; exopods 8–9 segmented. Sixth to eighth female thoracopods with a pair of developed oostegites, first pair smaller than posterior pairs. Genital appendage of male short and cylindrical, armed distally with six simple setae (fig. 3J). Pleopods of the female (fig. 4A–E) uniramous, unjointed, the first shorter than posterior pairs. Pleopods of the male (fig. 4F–J) well developed; first pleopod with plate–like endopod and 9–segmented exopod, second to fifth pleopods biramous with both rami

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subequal in length, 9–10 segmented; endopods with side pseudobranchial lobe rectangular, exopod of the fourth pair without modified setae. Telson (fig. 4K–L) linguiform, one and half times as long as broad at base; lateral margins armed with 8–10 short cuspidate setae similar in size; apex rounded–truncate, armed with two median cuspidate setae. Uropod (fig. 4K, M) short and robust; endopod extending beyond apex of telson for 1/5 of its length, armed on the medium inner ventral margin with 6–8 short cuspidate setae (seven in adult males, 6–8 in adult females and immature individuals); exopod longer than endopod, having entire margin setose, outer margin straight, inner margin convex. Colour (in preserved specimens): almost opaque white tegument with brown pigmentation irregularly distributed on the abdomen and some appendages; two black spots on the dorsal anterior part of the telson (fig. 5). Distribution and habitat The known distributional area of the new Mysidopsis species is at present restricted to the sandy edges of the Posidonia oceanica meadow from the coast of Mataró (north–western Mediterranean). Remarks M. iluroensis n. sp. represents the fifth species of the genus currently known from NE–Atlantic and Mediterranean. Previously known species of the genus are M. didelphys (Norman, 1863), M. angusta Sars, 1864 and M. gibbosa Sars, 1864, reported from northern Europe and Iceland to the Mediterranean, between coastal waters (known depth range: 1–220 m) and the recently described M. cachuchoensis San Vicente et al., 2012, reported from the 'Le Danois' Bank (Bay of Biscay) at 828 m depth. Although the genus Mysidopsis probably represents a polyphyletic melange of taxa (Price et al., 1994), the new species described in this paper undoubtedly shares the morphological diagnosis of the genus sensu Sars (1872), Tattersall & Tattersall (1951) and Tattersall (1969). The most distinctive characters whereby members of this genus may be recognized is the 4–articulate endopod of first thoracic appendages which was due to the fusion of the preischium and ischium, the fused carpopropodus divided into three subsegments in thoracopods 3–8, only setae along the inner margin of the uropodal endopod, not modified mandibular palps and the presence of an expodod in the maxilla. The pair of dark spots at the base of the telson is also a characteristic feature of the genus (Tattersall, 1969). All these characters are present in the new described species. In addition, the overall appearance of the individuals examined is very close to the type species of the genus: Mysidopsis didelphys (Norman, 1863). The new species can be distinguished from M. didelphys and M. angusta by the telson shape and the uropod armature. The rounded apex of the telson

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C D B

F G

1 mm 0.5 mm 0.1 mm

A B, C, D, F, G E, H, I

Fig. 2. Mysidopsis iluroensis n. sp., holotype, mature male (A–E, I) and paratype, mature male (F–H): A. Habitus in dorsal view; B. Right eye in dorsal view; C. Antennule in dorsal view; D. Antenna in ventral view; E. Labrum; F, G. Mandibles; H. Terminal segment of mandibular palp; I. Maxillule. Fig. 2. Mysidopsis iluroensis sp. n., holotipo macho adulto (A–E, I) y paratipo macho adulto (F–H): A. Vista dorsal; B. Vista dorsal del ojo derecho; C. Vista dorsal de la anténula; D. Vista ventral de la antena; E. Labro; F, G. Mandíbulas; H. Segmento terminal del palpo mandibular; I. Maxílula.

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H F I

1 mm 0.5 mm

A B, D

0.1 mm

C, J

0.5 mm

E–I

Fig. 3. Mysidopsis iluroensis n. sp., holotype, mature male (A–D, G–J) and other material, mature female, TL 4.1 mm (E–F): A. Maxilla; B. First thoracopod; C. Posterior articles of endopod of first maxilliped; D. Second thoracopod; E. Third thoracopod; F. Fifth thoracopod; G. Sixth thoracopod; H. Endopod of 7th thoracopod; I. Endopod of 8th thoracopod; J. Male genital organ. Fig. 3. Mysidopsis iluroensis sp. n., holotipo macho adulto (A–D, G–J) y otro material, hembra adulta, longitud total 4,1 mm (E–F): A. Maxila; B. Primer toracópodo; C. Artejos posteriores del endópodo del primer maxilípedo; D. Segundo toracópodo; E. Tercer toracópodo; F. Quinto toracópodo; G. Sexto toracópodo; H. Endópodo del 7º toracópodo; I. Endópodo del 8º toracópodo; J. Órgano genital masculino.

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F G

I J

K M

1 mm 0.5 mm 0.1 mm 0.5 mm

A–E F–J L K, M

Fig. 4. Mysidopsis iluroensis n. sp., other material, mature female, TL 4.1 mm (A–E) and holotype, mature male (F–M): A–E. First to 5th female pleopods; F–J. First to 5th male pleopods; K. Telson and uropods in dorsal view; L. Posterior end of telson; M. Uropod in ventral view. Fig. 4. Mysidopsis iluroensis sp. n., otro material, hembra adulta longitud total 4,1 mm (A–E) y holotipo macho adulto (F–M): A–E. Del 1º al 5º pleópodos femeninos; F–J. Del 1º al 5º pleópodos masculinos; K. Vista dorsal del telson y los urópodos; L. Extremo posterior del telson; M. Vista ventral del urópodo.

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1 mm

Fig. 5. Mysidopsis iluroensis n. sp., photograph of ethanol preserved specimens: A. Paratype, mature female in lateral view; B. Holotype, mature male in lateral view. Fig. 5. Mysidopsis iluroensis sp. n., fotografía de los ejemplares conservados en etanol: A. Paratipo, vista lateral de la hembra adulta; B. Holotipo, vista lateral del macho adulto.

of M. iluroensis n. sp. is armed with two median cuspidate setae whereas in M. didelphys the apex of the telson is truncate and armed at the outer corners with two cuspidate setae, and in M. angusta the apex of the telson has a median cleft. M. iluroensis sp. nov. differs from both latter species in the inner endopod armature of the uropod (with 6–8 cuspidate setae in M. iluroensis n. sp. versus only one in M. didelphys and M. angusta). M. iluroensis n. sp. can be easily distinguished from M. cachuchoensis by the ornamentation of its eyestalk (with dorsal finger–like papilla in the new species, absent in M. cachuchoensis), the structure of its antennal scale (without posterior article in M. cachuchoensis), the armature of the inner margin of its uropodal endopods (more than 30 graduated cuspidate setae from statocyst to sub–apex in M. cachuchoensis) and the armature of the lateral margins of its telson (anterior half naked in M. cachuchoensis). M. iluroensis sp. nov. shows the closest morphological similarity to M. gibbosa. The new species can be distinguished from M. gibbosa by its largest rostrum and the armature of the antennules and the uropod endopod. The dorsal surface of the carapace (without dorsal nodules in the new species versus two quite conspicuous and large nodules in M. gibbosa) is ano-

ther morphological difference. However, as mentioned by Bacescu (1941) and Ariani (1967), these nodules are sometimes poorly visible or even absent in some Mediterranean specimens. The armature of the dorsal surface of the third article of the antennular peduncle, with one and three small setae on the dorsal medium and posterior margin, respectively in M. iluroensis n. sp. distinguished also the new species from M. gibbosa (without any setae on the dorsal surface of the antennule). M. iluroensis sp. nov. differs from M. gibbosa in the inner endopod armature of the uropod (with 6–8 cuspidate setae in M. iluroensis versus only three–five small setae in M. gibbosa). After Tattersall (1909), Mauchline (1971) and Lagardère & Nouvel (1980) two forms of M. gibbosa may be found; one living in shallow bays, usually near the outflow of a freshwater stream, the other living at depths of 40–70 m. The general body shape of the shallow water form is as described by Tattersall & Tattersall (1951) for this species. The deep water form usually resembles young M. didelphys, the abdomen not being sigmoid to any great degree in lateral view but examination of the telson immediately distinguishes it from young M. didelphys. Neither of these two forms is morphologically close to M. iluroensis sp. nov.

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Key to species of Mysidopsis G. O. Sars, 1864 recorded in European waters. Clave de las especies de Mysidopsis G. O. Sars, 1864 registrados en aguas europeas.

1 Apex of telson with small median V–shaped cleft M. Apex of telson entire 2 2 Eyes with a finger–like papilla on the inner dorsal side of the eyestalk. Antennal scale with posterior article. Inner margin of uropodal endopod with no more than 8 cuspidate setae near statocyst. Anterior lateral margin of telson with cuspidate setae 3 Eyestalk without dorsal finger–like papilla. Antenna scale without posterior article. Uropodal endopod with a row of more than 30 cuspidate setae along inner margin. Anterior lateral margin of telson unarmed M. 3 Telson linguiform, apex with outer corners rounded and naked, only armed with a pair of median cuspidate setae. Inner margin of the uropod endopod armed with 4–8 short cuspidate setae 4 Telson triangular with narrow truncate apex, armed with a pair of large cuspidate setae at the outer corners. Inner margin of the uropod endopod with a single long cuspidate seta near the statocyst M. 4 Rostrum small, obtusely triangular. Inner margin of the uropod endopod armed with 3–5 cuspidate setae M. Rostrum well produced, broadly triangular and acutely pointed. Inner margin of the uropod endopod armed with 6–8 cuspidate setae M.

None of the morphological peculiarities described for the new species have previously been described in M. gibbosa even in the Mediterranean specimens (Sars, 1877; Carus, 1885; Lo Bianco, 1903; Tattersall, 1909; Colosi, 1929; Bacescu, 1941; Ariani, 1967; Barberá Cebrián et al., 2001). Moreover, re–examination of specimens of M. gibbosa sampled on sandy bottoms near the coast of Mataró (Masnou beach between 5–10 m depth) by San Vicente & Munilla (2000) confirms the differences between the two species, especially in the armature of the antennule (without any setae on the dorsal surface) and the uropod endopod (with four setae on all M. gibbosa specimens analyzed). Adapted from Tattersall & Tattersall (1951), Tattersall (1969) and San Vicente et al. (2012), an identification key is proposed to include the new Mediterranean species herein described. Acknowledgements Thanks to Gregori Muñoz and Jordi Corbera for providing the opportunity to study the mysid material sampled in the Mataró coast.

angusta G. O. Sars, 1864

cachuchoensis San Vicente et al., 2012

didelphys (Norman, 1863) gibbosa G. O. Sars, 1864 iluroensis n. sp.

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bosa, M. didelphys and M. angusta (Crustacea, Mysidacea). Journal of the Marine Biological Association of the United Kingdom, 50: 381–396. – 1971. The fauna of the Clyde Sea area. Crustacea: Mysidacea. Oban: Scottish Marine Biological Association: 1–26. – 1980. The biology of mysids and euphausiids. Advances in Marine Biology, 18: 1–680. Mauchline, J. & Murano, M., 1977. World list of the Mysidacea, Crustacea. Journal of the Tokyo University of Fisheries, 64: 39– 88. Mazzella, L., Scipione, M. B. & Buia, M. C., 1989. Spatio–temporal distribution of algal and animal communities in a Posidonia oceanica meadow. Marine Ecology, 10: 107–129 Mees, J. & Meland, K., Eds., 2012. World List of Lophogastrida, Stygiomysida and Mysida. Available online at http://www.marinespecies.org/mysidacea (Accessed on November 16, 2012). Müller, H. G., 1993. World Catalogue and Bibliography of the Recent Mysidacea. Wissenschaftler Verlag, Tropical Products Trading Center, Germany: 1–421. Price, W. W., Heard, R. W. & Stuck, L., 1994. Observations on the genus Mysidopsis Sars, 1864 with the designation of a new genus, Americamysis, and the descriptions of Americamysis alleni and A. stucki (Peracarida: Mysidacea: Mysidae), from the Gulf of Mexico. Proceedings of the Biological Society of Washington, 107(4): 680–698. San Vicente, C., 2004. Curso Práctico de Entomología. Capítulo 22. Misidáceos: 363–378 (J. A. Barrientos, Ed.). Asociación Española de Entomología, CIBIO (Centro Iberoamericano de la Biodiversidad). Universitat Autònoma de Barcelona, Servei de Publicacions. San Vicente, C., Frutos, I. & Sorbe, J. C., 2012. Mysidopsis cachuchoensis sp. nov. (Crustacea: Mysida: Mysidae), a new suprabenthic mysid from bathyal soft bottoms of the Le Danois Bank (southern Bay of Biscay). Journal of the Marine Biological Association of the United Kingdom: 1–12. (Available on CJO. DOI:10.1017/S0025315412000987). San Vicente, C. & Munilla, T., 2000. Misidáceos suprabentónicos de las playas catalanas (Mediterráneo nordoccidental). Orsis, 15: 45–55. Sánchez–Jerez, P., Barberá–Cebrián, C. & Ramos– Esplá, A. A., 1999. Comparison of the epifauna spatial distribution in Posidonia oceanica, Cymodocea nodosa and unvegetated bottoms: importance of the meadow edges. Acta Oecologica, 20: 391–405 Sars, G. O., 1872. Carcinologiske Bidrag til Norges Fauna. I. Monographi over de ved Norges Kyster forekommende Mysider. Pt. 2: 1–34. Det Kongl. Norske Videnskabsselskab i Trondhjem. Brøgger & Christies Bogtrykkeri, Christiania. – 1877. Nye Bidrag til Kundskaben om Middelhavets Invertebratfauna I. Middelhavets Mysider. Archiv for Mathematik og Naturvidenskab Christiania, 2: 10–119. Scipione, M. B., Gambi, M. C., Lorenti, M., Russo, G. F. & Zupo, V., 1996. Vagile fauna of the leaf

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by the Maia and Puritan in the Mediterranean. Mittheilungen aus der Zoologischen Station zu Neapel, 19: 117–143. Tattersall, W. M. & Tattersall, O. S., 1951. The British Mysidacea. Ray Society, London. Watling, L., 1989. A classification system for crustacean setae based on the homology concept. In: Functional morphology of feeding and grooming in Crustacea: 15–26 (B. Felgenhauer, L. Watling, A. B. Thistle, Eds.). A. A. Balkema, Rotterdam. Wittmann, K. J., 2001. Centennial changes in the near–shore mysid fauna of the Gulf of Naples (Mediterranean Sea), with description of Heteromysis riedli sp. n. (Crustacea, Mysidacea). Marine Ecology, 22: 85–110.

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Effects of migratory status and habitat on the prevalence and intensity of infection by haemoparasites in passerines in eastern Spain J. Rivera, E. Barba, A. Mestre, J. Rueda, M. Sasa, P. Vera & J. S. Monrós

Rivera, J., Barba, E., Mestre, A., Rueda, J., Sasa, M., Vera, P. & Monrós, J. S., 2013. Effects of migratory status and habitat on the prevalence and intensity of infection by haemoparasites in passerines in eastern Spain. Animal Biodiversity and Conservation, 36.1: 113–121. Abstract Effects of migratory status and habitat on the prevalence and intensity of infection by haemoparasites in passerines in eastern Spain.— The Iberian peninsula is a suitable place to study the effects of migratory condition on the prevalence of blood parasites in avian communities as resident, local populations cohabit with migratory species and with abundant vector populations. In this study we examined the incidence of avian blood parasites in three localities in the Mediterranean region (east Spain), in relation to the migratory status of the species. We analyzed 333 blood smears from 11 avian species, and obtained an overall prevalence of 9.6%. The prevalence of parasites varied among the different species studied, although intensity of infection did not. Our results are discussed in terms of population dynamics and abundance of Diptera vectors able to transmit blood parasites to other birds. Key words: Blood parasites, Trypanosoma ssp., Haemoproteus spp., Passeriformes, Diptera vectors. Resumen Efectos del estatus migratorio y del tipo de hábitat sobre la prevalencia y la intensidad de la infección por hemoparásitos en paseriformes en el este de España.— La península ibérica es un sitio idóneo para estudiar los efectos de la condición migratoria en la prevalencia de hemoparásitos en comunidades de aves, dado que convergen poblaciones residentes locales con especies migratorias y abundantes poblaciones de vectores. En este trabajo examinamos la incidencia de hemoparásitos presentes en aves de tres localidades de la región mediterránea (este de España), con respecto del estatus migratorio. Examinamos 333 frotis sanguíneos de 11 especies, y encontramos una prevalencia global del 9,6%. A diferencia de la intensidad de la infección, la prevalencia de parásitos mostró variación entre las distintas especies estudiadas. Nuestros resultados se interpretan en relación con la dinámica de poblaciones y la abundancia de dípteros vectores capaces de transmitir los hemoparásitos a otras aves. Palabras clave: Hemoparásitos, Trypanosoma spp., Haemoproteus spp., Paseriformes, Dípteros vectores. Received: 25 IV 12; Conditional acceptance: 25 IX 12; Final acceptance: 22 IV 13 Jennifer Rivera, Mahmood Sasa, Inst. 'Clodomiro Picado', Univ. de Costa Rica, San José, Costa Rica.– Emilio Barba, Pablo Vera & Juan S. Monrós, Inst. 'Cavanilles' of Biodiversidad y Biología Evolutiva, Univ. de Valencia, AC 22085, E–46071 Valencia, España (Spain).– Alexandre Mestre & Juan Rueda, Dept. de Microbiología y Ecología, Univ. de Valencia, c/ Dr. Moliner 50, E–46100 Burjassot, España (Spain). Corresponding author: J. S. Monrós. E–mail: monros@uv.es

ISSN: 1578–665X

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Introduction Avian hematozoa parasites (Protista) are a heterogeneous group of organisms widely distributed worldwide (Peirce, 1981; Valkiūnas, 2005). Atkinson & Van Riper (1991) noted that haemoparasites have been recorded in almost 70% of the avian species examined, although prevalence estimates may depend on the method used in their detection (Fallon et al., 2005). Parasites from the genus Haemoproteus are among the most common avian haematozoa (two thirds of the described blood parasite morphospecies, Valkiūnas, 2005). Parasites of the genus Leucocytozoon, Plasmodium (Bennett et al., 1993) and some Trypanosoma species (Kučera, 1982) are also common in avian species. These haemoparasites exert selective pressure on their hosts (Hamilton & Zuck, 1982), negatively affecting the efficiency of metabolism (Chen et al., 2001), survival, breeding success, and physical aptitude (Marzal et al., 2008; Stjernman et al., 2008; Ruiz de Castañeda et al., 2009; Martínez de la Puente et al., 2010), and body growth (Soler et al., 2003). The incidence of haemoparasites in avian communities varies geographically (Sol et al., 2000). This variation has been linked to habitat characteristics, species composition in the community, vector–host specificity and ecological requirements of the vectors (Deviche et al., 2005). Prevalence and intensity of parasitic infections on birds may also depend on the migratory status of the host species. The probability of being infected would thus be higher in migratory species than in sedentary species, as they are exposed to more than one parasitic fauna during their life cycle (Figuerola & Green, 2000). Migration may also limit the transmission of parasites to new host species, due to the vector–host specificity (Hellgren et al., 2008). Habitat features affect the incidence of infections in birds (Martínez–Abraín et al., 2004) due to differences in vector abundance and behavior (Bennett et al., 1982). The incidence of parasitemia can thus be expected to be lower in semi–arid regions (Little & Earlé, 1995) than in humid regions with aquatic environments (Moyer et al., 2002). A latitudinal gradient related to climatic conditions and their effect on vectors could be involved in the prevalence of blood parasites in birds (Bensch & Åkesson, 2003). Several studies carried out in the north and center of Europe, where seasonal climatic changes are severe, found that the prevalence of haemoparasites was relatively high (i.e. Kučera, 1981; Valkiūnas et al., 2003; Shurulinkov & Golemansky, 2003), but other studies found a higher prevalence in the south (Marzal el al., 2011). Thus, in southern Europe no clear pattern in the prevalence of blood parasites has been observed (i.e. Merino et al., 1997; Valera et al., 2003). Depending on latitude, the Iberian Peninsula shows peculiar climatic characteristics that make it suitable to host high numbers of migratory birds during the winter (Tellería, 1988). Mild peninsular winters thus provide a great variety of resources both for short distance migrants and resident bird populations (Senar & Borras, 2004).

Rivera et al.

In Spain, a number of significant studies have been carried out to describe and understand the patterns of haemoparasite infections in birds (Merino et al., 1997; Tomás et al., 2007). Studies in the center of the country have shown that the higher the vector abundance, the higher the haemoparasitic prevalence (Merino & Potti, 1995). Preliminary surveys on the Mediterranean coast showed that haemoparasites are almost absent in the Passeriformes species (Parus major, Periparus ater, Lophophanes cristatus; E. Barba, non–published data). Similarly, blood parasites were absent in nigh–jars, Caprimulgus ruficollis (Forero et al., 1997) and in storks Ciconia ciconia (Jovani et al., 2002) in Doñana National Park, a patchy region with wetland and Mediterranean forests. Absence of blood parasites was also noted in Kentish plover and gulls breeding on the Mediterranean coast of Spain (Figuerola et al., 1996; Martínez–Abraín et al., 2002). However, a high prevalence of infection by blood parasites has been found in both migratory and resident species in the south of the Iberian peninsula (Marzal et al., 2008, 2011; López et al., 2011). Most parasitic infections are transmitted by Diptera (Ceratopogonidae, Culicidae and Simuliidae) and the abundance of these vectors depends on the local climate and water conditions in each season (Valkiūnas et al., 2003). The present study aimed to determine the haemoparasitic infection prevalence and intensity in Passeriformes in three localities in Eastern Spain, to analyze the differences in prevalence and intensity of infection between resident and wintering species, and to relate these measures with the presence of vectors that are potential transmitters of the parasitic infections. Material and methods Birds included in this study were trapped between September and December 2008 in three localities close to the Mediterranean coast in eastern Spain (fig. 1). The first locality was the Marjal Pego–Oliva Natural Park (Pego–Oliva; 38° 52' N, 0° 3' W), on the border between the provinces of Valencia and Alicante. The birds were trapped in a wetland with large reed bed areas with mixed patches of cattails and sedge, next to rice fields. The second study site was an orange grove (Citrus sinensis) in Sagunto, province of Valencia (Sagunto; 39º 42' N, 0º 15' W), 4 km from the coast. The third locality was L’Albufera Natural Park (Albufera; 39º 19' N, 0º 21' W), a wetland in the south of the Valencia city, dominated by rice fields and some patches of marshland natural vegetation. In the three study areas, birds were trapped using mist–nets, operating weekly as part of the constant effort ringing programs. In all three sites, 60 m of mist–nets were set at dawn and were operated for 4 hours, following Belda et al. (2007). Each bird was banded with an individual metal ring. Each species was catalogued as resident (species present throughout the year) or wintering (migratory species that winter but do not breed in the study area).

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Sagunto

Albufera

Pego–Oliva

Fig. 1. Location of the study localities in the Iberian peninsula. Fig. 1. Situación geográfica de las localidades en la península ibérica.

We extracted a drop of blood from the brachial vein of each trapped bird. The drop was placed on a glass slide and dried air. In the laboratory, samples were fixed with absolute methanol and dyed with Giemsa for 45 minutes, following the protocol of Merino et al. (1997). We randomly chose one half of the slide and quantified at 400x the presence of extracellular parasites (Trypanosoma spp.) or intracellular parasites (Leucocytozoon spp.) along the longitudinal axis. The number of haemotozoa observed in 100 optic fields was recorded. Infection intensity by intracellular parasites (Haemoproteus spp. o Plasmodium spp.) was obtained as the number of parasites per 2,000 erythrocytes, following Merino & Potti (1995). All the slides were revised by J. R. Parasite identification was based in the morphological characteristics after Valkiūnas (2005). Complementarily to the bird sampling, water samples were collected in the three study areas to determine the composition of the potential vector community (related to haemoparasite transmission) in different water bodies (such as irrigation ponds, natural springs, and channels). We sampled 52 water bodies, 39 in natural habitats and 13 in artificial ponds. The sampling was performed using a hand net with a square frame of 25 cm per side and a net with pore diameter of 250 µm Each sub–sample was concentrated on a 30 x 40 cm plastic plate. Sampling was concluded when no new taxa were found in the sub–sample. The whole sample (as the set of sub–samples) was stored in a plastic 1 l bottle in 70º alcohol. In the laboratory, samples were washed with water in a 250 µm pore diameter sieve to remove the silt. Species were

identified following Tachet et al. (2000) and Rueda & López (2003) using a Motic Digital Microscope DM 143 stereoscopic microscope and a Bresser TrinoLab 40–1,600x microscope. The prevalence and infection intensity were analyzed at two levels: migratory status and habitat (locality). We analyzed the differences between groups using mixed generalized linear models (GLMMs) fitted by Laplace approximation. Two analyses were made for both Trypanosoma spp. and Haemoproteus spp. using locality as a fixed factor in the first analysis and migratory status in the second, and individual and species as random factor in both analyses. The individual–level represents a per–observation error term, which captures over–dispersion (Elston et al., 2001; Atkins et al., 2013). We were unable to analyse the two effects together because of zero inflation in the results; and one for Trypanosoma sp. only with data obtained in Chiffchaffs (Phylloscopus collybita) as it was the only species that was trapped in the three localities (using locality as fixed factor) (Brew & Maddy, 1995). We used binomial models with a logit link function and Poisson models with a logarithmic link function. All tests were performed using the lme4 package v.0.999375–42 (Bates et al., 2012) for R version 2.14.1 (R Development Core Team, 2009). Results A total of 333 birds were trapped, belonging to 11 species and five families. In all three localities, both migratory and sedentary birds were trapped and

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Table 1. Infection status for the individuals of the 11 species included in this study: ITry. Infected by Trypanosoma spp.; IHae. Infected by Haemoproteus spp. Migratory status and locality of capture are given for each species. Tabla 1. Situación de los individuos de las 11 especies incluidas en el presente estudio respecto de la infección: ITry. Infectado por Trypanosoma spp.; IHae. Infectado por Haemoproteus spp. Para cada especie se indican el estatus migratorio y la localidad de captura.

Number of birds Infected by

Sampled

ITry

Migratory

IHae

status

Locality

Sylviidae Acrocephalus melanopogon

3(10)

0(0)

Resident

Pego–Oliva

Cettia cetti

0(0)

Resident

Pego–Oliva

5(5.5)

0(0)

Wintering

Pego–Oliva,

Phylloscopus collybita

Sagunto, Albufera

Sylvia atricapilla

1(3.3)

21(70)

Wintering

Sagunto

Sylvia melanocephala

1(4.8)

0(0)

Resident

Sagunto

0(0)

Wintering

Albufera

Emberizidae Emberiza schoeniclus

Passeridae Passer domesticus

0(0)

Resident

Albufera

Passer montanus

0(0)

Resident

Albufera

Turdidae Turdus merula

0(0)

Resident

Sagunto

Erithacus rubecula

0(0)

2(6.7)

Wintering

Sagunto

Wintering

Albufera

Fringillidae Fringilla coelebs

0(0)

Total

333

10(3)

23(6.9)

sampled (table 1). Only 32 birds were infected, with a global prevalence of 9.6% (table 1). The most common parasite was Haemoproteus spp., which was identified in 23 birds (6.9%). Trypanosoma spp. was detected in 10 birds (3.0%). Only one individual of Blackcap Sylvia atricapilla showed both parasites (0.3%). No individual showed infection by Plasmodium spp. The prevalence of infected birds differed between species (x2 = 216.5, p < 0.001, df = 10, table 1). Haemoparasites were not detected in six species (Cettia’s Warbler Cettia cetti, Reed Bunting Emberiza schoeniclus, Chaffinch Fringilla coelebs, House Sparrow Passer domesticus, Tree Sparrow Passer montanus and Blackbird Turdus merula; table 1). Prevalence did not correlate with the number of samples collected for each species (Spearman’s rho; Trypanosoma spp.: r = 0.549, p = 0.080; Haemoproteus spp.: r = –0.299, p = 0.371). The species that showed the highest abundance of parasites and the highest proportion of infected individuals (n = 22) was the Blackcap,

a species that winters in this area, mainly in shrubs and croplands rather than wetlands. Trypanosoma spp. infections were detected in Sardinian Warbler Sylvia melanocephala (n = 1), Chiffchaff (n = 5), Blackcap (n = 1) and Moustached Warbler Acrocephalus melanopogon (n = 3), although in an overall analysis no significant differences were found in the infection prevalence between species (x2 = 12.20; p = 0.27; df = 10). Taking Pego–Oliva as reference locality level to calculate the estimators of locality effects, we did not find any statistical differences between localities (Wald x2 < 0.001; p > 0.999; df = 2), or between migratory status (Wald x2 < 0.001; p = 0.988; df = 1) (see the lower coefficient values of each level compared with the higher SE values showed in table 2). In a partial analysis with data collected on Chiffchaffs (the only species found in the three sampling localities), we did not find differences in the prevalence of Trypanosoma spp. between localities (Wald x2 < 0.001; p > 0.999; df = 2).

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Table 2. Results of the GLMMs used to analyze the effects of locality and migratory status on the prevalence of Trypanosoma spp. and Haemoproteus spp. Tabla 2. Resultados de los modelos lineales generalizados mixtos utilizados para analizar los efectos de la localidad y el estatus migratorio en la prevalencia de Trypanosoma spp. y Haemoproteus spp.

Effect

Estimate

Trypanosoma spp.

Intercept

–14.28

28.75

Locality

–0.497

0.619

Sagunto

–0.14

37.05

0.004

0.997

Albufera

–17.53

6.3·106

< 0.001

> 0.999

Pego–Oliva

0.00

–

Intercept

–15.14

24.97

–0.606

0.544

Resident

0.60

39.30

0.015

0.988

Wintering

0.00

–

Haemoproteus spp. Locality

Intercept

–21.80

6.7·10

–0.003

0.997

Sagunto

17.79

6.7·10

0.003

0.998

Albufera

–0.000001

8.7·10

< 0.001

> 0.999

Pego–Oliva

0.00

–

Intercept

–3.99

1.60

–2.497

0.012

Resident

–17.40

4.2·10

–0.004

0.997

Wintering

0.00

–

Migratory status

Haemoproteus spp. infections were found only in Blackcaps (n = 21) and European Robin Erithacus rubecula (n = 2), and the prevalence differed between all the species (x2 = 205.99; p < 0.001; df = 10). A posteriori test showed that the differences were due to Blackcaps (x2 = 18.30; p = 0.05; df = 10). Again taking Pego–Oliva as reference locality level to calculate the estimators of locality effects, we did not find statistical differences between localities (Wald x2 < 0.001; p > 0.999; df = 2) or between migratory status (Wald x2 < 0.001; p = 0.997; df = 1) (table 2). Taking Pego–Oliva as reference locality level to calculate the estimators of locality effects, we did not find statistical differences in the intensity of Trypanosoma parasitism between localities (Wald x2 < 0.001; p > 0.999; df = 2), or between migratory status (Wald x2 = 0.006; p = 0.936; df = 1) (see coefficient values of each level compared with the higher SE values in table 3). Neither did we find differences in the infection intensity between localities when we considered only the data collected on Chiffchaffs (Wald x2 = 0.169; p = 0.919; df = 2). In the case of Haemoproteus infection, intensity results were similar, showing no differences between localities (Wald x2 < 0.001; p > 0.999; df = 2), or migratory status (Wald x2 < 0.001; p > 0.999; df = 1) ( table 3). Table 4 shows the composition of the community of Diptera species potentially acting as a vector for haemoparasite transmission in the three sampling areas. The Pego–Oliva area had the highest richness

3 3 3

(13 species and eight genera). In Sagunto, only two species (from two different genera) were detected. Unfortunately, our sampling strategy did not allow comparison of species abundance between species or localities. Discussion The intensity and prevalence of infection caused by Trypanosoma spp. did not differ between species, migratory status, or locality, as shown previously in studies carried out in the center of Spain (Merino et al., 1997) and north of Europe (Hauptmanova et al., 2006). These results may be attributed to several factors: i) low frequency of individuals infected with this parasite (i.e. only five individuals of 91 Chiffchaffs sampled showed infection by Trypanosoma spp.); ii) problems with the methodology used for the detection could increase the number of false negatives; Apanius (1991) showed that Trypanosomas are not commonly found in peripheral blood, but are abundant in the bone marrow of the infected bird; and iii) as sampling was done in autumn, birds may have had low intensity of infection as they successfully passed the peak period of parasitic infection and thus present residual infection rates (Pérez–Tris & Bensch, 2005; Arizaga et al., 2009). Haemoproteus spp. was the most prevalent infection, with values similar to those reported in other

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Table 3. Results of the GLMMs used to analyze the effects of locality and migratory status on the infection intensity of Trypanosoma spp. and Haemoproteus spp.: PML. Parasite mean load. Tabla 3. Resultados de los modelos lineales generalizados mixtos utilizados para analizar los efectos de la localidad y el estatus migratorio en la intensidad de la infección causada por Trypanosoma spp. y Haemoproteus spp.: PML. Carga parasitaria media.

Effect

Estimate

PML

Trypanosoma spp.

Intercept

–8.18

3.22

–2.54

0.011

–

Sagunto

–0.006

4.16

–0.001

0.999

0.04

Albufera

–18.03

Pego–Oliva

Locality

Migratory status Intercept

4.5·10 < 0.001 > 0.999 4

0.00

–

0.05

–9.06

2.93

–3.09

0.002

–

Resident

0.37

4.62

0.080

0.936

0.04

Wintering

0.00

–

0.02

Haemoproteus spp. Locality

Intercept

–22.76

1.0·10 –0.002

0.998

–

Sagunto

15.89

1.0·104

0.999

0.90

Albufera

Pego–Oliva

Migratory status Intercept

0.002

–0.0003 1.3·104 < 0.001 > 0.999 0.00 –8.36

Resident

–17.81

Wintering

0.00

studies, with prevalence around 40% during the autumn (Merino et al., 2000; Pérez–Tris & Bensch, 2005; Arizaga et al., 2009). Haemoproteus infection is transmitted to birds by Culicoides (Diptera: Ceratopogonidae) (Garvin et al., 2006). The life cycle of this parasite develops rapidly, with asexual reproduction stages in the host (Merino et al., 2004), increasing the probability of infection transmission. Our data suggest that prevalence differed between species, although intensity of infection did not differ between species. Differences could also be attributed to the presence of a species with high infection values (Blackcap), although the importance of other variables such as age, sex, immune state, and season could not be tested due to the low sample size. Migratory status had a significant effect on the prevalence of Haemoproteus. Several studies show that Haemoproteus infections in migratory birds are common due to the wide distribution range of the parasite (Waldenström et al., 2002; Pérez–Tris & Bensch 2005). Waldenström et al. (2002) also found evidence of blood parasites as a cost of migration in birds, which may have a considerable impact on the evolution of migration. The highest prevalence of haemoparasites was recorded in Sagunto, although it was the locality with the lowest richness of vectors. This can be explained by the fact that the vector community is not rich but shows high abundance for some species. It is of

–

3.91

–2.14

0.032

4.6·104 < 0.001 > 0.999 –

–

0.00 0.00 – 0.00 0.52

note that some authors found that the incidence of haemoparasites is correlated with local abundance of vectors (Merilä et al., 1995; Sol et al., 2000). Therefore, if dipteral vectors have a wide distribution and small habitat restrictions, their local distribution and abundance could increase the presence of haemoparasites in different bird populations. According to this hypothesis, we would also expect a high parasitemia in the resident species. However, our results do not show this parasitemia, so we think that in this case, abundance of vectors does not explain the prevalence of haemoparasites. We think that these results are due to an effect of the host community and the migratory status of the hosts. Our results show that during the winter, Sagunto hosts several species with high blood parasite prevalence, particularly Blackcaps, a wintering species that was only trapped and sampled in Sagunto. In addition, some studies show that the prevalence of haemoparasites is related to macrohabitat characteristics. For example, Tella et al. (1999) noted that species of birds of prey nesting in forests showed a high prevalence of blood parasites. We did not find a high level of parasitemia in the other two localities, possibly due to the different habitat characteristics, as both were wetlands. Our results highlight the importance of considering migratory status as a possible factor influencing the prevalence of haemoparasites in bird communities.

Animal Biodiversity and Conservation 36.1 (2013)

Table 4. Composition of the community of Diptera vectors in the three study areas: S. Sagunto; P–O. Pego–Oliva; A. Albufera. Tabla 4. Composición de la comunidad de dípteros vectores en las tres zonas del estudio: S. Sagunto; P–O. Pego–Oliva; A. Albufera.

P–O

Anopheles spp.

Culex modestus

Culex theileri

Culex pipiens

Culicoides spp.

Culiseta subochrea

Dasyhelea spp.

Forcipomyia spp.

Ochlerotatus caspius

Ochlerotatus detritus

Simulium reptans

Simulium ruficorne

Simulium velutinum

Total species

Culiseta longiareolata

Acknowledgements We sincerely thank Santiago Merino and Josué Martínez de la Puente for their helpful collaboration in identifying the haemoparasites, and Rubén Piculo and José Luis Greño for their assistance in the fieldwork. We are also grateful to Aarón Gómez and Fabián Bonilla for providing the sampling equipment and for their comments and revision of the manuscript. Thanks too to Alfonso Marzal, Jordi Figuerola and an anonymous reviewer for valuable comments on this manuscript. J. R. received a grant from the Fundació General de la Universitat de València (Jóvenes Investigadores de Países en Vías de Desarrollo) for this study. References Apanius, V., 1991. Avian trypanosomes as models of hemoflagellate evolution. Parasitology Today, 7: 87–90. Arizaga, J., Barba, E. & Hernández, M. A., 2009. Do Haemosporidians affect fuel deposition rate and fuel load in migratory Blackcaps? Ardeola, 56: 41–47. Atkins, D. C., Baldwin, S. A., Zheng, C., Gallop, R. J.

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Three new species of Bryconamericus (Characiformes, Characidae), with keys for species from Ecuador and a discussion on the validity of the genus Knodus C. Román–Valencia, R. I. Ruiz–C., D. C. Taphorn B. & C. García–A. Román–Valencia, C., Ruiz–C., R. I., Taphorn B., D. C. & García–A., C., 2013. Three new species of Bryconamericus (Characiformes, Characidae), with keys for species from Ecuador and a discussion on the validity of the genus Knodus. Animal Biodiversity and Conservation, 36.1: 123–139. Abstract Three new species of Bryconamericus (Characiformes, Characidae), with keys for species from Ecuador and a discussion on the validity of the genus Knodus.— Three new species of characid fishes of the genus Bryconamericus are described from the Pacific coast and Amazon Basin in Ecuador, based on pigmentation and morphometric, meristic and osteological characters. B. bucayensis (n = 48) is distinguished by the number of scales between the lateral line and the pelvic–fin insertions (7–8 vs. 2–7, except B. terrabensis with 7–8 and B. arilepis with 9–10), the number of branched anal–fin rays (33–38 vs. 31 or fewer), the number of anterior anal–fin rays covered by a row of scales at their bases (28–31 vs. 4–26), and its wide anterior maxillary tooth being at least twice the width of the posterior tooth, both of which are pentacuspid (vs. maxillary teeth of same size). B. zamorensis (n = 126) is distinguished from congeners by having five teeth on the maxilla (vs. 1 or 2 teeth on maxilla), except B. rubropictus and B. thomasi, from which it differs in a reticulated pattern over the lateral stripe, generated by the concentration of melanophores, the scale margins, all along the sides of the body, the high number of branched anal–fin rays and vertebras, and the low branched dorsal–fin rays. The dorsal expansion of the rhinosphenoid forms a bony wall between olfactory nerves (vs. dorsal expansion of rhinosphenoid between olfactory nerves absent). Lateral process of palatine surpasses anterolateral margin of ectopterygoid (vs. palatine without lateral processes that laterally surpass the ectopterygoid), and the distal tip of sphenotic spine is laterally wide and undulated (vs. narrow). B. oroensis n. sp. (n = 124) is distinguished by having a dark lateral stripe overlaid with a peduncular spot and a reticulated pattern on the sides of the body (vs. peduncular spot and other body pigments not superimposed over a dark lateral stripe). It has three simple dorsal–fin rays, the first only visible in cleared and stained material and articulated, along with the second simple ray, with the first dorsal pterygiophores. The third simple ray is longer, and articulated with second dorsal pterygiophores (vs. only two simple dorsal–fin rays, both articulated with first dorsal pterygiophores). The anterior frontal is separated and so the fontanel front parietal is continued on the mesethmoids (vs. anterior tips of frontals united, and not separated by mesethmoids). Keys for identification of the species of Bryconamericus known to occur in Ecuador are included and the validity of the genus Knodus (vide Knodus carlosi) is discussed for cis Andean species. Key words: Bryconamericus n. sp., Taxonomy, Characid Fish, South America, Freshwater. Resumen Tres nuevas especies de Bryconamericus (Characiforme, Characidae), con claves de identificación para las especies de Ecuador y discusión sobre la validez del género Knodus.— Se describen tres especies nuevas de peces carácidos pertenecientes al género Bryconamericus en la costa del Pacifico y la cuenca del Amazonas en Ecuador que se basan en la pigmentación y caracteres morfométricos, merísticos y osteológicos. B. bucayensis (n = 48) se distingue por el número de escamas existentes entre la línea lateral y el origen de las aletas pélvicas (7–8 en lugar de 2–7, excepto B. terrabensis con 7–8 y B. arilepis con 9–10); por el número de radios ramificados de la aleta anal (33–38 en lugar de 31 o menos); por el número de radios anteriores de la aleta anal cuya base está cubierta por una hilera de escamas (28–31 en lugar de 4–26), por presentar un diente anterior del maxilar superior, al menos el doble de ancho que el diente posterior, ambos pentacúspides (en lugar de que los dientes del maxilar superior sean de igual tamaño). B. zamorensis (n = 126) se ISSN: 1578–665X

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distingue de sus congéneres por presentar cinco dientes en el maxilar superior (en lugar de 1 o 2), excepto de B. rubropictus y B. thomasi, de los que se distingue por el diseño reticulado de la parte lateral del cuerpo, generado por la concentración de melanóforos, por los bordes de las escamas, a los lados del cuerpo, por el mayor número de radios ramificados en la aleta anal y de vertebras, así como por el menor número de radios ramificados en la aleta dorsal. La extensión dorsal del rinosfenoide forma una pared entre los nervios olfatorios (en lugar de que la extensión dorsal del rinosfenoide entre los nervios olfatorios esté ausente). El proceso lateral del palatino sobrepasa el borde anterolateral del ectopterigoide (en lugar de que el palatino no tenga un proceso lateral y el extremo distal de la espina del esfenótico en vista lateral es ancha y ondulada (en lugar de ser estrecha). B. oroensis (n = 124) se distingue por presentar una banda lateral oscura que se solapa con la mancha del pedúnculo caudal y un diseño reticulado en los laterales del cuerpo (en lugar de que la mancha del pedúnculo caudal y otros pigmentos corporales no se solapen con una banda lateral). Tiene tres radios simples en la aleta dorsal el primero de los cuales sólo es visible cuando el material se ha limpiado y teñido, y se articula, junto con el segundo radio simple, con el primer pterigióforo dorsal. El tercer radio simple es más largo y se articula con el segundo pterigióforo dorsal (en lugar de tener solo dos radios simples en la aleta dorsal, ambos articulados con el primer pterigióforo dorsal). El frontal anterior está separado de manera que la fontanela frontoparietal se prolonga sobre el mesetmoides (en lugar de que los extremos anteriores del frontal estén unidos y no separados por el mesetmoides). Se incluyen las claves para identificar a las especies conocidas de Bryconamericus en Ecuador y se debate sobre la validez del género Knodus (véase Knodus carlosi) para las especies de la parte cisandina de Sudamérica. Palabras clave: Bryconamericus n. sp., Taxonomía, Pez carácido, Sudamérica, Agua dulce. Received: 27 VI 12; Conditional acceptance: 18 XII 12; Final acceptance: 20 V 13 C. Román–Valencia, R. I. Ruiz–C., D. C. Taphorn B. & C. García–A., Univ. del Quindío, Lab. de Ictiología, A. A. 2639, Armenia. Colombia.– D. C. Taphorn, 1822 N. Charles St., Belleville, Illinois. USA.– C. García–A., Univ. del Atlántico, Depto. de Biología, Barranquilla, Colombia. Corresponding autor: C. Román–Valencia: ceroman@uniquindio.edu.co

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Introduction

Comparative material

Systematically, Bryconamericus species from Central America have now been clearly resolved (Román–Valencia, 2002a; Román–Valencia & Vanegas–Rios, 2009) but South American species in most countries are still poorly understood, as is the case for both Pacific and Amazon drainages from Ecuador, Perú and Bolivia (Eigenmann, 1927; Böhlke, 1958; Vari & Siebert, 1990; Malabarba & Kindel, 1995; Silva, 2004; Langeani et al., 2005; Román–Valencia, 2005; Serra & Langeani, 2006; Román–Valencia et al., 2008a, 2008b, 2011). Available keys and species descriptions (Eigenmann, 1927; Böhlke, 1958; Gery, 1977) for these countries are of little use to identify the nominal species reported. We consider five species as valid from Ecuador: B. brevirostris (Günther), B. dahli Román–Valencia, B. pachacuti Eigenmann, B. phoenicopterus (Cope) and B. simus (Boulenger), in addition to the three new species described herein. B. peruanus is not a valid record from Colombia or Ecuador (Reis et al., 2003; Román–Valencia, 2011). These descriptions of new species of Bryconamericus from Ecuador stem from the authors ongoing revision of the genus Bryconamericus and are further proof of the as yet undocumented biodiversity of the genus. We provide keys to help identify species from Ecuador and separate them into the following regions: western coastal zone, eastern zone, Rio Napo drainage, Morona and Santiago River drainages, and the Conambo and Marañon River drainages in the Amazon River Basin.

In addition to the specimens listed below, see also those listed in Román–Valencia (1998, 2000, 2001, 2002a, 2002b, 2003a, 2003b, 2003c, 2003d, 2003e, 2005), Román–Valencia et al. (2008a, 2008b, 2009a, 2009b, 2011). Bryconamericus bolivianus: all from Bolivia: CAS 39506 (1), syntypes, La Paz, R. Amazon Basin, R. Colorado, tributary to lower Rio Bopi 10 miles above Huachi (= San Miguel de Huachi), 1 Sep 1921. CAS 39508 (2), paratype, El Beni, Rio Amazon Basin. Upper Rio Beni, Pena Colorado, four miles below Huachi, 1 IX 1921. CBF 06027 (3), Santa Cruz. Ichilo county, PN–AMI Amboro, Santa Rosa, 20 V 1996. CBF 06934 (3), la Paz, San Pedro county, Rio San Pedro. Coroico, Kaka, Beni. Madera, 12 XI 1996. CBF 08042 (10), Santa Cruz. Ichilo county, PN–AMI Amboro, Serrania Volcanes, 15 V 1996. CBF 07178 (10), Santa Cruz, Ichilo county, Parque National and integrated management area Amboro, 16 V 1996. CBF 08832 (50), Amazon. Madera, Beni. Cotacajes, afluente del Cotacajes, San Miguel de Huachi, 20 X 2008. CBF 06015 (10), Santa Cruz. Ichilo county, PN–AMI Ambooro, Tambo, 2 V 1996. CBF 07212 (5), Santa Cruz. Ichilo county, PN–AMI Amboro, Mairana, 19 V 1996. CBF 07239 (5), Santa Cruz. Ichilo municipality, PN–Ami Amboro, 10 Km Río debajo de Mairana, 19 V 1996. CBF 0021 (10). Cochabamba. Chapare county, Villa Tunari, 20 VI 1983. CBF 06023 (10), Santa Cruz. Ichilo county, PN–AMI Amboro, San Juan del Potrero, 23 V 1996. CBF 07879 (1), La Paz, F. Tamayo county, Rio Eslabòn, 21 VIII 1998. CBF 06695 (2), La Paz, Sudyungas county. Covendo, 16 XI 1996. UMSS 09850 (50), Amazon. Madera, Beni, Bopi, Rio Irpa Chuqui, San Pablo. UMSS 10002 (11), Amazon, Beni. Madera. Cotacajes, afluente del Cotacajes, San Miguel de Huachi, 1 V 2009. UMSS 05944 (38), Amazon. Madera. Mamore, Blanco, Rio San Joaquin, 10 XII 2004. UMSS (6), Amazon. Madera, Beni, Bopi, Rio Carrasco. Carrasco, 15 X 2009. UMSS 09900 (50), Amazon, Beni. Madera, Cotacajes, afluente del Cotacajes, San Miguel de Huachi, 5 XI 2009. UMSS10080 (50), Amazon. Madera, Beni, Kaka, Rio Tajilihui, 16 X 2009. UMSS 10093 (50), Amazon. Madera, Beni, Kaka, Rio Naranjakata, San José, 16 X 009. UMSS 10078 (4), Amazon. Madera, Beni, Kaka, Rio Tajilihui Alcoch, 16 X 2009. UMSS 01227 (50), Amazon. Itenez. Izozog, parapeti, Rio Parapeti, arriba de Camiri, 22 X 2005. UMSS 08831 (59), Amazon. Madera, Beni. Cotacajes, afluente del Cotacajes, San Miguel de Huachi 20 X 2008. UMSS 08921(50), Amazon. Madera, Alto Beni. Cuartel Tohomonoco, 9 XII 2008. UMSS 09827 (3), Amazon. Madera, Beni, Bopi, Río Yanamayu, Yanamayu, 23 X 2009. UMSS 05691 (8), Amazon. Madera. Mamore, Grande, Rio Mina Asientos, 6 IX 1998. Bryconamericus brevirostris: MUSM 3393 (3), Perú, Tumbes, San Jacinto, La Peña, Rio Tumbes, canal de Boca toma, 6 VII 1992. MUSM 3306 (3), Perú, Tumbes, Rio Tumbes to 500 m del puente, 5 II 1992. MUSM 6889 (50), Ecuador, Guayas, Rio Guayas basin. Cotimes, Rio Daule. MUSM 5732 (7), Perú, Tumbes, San Jacinto Bocatoma, Rio Tumbes, 16 VIII 1994. MUSM 3394 (9), Perú, Tumbes, Rio Tumbes en caudal

Material and methods Measurements were taken with digital calipers, recorded to hundredths of millimeters and usually expressed as percentages of standard (SL) or head length (HL) (table 1). Counts were made using a stereoscope with a dissection needle to extend the fins. Counts and measurements were taken from the left side of specimens when possible and were taken according to guidelines in Vari & Siebert (1990) and Armbruster (2012). Counts for the holotype are indicated with an asterisk (*). In the lists of types, the number of individuals is given immediately after the catalog number, which is followed by the range of standard length in mm (SL) for that lot; for example: MEPN 59236 (6), 48.5–72.9 mm SL indicates six individuals in lot MEPN 59236, with the smallest fish 48.5 mm SL and the largest 72.9 mm SL. All collections were from Ecuador. Acronyms used follow Sabaj–Pérez (2010). Meters above sea level is abbreviated as m a.s.l. Municipio is translated as county. Observations of bones and cartilage were made on cleared and stained specimens (C&S) prepared according to techniques outlined in Taylor & Van Dyke (1985) and Song & Parenti (1995). Bone nomenclature follows Weitzman (1962), Vari (1995), and Ruiz–C. & Román–Valencia (2006).

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de Riego, 6 II 1992. MUSM 3058 (1), Perú, Tumbes, Zarumilla. Matapalo, Rio Zarumilla, 11 XII 1990. MUSM 5765 (26), Perú, Tumbes, Zarumilla, Lepanga, Rio Zarumilla, 15 VIII 1994. MUSM 2582 (20), Tumbes, Rio Tumbes. Ca. La irrigación bocatoma, 10 VIII 1986. MUSM 1983 (20); Tumbes, Rio Zarumilla and pozo bajos, 12 VIII 1986. Bryconamericus dahli: all from Ecuador: MEPN 4023 (15), Esmeraldas, estero Sabalero to 600 m del campamento P. Chiquira, 22 X 1985. IUQ 3138 (1C&S.), Esmeraldas, Rio Matajita afl. Rio Mataje media hora centro comunal Mataje to Rio Mira and estero pensamiento to 800 m aguas sitio Pan, 22 X 1985. MEPN 4129 (83). Carchi, bahía pequeña de la quebrada Negra. MEPN 16–3979 (15), Esmeraldas, estero Boca del Onsole margen derecha del Río Guayllabamba River to 450 m Golondrina. MEPN 11150 (90), Esmeraldas, Estero Guavina, km 12, road orilla Río Esmeraldas, Valle de Sade, 8 III 1985. MEPN 11151 (19), Esmeraldas, Estero Pospi a 2 km de la desembocadura del Chimbogal river, on the Bravo river, 11 VIII 1985. MEPN 4309 (4) 55.1–68.5 mm SL, Esmeraldas, Río Mataje to 2 km del centro comunal Mataje, 8 II1988. MEPN 4022 (29), Esmeraldas, estero Claro 5 km antes de la caída del Río Bravo to 10 minutos del Río Chimbagol, 10 VIII 1985. IUQ 3140 (1C&S), Esmeraldas, estero Pistolosa media hora aguas debajo de Vargas Torres, IV 1984. Bryconamericvus cristiani: (Román–Valencia, 1998). Bryconamericus diaphanus: all from Perú: MUSM 17760 (50), Ucayali, Coronel, Portillo, cuenca Río Sheshe, RioTahuayo, 8º 06' 21'' S & 73º 55' 33,7'' W, 225 m a.s.l. MUSM 33585 (1), San Martin. Moyobamba, Rio Negro. MUSM 33475 (40), San Martin. Moyobamba, Rio Indeche, 826 m a.s.l. MUSM 17855 (50), Ucayali, Coronel Portillo, Rio Shesha, 8º 05' 46'' S, 73º 51' 36.4'' W, 230 m a.s.l., 12 II 2001. MUSM 10248 (1), Amazon, Bagua, Río Chiriaco, Nazareth, 15 II 1978. MUSM 32567 (5), San Martin. Caceres, Huicunga, Río Abiseo, Quebrada Machaco, 358 m a.s.l., 21 V 2005. Bryconamericus grosvenori: all from Perú: MUSM 32229 (3), Cusco, La Convencion, Echarate, Urubambva, Río Napitoniari, Monte carreto, 11 XI 2007. MUSM 12227 (10), Junin, Satipo, Quebrada La Florida, 11º 17' 06.36'' S, 74º 40' 45'' W, 23 IX 1995. MUSM 32340 (50), Cusco, La Convencion, Echarate. CCA Urubamba, Rio Shimaa, 21 XI 2007. MUSM 31314 (4), Cusco, La Convencion, Echarate, CCA, Alto Urubamba, Monte Camelo, quebrada Igorotoshiari. MUSM 36033 (8), Cusco, La Convencion, Echarate, Urubamba, Rio Piroton, quebrada Parotori, 21 V 2009. MUSM 32348 (10), Cusco, La Convencion, Echarate, Urubamba, 12 XI 2007. MUSM 12163 (1), Cusco, La Convencion, Río Urubamba, Río Picha, Pto. Huallana, 29 V 1997. MUSM 12115(2), Cusco, La Convencion, Rio Urubamba, Rio Picha. Mayapo, Rio Mayapo, 25 V 1997. MUSM 12072 (3). Cusco, La Convencion. Urubamba, 23 V 1997. Bryconamericus pachacuti: CAS 40829 (22), paratype, Perú. Cuzco, Rio Amazon, Rio Urubamba. IUQ 3155 (1C&S), Ecuador. Morona–Santiago, Rio Yapapa afl. Santiago River, 9 V 1991. MUSM 12268 (5),

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Perú, Junin, Chanchan, Riachuelo entre Rio Colorado, 7 XII 1987. MUSM 34001 (17), Perú, Ucayali, P. abajo, Rio Aguaytia entre Meihuga, Rio Tahuayo, 8 X 1986. MUSM 26975 (100), Perú, Junin, Satipo em Poyeni, Rio Tambo, Margen izquierda, 254 m a.s.l., 27 X 2005. MUSM 30529 (19), Perú, Cusco, Echarate, bajo Urubamba, Rio Camisea, Playa Paisita, 370 m a.s.l., 30 IV 2005. MUSM 32466 (4), Perú, Cusco, La Convencion, Echarate, quebrada Iherimpituari, Río Paratori, 16 III 2008. MUSM 11120 (24), Perú, Puno, Sandia, Rio Candamo, 358 m a.s.l., 2 IV 1997. MUSM 11061(9), Perú, Puno, Sandia, quebrada Candamo. MUSM 30199 (3), Perú, Pasco, Oxapampa, Pto. Bera afluente Río Apunmacayali, 26 V 2004. MUSM 35771 (9), Perú. Ucayali, Atalaya, Sepahua,quebrada Huayashi, 26 VII 2007. MUSM 12329 (1), Junin, Rio Perené via a Satipo, 21 IX 1995. MUSM 27058 (13), Perú, Junin, Satipo ccnn cheni, quebrada Pijireni, 269 m a.s.l., 26 X 2005. MUSM 37348 (22), Perú. Ucayali, Padre Abad, Rio Aguaytia, Rio Shamabo, 8º 50' 03'' S, 75º 34' 10'' W, 258 m a.s.l., 26 V 2009. MUSM 16144 (30), Perú. Ucayali, Padre Abad, Rio Aguaytia, quebrada Huiango km 18 via Curimana, 14 V 1997. MUSM 30363 (50), Perú, Pasco, Oxampa, Villa Rica. Caserío San Pedro de Pichanos, quebrada Pichanos, 3 VI 2004. MUSM 9155 (1), Perú, Puno. Carabajo, Rio Inambari, Sangaban, Rio Elcamayo, 24 VI 1994. MUSM 15866 (30), Perú. Ucayali, Padre Abad, Aguaytia, Rio Aguaytia, Rio Negro, 9º 02' 4.2'' S, 75º 30' 45.5'' W, 2 XI 1199. MUSM 3693 (11), Perú, Puno, Sandia, Rio Tambopata, quebrada a 500 m del campamento, 26 VIII 1992. MUSM 16720 (20), Perú, Junin. Chanchamayo, Rio Perené cerca al Rio Pancartambo, 1 V 1199. MUSM 26592 (8), Perú, Pasco, Oxampa, Pozazo, río Huancabamba, 21 X 2005. MUSM 37818 (30), Perú, Junin, Satipo. Mashira,quebrada Marado, 6 VI 2009. MUSM 18017 (50); Perú, Huanuco. CCA, Rio Pachitea, Honoris. Islãs Sargento Lores, 4 VII 2005. MUSM 33969 (34), Perú. Ucayali, Padre Abad, Aguaytia Rio Aguaytia, 19 X 2000. MUSM 20562 (50), Perú, Pasco, Oxapampa Pto. Bermudez, quebrada Ataz, 9 VIII 2002. MUSM 34324 (27), Perú. Cusco. Convencion, Echarate. CCNN Camaná, Alto Urubamba, 29 IX 2008. MUSM 32582 (24), Perú, Loreto, Andeas, Rio Pastaza, quebrada Thiyacil, 26 VIII 2007. MUSM 26590 (23), Perú, Pasco, Oxampa, Rio Huancabamba puente, 21 X 2005. MUSM 29125 (1), Perú. Madre de Dios, Tambopata, Rio Tambopata, playa Botafogo, 13 VI 2006. MUSM 25433 (2). Madre de Dios, Tambopata Mazuko, Rio Inambari, Quenque Creek, 8 IX 2009. Bryconamericus osgoodi: all from Perú: MUSM 14966 (5), Huanuco, Trigo Maria, Rio Huallaga, Rio Cueva de las Lechuzas, 13 VII 1998. Bryconamericus pectinatus: all from Perú: MUSM 3821 (1). Madre de Dios, seto, quebrada de Calli, 5 IX 1988. MUSM 3809 (1). Madre de Dios. Manú National Park, Rio Manu playa cerca de Cucha, 8 IX 1988. MUSM 3815 (1). Madre de Dios. Manú National Park, Rio Manu, 8 IX 1988. MUSM 29947 (3), Pasco, Okapampa. Icozacia, Rio Mayo, 20 V 2004. Bryconamericus phoenicopterus: MEPN 2120 (200), Ecuador, Zamora Chinchipe, playa frente al destaca-

Animal Biodiversity and Conservation 36.1 (2013)

mento militar. Mayaicu bajo, 3° 58' 15'' S, 78° 41' 15'' W, 18 VIII 1993. MEPN 44 (11), Ecuador, Zamora Chinchipe, Nangantza Río, playa frente al destacamento militar. Mayaicu, 18 VII 1993. IUQ 3135 and 3153 (2C&S), Ecuador. Morona Santiago, Rio Gualaquiza, 22 IX 1978. MUSM 20722 (4), Perú, San Martin, Tarapota. Morales, San Antonio, Rio Cuerabraza, 18 IX 1998. B. simus: all from Ecuador: BMNH 1898.11.4.71–7 (3), syntype. Carchi, Valle de Chotá, Norte de Ecuador. MEPN 4128 (83), Carchi, Bahía pequeña de la Quebrada Negra que rodea al sitio San Marcos, 8 XI 1987. Bryconamericus rubropictus (Braga, 2000). Bryconamericus terrabensis: Meek (Román–Valencia et al., 2008b). Bryconamericus turiuba (Román–Valencia et al., 2008b). Bryconamericus stramineus: MUSM 17039 (15), Brazil, n.s. Rio Formozinho a 17 km de R. Bonito, 21o 5' 14.6'' S, 56o 33' 35.7'' W, 6 IX 1998. Bryconamericus thomasi (Miquelarena & Aquino, 1995): all from Bolivia: CBF 01228 (10), Tarija, Gran Chaco county, 1.5 km. En linea recta al SO de Villamontes, 2 X 1988. UMSS 00806 (2), La Plata, Pilcomayo, Rio Pilaya, 12 VII 2005. UMSS 00740 (12), La Plata, Bermejo, Rio Emborozu, 12 VII 2005. UMSS (35), La Plata, Bermejo, Rio Orosas, 12 VIIl 2005. UMSS 03131 (14), La Plata, Bermejo, Rio Guadalquivir, 10 VII 2006. UMSS 04945 (1), La Plata, Bermejo, Gran de Tarija, Rio Tarija, 21 XI 2006. UMSS 04530 (3), La Plata, Bermejo, Grande de Tarija, Tarija, Rio Salinas, 5 X 2004. UMSS 5106 (11), La Plata, Bermejo, Arroyo Toro, 1 VII 2006. UMSS 00719 (1), Amazon, Mamoré, Rio Salado, 11 VII 2005. UMSS 00891 (8), Amazon, Itenez, San Pablo, Parapeti, Rio Heredia, 23 X 2005. UMSS 04968 (3), La Plata, Bermejo, Grande de Tarija, Rio Tarija, Rio Saycan, 6 X 2004. Bryconamericus sp. 1: CBF 05923 (2), Bolivia, Potosi, Linares municipality, Rio Mata, 11 X 1996. Bryconamericus sp. 2: MUSM 31598 (24), Perú, Lambayeque, Tenerife, Pacifico, Kañanis, Rio Huancdocinibe, Río Cañariaco, 19 IX 2007 Bryconamericus sp. 3: MUSM 19564 (65), Perú, Ayacucho, Huamanga, Rio Yucay, 20 VIII 2004. Bryconamericus sp. 4: MUSM 0832 (15), Perú. Ucayali, Rio Neshuya, 6 VII 1981. Knodus carlosi: all from Ecuador: MEPN 11149 (6), Orellana, Rio Jivino ca. 1,600 m del pozo Chontayacu. IUQ 3137 (1C&S), Ecuador, Sucumbios, Tipotini River, sector Mondaña 1 km aguas abajo, 28 I 1998. MUSM 21106 (40), Perú, Amazon. Condorcanqui. Cenepa, Rio Alto Cenepa, quebrada Capitan, 12 XI 2003. MUSM 8332 (70), Perú, Madre de Dios, Tambopata, Rio Tambopata, quebrada Garza, (13º 10' 21'' S, 69º 37' 41'' W), 2 X 1985. MUSM 16088 (30), Perú. Ucayali, Padre Abad, Rio Aguaytio, quebrada Moronal km 23.5 via Curimaná, 14 V 1997. MUSM 19013 (15), Perú, Pasco, Oxampa. Constitucion, Rìo El Dorado, 22 VII 2001. MUSM 33546 (2), Perú, San Martin, Moyomba, Rìo Alto Mayo, 810 m a.s.l., 24 VI 2006. MUSM 21284 (100), Perú, Amazon. Condorcanqui. Cenepa, quebrada Capitan, 75 m a.s.l., 13 XI 2005. MUSM 38111 (5), Perú, Loreto. cuenca del Marañon, Andoas, 201 m a.s.l., 27 IX 2008. MUSM 13590 (30),

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Perú. Cusco, La Convencion, Echarate, Rio Camisea quebrada Yopucuriari, 11 X 1998. MUSM 37638 (80), Perú, Loreto, Rio Corrientes, to 35 km de Jobano, 170 m a.s.l., 29 VI 2008. MUSM 35753 (50), Perú. Ucayali, Atalaya, Sepahua,quebrada Las Piedras, 25 VII 2009. MUSM 10601 (20), Perú, Amazon, Bagua. Imazita. Marañon, 11 XI 1996. MUSM 37696 (40), Perú, Loreto, Rio Corrientes, quebrada a 1 km de Río Corrientes, 30 VI 2008. MUSM 15654 (31), Perú, San Martin, Tarapota, Ahuashiyacu, Parte alta, 29 XI 1997. MUSM 32645 (4), Perú, San Martin. M. Caceres, Huicuago, Rio Huallabamba, 340 m a.s.l., 25 VI 2008. MUSM 31390 (13), Perú, Cusco, La Convencion, bajo Urubamba, 8 VI 2004. MUSM 30912 (30), Perú, Cusco, La Convencion, Echarate cca. Bajo Urubamba, 28 I 2005. MUSM 32564 (19), Perú, San Martin. M. Caceres, Huicungo, PNRA, abisco, 358 m a.s.l., 21 V 2008. MUSM29963 (50), Perú, Huanuco, Pto. Inca. Codo del Pozas o Ca Paliazu, quebrada Charepa, 22 V 2004. MUSM 26075 (90), Perú, Cusco, La Convencion, Echarate, Rio Camisea, 30 IX 2005. MUSM 30127 (10), Perú, Huanuco, pto. Inca. Codo de Rozuzo, 22 V 2004. MUSM 30677 (20), Perú. Cusco, La Convencion, Echarate, bajo Urubamba, 28 IV 2005. MUSM 30055 (20), Huanuco, pto. Inca. Codo de Pozuso, Pucacurga, Pucacurgacreek, 22 V 2004. MUSM 31647 (26), Perú. Cusco, La Convencion, Echarate, bajo Urubamba, 9 X 2007. MUSM 31463(24), Cusco, La Convencion, Echarate. Ca. del bajo Urubamba to Katshingari, 478 m a.s.l., 21 I 200. MUSM 30089 (25), Perú, Huanuco, pto. Inca codo del pozuzo, Huanpumayo, 290 m a.s.l., 22 V 2005. MUSM 33455 (50), Perú, San Martin, Moyobamba, Rio Negro, 21 X 2001. MUSM 32822 (50), Perú, Huanuco, Aucayacu, Jose Crespoy Castillo, CP Consuelo, Rios Huallaga and Aucayacu, 26 I 2008. MUSM 34323 (100), Perú. Cusco, La Convencion, Echarate, Alto Urubamba, boca Quebrada Kcishingar, 29 IX 2008. MUSM 29441 (40), Perú, Pasco, Oxapampa, Iscozacin, Rio Chuchurras, quebrada Helgo, 20 VI 2000. MUSM 32106 (50); Perú, Loreto, Andaces, CCA, Rio Corrientes, R. Macusari, 7 VII 2006. MUSM 31366 (29); Perú. Cusco, La Convencion, Echarate, bajo Urubamba, 22 VI 2004. Knodus delta: (Román–Valencia, 2003a, 2003b), all from Ecuador: MEPN 28 (11), Sucumbios, Rio Tiputini. Mondaña sector, 1 km downstream. MEPN 29 (50), Napo, Rio Huataracu. MEPN 35 (40), Sucumbios, Duguno River, 2 km from comuna Cofan. IUQ 3139 (2C&S), Sucumbios, Dugunom Rio 2 km de la comuna Cofan. Knodus heteresthes: IUQ 1166 (1C&S), 49.94 mm SL, Colombia, Guaviare, Retorno county, Amazon, Caño el Tigre en la vía a San José del Guaviare municipality. IUQ 1170 (1C&S), 40.82 mm SL, Colombia, Vaupés, Amazon, bajo Río Apoporis, pequeño drenaje del lago Taraira. IUQ 400 (3), Colombia, Vaupes, Amazon, drenaje del lago Taraira, bajo Río Apaporis. Knodus hypopterus: IUQ 1650 (1C&S), 35.34 mm SL, Colombia, cuenca Río Caqueta, Quebrada Manigua en el puente via Florencia–Belen. IUQ 395 (2C&S), 40.54–49.55 mm SL, Colombia, cuenca Río Caqueta, Quebrada Manigua en el puente via Florencia–Belen. IUQ 398 (9), Colombia. Caquetá, Amazon,

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Table 1. Morphometric and meristic data of Bryconamericus zamorensis n. sp., B. oroensis n. sp. and B. bucayensis n. sp. (standard and total lengths in mm, mean values in parenthesis). Tabla 1. Datos morfométricos y merísticos de Bryconamericus zamorensis sp. n., B. oroensis sp. n. y B. bucayensis sp. n. (longitudes estándar y total en mm, medias entre paréntesis).

B. zamorensis n. sp.

B. oroensis n. sp.

B. bucayensis n. sp.

Paratype Holotype

Standard length Total length

35.09–64.76 52.61

34.91–74.51 74.51

(45.65)

(49.82)

43.96–82.02 64.38

42.37–88.92 88.82

(56.80)

(60.33)

31.97–39.76 31.97

27.84–33.37 30.48

(35.25)

(30.64)

52.41–57.70 53.79

50.09–53.66 51.03

49.48–72.88 87.05 (59.15) 63.72–90.41 108.49 (74.70)

Percentages of SL Body depth Snout–dorsal fin distance

(55.60)

Snout–pectoral fin distance 23.71–27.92 26.29 Snout–pelvic fin distance Snout–anal fin distance

(25.65)

(24.89)

42.84–50.12 45.18

43.70–47.82 43.7

(46.89)

(45.83)

57.54–65.14 57.93

58.85–63.74 58.95

(61.66)

(61.11)

Dorsal fin–hypural distance 45.20–52.87 50.83 Dorsal–fin length Pectoral–fin length Pelvic–fin length Caudal peduncle depth Caudal peduncle length Head length Dorsal–anal fin distance Dorsal–pectoral distance Anal–fin length

(51.60) 23.07–26.87 24.56

(49.00) 30.0–37.55

31.51

47.94–52.73 52.09 (49.96) 28.04–33.60 29.23

31.73–38.31 38.11 (34.59) 51.55–53.05 52.69 (52.24) 23.89–26.27 23.89 (25.04) 44.63–47.04 45.03 (45.61) 59.73–64.64 63.35 (61.86) 49.72–52.65 51.87 (51.05) 20.91–37.19 22.4

(33.92)

(30.87)

39.16–44.83 42.41

32.03–39.32 38.38

(42.34)

(37.91)

18.45–26.45 24.33

22.04–25.92 22.04

(22.64)

(23.76)

18.10–24.95 20.49

16.70–23.82 18.47

(21.27)

(19.71)

(10.71)

11.77–16.24 14.41

12.11–14.86 12.74

7.85–10.68

(14.34)

(13.66)

(10.68)

13.26–20.23 16.84

16.47–21.02 17.8

(16.78)

(18.01)

9.81–13.74

11.88

(11.59) 8.12–13.82

10.04

10.83–12.98 12.98 (11.85) 9.37–13.45 10.67

(11.29)

(11.13)

22.45–28.08 23.97

23.43–27.75 23.43

(25.09)

(25.25)

(25.54) 20.35–34.54 21.56 (23.59) 12.53–28.67 14.53 (16.03) 9.35–11.69 11.43 7.85

22.35–25.86 22.8 (24.15) 32.16–46.37 38.51 (37.06) 37.22–52.11 42.97 (41.75) 13.77–18.09 13.77 (16.15)

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Table 1. (Cont.)

B. zamorensis n. sp.

B. oroensis n. sp.

B. bucayensis n. sp.

Paratype Holotype

Percentages of HL Snout length

21.01–32.19 24.98

Orbital diameter

28.32–39.35 38.14

26.12–38.10 26.11

(34.41)

(32.72)

41.33–53.04 47.51

40.02–51.55 51.55

(46.93)

(43.98)

Maxilla length

26.31–36.99

Interorbital distance

29.9

(26.32)

Postorbital distance

20.0–29.90

(26.32)

34.5

22.26–36.18 27.32

(31.31) 29.16–39.08

(29.23)

31.4

32.76–39.84 36.2

16.63–22.92 22.72 (21.14) 30.82–40.00 32.29 (34.91) 37.35–47.88 45.94 (41.84) 24.57–27.42 27.3 (26.02) 29.79–35.10 32.1

(32.87)

(36.99)

(31.83)

33–37

38–40

41–49

Meristics Lateral–line scales

Scale row between dorsal–fin origin and lateral line 6–8

6–9

Scale rows between anal–fin origin and lateral line

4–7

6–8

6–10

Scale rows between pelvic–fin and lateral line

5–6

Dorsal–fin rays Anal–fin rays Pelvic–fin rays Pectoral–fin rays

6–8

7–9

ii, 7, i

iii, 9

ii, 9

iii, 22–25

iii, 25

iii–iv, 24–28 iv, 24

iv–v, 33–38 v, 35

ii, 6–7

ii, 6

i, 7

ii, 9–10

ii, 10

ii, 9–12

ii, 12

Quebrada Pompella sobre el puente via Florencia– Belem, 14 XII 1998. IUQ 405 (23), Colombia, Florencia, Amazon, Río Caquetá basin, Quebrada Manigua en el puente via Florencia–Belem. Knodus meridae: all from Venezuela: AUM 44950 (5), Zulia, Lago Maracaibo basin, Rio Chama, slightly E of bridge on Hwy 7 SW of Lagunillas. IUQ 391 (30), Zulia, Lago Maracaibo basin, Río Aguas Calientes. IUQ 39 (9), Portuguesa, Portuguesa system, Rio Las Marias, cerca de Guanare. IUQ 393 (15). Mérida, Rio Chama, puente No. 4 a 20 km N–O de Mérida. IUQ 1207 (2C&S), 45.84–47.11 mm SL, Mérida, Rio Chama, puente # 4 a 20 km N–O de Mérida. IUQ 1164 (1C&S), 28.96 mm SL, Zulia, Rio Aguas Calientes. IUQ 2196 (1C&S), 36.81 mm SL, Portuguesa, Rio Portuguesa system, Rio Las Marias, cerca de Guanare. Knodus orteguasae: all from Colombia: IUQ 408 (13), Putumayo, Amazon, Rio Orito en el puente via

a Caldero. IUQ 1209 (3C&S), 37.33–48.58 mm SL, Putumayo, Oríto county, Amazon Rio Putumayo basin, Rio Orito en el puente vía a Caldero. Knodus pasco: all from Perú: MUSM 10308(2). Ucayali, Amazon, Pedro Abad, Rio Huacamayo, a 5 km de Aguaytía y 155 km desde Pucallpa,13 IX 1994. MUSM 10030 (4). Ucayali, Amazon, Purús, Esperanza, quebrada o arroyo Esperancillo, 09º 42' S, 70º 40' W, 4 IX 1994. MUSM 13655 (27), Cusco, Amazon, La Convencion, Echorate, San Martin, 11 VII 1998. MUSM 13907(4), Cusco, Amazon, La Convencion. MUSM 13664 (7), Cusco, Amazon, La Convención, Echarate, San Martin, quebrada Natsiringari, 10 X 1998. MUSM 13591 (1). Cusco, Amazon, La Convencion. MUSM 13914 (4), Cusco, Amazon, La Convencion. MUSM 13549 (6). Cusco, Amazon, La Convencion. Cashihari, 9 I 1998. MUSM 13576 (3). Cusco, Amazon, La Convencion, Pagoreni, 2 IX 1998. MUSM 13591 (3). Cusco, Amazonia, La Convención. MUSM 13906 (18), Cusco,

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Amazon, La Convencion, Echarate, Pagoreni,quebrada Oshetoato. MUSM 31700 (6). Cusco, Amazon, La Convencion, Echarate, Río Urubamba basin, quebrada Choro, Rio Camisea, 365 m a.s.l., 5 X 2007. MUSM 11081 (29), Puno, Amazon, Sandia, Rio Candano,quebrada Ebcbahuacji, 31 III 1997. MUSM 17379 (26), San Martin, Amazon, Rioja, Rio Trayaca, 3 IV 1998. MUSM 10351 (49), Puno, Amazonas, Sandia. Candamo, 11 XII 1996. Knodus moenkhausii: CBF07898 (2), Bolivia, La Paz, F. Tamayo, Rio Eslabon, 21 VIII 1998. CBF 3129 (5), Bolivia, Beni, Yacuma, reserva de la Biosfera estación Biológica Beni en la estancia 08, 5 IV 1986. CBF 4793 (8), Bolivia, Pando. Manuripi, Rio Nareuda, arriba del campamento Nareuda en la playa, 4 IX 1996. CBF08067 (8), Bolivia, Beni, J. Ballivian, puente camino San Borja–Santarosa (Achaparina), 8 VIII 1997. MUSM 24668 (32), Perú. Madre de Dios, Amazon, Tahuamanu, RioTahuamanu, Rio Maymanu, 263 m a.s.l., 23 VII 2004. MUSM 26419 (16), Perú. Cusco, Amazon, Quispicanchi, Camananti. Cueva Araza, 1 VIII 2005. MUSM 29882 (149), Perú, Pasco, Oxapampa, Pto. Bermúdez,constitución, plaza 3, R. Palcazu, 254 m a.s.l., 26 IX 2004. MUSM 13691 (111), Perú. Ucayali, Amazon, Atalaya, Sepahua, Rìo Urubamba, Quebrada Comarillo, 4 XI 1998. Knodus shinahota: all from Perú: MUSM 15576 (20); Amazon. Madre de Dios, Tambopata, Terzdacolpa, Rio Tambopata, 21 VIII 1992. MUSM 25439 (127). Madre de Dios, Tambopata. Mazuka CCA Rio Inambari, Rio Inambari, 311 m a.s.l., 27 VII 2004. Knodus victoriae: CBF 4834 (3), Bolivia, Pando. Manuripi, Rio Nareuda, arriba del campamento Nareuda en la playa, 4 IX 1996. MUSM 16368 (1), Perú. Madre de Dios, Tambopata, quebrada Jayave km 127, 20 II 1998. Knodus mizquae: UMSS 00700 (7), Bolivia, Amazon. Mamore río Salado, 11 VII 2005. Knodus sp. 1: CBF 6736 (2), Bolivia, La Paz, Sudyungas, Amazon, San Juan de Piquendo, 19 Nov. 1996. CBF 7423 (10), Bolivia, La Paz. Iturralde county. Campamento Candelaria (PN AMI Madidi), 24 IV 2001. CBF 07897 (2), Bolivia, La Paz, F. Tamayo county, Rio Eslabon, 21 VIII 1998. Knodus sp. 2: CBF 06341 (5), Bolivia, alto Paraguay, varias localidades: Rios Paraguay, Negro, Apa and Riacho La Paz, IX 1997. CBF 06368 (8), Bolivia, Alto Paraguay, varias localidades: de los Rios Paraguay, Negro, Apa and riacho La Paz, IX 1997. Knodus sp. 3: MUSM 17612 (25), Perú, Loreto, Amazon. Ucayali. Contamana, sierra de la Contamana, Río Ucayali, 7o 10' 54.4'' W, 74o 57' 10.4'' S, 16 XI 2000. MUSM 32466 (4). Knodus sp. 4: UMSS 00699 (21) Bolivia, Beni, Amazon. Mamore, Rio Salado, 11 VIIl 2005. UMSS 04413 (23), Bolivia, Beni, Amazon, Madera. Mamore, Ichilo, Rio Bolivar, 25 VI 2003. UMSS 07345 (17), Bolivia, Amazon. Itenez, Rio Blanco, 12 XII 2004. Knodus sp. 5: MUSM 10335 (31), Perú, Puno, Sandia, Rio Candamo, quebrada Unión, 7 XII 1996. MUSM 10351 (50), Perú, Puno, Sandia Candamo, 11 XII 1996. MUSM 6158 (25), Perú, Amazon, Condorcanqui, Marañon, Rio Comainas, 19 VII 1994. MUSM 0160 (30), Perú. Ucayali, Pucallpa. Ivita, Pisci granja, 31 V 1983.

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MUSM 3569 (30), Perú, Junin. Chanchamaya, elcimo, Rio Poncantamanbo. MUSM 10335 (31), Puno, Sandi, Rio Candamo. Union creek, 7 XII 1996; Perú, Cusco, Amazon, subcuenca Rio Paratori, La Convencion, Echarate, quebrada Iherimpituari, 16 III 2008. Bryconamericus bucayensis n. sp. (table 1, figs. 1–2) Holotype: MEPN 11125, 87.1 mm SL, Ecuador, Guayas, Rio Bucay 3 km. upstream from bride on Naranjal–Machala road, 79º 42' 31'' W, 02º 39' 45'' S, 80 m a.s.l., 24 IX 1992. Paratypes: all from Guayas, Ecuador: MEPN 59236 (6), 48.5–72.9 mm SL, Rio Bucay 3 km. upstream from bride on Naranjal–Machala road, 79º 42' 31'' W, 02º 39' 45'' S, 85 m a.s.l., 24 IX 1992. IUQ 3143 (1C&S), 56.9 mm SL. Collected with holotype. MEPN 11126 (5), 38.2–75.3 mm SL, Guayas, Rio Minas. Cooperativa 23 XI, 9 km south of Naranjal, R. Barriga. 22 IX 1992. MEPN 9847 (18), 24.0–44.3 mm SL, Río Minas, at Cooperativa 23 XI, 9 km south of Naranjal, 79º 39' 16'' W, 02º 41' 26'' S. MEPN 5980 (14), 64.3–74.3 mm SL, Rio Tenguel, 180 m upstream from eva Esperanza bridge. Machala Road, 79º 43' 56'' W, 03º 00' 24'' S, 110 m a.s.l., MEPN 6029 (3), 68.5–74.1 mm SL, Rio Tenguel, La Esperanza sector, 79º 44' 22'' W, 02º 59' 29'' S, 110 m a.s.l., 23 IX 1992. Diagnosis Bryconamericus bucayensis n. sp. is distinguished from congeners by: the number of scales between the lateral line and the pelvic–fin insertions (7–8 vs. 2–7, except B. terrabensis with 7–8 and B. arilepis with 9–10); the number of branched anal–fin rays (33–38 vs. 31 or fewer);the number of anterior anal–fin rays covered by a row of scales at their bases (28–31 vs. 4–26); presence of a wide anterior maxillary tooth, at least twice as wide as the posterior tooth, both of which are pentacuspid (vs. maxillary teeth of same size) (fig. 2). Description Morphometric data in table 1. Greatest body depth at dorsal–fin origin (mean maximum body depth about 34.6% SL). Area above orbits flat. Dorsal profile of head and body oblique from supraoccipital to dorsal–fin origin and from last dorsal–fin ray to caudal–fin base. Ventral profile of body rounded from snout to anal–fin base. Caudal peduncle laterally compressed. Head and snout short, mandibles equal, mouth terminal, lips soft and flexible and not covering outer row of premaxilla teeth; ventral border of upper mandible curved; posterior edge of maxilla reaching anterior edge of orbit; opening of posterior nostrils vertically ovoid; opening of anterior nostrils with a membranous flap. Distal tip of pectoral fin surpassing pelvic–fin insertions. Distal tip of pelvic fin not reaching anal–fin origin. Premaxilla with two rows of teeth. Four teeth of outer row pentacuspid, lateral teeth anteriorly displaced in relation to medial teeth, together forming an arc from ventral view, with central cusp larger. Inner premaxilla row with four tetra– or pentacuspid teeth that diminish gradually in size. Maxilla long. More than three–quarters length of second infraorbital, anterior margin with

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Fig. 1. Bryconamericus bucayensis n. sp., holotype, MEPN 11125, 87.1 mm SL, Guayas state, Bucaya River 3 km, upstream from Camboya Naranjal, Module 2. Fig. 1. Bryconamericus bucayensis sp. n., holotipo, MEPN 11125, 87,1 mm de longitud estándar, Estado de Guayas, río Bucaya 3 km, antes de llegar a Camboya Naranjal, Módulo 2.

notches, with two pentacuspid teeth, dorsal tooth wider than ventral. Dentary with five large front pentacuspid teeth with central cusp largest, followed by four or five small teeth, anterior–most tricuspid and subsequent ones unicuspid. Foramina on anterior ventral surface of sphenotic channel absent. Cleithrum with pointed dorsal process that surpasses entire supracleithrum, which is joined to postemporal. Lateral line complete, perforated scales 41–49 (44*, mean = 46.44). Scale rows between dorsal–fin origin and lateral line 6–9 (8*, mean = 6.76); scale rows between lateral line and anal–fin origin 6–10 (10*, mean = 7.76); scale rows between lateral line and pelvic–fin insertion 7–9, (8*, mean = 7.29). Predorsal scales arranged in regular series. Dorsal–fin rays ii, 9 (n = 40); first unbranched ray approximately one–half

length of second unbranched ray. Dorsal–fin origin located near middle of body and posterior to vertical through pelvic–fin origin. Anal–fin rays iv–v, 33–38 (v, 35*, n = 48). Anal–fin origin posterior to vertical through base of first dorsal–fin ray. Pectoral–fin rays ii, 9–12 (ii, 12*, n = 40). Pelvic–fin rays i, 7 (n = 48). Pelvic–fin origin anterior to vertical through dorsal–fin origin. Total number of vertebra 37–38. Secondary sexual dimorphism Sexually mature males with seven to twelve spines present on anterior branched anal–fin rays. Including longest anal–fin ray; one row of large hooks on 1st to 11th, each with 5–14 hooks. From 11–16 spines present along ventral surface of branched pelvic–fin rays; from 13–19 large hooks, located on both branches of rays, and extending on to anterior most part.

1 mm Fig. 2. Lateral view of maxillary bone of B. bucayensis n. sp. Fig. 2. Vista lateral del hueso maxilar de B. bucayensis sp. n.

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Color in alcohol Dorsum dark, greenish. Body with silvery lateral band from posterior edge of opercle to base of caudal fin. Humeral spot diffuse, round with faint ventral projection, second humeral spot faint, transverse. Peduncle spot arrow–head shaped, not extending anteriorly beyond caudal peduncle. Continued on to middle caudal–fin rays. Ventral–lateral region of body, from snout tip to caudal peduncle yellow. Fins hyaline. Distribution This species is so far known from the Bucaya, Tengui, Tenguel, and Guayas Rivers, Guayas, Pacific basin, Ecuador. Etymology Bryconamericus bucayensis is named after the Rio Bucaya, where the type series was collected. Bryconamericus zamorensis n. sp. (table 1, figs. 3–5) Holotype: MEPN 11134. Male 51.2 mm SL, Ecuador, Zamora Chinchipe, Rio Nangaritza at confluence with Rio Numpatakaime, 78º 31' 20'' W, 04º 20' 31'' S, 950 m a.s.l., 21 VII 1993. Paratypes: all Zamora Chinchipe, Ecuador: MEPN 11135 (37), 32.0–39.8 mm SL, Rio Nangaritza at confluence with Rio Numpatakaime, 78º 31' 20'' W, 04º 20' 31'' S, 950 m a.s.l., 21 VII 199. IUQ 3144 (2C&S), 48.0–55.1 mm SL collected with holotype. IUQ 3199 (2) 52.5–62.2 mm SL collected with holotype. MEPN 11136 (30), 37.4–72.9 mm SL, Rio Chicaña upper tributary of Rio Zamora, 78º 44' 35'' W, 03º 41' 48'' S, 1,150 m a.s.l., 29 III 1979. MEPN 7785 (3), 33.9–54.7 mm SL, Rio Salado South of Los Encuentros, tributary of Rio Zamora, 77º 37' 20'' W, 03º 45' 30'' S, 900 m a.s.l., 29 III 1979. MEPN 11136

(8), 47.4–64.5 mm SL, San Antonio de Guadalupe Creek, 78º 53' 45'' W, 03º 50' 8'' S, 1,250 m a.s.l., 21 III 1979. MEPN 4612, (40), 36.5–65.0 mm SL, Rio Chicaña. Upper tributary of Río Zamora, (78º 44' 35'' W, 03º 41' 8'' S), 1,750 m a.s.l., 29 III 1979. MEPN 11139 (3), 33.5–62.0 mm SL, at confluence of Numpatakaime and Nangaritza rivers, 78º 31' 20'' W, 04º 20' 31'' S, 950 m a.s.l., 21 VII 1999. Diagnosis Bryconamericus zamorensis n. sp. is distinguished from congeners by having five teeth on the maxilla (vs. 1 or 2 teeth on maxilla), except B. rubropictus and B. thomasi, from which it differs in a reticulated pattern over the lateral stripe, generated by the concentration of melanophores, the scale margins, all along the sides of the body, by the number high of branched anal fin ray, vertebra, and the low number of branched dorsal fin ray. Dorsal prolongation of rhinosphenoid forming a bony wall between olfactory nerves (vs. dorsal prolongation of rhinosphenoid between olfactory nerves absent). Lateral process of palatine over the anterolateral margin of ectopterygoid (vs. palatine without lateral processes that laterally surpass the ectopterygoid) (fig. 4). Distal tip of sphenotic spine, see laterally, wide and undulated (vs. narrow) (fig. 5). Description Morphometric data in table 1. Maximum body depth at dorsal–fin origin (mean maximum body depth about 35.3% SL). Area above orbits convex. Dorsal profile of head curved from supraoccipital to dorsal origin and oblique from last dorsal–fin ray to base of caudal fin. Ventral profile of body curved from snout to base of anal fin. Caudal peduncle laterally compressed. Head and snout short, mandibles equal, mouth terminal, lips soft and flexible, and not covering outer row of premaxilla teeth; ventral border of upper mandible straight; pos-

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Fig. 3. Bryconamericus zamorensis n. sp., holotype, MEPN 11134, male 51.2 mm SL, Zamora Chinchipe state, Rio Narigaritza at its confluence with the Numpatakaime River. Fig. 3. Bryconamericus zamorensis sp. n., holotipo, MEPN 11134, macho 51,2 mm de longitud estándar, Estado de Zamora Chinchipe, río Narigaritza en su confluencia con el río Numpatakaime.

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terior edge of maxilla reaching anterior edge of orbit; opening of posterior nostrils vertically ovoid; opening of anterior nostrils with membranous flap. Distal tip of pectoral fin not surpassing pelvic–fin insertion. Distal tip of pelvic fin not surpassing anal–fin origin. Premaxilla with two rows of teeth. Four or five teeth of outer row tricuspid and arranged in zigzag. Inner premaxilla row with four tetra or pentacuspid teeth that diminish gradually in size. Maxilla large, reaching posterior margin of second infraorbital; with five tricuspid teeth, the central cusp widest. Dentary with five large front pentacuspid teeth with central cusp largest, followed by eight to 12 small teeth, anterior–most tricuspid and subsequent ones unicuspid. Foramina on anterior dorsal surface of sphenotic channel absent. Pectoral girdle with pointed dorsal process above cleithrum whose end dorsal surpasses transversally supracleithrum. Lateral line complete, perforated scales 33–37 (37*, mean = 36.44). Scale rows between dorsal–fin origin and lateral line 6–8 (8*, mean = 6.76); scale rows between lateral line and anal–fin origin 4–7 (5*, mean = 6.76); scale rows between lateral line and pelvic–fin insertion5–6 (6*, mean = 5.29). Predorsal area with scales arranged in regular series only in anterior half; posterior half with six medial scales that alternate with lateral scales that extend up onto predorsal midline with scales. Dorsal–fin rays ii 7. I (n = 40); first unbranched ray approximately one–half length of unbranched second ray; dorsal–fin origin located near middle length of body and posterior to vertical through pelvic–fin origin.

1 mm

RSp

Vom

Lp–Pal

Fig. 4. Lateral view of grave nasal in B. zamorensis (see that the lateral process of palatine surpass the anterolateral margin of ectopterigoid): Rsp. Rhinosphenoid; Vom. Vomer; Lp–Pal. Lateral process of palatine. Fig. 4. Vista lateral de la fosa nasal en B. zamorensis (se ve que el proceso lateral del palatino cubre el borde anterolateral del ectopterigoide): Rsp. Rinosfenoide; Vom. Vómero; Lp–Pal. Proceso lateral del palatino.

Ptt

Tsc

Pte Scl Sph–Cnl

LL1 Cle

Sph–Spn 1 mm

Pte

Pc1

Sph–Spn

Sph–Cnl Sph–For

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Fig. 5. A. See lateral of canal and spine of sphenotic; dorsal end of cleitrum transversally surpasses the supracleitrum in B. zamorensis. B. Lateral view of canal, spine and foramen on sphenotic in B. phoenicopterus: Sph–Spn. Spine of sphenotic; Sph–Cnl. Canal of sphenotic; Sph–For. Foramen of sphenotic; Pte. Pterotic; Tsc. Canal semicircular; Ptt. Pos temporal; Scl. Supracleitrum; Cle. Cleitrum; LL1. First scale with pore of canal later sensorial; Pc1. Postcleitrum 1. Fig. 5. A. Vista lateral del canal y la espina del esfenótico; el extremo dorsal del cleitro sobrepasa transversalmente el supracleitro en B. zamorensis. B. Vista lateral del canal, la espina y el foramen del esfenótico en B. phoenicopterus: Sph–Spn. Espina del esfenótico; Sph–Cnl. Canal del esfenótico; Sph– For. Foramen del esfenótico; Pte. Pterótico; Tsc. Canal semicircular; Ptt. Postemporal; Scl. Supracleitro; Cle. Cleitro; LL1. Primera escama con poro del canal laterosensorial; Pc1. Poscleitro 1.

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Anal–fin rays iii, 22–25 (iii, 25*, n = 126); anal–fin origin posterior to vertical through base of first dorsal–fin ray. Pectoral–fin rays ii, 9–10 (ii, 10*, n = 40). Pelvic–fin rays ii, 6 (n = 126); pelvic–fin origin anterior to vertical through dorsal–fin origin. Caudal fin not scaled, forked with short pointed lobes, principal caudal rays 1/17/1 with 11/11 procurrents. Total number of vertebra 36–37. Secondary sexual dimorphism Sexually mature males have no spines on simple rays; one row of large spines is present on 1st to 14th branched anal–fin rays, each ray with 6–13 hooks, located on most medial branch. From 12–18 large spines on branched rays of pelvic fin, located on both branches of rays, and extending on to anterior–most part. Color in alcohol Dorsum dark brown. Body without dark lateral stripe from posterior edge of opercle to base of caudal fin. Humeral spot dark, round with a vertical elongate on ventral margin, second spot humeral absent; lateral band with a reticulated pattern, which extends from the humeral region to the caudal region. Fins hyaline except anal fin, which has dark band. Ventral–lateral region of body between snout tip and caudal peduncle yellow. Distribution This species is so far known from the Narigaritza River at its confluence with the Numpatakaime River, Zamora River drainage, Zamora Chinchipe, Amazon Basin, Ecuador. Etymology Bryconamericus zamorensis is named for the Zamora Chinchipe state, where the type series was collected. Bryconamericus oroensis n. sp. (table 1, figs. 6–7) Holotype: MEPN 11140, 74.5 mm SL, male, Ecuador, El Oro, pond in Ortega, 1.5 km NE of Zaruma, 79º 34' 05'' W, 03º 39' 46'' S, 930 m a.s.l., 14 V 2004. Paratypes: MEPN 11141 (8), 34.9–74.5 SL, El Oro, pond in Ortega, 1.5 km NE of Zaruma, 79º 34' 05'' W, 03º 39' 46'' S, 930 m a.s.l., 14 V 2004.IUQ 3145 (1C&S), 66.1 mm SL collected with holotype. MEPN 5445 (42), 38.8–69.8 mm SL, El Oro, Quebrada Digitamo, 120 m south of Trapiche, 18 km from Portovelo, 79º 40' 29'' W, 03º 36' 49'' S, 980 m a.s.l., 24 VIII 1994. MEPN 2886 (11), 55.6–88.7 mm SL, El Oro, Rio Playón en Santa Rosa, 79º 55' 20'' W, 03º 35' 00'' S, 100 m a.s.l., 13 IV 1979. IUQ 3651 (4), 55.3–70.1 mm SL, El Oro, Rio Playón en Santa Rosa, 79° 55' 20'' W, 03º 35' 00'' S, 100 m a.s.l., 13 IV 1979. MEPN 2871 (9), 40.1–87.7 mm SL, El Oro, Ortega, 15 km NE of Zaruma, 79º 36' 41'' W, 03º 42' 51'' S, 950 m a.s.l., 14 V 2004. IUQ 3150 (1C&S), 45.9 mm SL, El Oro, Río Negro, tributary of Amarillo River near Portobello. IUQ 3148 (1C&S), 60.6 mm SL, El Oro, Huertas at Jubones. MEPN 11144 (1), 55.3 mm SL, El Oro, Rio Calera at Huertas, 79º 40' 20'' W, 03º 5' 40'' S, 1,450 m a.s.l., 12 VIII 1978. MEPN 11145 (10), 41.2–68.9 mm SL, Loja, Rio Macará near international highway bridge and the city of Macará, 79º 57' 05'' W 04º 24' 30'' S), 800 m a.s.l., 20 VII 1978. MEPN 11146 (2), 43.8–63.8 mm SL, El Oro,

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Rio Negro, tributary of Amarillo River near Portobelo, (79º 43' 24'' W, 03º 45' 18'' S), 1,180 m a.s.l., 24 VIII 1994. MEPN 11147 (9), 39.4–96.6 mm SL, El Oro, artificial pond NE of Zaruma in Huertas parish, (79º 22' 40'' W, 04º 03' 50'' S), 1,500 m a.s.l., 16 VIII 1978. MEPN 2877 (29), 41,9–65.6 mm SL, Loja, Rio Catamayo at Arenal bridge, (79º 22' 40'' W, 04º 03' 50'' S), 1,500 m a.s.l., 18 VIII 1978. Diagnosis Bryconamericus oroensis n. sp. is distinguished from congeners by having a dark lateral stripe overlaid with a peduncular spot and a reticulated pattern on the sides of the body (vs. peduncular spot and other body pigments not superimposed over a dark lateral stripe). It has three simple dorsal–fin rays, the first only visible in cleared and stained material and articulated, along with the second simple ray, with the first dorsal pterygiophores; third simple ray longer, and articulated with second dorsal pterygiophores (vs. only two simple dorsal–fin rays, both articulated with first dorsal pterygiophores); anterior frontals both separated and so the fontanel front parietal continues on the mesethmoid (vs. anterior tips of frontals united and fontanel front parietal no continued on the mesethmoids). Description Table 1 shows morphometric data. Body depth on dorsal–fin origin (mean maximum body depth about 30.6% SL). Area above orbits convex. Dorsal profile of head presents variation, curved from snout to dorsal–fin origin or straight, oblique from last dorsal–fin ray to base of caudal fin. Ventral profile of body curved from snout to base of anal fin. Caudal peduncle laterally compressed. Head and snout short, mandibles equal, mouth terminal, lips soft and flexible, and not covering outer row of premaxilla teeth; posterior edge of maxilla reaching anterior edge of orbit; opening of posterior nostrils vertically ovoid; opening of anterior nostrils with membranous flap. Distal tip of pectoral fin may or may not surpass pelvic–fin insertion; distal tip of pelvic fin reaching anal–fin origin. Premaxilla with two rows of teeth. Fourth to five teeth of outer row tricuspid, and arranged in zigzag. Internal row with four heptacuspid teeth. Maxilla long, reaching posterior margin of second infraorbital, anterior margin of maxilla with notches, with two–three teeth. Most anterior pentacuspid and two following tricuspid with central cusps widest; teeth internally inclined. Dentary with five large front heptacuspid teeth with central cusp largest, followed by seven to nine small teeth, anterior–most tricuspid and subsequent teeth unicuspid. Foramina on anterior ventral surface of sphenotic channel absent. Cleithrum with pointed dorsal process that surpasses transversally supracleithrum. Lateral line complete, perforated scales 38–41 (41*, mean = 38.75). Scale rows between dorsal–fin origin and lateral line 6; scale rows between lateral line and anal–fin origin 6–8 (6* mean = 6.25); scale rows between lateral line and pelvic–fin insertion 6–8, (6*, mean = 7.29). Predorsal scales arranged in regular series. Dorsal–fin rays iii, 9; second unbranched ray approximately one–half length of third unbranched ray.

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Fig. 6. Bryconamericus oroensis n. sp., holotype, MEPN 11140, male 74.51 mm SL, El Oro state, Ortega, 1.5 km NE of Zamora. Fig. 6. Bryconamericus oroensis sp. n., holotipo, MEPN 11140, macho 74,51 mm de longitud estándar, Estado de El Oro, Ortega, a 1,5 km al NE de Zamora.

1 cm B

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Fig. 7. Sexual dimorphism of Bryconamericus oroensis n. sp., paratypes: A. Male; B. Female. Fig. 7. Dimorfismo sexual de Bryconamericus oroensis sp. n., paratipos: A. Macho; B. Hembra.

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Dorsal–fin origin located near middle of body and posterior to vertical through pelvic–fin origin. Anal–fin rays iii–iv, 24–28 (iv, 24* n = 124). Anal–fin origin posterior to vertical through base of first dorsal–fin ray. Pectoral– fin rays ii, 10 (n = 124). Pelvic–fin rays ii, 6 (n = 124). Pelvic–fin origin anterior to vertical through dorsal–fin origin. Pelvic fin short, not reaching origin of the anal fin. Caudal fin not scaled, forked with short pointed lobes, principal caudal rays 1/17/1 with 12/11 procurrents. Total number of vertebra 38–39. Secondary sexual dimorphism Sexually mature males have row of spines on first to tenth branched anal–fin rays, no spines on simple rays, each ray with 5–11 spines, located on basal. Middle and posterior–most branches,11–23 large spines on all branched rays of pelvic fin, located on both branches of rays extending on to anterior–most part. Males with lateral stripe darker and more prominent than females. Males with caudal peduncle more swollen and caudal–fin rays thicker than females (fig. 5). B. oroensis shows sexual dichromatism, in males, related to higher concentration of melanophores along posterior margins of scales that, strongly concentrated along the lateral band and also along infraorbital, opercular and lateral surface of the cranium, whereas females have a lighter pigmentation pattern, with the humeral and peduncular regions pigmented. Color in alcohol Dorsum dark brown. Body with dark lateral stripe from posterior edge of opercle to base of caudal fin on male. More diffuse in females, especially anterior to level of dorsal fin. Humeral spot dark, round with faint ventral projection and followed by second transverse humeral spot; dark lateral band continued onto middle caudal–fin rays. Pectoral, pelvic and anal fins hyaline, but dorsal and caudal fin rays with dark bands. Ventro lateral region of body between snout tip and caudal peduncle yellow. Distribution This species is so far known from the Negro, Amarillo. Catamayo and Macará Rivers, Amazon basin; in the states of Loja, El Oro and Azuay, Ecuador. Etymology Bryconamericus oroensis is named after the El Oro province, where the type series was collected. Discussion The taxonomic validity of Bryconamericus and Knodus has been under discussion since Schultz (1944), who proposed Knodus was a modern synonym of Bryconamericus. Román–Valencia (2000, 2003a, 2005), Román–Valencia et al. (2008b) as part of a revision of the genus Bryconamericus, and Taphorn (1992) maintained the proposal of Schultz (1944) largely because the traditional character used to separate the two genera (caudal fin scaled or not) is not reliable. However, others authors Géry (1977), Lima et al. (2004), Weitzman et al.

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(2005), Ferreira & Lima (2006), Zarske & Géry (2006), Zarske (2008), Ferreira & Carvajal (2007), and Ferreira & Netto–Ferreira (2010) did not accept the position of Schultz (1944), Román–Valencia (2000, 2003a, 2005) or Taphorn (1992). Instead, they treated Knodus as valid as diagnosed by Eigenmann (1927), but they did not providea solution to the taxonomic problems this causes. Weitzman et al. (2005) ascertained that a large part of the phylogenetic relationships of clade A of Characidae depends upon a better phylogenetic knowledge of these two genera. Our recent observations confirm that the caudal scalation of Knodus species is an informative character of taxonomic and phylogenetic usefulness. This was already observed by several other authors (Ferreira & Carvajal, 2007; Ferreira & Netto–Ferreira, 2010).We therefore recognize the genus Knodus as valid, and relocate species we treated previously as Bryconamericus (see comparative material). As we now understand it, Knodus differs in the type of caudal scalation. In Bryconamericus, there are one or two larger, rounded scales located at the base of the caudal lobes. Furthermore, scalation does not extend beyond one–third of the length of the caudal–fin rays, and when well preserved, scales do not cover the procurrent caudal–fin rays. In Knodus, the caudal scales are smaller and sometimes horizontally elongated, and they cover at least two–thirds of the length of the caudal–fin rays as well as the procurrent caudal–fin rays. In addition, males of Knodus do not have a thickening of the interradial tissue of the anterior portion of the anal fin, as is observed in males of Bryconamericus. The monophyly of Bryconamericus has been discussed by Vari & Siebert (1990). Malabarba & Malabarba (1994), Silva & Malabarba (1996). Malabarba & Weitzman (2003) and Silva (2004). Calcagnotto et al. (2005) hypothesized that Bryconamericus should be classified within clade A (Malabarba & Weitzman, 2003), but they later concluded that the relationships between Bryconamericus and Knodus within clade A are not resolved and that neither genus is part of Stevardiinae: 'genera considered to be non–inseminating Clade A characids' (Menezes & Weitzman, 2009). Although Mirande (2010) suggests characters and proposes that Glandulocaudinae and Stevardiinae are members of one same clade, characters described by Mirande (2010) were not found in the species of Bryconamericus described herein: the epiphyseal branch of the epiphyseal bar opens just lateral to the cranial fontanel in B. oroensis n. sp. (vs. in most members of the Stevardiinae and some other species this branch of the canal system is instead completely absent), nine branched dorsal–fin rays in B. oroensis n. sp. and B. bucayensis n. sp. (vs. eight branched dorsal–fin rays); ten pterygiophores in the dorsal fin of B. oroensis n. sp. and B. bucayensis n. sp. (vs. nine dorsal–fin pterygiophores),three simple anterior dorsal–fin rays, the first two articulated with the first dorsal–fin pterygiophores in B. oroensis n. sp. (vs. two simple dorsal–fin rays, each articulated respectively with the first two dorsal–fin pterygiophores. (except in B. dahli: observation personal); folds present along the posterior margin of the pterosphenoid in B. oroensis n. sp. and B. zamorensis n. sp. (vs. foramina present in posterior margin of pterosphenoid); this last character

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Identification key to the species of Bryconamericus from Ecuador Coastal region (Pacific Basin). Clave de identificación de las especies de Bryconamericus de la región costera de Ecuador (cuenca del Pacífico).

1 Humeral spot absent Humeral spot present

B. brevirostris 2

2 41–49 lateral–line scales; 33–38 branched anal–fin rays

B. bucayensis n. sp.

27–40 lateral–line scales; 22–32 branched anal–fin rays 3 Dark caudal peduncle spot absent

3 B. simus

Dark caudal peduncle spot present

B. dahli

Identification key to the species of Bryconamericus from Rios Morona–Santiago, Napo and Conambo Marañon (Amazon Basin). Clave de identificación de las especies de Bryconamericus de los ríos Morona–Santiago, Napo Conambo Marañon (cuenca del Amazonas).

1 Small scales present on both caudal–fin lobes for at least one third of their length. Caudal peduncle spot not conspicuous, that is, not distinguishable from band through middle caudal–fin rays; first procurrent caudal–fin rays visible

B. pachacuti

Caudal–fin lobes without small scales, or present only near base, not covering one third of their length. Caudal peduncle spot conspicuous as separate from band through middle caudal–fin rays; first procurrent caudal–fin rays not visible

2 Lateral line with 37–41 pored scales. Male’s pelvic–fin rays with long, thin spines

Lateral line with 33–37 pored scales. Male’s pelvic–fin rays without long thin spines 3 Dark, arrowhead shaped caudal peduncle spot present, sometimes with anterior portion slightly widened; dark pigment of caudal peduncle spot extending anterior along sides of body only as diffuse melanophores. Males with intense dark lateral stripe extending from posterior margin of humeral spot to caudal peduncle and extending onto middle caudal–fin rays No arrowhead shaped caudal peduncle spot. Males without intense dark lateral stripe

belongs to alternative state 43 (0) Mirande (2010, in fig. 10) and relationships with the developed laminar on the area half of pterosfenoids. However, although they do not have all the characters proposed as diagnostic of Stevardiinae, males of the species of Bryconamericus examined in this study showed notable thickening of the interradial membrane tissue of the anterior part of the anal fin. This coincides with observations for other members of clade A such as Bryconadenos, that 'have

B. zamorensis n. sp.

B. oroensis n. sp. B. phoenicopterus

glandular club cells at the surface of the epidermis on the anterior part of the anal fin' (Weitzman et al., 2005). The relationship of Bryconamericus with other members of clade clado A is evident (Malabarba & Weitzman, 2003) but its relationship with Knodus is not yet resolved. Although the dorsal–fin formula of ii, 8 is not present in B. bucayensis and B. oroensis these two species do share a synapomorphy: the dorsal tip of the cleithrum surpasses the supracleithrum and contacts the pore

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of the first scale of the lateral line (vs. cleithrum not surpassing supracleithrum) (fig. 5), perhaps indicating that a separate lineage has developed in transandean Ecuador. Genus Bryconamericus is shown to have high species diversity in both trans and cis–andean drainages, whereas Knodus is only known from cis–andean river systems, and none of its species are reported from interior Andean drainages such as the Magdalena–Cauca Basin, the Pacific coastal drainages of Colombia, Ecuador and Perú. We note that no other Trans Andean characid species has the caudal and anal scalation pattern observed in Knodus, which is common in many Cisandean species of the genera Moenkhausia, Hemigrammus and Tetragonopterus. Acknowledgements We extend our sincere thanks to the University of Quindío, Vicerrectoria de Investigaciones, for financial assistance to carry out this study (to C. R.–V.). We thank James Maclaine and Harry Taylor (BMNH) for generously providing photographs of type material. To Jonathan W. Armbruster (AUM), David Catania (CAS), Soraya Barrera, Jaime Sarmiento (CBF), Ramiro Barriga (MEPN), Hernàn Ortega (MUSM) and Mabel Maldonado (UMSS) for the loan of specimens. We thank Marcos Mirande who read the manuscript and gave many valuable suggestions. References Armbruster, J. W., 2012. Standardized measurements, landmarks, and meristic counts for cypriniform fishes. Zootaxa, 3586: 8–16. Böhlke, E. J., 1958. Studies of the family Characidae. A report on several extensive recent collections from Ecuador. Proceeding Philosophy Academy Sciences, CX: 1–121. Braga, L., 2000. Redescription of Bryconamericus rubropictus (Berg) n. comb. (Ostariophysi, Characidae) and reference to its secondary sexual dimorphism. Revista Museo Argentino de Ciencias Naturales n.s., 2: 145–150. Calcagnotto, D., Schaefer S. A. & DeSalle, R., 2005. Relationships among characiform fishes inferred from analysis of nuclear and mitochondrial gene sequences. Molecular Phylogenetics and Evolution, 36: 135–153. Eigenmann, C. H., 1927. The American Characidae. Memoirs of the Museum of Comparative Zoology, 43: 311–428. Ferreria, K. M. & Carvajal, F. M., 2007. Knodus shinahota (Characiformes: Characidae) a new species from the río Shinahota, río Chapare basin (Mamoré system), Bolivia. Neotropical Icthyology, 5: 31–36. Ferreira, K. M. & Lima, F. C. T., 2006. A new species of Knodus (Characiformes: Characidae) from the Rio Tiquié upper Rio Negro system, Brazil. Copeia, 4: 630–639. Ferreira, K. M. & Netto–Ferreira, A. L., 2010. Knodus

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El primer autor rebrà 50 separates del treball sense càrrec a més d'una separata electrònica en format PDF. Manuscrits Els treballs seran presentats en format DIN A–4 (30 línies de 70 espais cada una) a doble espai i amb totes les pàgines numerades. Els manuscrits han de ser complets, amb taules i figures. No s'han d'enviar les figures originals fins que l'article no hagi estat acceptat. El text es podrà redactar en anglès, castellà o català. Se suggereix als autors que enviïn els seus treballs en anglès. La revista els ofereix, sense cap càrrec, un servei de correcció per part d'una persona especialitzada en revistes científiques. En tots els casos, els textos hauran de ser redactats correctament i amb un llenguatge clar i concís. La redacció del text serà impersonal, i s'evitarà sempre la primera persona. Els caràcters cursius s’empraran per als noms científics de gèneres i d’espècies i per als neologis mes intraduïbles; les cites textuals, independentment de la llengua, seran consignades en lletra rodona i entre cometes i els noms d’autor que segueixin un tàxon aniran en rodona. Quan se citi una espècie per primera vegada en el text, es ressenyarà, sempre que sigui possible, el seu nom comú. Els topònims s’escriuran o bé en la forma original o bé en la llengua en què estigui escrit el treball, seguint sempre el mateix criteri. Els nombres de l’u al nou, sempre que estiguin en el text, s’escriuran amb lletres, excepte quan precedeixin una unitat de mesura. Els nombres més grans s'escriuran amb xifres excepte quan comencin una frase. Les dates s’indicaran de la forma següent: 28 VI 99 (un únic dia); 28, 30 VI 99 (dies 28 i 30); 28–30 VI 99 (dies 28 a 30). S’evitaran sempre les notes a peu de pàgina. Format dels articles Títol. Serà concís, però suficientment indicador del contingut. Els títols amb designacions de sèries numèriques (I, II, III, etc.) seran acceptats previ acord amb l'editor. Nom de l’autor o els autors Abstract en anglès que no ultrapassi les 12 línies mecanografiades (860 espais) i que mostri l’essència del manuscrit (introducció, material, mètodes, resultats i discussió). S'evitaran les especulacions i les cites bibliogràfiques. Estarà encapçalat pel títol del treball en cursiva. Key words en anglès (sis com a màxim), que orientin sobre el contingut del treball en ordre d’importància. Resumen en castellà, traducció de l'Abstract. De la traducció se'n farà càrrec la revista per a aquells autors que no siguin castellanoparlants. Palabras clave en castellà.

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

studies to conservation biology: a review. Animal Biodiversity and Conservation. La relació de referències bibliogràfiques d’un tre ball serà establerta i s’ordenarà alfabèticament per autors i cronològicament per a un mateix autor, afegint les lletres a, b, c,... als treballs del mateix any. En el text, s’indicaran en la forma usual: "... segons Wemmer (1998)...", "...ha estat definit per Robinson & Redford (1991)...", "...les prospeccions realitzades (Begon et al., 1999)...". Taules. Es numeraran 1, 2, 3, etc. i han de ser sempre ressenyades en el text. Les taules grans seran més estretes i llargues que amples i curtes ja que s'han d'encaixar en l'amplada de la caixa de la revista. Figures. Tota classe d’il·lustracions (gràfics, figures o fotografies) entraran amb el nom de figura i es numeraran 1, 2, 3, etc. i han de ser sempre ressen yades en el text. Es podran incloure fotografies si són imprescindibles. Si les fotografies són en color, el cost de la seva publicació anirà a càrrec dels au tors. La mida màxima de les figures és de 15,5 cm d'amplada per 24 cm d'alçada. S'evitaran les figures tridimensionals. Tant els mapes com els dibuixos han d'incloure l'escala. Els ombreigs preferibles són blanc, negre o trama. S'evitaran els punteigs ja que no es reprodueixen bé. Peus de figura i capçaleres de taula. Seran clars, concisos i bilingües en la llengua de l’article i en anglès. Els títols dels apartats generals de l’article (Intro ducción, Material y métodos, Resultados, Discusión, Conclusiones, Agradecimientos y Referencias) no aniran numerats. No es poden utilitzar més de tres nivells de títols. Els autors procuraran que els seus treballs originals no passin de 20 pàgines (incloent–hi figures i taules). Si a l'article es descriuen nous tàxons, caldrà que els tipus estiguin dipositats en una institució pública. Es recomana als autors la consulta de fascicles recents de la revista per tenir en compte les seves normes.

Animal Biodiversity and Conservation 36.1 (2013)

Animal Biodiversity and Conservation Animal Biodiversity and Conservation es una revista interdisciplinar, publicada desde 1958 por el Museo Ciencias Naturales de Barcelona. Incluye artículos de investigación empírica y teórica en todas las áreas de la zoología (sistemática, taxonomía, morfología, biogeografía, ecología, etología, fisiolo gía y genética) procedentes de todas las regiones del mundo, con especial énfasis en los estudios que de una manera u otra tengan relevancia en la biología de la conservación. La revista no publica compila ciones 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 siem pre que es posible. La publicación de los trabajos aceptados se realiza con la mayor rapidez posible, normalmente dentro de los 12 meses siguientes a la recepción del trabajo. Una vez aceptado, el trabajo pasará a ser propie dad de la revista. Ésta se reserva los derechos de autor, y ninguna parte del trabajo podrá ser reprodu cida sin citar su procedencia.

Normas de publicación Los trabajos se enviarán preferentemente de forma electrónica (abc@bcn.cat). El formato preferido es un documento Rich Text Format (RTF) o DOC, que incluya las figuras y las tablas. Las figuras deberán enviarse también en archivos separados en formato TIFF, EPS o JPEG. Si se opta por la versión impresa, deberán remitirse cuatro copias juntamente con una copia en disquete a la Secretaría de Redacción. Debe incluirse, con el artículo, una carta donde conste que el trabajo versa sobre investigaciones originales no publicadas anteriormente y que se somete en ex clusiva 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 necesa rios y que han cumplido la normativa de protección animal vigente. Los autores pueden enviar también sugerencias para asesores. Cuando el trabajo sea aceptado los autores de berán enviar a la Redacción una copia impresa de la versión final junto con un disquete del manuscrito preparado con un procesador de textos e indicando

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el programa utilizado (preferiblemente Word). Las pruebas de imprenta enviadas a los autores deberán remitirse corregidas al Consejo Editor en el plazo máximo de 10 días. Los gastos debidos a modifica ciones sustanciales en las pruebas de imprenta, intro ducidas por los autores, irán a cargo de los mismos. El primer autor recibirá 50 separatas del trabajo sin cargo alguno y una copia electrónica en for mato PDF. Manuscritos Los trabajos se presentarán en formato DIN A–4 (30 lí neas de 70 espacios cada una) a doble espacio y con las páginas numeradas. Los manuscritos deben estar completos, con tablas y figuras. No enviar las figuras originales hasta que el artículo haya sido aceptado. El texto podrá redactarse en inglés, castellano o catalán. Se sugiere a los autores que envíen sus trabajos en inglés. La revista ofrece, sin cargo ningu no, un servicio de corrección por parte de una persona especializada en revistas científicas. En cualquier caso debe presentarse siempre de forma correcta y con un lenguaje claro y conciso. La redacción del texto deberá ser impersonal, evitándose siempre la primera persona. Los caracteres en cursiva se utilizarán para los nombres científicos de géneros y especies y para los neologismos que no tengan traducción; las citas textuales, independientemente de la lengua en que estén, irán en letra redonda y entre comillas; el nombre del autor que sigue a un taxón se escribirá también en redonda. Al citar por primera vez una especie en el trabajo, deberá especificarse siempre que sea posible su nombre común. Los topónimos se escribirán bien en su forma original o bien en la lengua en que esté redactado el trabajo, siguiendo el mismo criterio a lo largo de todo el artículo. Los números del uno al nueve se escribirán con letras, a excepción de cuando precedan una unidad de medida. Los números mayores de nueve se escribirán con cifras excepto al empezar una frase. Las fechas se indicarán de la siguiente forma: 28 VI 99 (un único día); 28, 30 VI 99 (días 28 y 30); 28–30 VI 99 (días 28 al 30). Se evitarán siempre las notas a pie de página. Formato de los artículos Título. Será conciso pero suficientemente explicativo del contenido del trabajo. Los títulos con designacio nes de series numéricas (I, II, III, etc.) serán aceptados excepcionalmente previo consentimiento del editor. Nombre del autor o autores Abstract en inglés de 12 líneas mecanografiadas (860 espacios como máximo) y que exprese la esen cia del manuscrito (introducción, material, métodos, resultados y discusión). Se evitarán las especulacio nes y las citas bibliográficas. Irá encabezado por el título del trabajo en cursiva.

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

Merilä, J., 1996. Genetic and quantitative trait vari ation in natural bird populations. Tesis doctoral, Uppsala University. * Los trabajos en prensa sólo se citarán si han sido aceptados para su publicación: Ripoll, M. (in press). The relevance of population studies to conservation biology: a review. Animal Biodiversity and Conservation. Las referencias se ordenarán alfabéticamente por autores, cronológicamente para un mismo autor y con las letras a, b, c,... para los trabajos de un mismo autor y año. En el texto las referencias bibliográficas se indicarán en la forma usual: "... según Wemmer (1998)...", "...ha sido definido por Robinson & Redford (1991)...", "...las prospecciones realizadas (Begon et al., 1999)...". Tablas. Se numerarán 1, 2, 3, etc. y se reseñarán todas en el texto. Las tablas grandes deben ser más estrechas y largas que anchas y cortas ya que deben ajustarse a la caja de la revista. Figuras. Toda clase de ilustraciones (gráficas, figuras o fotografías) se considerarán figuras, se numerarán 1, 2, 3, etc. y se citarán todas en el texto. Pueden incluirse fotografías si son imprescindibles. Si las fotografías son en color, el coste de su publicación irá a cargo de los autores. El tamaño máximo de las figuras es de 15,5 cm de ancho y 24 cm de alto. Deben evitarse las figuras tridimensionales. Tanto los mapas como los dibujos deben incluir la escala. Los sombreados preferibles son blanco, negro o trama. Deben evitarse los punteados ya que no se reproducen bien. Pies de figura y cabeceras de tabla. Serán claros, concisos y bilingües en castellano e inglés. Los títulos de los apartados generales del artículo (Introducción, Material y métodos, Resultados, Dis cusión, Agradecimientos y Referencias) no se nume rarán. No utilizar más de tres niveles de títulos. Los autores procurarán que sus trabajos originales no excedan las 20 páginas incluidas figuras y tablas. Si en el artículo se describen nuevos taxones, es imprescindible que los tipos estén depositados en alguna institución pública. Se recomienda a los autores la consulta de fascículos recientes de la revista para seguir sus directrices.

Animal Biodiversity and Conservation 36.1 (2013)

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, Ethology, Physiology and Genetics). It gives special emphasis to studies related to Conservation Biology. The journal does not publish bibliographic compila tions, listings, catalogues or collections of species, or isolated descriptions of a single specimen. Stud ies 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 assuring world–wide access to articles published therein. All manuscripts are screened by the Executive Editor, an Editor and two independent reviewers so as to guarantee the quality of the papers. The review process aims to be rapid and constructive. Once ac cepted, papers are published as soon as is practicable. This is usually within 12 months of initial submission. Upon acceptance, manuscripts become the pro perty of the journal, which reserves copyright, and no published material may be reproduced or cited without acknowledging the source of information.

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

Information for authors Electronic submission of papers is encouraged (abc@bcn.cat). The preferred format is DOC or RTF. All figures must be readable by Word, embedded at the end of the manuscript and submitted together in a separate attachment in a TIFF, EPS or JPEG file. Tables should be placed at the end of the docu ment. If a printed version is sent, four copies should be forwarded to the Editorial Office, together with a copy on computer disc. A cover letter stating that the article reports original research that has not been published elsewhere and has been submitted exclusively for consideration in Animal Biodiversity and Conservation is also necessary. When animal manipulation has been necessary, the cover letter should also specify that the authors follow current norms on the protection of animal species and that they have obtained all relevant permits and authorisa tions. Authors may suggest referees for their papers. Once an article has been accepted, authors should send a paper copy and an electronic copy of the final version. Please identify software (preferably Word). Proofs sent to the authors for correction should be returned to the Editorial Board within 10 days. Expenses due to any substantial alterations of the proofs will be charged to the authors. The first author will receive 50 reprints free of charge and an electronic version of the article in PDF format. ISSN paper: 1578–665 X ISSN digital: 2014–928 X

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. © 2013 Museu de Ciències Naturals de Barcelona

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

logical order for each author, adding the letters a, b, c,... to papers of the same year. Bibliographic citations in the text must appear in the usual way: "...according to Wemmer (1998)...", "...has been defined by Robinson & Redford (1991)...", "...the prospections that have been carried out (Begon et al., 1999)..." Tables. Must be numbered in Arabic numerals with reference in the text. Large tables should be narrow (across the page) and long (down the page) rather than wide and short, so that they can be fitted into the column width of the journal. Figures. All illustrations (graphs, drawings, photo graphs) should be termed as figures, and numbered consecutively in Arabic numerals (1, 2, 3, etc.) with reference in the text. Glossy print photographs, if essential, may be included. The Journal will publish colour photographs but the author will be charged for the cost. Figures have a maximum size of 15.5 cm wide by 24 cm long. Figures should not be tridimen sional. Any maps or drawings should include a scale. Shadings should be kept to a minimum and preferably with black, white or bold hatching. Stippling should be avoided as it may be lost in reproduction. Legends of tables and figures. Legends of tables and figures should be clear, concise, and written both in English and Spanish. Main headings (Introduction, Material and methods, Results, Discussion, Acknowledgements and Referen ces) should not be numbered. Do not use more than three levels of headings. Manuscripts should not exceed 20 pages including figures and tables. If the article describes new taxa, type material must be deposited in a public institution. Authors are advised to consult recent issues of the journal and follow its conventions.

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Arizaga et al.

Animal Biodiversity and Conservation 36.1 (2013)

Welcome to the electronic version of Animal Biodiversity and Conservation

Re co se lec mme nd tro to nic yo ur ac ce lib ss rar y!

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www.abc.museucienciesjournals.cat

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

Arizaga et al.

Animal Biodiversity and Conservation 36.1 (2013)

Erratum En el volum 35.2 (2012) es van publicar els abstracts del XXXth IUGB Congress and Perdix XIII Conference. A la pàgina 151 vam cometre una omissió de la informació que havíem de presentar. Adjuntem aquí la pàgina tal com s'hauria d'haver publicat i demanen disculpes als editors per aquesta omissió. En el volumen 35.2 (2012) se publicaron los abstracts del XXXth IUGB Congress and Perdix XIII Conference. En la página 151 se omitió parte de la información que debía contener. Adjuntamos aquí la página tal y como debería haberse publicado y pedimos disculpas a los editores por esta omisión. In volume 35.2 (2012) we published the abstracts from the XXXth IUGB Congress and Perdix XII Conference. Part of the information on page 151 was erroneously omitted. We here attach this page as it should have appeared and apologise to the editors for this erratum.

XXXth IUGB Congress and Perdix XIII Conference Editors executius / Editores ejecutivos / Executive Editors Manel Puigcerver Univ. de Barcelona, Spain José Domingo Rodríguez Teijeiro Univ. de Barcelona, Spain Joan Carles Senar Museu de Ciències Naturals de Barcelona, Spain Secretària de redacció / Secretaria de redacción / Managing Editor Montserrat Ferrer Museu de Ciències Naturals de Barcelona, Spain Comité organitzador / Comité organizador / Organizing Comittee Francis Buner Game and Wildlife Conservation Trust, UK Chair of Perdix XIII Manel Puigcerver Univ. de Barcelona, Spain Chair of XXXth IUGB Congress Nicholas Aebischer Game and Wildlife Conservation Trust, UK Antonio Bea Ekos estudios ambientales,Spain. Ricard Casanovas Generalitat de Catalunya, Spain Jorge Cassinello Inst. de Investigación en Recursos Cinegéticos, CSIC, Spain Miguel Delibes Estación Biológica de Doñana, CSIC, Spain Xavier Ferrer Univ. de Barcelona, Spain Francesc Llimona Consorci del Parc de Collserola, Spain José María López Generalitat de Catalunya, Spain Santi Mañosa Univ. de Barcelona, Spain Francesc Piera Federación Catalana de Caza, Spain José Domingo Rodríguez Teijeiro Univ. de Barcelona, Spain Carme Rosell Minuartia, Spain Jordi Ruiz Generalitat de Catalunya, Spain Francesc Sardà Centre Tecnològico Forestal de Catalunya, Spain José Mari Usarraga Federación de Caza de Euskadi, Spain Javier Viñuela Inst. de Investigación en Recursos Cinegéticos, CSIC, Spain Sessions plenàries / Sesiones plenarias / Plenary Sessions Veterinary aspects of wildlife and conservation: Peter D. Walsh Species extinctions and population dynamics: Philip K. J. McGowan Wildlife law and policy: Borja Heredia Conservation and management of migratory species: Manel Puigcerver Wildlife biology, behaviour and game species management: Nicholas Aebischer Interactions humans–wildlife: Steve Redpath Methodologies, models and techniques: Lisette Waits Human dimensions of game wildlife management: John Linnell ISSN: 1578–665X

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Arxius de Miscel·lània Zoològica vol. 10 (2012) Museu de Ciències Naturals de Barcelona ISSN: 1698–0476 www.amz.museucienciesjournals.cat

Índex / Índice / Contents Fontana–Bria, L. & Selfa, J., 2012. Revisió dels odonats valencians de la col·lecció d’artròpodes del Museu de Ciències Naturals de Barcelona. Arxius de Miscel·lània Zoològica, 10: 1–8. Abstract Revision of valencian dragonflies in the arthropod collection at the Museu de Ciències Naturals de Barcelona.— The current work shows the species of valencian dragonflies preserved in the arthropod collection at the Museu de Ciències Naturals de Barcelona (Spain). A total of 33 specimens have been found belonging to 12 species, which represent the 18% of the known species in the Valencian Country. Key words: Odonata, Revision, Valencian Country, Museu de Ciències Naturals de Barcelona, Spain. Jawad, L. A., Al–Shogebai, S. & Al–Mamry, J. M., 2012. First record of Atractoscion aequidens (Sciaenidae) from the Arabian Sea Coasts of Oman and Acanthopagrus catenula (Sparidae) from the Oman Sea (Gulf of Oman), northwestern Indian Ocean (Teleostei, Sciaenidae, Sparidae). Arxius de Miscel·lània Zoològica, 10: 9–15. Abstract First record of Atractoscion aequidens (Sciaenidae) from the Arabian Sea Coasts of Oman and Acanthopagrus catenula (Sparidae) from the Oman Sea (Gulf of Oman), northwestern Indian Ocean (Teleostei, Sciaenidae, Sparidae).— The first record of Atractoscion aequidens from the Arabian Sea coasts of Oman and Acanthopagrus catenula from waters around City of Muscat on the Sea of Oman is reported based on one (671 mm in SL) and ten specimens (111–257 mm SL), respectively. This account represents the second record of A. aequidens in the northern Indian Ocean and the northernmost record of A. catenula in the same ocean. Morphometric and meristic data are provided for the two species and compared with those from specimens collected from other parts of the world. The southern distribution of A. catenula is corrected in this report. Key words: New record, New range extension, Sciaenidae, Sparidae, Oman Sea, Arabian Sea. Agulló, J. & Prieto, M., 2012. Nuevos datos sobre la distribución de Stenostoma rostratum (Fabricius, 1787) en Cataluña (nordeste de la península Ibérica) (Coleoptera, Oedemeridae). Arxius de Miscel·lània Zoològica, 10: 17–28. Abstract New data on the distribution of Stenostoma rostratum (Fabricius, 1787) in Catalonia (northeastern Iberian peninsula) (Coleoptera, Oedemeridae).— Distribution of Stenostoma rostratum (Fabricius, 1787) in Catalonia is updated. New data were obtained from recent prospections at two natural reserves of the Mediterranean coast, the Espais Naturals del Delta de Llobregat (province of Barcelona) and the Parc Natural de les Illes Medes, Massís del Montgrí i Plana del Baix Ter (province of Girona). The species is found exclusively in sand dunes, ocupping a reduced area in the catalonian coast. Stenostoma rostratum is considered a threatened species in Catalonia, the last records published up to now being obtained 40 years ago. Some remarks about its habitat, ecology and conservation in both protected areas are also provided. Key words: Stenostoma rostratum, Oedemeridae, Coleoptera, Catalonia, Distribution, Sand dunes, Conservation.

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Animal Biodiversity and Conservation 36.1 (2013)

Núñez, R., 2012. The butterflies of Turquino National Park, Sierra Maestra, Cuba (Lepidoptera, Papilionoidea). Arxius de Miscel·lània Zoològica, 10: 29–49. Abstract The butterflies of Turquino National Park, Sierra Maestra, Cuba (Lepidoptera, Papilionoidea).— Between February and November 2011, we conducted a species inventory, created a natural history database and a made a first approach to the composition and structure of the butterfly communities present at several vegetation types in the Turquino National Park. The inventory included 83 species, 29 of them endemic. We recorded 57 species (18 endemic) in transects along main vegetation pathways. In disturbed vegetation, species richness was higher (48) and abundance was better distributed, but the proportion of endemism was lower (23%). Species richness decreased and the dominance and proportion of endemism increased with altitude. Numbers of species and the proportions of endemism at natural habitats sampled were: 19 and 58% for evergreen forest, 10 and 60% for rainforest, eight and 100% for cloud forest, and four and 100% for the elfin thicket. Flowers of 27 plants were recorded as nectar sources for 30 butterfly species, and host plants were recorded for nine species. Key words: Communities, Conservation, Natural history, Endemism, Caribbean. Fernández, A. M., Lloris, D., Pérez Gil, J. L. & Esteban, A., 2012. On the occurrence of Zenopsis conchifer (Lowe, 1852) (Osteichthyes, Zeidae) in the Mediterranean Sea. Arxius de Miscel·lània Zoològica, 10: 50–54. Abstract On the occurrence of Zenopsis conchifer (Lowe, 1852) (Osteichthyes, Zeidae) in the Mediterranean Sea.— The capture of four specimens of Silvery John Dory (Zenopsis conchifer), a species recorded in the Mediterranean Sea for the first time in 2006, is reported from the Iberian coast (western Mediterranean). One of the specimens was caught near the Strait of Gibraltar and is probably a vagrant. Despite these catches, there is no evidence of a self–sustaining population, so this species should be considered as alien in the Mediterranean. Key words: Zenopsis conchifer, Fish invasions, Exotic species, Mediterranean. Godoy, M. C., Laffont, E. R., Coronel, J. M. & Etcheverry, C., 2012. Termite (Insecta, Isoptera) assemblage of a gallery forest relic from the Chaco province (Argentina): taxonomic and functional groups. Arxius de Miscel·lània Zoològica, 10: 55–67. Abstract Termite (Insecta, Isoptera) assemblage of a gallery forest relic from the Chaco province (Argentina): taxonomic and functional groups.— Termite fauna of the gallery forest in the Colonia Benitez Reserve (Chaco province, Argentina) were analyzed using the rapid diversity assessment protocol (100 x 2 m transects). Twelve species, 10 genera and two families (Kalotermitidae and Termitidae), were detected, comprising the four feeding groups recognized for termites. True soil–feeders (IV) showed the highest species richness, and dead wood and grasses feeders (II) had the highest relative abundance. The most frequently occupied microhabitats were dead wood pieces lying on the ground. These results indicate that the Reserve harbors a diverse termite community similar to the ‘monte fuerte’ isopteran fauna (91.6% shared species). Our findings also support the Reserve´s value as a well–preserved fragment of the original gallery forest and emphasize the need to promote its conservation. Key words: Termite community, Neotropical region, Feeding groups.

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Índex / Índice / Contents Animal Biodiversity and Conservation 36.1 (2013) ISSN 1578–665X 1–11 J. Arizaga, E. Unamuno, O. Clarabuch & A. Azkona The impact of an invasive exotic bush on the stopover ecology of migrant passerines 13–31 M. Kulfan, M. Holecová & P. Beracko Dalechampii oak (Quercus dalechampii Ten.), an important host plant for folivorous lepi� doptera larvae 33–36 A. Martínez–Abraín Why do ecologists aim to get positive results? Once again, negative results are necessary for better knowledge accumu� lation 37–46 J. Lozano, J. G. Casanovas, E. Virgós & J. M. Zorrilla The competitor release effect applied to carnivore species: how red foxes can increase in numbers when persecuted 47–57 L. Klemann Jr. & J. S. Vieira Assessing the extent of occurrence, area of occupancy, territory size, and population size of marsh tapaculo (Scytalopus iraiensis) 59–67 A. Martínez–Ortí & V. Borredà Drusia (Escutiella) alexantoni n. sp. �� (Gas� tropoda, Pulmonata, Parmacellidae), a new terrestrial slug from the Atlantic coast of Morocco

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69–78 J. Arizaga & I. Tamayo Connectivity patterns and key non–breeding areas of white–throated bluethroat (Luscinia svecica) European populations 79–88 M. Sarasa & J.–A. Sarasa Intensive monitoring suggests population oscillations and migration in wild boar Sus scrofa in the Pyrenees 89–99 H. V. Bogdan, S.–D. Covaciu–Marcov, O. Gaceu, A.–.S. Cicort–Lucaciu, S. Ferenţi & I. Sas–Kovács How do we share food? Feeding of four amphibian species from an aquatic habitat in south–western Romania 101–111 C. San Vicente A new species of Mysidopsis (Crustacea, Mysida, Mysidae) from coastal waters of Catalonia (North–western Mediterranean) 113–121 J. Rivera, E. Barba, A. Mestre, J. Rueda, M. Sasa, P. Vera & J. S. Monrós Effects of migratory status and habitat on the prevalence and intensity of infection by haemoparasites in passerines in eastern Spain 123–139 C. Román–Valencia, R. I. Ruiz–C., D. C. Taphorn B. & C. García–A. Three new species of Bryconamericus (Characiformes, Characidae), with keys for species from Ecuador and a discussion on the validity of the genus Knodus