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en J. Hatchwell, Univ. of Sheffield, UK

Dibuix de la coberta / Dibujo de la portada / Drawing of the cover: Vulpes vulpes, guilla/guineu, zorro común/zorro rojo, red fox (Jordi Domènech) Localització / Localización / Locality: Als voltants de Montserrat, en los alrededores de Montserrat, around Montserrat

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Animal Biodiversity and Conservation 42.2, 2019 © 2019 Museu de Ciències Naturals de Barcelona, Consorci format per l'Ajuntament de Barcelona i la Generalitat de Catalunya Autoedició: Montserrat Ferrer Fotomecànica i impressió: CEVAGRAF SCCL ISSN: 1578–665 X eISSN: 2014–928 X Dipòsit legal: B. 5357–2013

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

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Poaching in non–volant mammals in the Neotropical region: the importance of a metric to assess its impacts A. C. Ferreguetti, C. F. Duarte Rocha, H. Godoy Bergallo

Ferreguetti, A. C., Duarte Rocha, C. F., Godoy Bergallo, H., 2019. Poaching in non–volant mammals in the Neotropical region: the importance of a metric to assess its impacts. Animal Biodiversity and Conservation, 42.2: 203–211, Doi: https://doi.org/10.32800/abc.2019.42.0203 Abstract Poaching in non–volant mammals in the Neotropical region: the importance of a metric to assess its impacts. Much of the information on the hunting of mammals in natural environments is not performed in a standard way and is usually dispersed by different areas or regions that have different environmental structures. This limitation prevents the detection of trends and patterns such as which biomes are under more pressure and what are the rates and level of impact. We aimed to review the scientific literature on poaching of non–volant mammals to evaluate the impact at different study sites in the Neotropical region. We found that in more than half of these studies (66/112, 59 %), the main objectives were related to characterizing hunting activity while the potential impact of the hunting was not assessed. Evaluating the poaching through a metric assessment using qualitative and quantitative variables was the main objective in only 58 articles. We classified the hunting events as subsistence in most cases (46/58, 79 %), as illegal in a few case (12/58, 21 %) and as legal in one study only (1/58, 2 %). Based on this extensive review of scientific literature, we propose a metric assessment that can be performed in natural reserves and can lead to extensive monitoring on mammal populations through training on how to gauge this geo–referenced data. Key words: Conservation, Extinction, Hunting, Mammals, Standard monitoring Resumen La caza ilegal de mamíferos no voladores en la región neotropical: la importancia de evaluar sus repercusiones con un parámetro. Gran parte de la información sobre la caza de mamíferos en ambientes naturales no se recaba de forma estandarizada y generalmente se dispersa en zonas o regiones distintas que tienen estructuras ambientales diferentes. Esta limitación impide la detección de tendencias y pautas como las relacionadas con los biomas que padecen más presión o los índices y el grado de repercusiones. La finalidad de este trabajo es examinar las publicaciones científicas sobre la caza ilegal de mamíferos no voladores, con vistas a evaluar las repercusiones en diferentes sitios de estudio de la región neotropical. Encontramos que en más de la mitad de estos estudios (66/112; 59 %), los objetivos principales estaban relacionados con la caracterización de la actividad cinegética, pero no se evaluaban las posibles repercusiones de la caza. Solo 58 artículos tenían el propósito de evaluar la caza ilegal mediante una evaluación paramétrica utilizando variables cualitativas y cuantitativas. En el presente estudio clasificamos los episodios de caza como de subsistencia (46/58; 79 %), ilegales en unos pocos casos (12/58; 21 %) y legales en un único estudio (1/58; 2 %). Sobre la base de este amplio examen de las publicaciones científicas, proponemos una evaluación métrica que puede llevarse a cabo en reservas naturales y que permite hacer un seguimiento exhaustivo de las poblaciones de mamíferos gracias a la formación impartida sobre cómo analizar estos datos georreferenciados. Palabras clave: Conservación, Extinción, Caza, Mamíferos, Seguimiento estandarizado Received: 11 IV 18; Conditional acceptance: 07 IX 18; Final acceptance: 29 X 18 Atilla Colombo Ferreguetti, Carlos Frederico Duarte Rocha, Helena Godoy Bergallo, Department of Ecology, Rio de Janeiro State University, Rua São Francisco Xavier 524, Pavilhão Haroldo Lisboa da Cunha, 2º andar, sala 224, Bairro Maracanã, CEP 20550–013 Rio de Janeiro, RJ–Brazil. * Corresponding author: Atilla Colombo Ferreguetti. E–mail: atilla.ferreguetti@gmail.com ISSN: 1578–665 X eISSN: 2014–928 X

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Introduction Human activity has deeply changed most ecosystems in many regions of the world (Steffen et al., 2015), causing widespread loss of biodiversity (Vellend et al., 2007; Arroyo–Rodriguez et al., 2013; Newbold et al., 2015), changes in community structure (Dornelas et al., 2014), and loss of ecosystem functions and services (Mitchell et al., 2015). Tropical forests are one of the biomes most threatened by human activities, and each year about 13 million hectares of these forests around in the world have been devastated (Myers et al., 2000). Exploitation of plant and animal resources in a non–sustainable approach in the natural landscape have led to biodiversity loss, pollution, invasion of exotic species, local extinction of native species (Cardinale et al., 2012), deforestation and habitat fragmentation (Laurance and Bierregaard Jr., 1997; Laurance, 1999). Tourism, hunting, agriculture and livestock practices also affect biodiversity and the survival of species (Cullen et al., 2000). Loss of habitats and hunting of species are considered the main threats to the maintenance of non–volant mammal populations (Redford, 1992; Peres, 2001; Milner–Gulland and Bennett, 2003). Excessive removal of specimens from nature is a major threat to world fauna (Robinson and Redford, 1991; Bennett and Robinson, 2000a; Alves et al., 2012). Several studies show that hunting activities in the Neotropics are generally carried out in an uncontrolled manner, the impact of which makes populations unviable and natural resources unsustainable for ecosystem function (Hill and Padwe, 2000; Bodmer and Robinson, 2006; Fernandes–Ferreira et al., 2012). Much information on the hunting of mammals in natural environments is focuses on one or few species. In addition, this information is not standardized through a general protocol, and is dispersed from locations or regions with different environmental structures. This lack of standardization prevents the detection of trends and patterns concerning those biomes that are likely under highest pressure, and the quantification of the rate and level of the hunting impact. We performed a review based on the information published in scientific journals on hunting in non–volant mammals in the Neotropical region. This review of the literature aimed to evaluate the use of metric assessment and the impact of hunting at several study sites in the Neotropical region. We sought to answer the following questions: i) for which biome have most studies been performed to evaluate the impact of hunting on mammals?; (ii) how many studies have evaluated the events and classified illegal or subsistence hunting?; iii) which metric assessment was used to evaluate the hunting impact in each study?; iv) has the metric assessment used to test the impact of hunting produced a statistically significant result?; and (v) can hunting records help to build a metric assessment to monitor impact of hunting? Three electronic databases were used to search the scientific literature: ISI Web of Science, Google Scholar and Scielo. The search terms used were entered in the categories 'Title, abstract and keywords' and 'Topic' (TS). The search was based on seven sets

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of keywords, equally applied to the three databases. The main set referred to variations in hunting terms (impact studied) and included 'Hunt*' OR 'poach*' OR 'bushmeat'. The main set was crossed separately with five other sets referring to the object of the study (mammals) and locality (Neotropical Region) through the Boolean operator AND: ('mammal*') AND ('Neotropic*'). We restricted our search to articles published in three languages: English, Portuguese and Spanish. We considered only studies published from 1920 until 20 XII 2017, the date the search was conducted. Humans and hunting: contextualizing this interaction Wildlife has been a major resource for humans for the past six million years (Stanford and Bunn, 2001). Throughout our history, humans have interacted with wild mammal species of many different forms (Happold, 1995). Relationships thus vary according to different human cultures and are reflected in the negative or positive effects on the wild mammals involved (Leopold, 1959; Bodmer et al., 1997; Alves et al., 2009). Animals have been used over time for multiple purposes. They have not only provided food, but have also been used in the creation of artifacts, for transportation, as a source of beauty and inspiration, and as symbols of gods in religious rituals (Ripple and Perrine, 1999; Alves et al., 2012). Some species, such as felines, are hunted and killed because they represent risks to human life or domestic livestock, while others, such as rodents and some species of medium–sized mammals, pose a threat to crops (Treves et al., 2006; Mendonça et al., 2011; Macedo et al., 2015). This ambiguity in the interaction between human and animals is common in many cultures and depends on the species involved (Antonites and Odendaal, 2004; Alves et al., 2012; Alves and Souto, 2015). Indeed, in agricultural societies, hunting involves a dual relationship of familiarity and friendship with domestic animals, and hostility and aggression with the wild and mysterious world (Macedo et al., 2015). Hunting, especially in rural areas, tends to promote a rapprochement or rejection relationship with wild animals and tends to be transmitted over generations of human settlements in natural environments. Hunting in the Neotropical region The Neotropical region extends from Central America (including Tropical Mexico) to southern South America. This biogeographic region is characterized by significant biotic and climatic diversity (Morrone, 2014). It comprises 78 ecoregions formed predominantly by tropical and subtropical forests and open formations interrupted by rivers (Morrone, 2014). Hunting of wild animals occurs throughout the Neotropical region, being carried out by indigenous, rural, and urban populations (Becker, 1981; Cullen et al., 2000; Fernandes–Ferreira et al., 2012). Hunting can be considered a cultural trait that is strongly

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Number of studies published

45 40 35 30 25 20 15 10 0

Amazon Atlantic Neotropical Chaco rainforest rainforest forest

Chiquitana Caatinga forest

Fig. 1. Scientific publications that evaluated the impact of hunting on non–volant Neotropical mammals classified by types of environment. This figure was based on the 58 scientific publications evaluated. Fig. 1. Publicaciones científicas que analizaron las repercusiones de la caza de mamíferos no voladores en la región neotropical, clasificadas por tipo de ambiente. Este gráfico se basa en las 58 publicaciones científicas estudiadas.

rooted in the Neotropics; it involves several aspects, depending on the human community in question and the region considered. The considerable progress in living conditions recorded in the second half of the twentieth century resulted in unprecedented urbanization, as well as an improvement and dynamization of the productive processes of animal protein (meat) and its derivatives. Together with the advancement in the perception of values concerning the importance of preserving natural resources, society has begun intensive discussion on hunting. Many groups have advocated an unrestricted ban on hunting, especially sport hunting (Leopold, 1959; Collazos et al., 1960; Pierret and Dourojeanni, 1966). In the Neotropical region, hunting began to be studied at the beginning of the 20th century in order to characterize the activity with a cultural focus (Leopold 1959; Collazos et al., 1960; Pierret and Dourojeanni, 1966). However, it was not until the end of this century that studies began to focus on the hunting impact on wildlife (Bodmer et al., 1988; Paz y Miño, 1988; Peres, 1990). Of the 112 scientific articles reviewed, the main objective in more than half (66/112, 59 %) was to characterize the hunting activity only; the potential impact was not evaluated. Only 58 of the articles used a metric be it qualitative or quantitative –as the main objective to evaluate the hunting (table 1s in supplementary material). Of these, 38 studies published were carried out in the Amazon (about 70 %), followed by 10 studies in Neotropical Forest in general (17.3 %), eight studies in the Atlantic Forest (13.7 %), and only one study in the Bolivian Chaco and one in the Brazilian Semi–Arid region (1.7 %) (fig. 1).

The importance of hunting as a source of animal protein was evidenced in the first reports about the Amazon. In 1864, naturalist Henry Bates described hunts and the habit of local populations along the Amazon River to consume wild animals (Bates, 1864). Many studies on hunting among mestizo and indigenous populations have been carried out in the Amazon, especially since 1970. In that decade the availability of protein foods was already discussed as a limiting factor for human groups (Gross, 1975) as was the importance of hunting as a source of protein and fat for the Amazon populations (Ayres and Ayres, 1979). The hunting practiced by mestizo and indigenous populations of the Amazon was compared at the end of the 1980s, as biological factors such as density and abundance of species, and cultural factors, such as food and technical restrictions of hunting, were crucial to differentiate between these human groups (Redford and Robinson, 1987). In the 1990s, it was suggested that human population growth and settlement age (a supposed index of time to familiarize with the local environment and fauna) were associated with the negative effects of hunting on vertebrate fauna (Vickers, 1991; Redford, 1992). Since 2000, several aspects related to the sustainability of hunting in tropical forests have been studied (e.g. Bennett and Robinson, 2000a), although most of these studies have addressed subsistence hunting and few have addressed poaching (illegal hunting). Data from the available hunting studies classified the events as subsistence (46,78 %), while 12 (20 %) classified hunting as illegal and only one (2 %) as legal (supplementary material).

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Currently, Amazonian rural communities continue to hunt, although the commercial exploitation of wildlife has become an illegal activity in Brazil since 1967 under the Wildlife Protection Act. (Law No. 5,197, of February 3, 1967). According to this law, hunting was prohibited even for human populations that depended on wildlife for food. Only in 1998, with the advent of the Environmental Crimes Law (Law No. 9,605, of February 12, 1998), was subsistence hunting recognized as a non–criminal activity provided that it was carried out 'in a state of need to quench hunger of the agent or his family'. However, this law does not correspond to the reality in the Amazon region, where the barter of hunting products for primary necessities is characterized as commercial hunting and is therefore considered illegal (Caughley and Gunn, 1996).

species density, promoting alteration of community structure and overexploitation of resources by herbivores that previously had their populations controlled by these predators (Terborgh et al., 2001). In addition, human hunters often tend to hunt those species that top predators select as prey, such as ungulates and rodent species (Leite and Galvão, 2002), and this may reduce the capacity of a habitat to sustain populations of large carnivores. In the Neotropical region, primates, tapirs and carnivores are particularly vulnerable to overhunting due to their low intrinsic rates of natural growth, high longevity, long generation time, and low population densities (Bodmer et al., 1997; Cardillo et al., 2004). Populations of ungulates and large primates decline as soon as hunting becomes a chronic process (Peres, 2000b).

Mammals and hunting: impacts

How has the hunting impact been assessed in the Neotropical region?

Loss of habitat and overhunting of species are considered the main threats to the survival of many species of large forest vertebrates (Redford, 1992; Milner–Gulland and Bennett, 2003; Dirzo et al., 2014). Increased human density (Brook et al. 2006), the growth of access to new technologies (Vickers, 1991; Mena et al., 2000; Stearman, 2000), and the loss of traditional hunting practices (Leeuwenberg and Robinson, 2000; Mena et al., 2000; Stearman, 2000) have promoted the overhunting of populations of Neotropical mammals (Bennett and Robinson, 2000a, 2000b; Silvius et al., 2004). The overhunting of tropical forest vertebrates has led to the decline in population of many species (Bennett and Robinson, 2000b), causing extinctions of local and global species (Peres, 1990; Ulloa et al., 2004). Hunting can affect mammalian populations (Chiarello, 2000; Peres, 2000b; Crawshaw et al., 2004) and change communities (Peres, 1990, 2001; Naughton– Treves et al., 2003), but it tends to be underestimated (Redford, 1992) due to lack of standardization and difficulties in detection (Peres et al., 2006). This occurs both in areas where there is anthropogenic habitat disturbance (Daily et al., 2003; Naughton–Treves et al., 2003) and in areas with little or no forest change (Redford, 1992; Peres, 1996; Peres and Lake, 2003), including within protected areas (Chiarello, 2000; Altrichter and Almeida, 2002; Olmos et al., 2004). Most hunted species are frugivorous and/or herbivorous (Peres, 2000a, 2000b; Townsend, 2000), and they play an ecological role in the dynamics of natural environments (Dirzo and Miranda, 1991; Wright et al., 2000; Stoner et al., 2007). The overhunting of large forest vertebrates can compromise important ecological processes for the maintenance of forest structure and species composition (Dirzo and Miranda, 1991; Wright et al., 2000; Dirzo et al., 2014), reducing long–term biodiversity (Terborgh, 1992, 2000). Extirpation of species tends to compromise the ecosystem functionally and may result in the depletion of forest environments (Harrison, 2011). Population reduction of top–predators (e.g. Panthera onca and Puma concolor) due to systematic killing by hunting (Crawshaw et al., 2004) may result in increased prey

One of the most cited hypotheses in the field of Conservation Biology is undoubtedly Kent Redford's 'Empty Forest' (Redford, 1992). It has been proposed that we are moving towards a situation where extensive, seemingly intact forest areas present a series of ecological extinctions as a result of hunting and a supposed defaunation. Large species, especially mammals, could have such small populations that vital functions for the maintenance of ecosystems would be highly affected. In the long–term, therefore, the preservation of tropical forest vegetation would not be possible if the fauna were not also preserved (Redford, 1992). The question of 'empty forest' has also been evaluated considering the effects of hunting, showing the potential association between hunting and negative effects on the vegeation (Harrison, 2011). The species most appreciated by subsistence hunters are generally responsible for ecological interactions that directly influence plant regeneration (Dirzo, 2001; Wright et al., 2007; Terborgh et al., 2008). These interactions include predation of seeds before and after dispersion, primary and secondary seed dispersal, and leaf and grass herbivory (Wright et al., 2007). The consequences of deforestation from fauna hunting in forest dynamics include reductions in predation and dispersal of seeds, which may lead to changes in total recruitment of seedlings, composition, decreases in diversity of flora (Dirzo and Miranda, 1991; Terborgh et al., 2008; Dirzo et al., 2014), and even alterations of carbon stocks in tropical forests (Bello et al., 2015; Kurten et al., 2015). Many studies based on the ‘empty forest’ hypothesis qualitatively compared the impact of hunting on wildlife in areas without hunting or hunting classified at different intensities. Of these 58 studies evaluated, 39 used a qualitative approach to characterize hunting and assess the impact on mammals (fig. 2A). The methods used to characterize the impact of hunting used in 95 % of the studies were: 20 studies used hunting intensity classes (low, medium and high) by locality and 18 relied on presence/absence data (i.e. with and without hunting) (fig. 2A). However,

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Method applied Number of studies published

Significant result?

20 15 10 0

C Number of studies published

207

Intensity classes

Presence/ absence

Others

Slaughtered Evidences animals

Others

Yes

Partial

10 8 6 4 2 0

Yes

Fig. 2. Classification of scientific publications evaluating the impact of hunting on non–volant Neotropical mammals: A, metric estimated by the qualitative method; B, significance of the result found in each study that used the qualitative method; C, metric estimated by the quantitative method; and D, significance of the result found in each study that used the quantitative method. Fig. 2. Clasificación de las publicaciones científicas que analizan las repercusiones de la caza de mamíferos no voladores en la región neotropical: A, parámetro estimado por el método cualitativo; B, significación del resultado obtenido en cada estudio que utilizó el método cualitativo; C, parámetro estimado por el método cuantitativo; y D, significación del resultado obtenido en cada estudio que utilizó el método cuantitativo.

using a qualitative approach to evaluate the impact of hunting, almost half of the studies did not find statistically significant results (fig. 2B). Concomitantly, models were developed to quantitatively measure the sustainability of hunting in tropical areas, representing about 33 % (19/58) of the studies as shown in figure 2C (Robinson and Redford, 1991; Robinson and Bennett, 1999; Bodmer and Robinson, 2004). Of the 19 studies that assessed the impact of hunting quantitatively, 11 were for subsistence hunting in the Amazon Forest using the number of slaughtered animals as metric. This assessment is possible for subsistence hunting because the communities that practice hunting report the number of individuals that are extracted from nature. This metric cannot be applied to measure illegal hunting, however. Therefore, the eight studies that evaluated poaching used evidence of hunting as an indicator, but continued ranking the intensity of hunting. Some studies assume that the density of huntable species in non–hunting areas represents a precise estimate of the support capacity in a region, thus concluding the number of individuals an area could harbor (Caughley, 1977;

Caughley and Sinclair, 1994). All 19 studies using a quantitative metric found a statistically significant result on the impact of hunting on mammals (fig. 2D). The importance of a quantitative metric method to detect the poaching impact and long–term standardized monitoring As previously reported, most studies that evaluated the impact of hunting considered subsistence hunting. To quantify the impact of hunting, the number of animals slaughtered (fig. 2C) was used as a metric assessment in most studies. For subsistence hunting, this metric may indicate an estimate of how species are being affected (Aquino and Calle, 2003; Peres and Nascimento, 2006; Parry et al., 2009), but for poaching it would not be possible to quantify, since there is no access to the actual number of animals killed. Quantifying the impact of illegal hunting is therefore challenging. A few studies have used hunting evidence as a metric to quantify impact (Chiarello, 2000; Wright et al., 2000), but they have used this

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evidence as a general value, not considering that such evidence was not spatially distributed uniformly. Neither was the temporal distribution of this evidence considered. Although a quantitative metric was used, the studies were not performed in a standardized way that allowed comparison between different Neotropical regions. This emphasizes that in addition to a quantitative metric, it is necessary to have a minimum of possible standardization that provides a bigger picture of the impact on the mammals. In this context, a quantitative metric assessment has been proposed. This approach considers the spatial distribution of hunting evidence per km2 and allows the trends of this impact to be monitored over time (Ferreguetti et al., 2015, 2016, 2017). This metric will collect the evidence of illegal hunting in a standardized way over time. The metric can be generated by considering each poaching event separately (date, reserve where the event was recorded, location/region of the event, geographic coordinates and type of evidence collected). Any evidence of hunting can be georeferenced over time. Examples that can be considered as evidence of hunting to georeferenced: (1) hunting elements found such as traps or baited sites: leg–hold traps, snare traps, crushing or weight traps, fall–and–apprising traps ('arapucas'), cage traps, cartridges and archery traps, corral, pitfall, among others kind of traps; (2) direct evidence of the presence of hunters, such as encounters, slaughtered animals, and camps. Together with this georeferenced database, it is recommended to use the poacher's records by using camera traps to calculate the metric. Based on the construction of this database of georeferenced hunting events it is possible to calculate a quantitative metric that consists of dividing the study area into 1–km2 grids by positioning on a digital map of the target Reserve and identifying sample sites by each area size. For example, a Reserve of 100 km2 will result in 100 grids with an intensity of hunting events per km2. Moreover, it is important to avoid counting the same record twice by removing the evidence found. Monitoring should be done on a regular basis, not exceeding a period of three months without monitoring. The metric proposed can be carried out in protected areas and can still rely on the population for a monitoring performance through training on how to gauge this georeferenced data and how to pursue conservation actions to mitigate the impact of hunting on mammalian species. References Altrichter, M., Almeida, R., 2002. Exploitation of white– lipped peccaries Tayassu pecari (Artiodactyla: Tayassuidae) on the Osa Peninsula, Costa Rica. Oryx, 36(2): 126–132. Alves, R. R. N., Souto, W. M. S., 2015. Ethnozoology: a brief introduction. Ethnobiology and Conservation, 4: 1–13. Alves, R. R., Mendonça, L. E., Confessor, M. V., Vieira, W. L., Lopez, L. C., 2009. Hunting strategies used in the semi–arid region of northeastern Brazil. Jour-

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Studies of the mesovoid shallow substratum can change the accepted autecology of species: the case of ground beetles (Coleoptera, Carabidae) in the Sierra de Guadarrama National Park (Spain) V. M. Ortuño, E. Ledesma, A. Jiménez–Valverde, G. Pérez–Suárez Ortuño, V. M., Ledesma, E., Jiménez–Valverde, A., Pérez–Suárez, G., 2019. Studies of the mesovoid shallow substratum can change the accepted autecology of species: the case of ground beetles (Coleoptera, Carabidae) in the Sierra de Guadarrama National Park (Spain). Animal Biodiversity and Conservation, 42.2: 213–226, Doi: https://doi.org/10.32800/abc.2019.42.0213 Abstract Studies of the mesovoid shallow substratum can change the accepted autecology of species: the case of ground beetles (Coleoptera, Carabidae) in the Sierra de Guadarrama National Park (Spain). The family Carabidae is of particular interest not only due to its great specific diversity but also due to the geophilic nature of many of its members, which makes them good bioindicators of soil characteristics. The diversity of the epigean Carabidae is relatively well studied. However, there are no robust data on the presence of these beetles in hypogean habitats of non–karstic substrate and, therefore, without the development of caves. In the present study, we sampled the mesovoid shallow substratum (MSS) at various sites in the Sierra de Guadarrama National Park, with the aim of characterising the Carabidae hypogean fauna. Among many other organisms, we collected 12 species of Carabidae. Of these, despite being known from epigean/edaphic habitats, Leistus (Leistus) constrictus Schaufuss, 1862, Nebria (Nebria) vuillefroyi Chaudoir, 1866, Trechus (Trechus) schaufussi pandellei Putzeys, 1870, and Laemostenus (Eucryptotrichus) pinicola (Graells, 1851) are consistently reported from MSS habitats, albeit with certain differences as regards their occupation of subterranean spaces. The species from forest–dwelling (thermophilous) lineages, T. (T.) schaufussi pandellei and L. (E.) pinicola, presented a higher prevalence in subsoil cavities at lower altitudes, whereas those from orobiont (psychrophilic) lineages, L. (L.) constrictus and N. (N.) vuillefroyi, predominated in subsoils at higher altitudes. As regards the presence of these four species during their different life cycle stages, we found that N. (N.) vuillefroyi was present and abundant as both larval (in the three preimaginal stages) and imago stages, showing the most evident trend towards an hypogean lifestyle. In contrast, for the other three species, only one of the two stages showed a high presence on hypogean habitats. The facultative hypogean capabilities of N. (N.) vuillefroyi and L. (L.) constrictus calls into question the protected status conferred on both species when it was thought that they were exclusively epigean. Key words: Mesovoid shallow substratum, Hypogean, Orobiome, Autoecology, Carabidae, Sierra de Guadarrama National Park, Iberian peninsula Resumen El estudio del medio subterráneo superficial puede cambiar la autecología aceptada de las especies: el caso de los carábidos (Coleoptera, Carabidae) en el Parque Nacional de la Sierra de Guadarrama (España). La familia Carabidae es de especial interés debido a la gran diversidad específica que atesora y al carácter geófilo de muchas de sus especies, lo que convierte a los integrantes de esta familia en buenos bioindicadores de las características del suelo. La diversidad de los carábidos de hábitos epigeos está relativamente bien estudiada; sin embargo, no hay datos sólidos sobre la presencia de estos coleópteros en el medio hipogeo de sustrato no kárstico y, por consiguiente, sin la formación de cuevas. En este estudio se realizaron capturas en el medio subterráneo superficial (MSS, en su sigla en inglés) de varios lugares del Parque Nacional de la Sierra de Guadarrama, con la finalidad de determinar las características de los carábidos hipogeos. Entre otros muchos organismos, se capturaron 12 especies de Carabidae, de las cuales Leistus (Leistus) constrictus Schaufuss, 1862; Nebria (Nebria) vuillefroyi Chaudoir, 1866; Trechus (Trechus) schaufussi pandellei Putzeys, 1870 y Laemostenus ISSN: 1578–665 X eISSN: 2014–928 X

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(Eucryptotrichus) pinicola (Graells, 1851) se capturaron frecuentemente en el medio subterráneo superficial pese a conocerse de medios epigeos y edáficos. No obstante, se observaron ciertas diferencias en la ocupación de los espacios subterráneos. Las especies provenientes de linajes de hábitos forestales (termófilos), como T. (T.) schaufussi pandellei y L. (E.) pinicola, se encontraron más frecuentemente en las oquedades del subsuelo de cotas más bajas, mientras que las especies procedentes de linajes con hábitos orobiontes (psicrófilos), como L. (L.) constrictus y N. (N.) vuillefroyi, predominaban en el subsuelo de cotas más elevadas. En cuanto a la presencia de estas cuatro especies durante sus diferentes fases del ciclo de vida, encontramos que N. (N.) vuillefroyi era abundante tanto en forma larvaria (en los tres estadios preimaginales) como en fase de imago, lo que muestra la clara tendencia hacia la adopción de un estilo de vida hipogeo. Por el contrario, en las otras tres especies solo una de las dos fases tiene una elevada presencia en el medio hipogeo. Dadas las capacidades hipogeas facultativas de N. (N.) vuillefroyi y L. (L.) constrictus, se cuestiona la figura de protección que se atribuyó a ambas especies cuando se asumía que eran de actividad totalmente epigea. Palabras clave: Medio subterráneo superficial, Hipogeo, Orobioma, Autoecología, Carabidae, Parque Nacional de la Sierra de Guadarrama, Península ibérica Received: 09 V 18; Conditional acceptance: 20 IX 18; Final acceptance: 03 XII 18 V. M. Ortuño, E. Ledesma, A. Jiménez–Valverde, G. Pérez–Suárez, Research Team on Soil Biology and Subterranean Ecosystems, Departamento de Ciencias de la Vida, Facultad de Ciencias, Universidad de Alcalá, A. P. 20 University Campus, Alcalá de Henares, E–28805 Madrid, Spain. Corresponding author. V. M. Ortuño. E–mail: vicente.ortuno@uah.es

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Introduction The mesovoid shallow substratum (MSS) was discovered as a hypogean habitat in the 1980s (Juberthie et al., 1980, 1981; Uéno, 1980, 1981), and is probably one of the most extensive but least researched subterranean habitats on the planet. The MSS consists of a network of cracks, fissures and interstices in the subsoil which occurs in very different lithologies and is formed by different processes (Juberthie et al., 1980, 1981; Oromí et al., 1986; Juberthie, 2000; Ortuño et al., 2013). In some respects, the environmental conditions of the MSS resemble those of caves (absence of light, high humidity and limited temperature variations) and the MSS can therefore host strictly hypogean species which have adapted to this habitat in accordance with the hygrophilous and stenoic nature of their lineages (Gers, 1992, 1998; Culver and Pipan, 2008; Nitzu et al., 2010, 2014; Pipan et al., 2011; Rendoš et al., 2012; Ortuño et al., 2013; Gilgado et al., 2015). However, as the MSS has a close association with the surface and soil horizons (Giachino and Vailati, 2010; Nitzu et al., 2014; Jiménez–Valverde et al., 2015), the ease with which organic matter enters the system is a substantial ecological difference from caves (Gers, 1998). This characteristic favours the presence of high densities of Arthropoda, many of them epigean or endogean, which encounter temporarily appropriate conditions in the MSS. The MSS acts as a climatic refuge, and it is therefore not surprising to find relict species sheltering in this habitat in response to major past climate changes (Christian, 1987; Hernando et al., 1999; Růžička, 1999; Ortuño et al., 2014a, 2014b). From an ecological, evolutionary and conservationist perspective, the MSS is a very important habitat that has remained unknown for a long time (Pipan et al., 2011; Ortuño et al., 2013; Jiménez–Valverde et al., 2015), partly due to the extreme difficulty in accessing and sampling its biocoenosis (Mammola et al., 2016). Carabids are one of the most intensively studied groups of arthropods, constituting one of the megadiverse families of Coleoptera (Gaston, 1991). More than 30,000 species have been described (Niemelä, 1996; Lorenz, 2005), 1,450 of which appear in the Iberian context (Serrano, 2017). Carabidae form a taxonomic group that has been widely used as a valuable bioindicator species (Rainio and Niemelä, 2003) and as a key element in biogeographical studies (Noonan, 1979). The east to west line of many Iberian mountain ranges has fostered the isolation and speciation of hypsobiont forms (Ortuño, 2002). One of these ranges is the Central System, a mountainous region which hosts a remarkable diversity of Carabidae: close to 400 species/subspecies (Serrano et al., 2003). The Sierra de Guadarrama, located in the eastern half of the Central System, is perhaps one of the most outstanding mountain sectors, with almost 250 known species (Novoa, 1975; Serrano, 1989; Ortuño and Toribio, 1996, 2002). However, almost nothing is known of its hypogean habitats because the rocky substrate is mostly siliceous and therefore lacks caves, habitats that have traditionally provided information on hypogean life.

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The objective of the present study was to increase our knowledge about the fauna of carabids of siliceous MSS, taking into consideration information on species distribution, prevalence and altitudinal range. To this end, underground spaces were sampled in the Sierra de Guadarrama, a mountainous sector that is ideal for this type of study due to the nature of its rocks and the glacial and periglacial landforms of its peaks and slopes that contain large colluvial and glacial deposits (Sanz, 1986; Pedraza and Carrasco, 2005; JCL and CAM, 2010) which host MSS. Study of MSS can contribute to the discovery of unknown preimaginal stages (instar) of many species and reveal the hypogean behavior of imagoes stages traditionally considered epigean. In sum, this would require reassessment of the currently accepted autecological knowledge of many species of Carabidae. Material and methods Study area The study was conducted in the Sierra de Guadarrama, a mountainous sector which forms a large part of the Central System, the mountain range that divides the centre of the Iberian Peninsula into two sub–plateaus. Sampling was conducted within the geographical limits of the 33,960 hectare Sierra de Guadarrama National Park (BOE, 2013) and also in part of the Peripheral Protection Area covering 62,687.26 hectares (MAPAMA). There are three mountain belts within the protected area of this national park (fig. 1A): the Montes Carpetanos, the Siete Picos–La Mujer Muerta and the Cuerda Larga, the latter being the most complex of the three since it is associated with two other important belts, La Pedriza and the Sierra de los Porrones. The three mountain belts converge at two mountain passes, those of Navacerrada and Los Cotos, where there are two non–protected areas (fig. 1A) due to the presence of ski slopes (incompatible with the conservation policies of a national park). The lithology in these sectors of the Sierra de Guadarrama is dominated by the presence of orthogneiss (Vialette et al., 1987; PNSG a). Abundant scree slopes (colluvial or glacial deposits) have been generated by fragmentation of metamorphic rocks into smaller blocks due to glacial (Pedraza and Carrasco, 2005) and peri–glacial events (Sanz, 1986). Plutonic rocks such as granite are limited to a substantial part of Siete Picos and La Pedriza. These substrates were excluded from subterranean sampling because they are broken down during erosion processes and are thus not conducive to the formation of scree slopes (fig. 1A–1B). The Sierra de Guadarrama has a Mediterranean climate, with marked continentality, characterized by dry, cool summers and cold winters. However, the diverse topography of the three mountain belts favours a considerable variety of microclimates (PNSG b; Salazar Rincón and Vía García, 2003; JCL and CAM, 2010; Palomo Segovia, 2012). The studied area is divided into three bioclimatic zones: supramediter-

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ranean, oromediterranean and crioromediterranean (Rivas–Martínez, 1984; Rivas–Martínez et al., 1987). The supramediterranean zone extends from approximately 1,300 to 1,700 m a.s.l., dominated by the Pyrenean oak (Quercus pyrenaica Willd.), a species whose presence has been reduced by human activity in favour of the Scots pine (Pinus sylvestris L.). The oromediterranean zone extends from approximately 1,700 to 2,100 m a.s.l., although upper limits can reach higher altitudes depending on the orographic characteristics of the terrain in each place. This zone hosts the most extensive plant formations of the Sierra de Guadarrama National Park (JCL and CAM, 2010), divided into two belts: a lower belt (1,700 to 1,950 m a.s.l.) of forest dominated by P. sylvestris (fig. 1B) and an upper belt (1,950 to 2,100 m a.s.l.) characterized by scrub supra–forest (montane scrubland), composed primarily of two species (Novoa, 1977), Cytisus oromediterraneus Rivas Mart. et al. and Juniperus communis alpina (Suter) Celak., as well as Adenocarpus hispanicus (Lam.) DC. or Erica arborea L. to a lesser extent. The scrub alternates with pastureland, rocky outcrops and scree slopes, forming part of the high altitude landscapes of the Sierra de Guadarrama (fig. 1B). The crioromediterranean zone is comprised in the highest areas of the Sierra de Guadarrama (approx. 2,100 m a.s.l. up to the maximum altitude, 2,428 m a.s.l., at Peñalara peak). Typical vegetation consists of psychroxerophilic pastureland (with abundant presence of Festuca curvifolia Lag. ex Lange), sub–hygrophilous pastureland (dominated by Nardus stricta L. grasslands, typical of oligotrophic montane soils) (Rivas Martínez, 1963; Rivas Martínez et al., 1990) and peat bog (hosting species of the genus Carex L. 1753), although the presence of P. sylvestris has been documented in some south–facing areas (Muñoz Municio et al., 2004). The scree slopes located in the crioromediterranean and oromediterranean zones present a variety of rupicolous plant species characteristic of rocky substrates (JCL and CAM, 2010). Precipitation in the highest bioclimatic zones generally occurs in the form of snow, which remains on the ground throughout the winter and much of the spring. In basins or areas protected from strong sun, snow deposits persist for longer and are known as snowfields. Sampling After evaluating the amount of sampling effort that could be realized, we selected 33 scree slopes across the Sierra de Guadarrama National Park with the intention of covering most of the geographic area and bioclimatic zones of the Park, while taking into account the structural quality of the MSS and its accessibility (fig. 1). We installed 33 subterranean sampling devices (SSD) which were slightly modified from the pilot model developed for the first Iberian sampling campaign of these characteristics (Barranco et al., 2013; Ortuño et al., 2013). These SSDs consisted of a pitfall trap and a PVC cylinder measuring 11 cm in diameter and 1 m long, and contained numerous perforations (8 mm in diameter) from the middle al-

most to the bottom of the cylinder (sampling depth: –0.5 m to –0.9 m) (see fig. 2 in Baquero et al., 2017). Another 4 SSDs measuring 0.5 m long and containing perforations in the lower 30 cm (sampling depth: –0.2 m to –0.4 m) were installed in sampling sites 1 to 4 (Siete Picos–La Mujer Muerta mountain belts) next to the 1 m long SSDs. Only four short SSDs were installed due to the enormous extra amount of work that additional traps implied. The pitfall trap, which fitted perfectly inside the cylinder, contained a liquid preservative (1.2–propanediol) and was baited with a vial containing very strong–smelling cheese, a standard procedure in this kind of studies (Gers, 1992; Giachino and Vailati, 2010; López and Oromí, 2010). The cylinder was buried vertically in the appropriate substrate, and when the top was level with the soil surface, the trap was lowered inside on a nylon thread to the bottom; then the cylinder was covered and isolated from the external environment using waterproof material covered by a layer of soil or stones (see figure 2 in Baquero et al., 2017). The sites sampled fell within the National Park (31 locations) and, to a lesser extent, the Peripheral Protection Area (2 locations) (fig. 1B; see table 1 in Baquero et al., 2017). The SSDs were installed between 20/05/2015 and 09/07/2015, and collections were made at three different times between 20/05/2015 and 14/10/2016. Samples were taken to the laboratory, and specimens were identified to the species level and stored in the entomological collection of the Department of Life Sciences, Faculty of Sciences, University of Alcalá, Alcalá de Henares, Madrid, Spain (Collection V. M. Ortuño). The relative percentage (prevalence) of each dominant species was calculated for each sector and for each bioclimatic zone. Results We collected a total of 12 Carabidae species in the MSS (with the 1 m SSDs).These samples were unevenly represented: Carabus (Oreocarabus) guadarramus Laferté, 1847 [larva: 1, imagoes: 3]; Leistus (Leistus) constrictus Schaufuss, 1862 [L: 237, I: 32]; Nebria (Nebria) vuillefroyi Chaudoir, 1866 [L: 148, I: 203]; Nebria (Nebria) salina Fairmaire and Laboulbène, 1854 [I: 16]; Trechus (Trechus) quadristriatus (Schrank, 1781) [I: 1]; Trechus (Trechus) schaufussi pandellei Putzeys, 1870 [L: 1, I: 234]; Cryobius nemoralis nemoralis (Graells, 1851) [I: 1]; Steropus (Iberocorax) ghilianii (Putzeys, 1846) [I: 1]; Platyderus (Platyderus) varians Schaufuss, 1862 [I: 6]; Laemostenus (Eucryptotrichus) pinicola (Graells, 1851) [L: 14, I: 572]; Synuchus vivalis (Illiger, 1798) [I: 1]; and Cymindis (Cymindis) coadunata monticola Chevrolat, 1866 [I: 1]. Only four species, L. (L.) constrictus, N. (N.) vuillefroyi, T. (T.) schaufussi pandellei and L. (E.) pinicola, were very abundant in different MSS in the National Park. We collected a total of 269 specimens of Leistus (Leistus) constrictus with the 1 m SSDs, 237 (88 %) of which were preimaginal stages and 32 (12 %)

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ia ov id g Se adr M

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Peripheral Protection Area Pinus sylvestris High mountain: pastures, scrubs and scree slopes

Fig. 1. A, basic orography of the Sierra de Guadarrama National Park; B, location of the subterranean sampling devices (SSD) in the Sierra de Guadarrama National Park, with the distribution of the two most extended ecosystems and indication of the bioclimatic zone for each device: SMZ, supra–mediterranean zone; OMZ–F, oro–mediterranean zone (forest); OMZ–S, oro–mediterranean zone (scrub); CMZ, crioro–mediterranean zone. Fig. 1. A, orografía básica del Parque Nacional de la Sierra de Guadarrama; B, ubicación de los dispositivos de muestreo subterráneo (SSD por su sigla en inglés) en el Parque Nacional de la Sierra de Guadarrama, con la distribución de los dos ecosistemas más extensos e indicando el piso bioclimático para cada dispostivo: SMZ, zona supramediterránea; OMZ–F, zona oromediterránea (bosque); OMZ–S, zona oromediterránea (matorral); CMZ, zona crioromediterránea.

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were imagoes. Similar percentages of the three preimaginal stages were collected (26 % instar–I; 41% instar–II; 33 % instar–III). The sex ratio (based always on imagoes) was highly skewed to males (87.5:12.5). This species was widely distributed throughout the MSS in the study area, but was particularly abundant in the Loma de Pandasco (SSD–30). Compared to the other three species, L. (L.) constrictus showed an uneven prevalence: 9 % being found in the Siete Picos–La Mujer Muerta, 17 % in the Montes Carpetanos, 24 % in the Cuerda Larga and associated mountainous complex, and 27 % in the area of convergence (the mountain passes of Los Cotos and Navacerrada). Leistus (L.) constrictus comprised 7 % of the carabid fauna present in MSS at levels below 1,700 m a.s.l., but increased significantly in the oromediterranean forest zone (16 %) and scrub supra–forest (21 %), reaching a maximum value of 26 % above 2,100 m a.s.l., in the crioromediterranean zone (fig. 2). Nebria (Nebria) vuillefroyi was collected in large numbers in the 1 m SSDs, capturing 351 specimens: 148 (42 %) preimaginal stages and 203 (58 %) imagoes. Very dissimilar percentages of the three preimaginal stages were collected (55 % instar–I; 34% instar–II; 11 % instar–III). The sex ratio (imagoes) was slightly skewed to males (59:41). The 14 MSS sites where it was present considerably expand our chorological knowledge of this species, which was especially abundant in the Cerro de Navahonda (SSD–20) and Collado de Valdemartín scree slopes (SSD–28). This species accounted for 7 % of specimens collected in the area of convergence (mountain passes of Los Cotos and Navacerrada), 26 % in Montes Carpetanos and 36 % in Cuerda Larga and associated mountainous complex, whereas it was not found in Siete Picos–La Mujer Muerta. Nebria (N.) vuillefroyi was not present in the MSS below 1,700 m a.s.l. but showed a notable presence (15 %) in the oromediterranean forest zone, rising slightly to 20 % in the scrub supra–forest, and very significantly to 52 % at above 2,100 m a.s.l., in the crioromediterranean zone (fig. 2). The 1 m SSDs also collected a high number of Trechus (Trechus) schaufussi: 235 specimens, all imagoes except for one larva. The sex ratio was skewed in favour of females (28:72). This species was collected in the MSS at 33 sampling sites, and was especially abundant in Cancho del Río Peces (SSD–1) (Siete Picos–La Mujer Muerta) and the scree slope at La Najarra–Cuatro Calles (SSD–26) (Cuerda Larga and associated mountainous complex). The geographical prevalence of this species in the MSS was uneven, accounting for 39 % of specimens in the Siete Picos– La Mujer Muerta, 6 % in Montes Carpetanos, 14 % in Cuerda Larga and associated mountainous complex, and 11% in the area de convergence (the mountain passes of Los Cotos and Navacerrada). The species was highly represented in the supramediterranean zone, with prevalence values of 50 %. As altitude increased, its prevalence in the MSS decreased, accounting for 13 % and 20 % in the oromediterranean forest and the scrub supra–forest, respectively, and only 6 % in the crioromediterranean zone (fig. 2).

Laemostenus (Eucryptotrichus) pinicola was very abundant, with 586 specimens being collected in the 1 m SSDs, mainly imagoes (572; 98 %), and only 14 (2 %) preimaginal stages. The sex ratio was slightly skewed to females (40:60). This species was widely distributed throughout the study area, and was especially abundant in the scree slopes near El Paredón (SSD–21) and Las Revueltas–Los Horcos (SSD–16) (Montes Carpetanos), and Majada Conejo (SSD–4) (Siete Picos–La Mujer Muerta). Laemostenus (E.) pinicola had a high prevalence in the MSS of all mountainous sectors, exceeding 50 % in all cases except for Cuerda Larga and its associated mountainous complex (with 26 %). In terms of altitudinal range, this species presented a high prevalence of 42 % in the MSS of the supramediterranean zone, 53 % in the oromediterranean forest, 39 % in the scrub supra–forest, and 16 % the crioromediterranean zone (fig. 2). The high number of T. (T.) schaufussi pandellei and L. (E.) pinicola collected made it possible to try to compare the abundance of these species according to sampling depth. Trechus (T.) schaufussi pandellei was much less abundant – and even absent in the 0.5 m SSDs than in the 1 m SSDs (fig. 3A). However L. (E.) pinicola was more abundant in two of the 0.5 m SSDs (fig. 3B). Nevertheless, these results must be interpreted with caution due to the small number of short SDDs. Discussion Although the Iberian epigean carabid fauna can be considered relatively well known (Serrano, 2017), there is little information on their presence in the MSS (e.g., Ortuño and Toribio, 1994; Pons and Palmer, 1996; Ortuño, 1996, 2004; Fresneda et al., 1997; Toribio and Rodríguez, 1997; Carabajal, 1999; Hernando et al., 1999; Jeanne, 2000; Faille et al., 2012; Toribio, 2014; Ortuño et al., 2014a, 2017), and taxonomic publications predominate. Novoa (1977) listed 50 species present in the Sierra de Guadarrama oak woods of the supramediterranean zone, highlighting 7 species as particularly frequent (fig. 2). None of these species was collected in the MSS sampled in this bioclimatic zone, but instead finding T. (T.) schaufussi pandellei and L. (E.) pinicola, and to a lesser extent, L. (L.) constrictus (fig. 2) in the subsoil. With regard to the oromediterranean zone, Novoa (1977) distinguished between the Carabidae fauna of Pinus sylvestris pine woods and the carabid fauna of scrub supra–forest. The pine woods hosted 45 species of Carabidae, 11 of which were especially representative of this forest environment (Novoa, 1977) (fig. 2). Of these species, only T. (T.) schaufussi pandellei and L. (E.) pinicola seem to be established in subterranean habitats at this altitude, constituting somewhat less than 75% of the Carabidae specimens found in the sampled MSS (fig. 2). Laemostenus (E.) pinicola was the dominant species (with more than 50 % of specimens, followed far behind by L. (L.) constrictus (16 %), although both species had an important presence in the samples

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Harpalus (Harpalus) decipiens Trechus (Trechus) schaufussi pandellei / Zabrus (Iberozabrus) seidlitzi / Platyderus (Platyderus) varians Steropus (Iberocorax) ghilianii Calathus (Calathus) vuillefroyi / Cymindis (Cymindis) coadunata Elaphrus (Neoelaphrus) pyrenoeus Leistus (Leistus) constrictus Nebria (Nebria) vuillefroyi Bembidion (Trepanedoris) doris Bembidion (Philochthus) guadarramus Bembidion (Nepha) lateralis Bembidion (Nepha) ibericum Bembidion (Testediolum) carpetanum Cryobius nemoralis nemoralis Agonum sexpunctatum Agonum viridicupreum Harpalus (H.) contemptus

Cymindis (Menas) miliaris Carabus (Oreocarabus) ghilianii Notiophilus biguttatus Cryobius nemoralis nemoralis Calathus (Calathus) hispanicus hispanicus Laemostenus (Eucryptotrichus) pinicola Syntomus foveatus

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Fig. 2. Altitude distribution of the most conspicuous species of Carabidae in the epigeous environment (according to Novoa, 1977) of the Sierra de Guadarrama, and their correspondence in the mesovoid shallow substratum (according to results of this study). Fig. 2. Distribución altitudinal de las especies más notables de Carabidae en el medio epigeo (según Novoa, 1977) de la Sierra de Guadarrama y su correspondencia en el medio subterráneo superficial (según los resultados de este estudio).

of supramediterranean subsoil. Notably, N. (N.) vuillefroyi showed a high presence (15 %) at altitudes well below its known optimal epigean habitat, while T. (T.) schaufussi pandellei had a reduced presence, turning from being the most abundant species in the sampled supramediterranean MSS to being the least abundant in the oromediterranean forest MSS (fig. 2). The carabid fauna of the sampled oromediterranean scrub supra–forest MSS did not correspond to that observed by Novoa (1977) in epigean habitats, with the exception of T. (T.) schaufussi pandellei. Novoa described the presence of 30 species, eight of which were particularly frequent and abundant, and five of which also occupied the pine wood belt (fig. 2). Of these, T. (T.) schaufussi pandellei was collected in some abundance in the MSS (17 % of specimens), close to the values for N. (N.) vuillefroyi and L. (L.) constrictus (20 % and 21 %, respectively). Laemostenus (E.) pinicola was the most abundant species, as in the subsoil of the pine wood belt (fig. 2). The crioromediterranean zone hosts a remarkable num-

ber of Carabidae species: Novoa (1977) described 42, 19 of which were particularly frequent at high altitudes in the Sierra de Guadarrama (fig. 2). However, given the different hygrophilous nature of these species, they were unevenly distributed in habitats such as psychroxerophilic pasturelands, Nardus stricta grasslands, peat bogs and snowfields. At this altitude, the sampled MSS hosted three species which also formed part of the most representative species of the typically orophilous epigean fauna in the Sierra de Guadarrama (fig. 2): N. (N.) vuillefroyi, L. (L.) constrictus and T. (T.) schaufussi pandellei. A fourth species, L. (E.) pinicola, which is dominant in the sampled subsoil of lower bioclimatic zones but not recognised as characteristic of crioromediterranean epigean habitats, also appeared in the sampled crioromediterranean MSS, although its presence was very low (fig. 2). At such high altitudes, the presence of Carabidae in the MSS changed dramatically from that observed at lower altitudes. Nebria (N.) vuillefroyi became the dominant species (with more than 50 %

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Fig. 3. Abundance of Trechus (Trechus) schaufussi pandellei (A) and Laemostenus (Eucryptotrichus) pinicola (B) in double installation in the 'Siete Picos– La Mujer Muerta' chain: a, SSD (1 m); b, SSD (0.5 m). Fig. 3. Abundancia de Trechus (Trechus) schaufussi pandellei (A) y Laemostenus (Eucryptotrichus) pinicola (B) en doble instalación en el cordal montañoso de Siete Picos–La Mujer Muerta: a, SSD (1 m); b, SSD (0.5 m).

of the specimens), followed far behind by L. (L.) constrictus (26 %), L. (E.) pinicola (16%) and T. (T.) schaufussi pandellei (6 %) (fig. 2). Some species which are very frequent in epigean habitats in the Sierra de Guadarrama appeared only very occasionally in the MSS samples, namely Cryobius nemoralis nemoralis, Steropus (Iberocorax) ghilianii, Platyderus (Platyderus) varians and Cymindis (Cymindis) coadunata monticola. Others, also very frequent, such as Zabrus (Iberozabrus) seidlitzi and Calathus (Calathus) vuillefroyi, were not observed at all in the MSS. This finding, coupled with the prevalence of the four dominant species in the hypogean habitat (fig. 2), suggests that occupation of this hypogean environment by Carabidae fauna in the Sierra de Guadarrama does not depend on the greater or lesser penetrability of colluvial and glacial deposits. The presence of L. (L.) constrictus, N. (N.) vuillefroyi, T. (T.) schaufussi pandellei and L. (E.) pinicola in the MSS, but not of other taxa inhabiting the epigean environment, must be sought in their autecological characteristics. The data obtained in the present study suggest that the MSS acts as a filter for epigean species in such a way that only a few manage to achieve an appreciable abundance in the subsoil. Preliminary data from a second year of survey suggest inter–annual consistency of these findings (Vicente M. Ortuño, unpublished). Leistus (Leistus) constrictus is an endemic species restricted to the Guadarrama and Ayllón mountains (Perrault, 1979; Serrano, 2003, 2013), although known records indicate that it is not rare (Jeanne, 1966; Novoa, 1975; Perrault, 1979; Serrano, 1989; Ortuño and Toribio, 1996). It forms part of a group of orophilous species of Leistus Frölich, 1799, which

inhabit a peri–plateau ring in the northern half of the Iberian Peninsula (Jeanne, 1976; Perrault, 1979). The data available to date indicate that it preferentially selects habitats at oromediterranean altitudes, leading a sublapidicolous life on damp soils in the pine wood belt (Novoa, 1975). Nevertheless, it is not one of the most frequent carabids in these woods (Novoa, 1977) (fig. 2). At higher altitudes, where soil xericity increases, it has been found in Nardus stricta grasslands and on the edges of snowfields (Novoa, 1977). Available evidence indicates that L. (L.) constrictus encounters difficulties inhabiting dry soil habitats. Nevertheless, it does not appear to seek alternatives to meet its hygrophilous needs, instead presenting ripicolous behaviour, as evidenced by Novoa (1980). When found on the banks of a water course, it is normally protected by woodland (see collection data in Serrano, 1989; Ortuño, personal observation). The data reported here represent the first record of L. (L.) constrictus as an inhabitant of the MSS. The hygrophilous nature of the species in this genus, coupled with their predation of Acari and Collembola (Lindroth, 1985) —both groups very abundant in the subsoil— would explain why L. (L.) constrictus is so abundant and widely distributed throughout the MSS in the Sierra de Guadarrama. Although some imagoes were collected in the MSS, most specimens (88 %) were preimaginal stages (previously unknown), a finding which suggests that this species is eminently hypogean during larval stages (fig. 4). This indicates that the known autecology for this species requires revision, and calls into question its vulnerability status due to habitat alteration (BOCM, 1992), to date based solely on knowledge of its epigean behaviour. This is not the first time that hypogean behaviour has been

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L a r v a

I m a g o

Fig. 4. Representation of the hypogean activity (larva and imago phases) of the four species of Carabidae predominant in the mesovoid shallow substratum samples from the Sierra de Guadarrama National Park. Fig. 4. Representación de la actividad hipogea (fase larvaria y de imago) de las cuatro especies de Carabidae predominantes en las muestras del medio subterráneo superficial del Parque Nacional de la Sierra de Guadarrama.

reported in Leistus, previously noted by Assmann (1997) in relation to Leistus (Leistus) starkei Assmann, 1977, a species closely related to L. (L.) constrictus. The presence of Leistus (Pogonophorus) puncticeps Fairmaire and Laboulbene, 1854, in the MSS has also been observed in the eastern Iberian Peninsula (Jiménez–Valverde et al., 2015). Nebria (Nebria) vuillefroyi is an orobiont species endemic to the Guadarrama (Bruneau de Miré, 1964) and Ayllón (Ortuño and Toribio, 1994) mountains. Although it has also been in the Sierra de Béjar (Ledoux and Roux, 1992), the Sierra de Gredos (Serrano, 2003, 2013) and, as a result of the proposed synonymy with Nebria (Nebria) urbionensis Arribas, 1991, also in the Sierra de Urbión (Ledoux and Roux, 1992). This widespread distribution requires corroboration and further more detailed evidence to support the synonymy proposed by Ledoux and Roux (1992), since N. (N.) urbionensis may be a cryptic species that resembles N. (N.) vuillefroyi (Arribas and Ortuño, in prep.). It has been known for more than a century that this species inhabits the highest altitudes in the Sierra de Guadarrama, although in a description of the species, Chaudoir (1866) did

not specify its geographical origin. Its presence has been documented at several sites in this mountain range, based on imagoes observed on the edges of snowfields (Novoa, 1975; Ortuño and Toribio, 1996), and on instar–III (Vives, 1978). Snowfields are important habitats for the survival of some hygrophilous species of an orophilous nature such as N. (N.) vuillefroyi, since they provide soil moisture (due to the effect of melting) and are a good food source due to accumulated biomass from aerial plankton (Palanca and Castán, 1995). This species has been reported at altitudes below 2,000 m a.s.l., albeit sporadically, on stream banks in the Sierra de Guadarrama (between 1,700 and 1,769 m a.s.l.) and in the MSS of the Sierra de Ayllón at 1,650 m a.s.l. (Ortuño and Toribio, 1994). Here N. (N.) vuillefroyi is reported for the first time in the MSS of the Sierra de Guadarrama, revealing that larvae and imagoes maintain a constant presence in this habitat (fig. 4). The greater vulnerability of preimaginal stages coupled with the numerous specimens collected in the MSS suggest that this species mainly inhabits hypogean habitats. Imagoes seem to be facultative hypogean, although this lifestyle may become obligatory during much

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of the year when the snowfields disappear, and the summits become high altitude deserts. The MSS at high altitude provides high humidity, cool temperatures with fewer variations than outside, and a good supply of prey, facilitating occupation by this species. These environmental characteristics are also repeated in the MSS at lower altitudes, facilitating the presence of N. (N.) vuillefroyi at sites which do not normally host snowfields and therefore do not present optimal epigean characteristics (fig. 2). These observations indicate that the autecology assigned to date to this species is not complete, and that there is a need to reduce its vulnerability status due to habitat alteration (BOCM, 1992), since its optimum habitat is not restricted exclusively to nivicolous habitats or high altitude epigean environments. Trechus (Trechus) schaufussi is a species endemic to the Iberian Peninsula, whose distribution is limited to various mountain sites (Jeanne, 1976). Although it may form part of the nivicolous insects (Novoa, 1975; Ortuño and Toribio, 1996), its hygrophilous needs also appear to be satisfied in forest habitats (Jeanne and Zaballos, 1986). It has also been observed in different subterranean habitats (Ortuño and Arribas, 2010). Nine subspecies are currently recognised (see Serrano, 2013), of which T. (T.) schaufussi pandellei represents the population in Guadarrama (Putzeys, 1870) and Ayllón mountains (Ortuño and Arribas, 2010). In the Sierra de Guadarrama, it is not confined exclusively to epigean habitats but also inhabits the subsoil (Ortuño and Arribas, 2010). The present study demonstrates that it is widespread throughout the sampled MSS in the National Park, in all the bioclimatic zones sampled (fig. 2). The near absence of preimaginal stages in the MSS compared to the remarkable abundance of imagoes was notable. This might be due to edaphobiont lifestyles as larvae, whereas the imagoes display hypogean activity (fig. 4). It is not surprising that without any apparent adaptation to hypogean habitats, species of the genus Trechus Clairville, 1806, seek shelter in the MSS driven by their geophilous and lucifugous habits and hygrophilous needs (Ortuño, 2004). Laemostenus (Eucryptotrichus) pinicola is endemic to the Central System (Serrano, 2003), discovered in the Sierra de Guadarrama (Graells, 1851), and also found later in Gredos (Jeanne, 1968), Ayllón (Serrano, 1981), Béjar (Zaballos, 1986) and Estrella (Mateu, 1996) mountains. It has been observed in several montane habitats, although it has been considered a forest species (Jeanne, 1968) typical of rainforest (Vives and Vives, 1982), where it is sublapidicolous (Novoa, 1975) or lives under the bark and wood of dead pines (Graells, 1851). Its known altitudinal range is between 1,200 and 2,200 m a.s.l. (Zaballos, 1985), crossing the upper limit of the tree line, where it becomes less frequent (Novoa, 1975; Ortuño and Toribio, 1996). The lucifugous nature of this species (Ortuño and Toribio, 1996) is a widespread ecophysiological trait found to a varying degree in other Sphodrina species that have adopted subterranean habitats (Jeannel, 1937; Casale, 1988; Casale et al., 1998). This trait explains why it has been found not only under large sunken

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stones (Novoa, 1975; Zaballos, 1986) but also and especially in the MSS in all bioclimatic zones present in the study area (fig. 2). The specimens collected in the MSS suggest that imagoes but not larvae present intensive hypogean activity (fig. 4). This is the first time that this Sphodrini has been described in the MSS. In the light of the above results, these four Carabidae species should be considered facultatively hypogean. However, the data suggest differences in subterranean occupation according to the altitude of the MSS studied and to the thermophilous or psycrophilic nature of the lineages of each species (fig. 2). The more thermophilous, T. (T.) schaufussi pandellei and L. (E.) pinicola, both from forest–dwelling lineages, were less prevalent in high altitude subterranean spaces, whereas the psychrophilic, L. (L.) constrictus and N. (N.) vuillefroyi, from orobiont (glacial/nivicolous) lineages, became more prevalent in the subsoil as altitude increased. Regarding autecological and life cycle aspects, each of these four species shows a different interaction with the MSS (fig. 4). In Nebria (Nebria) vuillefroyi, the hypogean lifestyle seems to be an integral part of the life cycle (fig. 4) as evidenced by the numerous imagoes and larvae collected in the sampled MSS. This was very different to L. (L.) constrictus, also a Nebriini, in which the hypogean lifestyle was clearly attributable to the larvae, whereas imagoes were only sporadically present in the MSS (fig. 4). Some Nebriini have already been described in hypogean habitats in montane landscapes and areas subject to snow cover (Bruneau de Miré, 1985; Casale et al., 1998; Galán, 1993). These observations are consistent with the longstanding idea of a close interrelationship between some hypogean lineages and nivicolous fauna (Jeannel, 1943; Vandel, 1964; Bellés, 1987). Laemostenus (E.) pinicola showed yet another pattern: the hypogean activity was clearly evident in the case of imagoes but merely sporadic in that of larvae (fig. 4), which may have the soil preferences commonly found in many Carabidae. Lastly, T. (T.) schaufussi pandellei imagoes were present in the MSS, but not the larvae (fig. 4). Thus, clear and substantial differences between these four species were found with regards to subterranean occupation of imagoes and larvae. Trechus (T.) schaufussi pandellei and L. (E.) pinicola were more frequent at sampling sites in Siete Picos–La Mujer Muerta and in supramediterranean and oromediterranean forest bioclimatic zones (fig. 1B). This is probably because they are species from forest–dwelling lineages. The results of sampling at different depths suggested a preference of the imagoes of T. (T.) schaufussi pandellei for deeper subterranean spaces (fig. 3A). However, the data on L. (E.) pinicola indicate that this species was not as demanding (less stenoic) with respect to subterranean occupation depth (fig. 3B). This different response to hypogean habitats could correspond to the more stenoic nature of the Trechini lineage, favouring a more intensive and successful colonisation of subterranean environments than its Sphodrini counterparts (see Bellés, 1987; Casale et al., 1998).

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Conclusions Of the near 250 Carabidae species considered epigean and present in the Sierra de Guadarrama, only four, L. (L.) constrictus, N. (N.) vuillefroyi, T. (T.) schaufussi pandellei and L. (E.) pinicola, were systematically recovered in the sampled colluvial deposits. This suggests that penetrability and occupation of this habitat largely corresponds to the ecophysiological characteristics of the species, and that the MSS acts as a filter for the epigean Carabidae accessing hypogean environments. The Carabidae species inhabiting the MSS in the Sierra de Guadarrama National Park differed in the subsoil occupation according to altitude. The species from forest–dwelling lineages (thermophilous species), T. (T.) schaufussi pandellei and L. (E.) pinicola, presented a higher prevalence at lower altitudes than those from orobiont (psychrophilic species) lineages, L. (L.) constrictus and N. (N.) vuillefroyi, which predominated in subsoils at higher altitudes. Regarding the presence of the different life cycle stages of these four species in the MSS, we found that Nebria (N.) vuillefroyi is abundantly present in the subsoil at both larval and imago stages. In contrast, the other three species were only abundant in hypogean habitats in one of the two stages of the life cycle. We found that L. (L.) constrictus larvae were abundant in the MSS, thus forming part of the hypogean contingent, whereas the imago stage was rarely present. L. (E.) pinicola and T. (T.) schaufussi pandellei imagoes were abundant in the MSS, but the larval stages of both species were rare or absent, especially in the case of T. (T.) schaufussi pandellei. Nebria (N.) vuillefroyi and L. (L.) constrictus are both protected species and have been classified as 'vulnerable to habitat alteration' (according to the Community of Madrid Regional Catalogue of Endangered Species). The present discovery of facultative hypogean behaviour in both species suggests that their protection status is worthy of revision. We observed preliminary differences between the vertical occupations of subterranean spaces in the two more thermophilous Carabidae species collected. L. (E.) pinicola occupied both deeper and shallower spaces, whereas T. (T.) schaufussi pandellei predominated in deeper spaces. This finding suggests that T. (T.) schaufussi pandellei is more stenoic than L. (E.) pinicola. Acknowledgements This work was funded by the project 'Estudio de la diversidad y distribución de las especies animales residentes en el Medio Subterráneo Superficial de enclaves de Alta Montaña (P. N. de la Sierra de Guadarrama)' [Study of the diversity and distribution of the animal species of the Mesovoid Shallow Substratum in enclaves of high Mountain (Sierra de Guadarrama National Park)], conceded by the Autonomous Organism of National Parks of Spain. Ref.(1143/2014). It was also funded by the Program for Young Researchers

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of the University of Alcalá 'Contratos Predoctorales de Personal Investigador en Formación' under the budget implementation 30400M000.541.A 640.06, of which Enrique Ledesma is a beneficiary. A. Jiménez–Valverde was supported by the MINECO Ramón y Cajal Program (RYC–2013–14441). We would like to thank the staff at the National Park who kindly helped us with the permission applications and other formalities, and to those who also helped us with the fieldwork, especially Patricia Riquelme, Pablo Sanjuanbenito, Juan A. Vielva, Javier Donés, Marisol Redondo, Ignacio Granados, Ángel Rubio, César Martín, José Carrillo, Miguel Ángel Palomar, Ángel Velasco, Germán Mato, Manuel Criado, Enrique Calvo, Federico Madejón, Montserrat Sanz, and forestry agents of Buitrago de Lozoya. Thanks also to our colleagues who collaborated in the design of the samplings and the fieldwork, such as José D. Gilgado, Enrique Baquero, Alberto Sendra, Pablo Barranco, Alberto Tinaut, Rafael Jordana, Luis Subías, Juan José Herrero–Borgoñón, Douglas Zeleppelini and Javier Ledesma. Special thanks too to our colleagues and students who helped us in fieldwork and laboratory work: Joaquín Calatayud, David Cabanillas, Sara de Lope, and Daniel Méndez. References Assmann, Th., 1997. A new species of Leistus Frölich from the Picos de Europa, Cantabrian mountains, Spain (Coleoptera: Carabidae). Koleopterologische Rundschau, 67: 1–4. Baquero, E., Ledesma, E., Gilgado, J. D., Ortuño, V. M., Jordana, R., 2017. Distinctive Collembola communities in the Mesovoid Shallow Substratum: first data for the Sierra de Guadarrama National Park (Central Spain) and a description of two new species of Orchesella (Entomobryidae). PLOS One, 12(12): e0189205, https://doi.org/10.1371/journal. pone.0189205 Barranco, P., Gilgado, J. D., Ortuño, V. M., 2013. A new mute species of the genus Nemobius Serville (Orthoptera, Gryllidae, Nemobiinae) discovered in colluvial, stony debris in the Iberian Peninsula: A biological, phenological and biometric study. Zootaxa, 3691(2): 201–219, http://dx.doi. org/10.11646/zootaxa.3691.2.1 Bellés, X., 1987. Fauna cavernícola i intersticial de la península ibérica i les illes balears. CSIC, Ed. Moll, Mallorca. BOCM, 1992. Decreto 18/92, de 26 de marzo, por el que se aprueba el Catálogo Regional de especies amenazadas de fauna y flora silvestres y se crea la categoría de árboles singulares. Boletín Oficial de la Comunidad de Madrid, 85 (9 de abril de 1992): 5–11. BOE, 2013. Ley 7/2013, de 25 de junio, de declaración del Parque Nacional de la Sierra de Guadarrama. Boletín Oficial del Estado, 152 (26 de junio de 2013). Bruneau de Miré, P., 1964. Essai d’interprétation de la variation géographique et la spéciation chez les Nebria orophiles du nord–ouest de la Péninsule Ibérique. Revue Française d’entomologie, 31(1):

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Monitoring black grouse Tetrao tetrix in Isère, northern French Alps: cofactors, population trends and potential biases L. Dumont, E. Lauer, S. Zimmermann, P. Roche, P. Auliac, M. Sarasa

Dumont, L., Lauer, E., Zimmermann, S., Roche, P., Auliac, P., Sarasa, M., 2019. Monitoring black grouse Tetrao tetrix in Isère, northern French Alps: cofactors, population trends and potential biases. Animal Biodiversity and Conservation, 42.2: 227–244, Doi: https://doi.org/10.32800/abc.2019.42.0227 Abstract Monitoring black grouse Tetrao tetrix in Isère, northern French Alps: cofactors, population trends and potential biases. Wildlife management benefits from studies that verify or improve the reliability of monitoring protocols. In this study in Isère, France, we tested for potential links between the abundance of black grouse (Tetrao tetrix) in lek–count surveys and cofactors (procedural, geographical and meteorological cofactors) between 1989 and 2016. We also examined the effect of omitting or considering the important cofactors on the long–term population trend that can be inferred from lek–count data. Model selections for data at hand highlighted that the abundance of black grouse was mainly linked to procedural cofactors, such as the number of observers, the time of first observation of a displaying male, the day, and the year of the count. Some additional factors relating to the surface of the census sector, temperature, northing, altitude and wind conditions also appeared depending on the spatial or temporal scale of the analysis. The inclusion of the important cofactors in models modulated the estimates of population trends. The results of the larger dataset highlighted a mean increase of +17 % (+5.3 %; +29 %) of the abundance of black grouse from 1997 to 2001, and a mean increase in population of +47 % (+16 %; +87 %) throughout the study period (1989–2016). We discuss the hypothesis of plausible links between this past increase in the number of black grouse and the ecological impact of the winter storm 'Vivian'. Findings from our study and the ecological phenomena that were concomitant with the observed population trend provide opportunities to strengthen the monitoring and management of black grouse in the Alps. Key words: Abundance estimates, Protocol cofactors, Lek–count, Wildlife management, Alpine bird, Galliformes Resumen Seguimiento del gallo lira, Tetrao tetrix, en Isère, en el norte de los Alpes franceses: factores, tendencias demográficas y posibles sesgos. La gestión de la fauna silvestre se beneficia de estudios que verifican o mejoran la fiabilidad de los protocolos de seguimiento. En este estudio, realizado en Isère (Francia) entre 1989 y 2016, hemos analizado la posible relación entre la abundancia de gallo lira (Tetrao tetrix), determinada en estudios de conteo en cantaderos, y una serie de factores (geográficos, meteorológicos y de procedimiento). Después, hemos examinado también el efecto de omitir o considerar los factores importantes que inciden en la tendencia demográfica a largo plazo que puede ser inferida de los datos sobre conteos en cantaderos. Los modelos seleccionados para los datos disponibles subrayaron que la abundancia de gallo lira estaba principalmente relacionada con factores de procedimiento, como el número de observadores, la hora de la primera observación de un macho exhibiéndose, el día y el año del conteo. Según la escala espacial y temporal de los análisis, también se observó alguna relación con otros factores relacionados con la superficie del sector censado, la temperatura, la orientación respecto al norte, la altitud y las condiciones de viento. La inclusión de los factores importantes en los modelos modificó las estimaciones de las tendencias demográficas. Los resultados del conjunto de datos más grande indicaron un incremento medio de +17 % (+5,3 %; +29 %) y de +47 % (+16 %; +87 %) de la abundancia de gallo lira durante el período 1997–2001 y el período del estudio completo (1989–2016), respectivamente. Analizamos la hipótesis de que pudiera existir una relación entre este incremento pasado del número de gallos lira y el impacto ecológico del temporal de invierno "Vivian". Nuestro estudio, y el fenómeno ecológico que fue concomitante con la tendencia demográfica observada, son una oportunidad de reforzar las iniciativas de seguimiento y gestión del gallo lira en los Alpes. ISSN: 1578–665 X eISSN: 2014–928 X

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Palabras clave: Estimaciones de abundancia, Factores de protocolo, Conteo en cantaderos, Gestión de fauna silvestre, Ave alpina, Galliformes Received: 10 IV 18; Conditional acceptance: 30 VII 18; Final acceptance: 18 I 19 Lucas Dumont, Estelle Lauer, Sébastien Zimmermann, Fédération Départementale des Chasseurs de l’Isère, 2 Allée de Palestine, 38610 Gières, France.– Pascal Roche, Fédération Départementale des Chasseurs de Haute–Savoie, 142 Impasse des Glaises, 74350 Villy–le–Pelloux, France.– Philippe Auliac, Fédération Départementale des Chasseurs de Savoie, Allée du Petit Bois, 14 Parc de l’Étalope, Bassens, 73025 Chambéry, France.– and Mathieu Sarasa, Fédération Nationale des Chasseurs, 13 rue du Général Leclerc, 92136 Issy les Moulineaux, France (former address); BEOPS, 1 Esplanade Compans Caffarelli, 31000 Toulouse, France (present address). Corresponding author: Mathieu Sarasa. E–mail: msarasa@beops.fr

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Introduction Evidence–based management of wild species requires reliable monitoring protocols and accessible scientific information (Walsh et al., 2004; Walsh et al., 2015). Many census methods have been developed, some to provide abundance estimates and others to measure population trends (Gregory et al., 2004; Meriggi et al., 2008; Schwarz and Seber, 1999). Trend evaluations may give some advantages in reducing interpretation biases that are sometimes associated with using abundance indices (Thompson, 2002). However, abundance estimates are still essential in standard management practices of many exploited species, such as evaluating the level of sustainable harvesting in game species. The methods producing abundance estimates are nonetheless based on key assumptions. They commonly assume that all individuals of a given territory are counted, which represents a detection probability equal to 1, constant over time. It is widely agreed that such a probability is nearly impossible to reach in standard monitoring fieldwork (Walsh et al., 2004; Simons et al., 2007; Jacob et al., 2010; Fremgen et al., 2016; Baumgardt et al., 2017). The black grouse (Tetrao tetrix) has a wide distribution range, being found in numerous countries (Storch, 2007). When present, it is a flagship species in the mountainous and boreal habitats of Europe, such as the Alps and Scandinavia, although some populations are also present in lowland habitats like Belgium, the Netherlands, Great Britain, Poland and north–west Germany (Storch, 2000). Population decreases were recorded in the last half–century, particularly in lowland habitats and in its south–western distribution range (e.g. Belgium: Keulen et al., 2005; Netherlands: Larsson et al., 2008; Great Britain: Baines and Hudson, 1995; Sim et al., 2008; Poland: Merta et al., 2009; Germany: Ludwig et al., 2008, 2009). Although some local recoveries have recently been described (Scridel et al., 2017), in Europe, the available estimates of several demographic traits, such as survival rates and fledglings per hen, are consistent with these previous population decreases (Jahren et al., 2016; Gée et al., 2018). Male lek–count is a widespread method for monitoring bird species such as the black grouse where males display for females during the breeding season (Cayford and Walker, 1991; Baines, 1996; Gregory et al., 2004; Chamberlain et al., 2012). Population monitoring of black grouse using lek–count data in the French Alps started in the late 1970's for management purposes (Ellison et al., 1988; Gehin and Montadert, 2016). Data from this standard monitoring of tetraonid species in France are managed by the 'Observatoire des Galliformes de Montagne' (Mountain Galliformes Observatory; e.g. Gehin and Montadert 2016). The monitoring of black grouse in the French Alps suggests contrasted results, i.e. sectors with decreasing, stable or increasing abundance of birds, depending on the considered local area (Ellison and Magnani, 1985; Bernard–Laurent, 1994; Amblard and Montadert, 2017). However, these patterns of abundance of birds are based on

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ambitious methodological assumptions and implicitly rule out residual links between the number of counted males and cofactors (procedural, geographical and meteorological cofactors, detailed hereafter). Some studies of Galliformes and other species have emphasized that the results of counts were affected by cofactors due to links between a visibility bias and the detection probability of the focal species (Baumgardt et al., 2017). These cofactors may be, for example, meteorological factors such as wind (Baines, 1996; Simons et al., 2007; Drummer et al., 2011; Sadoti et al., 2016), precipitation (Anderson et al., 2015; Baines, 1996), temperature (O’Connor and Hicks, 1980; Zimmerman and Gutiérrez, 2007) and snow (Brubaker et al., 2013; Fremgen et al., 2016). Cofactors affecting the results of counts in wild species may also include procedural factors such as the number of observers (Sarasa and Sarasa, 2013), the observers' experience (Alldredge et al., 2007; Garel et al., 2005; Hancock et al., 1999; Jiguet, 2009), timing organization of counts (Cayford and Walker, 1991; Sim et al., 2008; Monroe et al., 2016; Baumgardt et al., 2017), count day (Baines, 1996; Cayford and Walker, 1991; Gregersen and Gregersen, 2014), count year (Hovi et al., 1996) and count site (Anderson et al., 2015; Baines and Richardson, 2007). Under–estimating key cofactors might thus expose the interpretations derived from count data to potential bias in abundance and trend estimates (Simons et al., 2007; Monroe et al., 2016), particularly when tracking population changes across annual intervals (Blomberg et al., 2013). In this study we tested for potential links between lek–count data from black grouse monitoring and several recorded cofactors in the northern French Alps. As the guidelines for standard monitoring of black grouse are partly focused on minimizing heterogeneity associated with meteorological and procedural cofactors (Leonard, 1989), only weak and negligible links between lek–count data and cofactors might be expected. Alternatively, if the goal of minimizing the heterogeneity linked to the procedural, geographical and meteorological cofactors is not fully met when applying the guidelines for standard monitoring of black grouse, substantial links between lek–count data and cofactors might be observed. Abundance trends that are inferred from lek–counts may also be affected if cofactors are not controlled for in models. Material and methods Study area We analysed data from spring counts of black grouse in Isère, France, between 1989 and 2016. Isère is a department in the French Alps, located at the south–west limit of the distribution area for black grouse in Europe (Bernard–Laurent, 1994). It is one of the nine departments where black grouse are monitored in France (Bernard–Laurent, 1994; Amblard and Montadert, 2017). To our knowledge,

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this is the only study area combining two crucial points: first, the availability of detailed raw data over a long period in numerous sectors and conserved for years in a standardized format after counts, and second, the financial support required to enable the development of this study. The sectors for monitoring black grouse were designed according to the topography within the altitudinal limits of black grouse distribution recorded in the French Alps (1,400 to 2,300 m a.s.l.; Bernard–Laurent, 1994). This altitudinal range is the subalpine zone where black grouse males display reproductive behaviour in open microhabitats such as ridges, clearings and meadows with few or no trees (Magnani, 1988; Bernard–Laurent, 1994). The mean size (± SD) of the monitored sectors was 230 ha (± 10.6; n = 133). Their size and delimitation combine the need for their surveillance within the two first hours of the males' display (OGM, 2004) and the accessibility and natural barriers of the sectors. The sectors were assumed to reveal the abundance of black grouse (OGM, 2004) and showed highly variable abundances, suggesting they may also have highly variable quality for black grouse, within the theoretical black grouse distribution in Isère. Lek–count data All the prospected sectors were monitored by applying the usual protocols for black grouse monitoring in France (described in Leonard, 1989; Montadert, 2016). Counts were conducted between the last week of April and the end of May. One or more observers, assumed to be the 'right' number of observers to survey the considered area (Ellison and Magnani, 1985), patrolled the sector, or remained in a fixed position if the layout of the area allowed, performing the count from early morning, before sunrise, and until no later than two hours after sunrise (Montadert, 2016). Male black grouse are known to arrive at lek site in the Alps at about three quarters of an hour before sunrise (Couturier and Couturier, 1980). In practice, logistical constraints resulted in some variability in some sectors during the study period. The spread of observers over the sectors aimed to ensure a survey of the entire area, although exact details were not available in a standardized format after counts. Neither was information available on the experience of observers. All cocks seen and/or heard were counted to obtain a total estimate of the number of male black grouse, each associated with a monitored sector on a monitoring day. At the end of the counts, the location and time of each display were noted to reduce or avoid the risk of double counting of birds, although the spatial information of displays was not conserved in a standardized format after counts. Spatially joined sectors were generally counted on the same day (Montadert, 2016) to minimize the risk of double counting at the edge of sectors. Counting was avoided during bad weather conditions i.e. heavy rain, strong winds, or dense fog (Leonard, 1989; Baines, 1996; Montadert, 2016).

All the lek–count surveys considered in this study were organized and performed by the hunter's federation of Isère (FDC 38). Cofactors Procedural cofactors Several procedural cofactors were noted on summary sheets during the lek–count: identification of the monitored sector (to connect the data to geographical cofactors; detailed hereafter), day and year, hour of first observation of a displaying male, counting time, and number of observers. The diversity of the cofactor values revealed the variability of fieldwork in practice (see results). Geographical cofactors Using the identification of the monitored sectors, GIS layer and Quantum GIS software (QGIS Development Team, 2017), we associated geographical cofactors with the lek–count data: surface of the sector, latitude, longitude, altitude, and exposure of the centroid of each sector. Exposure was included by two cofactors in the models (Pedersen et al., 2014): (1) northing, which was expressed as an index of 'north–facing–ness' (Nor) using the formula Nor = cos[radians(angle)], such that Nor varied from 1 (due north) to –1 (due south); and (2) easting, which was expressed as an index of 'east–facing–ness' (Ea) using the formula Ea = sin[radians(angle)], such that Ea varied from 1 (due east) to –1 (due west). The centroid of each sector was a square of 30 x 30 m. Meteorological cofactors Meteorological data from Météo France was also added to our database. Detailed local meteorological data were not available because not all monitored sectors have a meteorological station. Nevertheless, two meteorological stations of Météo France are located within the altitudinal stratum of the distribution of black grouse in Isère and have been recording data since 1989: Saint Christophe (1,570 m a.s.l.; 44º 56' 41.971200'' N, 6º 11' 16.764000'' E, WGS84) and Alpe d’Huez (1,860 m a.s.l.; 45º 5' 13.189200'' N, 6º 5' 5.960400'' E, WGS84). The available data from these meteorological stations was averaged per day to obtain a mean value, proxies of the meteorological parameters at the altitudinal stratum of the distribution of black grouse in Isère. For both meteorological stations, the available data are from 1989 to 2016 for the following meteorological parameters: accumulated daily rainfall (in mm), minimum daily temperature (in ºC), accumulated daily fresh snow (in cm), and total snow depth (in cm). For temperature, we used the minimum daily temperature, first, because it usually occurs just before sunrise (Reicosky et al., 1989), the period of the day at which the display of black grouse starts (Couturier and Couturier, 1980), and second, because the time spent displaying is negatively correlated with temperature (Baines,

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France

Alpe d'Huez Saint Christophe

Meteorological stations References sites FDC38 Census sectors FDC38 Altitudinal stratum

20 km

Fig. 1. The study area in Isère France (right panel), with the census sectors, the reference sites and the meteorological stations considered in this study, within the altitudinal range of the black grouse (1,400–2,300 m a.s.l.; Bernard–Laurent, 1994). Fig. 1. Zona del estudio en Isère (Francia) (panel derecho) donde se muestran los sectores censados, las zonas de referencia y las estaciones meteorológicas consideradas en este estudio, en el rango altitudinal del gallo lira (1.400–2.300 m s.n.m.; Bernard–Laurent, 1994).

1996). During the period of black grouse lek–count, more than 99 % of the data for snow was equal to zero. Thus, the analysis of a potential link between the snow factor and the monitoring of black grouse was beyond the scope of our dataset. Wind data from the considered meteorological stations was only available from 2005 onwards. In our analysis we considered the maximum mean wind speed recorded in a 10–minute period between 6 a.m. and 7 a.m., the time of sunrise at which the first observations of black grouse generally occurred in our study. Strong windy days were avoided in accordance with the methodological recommendations (Leonard, 1989; Montadert, 2016). Wind data thus express breeze conditions but not windy weather. Analysis Standardization of data To facilitate data analysis, the days were converted into Julian days (number of days after January first), the hour of first observation of a displaying male was converted into number of minutes after midnight, and counting time was converted into minutes. Data were not available for at least one parameter required for analysis on nearly half of the summary sheets. As the number of complete summary sheets was high (57 %) and to avoid the use of potentially question-

able inferences of missing data, incomplete summary sheets were removed from our database. Thus, our database contains data from 1,040 counts made in 133 sectors monitored by the FDC 38, with a mean of 4.56 counted cocks per sector (SD = ± 3.38). Thirty–eight of these 133 sectors, constitute six reference sites that are used in Isère for the usual monitoring of black grouse (fig. 1). To assess the potential effects of the spatial scale and the abundance of available information on inferred biological signals and thus on the monitoring of black grouse, we built two datasets, allowing direct comparison of the results between two spatial scales. The first, ‘Total Isère’, included the complete information of lek–count available, monitored by FDC 38 in Isère from 1989 to 2016 (n = 1,040). The second, ‘Reference Sites’, included only the complete information of lek–counts available from 1992 to 2016 for the sectors located in the six reference sites monitored by FDC 38 (n = 349). Statistical analysis For both databases (Total Isère and Reference sites) we built and compared generalized additive mixed models (GAMMs, Wood, 2006) of the number of counted cocks using the variables listed in table 1. Sectors were included as random effect (Wood, 2006). Lindsey (1999) highlighted that one should

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Table 1. List of factors (F), with available data, potentially associated with lek–counts of black grouse in Isère, France. Tabla 1. Lista de factores (F), con disponibilidad de datos, potencialmente asociados con los recuentos en cantaderos de gallo lira en Isère, Francia.

F Description

Year of the count

Alt

Altitude of the prospected sector’s

Julian day of the count (number of

centroid (m)

days after first January)

Ea Easting: cos of the exposition of the

Hour of first observation of the count

(min after midnight)

Nor Northing: sin of the exposition of the

sector’s centroid (rad)

CT Count time (min)

sector’s centroid (rad)

NO Number of observers during the count

Daily rainfall (mm)

Surface of the prospected sector (ha)

Daily minimal temperature (°C)

Latitude of the prospected sector’s

Maximum mean wind speed between

centroid (WGS84)

6 a.m. and 7 a.m. (m/s)

Longitude of the prospected sector’s

centroid (WGS84)

only look for overdispersion if the deviance is at least twice the number of degrees of freedom. In our case study, the ratio estimating the degree of overdispersion was 1.66, thus lower than 2. We therefore used Poisson distribution models without problematic overdispersion and the analyses displayed an acceptable fit to the data. Model selection was based on Akaike Information Criterion corrected for small sample size (AICc; Burnham and Anderson, 2002). The relative importance (RI) of explanatory variables was estimated to quantify the importance of each factor within the set of models (Burnham and Anderson, 2002; Anderson and Burnham, 2002). Analysis, including breeze (Br), was performed using a short–term subset of the data because this factor has only been available since 2005. The main variables and their links with the counts of black grouse in our study area were already identified in the results from this first analytical step (Wood, 2006; see results). Nonetheless, to provide the results on the trends in a more reader–friendly format, for illustrative purposes we performed a trend analysis (Knape, 2016) as a second analytical step, once the best model was identified. We used the long–term dataset of counted black grouse and the default parameters of the models (Knape, 2016) to focus the illustration on the most reliable signs of population change (Knape, 2016) and not on short–term stochastic fluctuations that are beyond the scope of this study. More specifically, it allowed a graphic comparison of the population trends that would be inferred considering, or not, the factors

underlined by the best model for the data at hand. All analysis were performed using the 'R' statistical software (R Core Team, 2017), the 'mgcv' (Wood, 2006), and 'poptrend' (Knape, 2016) packages. Results Total Isère For Total Isère, the number of black grouse counted was linked to procedural and meteorological factors. Model selection suggested that the best model for the long–term data at hand should include as explanatory factors: the surface of the sector, the counting time, the number of observers, the hour of first observation of a displaying male, the day of the count, the year, the daily minimum temperature and the 'north–facing–ness' (R² = 0.32; table 2; fig. 2). The estimates of cocks increased with the surface of the considered sector up to a threshold of between 350 ha to 400 ha, beyond which it remained averagely stable (fig. 2F). The estimates also increased with counting time (fig. 2E) and with the number of observers up to a threshold of five or six observers, beyond which it appeared averagely stable (fig. 2A). The number of counted cocks decreased when the hour of first observation increased (fig. 2B) and when ‘north–facing–ness' increased (fig. 2G). The estimates of cocks appeared to decrease from the 10th to the 20th of May and then to increase up to the end of May (fig. 2C). The observed abundance of cocks varied non–linearly with the year (fig. 2D) and increased with

Animal Biodiversity and Conservation 42.2 (2019)

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Table 2. Model selection for factors associated with the number of black grouse Tetrao tetrix counted during lek–count surveys for Total Isère between 1989 and 2016: n, sample size; K, number of estimated parameters; AICc, Akaike's Information Criterion corrected for small sample size; ΔAICc, difference of AICc between the model and the most–parsimonious model of the set; L(gi/x), likelihood of the model to be the best model of the tested models; Wi, Akaike weight of the model; R², proportion of the variance in the data explained by the model; RI, relative importance of each factor. Only the ten best models are reported (Burnham and Anderson, 2002). (For abbreviations see table 1). Tabla 2. Selección de modelos para los factores asociados con el número de gallos lira Tetrao tetrix contados durante los estudios de seguimiento en cantaderos de todo Isère entre 1989 y 2016: n, tamaño muestral; K, número de parámetros estimados; AICc, criterio de información de Akaike corregido para pequeños tamaños muestrales; ΔAICc, diferencia de AICc entre el modelo y el modelo más parsimonioso del conjunto; L(gi/x), probabilidad de que el modelo sea el mejor modelo de los modelos estudiados; Wi: peso de Akaike del modelo; R², proporción de la varianza en los datos explicada por el modelo; RI, importancia relativa de cada factor. Solo se muestran los diez mejores modelos (Burnham y Anderson, 2002). (Para consultar las abreviaciones, véase la tabla 1).

Model

AICc

NO+H+Y+D+T+CT+S+Nor

1040

1804.92

0.33

0.32 NO

NO+H+Y+D+T+CT+S+Nor+R

1040

1806.79

1.87

0.39

0.13

0.32 H

NO+H+Y+D+T+CT+S

1040

1806.80

1.88

0.39

0.13

0.30 Y

NO+H+Y+D+T+CT+S+Nor+Ea

1040

1807.73

2.81

0.25

0.08

0.33 D

0.99

NO+H+Y+D+T+CT+S+R

1040

1808.66

3.74

0.15

0.05

0.30 T

0.99

NO+H+Y+D+T+CT+S+Nor+Alt

1040

1809.51

4.59

0.10

0.03

0.32 CT 0.97

NO+H+Y+D+T+CT+S+Nor+R+Ea 1040

1809.63

4.71

0.10

0.03

0.33 S

NO+H+Y+D+T+CT+S+Ea

1040

1810.04

5.12

0.08

0.03

0.31 Nor 0.72

NO+H+Y+D+T+CT+S+Nor+R+Alt 1040

1810.75

5.83

0.05

0.02

0.32 R

NO+H+Y+D+T+CT+S+Alt

1810.80

5.88

0.05

0.02

0.29 Ea 0.19

1040

ΔAICc L(gi/x) Wi

R²

0.95 0.28

Alt 0.10 La 0.01 Lo 0.01

the daily minimum temperature (fig. 2F). The tested factors that were not included in the best model for the long–term data at hand were: daily rainfall, latitude, longitude, altitude and ‘east–facing–ness’, although the daily rainfall factor was included in the second best model with substantial evidence support (table 2). A simplified version of the best model without a 'north–facing–ness' was the third best model, with a ∆AICc lower than 2 units. The analysis of the subset that included the breeze factor provided a nuanced version of these results. Model selection suggested that the best model for the short–term data at hand should include as explanatory factors the surface of the sector, the number of observers, the hour of first observation of displaying male, the day of the count, the year, the breeze factor and the 'north–facing–ness' (R² = 0.37; table 3). The patterns of counted cocks globally matched with those detailed above for long–term data (fig. 2–3). In addition, the number of counted cocks appeared positively linked to the breeze factor up to a threshold

of 3–4 m/s, beyond which it remained stable (fig. 3G). The number of counted cocks decreased when ‘north– facing–ness’ increased (fig. 3F). The factors counting time and daily minimum temperature, among others, were not included in the best model for short term data at hand (table 3). Reference sites For the Reference Sites, the number of counted black grouse was also linked to procedural and meteorological factors. Model selection suggested that the best model for the long–term data at hand should include as explanatory factors the number of observers, the hour of first observation of the displaying male, the day of the count, the year, and the minimum temperature (R² = 0.41; table 4; fig. 2). The links between the number of counted cocks and the associated factors closely match those detailed above for long–term data in Total Isère (fig. 2I–2M). Nonetheless, the link to daily minimum

Dumont et al.

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temperature suggested a potential non–linearity associated with a larger confidence interval for the higher values of temperatures (fig. 2M). In addition, the abundance of cocks with years in Reference Sites exhibited stronger variations than those in Total Isère (fig. 2L, 2D). Analysis of the subset for short–term data at hand that included the breeze factor provided a nuanced version of said results albeit with a lower capacity to distinguish the best model. The best model included as explanatory factors the number of observers, the hour of first observation of a displaying male, the day of the count, the year, the surface of the sector, the 'north–facing–ness', and altitude (R² = 0.63; table 5; fig. 3). Nevertheless, six simplified versions of this model, with and without the surface, the day, the ‘'north–facing–ness' and the altitude appeared as having substantial support for the data at hand (table 5; fig. 3). Population trends When omitting the important cofactors for the long–term dataset, the population trends for both Total Isère and Reference Sites appeared without any significant increase or decrease in abundance of counted cocks (fig. 4A, 4B). However, when considering the cofactors highlighted in the best models, different patterns emerged (fig. 4C, 4D). In Total Isère, the number of male black grouse showed two periods of overall stability separated by a significant mean increase in population of +17 % between 1997 and 2001 (CI 95 %: +5.3 % – +29 %). In Reference Sites, a decrease in the number of male black grouse appeared from 1992 to 1995 but this result is supported by the data that was recorded for 1992 only. Two periods of overall stability (1995–98 and 2001–16) were separated by a significant mean increase in population of +46% between 1998 and 2001 (CI 95 %: +18 % – +80 %). Over the entire period of the study (1989–2016), the larger dataset, on Total Isère, suggested a mean increase in population of +47 % (CI 95 %: +16 % – +87 %).

Discussion This study provides the first quantitative evidence, based on standard monitoring, that procedural and environmental cofactors modulate the estimates of abundance and population trends of the black grouse, a flagship species of alpine and northern environments. Black grouse monitoring and cofactors On the basis of the available information, our models suggest that the numbers of counted males in Isère are predominantly associated with procedural factors, and more specifically, with the number of observers, the hour of first observation of a displaying male, and the day of count, in addition to the year of the count. Additional details appeared, depending on the available information, the temporal scale and the spatial scale of the analysis. The analysis with the larger data set on Total Isère suggests that counting time, surface of the monitored sector, northing and temperature are also associated with the number of counted males. Nevertheless, the role of counting time, northing and the surface of sectors was not detected when using the smaller data set on Reference Sites. Furthermore, when considering the breeze factor, available on a smaller temporal scale, the relationships with counting time and temperature were not highlighted while the breeze factor was included in the best model of Total Isère, and altitude in the best model of Reference Sites. The amount of available information, the spatial scale and the temporal scale are thus important determinants of links between variables and of observed trends. When the sample size (n) is small compared to the number of estimable parameters in the approximating model (K) (see Burnham and Anderson, 2002), secondary, although important, cofactors are not necessarily detected, in accordance with the smaller spatial and temporal scales in our case study. Thus, sufficient data must be collected for a considered question so as to integrate a reasonable number of potential cofactors in the analyses and to analyze data at the

Fig. 2. Evolution of the number of black grouse cocks counted for Total Isère (left) and in Reference Sites (right) between 1989 and 2016, in relation to each cofactor of their respective best models (models with the lowest AICc, see tables 2 and 4). The solid lines represent the estimated patterns and shaded areas indicate 95 % confidence intervals. The left–hand y–axis represents the centred values and the right–hand axis represents the estimated abundance of male black grouse. Mean of predicted values: Total Isère = 4.32 birds; Reference Sites = 4.32 birds. Fig. 2. Evolución del número de machos de gallo lira contados en Total Isère (izquierda) y en las zonas de referencia (derecha) entre 1989 y 2016, en relación con cada cofactor de sus respectivos mejores modelos (modelos con el menor AICc, véanse tablas 2 y 4). Las líneas continuas representan los patrones estimados y las zonas sombreadas en gris indican el intervalo de confianza del 95 %. El eje de las Y de la izquierda representa los valores centrados y el de la derecha, las abundancias estimadas de machos de gallo lira. Media de los valores previstos: Total Isère = 4,32 aves; Zonas de Referencia = 4,32 aves.

Animal Biodiversity and Conservation 42.2 (2019)

Total Isère

0.6

4.9

0.5

4.8

0.3

4.6

0.2

4.5

4.3

–0.1

6 8 10 Number of observers

4.8

4.4

0.1

–0.4

4.0

–0.5

–0.8

3.6 250 300 350 400 450 500 550 Hour of first observation

0.6

–0.3 115

125

0.3

135 Day

1.0

5.5

4.4

0.2

4.5

4.1

–0.2

1990 1995 2000 2005 2010 2015 Year

0.3

4.6

0.1

4.4

–0.1

4.2

4.0 120 125 130 135 140 145 Day

0.5

4.9

0.2

4.6

–0.1

4.3

–0.4

4.0 1995

2000

2005 2010 2015 Year

300

0.5

4.9

0.2

4.6

–0.1

4.3

–0.4

350 400 450 500 Hour of first observation

5.0

150 200 250 Count time

3.5

0.6

4.1

100

4.0

–1.1

5.0

4.7

4.3

4.0 100

G 0.2

300 500 Surface

700 4.5 4.2

–0.1 –0.4

–1.0

5.0

4 6 8 Number of observers

4.5

300

4.5

–0.1 –0.3

0.7

145

0.1

4.2 2

0.4

5.1

5.2

0.3

Reference Sites 0.8

0.9

235

–0.5

0.0 Northing

0.5

1.0

3.9

0.2

4.5

4.3

–0.2

4.1

–5

0 5 Temperature

M 0.2

4.5

4.3

–0.2

4.1

–0.4

5 Temperature

3.9

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236

appropriate spatial and temporal scale. Nonetheless, as in other species (Sarasa and Sarasa, 2013; Monroe et al., 2016), our results underlined that the counts of black grouse are associated with procedural, meteorological and geographic cofactors, in addition to the expected links to the interannual variations of birds. These results complement those of previous studies on counts of black grouse (Baines 1996). It should be noted that in this study, the time of first observation might be a cofactor more connected to the timing count and logistical constraints than to the timing of the bird’s behaviour. The high estimates of counted males are preponderantly associated with counts performed by 6–7 observers per sector, with an early–morning first observation (between 4 a.m. and 5 a.m.), during long counting times (2h 30'–3h), in moderately large sectors (300–400 ha) and usually on days with mild temperatures (5°C or above) or a light breeze (3–5 m/s). The high estimates of counted males were usually recorded early in the counting season (during the last days of April) although a few counts occurred in the late counting season (later than 20th of May). These few counts in the late season might be too scarce to accurately reflect the range values of this procedural cofactor, highlighting a need for further studies on the phenology of reproductive displaying of black grouse in the Alps. The results referring to time and day mainly agree with those described in Sweden and Wales (Cayford and Walker, 1991; Borecha et al., 2017) even though the premating and peak mating periods of black grouse in the Alps must be further studied (perhaps in relation to the Normalized Difference Vegetation Index) and compared to northern populations. The predominant importance in our results of procedural cofactors rather than meteorological cofactors suggests that the methodological recommendations to observers (who must avoid adverse meteorological conditions during lek–counts) might successfully reduce potential noise related to meteorology in lek–count data. Another hypothesis suggests that the meteorological data available (only at two meteorological stations within the altitudinal range of the black grouse) may reveal macro–variations at the scale of Isère but might be too limited to allow a full analysis of the potential effects

of very local weather on lek–count estimates. The fact that most of the geographical cofactors (altitude, longitude, latitude, easting) were not selected in the best models also suggests a predominant role of procedural and very local factors on local estimates of black grouse rather than biogeographical factors. The negative link observed between 'north–facing–ness' and the number of black grouse cocks might appear paradoxical in a flagship species of mountainous and boreal habitats. However, further studies are required to explore potential associations to key resources at a very local level, as already reported in other grouse species (Storch, 1993). Although beyond the scope and the potentiality of our dataset, further studies might go into other factors able to induce over–estimations (e.g. double counting) or sub–estimations of abundance (e.g. variable experience and age of observers, Hancock et al., 1999; Farmer et al., 2014; variable male lek count attendance in black grouse, Baines, 1996; acoustic limitations, Simons et al. 2007; low or variable probability of detection in tetraonids, Zimmerman et al., 2007, Fremgen et al., 2016). Several observers are often required for fieldwork and current protocols already tend to minimize or avoid double counts (see Method section). However, the potential problems related to the incomplete probability of detection or to the variability of observers are to date unmitigated. The relative importance of the heterogeneity of the sites, in particular to local weather and microhabitats that modulate the probability of detection and the suitability of the habitat for birds is also an open question. Thus, in addition to integrating the important cofactors that were highlighted in our results into monitoring programs, further studies on other potential cofactors might be required to continue improving the monitoring program of the black grouse. Population trends Our results on the long–term population trends of abundance of counted males highlight that omitting cofactors creates a bias in population trends. When considering cofactors, a marked increase of +17 % (CI 95 %: +5.3 %; +29 %) of the counted black grouse appeared between 1997 and 2001, followed by a period

Fig. 3. Evolution of number of black grouse cocks counted for Total Isère (left) and in Reference Sites (right) between 2005 and 2016 in relation to each cofactor of their respective best models (models with the lowest AICc, see tables 3 and 5). The solid lines represent the estimated patterns and shaded areas indicate 95 % confidence intervals. The left–hand y–axis represents the centred values and the right–hand axis represents the estimated abundance of male black grouse. Mean of predicted values: Total Isère = 4.54 birds; Reference Sites = 4.64 birds. Fig. 3. Evolución del número de machos de gallo lira contados e Total Isère (izquierda) y en las Zonas de Referencia (derecha) entre 2005 y 2016, en relación con cada cofactor de sus respectivos mejores modelos (modelos con el menor AICc, véanse las tablas 3 y 5). Las líneas continuas representan los patrones estimados y las zonas sombreadas en gris indican el intervalo de confianza del 95 %. El eje de las Y de la izquierda representa los valores centrados y el de la derecha, las abundancias estimadas de machos de gallo lira. Media de los valores previstos:Total Isère = 4,54 aves; Zonas de Referencia = 4,64 aves.

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Reference Sites

Total Isère

A 1.0

5.6

1.0

5.6

0.6

5.2

0.6

5.2

0.2

4.8

0.2

4.8

4.4

–0.2

2 3 4 5 6 7 Number of observers

0.5

5.0

4.0

–1.0

3.5

200 300 400 500 Hour of first observation

0.4

4.9

0 –0.4

0.4

–0.4 –0.8 2006

2010 Year

–0.2

4.4 120

125

130 Day

135

140

4.1

2006 2008 2010 2012 2014 2016 Year 0.6

5.2

4.6

–0.6

3.8

4.0 100

4.7

M 0.3

4.5

300 500 Surface

700 4.9 4.6

–0.3

4.3

–0.6

4.0 –1.0

–0.5

0.0 Northing

0.5

1.0

0.5

5.1

0.1

4.7

–0.3

4.3

–0.7

3.9

1400 1500 1600 1700 1800 1900 Altitude

G 0.4

4.9

4.5

–0.4

4.1 2

5 Wind

4.1

–0.5

4.1

1.0

4.7

4.4

–0.4 –1.0

0.5

0.1

4.0

4.3 0.0 Northing

5.0

4.7

–0.2 –0.5

3.5

–0.2

4.2

F 0.2

0.4

350 400 450 500 Hour of first observation

5.0

–0.3 700

300

0.1

4.6

300 500 Surface

4.0

4.4

0 –0.6

4.5

0.4

5.0

0.3

100

4.4

5.0

–0.5

3.6

2014

0.5

–1.0

4.1

4.8

–0.5

4.5

115 120 125 130 135 140 145 Day

2 3 4 5 6 7 Number of observers

4.5

–0.5

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Table 3. Model selection for factors associated with the number of black grouse Tetrao tetrix counted during lek–count surveys for Total Isère between 2005 and 2016. Only the ten best models are reported (Burnham and Anderson, 2002). (For abbreviations see tables 1 and 2). Tabla 3. Selección de modelos para los factores asociados con el número de gallos lira Tetrao tetrix contados durante los estudios de seguimiento en cantaderos de todo Isère entre 2005 y 2016. Solo se muestran los diez mejores modelos (Burnham y Anderson, 2002). (Para consultar las abreviaciones, véase las tablas 1 y 2).

Model

AICc

NO+H+Y+D+Br+S+Nor

447

744.92

ΔAICc L(gi/x) 0

0.47

0.37

R²

NO+H+Y+D+Br+S

447

747.64

2.73

0.26

0.12

NO+H+Y+D+Br+CT+S+Nor

447

749.16

4.25

0.12

0.06

NO+H+Y+D+Br+CT+S

447

750.01

5.09

0.08

0.04

NO+H+Y+D+Br+S+Nor+Alt

447

750.20

5.29

0.07

0.03

NO+H+Y+D+Br+Nor

447

750.56

5.64

0.06

0.03

0.30

S 0.92

NO+H+Y+D+Br+T+S

447

750.77

5.85

0.05

0.03

0.34

Nor 0.72

NO+H+Y+D+Br+S+Nor+Ea

447

751.25

6.33

0.04

0.02

0.37

CT 0.17

NO+H+Y+D+Br+S+Ea

447

751.65

6.74

0.03

0.02

0.34

NO+H+Y+D+Br+S+R

447

751.91

7.00

0.03

0.01

0.34

0.37

0.34

0.37

Br 0.98

0.08

Ea 0.07

Alt 0.07 R 0.06 La 0.01 Lo 0.01

Table 4. Model selection for factors associated with the number of black grouse Tetrao tetrix counted during lek–count surveys in Reference Sites between 1992 and 2016. Only the ten best models are reported (Burnham and Anderson, 2002). (For abbreviations see tables 1 and 2). Tabla 4. Selección de modelos para los factores asociados con el número de gallos lira Tetrao tetrix contados durante los estudios de seguimiento en cantaderos de las zonas de referencia entre 1992 y 2016. Solo se muestran los diez mejores modelos (Burnham y Anderson, 2002). (Para consultar las abreviaciones, véase las tablas 1 y 2). Model

AICc

ΔAICc

L(gi/x)

NO+H+Y+T+D

349

628.32

NO+H+Y+T+Nor

349

629.61

1.29

NO+H+Y+T+D+Ea

349

630.99

2.67

NO+H+Y+D

349

631.14

NO+H+Y+T+D+Nor+Alt

349

631.20

NO+H+Y+T+D+S+Nor

349

NO+H+Y+T+D+S

349

NO+H+Y+T+D+CT

349

NO+H+Y+T+D+R

349

NO+H+Y+D+Nor

349

R²

0.21

0.41 NO

0.53

0.11

0.43 H

0.26

0.05

0.42 Y

2.81

0.24

0.05

0.41 D

2.88

0.24

0.05

0.48 T

0.79

631.61

3.29

0.19

0.04

0.52 Nor 0.40

631.83

3.50

0.17

0.04

0.47 Alt 0.20

632.16

3.84

0.15

0.03

0.41 S

632.34

4.01

0.13

0.03

0.41 Ea 0.19

632.36

4.04

0.13

0.03

0.42 CT 0.13

0.20

0.11

La 0.02 Lo 0.02

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Reference Sites

Total Isère

1.8 1.4

1.6 1.4

Trend

1.2

1.0

1.0 0.8

0.8

0.6

0.6 1990 1995 2000 2005 2010 2015 Year

1995

2000

2005 2010 Year

1995

2000

2005 Year

2015

B 1.8

1.4

1.6

1.2 Trend

1.4 Trend

1.2

1.0

0.8

0.6

0.6 1990 1995 2000 2005 2010 2015 Year

2010 2015

Fig. 4. Estimated long–term trends for black grouse from lek–count surveys between 1989 and 2016. A, omitting potential cofactors; B, the best model including cofactors.The solid line in these panels is the estimated long–term component of the trend, while the points are estimates of the trend with estimates of the random year effects superimposed. Estimates with random year effect and automatic degree of freedom, standardized with respect to the mean of the long–term component of the trends; confidence intervals (shaded area and vertical lines) are computed from the 2.5 % and 97.5 % quantiles of the bootstrap distributions; for periods where there is a significant increase or decrease in the trend at the 5 % level, the trend line is colored, respectively, in green and orange; periods where the curvature is significantly positive or negative are marked by green and orange rectangles at the bottom of the panels (Knape, 2016). For instance, an increase in the value from 1.0 to 1.2 suggests an increase of 20 % with respect to the mean. Fig. 4. Tendencia a largo plazo del gallo lira, estimada a partir de estudios de seguimiento por recuento en cantaderos entre 1989 y 2016. A, omitiendo posibles factores; B, el mejor modelo que incluye factores. La línea continua en estos paneles es el componente estimado a largo plazo de la tendencia, mientras que los puntos son estimaciones de la tendencia con estimaciones de los efectos aleatorios del año superpuestos. Estimaciones con efecto anual aleatorio y grado de libertad automático, estandarizado respecto al promedio de la componente a largo plazo de la tendencia; los intervalos de confianza (área sombreada y líneas verticales) se calculan a partir de los cuantiles del 2,5 % y del 97,5 % de las distribuciones de bootstrap; para los períodos donde hay un aumento o una disminución significativos de la tendencia en el nivel del 5 %, la línea de tendencia se indica en verde y naranja, respectivamente; los períodos en los que la curvatura es significativamente positiva o negativa están marcados con rectángulos verdes y naranjas en la parte inferior de los paneles (Knape, 2016). Por ejemplo, un incremento del valor de 1,0 a 1,2 sugiere un incremento del 20 % respecto al valor medio.

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Table 5. Model selection for factors associated with the number of black grouse Tetrao tetrix counted during lek–count surveys in Reference Sites between 2006 and 2016. Only the ten best models are reported (Burnham and Anderson, 2002). (For abbreviations see tables 1 and 2). Tabla 5. Selección de modelos para los factores asociados con el número de gallos lira Tetrao tetrix contados durante los estudios de seguimiento en cantaderos de las zonas de referencia entre 2006 y 2016. Solo se muestran los diez mejores modelos (Burnham y Anderson, 2002). (Para consultar las abreviaciones, véase las tablas 1 y 2). Model

AICc

ΔAICc L(gi/x)

R²

NO+H+Y+D+S+Nor+Alt

186

338.54

0.08

0.63

NO+H+Y+D+S

186

338.63

0.09

0.95

0.08

0.51

NO+H+Y+D

186

339.09

0.55

0.76

0.06

0.40

0.95

NO+H+Y+D+S+Nor

186

339.43

0.89

0.64

0.05

0.59

0.62

NO+H+Y

186

339.65

1.11

0.57

0.05

0.39

0.62

NO+H+Y+S

186

339.92

1.38

0.50

0.04

0.49

Nor 0.46

NO+H+Y+D+S+Nor

186

340.14

1.61

0.45

0.04

0.60

Alt 0.32

NO+H+Y+D+S+Nor+Br

186

341.12

2.58

0.28

0.02

0.59

Br 0.17

NO+H+Y+D+Nor+Alt

186

341.32

2.78

0.25

0.02

0.57

La 0.10

NO+H+Y+Nor+Alt

186

341.74

3.20

0.20

0.02

0.55

Lo 0.10

CT 0.10 T 0.08 R 0.07 Ea 0.07

of overall stability up to 2016. This trend differs from the trends reported in some parts of Italy (Viterbi et al., 2015), the UK (Warren and Baines, 2008; Summers et al., 2010; Scridel et al., 2017) and Finland (Ludwig et al., 2006) but corresponds to the trend highlighted in the Mont Avic Natural Park, Italy (Chamberlain et al., 2012). The increases in abundance of black grouse observed in Isère (France) and Mont Avic (Italy) during the second half of the 1990s were concomitant with a major ecological phenomenon that affected the habitat of the black grouse in the north–western Alps at this time. In February 1990, the winter–storm Vivian strongly affected the altitudinal range in the Alps, with gales of up to 270 km/h at the Italian–Swiss border (Schüepp et al., 1994). This storm was considered one of the most devastating wind–storms of the 20th century, causing about 100 million m3 of wind throw damage in Europe (Cinotti, 1992), including about 8 million m3 in France (Cinotti, 1992). Damage from Vivian also affected the Alpine forests in the northern French Alps (Dorren et al., 2004). Following the passage of Vivian, a proliferation of arthropods was reported in the Alps, mainly studied in Switzerland (e.g. Wermelinger et al., 2002). Arthropods, such as the saproxylic species, exhibited a boosted abundance that was observed as a successional phenomenon (Wermelinger et al., 2002). The increased abundance of several species that are associated with

late stages of disturbed forest, such as Cerambycidae (Wermelinger et al., 2002, 2003), occurred during the second half of the 1990s. This corresponded to the period of increased abundance of black grouse that was highlighted in our results and in the Mont Avic Natural Park, Italy (Chamberlain et al., 2012). Ants are one group of the species associated with dead wood and the late stages of disturbed forests (Lempérière et al., 2002; Lempérière and Marage, 2010). They are also a key resource in habitat selection (Schweiger et al., 2012) and reproduction of black grouse (Ponce and Magnani, 1988; Ponce, 1992; Baines et al., 2017). Consequently, a plausible hypothesis for the marked increase in abundance of black grouse in Isère and other areas in the Alps, such as Mont Avic, at this time, would be a bottom–up stimulation of grouse abundance during the 1990s, induced by the cascading effects of the winter–storm Vivian. Although further studies might be necessary to verify this hypothesis, other factors with potentially positive links to abundance of grouse did not appeared as exhibiting noteworthy variations only during that period in the northern French Alps: forestry (Office National des Forêts, 2001); climate (Bigot and Rome, 2010); hunting (Magnani, 2009; Lauer and Magnani, 2013); and husbandry (Chatelier and Delatre, 2003). For instance, the hunting of black grouse was reduced in the French Alps, including

Animal Biodiversity and Conservation 42.2 (2019)

our study area, to less than half, following a rather regular pattern between 2000–2011 (see page 18 of Lauer and Magnani, 2013). This decrease over a long period of time fits poorly with the marked increase in grouse numbers during 1997–2001 period only. Further analysis integrating additional demographic data could add to the understanding of the overall increase in grouse numbers during the full study period (1989–2016). Further studies in alpine habitats are needed to determine the cascading links between uncleared gaps of windthrow timber, abundance of arthropods –particularly ants– and abundance of grouse. Such studies could provide promising perspectives for the integrated management of black grouse and their habitat, as the management of dead wood in forests might be a key factor in the dynamics of birds such as the black grouse. Conclusions Several cofactors, in particular procedural cofactors, are linked to lek–counts of male black grouse, and the studied population exhibited an increase that was previously not detected in France. This corresponds, however, to an ecological phenomenon previously reported in the north–western Alps. Thus, these findings might stimulate substantial improvements in the monitoring and management of black grouse through greater integration of the important cofactors in monitoring protocols and models of abundance. The concomitance between the observed trends and other ecological phenomena in the Alps questions the need for integrated management of the habitat, specifically on the potential links between dead wood, arthropod populations and black grouse populations. Acknowledgements We would like to thank the anonymous associate editor and two anonymous reviewers for helpful comments on an earlier version of the manuscript. We thank all the participants in the monitoring of black grouse in France, particularly the volunteers and professional observers from the federations of hunters (FDC, FRC), the ONF, and the ONCFS for their major role in said monitoring. We would also like to thank Robin Paya for his technical assistance with GIS and Maygrett Maher and Agnès Sarasa for editing the English of this manuscript. This study was made possible thanks to the support provided by the Conseil Régional d'Auvergne– Rhône–Alpes, the Fédération Nationale des Chasseurs (FNC–PSN–PR18–2014), and the hunters' federations FDC38, FDC74, FDC73, and FRC–Auvergne–Rhône– Alpes (FDC38–BEOPS–PR1–2017). References Alldredge, M. W., Simons, T. R., Pollock, K. H., 2007. Factors affecting aural detections of songbirds. Ecological Applications, 17(3): 948–955.

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Effects of the invasive crayfish Procambarus clarkii on growth and development of Pelophylax perezi tadpoles in field conditions S. Saura–Mas, L. Benejam

Saura–Mas, S., Benejam, L., 2019. Effects of the invasive crayfish Procambarus clarkii on growth and development of Pelophylax perezi tadpoles in field conditions. Animal Biodiversity and Conservation, 42.2: 245–252, https://doi.org/10.32800/abc.2019.42.0245 Abstract Effects of the invasive crayfish Procambarus clarkii on growth and development of Pelophylax perezi tadpoles in field conditions. Introduced predatory aquatic invertebrates may contribute to the global decline of amphibians as their larval are extremely vulnerable to predators. The objective of this study was to examine the effects of the predatory invasive crayfish Procambarus clarkii on the growth and development of native Iberian green frog tadpoles, Pelophylax perezi, in field conditions. We hypothesized that P. clarkii might affect P. perezi development by (a) inducing a delay in its metamorphosis and (b) reducing survival and mass of metamorphs. The experiment was developed in two ponds (with and without P. clarkii’s presence) in the Natural Park of Aiguamolls de l’Empordà (NE of the Iberian Peninsula). For each pond, groups of 10 tadpoles were randomly assigned to 15 cylindrical field enclosures. These enclosures avoided direct contact (i.e. predation) between both species. Our results suggest that, in field conditions, the presence of P. clarkii accelerates metamorphosis of P. perezi tadpoles. The higher growth rate of P. perezi through shorter larval periods could be the result of behavioural plasticity in response to the strong pressure imposed by P. clarkii. This conclusion would be in accordance with the hypothesis that phenotypic plasticity plays an important role in the conservation of P. perezi in front of biological invasions. Key words: Tadpole, Predator, Invasion, Non–native, Larval development, Fitness Resumen Efectos del cangrejo exótico, Procambarus clarkii, en el crecimiento y el desarrollo de los renacuajos de Pelophylax perezi en condiciones de campo. Los invertebrados acuáticos depredadores introducidos pueden contribuir a la disminución general de los anfibios, cuyas larvas son extremadamente vulnerables a los depredadores. El objetivo de este estudio fue examinar los efectos del cangrejo exótico, Procambarus clarkii, en el crecimiento y el desarrollo de los renacuajos de la rana verde ibérica (Pelophylax perezi) en condiciones de campo. Concretamente, planteamos la hipótesis de que P. clarkii podría afectar al desarrollo de P. perezi de dos formas: (a) induciendo un retraso en su metamorfosis y (b) reduciendo la supervivencia y la cantidad de renacuajos en fase de metamorfosis. El experimento se realizó en dos estanques (con y sin presencia de P. clarkii), en el Parque Natural Aiguamolls de l’Empordà (NE de la península ibérica). En cada estanque se introdujeron aleatoriamente grupos de 10 renacuajos en 15 cilindros cerrados de malla. Estos cilindros evitaron el contacto directo (es decir, la depredación) entre ambas especies. Nuestros resultados sugieren que, en condiciones de campo, la presencia de P. clarkii podría acelerar la metamorfosis de los renacuajos de P. perezi. La plasticidad en el comportamiento de P. perezi como respuesta a la fuerte presión ejercida por P. clarkii permite acelerar la tasa de crecimiento reduciendo los períodos larvales. Además, este estudio avalaría la hipótesis de que la plasticidad fenotípica juega un papel importante en la conservación de P. perezi frente a las invasiones biológicas. Palabras clave: Renacuajo, Depredador, Invasión, Alóctono, Desarrollo larvario, Eficacia biológica Received: 28 VI 18; Conditional acceptance: 28 X 18; Final acceptance: 24 I 19 Sandra Saura–Mas, CREAF (Center for Ecological Research and Forestry Applications), 08193 Cerdanyola del Vallès, Catalonia, Spain; Unit of Ecology, Department of Animal and Plant Biology and Ecology, Autonomous University of Barcelona, 08193 Bellaterra, Catalonia, Spain.– Lluís Benejam, Aquatic Ecology Group, University of Vic–Central University of Catalonia, Vic 08500, Catalonia, Spain. Corresponding author: Sandra Saura–Mas. E–mail: s.sauramas@creaf.uab.cat ISSN: 1578–665 X eISSN: 2014–928 X

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Introduction Global change, including biological invasions, habitat fragmentation and destruction, acidity, pollution (such as insecticides and herbicides) and emerging infectious diseases, is known to be causing amphibian declines worldwide (Houlahan et al., 2000; Saura–Mas et al., 2002; Collins and Storfer, 2003; Kats and Ferrer, 2003; Stuart et al., 2004; Hayes et al., 2010). It is difficult to determine the exact causes of the decline in amphibian populations as their demography is characterized by fluctuations in the number of individuals due to their biological dependence on environmental conditions and their meta–population traits. However, the joint action of natural fluctuations in populations’ and anthropogenic might result in local extinction more easily than either alone (Pechmann et al., 1991). Biological invasions are a major factors in global change due to their effects on the natural ecosystems (Vitousek et al., 1996; Garcia–Berthou, 2010; Simberloff et al., 2013). Humans have deliberately introduced animals outside their natural range for a variety of motives (commercial and non–commercial). As a result, allochthonous species often cause declines and even extinctions of native amphibian populations around the world (Carey et al., 2003; Lever, 2003). Invasive species can affect amphibians in aspects such as predation, competition, parasitism, and habitat disruption. The bullfrog (Lithobates catesbiana) is a native of North–America and considered to be one of the most harmful invasive species worldwide since it negatively affects native amphibians through competition and predation (Kats and Ferrer, 2003; Beebee and Griffiths, 2005; GISD, 2018). Other invasive species such as introduced predatory aquatic invertebrates and fishes are a concern for amphibian populations and may contribute to global amphibian decline because larval amphibians are extremely vulnerable to vertebrate and invertebrate predators (Alford and Richards, 1999). In recent years, one of the most important pests in Iberian Peninsula wetlands, streams and ponds has been the crayfish Procambarus clarkii, which has affected the ecosystem dynamics and caused large socio–economic damage, mainly to rice fields in these areas (Beja, 1996; Gutiérrez–Yurrita, 1997; Twardochleb et al., 2013; Carreira et al., 2014). P. clarkii is a native of North–America and it has become invasive in many continental aquatic Mediterranean ecosystems (Taylorab et al., 1996; Gonçalves et al., 2015). Several studies show that anuran tadpoles can detect chemical cues coming from predators and injured prey during predation events (Petranka et al., 1987; Schoeppner and Relyea, 2005; Fraker et al., 2009), Sublethal effects can thus be produced in amphibians by altering habitat conditions, behaviour, and development and growth (Anholt and Werner, 1995; Relyea, 2001, 2002; Rodríguez et al., 2005; Tejedo et al., 2010). The eradication of P. clarkii from most Mediterranean wetlands is nearly impossible due to its reproductive and invasive traits. It has been suggested that the species plays an important role in the decline of local populations of amphibians in such ecosystems because

of its predatory role, and particularly its direct or indirect effects on tadpoles (Renai and Gherardi, 2004; Rodríguez et al., 2005; Cruz et al., 2006). Many authors have described changes in behaviour, morphology and growth during metamorphosis processes of native tadpoles in interaction with P. clarkii (Almeida et al., 2011; Gonçalves et al., 2011). Nunes et al. (2014a) observed that eight of nine species of tadpoles changed their morphology or life history when reared with the fed dragonfly, but only four made such changes when reared with the fed crayfish, suggesting among–species variation in the ability to respond to a novel predator. Nowadays, one of the most important challenges in animal ecology is to know more about phenotypic plasticity of species in front of environmental changes. Many studies have been conducted to obtain more information on phenotypic plasticity of prey induced by predators. Some studies have focused in anuran species report a clear lack of response to invasive predators (Smith et al., 2007; Gomez–Mestre and Díaz–Paniagua, 2011; Vázquez et al., 2017), while other studies report different types of behavioural and morphological responses after a relatively short period of coexistence with the invasive predator (Kiesecker and Blaustein, 1997; Pearl et al., 2003; Almeida et al., 2011; Gonçalves et al., 2011; Pujol–Buxó et al., 2013; Nunes et al., 2014a). The phenotypic plasticity of prey in front of invasive predator species is also an important factor to take into account. Heritable phenotypic plasticity to native species might thus be a key step to understanding the effects of global environmental changes, such as biological invasions. The main objective of this study was to examine the effects of the presence of the predatory invasive crayfish P. clarkii (with non–lethal effects) on the growth and development of the Iberian green frog Pelophylax perezi in field conditions in the Natural Park of Aiguamolls de l’Empordà (NE of the Iberian Peninsula). Most previous studies examining the effects of invasive predators to anuran development have been developed under lab or mesocosm conditions. An innovative aspect of our study is that the experiment was developed under field conditions. We aimed to answer the following questions: first, does the presence of P. clarkii hasten metamorphosis of P. perezi? We hypothesized that the time to metamorphosis would be delayed because of diminished activity and consequent lower energy intake (Tejedo et al., 2010; Touchon et al., 2015); and second, does the presence of P. clarkii affect survival, growth rate and mass of metamorphs? We hypothesize that detection of the predator P. clarkii could induce differences in feeding activity, resulting in smaller sizes at metamorphosis (Orizaola et al., 2012; Richter–Boix et al., 2004). Material and methods Study species P. perezi is the most common and widespread frog species in the Iberian Peninsula (Bosch et al., 2009; Masó and Pijoan, 2011). Adults are essentially aquatic,

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although they have a certain terrestrial dispersion capacity (Egea–Serrano, 2009). P. perezi is present in many types of Mediterranean and Eurosiberian habitats, such as wetlands, ponds, lakes, rice fields, and rivers. In the study area, it shares habitats with natural predators such as autochthon fish and dragonflies, as well as with invasive predators such as P. clarkii. The Natural Park of Aiguamolls de l’Empordà consists of wetlands that can be considered a long–term invaded area because P. clarkii has been present since the 1980s (Moreno–Amich and Vila–Gispert, 2000), and by the 1990s it had become abundant throughout the Park (Moreno–Amich and Vila–Gispert, 2000). The red swamp crayfish (P. clarkii) has a cylindrical body with a clearly marked abdomen and differentiated and segmented thoracic limbs. It is considered an opportunistic omnivore (Gutiérrez–Yurrita et al., 1998) and it tends to inhabit swamp areas with abundant vegetation (especially macrophytes) (Gherardi et al., 2002). It exhibits characteristics of an r–selected species, including early maturity at small body size (10 g), rapid growth rates (50 g in 3–5 months), large numbers of offspring at a given parental size (a female of an average size producing 400 pleopodal eggs), and relatively short life spans (Gherardi, 2006). Experimental design and study site Field work was conducted at the Natural Park of Aiguamolls de l’Empordà (North–East of the Iberian Peninsula). Two ponds were created for the experiment in the natural park wetlands area at the beginning of autumn 2003 (UTM: 507473, 4674382). The area of each pond was 4 x 4 m and 1 m in depth, and they were separated from each other by 5 m. After they were made, the ponds were surrounded by a plastic fence of 1 m in height. This fence was sunk 30 cm into the ground to avoid amphibians and crayfishes entering the ponds. The bottom of one of the ponds, considered a control pond (without P. clarkii), was lined with screen–mesh under the soil to avoid the entrance of individuals of P. clarkii. During autumn 2003 and spring 2004, the ponds were naturally colonized by aquatic vegetation with species such as Typha latifolia and Chara vulgaris. As the ponds were alongside each other, they were under the same environmental conditions with the same forest canopy and similar to sunlight. We therefore assume that the two ponds had similar physical–chemical water features with around 600 µS/cm conductivity, 8.24 pH, and near 100 % exposure of oxygen saturation (these variables were measured twice during the experiment, in June, and in August). Water temperature was measured 10 times during the experiment and was similar between ponds, ranging from the 29 ºC in July and August to a minimum of 21ºC in October. In 27th April of 2004 two egg masses of P. perezi were collected in the Natural Park of Aiguamolls de l’Empordà. Eggs hatched in the laboratory at 23–25 ºC and were held until tadpoles were free–swimming (Gosner stage 25; Gosner, 1960). We mixed tadpoles from the different clutches before use in the experiment to homogenize genetic variation. For each pond, groups of 10 tadpoles were randomly assigned to 15 cylin-

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drical field enclosures made of plastic screen mesh (2.0 mm mesh; 19 cm diameter, 46 cm height; depth of enclosure submergence was approximately 40 cm [11–liter volume]) and were divided among three spatial blocks. There was thus a total of 30 replicates, 15 for each treatment (i.e. for each pond), and 300 tadpoles. One of the ponds was a control, without P. clarkii (hereafter referred to as 'control pond'). The pond with the invasive predator (hereafter referred to as 'invaded pond') had a density of individuals of P. clarkii similar to that of natural conditions (6 crayfish/m2) (Gherardi and Acquistapace, 2007). P. clarkii individuals were collected with crayfish traps in the Natural Park of Aiguamolls de l’Empordà and we added predators, regardless of sex, to the experimental pond on the same day they were collected. We searched ponds daily for metamorphs, defined by the emergence of at least one forelimb (stage 42; Gosner, 1960). When metamorphs were detected, they were removed immediately from the enclosure and total weight was measured to the nearest 0.0001 g. On day 156 (10th of October 2004), we ended the experiment because most surviving animals had reached metamorphosis. Data analyses Mass at metamorphosis (weight of metamorph), larval period (number of days until metamorphosis, day of Gosner stage 42) and survival at metamorphosis (Sm, metamorphosed) were used to measure the response of tadpoles to the invasive predator P. clarkii. Total survival (Ts, number of metamorphosed individuals and tadpoles that survived at the end of the experiment) was also considered since some individuals did not metamorphose at the end of the experiment, but they were alive. Survival data was binomial: alive or death. A linear regression was performed to analyze mass at metamorphose and larval period linear dependence. This analysis was done at an individual level because working with the means rules out intravariability within the species and we could have lost information concerning the relationship between these two variables. Treatment effects (control pond or invaded pond) on mass of metamorph, larval period and growth rate were analysed using the SPSS program, and the univariant model, considering individuals as experimental units. The enclosure was nested within a particular pond; this term was considered random in an overall mixed GLM (Df are in table 1). Larval period, mass of metamorph, and growth rate were log transformed to achieve normality distribution. Survival at metamorphosis and total survival were analysed using a generalized linear model with a binomial distribution and a logit function. All statistical analyses were performed using SPSS 15 (SPSS Inc., Chicago, IL, USA, 1989–2006). Results Overall, our results showed that tadpoles reared under the presence of P. clarkii had a shorter larval period (pond with P. clarkii: 67.30 days (SE 1.62);

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Table 1. General linear model results for effects of treatments to mass at metamorphosis and larval period and growth rate (log transformed to assume normality). Tabla 1. Resultados obtenidos con el modelo lineal generalizado de los efectos de los tratamientos en la masa en la metamorfosis, el período larvario y la tasa de crecimiento (log transformado para suponer normalidad).

Df Mean square

Error

P–value

Larval period

Pond

0.72

0.15

104.48

< 0.0001

Pond * enclosure

0.01

0.27

2.80

0.001

Mass at metamorphosis

Pond

0.04

0.71

1.60

0.22

Pond * enclosure

0.02

1.28

1.36

0.18

Growth rate

Pond

1.11

0.57

43.6

< 0.0001

Pond * enclosure

0.03

1.06

2.61

0.002

pond without P. clarkii: 105.9 days (SE. 4.89); fig. 1, table 1). The metamorph mass and larval period were linearly correlated for both treatment groups (fig. 2; pond without P. clarkii, R2 = 0.348 (p = 0.001) and pond with P. clarkii, R2 = 0.314, p = 0.001). Therefore, the final weight of metamorph increased with the increase of the larval period, and growth rate was significantly higher in the pond with P. clarkii (table 1, pond with P. clarkii, 0.006 g/d, SE 0.0003; pond without P. clarkii: 0.004 g/d, SE 0.0002). No significant differences were found between treatments for the metamorph mass (table 1; pond with P. clarkii, 0.43 g, S.E. 0.19; pond without P. clarkii, 0.40 g, SE 0.0)). Figure 2 shows the segregation of the two groups during the larval period, but no segregation occurred between groups along the metamorph mass axis, indicating treatment had no effect on this variable. The presence of P. clarkii did not promote significant differences in total survival or in survival to metamorphosis. Discussion Our results suggest that the presence of the invasive species P. clarkii influences tadpole development since tadpoles reared in the presence of P. clarkii had a shorter larval period, reaching metamorphosis earlier than tadpoles reared in an environment without P. clarkii. We hypothesized that the time to metamorphosis would probably be longer because of diminishing activity and consequently less energy intake (Tejedo et al., 2010, Touchon et al., 2015). However, discussion on this issue continues and our results indicate other patterns could operate in this predatory–prey relation. Denver (1995, 1997a) showed that tadpoles in drying or stressed conditions initiated metamorphosis early due to activation of the corticotrophin–releasing hormone, known to be responsible to advance metamorphosis for tadpoles (Denver, 1997b). P. clarkii might act as a stressor to P. perezi, as it reportedly preys on egg masses, tadpoles and even adult amphibians (Gherardi et al., 2001). Therefore, in our study, P. clarkii might

have accelerated metamorphosis of tadpoles through such activation of corticotrophin–releasing hormone. Orizaola et al. (2012) and Richter–Boix et al. (2004) suggested that the presence of predators will result in smaller sizes at metamorphosis. Nevertheless, although mass at metamorphosis was not significantly different between treatments in our study, our results show that growth rate was significantly higher for individuals reared under P. clarkii presence. This agrees with Nunes et al. (2014b), who showed that P. perezi tadpoles tended to grow faster in the presence of crayfish than in non–predatory environments. A larger prey size might provide an advantage from predation, so that increasing growth rate could also be a direct and adaptive response to predation (Urban, 2007). Importantly, however, the growth/predation risk trade–off is a common constraint documented for many organisms, with higher growth rates coming at the expense of increased vulnerability to predators (Lima and Dill, 1990; McPeek, 2004). Nunes et al. (2014b) studied P. perezi and P. clarkii relations considering frog populations differing in historical exposure to the invasive predator. Tadpoles from non–invaded populations responded to the presence of P. clarkii with behavioural plasticity (they reduced behavioural activity), whereas long–term invaded populations showed canalized antipredator behavior (they presented a constant low activity level). Their results suggest that, while native P. perezi populations responded behaviourally to P. clarkii, the strong predation pressure imposed by the crayfish has induced the evolution of qualitatively different antipredator defences in populations with longer coexistence time. The Natural Park of Aiguamolls de l’Empordà consists of wetlands that can be considered a long–term invaded area (P. clarkii has been present for more than 20 years). In our experiment, therefore, tadpoles (from both treatments) might show a constant low activity level. As a result, tadpoles reared under the presence of the crayfish, might not present changes in behavioural activity levels but higher growth rates, while tadpoles without the presence of crayfishes might not accelerate growth

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Without P. clarkii With P. clarkii

Number of metamorphos

0 56

69 73 78 84 88 95 103 109 119 156 Time to metamorphosis (days)

Fig. 1. Relation between number of metamorphs and time to metamorphosis for the two treatments. Fig. 1. Relación entre el número de renacuajos en fase de metamorfosis y la duración de la metamorfosis para los dos tratamientos.

Log mass of metamorphosis (g)

0.20

Without P. clarkii With P. clarkii

0.00

–0.20

–0.40

–0.60

–0.80 1.70

1.80 1.90 2.00 2.10 Log larval period length (days)

2.20

Fig. 2. Tadpole mass at metamorphosis linearly dependent on the tadpole larval period for each treatment: pond without P. clarkii (R2 = 0.348, p = 0.001) and pond with P. clarkii (R2 = 0.314, p = 0.001). Fig. 2. Relación lineal entre el número de renacuajos en fase de metamorfosis y el período larval para cada tratamiento: estanques sin P. clarkii (R2 = 0,348, p = 0,001) y estanques con P. clarkii (R2 = 0,314, p = 0,001).

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since they did not detect cues and danger. Nunes et al. (2014b) stated that a negative correlation between growth and behaviour agrees well with the low activity levels found in long coexistence populations. The role of antipredator phenotypic plasticity might be crucial in population dynamics during biological invasions (Miner et al., 2005). Relyea (2005) first examined the heritability of predator–induced defences, and observed that predator–induced traits can frequently be heritable, although the magnitude of heritability can be wide ranging across environments. Beyond the direct predation impact, P. clarkii is known for its capacity to destroy macrophytes and to increase water turbitidy by digging tunnels, all actions that contribute to increase algal growth in invaded ponds and lakes (Rodriguez et al., 2005). As a result, tadpoles would feed much better in a more eutrophic and algal dominated pond than in an oligotrophic pond dominated by macrophyte. This could explain the higher growth rate among tadpoles in the pond with crayfish, but we cannot make this conclusion as we did not measure changes in the algae community or in turbidity. As some studies indicate how location (field or laboratory, for example) can affect survival (Mitchell, 1990; Saura–Mas et al., 2002), we hypothesized that presence or absence of P. clarkii could also affect survival rates by increasing stress variables related to growth. Nevertheless, we did not detect a significant impact on survival rates, probably because there was no direct contact between prey and predator. Our results suggest that stress promoted by the presence of P. clarkii might promote shorter life cycles but not changes in survival if there is no contact between prey and predator. A potential weakness in our study is that our statistical analysis did not consider the possibility that native predators could also induce defences in tadpoles. While tadpoles in this study have similar habitats and experience similar native predator regimes, we cannot exclude the possibility that these differences may result from adaptation to other local habitat features such as competitors or food availability (Relyea, 2002; Richter–Boix et al., 2010). Finally, we conclude that, in field conditions, the invasive species P. clarkii might accelerate a metamorphosis of P. perezi tadpoles in Mediterranean wetlands ecosystems. Here we show that in addition to direct predation, this invasive predator may also alter P. perezi populations by increasing their growth rate by decreasing the larval period and, as a result, maybe also decreasing the length of the life cycle. These P. perezi life cycle changes might not be synchronised with the food web in Mediterranean wetlands, causing effects at a community level. These results represent a preliminary approach to the study of changes that this invasive species can drive in P. perezi populations. Further studies with higher statistical power are needed to confirm our trends. Our findings, however, coincide with observations from other studies indicating that phenotypic plasticity in P. perezi may play an important role in population dynamics in the face of global changes such as those involving invasive predators.

Acknowledgements We would like to thank Sergi Romero and Josep Espigulé for their support and help during the creation of the ponds and M. D. Boone for helpful comments during the design of the experiment. The two ponds with research objectives were created under the supervision of the park manager. We would also like to thank J. M. Benejam, J. Carol, J. Guillamet, B. Minobis and M. Pifarrer for their volunteer field work, and the 'Servei de Control de Mosquits de la Badia de Roses i del Baix Ter' for the loan of the precision scale to measure mass of metamorphs. References Alford, R. A., Richards, S. J., 1999. Global amphibian declines: a problem in applied ecology. Annual Review of Ecology, Evolution and Systematics, 30: 133–165. Almeida, E., Nunes A., Andrade P., Alves S., Guerreiro C., Rebelo, R., 2011. Antipredator responses of two anurans towards native and exotic predators. Amphibia–Reptilia, 32: 341–350. Anholt, B. R., Werner, E. E., 1995. Interaction between food availability and predation mortality mediated by adaptive behavior. Ecology, 76: 2230–2234. Beebee, T. J. C., Griffiths, R. A., 2005. The amphibian decline crisis: a watershed for conservation biology? Biological Conservation, 125: 271–285. Beja, P. R., 1996. An analysis of otter Lutra lutra predation on introduced American crayfish Procambarus clarkii in Iberian streams. Journal of Applied Ecology, 33: 1156–1170. Bosch, J., Tejedo, M., Beja, P., Martínez–Solano, I., Salvador, A., García–París, M., Gil, E. R., Beebee, T., 2009. Pelophylax perezi. IUCN Red List of Threatened Species, . Version 2014.1. International Union for Conservation of Nature, http://dx.doi.org/10.2305/IUCN.UK.2009.RLTS. T58692A11812894.en [Accessed on 18 July 2014]. Carey, C., Bradford, D. F., Brunner, J. L., Collins, J. P., Davidson, E. W., Longcore, J. E., Ouellet, M., Pessier, A. P., Schock, D. M., 2003. Biotic factors in amphibian population dieclines. In: Multiple stressors and declining amphibian: evaluating cause and effect: 1–49 (G. Linder, D. W. Sparling, S. K. Krest, Eds.). Society of Environmental Toxicology and Chemistry (SETAC), Pensacola, Florida. Carreira, B. M., Dias, M. P., Rebelo, R., 2014. How consumption and fragmentation of macrophytes by the invasive crayfish Procambarus clarkii shape the macrophyte communities of temporary ponds. Hydrobiologia, 721: 89–98. Collins, J. P., Storfer, A., 2003. Global amphibian declines: sorting the hypotheses. Diversity and Distributions, 9: 89–98. Cruz, M. J., Rebelo, R., Crespo, E. G., Crespo, E. G., 2006. Effects of an introduced crayfish, Procambarus clarkii, on the distribution of south–western Iberian amphibians in their breeding habitats. Ecography, 29: 329–338.

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Conservation status of Paramuricea clavata (Risso, 1826) (Anthozoa, Alcyonacea) in the Chafarinas Islands (Mediterranean Sea) L. Sánchez–Tocino, A. de la Linde Rubio, M. J. López–Rodríguez, J. M. Tierno de Figueroa Sánchez–Tocino, L., de la Linde Rubio, A., López–Rodríguez, M. J., Tierno de Figueroa, J. M., 2019. Conservation status of Paramuricea clavata (Risso, 1826) (Anthozoa, Alcyonacea) in the Chafarinas Islands (Mediterranean Sea). Animal Biodiversity and Conservation, 42.2: 253–256, https://doi.org/10.32800/abc.2019.42.0253 Abstract Conservation status of Paramuricea clavata (Risso, 1826) (Anthozoa, Alcyonacea) in the Chafarinas Islands (Mediterranean Sea). The red gorgonian Paramuricea clavata (Risso, 1826) is affected by the combined effects of environmental stress factors and diseases in the Mediterranean area. Samplings at different depths in two sites of Chafarinas Islands (South–Western Mediterranean Sea) were carried out to quantify the degree of injuries on red gorgonian colonies. The results showed that shallow colonies displayed a higher rate of injuries than deep colonies. Overall, the conservation status of the population was worse than previously considered in this area. Key words: Red gorgonian, Paramuricea clavata, Health status, Northern Africa Resumen Estado de conservación de Paramuricea clavata (Risso, 1826) (Anthozoa, Alcyonacea) en las islas Chafarinas (mar Mediterráneo). En la zona del Mediterráneo, la gorgonia roja Paramuricea clavata (Risso, 1826) sufre los efectos combinados de factores de estrés ambiental e infecciones. Se extrajeron muestras a diferentes profundidades en dos sitios de las islas Chafarinas (mar Mediterráneo sudoccidental) para cuantificar el grado de daño que presentaban las colonias. Los resultados mostraron que las colonias de aguas más superficiales presentaban una mayor tasa de daño que las colonias de aguas más profundas. En conjunto, el estado de conservación de la población era peor de lo que se había considerado previamente en esta zona. Palabras clave: Gorgonia roja, Paramuricea clavata, Estado de salud, Norte de África Received: 25 X 18; Conditional acceptance: 14 I 19; Final acceptance: 29 I 19 L. Sánchez–Tocino, J. M. Tierno de Figueroa, Departamento de Zoología, Facultad de Ciencias, Universidad de Granada, Campus Fuentenueva, 18071 Granada, Spain.– A. de la Linde Rubio, Urbanización los Delfines, Pl 4, 2º, 11207 Algeciras, Cádiz, Spain.– M. J. López–Rodríguez, Departamento de Ecología, Facultad de Ciencias, Universidad de Granada, Campus Fuentenueva, 18071 Granada, Spain. Corresponding author: J. M. Tierno de Figueroa, E–mail: jmtdef@ugr.es

Mass mortality events of gorgonians and other marine macroinvertebrates have been repeatedly reported in the last decades in the Mediterranean Sea (e.g. Cerrano et al., 2000; Garrabou et al., 2009). These events, usually, but not only, have been associated with episodes of positive water temperature anomalies. One of most affected species is the red gorgonian ISSN: 1578–665 X eISSN: 2014–928 X

Paramuricea clavata (Risso, 1826) (e.g. Cerrano et al., 2000; Martin et al., 2002; Huete–Stauffer et al., 2011). Several authors have noted that the combined effect of environmental stress and microorganism infections may be jeopardizing the conservation status of the P. clavata populations (e.g. Vezzulli et al., 2013). Moreover, water warming anomalies are © 2019 Museu de Ciències Naturals de Barcelona Papers are published under a Creative Commons Attribution 4.0 International License

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generally related to a decrease of oxygen availability, enhancing the necrosis of coenechyme (Previati et al., 2010). During the sampling conducted in summer 2015 to evaluate a mass mortality of the white gorgonian Eunicella sigularis (Esper, 1791) in the Chafarinas Islands (North Africa, Western Mediterranean Sea), related to a high temperature event in 2014, some damaged colonies of P. clavata were also observed (de la Linde Rubio et al., 2018). Particularly, de la Linde Rubio et al. (2018) carried out a sampling in Tajo del Cementerio (Rey Island, 35º 10' 50.02'' N, 2º 25' 1.11'' W) between 20 and 30 m depth to assess the number of affected colonies of P. clavata (identified by being partially covered by epibionts). They found that more than 50 % of the colonies were affected (most of them displaying less than 50 % of their surface with epibionts) but none were dead. They noted that the studied population had been less affected than populations in other Mediterranean areas where mortality events of this species were reported (e.g. Cerrano et al., 2005; Crisci et al., 2011). In order to better evaluate the conservation status of P. clavata in the Chafarinas Islands, we carried out a new study in two localities and at different depths: (1) Rey Island, between Tajo del Pirata and Tajo del Cementerio (35º 10' 48.08'' N, 2º 25' 09.10'' W), 29 VIII 18, depths: 22, 20 and 17 m East exposure; (2) Isabel II Island, North, under the lighthouse (35º 11' 02.97'' N, 2º 25' 50.96'' W), 01 IX 18, depths: 25, 23, 20 and 18 m, North exposure. Records of water temperature from the buoy of Melilla (the nearest place with available data) at 15 m depth in 2017 showed a maximum temperature of 28.6 ºC in mid–August (Puertos del Estado webpage, http://www.puertos.es/es–es/oceanografia/ Paginas/portus.aspx). At each depth, we quantified

Fig. 1. Paramuricea clavata colony partially covered by epibionts. Fig. 1. Colonia de Paramuricea clavata parcialmente cubierta por epibiontes.

Table 1. Mean (± SD) number of healthy, non–healthy and dead colonies at each depth at both study sites. Note that the number of samples considered for their calculation has been reduced to five at each depth at Isabel II Island to balance the estimates. Tabla 1. Número medio de colonias (± DE) saludables, no saludables y muertas a cada profundidad en ambas localidades de estudio. Nótese que el número de muestras consideradas para su cálculo se han reducido a cinco en todas las profundidades de la isla de Isabel II para equilibrar las estimaciones.

Depth (m) N

Number of colonies

Rey Island 17 10 49

Mean ± SD Dead Non–healthy Healthy 1.40 ± 1.43 3.20 ± 1.81 0.3 ± 0.48

20 10 70

1.50 ± 1.27 4.80 ± 1.32 0.70 ± 1.25

22 10 71

0.20 ± 0.42 1.90 ± 1.66 6.10 ± 3.31

0.60 ± 0.55 3.80 ± 1.10 4.00 ± 2.12

Isabel II Island

20 5 37

1.20 ± 0.84 6.20 ± 3.19

23 5 28

0.60 ± 0.55 5.00 ± 1.87

25 5 28

0.60 ± 0.55 5.00 ± 2.74

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Rey Island

Percentage

70 60 50 40

% Healthy

% Non–healthy % Dead

20 10 0

20 Depth (m)

Percentage

Isabel II Island 100 90 80 70 60 50 40 30 20 10 0

% Healthy % Non–healthy % Dead

23 20 Depth (m)

Fig. 2. Percentage of healthy, non–healthy and dead colonies of Paramuricea clavata at each study site. Fig. 2. Porcentaje de colonias de Paramuricea clavata saludables, no saludables y muertas en cada lugar de estudio.

the number of healthy (with 0 % epibiosis or necrosis), affected (with 1–99 % epibioisis or necrosis), and dead colonies (with 100 % epibiosis or necrosis) of P. clavata (fig. 1) using 10 squares of 50 x 50 cm, except at Isabel II island at 20 and 18 m deep, where only five squares were sampled at each depth. A total of 202 colonies were assessed in Rey Island (83 colonies at 22 m depth, 70 colonies at 20 m depth, and 49 colonies at 17 m depth) and 196 colonies in Isabel II Island (62 colonies at 25 m depth, 55 colonies at 23 m depth, 37 colonies at 20 m depth, and 42 colonies at 18 m depth). The results showed that dead colonies were more frequent in shallower waters and the general conservation status of the colonies increased with depth (fig. 2; table 1), as in shallower waters the effects of high temperature episodes are more noticeable. For example, Martin et al. (2002) experimentally demonstrated that necrotic diseases significantly speed up from a certain tem-

perature value in two studied gorgonian species, one of them P. clavata, and Bally and Garrabou (2007) experimentally demonstrated the causal role of the thermodependent bacteria Vibrio coralliilyticus as an infectious agent in the Mediterranean P. clavata colonies. The results obtained for the Chafarinas Islands also show that, when studied in a shallower range, the conservation status of P. clavata colonies is worse than previously considered (de la Linde Rubio et al., 2018). This could also be related to the highest peak in temperature occurring in summer 2017 compared with that of summer 2014 after which de la Linde Rubio et al. (2018) carried out their study. Finally, it should be noted that, in addition to the effect of increases in temperature, the health of P. clavata populations in the Mediterranean have been affected by injuries caused by anchoring and fishing (Barvestrello et al., 1997), but this is not the case in the Chafarinas Is-

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lands where P. clavata colonies are found in vertical coastal slopes, relatively far from the fishing activity. The fate of the partially damaged colonies could be a total or partial recovery of the necrotic tissue or the fragmentation of affected branches. Further surveys could clarify how the affected population will recover and whether thermal anomalies are leading to a general loss of structural complexity of the benthic assemblages. Moreover, the application of restoration measures such as pruning could contribute to their faster recovery, as has been demonstrated in other gorgonian species as Ellisella paraplexauroides (Sánchez–Tocino et al. 2017). Acknowledgements The authors thank the military detachment in the Chafarinas Islands, personnel at the Organismo Autónomo de Parques Nacionales and particularly Javi Díaz for help and collaboration. We also thank two anonymous referees for their valuable comments that notably improved the manuscript. References Bally, M., Garrabou, J., 2007. Thermodependent bacterial pathogens and mass mortalities in temperate benthic communities: a new case of emerging disease linked to climate change. Global Change Biology, 13: 2078–2088. Barvestrello, G., Cerrano, C., Zanzi, D., Cattaneo– Vietti, R., 1997. Damage by fishing activities to the Gorgonian coral Paramuricea clavata in the Ligurian Sea. Aquatic Conservation: Marine and Freshwater Ecosystems, 7: 253–262. Cerrano, C., Arillo, A., Azzini, F., Calcinai, B., Castellano, L., Muti, C., Valisano, L., Zega, G., Bavestrello, G., 2005. Gorgonian population recovery after a mass mortality event. Aquatic Conservation: Marine and Freshwater Ecosystems, 15: 147–157. Cerrano, C., Bavestrello, G., Bianchi, C. N., Cattaneo–Vietti, R., Bava, S., Morganti, C., Morri, C., Picco, P., Sara, G., Schiaparelli, S., Siccardi, A., Sponga. F., 2000. A catastrophic mass–mortality episode of gorgonians and other organisms in the Ligurian Sea (NW Mediterranean), summer 1999. Ecology Letters, 3: 284–293.

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Crisci, C., Bensoussan, N., Romano, J. C., Garrabou, J., 2011. Temperature anomalies and mortality events in marine communities: insights on factors behind differential mortality impacts in the NW Mediterranean. Plos One, 6(9): e23814. de la Linde Rubio, A., Tierno de Figueroa, J. M., López–Rodríguez, M. J., Sánchez–Tocino, L., 2018. Mass mortality of Eunicella sigularis (Anthozoa: Octocorallia) in Chafarinas Islands (North Africa, Western Mediterranean Sea). Revista de Biología Marina y Oceanografía, 53: 285–290. Garrabou, J., Coma, R., Bensoussan, N., Bally, M., Chelvaldonné, P., Cigliano, M., Diaz, D., Harmelin, J. G., Gambi, M. C., Kersting, D. K., Ledoux, J. B., Lejeusne, C., Linares, C., Marschal, C., Perez, T., Ribes, M., Romano, J. C., Serrano, E., Torrents, O., Zabala, M., Zuberer, F., Cerrano, C., 2009. A new large scale mass mortality event in the NW Mediterranean rocky benthic communities: effects of the 2003 heat wave. Global Change Biology, 15: 1090–1103. Huete–Stauffer, C., Vielmini, I., Palma, M., Navone, A., Panzalis, P., Vezzulli, L., Misic, C., Cerrano, C., 2011. Paramuricea clavata (Anthozoa, Octocorallia) loss in the Marine Protected Area of Tavolara (Sardinia, Italy) due to a mass mortality event. Marine Ecology, 32: 107–116. Martin, Y., Bonnefont, J. L., Chancerelle, L., 2002. Gorgonians mass mortality during the 1999 late summer in French Mediterranean coastal waters: the bacterial hypothesis. Water Research, 36: 779–782. Previati, M., Scinto, A., Cerrano, C., Osinga, R., 2010. Oxygen consumption in Mediterranean octocorals under different temperatures. Journal of Experimental Marine Biology and Ecology, 390: 39–48. Sanchez–Tocino, L., de la Linde Rubio, A., Lizana Rosas, M. S., Pérez Guerra, T., Tierno de Figueroa, J. M., 2017. Pruning treatment: A possible method for improving the conservation status of a Ellisella paraplexauroides Stiasny, 1936 (Anthozoa, Alcyonacea) population in the Chafarinas Islands?. Mediterranean Marine Science, 18: 479–485. Vezzulli, L., Pezzati, E., Huete–Stauffer, C., Pruzzo, C., Cerrano, C., 2013. 16SrDNA pyrosequencing of the Mediterranean gorgonian Paramuricea clavata reveals a link among alterations in bacterial holobiont members, anthropogenic influence and disease outbreaks. Plos One, 8(6): e67745.

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Risk map of transmission of urogenital schistosomiasis by Bulinus truncatus (Audouin, 1827) (Mollusca Gastropoda, Bulinidae) in Spain and Portugal A. Martínez–Ortí, D. Vilavella, M. D. Bargues, S. Mas–Coma Martínez–Ortí, A., Vilavella, D., Bargues, M. D., Mas–Coma, S., 2019. Risk map of transmission of urogenital schistosomiasis by Bulinus truncatus (Audouin, 1827) (Mollusca Gastropoda, Bulinidae) in Spain and Portugal. Animal Biodiversity and Conservation, 42.2: 257–266, Doi: https://doi.org/10.32800/abc.2019.42.0257 Abstract Risk map of transmission of urogenital schistosomiasis by Bulinus truncatus (Audouin, 1827) (Mollusca Gastropoda, Bulinidae) in Spain and Portugal. We present a geographical distribution map of Bulinus truncatus based on historical and current localities in Spain and Portugal, that corresponds to the risk map of urogenital schistosomiasis for this freshwater snail. We reviewed samples of the species deposited at the Museu de Ciències Naturals of Barcelona and the Museo Nacional de Ciencias Naturales of Madrid, together with our own data, including some unpublished contributions. This map will help determine the optimal area for this species and identify areas of greatest risk for urogenital schistosomiasis in the two countries. We emphasize that global change and climate change may favour the presence of both the vector (B. truncatus) and the parasite (Schistosoma haematobium) in Spain and Portugal. Key words: Bulinus truncatus, Bulinidae, Urogenital schistosomiasis, Risk map, Spain, Portugal Resumen Mapa del riesgo de contraer esquistosomiasis urogenital provocada por Bulinus truncatus (Audouin, 1827) (Mollusca Gastropoda, Bulinidae) en España y Portugal. Se da a conocer el mapa de la distribución geográfica de Bulinus truncatus en España y Portugal en el que se recopilan las localidades históricas y actuales, que coincide con el mapa del riesgo de contraer esquistosomiasis urogenital provocada por este caracol de agua dulce. Se revisan las muestras de esta especie depositadas en el Museu de Ciències Naturals de Barcelona y en el Museo de Ciencias Naturales de Madrid, así como datos propios, incluidas algunas aportaciones inéditas. Este mapa permitirá conocer el área óptima de esta especie y determinar las zonas de mayor riesgo de contraer esquistosomiasis urogenital en los dos países. Se pone de manifiesto que el cambio global y el cambio climático pueden favorecer la presencia tanto del vector (B. truncatus) como del parásito (Schistosoma haematobium) en España y Portugal. Palabras clave: Bulinus truncatus, Bulinidae, Esquistosomiasis urogenital, Mapa de riesgo, España, Portugal Received: 16 XI 18; Conditional acceptance: 10 I 19; Final acceptance: 30 I 19 A. Martínez–Ortí, D. Vilavella, M. D. Bargues, S. Mas–Coma, Unit of Sanitary Parasitology, Department of FF & F and Parasitology, Faculty of Farmacy, Universitat de València, Burjassot, Valencia, Spain.– A. Martínez–Ortí, D. Vilavella, Museu Valencià d’Història Natural–i\Biotaxa, Alginet, Valencia, Spain. Corresponding author: A. Martínez–Ortí. E–mail: amorti@uv.es

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Introduction Bulinus O. F. Müller, 1781 is a genus of the Family Bulinidae P. Fischer and Crosse, 1880. It is of great health interest because most of its species are vectors of parasites that can cause serious diseases in humans, such as schistosomiasis (bilharziasis) (Malek and Cheng, 1974; Brown, 1980). Urogenital schistosomiasis was recently confirmed in the island of Corsica (France), with both the trematode Schistosoma haematobium (Bilharz, 1852) and its vector, the bulinid gastropod mollusc Bulinus truncatus (Audouin, 1827), coexisting (Berry et al., 2014; Holtfreter et al., 2014; Boissier et al., 2015, 2016). It should be noted that the epidemic in Corsica is not only due to the pure S. haematobium strain but also to a hybrid of S. haematobium and S. bovis (Sonsino, 1876), with which the epidemic acquires zoonotic connotations and great potential for dispersion, broadening the spectrum of reservoirs and vectors of the causal agent (Huyse et al., 2009; Boissier et al., 2016; Kincaid–Smith et al., 2017). Urogenital schistosomiasis can cause squamous bladder cancer and prostate cancer, acute appendicitis, and infertility in women (Gabbi et al., 2006; Figueiredo et al., 2015; Grácio, 2018). It was detected in the Iberian Peninsula at the beginning of the 20th century in the Algarve in Portugal (Sampaio Xavier and Fraga de Azevedo, 1966). It is currently known in Spain and Portugal, although the trematode has only been found in sub–Saharan migrants. No cases of autochthonous transmission have been reported to date (Martínez–Ortí et al., 2015). In Spain, the disease has been diagnosed in four hospitals: 1) Poniente Hospital in El Ejido (Almería) (Salas–Coronas in Ríos 2011, Salas–Coronas et al., 2018); 2) Hospital Universitario Materno–Infantil de Canarias, in Gran Canaria (Ramos Macías et al., 2010); 3) Vall d’Hebron Hospital in Barcelona (Bocanegra et al., 2014); and 4) Hospital Clinic de Barcelona (Calvo–Cano et al., 2015). The ages of patients ranged from 20 to 30 years. Several cases have also been detected in the São João Hospital in Porto, Portugal (Dr. Rogério Ruas, pers. comm.), and cases have also been documented in returned military personnel from the former Portuguese African colonies (Vieira et al., 2007). Several authors suggest that the lymnaeoid molluscs B. truncatus and Planorbarius metidjensis (Forbes, 1834) should be considered the intermediate hosts of S. haematobium and a S. haematobium –S. bovis hybrids, respectively, and therefore susceptible to the spread of the disease in Portugal and Spain (Fraga de Azevedo, 1965; Fraga de Azevedo and Xavier Sampaio, 1965, 1969; Ramajo–Martín, 1978; Grácio, 1983; Boissier et al., 2016; Kincaid–Smith et al., 2017). B. truncatus is also a transmitter of S. bovis and experimentally of S. margrebowiei Le Roux, 1933 (Southgate and Knowles, 1977) and of Paramphistonum cervi (Schrank, 1790) and P. microbothrium Fischoeder, 1901, parasitizing small and large rumiants (Malek and Cheng, 1974; Brown, 1980; Pampiglione et al., 1988). It has also been experimentally demonstrated that it is the second intermediate host of Echinostoma spp. (Brown, 1980; Christensen et al., 1980).

Martínez–Ortí et al.

B. truncatus presents a wider geographic area, extending along the coasts of the circum–Mediterranean region, much of Africa, the Middle East and the Irano–Turanian region. In Europe it is cited from various Mediterranean countries and Portugal (Martínez–Ortí et al., 2015), but it has not been mentioned in the Macaronesian islands (Bank et al., 2002). In North Africa it is known from Egypt to the south of Morocco and the Sahara, and from Angola to Malawi and Ethiopia in the south, as well as in the Middle East (Israel, Iran, Iraq, Jordan, Saudi Arabia, Syria and Yemen) (Germain, 1931; Larambergue, 1939; Brown, 1980; Schütt, 1987; Neubert, 1998; Martínez–Ortí et al., 2015). This species does not present a fossil record in Europe, although the genus Bulinus is present. The oldest fossil record in Europe is known from other species of the genus Bulinus: B. meneghinii (Sacco, 1886) cited from Fossano (Piamonte, Italy) (7,246– 5,552 Ma), B. corici Harzhauser and Neubauer, 2012 (15.97–13.65 Ma) from Jauring in the Aflenz basin (Austrian Alps) (Harzhauser et al., 2011) and B. matejici (Pavlović, 1931) from Miocene lacustrine deposits (ca. 15.97–11.63 Ma) of the Serbian Lake System at Ćerane near Kaona, Gornja Mutnica, Mađare and Pardik in central Serbia and the late Miocene of the Turiec Basin in Slovakia and Kosovo (Neubauer et al., 2017). Another five fossil species of Bulinus are present in Kosovo: B. bouei (Pavlović, 1931), B. ornatus (Pavlović, 1931), B. pavlovici (Atanacković, 1959), B. stevanovici (Atanacković, 1959) and B. striatus (Milošević, 1978). Kincaid–Smith et al. (2017) questioned the spatial repartition of schistosome intermediate hosts in Europe and emphasized the urgency of determining the geographical distribution of the snail’s intermediate host populations and monitoring their spatiotemporal dynamics. Morelet (1845) cites B. truncatus for the first time in the Iberian Peninsula, in Coimbra in Portugal. For Spain, Graells (1846) later cites its presence in the locality of Barcelona and Bofill (1917) reports findings in the Balearic Islands. Currently in Spain its geographical distribution is circumscribed to various Mediterranean and Atlantic regions: Andalusia, the Balearic Islands, Catalonia, Valencia and Galicia, although we have proof of its living presence only in Andalusia (Rolán et al., 1987; Bech, 1990; Pérez–Quintero et al., 2004). However, in continental Portugal it is documented throughout the entire length of the country (Morelet, 1845; Nobre, 1913, 1941; Fraga de Azevedo et al., 1969a, 1969b; Sampaio et al., 1975, 1977; Medeiros and Simões, 1979). In Spain, in Ibiza and Gavà, and in Malta, B. truncatus is known from the Holocene, relatively recent (–12,000 –10,000 years) (Gasull, 1965; Marqués–Roca, 1974; Giusti et al., 1995). The biogeographical data collected in this work is of great interest to public health before urogenital schistosomiasis possibly reaches the Iberian Peninsula, as recently occurred in Corsica. Therefore, knowing such as what location of the disease vector, B. truncatus, in the Iberian Peninsula and Balearic Islands, is of great interest to predict the possible transmissions of this disease and control the possible foci of transmis-

Animal Biodiversity and Conservation 42.2 (2019)

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Fig. 1–3. Bulinus truncatus: 1, vegetable garden in Bonanova, Barcelona (Spain) (MNCN 15.05–26320, h = 6.1 mm). 2, Adra (Almeria, Spain) (MZB 84–7124, h = 7.9 mm); 3, Porto–Vecchio, Corsica (France) (MVHN–240815AF01, h = 8.0 mm). Figs. 1–3. Bulinus truncatus: 1, huerto de Bonanova, Barcelona (España) (MNCN 15.05–26320, h = 6,1 mm); 2, Adra, Almería (España) (MZB 84–7124, h = 7,9 mm); 3, Porto Vecchio, Córcega (Francia) (MVHN–240815AF01, h = 8,0 mm).

sion. Kincaid–Smith et al. (2017) indicated the need for health authorities to better quantify the risks and prevent future outbreaks of urogenital schistosomiasis in Europe and to know more about the ecological features that cause the host–parasite interactions. Material and methods For a better knowledge of its geographical distribution and the places where it has been historically found in the Iberian Peninsula and the Balearic Islands, we performed an exhaustive review of the samples of this species deposited at the 'Museo Nacional de Ciencias Naturales' of Madrid (MNCN), the 'Museu de Ciències Naturals' of Barcelona (MCN), and our own data deposited in the Museu Valencià d’Història Natural (MVHN) of Alginet (Valencia), which has also allowed us to provide several new locations (Martínez–Ortí, 2017; Uribe and Agulló Villaronga, 2018). All known bibliographic locations of B. truncatus in the Iberian Peninsula and the Balearic Islands are listed, and numerous samples deposited in the MNCN of Madrid and the MCN of Barcelona have been reviewed. The localities whose samples are deposited in the MVHN are also provided (tables 1, 2). In these tables of localities in Spain and Portugal, the place and municipality is named the first time it is cited, along with the author and year of publication; the same places mentioned by the same author or other authors later are not repeated. Four samples from the MNCN and 29 from the MCN have been reviewed, some of them corresponding to unpublished localities (figs. 1–3; tables 1, 2). The geographical coordinates of locations in the cities of islands such as Ibiza, Mallorca or Menorca, as places where the

species are found, are not indicated as the precise localities are not known by some authors. The geographical distribution map was made using an Excel spreadsheet (Microsoft Office 2014) and the Qgis version 2.18 geographic information system program. Results For the last two centuries, B. truncatus has been present on the coasts of Spain and Portugal. However, it has not been reported from the centre of the Iberian Peninsula, the Pyrenees or Cantabrian areas despite numerous malacological studies conducted in these areas (fig. 4), In Spain it has been described by several authors from 63 localities near the coast (fig. 4; table 1). Of the 32 revised samples of B. truncatus deposited at the MNCN and the MCN, six are unpublished, and two new locations are provided (table 1).There are no recent reports of B. truncatus in Portugal. From 1845 onwards it has been reported from 17 localities in the regions of Minho, Douro Litoral, and Beira Litoral in the north and the Algarve in the south, but it has not been reported since (Mendez Simoes, 2016) (fig. 4; table 2). Similarly, since 1983 it has not been found in France, and Larembergue (1939), indicated that B. truncatus should be eliminated from the list of molluscs living in mainland France as he did not find it during his samplings. Discussion The geographical distribution of B. truncatus has been predicted to spread both in Spain and Portugal, and also

Martínez–Ortí et al.

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Table 1. List of localities of Bulinus truncatus in Spain. Only the author indicating the location for the first time is shown. The locations are transcribed exactly as they appear in the original publications: UTM, Universal Transverse Mercator. Tabla 1. Lista de las localidades de Bulinus truncatus en España. Solo se indica el autor que señala el lugar por primera vez. Se transcriben las localizaciones tal y como aparecen en las publicaciones originales: UTM, sistema de coordenadas universal transversal de Mercator. Nº Locality

Reference

Barcelona

Graells (1846)

31TDF28

Casa Antúnez, Barcelona

Chía (1887)

31TDF28

Between the hypodrome and the mouth

of the Llobregat River (Barcelona)

Fagot (1892)

31TDF27

Besos River, Barcelona

Martorell and Bofill (1888)

31TDF38

'Menorca' (Balearic Islands)

Martorell and Bofill (1888)

Hostalrich (Girona)

Chía (1893)

31TDG62

Pubol Pond (Girona)

Chía (1893)

31TDG95

Hostalrich, Tordera River (Girona)

Chía (1916)

31TDG62

Albufera, Alcudia Lagoon (Mallorca, Balearic Islands)

Bofill (1917)

31SEE10

Maluquer (1917)

31SEE10

11 Segles (irrigation canal) of de Deià (Mallorca, Balearic Islands) Maluquer (1917)

31TEE60

10 Albufera, Alcudia Lagoon (Mallorca, Balearic Islands) 12 'Ibiza' (Balearic Islands)

Bofill (1919)

UTM

–

13 Between the hypodrome and the mouth

Bofill and Haas (1920)

31TDF27

14 Castelldefels (Barcelona)

Bofill and Haas (1920)

31TDF17

15 Mouth of Besos River (Barcelona)

Bofill et al. (1921)

31TDF38

16 Mataró (Barcelona)

Bofill et al. (1921)

31TDF59

17 Remolar Pond, Viladecans (Barcelona)

Haas (1929)

31TDF27

18 Santa Galdana (Menorca, Balearic Islands)

Aguilar–Amat (1933)

31SEE82

19 Bottom of Mahón ria, Colársega

Sacchi (1954)

31SFE01

Sacchi (1954)

31SEE91

Sacchi (1954)

31SEE82

of the Llobregat River (Barcelona)

(Menorca, Balearic Islands)

20 Rice field irrigation canals of Son Canesias,

Son Bou (Menorca, Balearic Islands)

21 Ferrerías River, Barranc d’ Algendar River

(Menorca, Balearic Islands)

22 Irrigation canal of Son Saura, Ciutadella

Sacchi (1954)

31SEE72

23 'Menorca' (Balearic Islands)

(Menorca, Balearic Islands)

Sacchi (1957a)

–

24 'Mallorca' (Balearic Islands)

Sacchi (1957b)

25 Palma. Molinar de Levante (Mallorca, Balearic Islands)

Gasull (1965)

31SDD77

26 Palma. Mestre Pere Spring (Mallorca, Balearic Islands)

Gasull (1965)

31SDD78

27 La Pobla. 'Ca'n Roca' (Mallorca, Balearic Islands)

Gasull (1965)

31SEE00

28 La Pobla. 'Ca'n Pujolet' (Mallorca, Balearic Islands)

Gasull (1965)

31SEE00

29 La Pobla. 'Ca'n Blau' (Mallorca, Balearic Islands)

–

Gasull (1965)

31SEE00

30 Muro, Son San Juan Spring (Mallorca, Balearic Islands) Gasull (1965)

31SEE00

31 Ratjada cove. Son Moll, Torrente (Mallorca, Balearic Islands) Gasull (1965)

31SED39

32 Artá. Irrigation canal Molins Molinet (Mallorca, Balearic Islands) Gasull (1965)

31SED29

33 Son Servera. Son Jordi, Torrente (Mallorca, Balearic Islands) Gasull (1965)

31SED38

34 Alaior. Son Bou Beach (Menorca, Balearic Islands)

31SEE91

Gasull (1965)

Animal Biodiversity and Conservation 42.2 (2019)

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Table 1. (Cont.)

Nº Locality

Reference

UTM

35 Ferreries, Santa Galdana, Binissaid Spring

(Menorca, Balearic Islands)

Gasull (1965)

31SEE82

Gasull (1965)

31SEE72

36 Ciutadella, Son Saura Beach, irrigation canal

(Menorca, Balearic Islands)

37 Ciutadella, Cala Macarella, irrigation canal

Gasull (1965)

31SEE82

38 Mahón, María Spring (Menorca, Balearic Islands)

(Menorca, Balearic Islands)

Gasull (1965)

31SFE01

39 Fornells, Coves Noves (Menorca, Balearic Islands)

Gasull (1965)

31TEE93

40 'Cala en Porter' (Menorca, Balearic Islands)

Gasull (1965)

31TEE91

41 San Antonio, Flandriense mudflats in the bay

Gasull (1965)

31SCD51

42 Rocina Creek, El Rocio P. N. Doñana (Huelva)

(Ibiza, Balearic Islands)

Marazanof (1966)

29SQB21

43 El Saltillo Creek, P. N. Doñana (Huelva)

Marazanof (1966)

29SQB12

Vilella (1967)

31TCF47

Altimira (1969)

31TDF27

44 Water pond between Picamoixons and Alcover,

Alt Camp (Tarragona)

45 La Farola. Delta del Llobregat (Barcelona)

46 Beach by the lighthouse, detritus from alluvial river deposits.

Altimira (1969)

31TDF27

47 Red clay on Gavà sands (Barcelona)

Delta del Llobregat (Barcelona)

Marqués–Roca (1974)

31TDF17

48 Fornells (Menorca, Balearic Islands)

Hidalgo in Compte (1985)

31SEE93

49 Xuño Lake (La Coruña)

Rolán et al. (1987)

29TMH92

50 Ampùries, Alt Empordà (Girona)

Bech (1993)

31TEG06

51 'Veta del Martinazo', Guadalquivir river basin (Huelva)

Pérez–Quintero et al. (2004) 29SQB20

52 Prado river bed, Piedras River (Huelva)

Pérez–Quintero et al. (2004) 29SPB5220

53 Adra Lagoon (Almería)

Bayo (2005)

54 Villena, Rey irrigation canal (Alicante) (486 m)

Martínez–Ortí et al. (2015)

(MVHN–281214TY01) 55 El Ejido. Poniente Hospital (Almería)

30SXH8572

Martínez–Ortí et al. (2015)

(MVHN–071214TP01) 56 Adra (Almería). Gasull coll. MZB 84–7124 (fig. 2)

30SWF06

30SWF1867

unpublished

30SWF06

unpublished

31TDF28

Ortiz de Zárate coll. MNCN–15.05/26320) (02/1957) (fig. 1) unpublished

31TDF28

57 Mr Muntades' vegetable garden, neighbourhood

of Bonanova (Barcelona) (Bech coll. MZB 2009–0535)

58 Vegetable garden in Bonanova (Barcelona) (Altimira ex–coll.,

59 Palma de Mallorca, irrigation canal (Mallorca, Balearic Islands)

(Cobos coll.; Compte ex–coll., MCNM–15.05/40630)

unpublished

31SDD78

60 Barcelona, in a private estate, exhausted locality (Altimira excoll.)

unpublished

31TDF28

61 Menorca (Ortiz de Zárate coll. MCNM–15.05/26314)

(Cobos coll. MCNM–15.05/40640)

unpublished

–

62 O'Grove Lake (Pontevedra) (E. Rolán, pers. comm.)

unpublished

29TNH00

unpublished

30SUF2540

63 Mouth of the Verde River, Marbella (Málaga)

(06/2013) (S. Torres Alba pers. comm.)

Martínez–Ortí et al.

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Table 2. List of localities of Bulinus truncatus in Portugal. Only the author indicating the location for the first time is shown. Tabla 2. Lista de las localidades de Bulinus truncatus en Portugal. Solo se indica el autor que señala el lugar por primera vez. Nº 64 65 66 67 68 69 70 71 72 73 74 75 76 77 78 79 80

Portuguese localities Coimbra, environs Porto Matozinhos Leça da Palmeira Esmoriz Senhor da Pedra, Ovar Aveiro Buçaco Buarcos Vala de Ceria Monçao, Souza River Fanzeres Silves, Algarve Montemor–o–Vehlo, Algarve Monchique, Algarve Choupal, Mondego River, Coimbra Aldeia das Vinhas, Algarve

Reference Morelet (1845) Nobre (1913) Nobre (1913) Nobre (1913) Nobre (1913) Nobre (1913) Nobre (1913) Nobre (1913) Nobre (1913) Nobre (1941) Nobre (1941) Nobre (1941) Fraga de Azevedo et al. (1969a, 1969b) Sampaio et al. (1975) Sampaio et al. (1975) Medeiros and Simões (1979) Grácio (1983)

in the rest of Europe, due to the rise in global temperature, causing a potential increase in transmission of urogenital schistosomiasis in the continent (Mas–Coma et al., 1987; Mas–Coma et al., 2010; McCreesh and Booth, 2013; Abou– El–Naga, 2013; Martínez–Ortí et al., 2015; Kincaid–Smith et al., 2017). The introduction of the disease from Africa to Europe is a phenomenon related to global change (essentially immigration and tourism) and climate change (especially, in this case, the global warming experienced in southern Europe). Due to the large migratory movements currently occurring in southern Europe, and in Spain in particular, with a large number of sub–Saharan migrants arriving in different ways –mainly in small boats– , on the eastern and southern coasts of Spain, precise knowledge of the presence of the vector of the urogenital schistosomiasis, the bulinid B. truncatus, in our freshwaters, is of great importance. The suggested dispersion of B. truncatus is most likely due to humans and birds. Neolithic colonists started to migrate around 10,500 years BCE and colonized the European continent. The synchronicity of the demographic growth with this phase of early human expansion and the exclusion of other factors suggests that Neolithic settlers, or traders, acted as vectors for the snails. Genetic analysis of Sardinian human populations connect Sardinia and Northern Africa through early human migrations (Jesse et al., 2011). Neither can it be ruled out that the period of Arab colonization played a role regarding the introduction of B. truncatus in the Iberian Peninsula. Birds may also be implicated

UTM 29TNE4952 29TNF3555 29TNF2656 29TNF2561 29TNF3134 29TNF3123 29TNE2999 29TNE5766 29TNE1046 29TNE5247 29TNG4358 29TNF3957 29SNB4916 29TNE2647 29SNB3930 29TNE4752 29SNB7706

in the passage from Africa to Southern Europe as passive transporters of the snail on their wings, feathers or legs, (Nobre, 1941; Russell–Hunter, 1978; Giusti et al., 1995; Boyer and Audibert, 2007; Jesse et al., 2011; Valledor and González, 2014; Neubauer et al., 2017). The Spanish Society of Ornithology (SEO, 2017) has notified the arrival to Spain of numerous species of African birds as a result of global warming. The passive transport of clusters or juveniles by insects such as larger water beetles (Dytiscus spp.) is also possible (Russell–Hunter, 1978). B. truncatus is a ubiquitous species with a high capacity for self–fertilization. It lives in Spain in coastal environments such as lagoons, upwellings, and river mouths where there is little current or low velocity (Martínez–Ortí et al., 2015). When an individual reaches a new habitat it grows quickly and continuously, enabling it to expand its geographical area of occupation, even becoming a pest. Its eradication is practically impossible. Isolated snails in the natural environment grow even more quickly and produce many more clusters than when the population density increases and forms colonies, probably due to the absence of copulation (Bayomy and Joosse, 1987). In general, extinction in many of the known populations has been due to the destruction of the coast following the construction of urban areas, infrastructures, recreational areas and marinas, as well as the purposeful desiccation of water bodies to avoid the transmission of diseases transmitted by mosquitoes (fig. 2, tables 1, 2; Martínez–Ortí et al., 2015).

Animal Biodiversity and Conservation 42.2 (2019)

10º 0' W

8º 0' W

6º 0' W

4º 0' W

263

2º 0' W

0º0'W

2º 0' W

4º 0' W

42º 0' N

40º 0' N

38º 0' N

N W

E S

36º 0' N

100

200 km

Altitude > 1 375 m 750 m 1.125 m 1.500 m

Fig. 4. Risk map of urogenital schistosomiasis vectorized by Bulinus truncatus in Spain and Portugal. Dotted numbers correspond to the locality codes shown in table 1. Fig. 4. Mapa del riesgo de contraer esquistosomiasis urogenital transmitida por Bulinus truncatus en España y Portugal. Los números de los puntos corresponden con los códigos de las localidades que figuran en la tabla 1.

This disease is geographically endemic in subtropical and tropical areas, such as the Caribbean and the eastern coast of South America, Africa and the Middle East (Crompton and Peters, 2010). The recent Corsica epidemic shows how African immigrants from regions where this parasitic disease is endemic have introduced the disease in Europe. This could occur in Spain if eggs of the parasite from infected individuals spread through urine in bodies of fresh water where the appropriate vector snails are found, thus establishing an autochthonous transmission upon completion of their biological cycle. In view of the above, knowledge of the exact location of populations of B. truncatus in Spain and Portugal as shown in this paper confer exceptional interest from the point of view of public health. Knowing these locations will allow us to take specific and effective measures to try to eradicate these populations that are susceptible to transmitting urogenital schistosomiasis. The generated risk map will allow the development of models to predict the optimal area for this species in Spain and Portugal, and identify zones at greatest risk of establishment and expansion of the vector and the disease. Greater efforts are thus needed to sample the areas where it has been cited so as to confirm its persistence in these areas. It would be interesting to review its presence in tourist areas such as the Albufera de l'Alcudia in Mallorca (Balearic Islands), the

Natural Park of Doñana (Andalusia), the Albufera of Valencia, the deltas of rivers such as the Duero, Tajo, Guadiana and Ebro, and also the abundant coastal wetlands where B. truncatus can establish and thrive. Acknowledgements To Dr. Francesc Uribe, curator of molluscs, and Miguel Prieto, data manager, at the 'Museu de Ciències Naturals' of Barcelona, to Dr. Rafael Araujo, curator of molluscs at the 'Museo Nacional de Ciencias Naturales' of Madrid, for allowing us to review the samples deposited in their respective collections. Also to Dr. Emilio Rolán and Sebastián Torres Alba for information about new populations. The study was financed by Project nº RD12 / 0018/0013, Cooperative Research Network in Tropical Diseases–RICET, IV National Program of I+D+I 2008–2011, ISCIII–Subdirectorate General for Networks and Cooperative Research Centers and Funds FEDER, Ministry of Health and Consumption, Spain; PROMETEO nº2012/042 Aid Program for Research Groups of Excellence, Generalitat Valenciana (Spain) and Health Research Project No. PI16/00520, Subprograma Estatal de Generación de Conocimiento de la Acción Estratégica en Salud (AES), Plan Estatal de Investigación Científica y Técnica y de Innovación, ISCIII–MINECO, Madrid, Spain.

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Differential distribution of resources for females on a dioecious plant affects the small–scale distribution of male of an oligolectic bee C. Polidori, M. Federici

Polidori, C., Federici, M., 2019. Differential distribution of resources for females on a dioecious plant affects the small–scale distribution of male of an oligolectic bee. Animal Biodiversity and Conservation, 42.2: 267–277. Doi: https://doi.org/10.32800/abc.2019.42.0267 Abstract Differential distribution of resources for females on a dioecious plant affects the small–scale distribution of male of an oligolectic bee. Females of the solitary digger bee Andrena florea Fabricius, 1793 (Hymenoptera, Andrenidae) nest in aggregations and collect pollen almost exclusively on dioecious plants of the genus Bryonia, making this species a good model to study the relationship between nest density, male density, male behaviour and female lecty. At a study site in the valley of the river Serio, in Italy, an aggregation of this bee showed low density of randomly distributed nests and was closely surrounded by B. dioica plants. Female nectar foraging and male feeding and mate–searching activity, confined to the host plants, peaked at similar hours across the day, while female pollen foraging peaked earlier. Males fed on plants of both sexes but seemed to perch waiting for females more frequently on male B. dioica leaves. Individual males more often visited only one of the male plants, for up to four days; here they did not interact aggressively with conspecifics, suggesting scramble competition in resource–based home ranges and not territoriality. These findings are preliminarily in accordance with the predicted resource–based rendez–vous sites at low nest density for oligolectic bees and the predicted occurrence of scramble competition in case of high male density. Additionally, males would maximize their mating opportunity by mainly perching on male plants, the only source of the most limited resource for females (pollen). Key words: Mating strategy, Digger bee, Dioecious plant, Bryonia, Andrena, Andrenidae Resumen Las diferencias en la distribución de los recursos para las hembras, en una planta dioica, afectan a la distribución a pequeña escala de los machos de una abeja oligoléctica. Las hembras de la abeja excavadora solitaria Andrena florea Fabricius, 1793 (Hymenoptera, Andrenidae) anidan en agregaciones y recolectan polen casi exclusivamente de las plantas dioicas del género Bryonia, lo que hace de esta especie un buen modelo para estudiar la relación entre la densidad de nidos, la densidad de machos y su comportamiento y la especialización trófica de las hembras. En una localidad del valle del río Serio, en Italia, se estudió una agregación de esta abeja con una baja densidad de nidos distribuidos al azar y rodeada muy de cerca por plantas de B. dioica. Las actividades de alimentación y de búsqueda de pareja de los machos, que se realiza solo sobre las plantas hospedadoras, y la actividad de recolección de néctar de las hembras alcanzaron su máximo en horas similares a lo largo del día; no obstante, la recolección de polen, realizada por las hembras, alcanzó su máximo en horas más tempranas. Los machos se alimentaban en plantas de ambos sexos, pero aparentemente se posaban con mayor frecuencia en las hojas de las plantas masculinas de B. dioica para esperar a las hembras. Los machos visitaron más frecuentemente solo una de las plantas masculinas, donde llegaban a permanecer hasta cuatro días y donde no interactuaron entre ellos de manera agresiva, lo que sugiere competencia por acaparamiento en las áreas de distribución determinadas por los recursos y no territorialidad. En principio, estos resultados coinciden con la predicción de que los encuentros tendrían lugar en sitios donde se encuentran los recursos y la densidad de nidos es baja en el caso de las abejas oligolécticas, y con la predicción de competencia por acaparamiento cuando la densidad de machos es elevada. Asimismo, los machos aprovecharían al máximo su oportunidad de apareamiento principalmente visitando plantas masculinas, que son la única fuente del recurso más limitante para las hembras (polen). ISSN: 1578–665 X eISSN: 2014–928 X

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Palabras clave: Estrategia de apareamiento, Abeja excavadora, Planta dioica, Bryonia, Andrena, Andrenidae Received: 30 X 18; Conditional acceptance: 21 I 19; Final acceptance: 04 III 19 Carlo Polidori, Instituto de Ciencias Ambientales (ICAM), Universidad de Castilla–La Mancha, Avenida Carlos III s/n., 45071 Toledo, Spain.– Matteo Federici, Associazione 'APILOMBARDIA', Via Emilia 74, 27058 Voghera (PV), Italy. Corresponding author. E–mail address: carlo.polidori@uclm.es

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Introduction Behaviour of male bees (Apoidea) has been studied in many species, showing that they use a wide range of behavioural strategies to find mates (reviewed in Alcock et al., 1978; Eickwort and Ginsberg, 1980; Ayasse et al., 2001; Paxton, 2005). Such strategies are mainly driven by the fact that females select males to maximise their individual reproductive success, and by the fact that receptive females, which are generally prone to mate only for a short time after emergence, are a major limiting resource (Alcock et al., 1978). According to these considerations, the evolution of mating strategies in bees is expected to be driven, and thus somehow predicted, by three factors: the distribution in space and time of receptive females; the ratio of males over females; and female reproductive life–history (Ayasse et al., 2001; Shuster and Wade, 2003; Paxton, 2005). Regarding the first factor, if females nest in dense aggregations, males are expected to find a partner at nesting sites, while if females nest in low–density, males should look for a mate on flowers (resource sites) or non–resource sites (Alcock et al., 1978; Ayasse et al., 2001) (fig. 1). In reference to the second, male scramble competition (i.e. patrolling the areas visited by females without establishing territories) would be favoured if male density is high, while male territoriality would be favoured if male density is low (Paxton, 2005), with males in the former showing no fixed home ranges and males in the latter case showing defined territories that are actively defended from conspecific competitors (fig. 1). And in reference to the third factor, if the level of specialization of pollen resource by females is taken into account, at low female (nest) density, males of oligolectic bees would use resource–based rendez–vous sites while males of polylectic bees would use non–resource based sites (Paxton, 2005) (fig. 1). This is because receptive females of oligolectic species more predictably aggregate at their host flowers whereas females of polylectic species are more likely to disperse widely across flowering plant species (but see exceptions in Westrich, 1989; Danforth, 1991; Seidelmann, 1999). Thus, the different combinations of these traits lead to different outcomes in male behaviour: 1) when nest density is high and male density is low, male territoriality is expected at the nesting site; 2) when nest density is low and male density is high, male scramble competition at resource–based sites (in case of oligolectic species) or at other sites is more likely; and 3) when nest density is low and male density is low we can expect male territoriality at resource–based sites (in case of oligolectic species) or in other sites (in case of polylectic species) (fig. 1). Although these predictions have generally been confirmed by comparative analyses of literature data (see above), particular situations have been poorly studied in detail. For example, it is still little known whether, in the case of oligolectic bee species foraging on plant species presenting individuals of different sexes (dioecious), males would equally patrol/control both male and female plants, or whether there is a bias towards the male plants, because they reward the most limited

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resource for females (pollen) (fig. 1). Increasing our knowledge on specialized bee behaviour is important for conservation plans, particularly because pollinators that rely on few plants for their resources have declined impressively in the last years (Memmott et al., 2004). Here, we report observations on the behaviour of males and females of the solitary digger bee, Andrena florea Fabricius (Andrenidae), an oligolectic species whose females collect pollen almost exclusively from the delicate white blossoms of bryony (Bryonia, Cucurbitaceae), a genus of dioecious plants (Westrich, 1989; Schröder and Lunau, 2001) and –though more rarely– from species of Ecballium, the sister genus to Bryonia (Dukas, 1987). Andrena florea belongs to one of the richest bee genera (roughly 1300 species worldwide, Gusenleitner and Schwarz, 2002; Michener 2007). It is a univoltine species in Central and Southern Europe on flight from May to June, with males emerging some days earlier than females (Westrich, 1989). Females are solitary and dig nests in the soil at variable distances from their plant hosts (Edwards and Williams, 2004). While females are strictly associated with Bryonia, males have been observed on this plant but also on plants belonging to other genera, such as Echium, Rubus, Crataegus, Hieracium (Gadoum and Didier, 2008) and Lithospermum (Baldock, 2008). Because males have been sometimes observed in large numbers on these plants, it has been suggested that they could be mating sites (Baldock, 2008). Although the mating system of a number of species of Andrena Fabricius, 1775 was studied in detail (e.g. Westrich, 1989; Paxton and Tengö, 1996; Paxton et al., 1999), as far as we know, no detailed information is available on A. florea mating strategy, the object of the present study. Additionally, given the pronounced oligolecty of the studied bee species and the reproductive biology of Bryona, we argue that males would visit the male plants rather than the female plants, because female bee are likely to be more aggregated on plants that offer their most limiting resource (pollen). The aim of this study was to study behavioural aspects that could be relevant to the better understanding of the mating strategy of A. florea, and used for future tests of the predictions presented above on mating strategy. In particular, in order to evaluate the presence of territoriality or scramble–competition, we evaluated: 1) female nest aggregation size and spatial arrangement of nests; 2) female pollen and nectar foraging on plants throughout the day and in relation to the number of flowers; 3) male foraging and perching on leaves of plants throughout the day and in relation to the number of flowers; and 4) behaviour of marked males on plants, Material and methods Field data collection The field study was conducted at the 'Stazione Sperimentale per la Conservazione Della Flora di Pianura', in the 'Castelleone wood', an area of ca. 15 ha, 55–64 m a.s.l. in the valley of the river Serio

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(45º 18′ 0″ N, 9º 46′ 0″ E). The area has the typical climate of the Padana plain, with cold winters (1.5 ºC on average in January), hot summers (24 ºC on average in July) and generally high air humidity. Rainfall is 600–1,000 mm/year and rather evenly distributed during the year, with July and February being the months with the least rain and October the wettest month (100 mm). In a path leading into the wood, a small nest aggregation of A. florea (21 nests in about 15 m2 at the end of the study) was found in the proximity (about 30 m from its far northern edge) of about ten of B. dioica plants, arranged in a row. Field observations were carried out from 28 April to 2 May 2004 at the bee nesting site, and from 3 to 13 May 2004 in front of four B. dioica plants, two male plants and two female plants, at about 1–2 m distance from each other. At the nesting site, the number of nests of A. florea was counted and the nest coordinates were recorded in a Cartesian system (error: ± 0.5 cm) with the help of a 1 m x 1 m mesh mounted on the ground (Polidori et al., 2006, 2008). Regarding the plants, we counted the total number of flowers for each plant on the first and last day of the study. We also calculated the mean and used this as an estimate of plant size. A total of 20 male bees were netted on plants, individually marked on the thorax using a combination of colours with non– toxic paint (Uni–Posca®), and then released. During the observations on plants, consisting of 15–min observations for each plant each hour from 9:00 to 18:00 (GMT), we recorded the following events: 1) male feeding (marked and not marked males landing on a flower and feeding on nectar; 2) male perching (marked and unmarked males landing on a leaf and staying there for at least 5 seconds); and 3) female foraging (females landing on a flower and gathering pollen or nectar). While it was easy to score whether a female bee foraged on pollen because of its presence on the scopae, we could not distinguish whether pollen–foraging bees were also collecting nectar. Thus, we could only separate nectar foraging from pollen or pollen + nectar foraging (hereafter: pollen foraging). For marked males, we recorded the number of days they were recaptured on each of the four plants. In this case all the observations for each male (feeding and perching) were pooled for the analysis. Statistical analysis For each nest at the aggregation, we calculated the distance to the nearest nest (nearest neighbour distance, NND). We then obtained nest density by dividing the total number of nests by the nesting area (a rectangle including all nests and a buffer of 10 cm from the extreme nests on all directions). We used Ripley’s K–distribution (or K–function) (Ripley, 1976) to analyse the spatial distribution of nests. The K–distribution is the cumulative frequency distribution of observations at a given point–to–point distance (or within a distance class). Because it preserves distances at multiple scales, Ripley’s K can quantify the intensity of patterns at multiple (different distan-

ces/scales) scales. Ripley’s analysis was performed using the Duncan program (Duncan, 1991) and the results obtained were transformed following the function L(t) = (K(t)/π)1/2 – t, where t is the spatial step (scale; in this study 25 cm) with which the test was performed (Haase, 1995). If the observed L(t) curve stays above the upper envelope of the expected curve for randomness for a given distance (step), the distribution is clumped (nests are in clusters), while if it stays below the lower envelope, the distribution is regular (nests are dispersed). The behavioural data were analysed using non–parametric statistics because conditions for parametric tests were not met. The difference in male activity (number of records for feeding or perching) between the four observed plants was tested using the x2–test. Differences in the hourly distribution of the frequency of female nectar feeding, female pollen foraging, male nectar feeding and male perching were tested using Kruskall–Wallis followed by multiple pairwise comparisons using Dunn's procedure. We used the The Spearman test was used to verify the linear correlation between the number of perching records and the number of perched plants across marked males. Results Nest density in the aggregation was 1.5 nests/m2, and nests were randomly distributed within the area, as the L(t) function fell within the confidence envelope at all steps (fig. 2A–2B). Females were frequently seen at nest entrance (fig. 2C) or returning to the nests after a foraging trip, but no males were observed on the nesting site. Bryonia dioica plants observed in this study contained from 17 to 59.6 flowers each (average of the first and last counts) and were visited frequently by males and females of A. florea (119–429 contacts in total during the study, 19.8–71.2 contacts per day). The most commonly observed behaviour was male feeding (n = 429), followed by male perching (n = 188) and female foraging (n = 119), particularly nectar–only feeding (n = 98). The distributions of these activities across the hours of the day were roughly similar, starting at 10.00, peaking at mid–day, and stopping after 17:00 (fig. 3). A closer inspection revealed, however, that female pollen foraging peaked between 11:00 and 12:00 and female nectar foraging peaked between 13:00 and 15:00 (fig. 3). On the other hand, male perching activity spanned the largest temporal window during the day (12:00–16:00), while males fed more frequently between 13:00 and 15:00 (fig. 3). Median values for periods of activity were invariably 14:00 for all groups with the exception of female pollen foraging, which had a median of 12:00 (fig. 3), and a Kruskall–Wallis test revealed that overall activity of females and males largely overlapped during the day, but female pollen foraging occurred significantly earlier (K = 31.3, 21 < n > 429, df = 3, P < 0.0001; multiple pairwise comparisons using Dunn’s procedure: P < 0.0001 between pollen foraging and all the others categories, and 0.33 < P > 0.81 between all the other pairs of categories).

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M Fig. 1. Schematic representation of the expected relationships between female nest density, male density, male behaviour (S, scramble competition; T, territoriality), female lecty (O, oligolectic; P, polylectic) and plant reproductive biology (M, monoecious; D, dioecious). Rendez–vous can occur at the female nesting site (N), at resource–based sites (R), or in other sites (O). In case of R–sites, rendez–vous can occur mostly on male plants (m) or plants of both sexes (all). See text for details. Fig. 1. Representación esquemática de las relaciones esperadas entre la densidad de nidos de las hembras, la densidad de los machos, el comportamiento de los machos (S, competencia por acaparamiento; T, territorialidad), la especialización en el uso de las plantas por parte de las hembras (O, oligoléctica; P, poliléctica) y la biología reproductiva de la planta (M, monoica; D, dioica). Los encuentros entre machos y hembras pueden ocurrir en el sitio de anidación de las hembras (N ), en sitios donde se encuentran los recursos para las hembras (R) o en otros lugares (O). En el caso de sitios R, los encuentros pueden ocurrir principalmente sobre plantas masculinas (m) o sobre plantas de ambos sexos (all). Véase el texto para obtener información detallada.

We recorded a total of 103 and 16 female visits, respectively, on male plants and female plants, and a total of 472 and 145 male visits, respectively, on male plants and female plants. Male feeding and male perching (fig. 4A) both seemed to be most abundant on the male plant with the greatest number of flowers (about 59) (fig. 4B). The x2–test showed that the numbers of both perching and feeding were significantly higher on that male plant (x2 = 22.02 and χ2 = 35.07, respectively (df = 3)). The other three plants had a similar number of flowers (about 17–21) and were visited by males with similar frequency (fig. 4B). Males may prefer male plants over female plants because they are the only sex providing pollen for female bees, or because they simply aggregate on plants with more flowers to feed on. Although our sample size was too small to test for either possibility, males still seemed to perch more on male plants once having taken into account the number of flowers per plant; the number of perching records on male plants (per flower) was almost four times

higher than that on female plants (4.21 vs. 1.24). On the contrary, males seemed to feed more equally on both male (8.14) and female (5.82) plants. Some of the 20 marked males seemed to have a clear preference for one of the two male plants observed during the study (fig. 5). In particular, six males were recorded on a single plant (or almost on a single plant, > 90 % of records) for at least three days (up to four days) (fig. 5). Ten additional males were only recorded on one or two days, and always on a single plant, while the remaining four males were observed shifting plants, even in a single day (fig.4). Even in these cases males seemed to spend more time on one of the two plants (66 %–88 % of records) (fig.5). The number of visited plants by a marked male may also depend on the number of visits recorded for that male during the study, so that males that only visited one plant may be those with a low number of observations. However, this seems unlikely since the number of visited plants did not

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Fig. 2. A, map of nests in the studied aggregation of A. florea; B, Ripley's L(t) function for the A. florea nest aggregation (bold line: transformed Ripley's K function (L(t) = (K(t)/ᴨ)1/2 – t), calculated from observed data and for given spatial steps (scale); thin line: upper 95 % confidence envelope; dashed line: lower 95 % confidence envelope; step used was 25 cm); C, A. florea female protruding from its nest. Fig. 2. A, mapa de nidos en la agregación estudiada de A. florea; B, función L(t) de Ripley para la agregación de nidos de A. florea (línea en negrita: función K transformada de Ripley (L(t) = (K(t)/ᴨ)1/2 – t), calculada a partir de datos observados y para determinados pasos espaciales (escala); línea delgada: límite superior del intervalo de confianza del 95 %; línea discontinua: límite inferior del intervalo de confianza del 95 %; el paso espacial utilizado fue de 25 cm); C, A. florea hembra que se asoma desde su nido.

increase with the number of records across marked males (Spearman correlation test, ρ = 0.21, n = 13, P = 0.49) (we excluded the seven males with only one record). Discussion The relationship of A. florea and plants of the genus Bryonia is probably as old as the long history of Bryonia in Eurasia (L. Larkin, personal communication quoted in Volz and Renner, 2008) and there is great overlap between ranges of these bees and their host plants (Westrich, 1989; Schröder and Lunau, 2001). This strict relationship is thus likely to have influenced the behaviour of the bees, not only in terms of foraging

behaviour by females (e.g. pollen collection), but also in terms of male mating tactics. With this study, we provide quantitative data on the behaviour of males and females of A. florea in a first attempt to understand the relationship between male density, female density, and pollen specialization by females. These are, however, provisional results that can be biased due to the small sample size, so that it is necessary to obtain additional data to perform a formal test of the predictions on mating strategy based on these variables. In any case, our results at least suggest that the hypothesis that the evolution of reproductive systems and strategies is driven by the distribution in space and time of receptive females and by female reproductive life–history (Shuster and Wade, 2003) may also be valid for A. florea.

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Fig. 3. Daily patterns of A. florea female foraging (N, nectar; P, pollen; P+N, pollen and nectar), male foraging (N, nectar) and mate–searching behaviour (patrol, patrolling the plant; perch, perching on plant leaves). Fig. 3. Patrones diarios de recolección de las hembras de A. florea (N, néctar; P, polen; P+N, polen y néctar), de recolección de los machos (N, néctar) y de comportamiento de búsqueda de pareja (patrol, patrullando las plantas; perch, posándose en las hojas de la planta).

In our studied population of A. florea, females nested in a low–density, randomly arranged aggregation, and males did not mate at the nesting site. This agrees with the prediction that if females nest in low–density, males would search for a mate at flowers (resources) or non–resource sites rather than at nesting sites (Alcock et al., 1978; Ayasse et al., 2001; Paxton, 2005; see fig. 1). The small observed nest aggregation departed from the more common trend of a clumped arrangement of nests observed in fossorial bees and wasps (Potts and Willmer, 1998; Polidori, 2017; Polidori et al., 2008; Asís et al., 2014). However, at least a couple of Andrena species (A. wilkella (Kirby, 1802) and A. nigroaenea (Kirby, 1802)) were also observed nesting in a random pattern in loose aggregations (Tengö et al., 1990; Schiestl and Ayasse, 2000). The studied nest aggregation was found close to B. dioica plants, in accordance with the fact that foraging ranges of Andrena bees of comparable body size to A. florea (11–12 mm of body length) does not generally exceed 1000 m radius from nests (Zurbuchen et al., 2010). The distance between nests and host plants for another population of this species was indeed reported to be about 1 km (Edwards and Williams, 2004), though it is likely to be more common to find lower distances. Males were only observed to search for females exclusively on B. dioica flowers, according to the prediction that, at low female (nest) density, ma-

les of oligolectic bees would use resource–based, rendez–vous sites (Paxton, 2005). Female pollen specialization in this species can be visible even in the adjustment of the female bee foraging behaviour at a daily scale: females collect pollen considerably earlier in the morning (peak at 11:00–12:00) than their collection of nectar (peak at 14:00–15:00), i.e. when pollen presentation by the staminate flowers is also peaking (Schröder and Lunau, 2001). Previous observation on foraging behaviour by females of this species reported similar daily curves, with pollen collection peaking even earlier, at 10:00–11:00 (Schröder and Lunau, 2001). Differences between populations in the time of pollen foraging may be related to temperature or other climatic or biotic factors which are as yet unknown. Males of A. florea were abundant on plants and were not aggressive towards one another, with no evidence of territorial behaviour, likely reflecting the unpredictable distribution of females (Alcock et al., 1978). Theory predicts that male scramble competition would be favoured in case of high male density (Alcock et al., 1978; Ayasse et al., 2001; Paxton, 2005, see fig. 1). However, although a lack of territorial behaviour generally leads to no fixed home ranges in individual males (Paxton, 2005) it seems that a certain temporal constancy in the use of perched plants (up to several days) occurs in A. florea males. This would agree with findings in some

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Fig. 4. A, A. florea male perching on a B. dioica leaf on a male plant; B, number of flowers on the four studied plants plotted against the number of male activity records (perching and feeding) on the plants. Median values calculated across plants are shown. Fig. 4. A, macho de A. florea posado en una hoja de una planta masculina de B. dioica; B, número de flores en las cuatro plantas estudiadas y registros de actividad de los machos (posado y alimentación) en las plantas. Se muestran los valores medianos calculados por planta.

other bee species, where males may show temporally constant home ranges (at least within scales of some days) but do not defend them (thus they are not territorial) (Willmer et al., 1994). The home ranges of marked males seemed to be on male plants only, in keeping with our hypothesis that males would mainly patrol/control the plants that reward the most limited resource for females (pollen). The fact that males most often use male plants for mating purposes is supported by their nectar–feeding distribution, which seemed to depend more on flower number on the plant. If this pattern is confirmed in a larger dataset, this would be the first case of an oligolectic bee species, as far as we know, showing this bias of male mate–searching on male plants. Variation in male home range size is common in bees. Males of some species confine their search to one or only a few plants, while those of other species can use a much larger area (Free and Butler, 1959; Stiles, 1976). For example, large home ranges (> 49 m2) of male Andrena erigeniae Robertson, 1891 are set on the females' strictly–associated host plant Claytonia virginica, with individually marked males overlapping their areas (Barrows, 1978). However, the much larger home ranges recorded for males of this species

(compared to those found here for A. florea) may be due to the different reproductive biology of this plant, where each flower acts as a male (pollen available to pollinators) for approximately one day and then acts as a female for one day or more (Motten et al., 1981). Thus, there would be no benefit for males of A. erigeniae to concentrate on single plants while seeking a partner, but it would be beneficial for A. florea males to focus on male plants (leading to smaller ranges). As for A. florea, males of some other bee species spend most of their time perched or hovering, waiting for females (Alcock et al., 1978; Paxton, 2005). It is not known if A. florea males mark B. dioica leaves with female–attracting pheromones, as was observed in males of other bee species perching on plants waiting for females (Alcock et al., 1978; Paxton, 2005), or if they just increase the probability to intercept a receptive female by perching most often on male plants. Marking secretions from males have been reported from a few Andrena species (Priesner, 1973; Bergström et al., 1982; Tengö et al., 1990), opening the possibility that this behaviour also occurs in A. florea. It is also unknown whether virgin females possess a cuticular hydrocarbon profile that signals to (or is used as a cue by) males, thus guiding them

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Fig. 5. Percent frequency of recaptures on the two male plants (grey or black) for each of the 20 marked males and for five subsequent days. If no percentage is given, the male stayed that day only on one of the two plants (100% of recaptures). Numbers on the right represent the sample size (number of observations) for each of the marked males. Fig. 5. Frecuencia (%) de recapturas en las dos plantas masculinas (gris o negro) de cada uno de los 20 machos marcados y durante cinco días seguidos. El porcentaje no aparece si ese día el macho solo permaneció en una de las dos plantas (100 % de las recapturas). Los números a la derecha representan el tamaño de la muestra (número de observaciones) de cada uno de los machos marcados.

to copulate only with receptive females. Such a situation is known for A. nigroaenea, for which males were observed to search for females on the nesting site rather than on host plants (Schiesti and Ayasse, 2000). Other Andrena species show males searching for females on the nesting sites, such as A. erythronii Robertson, 1891 (Michener and Rettenmeyer, 1956), A. flavipes Panzer, 1799 (Butler, 1965) and A. vaga Panzer, 1799 (Rezkova et al., 2012). However, available, though often scattered, data seem to show that Andrena males patrol and mate on flowers more frequently than at other places (reviewed in Barrows, 1978; see also Paxton, 2005). Furthermore, territorial behaviour was observed in males of very few Andrena species (e.g. A. foxii Cockerell, 1898; Thorp, 1969; Linsley et al., 1973). Our findings on A. florea share some observations with those on other oligolectic Andrena species. For

example, a population of A. erigeniae nesting in a low– to moderate–density aggregation (1–21 nests/m2) has occasionally been collected on a number of plant species, but females appear to be entirely restricted for pollen to C. virginica, and data from collecting records indicate that mating probably takes place on flowers of this exclusive pollen–rewarding host plant (Davis and LaBerge, 1975). Andrena agilissima (Scopoli, 1770) is also known to have a resource– based rendez–vous site (Paxton, 2005) (and even intra–nidal mating upon adult emergence, Paxton et al., 1999), but the studied nest aggregation of this species was very large and dense (Giovannetti et al., 2003; Polidori et al., 2005). On the other hand, Stephen (1966) reported that males of Andrena perplexa Smith, 1853 patrol the nesting site and most copulas occurred there. Andrena nigrae Robertson, 1905 and A. mariae Robertson, 1891 were observed

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while mating on Salix spp., their only source of pollen (Miliczky, 1988). The association between lecty status and rendez–vous site does not seem, however, to be very clear, since cases of polylectic bee species using resource–based rendez–vous sites are also known (review in Paxton, 2005). An extreme example is Andrena prunorum Cockerell, 1896, whose males were observed patrolling cultivated cherry, peach, pear, and apple, according with the great polylecty (77 host plant genera in 32 families) shown by females (Miliczky, 2008). Classification of mating tactics in relation to bee biological traits itself is sometimes difficult, given that certain variation in mating tactics was reported even at an intra–specific level for several Andrena species (e.g. Paxton, 2005; Paxton et al., 1999, Paxton and Tengö, 1996). In conclusion, in this paper we provide novel information on the behaviour of A. florea, a European species of oligolectic bee strictly associated with Bryonia plants. Our data can be used for further analysis of the mating strategy of this species, including a more formal and quantitative test of the predictions arising from the interaction between female and male densities and resource specialization in bees. Interestingly, we found certain home range constancy exclusively on male plants, and we suggest that the reproductive biology of the host plant should be taken into account in new studies on bees in order to see how widespread this phenomenon is. Detailed information on the behaviour of wild bees is fundamental as an element to take into consideration while making conservation plans that aim to protect pollinators. Acknowledgements CP was funded by a post–doctoral contract from the Universidad de Castilla–La Mancha. We thank Cristina Papadia and Stefania Bevacqua for helping during the field observations. References Alcock, J., Barrows, E. M., Gordh, G., Hubbard, L. J., Kirkendall, L., Pyle, D. W., Ponder, T. L., Zalom, F. G., 1978. The ecology and evolution of male reproductive behaviour in the bees and wasps. Zoological Journal of the Linnean Society, 64: 293–326. Asís, J. D., Ballesteros, Y., Tormos, J., Baños–Picón, L., Polidori, C., 2014. Spatial nest–settlement decisions in digger wasps: conspecifics matter more than heterospecifics and previous experience. Ethology, 120: 340–353. Ayasse, M., Paxton, R. J., Tengö, J., 2001. Mating behavior and chemical communication in the order Hymenoptera. Annual Review of Entomology, 46: 31–78. Baldock, D. W., 2008. Bees of surrey. Surrey Wildlife Trust, Pirbright. Barrows, E. M., 1978. Male behavior in Andrena erigeniae (Hymenoptera: Andrenidae) with comparative notes. Journal of Kansas Entomological

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Society, 51: 798–806. Bergström, G., Tengö, J., Reith, W., Francke, W., 1982. Multicomponent mandibular gland secretions in three species of Andrena bees (Hym., Apoidea). Zeitschrift für Naturforschung C, 37c: 1124–1129. Butler, C. G., 1965. Sex attraction in Andrena flavipes Panzer (Hymenoptera: Apidae), with some observations on nest–site restriction. Proceedings of the Entomological Society of Washington, 40: 77–80. Danforth, B. N., 1991. Female foraging and intranest behavior of a communal bee, Perdita portalis (Hymenoptera: Andrenidae). Annals of the Entomological Society of America, 84: 537–548. Davis, L. R. Jr, LaBerge, W. E., 1975. The nest biology of the bee Andrena (Ptilandrena) erigeniae Robertson (Hymenoptera: Andrenidae). Illinois Natural History Survey Biological Notes, 95: 1–16. Dukas, R., 1987. Foraging behavior of three bee species in a natural mimicry system: Female flowers which mimic male flowers in Ecballium elaterium (Cucurbitaceae). Oecologia, 74: 256–263. Duncan, R. P., 1991. Competition and the coexistence of species in a mixed podocarp stand. Journal of Ecology, 69: 559–564. Edwards, M., Williams, P., 2004. Where have all the bumblebees gone and could they ever return? British Wildlife, 15: 305–312. Eickwort, G. C., Ginsberg, H. S., 1980. Foraging and mating behavior in Apoidea. Annual Review of Entomology, 25: 421–446. Free, J. B., Butler, C. G., 1959. Bumblebees. Macmillan, New York. Gadoum S., Didier, B., 2008. Andrena florea et al bryone. Insectes Pollinisateurs, 150: 23–24, https://studylibfr.com/doc/698502/andrena-florea-et-la-bryone Giovanetti, M., Scamoni, E., Andrietti, F., 2003. The multi–entrance system in an aggregation of Andrena agilissima (Hymenoptera; Andrenidae). Ethology Ecology & Evolution, 15: 1–18. Gusenleitner, F., Schwarz, M. 2002. Weltweite Checkliste der Bienengattung Andrena mit Bemerkungen und Ergänzungen zu paläarktichen Arten (Hymenoptera, Apidae, Andrenidae, Andrena). Entomofauna, Suppl. 12: 1–1280. Haase, P., 1995. Spatial pattern analysis in ecology based on Ripley’s K–function: introduction and methods of edge correction. Journal of Vegetation Science, 6: 575–582. Linsley, E. G., MacSwain, J. W., Raven, P. H., Thorp, R. W., 1973. Comparative behavior of bees and Onagraceae V. Camissonia and Oenothera bees of Cismontane California and Baja California. University of California Publications in Entomology, 71: 1–68. Memmott, J., Waser, N. M., Price, M. V., 2004.Tolerance of pollination networks to species extinctions. Proceedings of the Royal Society B, 271: 2605–2611. Michener, C. D., Rettenmeyer, C. W., 1956. The ethology of Andrena erythronii with comparative data on other species (hymenoptera, Andrenidae). The University of Kansas Science Bulletin, 37: 645–684.

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Michener, C. D., 2007. The Bees of the World (2nd Edition). The Johns Hopkins University Press, Baltimore and London. Miliczky, E. R., 1988. Observations on the bionomics of the bee Andrena (Tylandrena) erythrogaster Ashmead (Hymenoptera: Andrenidae) with notes on A. (Micrandrena) personata Robertson and A. (Holandrena) c. cressonii Robertson. Illinois Natural History Survey Biological Notes, 130: 1–18. – 2008. Observations on the nesting biology of Andrena (Plastandrena) prunorum Cockerellin Washington state (Hymenoptera: Andrenidae). Journal of Kansas Entomological Society, 81: 110–121. Motten, A. F., Campbell, D. R., Alexander, D. E., Miller, H. L., 1981. Pollination effectiveness of specialist and generalist visitors to a North Carolina population of Claytonia virginica. Ecology, 62: 1278–1287. Paxton, R. J., 2005. Male mating behaviour and mating systems of bees: an overview. Apidologie, 36: 145–156. Paxton, R. J., Giovanetti, M., Andrietti, F., Scamoni, E., Scanni, B., 1999. Mating in a communal bee, Andrena agilissima (Hymenoptera Andrenidae). Ethology Ecology & Evolution, 11: 371–382. Paxton, R. J., Tengö, J., 1996. Intranidal mating, emergence, and sex ratio in a communal bee Andrena jacobi Perkins 1921 (Hymenoptera: Andrenidae). Journal of Insect Behavior, 9: 421–440. Polidori, C., 2017. Interactions between the social digger wasp, Cerceris rubida, and its brood parasitic flies at a Mediterranean nest aggregation. Journal of Insect Behavior, 30: 86–102. Polidori, C., Casiraghi, M., Di Lorenzo, M., Valarani, B., Andrietti, F., 2006. Philopatry, nest choice, and aggregation temporal–spatial change in the digger wasp Cerceris arenaria (Hymenoptera: Crabronidae). Journal of Ethology, 24: 155–163. Polidori, C., Mendiola, P., Asís, J. D., Tormos, J., Selfa, J., Andrietti, F. 2008. Female–female attraction influences nest establishment in the digger wasp Stizus continuus (Hymenoptera: Crabronidae). Animal Behaviour, 75: 1651–1661. Polidori, C., Scanni, B., Scamoni, E., Giovanetti, M., Andrietti, F., Paxton, R. J., 2005. Satellite flies (Leucophora personata, Diptera: Anthomyiidae) and other dipteran parasites of the communal bee Andrena agilissima (Hymenoptera: Andrenidae) on the island of Elba, Italy. Journal of Natural History, 39: 2745–2758. Potts, S. G., Willmer, P., 1998. Compact housing in built–up areas: spatial patterning of nests in aggregations of a ground–nesting bee. Ecological Entomology, 23: 427–432. Priesner, E., 1973. Reaktionen von Riechrezeptoren männlicher Solitärbienen (Hymenoptera, Apoidea)

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auf Inhaltsstoffe von Ophrys–Blüten. Zoon Suppl, 1: 43–54. Rezkova, K., Žáková, M., Žáková, Z., Straka, J., 2012. Analysis of nesting behaviour based on daily observation of Andrena vaga (Hymenoptera: Andrenidae). Journal of Insect Behavior, 25: 24–47. Ripley, B. D., 1976. The second order analysis of stationary point processes. Journal of Applied Probability, 13: 255–266. Schiestl, F. P., Ayasse, M., 2000. Post–mating odor in females of the solitary bee, Andrena nigroaenea (Apoidea, Andrenidae), inhibits male mating behavior. Behavioural Ecology and Sociobiology, 48: 303–307. Schröder, S., Lunau, K., 2001. Die oligolektische Sandbiene Andrena florea und die Rote Zaunrübe Bryonia dioica – Schnittstelle zweier spezialisierter Fortpflanzungssysteme. Mitteilungen der Deutschen Gesellschaft für Allgemeine und Angewandte Entomologie, 13: 529–533. Seidelmann, K., 1999. The race for females: the mating system of the red mason bee, Osmia rufa (L.) (Hymenoptera: Megachilidae). Journal of Insect Behavior, 12: 13–25. Shuster, S. M., Wade, M. J., 2003. Mating Systems and Strategies. Princeton University Press, Princeton. Stiles, E. W., 1976. Comparison of male bumblebee flight paths: Temperate and tropical (Hymenoptera: Apoidea). Journal of Kansas Entomological Society, 49: 266–274. Tengö, J., Ågren, L., Baur, B., Isaksson, R., Liljefors, T., Mori, K., Konig, W., Francke, W., 1990. Andrena wilkella male bees discriminate between enantiomers of cephalic secretion components. Journal of Chemical Ecology, 6: 429–441. Thorp, R. W., 1969. Systematics and ecology of bees of the subgenus Diandrena (Hymenoptera: Andrenidae). University of California Publications in Entomology, 52: 1–146. Volz, S. M., Renner, S. S., 2008. Hybridization, polyploidy, and evolutionary transitions between monoecy and dioecy in Bryonia (Cucurbitaceae). American Journal of Botany, 95: 1297–1306. Westrich, P., 1989. Die Wildbienen Baden–Wuerttembergs. Volume 1 and 2. Eugen Ulmer, Stuttgart. Willmer, P. G., Gilbert, F., Ghazoul, J., Zalat, S., Semida, F., 1994. A novel form of territoriality: daily paternal investment in an anthophorid bee. Animal Behaviour, 48: 535–549. Zurbuchen, A., Landert, L., Klaiber, J., Müller, A., Hein, S., Dorn, S., 2010. Maximum foraging ranges in solitary bees: only few individuals have the capability to cover long foraging distances. Biological Conservation, 143: 669–676.

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Seasonality affects avian species distribution but not diversity and nestedness patterns in the urban parks of Vitoria–Gasteiz (Spain) I. de la Hera

De la Hera, I., 2019. Seasonality affects avian species distribution but not diversity and nestedness patterns in the urban parks of Vitoria–Gasteiz (Spain). Animal Biodiversity and Conservation, 42.2: 279–291, Doi: https://doi.org/10.32800/abc.2019.42.0279 Abstract Seasonality affects avian species distribution but not diversity and nestedness patterns in the urban parks of Vitoria–Gasteiz (Spain). Environmental seasonality leads to variation in the composition and structure of bird communities over the year that might alter biodiversity and nestedness patterns in urban parks and other fragmented habitats. These changes could have important implications in the management and conservation of urban green areas and their populations, but they are largely unexplored. In this study, the composition, diversity and nestedness of the breeding and wintering avian communities in 31 urban parks of Vitoria–Gasteiz (Spain) were analysed. Avian diversity was significantly greater during breeding than during the winter period, although the most diverse parks during breeding were also the most diverse during winter. Most of the among–park variation in diversity was explained by park size, while tree density had a marginal contribution that was only significant during winter. Avian communities showed a significant nested subset pattern that was similar between seasons, with these patterns being mainly mediated by park size. Although the distribution of seven out of the 16 species occurring all–year–round changed significantly from one season to the other, the park ranks in the nestedness matrices were strongly correlated between seasons. This was caused by the reduction in the park distribution of some species from one season to the other that was compensated by the expansion of other species that were initially less common. These results support the idea that, in small and medium–sized cities, park size is the main constraint on avian diversity, and the presence of relatively large parks (> 10 ha) should be encouraged to promote a rich avifauna all year round. Key words: Avian migration, Effective number of species, Environmental noise, European Green Capital, Mantel test. Resumen La estacionalidad afecta a la distribución de especies de aves, pero no a los patrones de diversidad y anidamiento en los parques urbanos de Vitoria–Gasteiz (España). La estacionalidad ambiental causa variaciones en la composición y estructura de las comunidades de aves a lo largo del año que podrían alterar los patrones de biodiversidad y anidamiento en los parques urbanos y otros hábitats fragmentados. Estos cambios podrían tener importantes implicaciones en la gestión y conservación de las áreas verdes urbanas y sus poblaciones de aves, que se han estudiado poco. En este estudio se analizaron la composición, la diversidad y el anidamiento de las comunidades de aves reproductoras e invernantes en 31 parques urbanos de Vitoria–Gasteiz (España). La diversidad de aves fue significativamente mayor durante la época reproductiva que en el período invernal, aunque los parques más diversos durante la reproducción también fueron los más diversos en invierno. La mayor parte de la variación de la diversidad entre parques se explicó por el tamaño del parque, mientras que la densidad del arbolado tuvo una contribución escasa que solo fue significativa en invierno. Las comunidades de aves mostraron un patrón de anidamiento significativo y similar en ambas estaciones, que estaba fundamentalmente determinado por el tamaño del parque. A pesar de que la distribución de siete de las 16 especies que están presentes todo el año cambió significativamente de una estación a otra, las posiciones de los parques en las matrices de anidamiento estuvieron estrechamente correlacionadas entre estaciones. Ello es debido a que la reducción de algunas especies en los parques de una estación a otra se vio compensada por el aumento de otras especies que inicialmente eran menos comunes. Estos resultados apoyan las ideas de que, en ciudades ISSN: 1578–665 X eISSN: 2014–928 X

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de pequeño y mediano tamaño, la superficie del parque es el principal factor limitante de la diversidad de aves, y que debería fomentarse la existencia de parques relativamente grandes (> 10 ha) para favorecer una rica avifauna durante todo el año. Palabras clave: Migración de aves, Número efectivo de especies, Ruido Ambiental, Capital Verde Europea, Prueba de Mantel Received: 01 X 18; Conditional acceptance: 10 XII 18; Final acceptance: 11 III 19 I. de la Hera, School of Biological, Earth and Environmental Sciences, University College Cork, Cork T23XA50, Ireland. E–mail: delaheraivan@gmail.com

Animal Biodiversity and Conservation 42.2 (2019)

Introduction Urbanisation is a major driver of environmental deterioration worldwide (McDonald et al., 2013). Apart from the direct damage to the land, urban areas are also primary consumers of energy and other external resources, producing pollutants that affect natural habitats elsewhere (Rickwood et al., 2008) and contributing significantly to what is known as Earth's sixth mass extinction (Ceballos et al., 2017). Urbanisation has also changed people's lifestyles dramatically. It is expected that the increasing number of people living in cities (around 70 % of world's population by 2050; United Nations, 2018) will also spend most of their time indoors. For example, North Americans spend more than 90 per cent of their time in buildings (87 %) or cars (6 %; Klepeis et al., 2001), drastically reducing their chances of interacting with nature. This is not of minor importance from a conservation perspective because direct experiences with the natural world trigger environmental awareness and conservation actions (Dearborn and Kark, 2010). In this scenario, urban ecosystems play a key role as a last–resort bond between urban human population and nature, contributing indirectly to global biodiversity conservation (see pigeon paradox; Dunn et al., 2006). Cities are actively developing strategies to attenuate their many environmental challenges (e.g. air pollution, noise, waste management; Price and Tsouros, 1996). The protection and enhancement of their green spaces and natural capital is a main area of environmental performance that is highly recognised within some leading initiatives, such as the European Green Capital Award (EGCA; Gudmundsson, 2015). However, and although urban ecology is a growing field in research, our knowledge of urban biodiversity remains limited in most cities, generally being descriptive and mainly focused on specific taxonomic groups during restricted periods of the year. Urban parks are hot–spots of biodiversity within the concrete matrix, and birds are among the most visible components (Caula et al., 2014). Avian communities are highly dynamic as a consequence of the environmental changes that occur between seasons across the globe, particularly in biogeographic areas subjected to strong seasonal regimes (Newton, 2008). Thus, avian communities in any geographic location can vary to some extent in species composition and structure throughout the year, with some species exclusively occurring during the breeding season, migration periods, and/or winter. Although less noticeable, the distributional patterns of some common species that are present year round in a region can also be dramatically altered between seasons in fragmented landscapes. This might be caused by seasonal changes in the characteristics of the fragments of habitat, but also because the abundance and/or behaviour (e.g. territoriality during the breeding season vs. gregariousness out of the breeding season) of these species differ between seasons. In any case, these seasonal changes might

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have consequences for avian diversity (Caula et al., 2014) and the species–park nestedness patterns (i.e. animal communities form 'nested subsets' if the most diverse fragments contain species that are not present in the least diverse ones; Patterson and Atmar, 1986; Blake, 1991) in urban environments that have been poorly explored (Murgui, 2010; Wang et al., 2013). Although birds are favourite study models in urban ecology (Marzluff, 2017), very few studies have analysed both the breeding and wintering avian communities in the same set of urban parks (Caula et al., 2014; Leveau and Leveau, 2016), let alone estimating and comparing their nestedness patterns between seasons (Murgui, 2010; Wang et al., 2013). This means that our knowledge of urban bird communities is biased towards one season (typically the breeding season). It also means that the factors that predict avian diversity patterns during breeding might not be valid during other periods of the year (e.g. winter), which would have important implications in the design of urban areas aiming to promote a rich avifauna all year–round (Nielsen et al., 2013). Likewise, it has been suggested that the nested patterns that avian breeding communities typically show during the breeding season could be eroded during winter (Murgui, 2010). These seasonal differences in nestedness could be caused by urban birds selecting alternative habitats and/or relaxing their ecological requirements during the non–breeding period (McClure et al., 2013). Identifying the existence of these patterns would be relevant for the long–term conservation of urban bird populations (Murgui, 2010). In this study, I surveyed the breeding and wintering avian communities of 31 urban parks in Vitoria–Gasteiz (EGCA 2012 holder) in order to expand our understanding of the factors explaining between–park and between–season variation in the avian communities of this city. The main aims were to: (1) identify the features of parks from among a set of potential candidates (e.g. size, vegetation characteristics, noise) that contributed the most to avian diversity, and to determine whether the contribution of these factors differed between the breeding and wintering period; (2) test the existence of nestedness patterns in breeding and wintering bird communities and the features of the parks that were correlated with their nestedness ranks (Patterson, 1987); and (3) explore how seasonality affected nestedness (i.e. whether nestedness is indeed disrupted in winter compared to the breeding season) and the park and species nestedness ranks. For this third purpose, I considered only the resident species (i.e. avian species occurring in both breeding and wintering periods). This approach allowed us to assess whether species and parks tended to maintain (i.e. ranks are correlated between seasons) or not to maintain (uncorrelated ranks) their position in the nestedness matrix between seasons. I considered this assessment would provide insight into how seasonality affects the avian communities of the urban parks in Vitoria–Gasteiz.

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Material and methods Study area Vitoria–Gasteiz (42º 50' 58.4'' N 2º 40' 14.8'' W; 525 m a.s.l.; c.a. 250,000 inhabitants) is located in Northern Iberia (Araba, Basque Country, Spain) on a sub–plateau dominated by a mosaic of agricultural fields and forest patches. The Portuguese oak Quercus faginea is the most representative tree species. This region lies in a transitional area between the Mediterranean and the Atlantic climate (continental supra–Mediterranean climatic territory; Font, 1983) that confers cool summers (maximum temperature of ± 24 ºC and ± 90 mm of rain between June and August), and cold, wet winters (minimum temperatures of around 1 ºC and ± 250 mm of rain between November and February; Ninyerola et al., 2005). The main body of the city (approximately 26 km2) is surrounded by an almost complete green belt composed of several streams with their associated riverine vegetation and a few large peri–urban parks. Inside this peripheral green belt, the city hosts around 100,000 trees distributed in innumerable tree–lined streets and urban green areas of different size. This green framework allows the city to surpass the minimum tree and green area per capita recommendations made by the Health World Organization, established in one tree every three inhabitants and, at least, 9 m2 of green public areas per citizen (Ayuntamiento de Vitoria–Gasteiz, 2010). Many of the urban parks of the city with an extension of over 0.5 ha were selected for this study, which rendered a sample size of 31 parks (fig. 1, table 1). Bird surveys and avian diversity estimates Each park was visited three times (three rounds) by the author throughout the study period. Completing a full 31–park round took four days. Two of the rounds were carried out in the traditional wintering season in the region (Gainzarain, 2006). The first winter round occurred on 10–13 I 14, and the second took place during the first fortnight of February (sampling days were 7, 12–14 II 14). The order of park visits was reversed in the February survey compared to the January survey to avoid unexpected time–of–the–day effects. The third visit to the parks took place in late spring, 5–8 VI 14, coinciding with the breeding season in this region (De la Hera et al., 2014). Implementing the same avian sampling method in parks with marked differences in size and shape is not a straightforward task (e.g. Jokimäki, 1999; Carbó–Ramírez and Zuria, 2011). In this study, a complete count method was selected over others (e.g. transects of fixed size or point counts) to reach all the potential microhabitats available within the parks, thereby maximizing the chances of detecting all the species occurring in each park, while obtaining estimates of the relative abundance of each species (Murgui, 2010). Thus, each park was surveyed by walking routes covering the whole park area. Winter surveys were carried out between 9:00 h and 14:00 h, while spring surveys took place between 6:30 h and

10:30 h, avoiding periods of rain and strong wind during the sampling. While detectability of birds during the breeding season is highest in the first hours of the day and surveys should be restricted to this period, avian detectability is more homogeneous during winter and surveys can be safely extended until early afternoon (Rollfinke and Yahner, 1990). Avian diversity was strongly intercorrelated between the two winter surveys (see Results), suggesting a relatively minor effect of sampling time on avian diversity estimates. It was also considered that the differences in detectability between species did not change significantly between the breeding and the winter period (e.g. large birds are more easy to detect than small ones; Johnston et al., 2014), a realistic assumption that would make the seasonal bias in detectability relatively homogenous between species (Anderson et al., 2015). The sampling time at each park varied between four and 70 minutes, with high and significant between–park repeatability in the sampling time invested (intraclass correlation coefficient (ri) for the logarithmically transformed sampling time: ri = 0.93, F30,62= 41.8, P < 0.001). Following this procedure, it was noticed that the sampling effort (time/hectare) was comparatively greater in small parks than in large parks (log[sampling time] = 0.59 + 1.01*log[park area]; r = 0.96, P < 0.001) but this potential bias made the results of this study conservative (see Discussion). During the surveys, all the visual and aural contacts with birds making an effective use of the park were annotated, trying to avoid double counting. Thus, birds flying over the tree canopy were excluded, which explains the complete absence of hirundinid and swift records in the database during the breeding period. Three anthropogenic avian species (house sparrow Passer domesticus, feral pigeon Columba livia and spotless starling Sturnus unicolor) and waterfowl were also ruled out from the final dataset given that their numbers strongly depend on stochastic supplementary feeding by the public and on the presence of permanent waterbodies to roost, respectively, which would have disrupted the avian diversity estimates. Avian diversity was estimated for each park and visit using the effective number of species (D), which is a more meaningful and correct index to express diversity. This metric is defined as the number of equally– common species needed to obtain a specific score of entropy or other traditional diversity surrogates (Jost, 2006). In the case of this study, the Shannon–Weiner entropy index was initially calculated and its values were transformed into D using the formula suggested by Jost (2006): D = exp(x), where x is the Shannon index value of each visit to a park. Given the relatively low number of bird records obtained in the smallest parks, their intrapolation/extrapolation curves could not be calculated reliably, making rarefaction methods impossible to apply under the sampling method used in this study (Chao and Jost, 2012). Urban park characteristics QGIS was used to digitalize the limits of the parks and to obtain their area and shape (i.e. park area divided

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Vitoria Gasteiz

500 m

Fig. 1. Spatial distribution of the 31 urban parks selected for studying their breeding and wintering avian communities. See table 1 for the park names. Fig. 1. Distribución espacial de los 31 parques urbanos seleccionados para estudiar las comunidades de aves que albergan en la época reproductiva y el período invernal. Véase la tabla 1 para consultar el nombre de los parques.

by its perimeter). Vitoria–Gasteiz council provided (fully transferred on the 16 I 15) the most updated digital information on urban vegetation. This enabled accurate estimates following the variables of the parks (table 1): grass cover, shrub cover, tree density, and tree diversity (measured as the effective number of tree species). Between 2 and 6 geographic locations within each park were selected randomly to visually estimate the tree height and trunk diameter at breast height of the 15 closest trees, since this information was not operational in the digital data provided by the city council. The calculation of the intraclass correlation coefficients for these two variables showed reasonably high and significant between–park repeatability (tree height: ri = 0.58, F30,41= 4.24, P < 0.001; trunk diameter: ri = 0.70, F30,41= 6.27, P < 0.001), suggesting that the mean values of these estimates of tree height and trunk diameter were representative of each park. From the noise maps of Vitoria–Gasteiz published in 2012 (Ayuntamiento de Vitoria–Gasteiz, 2012), mean environmental noise was also obtained for each park as an indicator of human–induced disturbance that might affect avian communities (González–Oreja et al., 2012). For this purpose, the grids of the full–day environmental noise map that intersected with each

park were extracted digitally, and the mean value of those grids was used as a surrogate of the human disturbance experienced by each park. Statistical analyses First, whether the occurrence of each avian species differed between seasons was tested for those species that were present in at least one park/visit in each season (table 2). For this purpose, generalized linear mixed models with binomial error distribution (0, absence; 1, presence, for each park and visit) were carried out including season (n = 62 for winter; n = 31 for breeding) as fixed effects and park as a random factor. Second, the way avian diversity varied between the two winter surveys and between seasons were explored using Pearson correlations and paired–t tests. To explore the existence of spatial autocorrelation in the data, a set of Mantel tests were performed. After confirming the lack of spatial autocorrelation (see Results), winter and breeding avian diversities were analysed, using multiple regression, in relation to three principal components (PCs) obtained from a principal component analysis (PCA) that included all the descriptive characteristics of the parks

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Table 1. Characteristics of the 31 urban parks studied in Vitoria–Gasteiz and their avian diversity estimates expressed as the effective number of species (D). See Material and methods section for more details on data collection: N, park number; P, park name; A, area (in ha); S, shape (in ha/km); Gc, grass cover (in %); Sc, shrub cover (in %); Td, tree density (no/ha); Mth, mean tree height (in m); Mttd, mean tree trunk diameter (in cm); D, tree diversity; dB, mean noise; MDw, mean winter avian diversity; Db, breeding avian diversity. Tabla 1. Características de los 31 parques urbanos estudiados en Vitoria–Gasteiz y estimaciones de la diversidad de aves que contienen, expresadas como el número efectivo de especies (D). Véase el apartado "Material and methods" para obtener información detallada sobre la recopilación de datos: N, número de parque; P, nombre del parque; A, área (en ha); S, forma (en ha/km); Gc, cubierta de plantas herbáceas (en %); Sc, cubierta de arbustos (en %); Td, densidad de árboles (no/ha); Mth, altura media del árbol (en m); Mttd, diámetro medio del tronco del árbol (en cm); D, diversidad de árboles; dB, ruido medio; MDw, diversidad aviar media en invierno; Db, diversidad aviar durante la reproducción.

N P

Urban park characteristics A

S Gc Sc Td Mth Mttd D

Avian diversity dB MDw

1 Adriano VI

1.21 2.43 33 10.4 31 8.1

28 2.4 63.7 3.47 1.75

2 Arana

3.09 4.08 77.6 7.3 108 15.4 42.7 13.4 65.2 4.34 8.09

3 Aranbizkarra

6.1 4.91 78.9 2.7 118 13.7 34.4 19.2 64.7 7.48 9.44

4 Ariznavarra

2.16 3.48 78 2.9 131 10.4 24.5 12.5 69.3 4.66 3.89

5 Arriaga

17.01 9.35 75 4.6 105 13.3 36.4 28 63.4 9.89 10.43

6 Astronomos

0.74 2.15 74.4 5.8 199 12.6 29.3 12.2 62.8 3.17 2.83

7 Campa Sansomendi 9.63 6.15 81.6 1.6 164 12.9 26.8 26.9 63.3 8.41 10.04 8

Castillo de Zaitegui 1.94

2.66 41.5 8.4

11.5

11.4

60.3 3.36

5.53

9 Conservatorio

3.23 3.63 43.7 7.1 104 15.3 38.7 14.2 59.7 6.32 7.1

10 Constitucion

1.15 2.57 38.8 2.6 125 15.1 39.3 17.8 68.5 2.89 3.79

11 Deba

0.95 2.03 52.9 21.9 40 12.4 29.5 4.8 69.7 1.44 1

12 Florida

2.71 3.67 33.5 2.9 146 15.6 45.7 9.8 64 5.23 7.96

13 Gazalbide

1.77 2.88 69.1 4.1 143 14.3 40 6.7 59.5 6.03 5.74

14 Gerardo Armesto

0.5 1.77 49.7 0.3 143 13.1 37.7 11.6 68.7 1

15 Gran Sol

0.94 1.83 44.7 3.1 225 6.7 12.3 13.2 55.4 1 2.87

16 Judimendi

2.33 3.1 60.3 1.2 131 13.9 27 10.3 58.4 3.71 5.67

17 Maria de Maeztu

3.02

18 Maurice Ravel

2.69 2.49 53 2.3 68 14.2 36.3 16.2 61 5.38 5.33

19 Molinuevo

4.83 3.85 51.5 2 144 14.6 38.7 31.9 64.1 5.8 7.3

20 Obispo Ballester

0.61 1.65 70.3 7.9 107 11.7 31.3 7.2 69.8 4.23 3

21 Parque del Este

2.76

3.46

22 Plaza de Llodio

2.44

2.57 61.4 0.5 102 13.2

23 Prado

3.36 4.08 76.9 0 118 15.5 44.3 13.5 64.6 3.44 6.83

2.44

1.3 153 13.9

5.1

7.3

12.4

64.4 6.46

6.24

11.7

56.8 1.91

33.7 14.4

60.4 3.85

5.66

24 Rosaleda Bolivia 1.98 2.12 54.9 7.1 134 13.9 40.7 8.3 55 5.32 4.34 25 Salvador Allende

1.02 1.54 29 13.2 72 11.5 35.3 10.4 65.4 2.88 3.37

26 San Martin

8.41

27 Sansomendi

2.69 3.71 72.3 3.4 130 12.3 31.7 16 60.5 7.87 7.4

28 Santa Barbara

0.74 2.14 39.5 8.1 82 9.7 31.3 5.8 62.9 3.29 4.75

29 Simon Bolivar

1.33 2.5 17.8 2.1 104 12.2 30 8.1 52.5 1 1.75

30 Zaldiaran

0.53 1.28 27.1 0 85 9.1 18.2 5.9 51.7 1

31 Zaramaga

0.98 1.81 38.2 5 141 12.4 28 4.9 63 2.84 3.36

60.6

134

9.9

27.8 15.3

56.9 8.07

5.34

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Table 2. Between–survey frequency of occurrence of the 21 avian species detected in the urban parks of Vitoria–Gasteiz and results of the generalized linear mixed models that tested the existence of differences between seasons in these frequencies for the 16 resident species, i.e. species occurring in at least one park/visit in both seasons: Jan, parks occupaied in January; Feb, parks occupied in February; Jun, parks occupied in June; V, variance ± SD; I, intercept (breeding); S, season (winter). (The statistical significance of the effects was simplified in the superscript: n.s. non–significant, * P < 0.05, ** P < 0.01, *** P < 0.001). Tabla 2. Frecuencia entre estudios de la presencia de las 21 especies de aves detectadas en los parques urbanos de Vitoria–Gasteiz y resultados de los modelos mixtos lineales generalizados con los que se analizó la existencia de diferencias en estas frecuencias entre estaciones para las 16 especies residentes, es decir, las especies observadas por lo menos en un parque o visita en ambas estaciones: Jan, parques ocupados en enero; Feb, parques ocupados en febrero; Jun, parques ocupados en junio; V, variancia ± DE; I, interceptar (críanza); S, estación (invierno). (La significación estadística de los efectos se simplificó en el superíndice: n.s. no significativo; * P < 0,05; ** P < 0,01; *** P < 0,001).

Jan Feb Jun

Random effects V

Fixed effects I

Avian species occurring in either winter or summer (linear mixed models) Long–tailed tit Aegithalos caudatus

1.78 ± 1.34 –2.20 ± 0.85** –0.91 ± 0.74n.s.

European goldfinch Carduelis carduelis

4 14 0.37 ± 0.61 –0.21 ± 0.40n.s. –2.37 ± 0.70***

European greenfinch Chloris chloris

4 12 1.80 ± 1.34 –0.63 ± 0.52n.s. –2.05 ± 0.71**

Short–toed treecreeper Certhia brachydactyla 10 10 8 152.4 ± 12.3 –9.24 ± 2.38*** 1.55 ± 1.18n.s. Blue tit Cyanistes caeruleus

17 16 15 5.19 ± 2.28 –0.11 ± 0.62n.s. 0.35 ± 0.55n.s.

European robin Erithacus rubecula

15 10 17 4.64 ± 2.16

Chaffinch Fringilla coelebs

15 8

Pied wagtail Motacilla alba

2 130.4 ± 11.4 –15.04 ± 4.6**

0.36 ± 0.64n.s. –1.06 ± 0.63n.s.

3.78 ± 1.95 –4.94 ± 1.62** 4.04 ± 1.46** 7.45 ± 3.12*

Great tit Parus major

22 19 8

2.59 ± 1.61

Black redstart Phoenicurus ochruros

0.00 ± 0.00 –2.23 ± 0.61*** –1.88 ± 1.18n.s.

Common magpie Pica pica

20 20 25 2.61 ± 1.62

Firecrest Regulus ignicapilla

13 14 15 3.19 ± 1.79 –0.11 ± 0.55n.s. –0.31 ± 0.54n.s.

European serin Serinus serinus

–1.58 ± 0.64* 2.59 ± 0.76*** 2.11 ± 0.74**

–1.18 ± .66n.s.

0 21 372.1 ± 19.2 8.36 ± 2.20*** –20.1 ± 5.51***

Eurasian collared dove Streptopelia decaocto 1

65.9 ± 8.12 –6.77 ± 2.34** –4.67 ± 2.25*

Eurasian blackcap Sylvia atricapilla

1.25 ± 1.12 –2.35 ± 0.90** –0.83 ± 0.78n.s.

Blackbird Turdus merula

22 21 23

16.9 ± 4.1

3.48 ± 2.66n.s. –0.59 ± 0.80n.s.

Avian species occurring in only one season Melodious warbler Hippolais polyglotta

0 0 2

–

Common wheatear Oenanthe oenanthe

0 0 1

–

Common chiffchaff Phylloscopus collybita

11 10 0

–

Eurasian wren Troglodytes troglodytes

0 0 1

–

Song thrush Turdus philomelos

2 3 0

–

mentioned above. Some of these variables had to be transformed to meet the normality requirements (table 3). Finally, it was examined whether the species–park matrices of Vitoria–Gasteiz showed a nested pattern

(Blake, 1991). For this purpose, the NODF index (Almeida–Neto et al., 2008) was calculated for each of the three presence/absence matrices independently (two for winter and one for breeding), and the associated P–values were obtained from binary null

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Table 3. Coefficients of correlation (factor loadings) between the three principal components (PC) derived from the principal component analysis and the urban park characteristics. Table also shows the eigenvalues, the percentage of variance explained by each component and the significance levels between each PC and the descriptive traits of the parks. (n.s. non–significant; * P < 0.05; ** P < 0.01; *** P < 0.001). Tabla 3. Coeficientes de correlación (cargas factoriales) entre los tres componentes principales (PC, en su sigla en inglés) derivados del análisis de componentes principales y las características de los parques urbanos. En la tabla también se indican las raíces latentes, el porcentaje de varianza explicado por cada componente y los niveles de significación entre cada componente principal y los rasgos descriptivos de los parques. (n.s. no significativo; * P < 0,05; ** P < 0,01; *** P < 0,001). Variables

PC1 PC2 PC3

Area (log–transformed)

0.44***

–0.20n.s. –0.37*

Shape (log–transformed)

0.45***

–0.15n.s. –0.38*

Grass cover

0.38***

–0.02n.s. –0.02n.s.

Shrub cover (arcsine–transformed)

–0.13

0.44*** –0.38*

n.s.

Tree density

0.20*

–0.16n.s. 0.66***

Tree height

0.34***

0.42***

0.29n.s.

Tree trunk diameter

0.27**

0.52***

0.16n.s.

Tree diversity (log–transformed)

0.44***

–0.20

0.09n.s.

n.s.

Noise 0.14n.s. 0.47*** –0.12n.s. Eigenvalue

3.44 2.00 1.20

Variance explained

0.38

models. Random matrices were generated, keeping the frequency of each species constant using the 'c0' method in order to control for the fact that some species are more common than others (Jonsson, 2001). The relationship between the nestedness ranks of the parks (i.e. the value of the richest park will be 31, while it will be one for the poorest park) and their PC scores were also assessed by Spearman correlation coefficients in order to identify candidate park characteristics that might promote the observed nested patterns. Three restricted presence/absence matrices of equal size (31 parks and 16 species) were additionally analysed. The number of species in these matrices was limited to the 16 (resident) species that were present in at least one park/visit in both summer and winter (table 2), as a way to assess whether park (range: 1–31) and species ranks (range: 1–16; value 16 for the most widely–distributed species and one for the least common) were correlated between seasons using comparable matrices. This approach helped to indirectly assess how seasonality affects the species–park nestedness matrices in Vitoria–Gasteiz. All analyses were performed with R version 3.4.3, using 'vegan' package for diversity calculations, Mantel tests and nestedness analyses (Oksanen et al., 2018). An α threshold of P = 0.05 was used in all statistical tests.

0.22

0.13

Results Avian composition and diversity between the breeding and winter period After excluding aquatic and anthropogenic avian species (see Material and methods), records of 1,514 birds from 21 species were gathered during the surveys (table 2, see also table 1s in Supplementary material). The common chiffchaff Phylloscopus collybita and the song thrush Turdus philomelos were only detected in winter in the parks of Vitoria–Gasteiz while the Eurasian wren Troglodytes troglodytes, common Northern wheatear Oenanthe oenanthe, and the melodious warbler Hippolais polyglotta appeared only during breeding (table 2) and in very low numbers. For species occurring in both seasons in at least one park/visit, the common chaffinch Fringilla coelebs, pied wagtail Motacilla alba and great tit Parus major were significantly more common during winter than during the breeding season, while the European goldfinch Carduelis carduelis, the European greenfinch Chloris chloris, the European serin Serinus serinus and the Eurasian collared dove Streptopelia decaocto were more widely distributed during the breeding season than during winter (table 2). The remaining species (n = 9) showed no significant differences between seasons.

Animal Biodiversity and Conservation 42.2 (2019)

287

Effective number of species (D)

A 10

Breeding Wintering

2 –2

0 Park size (PC1)

Residuals of D on PC1

2 1 0 –1 –2 –3

Breeding Wintering –2

–1 0 Tree diversity (PC3)

Fig. 2. Relationship between winter (in black) and breeding (in grey) avian diversity and (A) park size – PC1– (partial r2 was 0.76 and 0.61 for breeding and winter diversity) and (B) tree density –PC3– (partial r2 was 0.08 and 0.02, respectively). Fig. 2. Relación entre la diversidad de aves en el período invernal (en negro) y la época reproductiva (en gris), por un lado, y (A) la superficie del parque (CP1) (el coeficiente r2 parcial fue de 0,76 y 0,61 para la diversidad en la época reproductiva y en el período invernal) y (B) la densidad de los árboles (CP3) (el coeficiente r2 parcial fue de 0,08 y 0,02, respectivamente).

The diversity of the parks was estimated as the effective number of species and varied between one and 11 species (table 1). Winter diversity estimated in January was strongly correlated with the values obtained in February (Pearson r = 0.72, P < 0.001), with no significant differences between them (paired t–test: t = 1.48, P = 0.149), so the average values of these two surveys were used as an estimate

of winter diversity at each park (table 1). Parks with higher diversity in winter also had a higher effective number of species during the breeding period (Pearson r = 0.84, P < 0.001), although, on average, avian diversity during the breeding period (D = 4.99 ± 2.68 SD) was significantly higher than during the winter (D = 4.38 ± 2.39 SD; paired t–test: t = 2.34, P = 0.026).

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Avian diversity and spatial autocorrelation The Mantel tests did not show a significant association between the spatial distribution of the parks and their winter (Mantel statistic r = 0.10, P = 0.074) or breeding avian diversity (Mantel statistic r = 0.01, P = 0.405). Neither did the size of the studied parks show a spatial autocorrelation in Vitoria–Gasteiz (Mantel statistic r = 0.02, P = 0.382). For this reason, I did not apply any correction for spatial autocorrelation in the following analyses, and each park was treated as an independent data point. Relationships between avian diversity and urban park characteristics The PCA on the nine abovementioned characteristics of the urban parks provided three principal components (PCs) with eigenvalues higher than one (table 3). PC1 was interpreted as an index of park size, where high factor scores represented parks with a larger area, higher area–perimeter ratio, and more tree diversity than parks with low factor scores of PC1. Tree height, tree trunk diameter, shrub cover and noise were strongly and positively correlated with PC2 values, so that this component was considered a surrogate of the degree of vegetation development. Finally, tree density was the only variable strongly associated with PC3 (table 3). PC1 was the main component explaining the observed variation in both winter (partial r2 = 0.61) and breeding avian diversity (partial r2= 0.76; fig. 2A, see table 4 for the overall r2 of the models) with no between–season differences in the slope of this relationship (season × PC1 interaction: t = 1.84, P = 0.077). PC3 also had a minor significant contribution to winter avian diversity (partial r2 = 0.08), where more densely wooded parks had lower diversity indexes (fig. 2B). This effect was not significant for breeding avian diversity (partial r2 = 0.02), although the diversity–PC3 slopes did not differ significantly between seasons (effects of season × PC3 interaction: t = 1.13, P = 0.268). On the other hand, PC2 was not significantly associated with avian diversity in either the breeding or winter period (partial r2 ≈ 0, in both cases; table 4). Nestedness patterns and urban park characteristics Avian communities showed a significant nested subset pattern in the urban parks of Vitoria–Gasteiz under the null model selected, and the nestedness values were similar during the winter (January survey: NODF = 55.6, P < 0.001, NODF simulated values = 32.3–37.1 [95 % CI]; February survey: NODF = 57.0, P < 0.001, NODF simulated values = 34.9–40.7 [95 % CI]) and the breeding period (NODF = 61.9, P < 0.001, NODF simulated values = 35.7–40.8 [9 5 % CI]). For the three full presence/ absence matrices, park nestedness ranks were positively associated with PC1 (richer parks had higher PC1 scores; January survey: Spearman r = 0.67; February survey: r = 0.68; breeding survey: Spearman

r = 0.82; all P < 0.001), but in no case park ranks were correlated with PC2. Finally, PC3 only showed a significant negative association with the ranks for the February survey (Spearman r = –0.40, P = 0.025), but not for January (Spearman r = –0.18, P = 0.337) or June (Spearman r = –0.19, P = 0.308). The presence–absence matrices that only considered the 16 species occurring in both seasons also showed a significant nested pattern (January survey: NODF = 56.7, P < 0.001, NODF simulated values = 33.1–37.8 [95 % CI]; February survey: NODF = 54.5, P < 0.001, NODF simulated values = 31.3–36.7 [95 % CI]; June survey: NODF = 63.9, P < 0.001, NODF simulated values = 41.4–46.7 [9 % CI]). Species ranks of the two restricted winter matrices were correlated with each other (Spearman [January–February] r = 0.89, P < 0.001), but this association was not significant between seasons (Spearman [January– June] r = 0.32, P = 0.222; Spearman [February–June] r = 0.48, P = 0.062). In spite of this lack of association in the species ranks between seasons, the park ranks of the three matrices were significantly, and positively, inter–correlated (Spearman [January–February] r = 0.82, P < 0.001; Spearman [January–June] r = 0.79, P < 0.001; Spearman [February–June] r = 0.76, P < 0.001) and their associations with the PCs of the parks were qualitatively the same as those observed for the full matrices. Discussion In accordance with expectations in a city in a temperate zone with a marked seasonal regime, the distribution patterns of some species changed dramatically between the breeding and wintering seasons, the two most stable seasons of the year for avian assemblages. For example, seven out of the 16 species that occurred all–year round (resident species) showed significant differences in their park occupancy rates between the breeding and the wintering period: four of them were more common during the breeding season and three were more common during the winter (table 2). This led to a lack of association in the resident species nestedness ranks between seasons, which confirmed that the structure and composition of the avian community differed markedly in summer compared to winter. These dynamics were probably mainly determined by the seasonal changes in abundance that the migratory dynamics of each species promote in this region (Gainzarain, 2006; De la Hera et al., 2014), although other factors, such as changes in behaviour (gregariousness versus territoriality) or habitat selection, might also contribute, but to a lesser extent (Murgui, 2010). In an area where winter conditions are harsh and spring–summers are relatively mild (Ninyerola et al., 2005), avian communities are also expected to be more diverse during the more favourable season (Newton, 2008). Avian diversity was slightly higher during the breeding season than during the wintering period in the urban parks of Vitoria–Gasteiz. However, these differences were relatively small. This could be

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Table 4. Results of the multiple regression analyses exploring the contribution of park characteristics (PC1–3) to the variation in winter and breeding avian diversity, expressed as the effective number of species (D). Tabla 4. Resultados de los análisis de regresión múltiple para estudiar la contribución de las características de los parques (CP1–3) a la variación de la diversidad de aves entre el período invernal y la época reproductiva, expresada como el número efectivo de especies (D).

Winter avian diversity (R2 = 0.69) Estimate ± SE

t–value

Breeding avian diversity (R2 = 0.78)

P Estimate ± SE

t–value

Intercept

4.38 ± 0.25

17.38

< 0.001

4.99 ± 0.24

20.92

< 0.001

PC1

1.00 ± 0.14

7.25

< 0.001

1.26 ± 0.13

9.61

< 0.001

PC2

–0.05 ± 0.18

–0.25

0.805

–0.01 ± 0.17

–0.04

0.970

PC3

–0.63 ± 0.23

–2.69

0.012

–0.35 ± 0.22

–1.60

0.121

caused by two main reasons. Firstly, from the pool of approximately 18 migratory species that occur in the region exclusively for breeding but spend the winter in tropical Africa (Martí and Del Moral, 2003; Unanue–Goikoetxea, 2017), only two species were observed in spring within the city boundaries (i.e. melodious Warbler and Northern wheatear), and their occurrence was very low, if not anecdotal (table 2). These two species are insectivore specialists that, like other representatives of this group that are lacking from the urban matrix but are common outside of it (e.g. Iberian chiffchaff Phylloscopus ibericus), probably have serious difficulties to find suitable conditions and enough food resources (invertebrates) for breeding in urban green areas as a consequence of causes such as pollution, greater presence of exotic plants, and intensive management of urban vegetation (e.g. regular lawn mowing, tree/bush pruning; Jones and Leather, 2012; Leveau and Leveau, 2016). Secondly, diversity was balanced between seasons by the more frequent winter occurrence in the parks of several species that are present year–round at a regional scale, such as the chaffinch, the pied wagtail, the song thrush and the common chiffchaff. The population sizes of these four species are boosted in winter by the arrival of many migratory conspecifics from higher latitudes (Asensio, 1985; Pérez–Tris and Asensio, 1997; SEO/Birdlife, 2012). Additionally, anthropogenic food resources and higher temperatures have also been suggested as factors that would allow these species to use the urban parks for overwintering (Gainzarain, 2006). Ultimately, what makes breeding communities slightly more diverse is the usual presence of some finch species (i.e. European greenfinch, European goldfinch, and European serin) during the breeding period that are rare during winter. This is probably a consequence of the tendency of many finches to group together in alternative habitats of the periphery of the city during the non–breeding period, becoming less frequent in the urban matrix (Gainzarain, 2006). Murgui (2010) found this same pattern in the city of Va-

lencia (southeastern Spain), although the harsh winter conditions in Vitoria–Gasteiz suggest that the regional abundance of finches will be much lower than in Southern Iberia during winter (Gainzarain, 2006; SEO/ Birdlife, 2012). Interestingly, the consequences for the nestedness patterns of the seasonal rearrangements of finch populations were completely different between these two cities (Murgui, 2010). Thus, while winter avian communities did not show a nested pattern in the urban parks of Valencia, nestedness was significant and similar in both seasons in Vitoria–Gasteiz. In any case, the seasonal variation in abundance of many breeding and wintering birds in the urban parks of Vitoria–Gasteiz (see above) suggests that, during some parts of their annual cycle, many of them rely on habitats located outside the city. Expanding our knowledge of the spatiotemporal distribution of these urban birds is essential to understand their population dynamics and implement management practices that could favour their occurrence and abundance within the city boundaries. In spite of the structural and composition differences between the breeding and wintering avian communities, avian diversity and nestedness ranks were consistent between seasons, with the urban parks that were the most diverse during breeding also being the most diverse during winter. This supports the idea that the reduction in the park distribution of some species from one season to the other is compensated by the expansion of other species. Although previous research has explored the potential determinants of the richness of avian breeding species in the urban parks of Vitoria–Gasteiz (De la Hera et al., 2009), this is the first study associating avian diversity with park features using more suitable measures of diversity and orthogonal explanatory variables by means of PCA for both breeding and wintering assemblages, and the first to describe park–species nested patterns. The results of this study highlight park size (PC1) as the indisputably best predictor explaining between–park variation in avian diversity (r2 > 0.61) and the nested-

290

ness patterns in both seasons. Tree density (PC3) also contributed to avian diversity, although to a lesser extent: parks with less density of trees exhibited higher diversity values. As observed in Valencia, the nested patterns observed in the urban parks of Vitoria–Gasteiz might be mainly determined by selective extinction (Patterson and Atmar, 1986) and/or colonization (Cook and Quinn, 1995). This would be supported by the observed pervasive influence of park size and the fact that the parks studied have similar designs and management practices (Murgui, 2010). Thus, larger parks will be able to host most of the generalist urban species and a few specialists that are probably not able to persist in smaller parks in the long–term. It is important to note that the fact that small parks were relatively oversampled per unit area compared to large parks (see Material and methods) makes the patterns observed in this study more robust, because, in spite of these differences, large parks were still richer and more diverse than small parks. This study also showed that nestedness and avian diversity patterns were not greatly affected by seasonality in the urban parks of Vitoria–Gasteiz. This is probably a common pattern in relatively small cities, where the size of the parks rarely exceeds the threshold in which the species–area relationship maintains its positive slope, which has been established at approximately 10 hectares for urban parks (Nielsen et al., 2013). Promoting habitat heterogeneity within and between urban parks (e.g. implementing different management practices of the vegetation) could increase the number of urban species in the city and expand the distribution of some rare ones, but the benefits of these measures would be cushioned in small parks by edge effects and other factors (Fernández–Juricic, 2001). Thus, these results suggest that park size is the main constraint for avian diversity in small and medium–sized cities, so that favouring the existence of a few relatively large parks (i.e. over 10 ha) instead of many small ones would be a much more effective measure to maintain a diverse urban avifauna all year round. Acknowledgements I am very grateful to Iranzu Sanz de Galdeano, María Báez, Fernando de Juana, Andrés Alonso and Luis Lobo (Ayuntamiento de Vitoria–Gasteiz) for their help in compiling the parks data, Mónica Tomás (AACACUSTICA) for providing the noise maps, María Torres–Sánchez for meaningful discussions, and Luis M. Carrascal and three anonymous reviewers for providing constructive comments on an early version of the manuscript. References Almeida–Neto, M., Guimarães, P., Guimarães, P. R., Loyola, R. D., Ulrich, W., 2008. A consistent metric for nestedness analysis in ecological systems: reconciling concept and measurement. Oikos,

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Survival and longevity in neotropical damselflies (Odonata, Polythoridae) A. Cordero–Rivera, I. Sanmartín–Villar, M. Sánchez Herrera, A. Rivas–Torres, A . C. Encalada

Cordero–Rivera, A., Sanmartín–Villar, I., Sánchez Herrera, M., Rivas–Torres, A., Encalada, A. C., 2019. Survival and longevity in neotropical damselflies (Odonata, Polythoridae). Animal Biodiversity and Conservation, 42.2: 293–300, Doi: https://doi.org/10.32800/abc.2019.42.0293 Abstract Survival and longevity in neotropical damselflies (Odonata, Polythoridae). Longevity among insect orders varies greatly, and has mainly been studied in insects in temperate biomes, where seasonality determines high synchronization of reproductive activities and limits lifespan. Most forest damselflies in tropical regions have low population densities and are almost never observed in copula. We hypothesized that selection will favour a high survival rate and hence high lifespan, allowing the animals to be ready for the occasional events that favour reproduction. We studied two neotropical damselflies, Polythore mutata and P. derivata, in Ecuador, using mark–recapture methods. We found that sex affected the rate of recapture, but daily survival rate was affected by sex only in one population. We found evidence that suggests stabilizing or directional selection on body size. The maximum lifespan was 54–63 days. We conclude that the survival rate of Polythore damselflies in tropical forests is comparable to that of similar damselflies in temperate zones. Key words: Lifespan, Rainforest, Low density, Body size, Mark–recapture Resumen Supervivencia y longevidad de las libélulas del neotrópico (Odonata, Polythoridae). La longevidad, que entre los órdenes de insectos es muy variable, se ha estudiado principalmente en insectos de biomas templados, donde la estacionalidad determina una alta sincronización de las actividades reproductivas y limita la longevidad. La mayoría de las libélulas de las regiones tropicales vive en poblaciones con una densidad baja y casi nunca se observan en cópula. Nuestra hipótesis es que la selección favorecerá una alta tasa de supervivencia y, por lo tanto, una gran esperanza de vida, lo que permitiría que los animales estuvieran listos para los eventos ocasionales que favorecen la reproducción. Estudiamos dos libélulas neotropicales, Polythore mutata y P. derivata, en Ecuador, utilizando métodos de marcaje y recaptura. Constatamos que el sexo afectó a la tasa de recaptura, pero que solo afectó a la tasa de supervivencia diaria en una población. Hallamos indicios que sugieren la existencia de selección estabilizadora o direccional del tamaño del cuerpo. La longevidad máxima observada fue de 54–63 días. Concluimos que la tasa de supervivencia de las libélulas del género Polythore en los bosques tropicales es comparable a la de libélulas similares de las zonas templadas. Palabras clave: Longevidad, Selva tropical, Baja densidad, Tamaño corporal, Marcaje y recaptura Received: 30 X 18; Conditional acceptance: 11 I 19; Final acceptance: 12 III 19 Adolfo Cordero–Rivera, Iago Sanmartín–Villar, Anais Rivas–Torres, ECOEVO Lab, Escola de Enxeñaría Forestal, Universidade de Vigo, Campus A Xunqueira, 36005 Pontevedra, Galiza, Spain.– Melissa Sánchez Herrera, Biology Program, Faculty of Natural Sciences and Mathematics, Universidad del Rosario, Bogotá, Colombia.– Andrea C. Encalada, Laboratorio de Ecología Acuática, Instituto BIOSFERA, Universidad San Francisco de Quito, Diego de Robles y Vía Interoceánica, Campus Cumbayá, 17–12–841, Quito, Ecuador. Corresponding author: A. Cordero–Rivera. E–mail: adolfo.cordero@uvigo.es ISSN: 1578–665 X eISSN: 2014–928 X

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Introduction Over the last decades, ecological research on natural populations of insects has accumulated a large dataset about survival and reproductive behaviour, allowing a good understanding of their reproductive strategies (Cornell and Hawkins, 1995). Longevity and fecundity among insect orders vary greatly, and recent evidence indicates that some insects live long enough to experience reproductive senescence (Sherratt et al., 2010). However while strong statistical evidence has been presented for senescence in vertebrate populations in the wild, we know little about the rate and shape of senescence in wild populations of insects. Most of this information comes from studies of insects in temperate biomes (Cordero–Rivera and Stoks, 2008), where seasonal regulation (Tauber and Tauber, 1976) and reproductive synchronization (Cordero Rivera and Andrés Abad, 1999) are clearly adaptive. Nevertheless, in tropical regions, where temperature remains almost constant, there are no cues to reproductive synchronization, except those derived from rain regimes (Wikelski et al., 2000). Odonates (damselflies and dragonflies) have been widely studied using mark–recapture methods in temperate regions (Cordero-Rivera and Stoks, 2008). For these reasons, studies of longevity of tropical damselflies are clearly needed (Cordero–Rivera and Stoks, 2008). In tropical forests, some dragonflies are key species linking stream trophic networks to aerial networks, and are present at very low densities, particularly in South America. For instance, individuals of the family Calopterygidae can be found in their thousands in European streams but only a few scattered specimens can be found in the Amazon (for instance members of the genera Hetaerina, Mnesarete or Ormenophlebia) (Córdoba–Aguilar and Cordero–Rivera, 2005). In these tropical areas, reproductive activity is rarely observed. We hypothesized that animals must survive long periods to be able to reproduce when conditions are appropriate (e.g. Sanmartín–Villar and Cordero–Rivera, 2016). Due to its particular individual and population traits, the family Polythoridae (Zygoptera) is a relevant model in ecology and evolution (Sánchez Herrera et al., 2015). A recent molecular phylogeny of this family suggests the existence of two clades, one grouping the species distributed across the Amazon basin and the other grouping species in the Andes (Sanchez Herrera et al., 2018). Most species of this genus possess sexual dimorphism, with males and females displaying different colour patterns. For example, the Amazonian species Polythore mutata (McLachlan, 1881) shows female colour polymorphism in addition to sexual dimorphism. The males and one phenotype of females (androchrome, similar colour to males) display a bright milky white band on the wings while the other phenotype of females (gynochrome, different colour to males) shows a bright orange band in the forewing and a violet band on the hindwing (Sanmartín–Villar and Cordero–Rivera, 2016). Beccaloni (1997) described several mimicry rings for the Ithomiinae butterflies in Jatun Sacha

in Ecuador. This author suggested that P. mutata phenotypes converge in two of the seven mimicry rings he described for this geographic location. His work was mainly descriptive, but he also looked at the UV reflections which seemed to be consistent with the butterflies as well. Other researchers have anecdotally reported the similarities of the flight of Polythore damselflies with butterflies, and recent studies of the flight behaviour in the species Polythore procera, Euthore fasciata and Gretta andromica suggest that there is a mimicry signal among them (Outomuro et al., 2016a) wing shape, and flight style. The study species have wings with a subapical white patch, considered the aposematic signal, and a more apical black patch. The main predators are VS–birds, visually more sensitive to violet than to ultraviolet wavelengths. Using multiple mark and recapture methods, we estimated recapture rates, survival probability and life expectancy of two geographically distant populations of P. mutata in Ecuador. Our hypothesis was that these animals have high survival (see above). We aimed to answer the following questions: 1) is the survival rate different between males and females?; 2) are both sexes recaptured with similar probabilities?; 3) for P. mutata females, is the survival probability of colour morphs the same?; and 4) does body length affect survival? Given the presence of a small population of Polythore derivata in one of the field sites, we also give a first analysis of survival in this species. Material and methods Field data collection The field data collection and observations were conducted at two localities in Ecuador where P. mutata was previously recorded. The first location was the Tiputini Biological Station (TBS), at the border of Yasuni National Park (76.146041ºW, 0.635000ºS, Orellana province). We visited this location three times; from 4 to 12 December 2012, from 4 to 7 February 2013, and from 10 to 23 June 2013. However, at the third date, the density of P. mutata was extremely low and no marked individuals were resighted, which suggests that the time span of the study was appropriate for the expected lifespan of the species. A few individuals of Polythore derivata were observed during these sampling periods, but too few to be studied. The second field site was the Jatun Sacha Biological Station (JSBS), near Tena (77.615677 ºW, 1.067593 ºS, Napo province), the same place visited by Beccaloni (1997). Here the density of P. mutata was higher than in Tiputini, and was accompanied by some Polythore derivata and isolated specimens of Polythore concinna. Field work was done during the dry season, between 30 October and 16 December 2014 (see Sanmartín–Villar and Cordero–Rivera, 2016, for further details). In both localities, one to two observers walked over small streams and forest paths where the damselflies were found, for an average of 7 h per day. The animals

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Fig. 1. A P. mutata male (A) from TBS and a P. derivata male (marked) (B) from JSBS. The insert in (A) shows the tip of the abdomen of a male P. mutata with algae growing, which is suggestive of old age. Pictures by ACR. Fig. 1. Un macho de P. mutata (A) de la estación biológica de Tiputini (TBS) y uno marcado de P. derivata (B) de la estación biológica de Jatun Sacha (JSBS). El círculo en (A) muestra el extremo del abdomen de un macho de P. mutata con algas, lo que sugiere que es de edad avanzada. Fotografias de ACR.

were captured with an entomological net, measured with a calliper (body length to the nearest 0.1 mm, including anal appendages), marked with a white number in the wing (xylene–free white permanent marker; Pilot Super Color EF: www.pilotpen.com; fig. 1), photographed, and released. Re–sightings were carried out with the naked eye or using a Minox monocular, so that animals were only netted again if needed for specific reasons. Mark and recapture analyses To analyse mark–recapture histories we used the software Mark 8.1 (White and Burnham, 1999). Our analyses included sex and time (and their interaction) as factors to test for their effects on survival and recapture probabilities. Each population was analysed separately, because migration is extremely unlikely due to the high geographic distance between the

populations (the linear distance is 167.2 km, estimated from GoogleEarth). We used the Information Theory Approach to rank models by their Akaike’s Information Criterium (AIC), which is minimised in the models that are better supported by the data (Burnham and Anderson, 1998). However, this method does not allow to know whether the candidate models are good enough to explain the variability of the data. Therefore, we first tested the fit of the full time–dependent Cormack–Jolly–Seber model by groups using program Release. The model is defined by: Phi(g*t) p(g*t) where Phi and p represent the recapture and survival probability, g is sex and t is time. In the case of TBS, this model showed good adjustment for males of P. mutata (Goodness of fit results (Test 2 + Test 3): x218 = 15.71, p = 0.613), but due to the low recapture rate of females, this

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1,0 0,9

Survival probability

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

0,0 36 37 38 39 40 41 42 43 44 45 46 47 48 49 50 Body length (mm)

Fig. 2. The relationship between body length and survival probability of P. mutata at Tiputini (sexes combined). Note that the estimates suggest stabilizing selection. Grey lines indicate 95 % confidence intervals. Fig. 2. Relación entre la longitud corporal y la tasa de supervivencia de P. mutata en Tiputini (sexos combinados). Nótese que las estimaciones sugieren la existencia de una selección estabilizadora. Las líneas grises indican el intervalo de confianza del 95 %.

test could not be calculated. The adjustment was clear in the case of P. mutata from JSBS (Test 2 + 3 by groups: males x252 = 22.73, p = 1; females x227 = 8.21, p = 1). For P. derivata at JSBS, only males were recaptured, and again the saturated model showed good adjustment (Test 2 + 3: males x214 = 9.10, p = 0.825). In JSBS we ran an analysis of P. mutata females only to test for differences between morphs. The saturated model was also appropriate (Test 2 + 3 by morph: androchrome x27 = 0.00, p = 1; gynochrome x27 = 9.45, p = 0.977). The second step was to estimate the variance inflation parameter (c–hat) by two methods. We divided the c–hat obtained from model Phi(g*t) p(g*t) by the mean c–hat of the bootstrap simulations in Mark. With this method we always obtained c–hat = 1. The second method divided the deviance of the saturated model by the mean deviance of the bootstrap procedure. Values ranged between 0.98 and 1.27. No correction was done when c–hat was lower than one, but otherwise c–hat estimates were used to adjust parameter estimates and standard errors. In the next step, we ran models without individual covariates, to select the most supported model, as the one that minimizes Quasi Akaike's Information Criterion (QAIC). Once these models were identified, body length was included as an individual covariate, both as linear and quadratic terms. When there was statistical support for more than one model, estimates

of parameters were obtained by model averaging using Mark software and are shown as average probability ± SE. We estimated the expected longevity using the formula of Cook et al. (1967): Longevity = 1 / loge(survival) Results Overall, population density of both Polythore species was low. At TBS, we marked 21 males and 13 females of P. mutata in December 2012 and 16 males and 5 females in February 2013, and recaptured four males and one female from 2012. No marked animals were found in June 2013. At JSBS we marked 76 males and 35 females of P. mutata, 19 males and five females of P. derivata and three males and one female of P. concinna. This last species is not analysed here due to the lack of recaptures. Female polymorphism was detected in both populations of P. mutata. In TBS we only observed one androchrome female in the first period (that could not be marked) and marked another androchrome female in the second sampling period, with an overall frequency of 5.6 % androchromes (1 out of 18 females), while at JSBS, androchromes were 40 % (14 out of 35). No female polymorphism was detected in P. derivata (see also Sanmartín–Villar and Cordero–Rivera, 2016).

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1,0 0,9

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

41 42 43 44 45 Body length (mm)

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

0,0

Fig. 3. The relationship between body length and survival probability of P. mutata at Jatun Sacha. There was positive directional selection in both sexes: A, males; B, females. (Grey lines indicate 95% confidence intervals). Fig. 3. Relación entre la longitud corporal y la probabilidad de supervivencia de P. mutata en Jatun Sacha. Hubo selección direccional positiva en ambos sexos: A, machos; B, hembras. (Las líneas grises indican el intervalo de confianza del 95 %).

Considering both sexes, at TBS model selection by AICc indicates that the best model is {Phi(.) p(g)}, which means similar survival for males and females and a different recapture rate by sex. The inclusion of body size as a covariate (table 1s) indicates that survival is affected by body size, and that the best model includes the quadratic

term, suggesting stabilizing selection (fig. 2). Given that the best model has a D AICc = 3.8 (table 1s), no model averaging was needed, and we estimated daily survival rate from the best model as 0.974 ± 0.011 (expected longevity 38 days) and recapture rates 0.391 ± 0.047 for males and 0.081 ± 0.041 for females.

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At JSBS, the model selected by AICc is {Phi(g) p(.)}, the opposite to the previous case (table 2s), indicating higher survival rates for females, while the recapture rate was similar for males and females. Including body length as a linear covariate improved the fit of this model, but the quadratic term diminished fit. Therefore, our data for JSBS suggest a positive directional selection for larger body length (fig. 3) in both sexes. Using model averaging, the daily survival rate was estimated as 0.821 ± 0.032 for males (expected longevity 5.1 days) and 0.953 ± 0.019 for females (20.8 days), with a common recapture rate of 0.200 ± 0.024. The analysis of recapture histories of the JSBS polymorphic females of P. mutata was limited by the low number of specimens found. On some dates, only one morph was found, and therefore interactions morph*time could not be calculated at these dates, and models including these interactions were excluded. The best model, as selected by AICc, was {Phi(.) p(g)}, and the second {Phi(g) p(g)}, and both were much more supported than the next models (DAICc > 4.3; table 2s). Using model averaging, we found the daily survival rate was estimated as 0.927 ± 0.034 (expected longevity of 13.2 days) for androchromes and 0.943 ± 0.021 (longevity 17.0 days) for gynochromes, and recapture rates 0.088 ± 0.038 for androchromes and 0.247 ± 0.044 for gynochromes. Recapture rates were clearly different between morphs, and there is some evidence for higher survival by gynochromes. Finally, we analysed the recapture histories of males of P. derivata from JSBS (no females were recaptured). In this case, given the small sample size (N = 16 males) we could not use covariates. The best model was the simplest one, with constant survival and recapture probabilities ({Phi(.) p(.)} indicating a survival rate of 0.962 ± 0.018 and a recapture rate of 0.241 ± 0.048, with an expected longevity of 25.6 days. Some observations might suggest higher lifespan than our estimates. For instance, some individuals showed algae and even bryophytes over the body surface (fig. 1A) at TBS. In addition, five individuals marked in December at TBS, were found alive in February, with a lifespan between 54 and 63 days. The proportion of animals with algae on the abdomen was 15.6 % in December and 38.5 % in February, when most of the specimens were clearly older. Three of the animals surviving from December to February showed algae on the abdomen, but none of these had algae present when first marked. Discussion We found that Polythore mutata and P. derivata damselflies have a high survival rate and low recapture probabilities, particularly in females. Mark–recapture studies provide information based on the subset of animals that are recaptured, and therefore the estimates are valid only if temporal emigration does not occur, and when emigration is permanent, this cannot be disentangled from mortality (Lebreton et al., 1992). Examination of trade–offs between reproduction and

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survival rely on individually marked animals, for which the exact time of death is most often unknown, because marked individuals cannot be followed closely through time. These limitations are to be taken into account when interpreting our results. For other Polythoridae, like Polythore procera, estimates of survival are ~0.97 (Sánchez–Herrera and Realpe, 2010), and ~0.40 to 1, depending on the population for P. gigantea (Altamiranda–Saavedra and Ortega, 2012), which are similar to our estimates. The analysis of recapture rates suggests that females were less often resighted at TBS, whereas at JSBS, about 20% of individuals were resighted, irrespective of sex. In contrast, recapture probabilities for the Andean species previously studied yielded higher estimates (Sánchez–Herrera and Realpe, 2010; Altamiranda– Saavedra and Ortega, 2012) than for our Amazonian species. Andean species of Polythore showed higher density of individuals overall than those in the Amazonian field sites we evaluated here. The preferred habitat of Polythore (i.e. rocky waterfalls, small creeks) is not common in the Amazon in comparison with the Andean foothills. In particular, the difference between the TBS and JSBS may be explained by the fact that observations at TBS were mainly done near the stream, where males showed high site fidelity and territorial behaviour, whereas at JSBS the topography allowed a more extensive examination of the areas around the streams and in the forest. In fact, some females were found feeding at the same sunspots in the forest for several days. Therefore, our data suggest that the lower recapture rate of female damselflies (Cordero– Rivera and Stoks, 2008) is due to females remaining at larger distances from the water. This behaviour was also observed for the Andean P. procera, where the females seem to remain at highly dense forest areas, while males remained near the open stream area (Sánchez–Herrera et al., 2010). At TBS, we found that survival probabilities were not affected by sex, while at JSBS they were higher for females, and were in the interval 0.821–0.974. These survival values translate into longevities of five to 38 days, similar to the maximum longevity observed in other damselflies from temperate regions (Cordero–Rivera and Stoks, 2008). Therefore, we did not find evidence of long pre–reproductive periods or high survival, contrary to our expectations. However, the presence of algae and liverworts growing on the abdomen and wings of some specimens (fig. 1A) suggests that they have lived in the humid forest for months (Lücking et al., 2010) because these algae cannot be acquired during the larval stage (the cuticle changes after metamorphosis) and because liverworts have a long life cycle (Lücking et al., 2010). Our estimates of maximum lifespan are thus likely below the real values. To test this hypothesis, future studies should include longer periods of fieldwork to maximize the probability of detecting a particularly high lifespan. Records of algae growing on rainforest odonates are very rare and have been interpreted as indicators of old age (e.g. Fincke and Hadrys, 2001). Algae were also found in other Ecuatorian species such as Metaleptobasis sp, Argia oculata, Heteragrion cooki,

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Hetaerina fuscoguttata, and Uracis imbuta (ACR, ISV and ART personal observations, 2014, 2016), and some odonates in rainforests in Yunnan (China, ACR, pers. obs. 2016). However, even the longest–living specimens in our samples were unlikely to live for more than two months. Higher longevity could also be the result of the low activity levels of Polythore individuals, which remained perched immobile almost all the time (see Sanmartín–Villar and Cordero–Rivera, 2016), but nevertheless showed nervous behaviour (they quickly escaped when a human approached) which is probably an efficient antipredator strategy. Finally, their possible mimicry to poisonous butterflies could also contribute to high survival (Beccaloni, 1997). This topic is worth exploring in future field experiments. Female colour polymorphism for P. mutata was also evidenced in the TBS population, which constitutes the first evidence of this polymorphism out of JSBS. However, we only saw two andromorphs throughout the whole sampling dates for TBS. Therefore, even though there are polymorphic individuals in different populations, the selective forces maintaining the polymorphism might differ depending on the geographic location. Beccaloni's (1997) suggestion of convergence of these morphs within two Ithomiinae mimicry rings for JSBS could play a role as a selective force potentially maintaining higher survival probabilities for P. mutata. Our results show little evidence for an effect of female colour morph on survival rate in females of P. mutata. These results are to be taking with caution because the number of females resighted was low, but they agree with previous work on other polymorphic damselflies (e.g. Andrés and Cordero–Rivera, 2001) that suggest no effect of female colour morphs on survival. Few studies however have applied the modern techniques of capture–recapture analysis (Lebreton et al., 1992) to polymorphic damselflies, and further research is clearly needed on this topic. The conspicuousness of the wings in male and androchrome females of P. mutata might be higher than that in gynochrome females, at least for predators that perceive the UV range (Bick and Bick, 1965). In addition, no androchrome females were observed near the streams where males concentrated, even in JSBS, where androchromes were common. This suggests the two female phenotypes have different reproductive strategies, a common phenomenon in other damselflies (Van Gossum et al., 2008). Future experiments are needed to decipher the significance of this polymorphism. We found that adding body length as a covariate in the mark–recapture analysis increases the fit of the models. At TBS, we found evidence that suggests stabilizing selection for body size, whereas at JSBS larger body length was apparently under positive selection. In fact, the average size of males at JSBS (42.9 ± 0.27, SE, N = 72) was smaller than the size of males at TBS (43.9 ± 0.45, N = 34; t104 = –2.109, p = 0.037). A recent review of body size evolution in the Odonata found a consistent trend of positive selection for larger size, but at the same time no evidence for an increase in body size over evolutionary time. The authors suggested that this lack of concordance is due to the trade–off between larger body size as

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adults and longer larval life (Waller and Svensson, 2017). Survivorship seems more related to body size for JSBS males than for females (compare the slopes in fig. 3). This could be explained by the territorial behaviour and the positive relationship between body size and reproductive success in males and the lack of relationship for this trait and fitness in many females (Serrano–Meneses et al., 2007; Sanmartín–Villar and Cordero–Rivera, 2016). In the case of TBS our results suggest stabilizing selection (fig. 2). It is possible that small males were excluded from the best areas by the aggressive behaviour of larger males. However, survivorship also decreased in the largest body sizes. This might be due to the associated cost of large bodies (e.g. maintenance, decrease of manoeuvrability, conspicuousness towards predators) or/and the different directions of natural and sexual selection in those traits (Outomuro et al., 2016b). Our results indicate that survival can or cannot be positively related to larger body size in these damselflies, possibly depending on local ecological conditions. A previous study of another forest damselfly found that in a sunny stream, body size was negatively correlated with survival, but in a shaded environment it was not (Rivas–Torres et al., 2017). Another possibility is that survival may show interannual variation, a suggestion that also merits further study. It seems likely that contrasting selective regimes between natural and sexual selection in adults, between larvae and adults (Waller and Svensson, 2017), and between years and localities contributes to the stasis observed on body size over time. Acknowledgements Funding was provided by a grant from the Spanish Ministry of Economy and Competitiveness, including FEDER funds (CGL2014–53140–P). ISV and ART were supported by FPI grants (BES–2012–052005 and BES–2015–071965). References Altamiranda–Saavedra, M., Ortega O., 2012. Estructura poblacional de Polythore gigantea (Odonata: Polythoridae) en sistemas lóticos con diferentes estados de conservación. Revista de Biología Tropical, 60: 1205–1216. Andrés, J. A., Cordero–Rivera, A., 2001. Survival rates in a natural population of the damselfly Ceriagrion tenellum: Effects of sex and female phenotype. Ecological Entomology, 26: 341–346. Beccaloni, G. W., 1997. Ecology, natural history and behaviour of ithomiine butterflies and their mimics in Ecuador (Lepidoptera: Nymphalidae: Ithomiinae). Tropical Lepidoptera, 8: 103–124. Bick, G. H., Bick, J. C., 1965. Color variation and significance of color in reproduction in the damselfly Argia apicalis (Say) (Zygoptera: Coenagriidae). Canadian Entomologist, 97: 32–41. Burnham, K. P., Anderson, D. R., 1998. Model selec-

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tion and inference. A practical information–theoretic approach. Springer, New York. Cook, L. M., Brower, P. P., Crozer, H. J., 1967. The accuracy of a population estimation from multiple recapture data. Journal of Animal Ecology, 36: 57–60. Cordero–Rivera, A., Stoks, R., 2008. Mark–recapture studies and demography. In: Dragonflies and damselflies. Model Organisms for Ecological and Evolutionary Research: 7–20 (A. Córdoba–Aguilar, Ed.). Oxford University Press, Oxford, Doi: 10.1093/acprof:oso/9780199230693.003.0002 Cordero Rivera, A., Andrés Abad, J. A., 1999. Lifetime mating success, survivorship and synchronized reproduction in the damselfly Ischnura pumilio (Odonata: Coenagrionidae). International Journal of Odonatology, 2:105–114. Córdoba–Aguilar, A., Cordero–Rivera, A., 2005. Evolution and ecology of Calopterygidae (Zygoptera: Odonata): status of knowledge and research perspectives. Neotropical Entomology, 34: 861–879, Doi: 10.1590/S1519–566X2005000600001 Cornell, H. V., Hawkins, B. A., 1995. Survival patterns and mortality sources of herbivorous insects: some demographic trends. The American Naturalist, 145: 563–593. Fincke, O. M., Hadrys, H., 2001. Unpredictable offspring survivorship in the damselfly, Megaloprepus coerulatus, shapes parental behavior, constrains sexual selection, and challenges traditional fitness estimates. Evolution, 55: 762–772. Van Gossum, H., Sherratt, T. N., Cordero–Rivera, A., 2008. The evolution of sex–limited colour polymorphisms. In: Dragonflies and damselflies. Model organisms for ecological and evolutionary research: 219–229 (A. Córdoba–Aguilar, Ed.). Oxford University Press, Oxford. Lebreton, J. D., Burnham, K. P., Clobert, J., Anderson, D. R., 1992. Modeling survival and testing biological hypotheses using marked animals: a unified approach with case studies. Ecological Monographs, 62: 67–118. Lücking, R., Mata–Lorenzen, J., Dauphin, L., 2010. Epizoic liverworts, lichens and fungi growing on Costa Rican Shield Mantis (Mantodea: Choeradodis). Studies on Neotropical Fauna and Environment, 45: 175–186. Outomuro, D., Ángel–Giraldo, P., Corral–Lopez, A., Realpe, E., 2016a. Multitrait aposematic signal in Batesian mimicry. Evolution, 70: 1596–1608, Doi: 10.1111/evo.12963 Outomuro, D., Söderquist, L., Nilsson–Örtman, V., Cortázar–Chinarro, M., Lundgren, C., Johansson, F., 2016b. Antagonistic natural and sexual selection on wing shape in a scrambling damselfly. Evolution, 70: 1582–1595, Doi: 10.1111/evo.12951 Rivas–Torres, A., Sanmartín–Villar, I., Gabela–Flo-

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res, M., Cordero–Rivera, A., 2017. Demographics and behaviour of Heteragrion cooki, a forest damselfly endemic to Ecuador (Odonata). International Journal of Odonatology, 20: 123–135, Doi: 10.1080/13887890.2017.1336495 Sánchez–Herrera, M., Realpe, E., 2010. Population structure of Polythore procera at a Colombian stream (Odonata: Polythoridae). International Journal of Odonatology, 13: 27–37. Sánchez–Herrera, M., Realpe, E., Salazar, C., 2010. A neotropical polymorphic damselfly shows poor congruence between genetic and traditional morphological characters in Odonata. Molecular Phylogenetics and Evolution, 57: 912–917. Sanchez Herrera, M., Beatty, C., Nunes, R., Realpe,E., Salazar, C., Ware, J. L., 2018. A molecular systematic analysis of the Neotropical banner winged damselflies (Polythoridae: Odonata). Systematic Entomology, 43: 56–67, Doi: 10.1111/syen.12249 Sánchez Herrera, M., Kuhn, W. R., Lorenzo–Carballa, M. O., Harding, K. M., Ankrom, N., Sherratt, T. N., Hoffmann, J., Van Gossum, H., Ware, J. L., Cordero–Rivera, A., Beatty, C. D., 2015. Mixed signals? Morphological and molecular evidence suggest a color polymorphism in some Neotropical Polythore damselflies. Plos One, 10: e0125074, Doi: 10.1371/journal.pone.0125074 Sanmartín–Villar, I., Cordero–Rivera, A., 2016. Female colour polymorphism and unique reproductive behaviour in Polythore damselflies (Zygoptera: Polythoridae). Neotropical Entomology, 45: 658–664, Doi: 10.1007/s13744–016–0417–7 Serrano–Meneses, M. A., Córdoba–Aguilar, A., Méndez, V., Layen, S. J., Székely, T., 2007. Sexual size dimorphism in the American rubyspot: male body size predicts male competition and mating success. Animal Behaviour, 73: 987–997. Sherratt, T. N., Laird, R. A., Hassall, C., Lowe, C. D., Harvey, I. F., Watts, P. C., Cordero–Rivera, A., Thompson, D. J., 2010. Empirical evidence of senescence in adult damselflies (Odonata: Zygoptera). Journal of Animal Ecology, 79: 1034–1044. Tauber M. J., Tauber, C. A., 1976. Insect seasonality: diapause maintenance, termination, and postdiapause development. Annu. Rev. Entomol., 21: 81–107. Waller, J. T., Svensson, E. I., 2017. Body size evolution in an old insect order: no evidence for Cope’s Rule in spite of fitness benefits of large size. Evolution, 71: 2178–2193, Doi: 10.1111/evo.13302 White, G. C., Burnham, K. P., 1999. Program MARK: Survival estimation from populations of marked animals. Bird Study, 46 Supplem.: 120–138. Wikelski, M., Hau, M., Wingfield, J. C., 2000. Seasonality of reproduction in a neotropical rain forest bird. Ecology, 81: 2458–2472.

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Hidden diversity under morphology–based identifications of widespread invasive species: the case of the 'well–known' hydromedusa Craspedacusta sowerbii Lankester 1880 J. A. Oualid, B. Iazza, N. M. Tamsouri, F. El Aamri, A. Moukrim, P. J. López–González Oualid, J. A., Iazza, B., Tamsouri, N. M., El Aamri, F., Moukrim, A., López–González, P. J., 2019. Hidden diversity under morphology–based identifications of widespread invasive species: the case of the 'well–known' hydromedusa Craspedacusta sowerbii Lankester 1880. Animal Biodiversity and Conservation, 42.2: 301–316, Doi: https://doi.org/10.32800/abc.2019.42.0301 Abstract Hidden diversity under morphology–based identifications of widespread invasive species: the case of the 'well– known' hydromedusa Craspedacusta sowerbii Lankester 1880. A relatively scarce number of morphological features available for delimiting closely related species and an increasingly worrisome scenario on Global Climate Change causing the rapid dispersion of invasive alien species can lead to the rapid spread of reports of a given species around the world. Craspedacusta sowerbii Lankester, 1880 is considered the most widespread freshwater jellyfish species and has been reported in numerous locations on all continents except Antarctica. Recently, a few medusae attributed to C. sowerbii were collected from a water reservoir (Bin El Ouidan) in Morocco, this being the first confirmed record of the species from North Africa. The morphology of these newly collected specimens agrees well with previous descriptions, but mitochondrial (Cox1 and 16S) and nuclear ITS (ITS1–5,8S–ITS2) molecular data lead to a discussion of a more complex general view concerning the number of species, synonyms and nomenclatural problems hidden behind the reports of Craspedacusta sowerbii. Key words: Craspedacusta, Cryptic species, Medusa, Invasive species, NIS, Cox1, 16S, ITS, Morocco Resumen La diversidad oculta en las identificaciones basadas en la morfología de especies invasoras de amplia distribución: el caso de la "bien conocida" hidromedusa Craspedacusta sowerbii Lankester 1880. El número relativamente escaso de características morfológicas utilizadas para delimitar especies estrechamente relacionadas y el panorama cada vez más preocupante en el que el cambio climático global provoca la rápida dispersión de especies exóticas invasoras pueden conducir a la difusión precipitada por todo el mundo de informes sobre una especie determinada. Craspedacusta sowerbii Lankester, 1880, que se considera la especie de medusa de agua dulce más extendida, ha sido observada en numerosos lugares en todos los continentes, excepto en la Antártida. Recientemente, se recogieron algunas medusas atribuidas a C. sowerbii en un embalse artificial (Bin El Ouidan) en Marruecos, que representaron el primer registro confirmado de la especie en el norte de África. La morfología de estos especímenes recién recolectados concuerda con las descripciones anteriores, pero los datos moleculares mitocondriales (Cox1 y 16S) y nucleares ITS (ITS1–5,8S–ITS2) suscitan un debate general más complejo con respecto al número de especies, las sinonimias y los problemas nomenclaturales ocultos tras los informes de Craspedacusta sowerbii. Palabras clave: Craspedacusta, Especies crípticas, Medusa, Especies invasoras, NIS, Cox1, 16S, ITS, Marruecos Received: 07 IX 18; Conditional acceptance: 16 X 18; Final acceptance: 21 III 19 Jaouad Abou Oualid, Abdellatif Moukrim, Faculté des Sciences, Université Ibnou Zohr, Agadir, Maroc.– Badiaa Iazza, Pablo J. López–González, Biodiversidad y Ecología Acuática (BECA), Departamento de Zoología, Facultad de Biología, Universidad de Sevilla, Sevilla, Spain.– Naoufal M. Tamsouri, Fatima El Aamri, Institut National de Recherche Halieutique (INRH), Maroc. Corresponding author: Jaouad Abou Oualid. E–mail: j.abououalid@gmail.com ISSN: 1578–665 X eISSN: 2014–928 X

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Introduction Among the recommendations suggested by a wide panel of specialists to ensure progress in the management of aquatic NIS (non–indigenous species), the first was the availability of taxonomic expertise (see Ojaveer et al., 2014). Taxonomists produce the basic knowledge for understanding biodiversity (Agnarsson and Kuntner, 2007; Linse, 2017). The second was the need to use molecular tools, as classical taxonomy often requires additional sources of information for species description and identification (Goldstein and DeSalle, 2011), leading to an integrative taxonomy (DeSalle et al., 2005; Rubinoff et al., 2006a, 2006b; Pires and Marinoni, 2010; Chen et al., 2011). Despite this, reliable identification of organisms to species level is one of the greatest constraints. The lack of specialists and the inaccuracy of species identifications often result in an erroneous interpretation of the actual biodiversity and inadequate conservation policies, from local to global levels. This is the so–called 'taxonomic impediment' (Hoagland, 1996; Giangrande, 2003; Dar et al., 2012). Avoiding this problem by limiting OTUs data matrices to higher taxonomic levels or considering functional biodiversity (Cernansky, 2017) is not a viable solution, especially when working on NIS, whose influence at different levels on native ecosystems is well documented (e.g. Bax et al., 2003; Wallentinus and Nyberg, 2007; Walther et al., 2009; Poulin et al., 2011; González–Duarte et al., 2016; among others). The freshwater genus Craspedacusta includes a still uncertain number of hydromedusan species (Bouillon et al., 2006; Fritz et al., 2009; Jankowski et al., 2008). The most reported species around the world in this genus is Craspedacusta sowerbii Lankester, 1880, which can be found in all continents and subcontinents except Antarctica (Dumont, 1994). This species is native to the Yangtze River basin in China (Kramp, 1961) and is considered a cosmopolitan invasive species. C. sowerbii colonizes all types of freshwater habitats, i.e. streams, freshwater lakes, ponds, reservoirs and rivers (Raposeiro et al., 2011; Karaouzas et al., 2015). C. sowerbii was first described from specimens found in a water–lily tank in Regent's Park, London, England in 1880 (Lankester, 1880a). Later, the species was reported from many different localities: United States (Garman, 1916), Hawaii, South Australia (Thomas, 1950), New Zealand, the Philippines, China, Japan (Acker, 1976), France, Sweden, Portugal (Ferreira, 1985), Canada (McAlpine et al., 2002), Spain (Pérez–Bote et al., 2006; Medina–Gavilán and González–Duarte, 2018), Mexico (Moreno–Leon and Ortega–Rubio, 2009), Brazil (Silva and Roche, 2007), Uruguay (Mañé–Garzón and Carbonell, 1971), India (Riyas and Kumar, 2017), Italy (Schifani et al., 2018), Chile (Fraire–Pacheco et al., 2017; Fuentes et al., 2019), Turkey (Balik et al., 2001; Bekleyen et al., 2011), Israel (Gasith et al., 2011) and Greece (Karaouzas et al., 2015). From the African continent, it has been recorded with certainty only from South Africa (Rayner, 1988; Rayner and Appleton, 1989, 1992). The recent record of a Craspedacusta species in Lake Manzala (Delta

Nile), reported by Gasith et al. (2011: 147 and SM1), is based on a series of doubtful identifications (initially ascribed to the genus Limnocnida) and comments on a brief mass occurrence of medusae (see Elster et al., 1960; Elster and Vollenweider, 1961; Dumont and Verheye, 1984: 315; Dumont, 1994, 2009: 496 for additional information on that bloom event). The life cycle of Craspedacusta sowerbii includes both polyp (assuming asexual reproduction) and free–swimming stages (involved in sexual reproduction) (Bekleyen et al., 2011; Gasith et al., 2011). The appearance of the active medusa stage is related to an increase in water temperature (Bekleyen et al., 2011). Occurrences of this pelagic stage are sporadic, lasting only a few weeks, usually in the late summer and autumn (Minchin et al., 2016). The polyp stage is often overlooked because of its small size, having a wide capacity to tolerate different temperature and light conditions (see Payne, 1924; Boulenger and Flower, 1928; Acker, 1976; Acker and Muscat, 1976). The polyp and medusa stages are rarely reported together (see Failla–Siquier et al., 2017). Duggan and Eastwood (2012) established a protocol to find polyp stages that would be usable even in water reservoirs where the medusa stage had not been previously observed. These authors reached the conclusion that C. sowerbii is more common and widespread than is apparent from observations of medusae. Estimating the timing of introduction of this species in a given region is therefore difficult if it is only carried out after jellyfish findings have been recorded. A few individuals of a hydromedusa species were recently detected in a Moroccan reservoir. These specimens were initially identified (based on morphological characters) as the well–known alien widespread species Craspedacusta sowerbii. This record is the first confirmed finding of this species in North Africa. However, a molecular study of this material and its comparison to previously available information revealed a more complex scenario, with nomenclatural and biogeographic implications. As this hydromedusa species is often reported in lists of alien species, the correct specific identification of the different Craspedacusta lineages becomes an urgent challenge to correctly understand how many invasion events and species could be involved. The present paper aims to stress the risks linked to the current trend of exponentially increasing numbers of morphology–based reports of invasive species. An integrative view, including both morphology and molecular information, should be applied as a rule for checking the current identity of these 'well known' species as there are several examples of cryptic species that are difficult or impossible to delimit due to overlapping morphological characters. Material and methods Sample collection Bin El Ouidan reservoir is located in Azilal province (coordinates: 6º 27' 50'' W; 32º 6' 24'' N), at 810 m a.s.l.

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0º E er r

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Spain

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co oc or

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0º N N S1 S2

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E S 30º S 1,500 km

Fig. 1. Localisation of Bin El Ouidan reservoir in Morocco: S1, site 1; S2, site 2; open circle, Rayner and Appleton (1992); striped circle, Rayner (1988) and Rayner and Appleton (1989); solid circle, present record. Fig. 1. Localización del embalse Bin El Ouidan en Marruecos: S1, punto 1; S2, punto 2; círculo blanco, Rayner and Appleton (1992); círculo rayado, Rayner (1988) y Rayner and Appleton (1989); círculo negro, registro presente.

(fig. 1). It covers 3,740 ha and has a maximum depth of nearly 100 m. The total volume reaches 1,384 million m3. The reservoir was built between 1949 and 1953. In December 2015 we surveyed two sites (fig. 1), and found medusa stages attributable to the genus Craspedacusta at site S1 (6º 26' 13'' W; 32º 05' 33'' N). In total, four specimens were collected by scuba divers from the water–column between 1 and 5 m of depth. The water temperature, as indicated by the dive computer, was 20 ºC and the visibility was 3 m. Two specimens were fixed in absolute ethanol for the molecular study, while the other two were fixed in formalin 4 % for morphological observations. Nomenclatural remarks Despite the precise nomenclatural comments by Fritz et al. (2007: 54) about the discovery, first descriptions of this jellyfish species (see also Allman, 1880; Lankester, 1880a, 1880b), and ICZN decision (see also Allen, 1910; Stiles, 1910), several subsequent authors still reported the species with the specific epithet 'sowerbyi'. In the original description, Lankester (1880a: 148) used the spelling Craspedacusta sowerbii in honour of Mr. Sowerby, understanding the genitive singular of the complete latinization of Sowerby to Sowerbius. The use of the form 'sowerbyi' must be considered an

erroneous spelling (Zarazaga, pers. comm.). In this case, Article 33.3.1 of the ICZN (1999) about the predominant use of erroneous spellings cannot be applied (indeed, according to Fritz et al. (2009) it is about 40 % of all references). Thus, in order to avoid the use of 'sowerbyi', all references to the species of Lankester in this paper will be made with the original spelling. Morphological observations and measurements Observation and photography of different parts of the medusae were performed with a camera (ToupCam™) attached to light microscopy (Olympus CX41). A Panasonic Lumix FZ28 camera was used for macroscopic photography. Measurements of bell and gametogenic tissues and tentacle length were performed using ImageJ 1.46r software (NIH, Bethesda, MD, USA). Phylogenetic analysis Total genomic DNA was extracted from two EtOH– preserved specimens using the E.Z.N.A. DNA kit (OmegaBiotech) following the manufacturer's instructions. The Cox1 and 16S mitochondrial regions as well as the nuclear ITS region (ITS1–5,8S–ITS2) were sequenced as proposed by Fritz el al. (2009) and Karaouzas et al. (2015) for comparative purposes.

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The Cox1 region was amplified using the primers dgLCO1490 5'–GGT CAA CAA ATC ATA AAG AYA TYG G–3' and dgHCO2198 5'–TAA ACT TCA GGG TGA CCA AAR AAY CA–3' (Meyer et al., 2005), the 16S region was amplified using the primers 16S. Cunningham.F.1mod 5'–ACG GAA TGA ACT CAA ATC ATG TAA G–3' and 16S. Cunningham.R.2 5'–TCG ACT GTT TAC CAA AAA CAT A–3' (Bridge et al., 1995). An 800 bp partial sequence of the nuclear DNA was amplified using the forward primer 5'–CCCTTTGTACACACCGCCCGTCGCT–3' and the reverse primer 5'–CTTTGGGCTGCAGTCCCAAGCAACCCGACTC–3' (Odorico and Miller, 1997). This last partial sequence included parts of the 18S rDNA and 28S rDNA, the complete ITS1 and ITS2 regions as well as the 5.8S rDNA region (Odorico and Miller, 1997; Fritz et al., 2009). Each PCR used 1 U of MyTaq Red DNA Polymerase (Bioline), 10 µM of each primer, approximately 30 ng of genomic DNA, and was brought to a final volume of 25 µL with H2O. Cox1 PCR was carried out using the following cycle profile: initial denaturation at 95 ºC for 1 min, 40 cycles of denaturation at 95 ºC for 15 s, annealing at 42 ºC for 15 s, and extension at 72 ºC for 10 s, and a final extension at 72 ºC for 5 min. The 16S and ITS PCRs used the same cycle profile, but 58 ºC and 60 ºC as annealing temperatures respectively. PCR products were purified using ExoSAP–IT™ PCR Product Cleanup Reagent (ThermoFisher Scientific) following the manufacturer’s instructions. Purified products were electrophoresed on an ABI PRISM® 3730xl Genetic Analyzer and sequence traces were edited using Sequencher™ v4.0. The obtained sequences were compared with homologous sequences from EMBL–Bank of other Olinididae species. Available sequences of campanulariid Obelia species were used as out–group in the implemented phylogenetic analyses. The alignments of the different sets of sequences were carried out using MUSCLE, as implemented in MEGA6 (Tamura et al., 2013). After alignment, the best nucleotide substitution model was selected using Modeltest as implemented in MEGA6, according to the Akaike Information Criterion (AIC) and hierarchical likelihood ratio test (hLRT). The 16S dataset (40 olindiid + 1 campanulariid sequences) had 554 positions, with a total of 276 variable and 232 parsimony–informative sites. The Cox1 dataset (64 olindiid + 1 campanulariid sequences) had 680 positions, with a total of 270 variable and 232 parsimony–informative sites. The ITS dataset sequences (55 Craspedacusta + 1 campanulariid sequences) had 933 positions, with a total of 378 variable and 209 parsimony–informative sites. The phylogenetic reconstructions were obtained applying Maximum Likelihood (ML) and Bayesian inference methods. ML method was carried out in MEGA6, and based on the T92 + G model (16S), GTR + I (Cox1), and K2 + G model (ITS) (Kimura, 1980; Tamura, 1992; Nei and Kumar, 2000) using the NNI heuristic method (Nearest Neighbor Interchange) and 1000 bootstraps replications (Felsenstein 1985). The Bayesian Inference was carried out in MrBayes v3.1.2 (Huelsenbeck and Ronquist, 2001; Ronquist and Huelsenbeck, 2003), using the model GTR + G

(lset nst = 6 rates = gamma), 107 generations and discarding 25 % initial trees. The material studied here has been deposited in the Museu de Ciènces Naturals in Barcelona (MZB), the collection of the first author (JAO–UIZ) at the University Ibn Zohr of Agadir, Morocco, and the collection of the research team Biodiversidad y Ecología Acuática in the University of Seville, Spain (BECA). Results Systematics Phylum Cnidaria Class Hydrozoa Subclass Trachylinae Order Limnomedusae Family Olindiidae Haeckel, 1879 Genus Craspedacusta Lankester 1880 Craspedacusta sowerbii Lankester 1880 See Lewis et al. (2012) and Jankowski (2001) for a complete list of synonyms. Material examined MZB 2018–0758 one specimen formalin fixed. MZB 2018–0757 one specimen fixed in absolute ethanol. JAO–UIZ(H1) one specimen formalin fixed. BECA(H1) one specimen fixed in absolute ethanol. Moreover, total DNA extraction from the specimen in MZB 2018–0757 [in BECA as BECA(H2)] and from BECA(H1) are kept within the molecular DNA collection of BECA. All specimens with the same sampling data as above described in the section Material and methods. Morphological remarks The medusa was the only stage recovered (fig. 2A). No polyps were found, and all specimens are female. The average bell/umbrella diameter is 20 ± 1 mm (19–21), flattened form. The mouth has four slightly folded lips overpassing the umbrella margin. Four gametogenic tissues pouch–like structures (7–11 mm length) are hanging from the radial canals. They are opaque in the basal fold–like part and translucent and voluminous in the apical part, giving a triangular shape (fig. 2B). The tentacles have no organs of adhesion and are connected to the marginal end of the umbrella on the ring canal (fig. 2E). Four long perradial tentacles (7–9 mm in length) emerging from the end of the four radial canals at the umbrella margin. About 60 medium tentacles (2.0–4.5 mm in length) arising from the pole of the bell were counted between the four long tentacles. Approximately 420 shorter tentacles (0.5–1.5 mm in length) extend around the bell edge. The three different sizes of tentacles are organized in a regular distribution along the umbrella edge. Along the tentacles, nematocysts are grouped in patches that

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A m

5 mm

B v

rc 5 mm D

5 mm

0.2 mm F

0.2 mm G

0.2 mm

Fig. 2. Craspedacusta sowerbii Lankester, 1880: A, subumberllar surface. B, m, manubrium; C, gn, gamets; rc, radial canal; t, tentacles; v, velum. D, dispersion of tentacles. E, arrangement of nematocysts in per–radial tentacle. F, arrangement of nematocysts in the middle part of tentacle. G, arrangement of nematocysts in the apical part of tentacle. H, apical part of tentacle. . Fig. 2. Craspedacusta sowerbii Lankester, 1880: A, superficie subumberlar. B, m, manubrio. C, gn, gametos; rc, canal radial; t, tentáculos; v, velo. D, disposición de los tentáculos. E, ordenación de nematocistos en tentáculos perradiales. F, ordenación de nematocistos en parte media de un tentáculo. G, ordenación de nematocistos en el extremo apical del tentáculo. H, extremo apical del tentáculo.

are arranged in spaced parallel rings (fig. 2E, 2B), distances between consecutive rings are distinctly reduced distally (fig. 2G). Phylogenetic analyses Cox1 analyses (fig. 3) placed the sequences obtained in this work for the two Moroccan specimens in a well–supported clade (Bootstarp [Bts.] 99, posterior probability [PP.] 0.99) with the German sequences and a Chinese (Sichuan province) sequence constituted, with an internal p–distances (German–Moroccan to Chinese sequence) of 0.3 %. This last German–Moroccan–Chinese clade is the sister group of a relatively

poorly supported clade with two well defined groups, a Switzerland sequence and a well–supported clade (Bts. 100, PP. 1) including a conglomerate of Chilean–Italian–Indian–Grecian–Chinese sequences (all of which were also attributed to C. sowerbii), average uncorrected p–distance between these last two clades 17.5 %. Average uncorrected p–distance between German–Moroccan–Chinese clade and Switzerland sequence 13.6 %. Average uncorrected p–distance between German–Moroccan–Chinese clade and Chilean–Italian–Indian–Grecian–Chinese clade was 16.3 %. Phylogenetic hypotheses based in Cox1 suggest that there are at least three Craspedacusta species in Europe.

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The 16S analyses (fig. 4) indicate that the analyzed Moroccan individuals merge well among other Craspedacusta sequences. There is very little previous 16S information on Craspedacusta sowerbii, just a sequence from Lake Huato (USA), another from Uruguay, two sequences from Switzerland, and a sequence of unknown locality (KY077294). Moroccan sequences form a well–supported clade (Bts. 99, PP. 1) with the sequence KY077294 and the two Switzerland sequences. Uncorrected p–distance between Morocco–unknown locality and Switzerland sequences is 0.2 %. This last European clade (having in mind the unknown origin of one of the sequences) is the sister group of the clade formed by both American sequences (USA + Uruguay), which are identical. All these mentioned sequences, all identified as C. sowerbii, are related to another two Craspedacusta species (C. sinensis and C. ziguiensis) in a relatively well–supported clade (Bts. 82, PP. 0.95). The clade grouping all Craspedacusta species is the sister group of Limnocnida tanganjicae, with high support (Bst. 90, PP. 0.99). Mean uncorrected p–distance between American populations attributed to C. sowerbii (USA and Uruguay) and the Moroccan–Switzerland–unknown–origin sequences is 4.3 % ± 0.1 (range 4.2–4.5 %). The genetic distance between the other two Craspedacusta species (C. ziguiensis and C. sinensis) is 6.2 %, and the distance between the Moroccan specimens and the latter species is 6.8 % and 8.3 %, respectively. In general, all Olindiidae genera are well supported in this 16S phylogenetic hypothesis. Inter–genera genetic distances (uncorrected p–distances) in 16S seem to be around 13–30 %. According to 16S' knowledge, a single species occurs in Morocco and Switzerland that is different from that present in America (USA + Uruguay). The ITS phylogenetic analyses (fig. 5) benefits from a higher number of sequences; thus, the analyses is here focused on the genus Craspedacusta instead of the whole available olindiid taxa. Craspedacusta sequences are mainly obtained from central Europe and China, although recent sequences from Chile and Sicily (Italy) have been published. Three main clades can be detected, all of them including Chinese specimens. Clade I (Bts. 99, PP. 1), includes Chinese sequences attributed to C. kiatingi and C. sichuanensis, as well as all German sequences attributed to C. sowerbii and the sequences obtained in this study from Moroccan specimens. The sister group of Clade I is composed of a single sequence of C. ziguiensis from China [support between both sister groups (Bts. 99, PP. 1)]. On the other hand, Clade II (Bts. 99, PP. 1), includes Chinese sequences attributed to C. sinensis and C. brevinema, while Clade III (Bts. 90, PP. 0.56) includes Chinese sequences attributed to. C. sowerbii and C. xinyangensis, as well as sequences from Italy and Chile. Sequences from Clade I (where the Moroccan specimens are included) have genetic distances (uncorrected p–distances) between 0.0 and 0.9 % (mean and SD 0.1 % ± 0.2), while this Clade I is 3.8 % ± 0.1 (range 3.6–4.2 %) distant from its sister group (C. ziguiensis). Uncorrected p–distances between Clade I and Clade II are 19.9 %± 0.4 (range 18.9–20.6 %),

while distances between Clade I and Clade III are 11.3 % ± 0.31 (range 10.0–12.4 %), finally, distances between Clade II and Clade III are 17.7 % ± 0.37 (range 15.8–18.5 %). According to our current ITS knowledge, four main lineages (species) can be detected, two of them (Clades I and III) including specimens identified as C. sowerbii. At least two species are present in Europe, while the known American sequences (Chile) and those from Central Europe and North Africa are definitively different lineages (Clades I and III, respectively). Unfortunately, there is no homogenous knowledge of the three genetic markers here examined along the entire distributional area where specimens attributed to C. sowerbii have been reported. Figure 6 shows the worldwide distribution of the main clades detected in the separate analyses of the three markers (see also fig. 3, 4, and 5 for comparison). Discussion Morphological remarks Caraspedacusta sowerbii has been recorded in several localities around the world. However, many identifications are not fully reliable since the records do not give detailed morphological characters (Moreno–Leon and Ortega–Rubio, 2009; Jakovčev– Todorović et al., 2010; Stefani et al., 2010; Gasith et al., 2011; Souza and Ladeira, 2011; Galarce et al., 2013; Gomes–Pereira and Dionísio, 2013; Fraire–Pacheco et al., 2017). Moreover, many hydromedusae species have several similar morphological characters especially within the genus Craspedacusta (Jankowski, 2001), and only a few records gave more detailed descriptions of specific morphological characters (Kramp, 1950; Jankowski, 2001; Lewis et al., 2012). Indeed, up to eleven Craspedacusta species have been described, mostly recorded from China only (Jankowski, 2001). However, according to Bouillon et al. (2006) and Jankowski et al. (2008), many species may not be valid and are likely to be just morphological variations of the same species. Jankowski (2001) studied all the species recorded within Craspedacusta in detail and found that only three should be considered valid (C. sowerbii, C. iseanum Oka and Hara, 1922 and C. sinensis Gaw and Kung, 1939), and two are uncertain (C. sichuanensis He and Kou, 1984 and C. ziguiensis He and Xu, 1985); the rest seem to be synonyms of C. sowerbii, keeping in mind that two other species were synonymised [the marine species C. vovasi Naumov and Stepanjants, 1971 and the brackish water one C. marginata Modeer, 1791 (see Hummelinck, 1938). In the present paper, the morphological characteristics of our specimens coincide with the typical characters of the medusae belonging to the genus Craspedacusta (Russell, 1953; Bouillon and Boero, 2000; Bouillon et al., 2004, 2006). They have, apart from a well–developed marginal nematocysts ring, four simple radial canals from which pouch–like gametogenic tissues are hanging, and centripetal vesicles embedded in the

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LN901194 Craspedacusta sowerbii (locality unknown) MF177130 Craspedacusta sowerbii (Chile – Espejo) MF177129 Craspedacusta sowerbii (Chile – Espejo) MF177128 Craspedacusta sowerbii (Chile – Espejo) MF177127 Craspedacusta sowerbii (Chile – Espejo) MF177126 Craspedacusta sowerbii (Chile – Espejo) MF177125 Craspedacusta sowerbii (Chile – Espejo) MF177124 Craspedacusta sowerbii (Chile – Espejo) MF177123 Craspedacusta sowerbii (Chile – Espejo) MF177122 Craspedacusta sowerbii (Chile – Espejo) MF177121 Craspedacusta sowerbii (Chile – Espejo) MF177120 Craspedacusta sowerbii (Chile – Illahuapi) MF177120 Craspedacusta sowerbii (Chile – Illahuapi) MF177119 Craspedacusta sowerbii (Chile – Illahuapi) MF177118 Craspedacusta sowerbii (Chile – Illahuapi) MF177117 Craspedacusta sowerbii (Chile – Illahuapi) MF177116 Craspedacusta sowerbii (Chile – Illahuapi) MF177115 Craspedacusta sowerbii (Chile – Illahuapi) MF177114 Craspedacusta sowerbii (Chile – Illahuapi) MF177113 Craspedacusta sowerbii (Chile – Illahuapi) 61/0.93 MF177112 Craspedacusta sowerbii (Chile – Illahuapi) MF177111 Craspedacusta sowerbii (Chile – Illahuapi) MF177110 Craspedacusta sowerbii (Chile – Leu–Leu) MF177110 Craspedacusta sowerbii (Chile – Leu–Leu) MF177109 Craspedacusta sowerbii (Chile – Leu–Leu) MF177108 Craspedacusta sowerbii (Chile – Leu–Leu) MF177107 Craspedacusta sowerbii (Chile – Leu–Leu) MF177106 Craspedacusta sowerbii (Chile – Leu–Leu) MF177105 Craspedacusta sowerbii (Chile – Leu–Leu) 100/0.1 MF177104 Craspedacusta sowerbii (Chile – Leu–Leu) MF177103 Craspedacusta sowerbii (Chile – Leu–Leu) MF177102 Craspedacusta sowerbii (Chile – Leu–Leu) MF177101 Craspedacusta sowerbii (Chile – Leu–Leu) MF177131 Craspedacusta sowerbii (Chile – Ancapulli) MF177132 Craspedacusta sowerbii (Chile – Ancapulli) 42/0.73 MF177133 Craspedacusta sowerbii (Chile – Ancapulli) MH230079 Craspedacusta sowerbii (Italy – Sycily) MG924343 Craspedacusta sowerbii (india – Kerala) NC018537 Craspedacusta sowerbii (China – Hubei) 100/1 81/0.63 KP231217 Craspedacusta sowerbii (Greece – Lake Marathon) MF000493 Craspedacusta sowerbii (Switzerland – Rinwiler Weier) 64/0.50 KF510026 Craspedacusta sowerbii (China – Sichuan)

Cox1

MK600508 Craspedacusta sowerbii BECA–H1 (Morocco, Bin El Ouidan)

99/– MK600508 Craspedacusta sowerbii BECA–H2 (Morocco, Bin El Ouidan) FJ423620 FJ423619 FJ423618 FJ423617 38/– FJ423616 FJ423615 FJ423614 FJ423613

49/–

96/1

25/–

38/–

Craspedacusta Craspedacusta Craspedacusta Craspedacusta Craspedacusta Craspedacusta Craspedacusta Craspedacusta

sowerbii sowerbii sowerbii sowerbii sowerbii sowerbii sowerbii sowerbii

(Germany (Germany (Germany (Germany (Germany (Germany (Germany (Germany

– – – – – – – –

Diezsee) Hohwiesensee) Canyon Suplingen) Kligenberg) Lobejun) Matschelsee) Fluckinger See) Schonbach)

100/1 AF383927 Maeotias marginata AF383926 Maeotias marginata KF962130 KF962131 KF962132 100/1 KF962133 KF962134 KF962135 KF962136 KF962137 KF962138 KF962139

Gonionemus Gonionemus Gonionemus Gonionemus Gonionemus Gonionemus Gonionemus Gonionemus Gonionemus Gonionemus

sp. sp. sp. sp. sp. sp. sp. sp. sp. sp.

0.5 MK600508 Craspedacusta sowerbii BECA–H1 (Morocco, Bin El Ouidan)

MK600509 Craspedacusta sowerbii BECA–H2 (Morocco, Bin El Ouidan)

0.99

JX121605 Olindias phosphorica JN00942 Cuabaia aphrodite AY530418 Obelia geniculata 0.05

0.82

KF510026 Craspedacusta sowerbii (China – FJ423620 Craspedacusta sowerbii (Germany FJ423619 Craspedacusta sowerbii (Germany FJ423618 Craspedacusta sowerbii (Germany FJ423617 Craspedacusta sowerbii (Germany FJ423616 Craspedacusta sowerbii (Germany FJ423615 Craspedacusta sowerbii (Germany FJ423614 Craspedacusta sowerbii (Germany FJ423613 Craspedacusta sowerbii (Germany

Sichuan) – Diezsee) – Hohwiesensee) – Canyon Suplingen) – Kligenberg) – Lobejun) – Matschelsee) – Fluckinger See) – Schonbach)

Fig. 3. Molecular analysis by the ML method. Relationship of olindiid species using Obelia geniculata as outgroup; the analysis is based on Cox1. The tree is drawn to scale, with branch lengths measured in the number of substitutions per site. The white arrow indicates the clade where Moroccan sequences merge; note the slightly different arrangement of the sequences by Bayesian method (subtree at the bottom right of the figure). Symbols of groupings/clades correspond to the Cox1 map in figure 6. Fig. 3. Análisis molecular mediante el método de la máxima verosimilitud. Relación entre especies de olíndidos utilizando Obelia geniculata como grupo externo; el análisis se basa en Cox1. El árbol está dibujado a escala y la longitud de las ramas indica el número de sustituciones por sitio. La flecha blanca indica el clado en el que se insertan las secuencias marroquíes; nótese que existe una ligera diferencia en la ordenación de las secuencias obtenidas mediante el método bayesiano (subárbol en la esquina inferior derecha de la imagen). Los símbolos de las agrupaciones o clados son los mismos que aparecen en el mapa Cox1 de la figura 6.

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16S

MK600506 Craspedacusta sowerbii BECA–H1 (Morocco, Bin El Ouidan)

65/0.85 99/1 MK600507 Craspedacusta sowerbii BECA–H2 (Morocco, Bin El Ouidan) KY077294 Craspedacusta sowerbii (locality unknown) MF000530 Craspedacusta sowerbii (Switzerland – Lake Leman) 100/1 95/1 MF000530 Craspedacusta sowerbii (Switzerland – Ringwiler Weier) 44/– 82/0.95

78/0.97

26/–

AY512507 Craspedacusta sinensis (China) EU293974 Craspedacusta ziguiensis (China)

90/0.99

44/0.50

EU293971 Craspedacusta sowerbii (USA – Lake Huato) KX267739 Craspedacusta sowerbii (Uruguay – Del Medio Lagoon)

100/1KY077295 Limnocnida tanganjicae KY077295 Limnocnida tanganjicae EU293975 Astrohydra japonica

AY512508 Maeotias marginata AB720909 Scolionema suvaense

EU293976 Gonionemus vertens KX565923 Gonionemus vertens KF962471 Gonionemus sp. KF962472 Gonionemus sp. KF962473 Gonionemus sp. 67/0.65 KF962474 Gonionemus sp. KF962475 Gonionemus sp. 97/– 94/– KF962476 Gonionemus sp. KF962477 Gonionemus sp. KF962478 Gonionemus sp. 99/1 KF962479 Gonionemus sp. KF962480 Gonionemus sp. 97/0.97 EU293973 Aglauropsis aeora KF184031 Olindias formosus EU293978 Olindias phosphorica 100/1 AY512509 Olindias phosphorica KT266630 Olindias sambaquiensis 97/0.99 93/1 EU293977 Olindias sambaquiensis 100/1 94/1

99/1

75/0.85 AY512534 Moerisia sp KX355402 Moerisia sp. 89/0.62 KT266626 Moerisia inkermanica KF962504 Moerisia inkermanica 100/1 EU876555 Moerisia sp. KF962503 Moerisia inkermanica KF962501 Moerisia inkermanica 64/0.86 KF962500 Moerisia inkermanica FJ550503 Obelia bidentata 0.1

Fig. 4. Molecular analysis by the ML method. Relationship of olindiid species using Obelia bidentata as outgroup, the analysis is based on 16S. The tree is drawn to scale, with branch lengths measured in the number of substitutions per site. Symbols of groupings/clades correspond to the 16S map in figure 6. Fig. 4. Análisis molecular mediante el método de la máxima verosimilitud. Relación entre especies de olíndidos utilizando Obelia bidentata como grupo externo; el análisis se basa en 16S. El árbol está dibujado a escala y la longitud de las ramas indica el número de sustituciones por sitio. Los símbolos de las agrupaciones o clados son los mismos que aparecen en el mapa 16S de la figura 6.

velum as internal closed ecto–endodermal statocysts. Because these common similar characters within the Craspedacusta species lead to confusion and doubts when identifying a specimen and ascribing it to a determined species, the application of more specific characters is needed. The here–observed specimens identified as C. sowerbii have four prominent large perradial tentacles, which are clearly shorter in C. sinensis (Kramp, 1950; Jankowski, 2001). This latter species, also found in China, is very similar to C. sowerbii. According to Kramp (1950), it differs from C. sowerbii also in having a markedly irregular distribution of the different tentacle sizes (which are evenly distributed in our medusae), as well as a characteristic nematocyst distribution on tentacles. In our specimens, transverse belts of clustered groups of two to 10 nematocysts cover the tentacles, while in C. sinensis, nematocysts

are located at the end of elongated cylindrical papillae that are not arranged in transverse rings on tentacles (Kramp, 1950; Jankowski, 2001). Moreover, in active swimming specimens, C. sinensis is easily recognizable by its remarkable changes in the umbrella diameter. This species actually varies from 0.48 cm at systole (contracted bell–shaped umbrella) to 1.8 cm during diastole (maximum dilated flattered umbrella) (Kramp, 1950). On the other hand, the observed extended tubular statocysts of different lengths and embedded in the velum of the here–studied specimens confirm our identification and discard the possibility of ascribing our medusae to C. iseanum. This species, found in Japan, is also very similar to C. sowerbii. According to Uchida (1955), the statocysts in C. iseanum are oval–shaped. Moreover, adult specimens in this species vary from five to 18 mm of umbrella diameter and have up to

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FJ423621 Craspedacusta sowerbii (Germany – Gaensedrecksee) FJ423622 Craspedacusta sowerbii (Germany – Lusshardt) FJ423624 Craspedacusta sowerbii (Germany – Schoenbach) FJ423625 Craspedacusta sowerbii (Germany – Flueckger See) FJ423626 Craspedacusta sowerbii (Germany – Matschelsee) FJ423627 Craspedacusta sowerbii (Germany – Klingenberg) FJ423628 Craspedacusta sowerbii (Germany – Canyon Suplingen) FJ423629 Craspedacusta sowerbii (Germany – Hohwiesensee) FJ423630 Craspedacusta sowerbii (Germany – Willersinn) FJ423631 Craspedacusta sowerbii (Germany – Streikoepfle) 54/– FJ423632 Craspedacusta sowerbii (Germany – Donau) FJ423633 Craspedacusta sowerbii (Germany – Loebejuen) AY513615 Craspedacusta kiatingi (China – Deyang) AY513616 Craspedacusta kiatingi (China – Huyu) 99/1 AY513617 Craspedacusta kiatingi (China – Pixian) AY513618 Craspedacusta kiatingi (China – Wenzhou) AY513619 Craspedacusta kiatingi (China – Yuntaishan) 99/1 AY513620 Craspedacusta sichanensisi (China – Quingchengshan) FJ423623 Craspedacusta sowerbii (Germany – Quarry Lake Diez) AY513637 Craspedacusta ziguiensis (China – Zigui) 60/0.86 99/1

AY513621 Craspedacusta sinensis (China – Jiang an) AY730678 Craspedacusta sinensis (China – Chongqing) AY513614 Craspedacusta brevinema (China – Zigui) AY730676 Craspedacusta sinensis (China – Maoping) AY513622 Craspedacusta sinensis (China – Luzhou) 86/0.82 AY513623 Craspedacusta sinensis (China – Wuyanghe) AY730675 Craspedacusta sinensis (China – Jiuwanxi) 73/0.82 64/0.62 AY730677 Craspedacusta sinensis (China – Xiangxi) 99/1

Clade II

38/0.68

Clade I

MK600504 Craspedacusta sowerbii BECA–H1 (Morocco, Bin El Ouidan) MK600505 Craspedacusta sowerbii BECA–H2 (Morocco, Bin El Ouidan)

ITS

64/0.87

AY513630 Craspedacusta sowerbii (China – Xin anjiang) JN874930 Craspedacusta sowerbii (China – Yongkang o Jinhua) AY513625 Craspedacusta sowerbii (China – Hengxi) JN874927 Craspedacusta sowerbii (China – Jiangbei of Ningbo) 90/0.56 JN874929 Craspedacusta sowerbii (China – Jiangbei of Ningbo) AY513629 Craspedacusta sowerbii (China – Xiangshan) JN874928 Craspedacusta sowerbii (China – Pingyang of Wenzhou) KY994575 Craspedacusta sowerbii (Chile) KY947356 Craspedacusta sowerbii (Chile) KY947355 Craspedacusta sowerbii (Chile) KY947354 Craspedacusta sowerbii (Chile) MH500048 Craspedacusta sowerbii (Italy – Sicily) AY513633 Craspedacusta sowerbii (China – Zhijiang) Ay513634 Craspedacusta sowerbii (China – Zhuzhou) AY513635 Craspedacusta xinyangensis (China – Anji) 17/0.58 AY513632 Craspedacusta sowerbii (China – Zhelin) AY513641 Craspedacusta sp. (China – Zigui) AY513626 Craspedacusta sowerbii (China – Louta) AY513624 Craspedacusta sowerbii (China – Daye) AY513638 Craspedacusta sp. (China – Zigui) AY513636 Craspedacusta xinyangensis (China – Yuantouzhu) 62/0.83 AY513640 Craspedacusta sp. (China – Zigui) AY513631 Craspedacusta sowerbii (China – Yonhkang) AY513628 Craspedacusta sowerbii (China – Puyang) AY513627 Craspedacusta sowerbii (China – Nanyang) 0.1

Clade III

66/0.56

KM603474 Obelia dichotoma

Fig. 5. Molecular analysis by the ML method. Relationship of Craspedacusta species using Obelia dichotoma as outgroup; the analysis is based on ITS. The tree is drawn to scale, with branch lengths measured in the number of substitutions per site. Symbols of groupings/clades correspond to of the ITS map in figure 6. Fig. 5. Análisis molecular mediante el método de la máxima verosimilitud. Relación entre especies de Craspedacusta utilizando Obelia dichotoma como grupo externo; el análisis se basa en ITS. El árbol está dibujado a escala y la longitud de las ramas indica el número de sustituciones por sitio. Los símbolos de las agrupaciones o clados son los mismos que aparecen en el mapa ITS de la figura 6.

128 tentacles (Lewis et al., 2012), while C. sowerbii adult specimens, like in the here–observed medusae, can reach up to 25 mm of umbrella diameter and have more than 400 tentacles (Russell, 1953; Jankowski, 2001). Concerning the nematocysts, C. iseanum have scattered, and not clustered, nematocysts on the tentacles. Nevertheless, all these specific characters may be trustworthy only when dealing with the identification of adult living specimens or at least adult well–preserved ones.

Phylogenetic analyses Despite the abundant literature reporting the occurrence of Craspedacusta sowerbii around the world (see Dumont, 1994; Didžiulis and Żurek, 2013 for additional references), available molecular information is relatively scarce. Part of this information is published as a representation of the genus (or family) for general phylogenetic papers about different cnidarian taxa (e.g. Collins, 2002; Kayal et al., 2015; Grange

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et al., 2017), and there are few specific papers on the phylogeny, molecular systematics, diversity and distribution of the genus Craspedacusta (Zou et al., 2012; Fritz et al., 2009; Zhang et al., 2009). A number of sequences can be obtained in databases such as GenBank, ca. 55 of them for the ITS region (for seven putative species), ca. 50 sequences are for the Cox1 fragment (all of them attributed to C. sowerbii), and seven are for the 16S (ascribed to three different species). As previously commented, recent morphological (Jankowski 2001) or molecular (Fritz et al., 2009; Zhang et al., 2009) contributions drastically reduced the number of species to three (or maybe five). On one hand, for Jankowki (2001), only C. sowerbii, C. iseanum, and C. sinensis (and perhaps C. sichuanensis and C. ziguiensis) could be considered valid. On the other hand, the two simultaneous contributions by Fritz et al. (2009) and Zhang et al. (2009) pointed out the existence of different lineages (in the former) and species (in the later), thus stressing the lack of consensus on the real diversity and systematics of the genus. Fritz et al. (2009) considered that the morphology of their German and Austrian samples agrees with C. sowerbii, and hence the Chinese ITS sequences (identical to their European material) attributed to C. kiatingi should be considered as C. sowerbii var. kiatingi (Gaw and Kung, 1939; Kramp, 1950). Fritz et al., (2009) identified three main clusters within their dataset: 'sinensis' [for C. sinensis, and C. brevinema (considered by these authors as a variety of the former)], 'sowerbyi' [sic, for Chinese sequences of C. sowerbii and C. xianyangensis (considered by these authors as a variety of the former)], and “kiatingi” (for the German and Austrian sequences of C. sowerbii, and the Chinese sequences of C. sichuaensis and C. kiatingi (considered by these authors as a variety of their European C. sowerbii)], remaining as doubtful the status of C. ziguiensis. In short, for Fritz et al. (2009) the "data support the assumption that there are three valid species, with the possibility of C. ziguiensis being a fourth one, and several, morphological quite different sub–species or variations of the freshwater jellyfish C. sowerbii". Although the identification of C. sinensis and C. ziguiensis as different species seems to be clearly stated, the assignable different specific name to be used for the two other Clades ('kiatingi' and 'sowerbii') is not so clearly defined in this last paper. Zhang et al. (2009) analysed eight putative Craspedacusta species using the nuclear marker ITS. Obviously, the trees obtained by these authors show similar conclusions, as both research groups shared a similar set of sequences: C. xinyangensis should be the synonym of C. sowerbii, C. sichuanensis the synonym of C. kiatingi and C. brevinema the synonym of C. sinensis, while the taxonomic status of C. ziguiensis is still uncertain. However, the main difference between the two contributions is the implications of those Austrian and German sequences, defining a clade C. kiatingi–C. sowerbii. The Chinese authors were probably unaware at that moment that a number of European sequences could be attributed to their C. kiatingi.

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Recently, Schifani et al. (2018) and Fuentes et al. (2019) obtained sequences from specimens identified as C. sowerbii from Sicily (Italy) and Chile, respectively. In their Cox1 analyses, Schifani et al. (2018) identified a Sicilian–Grecian–Chinese clade as C. sowerbii, a second German–Chinese clade as C. kiatingi, and a third clade of unknown origin as Craspedacusta sp. (actually, this last sequence was deposited by Dr. P. Schuchert from a polyp stage specimen collected in Ringwiler Weier (Canton Zurich, Switzerland) (see also features part of this sequence in GenBank Accession number MF000493). The ITS analyses of these authors also identify a Sicilian–Chinese C. sowerbii clade and a German–Chinese C. kiatingi clade (the origin of the sequence FJ423632 is indicated to be German Donau (Danube) in GenBank. It is somehow ironic that the type locality of this considered–to–be invader alien species was a water–lily tank in the Botanical Garden in Regent's Park, London (Lankester, 1880a), while the origin of the species (and the diversity hot–spot of the genus Craspedacusta) is currently supposed to occur in the Yangtze River valley in China (e.g. Didžiulis and Żurek, 2013). Many changes have occurred since the late 19th century in the type locality, which finally disappeared in 1932 (C. Magdalena, pers. comm.). Anyway, the species was subsequently reported also from Southern England (Broom Water, Teddington), only a few kilometres away from the type locality (Green, 1998). Thus, it is plausible that future sequences obtained from a specimen collected in the London area could be considered a topotype (or a neotype could be established, since the original type material seems not to have been deposited in any institution), and could help to soundly and univocally define C. sowerbii. The nomenclatural consequences of the current absence of molecular data from the area of the type locality are in fact very important. For the moment, according to the ITS phylogenetic hypothesis, mainland Europe and Morocco share the same haplotype, which is also shared with specimens identified as C. kiaitingi and C. sichuanensis from China, the diversity hot–spot of this hydromedusa genus. Most Chinese specimens identified as C. sowerbii are in a well separated clade (see fig. 5). It is possible to speculate on two scenarios: 1) the South England sequences are identical (or similar) to those from mainland Europe and Morocco (Clade I); and 2) the South England sequences are different from those from mainland Europe and Morocco, but identical (or similar) to those identified as C. sowerbii by Zhang et al. (2009) and that from Sicily (Clade III). The direct consequence of the first scenario would be that C. kiatingi and C. sichuanensis sequences from Zhang et al. (2009) must instead be attributed to C. sowerbii, while another available name should be selected for those sequences identified as C. sowerbii by Zhang et al. (2009), the Clade 'sowerbyi' of Fritz et al. (2009), including the Sicilian and Chilean specimens (Clade III in this paper). Conversely, the consequences of the second scenario suggest that at least two species of Craspedacusta occur in Europe and North Africa, C. sowerbii in Southern England,

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Cox1

16S

ITS Fig. 6. Known worldwide distribution of the observed groupings (similar sequences, or clades) attributed to Craspedacusta sowerbii in the different phylogenetic analyses carried out in this paper (see comparatively fig. 3, 4, and 5). Fig. 6. Distribución mundial conocida de las agrupaciones observadas (secuencias similares o clados) atribuidas a Craspedacusta sowerbii en los diferentes análisis filogenéticos llevados a cabo en este trabajo (compárense las fig. 3, 4 y 5).

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and C. kiatingi [or another available name according to the ICZN (1999)] in the rest of this area, as well as in China. For the moment we have no ITS information from the specimens collected in Switzerland, Greece, or India. The Cox1 phylogenetic hypothesis clearly shows that Europe experienced the invasion of at least three Craspedacusta species, one of them in central Europe (Germany) and Morocco, the second one is currently present in Greece and Sicily, while the third one is found in Switzerland (see fig. 3 and 6).The two first invasions are connected (or simply share similar sequence) with their respective Chinese populations. This was already detected by Karaouzas et al. (2015), suggesting that the phylogeny of the genus is in need of further investigations, as genetic distances between the C. sowerbii clades are around 15 %. In the available Cox1 information for olindiid species, all genera except Craspedacusta are represented by a single species or haplotype, making difficult to discuss about the expected range of genetic distances at species level. In our phylogenetic hypothesis uncorrected p– distances between olindiid genera varies between 15 and 26 %. As in the previous discussion, the name to be used for each Craspedacusta species will depend on the knowledge of a (still unknown) Cox1 sequence from a putative Southern England population, possibly after the establishment of a neotype. Information based on our 16S phylogenetic hypothesis about Craspedacusta species delimitation is scarce, but it is well defined that a single species that can currently be identified in America (Lake Huato, USA and Uruguay) is different from the one present in North Africa (Morocco), Switzerland, and an unknown locality (sequence KY077294, see Grange et al., 2017). The same problem already discussed in assigning the name of C. sowerbii to one or another clade is present here. For this marker it is possible to discuss about the relative genetic distances (uncorrected p–distances) that are recognized between species of another olindiid genus, the genus Olindias (see Bouillon et al., 2004: 206, 2006: 435). Genetic distances between the three species of Olindias, from which 16S sequences are available, vary from 5.5–5.7 % (O. phosphorica to O. sambaquiensis) to 9.8–10.0 % (O. formosus to O. sambaquiensis) (see also Collins et al., 2005, 2008). The genetic distance observed between North African–Switzerland and American sequences identified as C. sowerbii is 4.2–4.5 %, between North African–Switzerland and C. ziguiensis and C. sinensis it is 7.8–8.9 % and 6.8–7.4 %, respectively; and between the latter two species it is 6.2 %. This suggests that American specimens attributed to C. sowerbii should perhaps be considered a different species from the specimens analysed here from North Africa, as well as from those from Switzerland. Final remarks The described scenario could be much more complicated when considering that in the type locality of C. sowerbii, the aquatic plants of the water–lily tank in

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Regent's Park (i.e. the potential dispersal vector of this hydromedusa species) were imported from Brazil, and not from China (C. Magdalena, pers. comm.). At present, it is difficult to know when the dispersion of Craspedacusta species from Easter Asia began, and most of the proposed vectors are in part speculative. Perhaps there was a combination of initial introduction by trade of aquatic plants and a subsequent natural dispersion by aquatic animals (e.g. birds, insects). For this reason, to solve this unstable nomenclatural and biodiversity problem, it is highly desirable to start with an important (or at least representative) field and molecular sampling programme in Southern England, in order to decide which haplotype (or set of related haplotypes) could be considered the true Craspedacusta sowerbii. At this moment, for the present contribution, the most parsimonious solution would be the existence of a single species (clade) in England and the relatively close Central Europe (also shared by the Moroccan examined specimens). If this is correct, this clade should retain the specific epithet sowerbii, and then, as in the same group of sequences there are some attributed to C. kiatingi by Zhang et al. (2009), these sequences and individuals should also be assigned to C. sowerbii. According to these considerations, all materials included in the 'sowerbyi' clade of Fritz et al. (2009) and the clade that included all sequences attributed to C. sowerbii by Zhang et al. (2009) (including the Chilean and Sicilian sequences) should be assigned to a different species, which should be selected among the available names after a complete bibliographical and morphological review. It has also become clear that according to Cox1, at least three Craspedacusta species are present in Europe (see fig. 6, Cox1): one in central Europe (and Morocco), one in Greece and Sicily (for the moment ITS sequences from the Greek specimens are not available), and one in Switzerland (for the moment ITS information is not available). Unfortunately, no geographical information is currently available for a 16S sequence (KY077294, see Grange et al., 2017) and a Cox1 sequence (LN901194, see Kayal et al., 2015). In the future, knowing the origin of these and other additional sequences would provide important information on invasion events concerning this intriguing hydromedusa species. The possible introduction vectors of Craspedacusta sowerbii in the recorded new sites generally remain unidentified. Accordingly, several hypotheses about the possible introduction paths have been mentioned in different reports and works on this species (Dumont, 1994; Angradi, 1998; Karaouzas et al., 2015). They can mainly be resumed in two possible vectors: 1) vectors facilitated by human activities, and 2) natural vectors. With regard to the former, the most likely dispersal vector may be the transfer of the species polyp stage, or the result of a resistance structure in aquaria or exhibition tanks, or an association with commercial ornamental aquatic plants or animals (Oscoz et al., 2010; Gasith et al., 2011; Gomes–Pereira and Dionisio, 2013; Minchin et al., 2016). The minuscule and hardly recognizable resting forms of the species make its unintentional human–mediated

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dispersal likely. Some authors reported observations of C. sowerbii (medusa or polyp stage) coinciding with the introduction of stocked fish or aquatic plants (Parent, 1982). As for the second possible vector, desiccated podocysts attached to body parts of aquatic invertebrates and vertebrates (including birds) could have allowed this species to colonize near freshwater reservoirs (Jankowski, 2001; Jankowski et al., 2008). The cysts are able to survive for about 40 years while being completely desiccated (Bouillon and Boero, 2000; Bouillon et al., 2006; Lewis et al., 2012). These resting bodies may accidentally be transferred to new sites on bird’s feet or plumage. Then, in favorable conditions, cysts turn into medusae and podocysts become polyps that can lead to more budding. This makes the aerial passive dispersal a possible introduction path for C. sowerbii (Parent 1982; Dumont 1994; Oscoz et al., 2010; Didžiulis and Żurek 2013; Failla–Siquier et al., 2017). The characteristic drought–resistant forms of the species suggest that the natural aerial dispersal vector by migrating birds (see Reynolds et al., 2015; Green, 2016) may be an important factor in the introduction of alien species such as C. sowerbii into Bin El Ouidan, although in this reservoir, in order to limit the proliferation of algae as well as to enhance the biodiversity within the reservoir ecosystem, many exotic species including fish and aquatic plants (e.g. Oncorhynchus mykiss Walbaum, 1792; Barbus barbus Linnaeus, 1758; Micropterus dolomieu Lacepède, 1802; Sander lucioperca Linnaeus, 1758; Hypophthalmichthys molitrix Valenciennes, 1844; Ctenopharyngodon idella Valenciennes, 1844 and Cyprinus carpio Linnaeus, 1758) started to be introduced in the reservoir one year after its construction (Rabii Souilem, pers. comm.). Acknowledgements The authors thank Mustapha R. Habiballah and Rabii Souilem, Technical manager of the fish farming company 'Asmak Nile' in Béni Mellal (both members of Moroccan Royal Federation of Diving and Underwater Activities, FRMPAS) for their invaluable help and especially for the sampling trip carried out during December 2015. Thanks also to Dr. Allen Collins (Smithsonian Institution, National Museum of Natural History, in Washington) for kindly facilitating sampling information on some C. sowerbii sequences from USA localities, and Dr. Miguel Angel Alonso Zarazaga (Museo Nacional de Ciencias Naturales, Madrid) for helping us on nomenclatural questions relating to the correct specific epithet that should be used in the hydromedusan species here discussed. We are also very grateful to Dr. Carlos Magdalena (Royal Botanical Gardens, Kew, London) for useful comments on the vicissitudes of Regent’s Park water–lily tanks during the early 20th century, and additional literature on the presence of C. sowerbii around the type locality. We wish to thank the Editors and two anonymous reviewers for useful comments and criticisms on an early draft of this paper, and also Mr. Tony Krupa for reviewing the English version of the manuscript.

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Towards inclusion of genetic diversity measures into IUCN assessments: a case study on birds L. C. Vitorino, U. J. Borges Souza, T. P. F. A. Jardim, L. Ballesteros–Mejia

Vitorino, L. C., Borges Souza, U. J., Jardim, T. P. F. A., Ballesteros–Mejia, L., 2019. Towards inclusion of genetic diversity measures into IUCN assessments: a case study on birds. Animal Biodiversity and Conservation, 42.2: 317–335, Doi: https://doi.org/10.32800/abc.2019.42.0317 Abstract Towards inclusion of genetic diversity measures into IUCN assessments: a case study on birds. The IUCN Red List categorizes species based on their geographical distribution and population size. However, attributes such as genetic information are not yet considered. We compiled information on genetic diversity (HE, HO) and inbreeding coefficient (f) along with their ecological attributes (IUCN category, migratory habit, forest dependence and habitat type) from a literature survey to assess whether bird species categorized as being of highest conservation concern display the lowest genetic diversity. We used generalized linear mixed models (GLMM) to test whether avian species with less inclusive characteristics (e.g., taxa with small geographical distributions or low dispersal capability) display lower genetic diversity than those classified as Least Concern (LC). We used phylogenetic generalized least squares (pGLS) to account for phylogenetic independence of predictor variables and to verify robustness of GLMMs (generalized linear mixed models). In general, GLMM revealed more significant relationships among ecological attributes and genetic diversity patterns. After accounting for phylogenetic independence, the highest average heterozygosity values were observed in species falling under the LC category; non–migratory birds showed lower HO and HE average values than migratory birds, while non–forest birds showed lower heterozygosity than forest birds. Hence, we corroborate our hypothesis that genetic diversity of birds is lower in species of high conservation concern. We hope our results promote further studies on genetic diversity of bird populations. Lastly, we propose the incorporation of genetic data as metrics in the assessment of bird conservation status. Key words: International Union for Conservation of Nature, Red List, Expected heterozygosity, Observed heterozygosity, Inbreeding coefficient Resumen Lograr la inclusión de las medidas de diversidad genética en las evaluaciones de la Unión Internacional para la Conservación de la Naturaleza: un estudio monográfico sobre aves. La Lista Roja de la Unión Internacional para la Conservación de la Naturaleza (UICN) clasifica las especies según su distribución geográfica y el tamaño de población. Sin embargo, todavía no se tienen en cuenta algunos aspectos como la información genética. A fin de evaluar si las especies de aves clasificadas como de máximo interés para la conservación son las que presentan la menor diversidad genética, en este estudio compilamos información sobre la diversidad genética (HE, HO) y el coeficiente de endogamia (f), junto con sus características ecológicas (categoría de la UICN, hábitos migratorios, dependencia de los bosques y tipo de hábitat) a partir de un estudio de las publicaciones científicas. Utilizamos modelos mixtos lineales generalizados para determinar si las especies de aves con menos características inclusivas (por ejemplo, los taxones con una distribución geográfica reducida o con escasa capacidad de dispersión) presentan menor diversidad genética que las clasificadas como de Preocupación Menor. Utilizamos mínimos cuadrados generalizados filogenéticos para representar la independencia filogenética de las variables predictivas y para comprobar la robustez de los modelos mixtos lineales generalizados. En general, los modelos mixtos lineales generalizados revelaron la existencia de relaciones más significativas entre las características ecológicas y los patrones de diversidad genética. Al tener en cuenta la independencia filogenética, los valores máximos de heterocigosidad media se observaron en especies de la ISSN: 1578–665 X eISSN: 2014–928 X

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categoría Preocupación Menor; las aves no migratorias mostraron valores medios de HO y HE más bajos que los de las aves migratorias, mientras que las aves no forestales mostraron una heterocigosidad inferior a la de las aves forestales. Por consiguiente, corroboramos nuestra hipótesis de que la diversidad genética de las aves es inferior en especies de gran interés para la conservación. Esperamos que nuestros resultados promuevan nuevos estudios sobre la diversidad genética de las poblaciones de aves. Por último, proponemos que se incorporen datos genéticos como parámetros en la evaluación de la situación de la conservación de las aves. Palabras clave: Unión Internacional para la Conservación de la Naturaleza, Lista Roja, Heterocigosidad esperada, Heterocigosidad observada, Coeficiente de endogamia Received: 27 XI 18; Conditional acceptance: 07 III 19; Final acceptance: 25 III 19 Luciana Cristina Vitorino, Departamento de Ciências Biológicas, Instituto Federal Goiano, Campus Rio Verde, C.P. 66, 75901–970, Rio Verde, GO, Brazil.– Ueric José Borges Souza, Tatianne Piza Ferrari Abreu Jardim, Laboratório de Genética & Biodiversidade, Instituto de Ciências Biológicas, Universidade Federal de Goiás, C.P. 131, 74001–970, Goiânia, GO, Brazil.– Ueric José Borges de Souza, Instituto Nacional de Ciências e Tecnologia em Ecologia, Evolução e Conservação da Biodiversidade, Universidade Federal de Goiás, Goiânia, GO, Brazil.– Liliana Ballesteros–Mejia, Muséum national d’Histoire naturelle, Sorbonne Université, Institut de Systématique, Evolution, Biodiversité (ISYEB), UMR 7205 – CNRS, MNHN, UMPC, EPHE, Paris, France. Corresponding author: Luciana C. Vitorino. E–mail: luciana.vitorino@ifgoiano.edu.br

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Introduction The IUCN Red Lists of threatened species, hereafter Red Lists, is the source of best available information on the global conservation status of species worldwide, providing quantitative measures of extinction risks (Lamoreux et al., 2003) and associated drivers (Baillie et al., 2004). Red Lists are an important tool not only to prioritize species conservation and identification of key biodiversity areas, but also to guide conservation responses, provide support for planning or implementing biodiversity projects, and help understand potential impacts on biodiversity (Bennun et al., 2018). In this sense, Red List indicators bear the potential to quantify possible anthropogenic threats to species (e.g., Wraith and Pickering, 2018), and to strategically connect science and politics (Do et al., 2018; Rabaud et al., 2018). Once species are identified as being at risk in the Red Lists, it might be easier to induce willingness– to–pay for nature conservation in the broader public (Tisdell et al., 2007; Jin et al., 2018). Threatened species are typically prioritized in conservation policies because of the risk of their vanishing even before we can describe their characteristics, or before we know them as important parts of ecosystems. However, non–threatened species should also receive attention, as population–level analyses may reveal the local influence of anthropogenic changes, such as habitat loss and/or fragmentation. Such changes can increase selection pressures and culminate in genetic erosion, thereby endangering population persistence in the longer term (Bijlsma and Loeschcke, 2012). Assigning species to a given threat category is based on five data– driven criteria concerning: i) population size, ii) population fragmentation, iii) observed or projected declines in abundance, iv) geographic range size in combination with fragmentation, and v) a quantitative analysis of extinction probability (IUCN, 2001). However, there is still an important aspect of biodiversity that is largely ignored in conservation assessments of species: genetic diversity. Despite its importance for maintaining biological distinctiveness and evolutionary processes, measures of genetic diversity such as number of alleles, number of haplotypes and heterozygosity are not explicitly considered in the Red Lists. This is somewhat puzzling since there is consensus that conservation of endangered species requires deep knowledge of metapopulation dynamics and structure, which involves determining the degree of genetic diversity within and between populations. Variability estimated using molecular markers not only helps to distinguish genetically distinct populations that may be vulnerable to environmental changes (e.g., Lee and Mitchell–Olds, 2011; Hansen et al., 2012; Limborg et al., 2012; Munday et al., 2013; Razgour et al., 2018) but also infers phylogenetic relationships between individuals both within and between species, reconstructing genealogies and gathering information on inbreeding rates (e.g., Zollinger et al., 2012; McCormack et al., 2013; Lyu et al., 2018). The current use of microsatellite markers in biodiversity conservation studies is particularly useful to address

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issues related to the conservation genetics of various bird species (e.g., Moura et al., 2017; Houston et al., 2018; Moussy et al., 2018; Stojanovic et al., 2018). Conservation Genetics has been defined as the discipline that applies genetic concepts and tools, including molecular markers, to small populations to reduce their risk of extinction (Frankham et al., 2002; Allendorf et al., 2012). Among its many applications, it allows to detect potential bottlenecks, measure gene flow and hybridization between populations, assess paternity, assign individuals to their population of origin, and infer population structure (e.g., Contina et al., 2018; Coster et al., 2018; Haworth et al., 2018; Kangas et al., 2018). Birds are often used as bioindicators for various reasons (Furness et al., 1993; Herrera–Dueñas et al., 2014; Padoa–Schioppa et al., 2006; Silva et al., 2018; Souto et al., 2018), including the fact that their populations are heavily affected by air pollutants and pesticides. This has put many bird populations at risk, increasing the need for studies addressing loss of genetic diversity in metapopulations. Spielman et al. (2004) compared the heterozygosity (He) of species categorized as threatened vs non–threatened taxonomically related species and found that, on average, He was 35 % lower in threatened species. Similarly, Evans and Sheldon (2008) used Phylogenetic Independent Contrast (Felsenstein, 1985) to correlate heterozygosity with the increased extinction risk, showing that genetic diversity is relatively poor in the threatened bird species category. Here we chose to use a more robust statistical approach. Following Ballesteros–Mejia et al. (2016), we fitted GLMMs to test the influence of different ecological attributes and extinction risk on estimates of genetic diversity. GLMMs combines desirable properties of two statistical frameworks, namely linear mixed models, incorporating random effects, and GLM, which handles non–normal data (Bolker et al., 2009). Since phylogeny is known to influence ecological and morphological characteristics (Harvey and Pagel, 1991; Bennett and Owens, 2002), we used generalized least squares (pGLS) to account for phylogenetic relationships and verify the robustness of the results found by significant GLMMs. All species in a monophyletic group share a common ancestor and tend to resemble each other more than those randomly chosen across a phylogenetic tree. In light of their phylogenetic non–independence, the former cannot be considered as independent data points in statistical analyses (Garland et al., 1992) We addressed the relationship between genetic diversity and conservation status and tested the effect of ecological attributes on patterns of genetic diversity based on data from a literature survey. Specifically, we used GLMM to test the effect of migratory habits (migratory or non–migratory), forest dependence (high, medium or low), type of habitat (terrestrial vs non–restricted to it), and extinction risk (as classified in the IUCN Red List) on the patterns of genetic diversity (HO, observed heterozygosity; HE, expected heterozygosity; f, inbreeding coefficient). Additionally, we fitted Phylogenetic Least Squares (pGLS) to

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account for phylogenetic non–independence as well as to verify the robustness of GLMM predictions. Most taxa are adversely affected by genetic factors before being driven to extinction (Spielman et al., 2004), but since processes that compromise genetic diversity do not affect all bird species equally, habitat specialization may be a predictor of the demographic and genetic consequences of fragmentation (Khimoun et al., 2016). Thus, it is likely that species classified as endangered or critically endangered exhibit lower HO/ HE and higher f values. Therefore, we assessed whether the genetic diversity in birds based on microsatellite data available in the literature can be used as a proxy to define their risk category and inform conservation policies. Material and methods Obtaining bird genetic data and ecological attributes We conducted a survey of studies on avian population genetic based on microsatellite data. We used the Web of Science (http://www.isiknowledge.com) database with the following combinations of keywords: [birds* AND microsatellite* AND genetic diversity*], [birds* AND SSR* AND genetic diversity*], [bird* AND microsatellite* AND genetic diversity*] and [bird* AND SSR* AND genetic diversity*].We excluded studies that used less than four microsatellite loci. From the publications, we retrieved the following data: (i) title; (ii) year of publication; (iii) journal; (iv) study species; (v) number of individuals; (vi) number of loci; and (vii) mean values of genetic diversity; HE, HO and f. In several studies, f was not estimated. Thus, we estimate it using the equation: f = 1 – (HO / HE). We used the database of the IUCN Red List of Threatened Species (http://www.iucnredlist.org) to extract information about the conservation status of target species. We considered the following categories: Least Concern (LC), Near Threatened (NT), Vulnerable (VU), Endangered (EN), Critically Endangered (CR), Extinct in the Wild (EW) and Extinct (EX) as of December 2015 to July 2016. The Birdlife International database (http://www. birdlife.org/datazone/species) contains data on avian species worldwide, and we used it to obtain the following information for each bird species in the selected studies: (i) migratory habit (migratory or or low dependence or not occurring in forests) and (iii) habitat type (terrestrial or other). This information, as well as the conservation status of the species (IUCN Category), was considered here as the ecological attributes. Data analysis To evaluate the temporal trend in the number of articles published annually, and to correct the effect of the general increase in the number of articles over

time, we use the following equation (Eq. 1): Number of articles in year (x) Total number of articles in the Web of Science in year (x) we Initially used analysis of variance and t–student to explore genetic variation among populations. We included bird populations of the same species analysed with the same microsatellite markers (table 1) to test whether genetic parameters HO, HE and f varied significantly between population pairs. We fitted generalized linear mixed models (GLMMs) to investigate the effects of ecological attributes and conservation status on genetic diversity. IUCN category and forest dependence were treated as multistate categorical variables, whereas habitat type and migratory habit were treated as binary variables. Models were fitted for each genetic parameter (HO, HE and f) as response variables. Ecological attributes were fitted as fixed factors, and species identity was considered as a random factor because multiple variables were measured per species. Analyses were performed using MCMCglmm package (Hadfield, 2010) implemented in R version 3.4.4 (R Core Team, 2018). To account for phylogenetic non–independence of the effects of ecological attributes on genetic diversity, we first obtained the reference phylogenetic hypothesis of the species included in each analysis. We gathered ten thousand phylogenies sampled from a pseudo–posterior distribution (Jetz et al., 2012) deposited in BirdTree.org website (https://birdtree. org/). We made a consensus tree using Tree annotator 1.8.2 (Drummond et al., 2012) and dropped all species without data using the 'drop.tip' function in the package 'ape' (Paradis, 2004) implemented in R version 3.4.4. We then tested whether the studied ecological attributes showed a phylogenetic signal to account for phylogenetic relationships. We performed Abouheif’s proximity test of serial independence (Abouheif, 1999; Pavoinea et al., 2008) using the function 'abouheif. moran' from the R–package 'adephylo' (Jombart et al., 2010). We then fitted Phylogenetic Generalized Least Square Models (pGLS; Martins and Hansen, 1997) to the genetic parameters to verify whether GLMM models had resulted in robust inferences and hence the pattern persisted when accounting for phylogenetic relationships. We tested the solitary effect of each ecological attribute, as well as the effect of combining all of these in a complete model, on the genetic diversity parameters. When the same molecular marker was applied more than once to study a species the mean of the genetic parameters was used. The analyses were carried out using the package 'caper' (Orme, 2013) of the R. Finally, Pearson’s correlation analysis was performed to assess the effect of the number of individuals and the number of loci on the genetic diversity values; the number of loci and individuals were log–transformed (base 10) to reduce the discrepancies between values. Statistical tests were performed using the R statistical package.

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Table 1. Species investigated in multiple studies on different populations but using the same microsatellite panel. Tabla 1. Especies objeto de múltiples estudios sobre poblaciones distintas, pero utilizando el mismo grupo de microsatélites.

Species

Microsatellite loci

Aquila chrysaetos

Aa15, Aa26, Aa27, Aa36, Aa39 and Aa43 (Martínez-Cruz et al., 2002)

Bubo bubo

B101, B111, B126 and B11 (Isaksson and Tegelstrom, 2002)

Cyanistes caeruleus

Pca3, Pca4, Pca7, Pca8 and Pca9 (Dawson et al., 2000)

POCC1 and POCC6 (Bensch et al., 1997)

PATMP2-43 (Otter et al., 1998)

Ase18 (Richardson et al., 2000)

Pdoµ5 (Griffith et al., 1999)

Mcyµ (Double et al., 1998)

CcaTgu7, CcaTgu8, CcaTgu11, CcaTgu14, CcaTgu15, CcaTgu19,

CcaTgu25 and CcaTgu28 (Olano-Marin et al., 2010)

TG05-046, TG05-053, TG13-013 (Dawson et al., 2010)

Tgu07 (Slate et al., 2007)

Nipponia nippon

NnNF5 (Ji et al., 2004)

Passer domesticus

Pdoµ1 and Pdoµ4 (Neumann and Wetton, 1996)

Pdoµ5 (Griffith et al., 2007)

Pdoµ10 (Segelbacher et al., 2000)

Tetrao urogallus

TUD1, TUD3, TUD5 and BG15

(Segelbacher et al., 2000, Piertney and Hoglund, 2001)

Because of the cumulative effect of domestication and subsequent artificial selection, Gallus gallus was excluded from GLMM and pGLS analysis. Results Scientometrics The search revealed 359 published papers that met the criteria entered. They were published across 98 different journals between 1998 and 2015. Five of these journals (i.e., Molecular Ecology Resources, Conservation Genetics Resources, Conservation Genetics, Molecular Ecology and PLOS One) hosted 51.81 % of the articles; 35 journals published between four and eight articles, and the remaining 58 journals published only one article. The analysis of annual number of publications per journal revealed that 2014 was the year with the highest number of publications; 13 of the analysed journals published studies containing genetic diversity data of bird species. The second highest mean annual publication rate occurred in 2013 and in 2015; in each of these years, nine different journals published articles that analysed the genetic diversity of birds using microsatellite markers.

After correcting for the general trend with equation 1, the number of published papers with analyses of bird genetic diversity increased significantly increased over time (r = 0.740; p ≤ 0.01), especially between 2013 and 2014. Values of HE, HO and f were reported for 297 species (table 1s), 63 of which were represented more than once in the data set. Gallus gallus (7.80 % of the total species) was the most studied species followed by Passer domesticus (2.40 %), Aquila chrysaetos (1.20 %), and Tetrao urogallus (1.20 %). The total number of studied species was distributed among 94 families (table 1s) and 27 orders. the most highly represented orders were Passeriformes (44.10 %), Charadriiformes (8.10 %) and Galliformes (6.10 %) (fig. 1A). Most of the studied species were predominantly terrestrial (54.20 %) and exhibited no migratory behaviour (62.30 %). The analysis of species classification according to forest dependence showed that most species did not usually occur in forests (42.10 %), followed by species with low (23.20 %), medium (18.20 %) and high (16.50 %) forest dependence. Lastly, regarding classification according to IUCN category, the analyses showed that most of the studied bird species were categorized as LC (69.40 %) (fig. 1B).

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Gaviiformes Otidiformes Phoenicopteriformes Podicipediformes Tinamiformes Caprimulgiformes Dinornithiformes Piciformes Suliformes Apodiformes Columbiformes Pelecaniformes Sphenisciformes Cuculiformes Struthioniformes Procellariiformes Strigiformes Accipitriformes Ciconiiformes Coraciiformes Gruiformes Anseriformes Falconiformes Psittaciformes Galliformes Charadriiformes Passeriformes B

60 80 100 Total number of species

120

140

Extinct (EX) Extinct in the wild (EW) Critically endangered (CR) Endangered (EN) Vulnerable (VU) Near threatened (NT) Least concern (LC)

60 80 100 120 140 160 180 200 220 Total number of species

Fig. 1. A, bird orders evaluated in terms of articles addressing genetic diversity estimates and published between 1998 and 2015; B, classification of bird species that were the subject of articles published between 1998 and 2015 according to the IUCN category. Fig. 1. A, órdenes de aves que se evaluaron en cuanto a los artículos en los que se abordaban las estimaciones de la diversidad genética y que se publicaron entre 1998 y 2015; B, clasificación de las especies de aves que fueron objeto de artículos publicados entre 1998 y 2015, según la categoría de la UICN.

Genetic diversity of bird populations We compared the genetic diversity data obtained using the same microsatellite markers for three Aquila chrysaetos populations, two in Scotland and one in Slovakia, and found similar diversity patterns. HO values (0.52, 0.50 and 0.40, respec-

tively) were always lower than HE values (0.56, 0.54 and 0.43), whereas the f values were 0.34, 0.06 and 0.09 (fig. 2A). For the Bubo bubo, we compared data from four Spanish populations and one Norwegian population, and the highest average HO was observed in the latter (0.63), while the Spanish populations presented values of 0.4, 0.48,

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323

A Sweden

Norway

North Korea Scotland

China

Denmark

Ireland

Poland

England

Netherlands

780 km

Germany Belgium

Czech Republik Slovakia

France

Austria

Switzerland

Italy

Bosnia and Herzegovina Serbia Montenegro Macedonia Albania

f Aquila chysaetos

Hungary

Slovenia Croatia

Portugal

Japan

South Korea

Greece

Spain 780 km

Passer domesticus

f Bubo bubo

Cyanistes caeruleus

HE f

f Tetrao urogallus

Nipponia nippon

Fig. 2. Comparative map of the observed heterozygosity (HO), expected heterozygosity (HE) and inbreeding coefficient (f) values for different bird populations in different countries of Europe (A), China and Japan (B). Data for the same species were obtained using the same microsatellite loci. In the legend, the metrics with their colors refer to the species indicated on the right. Fig. 2. Mapa comparativo de los valores de heterocigosidad observada (HO), heterocigosidad esperada (HE) y coeficiente de endogamia (f) para distintas poblaciones de aves en diferentes países de Europa (A), China y el Japón (B). Los datos de las mismas especies se obtuvieron utilizando los mismos loci de microsatélites. En la leyenda, los parámetros con sus colores se refieren a las especies indicadas a la derecha.

0.38 and 0.51. The same pattern was observed for the mean HE with values of 0.6 in the Norwegian population and 0.39, 0.41, 0.37 and 0.55 for the Spanish populations. The f value was –0.02 for the Norwegian population and 0.01, 0.07, –0.14 and –0.03 for the Spanish populations, For Tetrao urogallus, we compared the average diversity values for three populations, one from the Czech Republic and two from Spain. The highest average HO was observed in the Czech population (0.67), while the Spanish populations showed HO values of 0.55 and 0.44. We observed the opposite pattern for the average HE and f values, with the lowest occurring in the Czech population (0.64 and –0.03) followed by 0.66 and 0.68 for HE and 0.33 and 0.15 for f in the Spanish populations. When comparing the Spanish and Czech populations we found significant differences for HO (t = 8.06, p < 0.0001) and HE (t = 19.81, p < 0.0001) according to the t–test. Significant differences were found between the average HO (F = 53.15, p < 0.001), HE (F = 11.57, p = 0.002) and f (F = 86.70, p < 0.001) values of four Passer domesticus populations from Scandina-

via, Belgium, France and England. The average HO was highest in the French population (0.87) followed by Scandinavian (0.85), Belgian (0.79) and English populations (0.71). The highest average HE values were found in the populations from France and Scandinavia (0.88 in both) followed by those from Britain (0.85) and Belgium (0.83). The f value was positive for all populations, and the lowest value was found in the population with the highest heterozygosity, the French population (0.01), followed by the Scandinavian (0.02), Belgian (0.05) and English populations (0.16). The t–test also revealed a significant difference between the mean HO (t = 11.49, p < 0.0001) and HE (t = 11. 84, p < 0.0001) values found in two populations of Cyanistes caeruleus, one Spanish (HO = 0.80 and HE = 0.79) and one Austrian (HO = 0.78 and HE = 0.76). The averages for f, however, were similar for the two populations with 0.03 in the Spanish population and –0.004 in the Austrian population. However, for Nipponia nippon, we found similar values when comparing a Japanese and a Chinese population: 0.48 and 0.46 for HO, 0.37 and 0.37 for HE, and 0.83 and 0.88 for f, respectively (fig. 2B).

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Table 2. Mean values of genetic diversity of microsatellite loci and ecological attributes of avian species: N, number of species analyzed; HO, observed heterozygosity; HE, expected heterozygosity; f, inbreeding coefficient. Tabla 2. Valores medios de la diversidad genética de los loci de microsatélites y características ecológicas de las especies de aves. N, número de especies analizadas; HO, heterocigosidad observada; HE, heterocigosidad esperada; f, coeficiente de endogamia. Ecological attribute

H E

201

0.59 ± 0.16

0.63 ± 0.16

0.07 ± 0.10

0.53 ± 0.17

0.54 ± 0.18

0.03 ± 0.20

0.52 ± 0.13

0.54 ± 0.14

0.04 ± 0.10

0.52 ± 0.19

0.54 ± 0.20

0.04 ± 0.07

0.53 ± 0.12

0.55 ± 0.11

0.03 ± 0.07

0.41 ± 0.00

0.40 ± 0.00

-0.02 ± 0.00

Migratory

109

0.59 ± 0.16

0.62 ± 0.16

0.06 ± 0.11

Non-migratory

176

0.56 ± 0.16

0.60 ± 0.16

0.06 ± 0.11

High

0.58 ± 0.16

0.61 ± 0.15

0.06 ± 0.12

Medium

0.60 ± 0.14

0.64 ± 0.14

0.05 ± 0.09

Low

0.56 ± 0.18

0.60 ± 0.18

0.06 ± 0.10

Not occurring in forests

120

0.56 ± 0.16

0.60 ± 0.16

0.06 ± 0.12

153

0.57 ± 0.17

0.60 ± 0.17

0.05 ± 0.11

the terrestrial environment 132

0.57 ± 0.15

0.61 ± 0.16

0.07 ± 0.11

IUCN category

Migratory habit

Forest dependence

Habitat type Terrestrial Not restricted to

Genetic diversity, ecological attributes and conservation status In general, we observed a large variation in the values of genetic diversity sampled, with HO, HE and f varying between 0.04 and 0.93, 0.08 and 0.91 and –0.61 and 0.64, respectively. When we evaluated the genetic diversity within orders, the lowest mean HO and HE values were observed in Otidiformes (HO = 0.44 and HE = 0.49), Accipitriformes (HO = 0.49 and HE = 0.50), Gaviiformes (HO = 0.45 and HE = 0.46) and Ciconiiformes (HO = 0.44 and HE = 0.47). Regarding the conservation status, we found variation in the values of genetic diversity among the IUCN categories (table 2). The LC category showed the highest genetic diversity values (HO = 0.59 and HE = 0.63). However, values decreased substantially across the categories of greater concern, with 0.5 and 0.54 being the averages observed for HO and HE of all other categories combined Birds with migratory habits presented higher heterozygosity values (HO = 0.59 and HE = 0.62) than those observed in non–migratory birds (HO = 0.56 and HE = 0.60), but no significant variations were found in

terms of f (f = 0.06 in both). Species that do not occur in forest environments or that have low dependence on forest fragments displayed the lowest averages of heterozygosity (HO = 0.56 and HE = 0.60 in both). In contrast, terrestrial birds or birds not restricted to terrestrial environments did not show differences in terms of heterozygosity (HO = 0.57 in both HE = 0.60 and HE = 0.61, respectively), although f was slightly higher for birds not restricted to terrestrial environments (table 2). When we accounted for phylogenetic relationships in the observed genetic patterns, the serial independence test showed that with the exception of the categorical variable IUCN category, all other attributes presented a significant phylogenetic signal (table 3). Phylogenetic generalized least squares performed on a complete model, with all variables combined, showed a significant relation between IUCN category and HO, HE and f values. We observed the highest HO values in the LC category (–0.021 ± 0.028) and the lowest for the EW category (–0.185 ± 0.089), which is the one of highest concern. The same pattern was observed for HE and f: the highest HE and f values were also observed in the LC category (0.079 ± 0.027

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and 0.172 ± 0.023, respectively) and the lowest in the EW category (–0.194 ± 0.086 and –0.009 ± 0.074). The migratory habit significantly affected HE and f values, equal to 0.042 ± 0.019 and 0.058 ± 0.016 for migratory and non–migratory birds, respectively. The forest dependence significantly affected only f, that turned out to be highest in birds with low dependence on forests (–0.021 ± 0.029) and lowest for species with medium forest dependence (–0.084 ± 0.030). Regarding habitat type, this significantly affected HO, HE and f, with the highest values observed in terrestrial species (–0.059 ± 0.023; –0.079 ± 0.022 and –0.041 ± 0.019) (table 4). When we evaluated the isolated effect of each variable, similar results were found for IUCN category and HO, HE and f, so that species of the category EW had significantly lower values of HO and HE (tables 2, 5), but the results differed for the other categorical variables. Migratory habit, forest dependence and type of habitat affected HO and HE. The pGLS confirmed the hypothesis that non–migratory species have lower diversity values. Likewise, HE values may be affected in non–forest bird species. We also observed that species in terrestrial habitat showed lower levels of HO and HE (table 5). A significant correlation was found between the number of loci used and the HO (r = –0.183; p ≤ 0.01) and HE (r = –0.191; p ≤ 0.01) estimates, but the correlations were negative (fig. 3). Conversely, no correlations were observed between the number of loci and f or between the number of individuals sampled and any estimate of genetic diversity. Discussion Scientometrics Microsatellite markers were developed in the 1980s (Tautz and Renz, 1984) and have since become increasingly popular in avian research. The observed annual increase in the number of published articles using estimates of genetic diversity for birds confirms this popularity. It should be noted that the accuracy of these indices is subject to the availability of study individuals, and is therefore favoured by larger sample sizes, which explains why most studies (69.40 %) were conducted with species in the Least Concern (LC) IUCN category. However, the IUCN system categorizes species as LC based on attributes such as a wide geographical distribution and large population size (IUCN, 2015), but global conservation status may not be representative of local trends, as indicated by national Red Lists. Garcia and Marini (2006) evaluated 494 threatened or near–threatened taxa of Brazilian birds finding that the classifications of only 26 % of these taxa were consistent with the global status, and revealing discrepancies between regional and global classifications. Such differences decrease the efficiency at which the IUCN list can be applied to establish national–scale conservation actions (Rodríguez et al., 2000), so studies should be conducted to evaluate the genetic diversity of bird species while

325

Table 3. Phylogenetic signal of ecological attributes for bird species included in the analyses of genetic diversity and differentiation using Abouheif's proximity test of serial independence: HO, observed heterozygosity; HE, expected heterozygosit; f, inbreeding coefficient. (Significant values are denoted in bold.) Tabla 3. Señal filogenética de las características ecológicas de las especies de aves incluidas en los análisis de la diversidad genética y la diferenciación utilizando la prueba de proximidad de la independencia serial, elaborada por Abouheif: HO, heterocigosidad observada; HE, heterocigosidad esperada; f, coeficiente de endogamia. (Los valores significativos se indican en negrita.) Ecological attribute

Observed Moran's I

p-value

IUCN category

–0.0457

0.9999

Migratory habit

0.0412

0.0010

Forest dependence

0.1519

0.0010

Habitat type

0.0590

0.0010

considering both state and national threat levels. In addition, the ability to perform studies involving estimates of genetic diversity is also apparently influenced by the behavioural traits of the study species. Indeed, such traits may either hinder or facilitate the sampling, which may explain why most studies have been performed on predominantly terrestrial (54.20 %) and non–migratory (62.30 %) species with low forest dependence (23.20 %) or species for which sampling was not restricted to forest environments (42.10 %). Most bird species in the studies (44.10 %) belonged to the order Passeriformes (passerines), the largest and most diverse avian order. The main Passeriformes lineages diversified on all continents and now occupy almost all terrestrial ecosystems (Barker et al., 2004), and they include approximately 5,700 species that account for nearly 60 % of all living birds. Passeriformes have been the focus of many ecological, behavioural, anatomical and evolutionary studies because of their ubiquity and enormous diversity (Ericson et al., 2014), generally driven by the colonization of new biogeographical regions (Kennedy et al., 2017). This order encompasses domestic species such as Poephila cincta and Serinus canaria as well as to the globally distributed Passer domesticus (2.40 %). The most predominant species in the literature was Gallus gallus (7.80 %), which has a large number of lineages distributed across the globe and has been widely used as a model organism in biochemical, molecular (e.g., Piekarski et al., 2015; Guizard et al., 2016) and genetic studies such as those describing

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Table 4. Phylogenetic generalized least squares for ecological attributes of bird species for each genetic parameter analysed. Data obtained considering the effect of the combination of all variables: HO, observed heterozygosity; HE, expected heterozygosity; f, inbreeding coefficient. T, T–value; P, P–value; C, coefficient ± SE. (Significant values are denoted in bold) Tabla 4. Mínimos cuadrados generalizados filogenéticos para las características ecológicas de las especies de aves para cada parámetro genético analizado. Datos obtenidos considerando el efectos de la combinación de todas las variables: HO, heterocigosidad observada; HE, heterocigosidad esperada; f, coeficiente de endogamia; T, valores de T; P, valores de P; C, coeficicente ± EE. (Los valores significativos se indican en negrita).

Ecological attribute Parameter

Adj–R = 0.0492 2

Adj–R = 0.135

Adj–R = 0.1942

IUCN category Intercept

0.636 ± 0.156 4.086 0.000

0.581 ± 0.149 3.890 0.000

–0.064 ± 0.128 –0.499 0.618

–0.021 ± 0.028 –0.755 0.451

0.079 ± 0.027 2.886 0.004

0.172 ± 0.023 7.354 0.000

–0.058 ± 0.044 –0.136 0.892

0.037 ± 0.042 0.890 0.374

0.088 ± 0.036 2.461 0.014

–0.058 ± 0.045 –1.277 0.202

0.018 ± 0.044 0.417 0.677

0.137 ± 0.037 3.670 0.000

–0.080 ± 0.042 –1.886 0.060

0.000 ± 0.041 –0.005 0.996

0.137 ± 0.035 3.929 0.000

CR EW

– – – – – –

–0.185 ± 0.089 –2.064 0.039 –0.194 ± 0.086 –2.256 0.025

– – –

–0.009 ± 0.074 –0.128 0.899

Migratory habit Migratory – – – – – –

– – –

Non–migratory

–0.005 ± 0.019 –0.269 0.787

0.042 ± 0.019 2.222 0.027

0.058 ± 0.016 3.639 0.000

Forest dependence High

– – – – – –

– – –

Medium –0.001 ± 0.037 –0.006 0.994 –0.031 ± 0.035 –0.890 0.374

–0.084 ± 0.030 –2.792 0.006

Low

–0.021 ± 0.029 –0.744 0.458

–0.005 ± 0.035 –0.163 0.870 –0.015 ± 0.034 –0.456 0.649

Not occurring in forests

0.034 ± 0.038 0.902 0.368

0.024 ± 0.037 0.652 0.515

–0.045 ± 0.031 –1.426 0.155

Terrestrial –0.059 ± 0.023 –2.572 0.011 –0.079 ± 0.022 –3.561 0.000

–0.041 ± 0.019 –2.149 0.032

Habitat type Not restricted to the terrestrial environment

– – – – – –

the genetic diversity of populations based on variations at microsatellite loci (Rajkumar et al., 2008; Zanetti et al., 2011; Babar et al., 2012). The large number of studies in Gallus gallus with its wide distribution obviates the obvious influence of its economic interest on the studies. Likewise, since Passer domesticus is widely distributed with generally large population sizes, it has been used as a model to predict cases of treatment bias given that individuals in a population differ in their susceptibility to capture (Simons et al., 2015).

– – –

Genetic diversity of bird populations The lack of significant differences between Aquila chrysaetos populations can be explained by the migratory nature of this species in Europe. It is categorized as LC in the European Red List with the expectation of increasing the size of local populations (European Red List, 2015). Native to mainland Europe and the Mediterranean, it currently occurs in the UK with a resident and most likely introduced population (Harrop et al., 2013). This would explain the limited differences

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327

Table 5. Phylogenetic generalized least squares for ecological attributes of bird species for each genetic parameter analysed. Data obtained by evaluating the isolated effect of each variable: HO, observed heterozygosity; HE, expected heterozygosity; f, inbreeding coefficient; T, T–value; P, P–value; C, coefficient ± SE. (Significant values are denoted in bold). Tabla 5. Mínimos cuadrados generalizados filogenéticos para las características ecológicas de las especies de aves para cada parámetro genético analizado. Datos obtenidos mediante la evaluación del efecto aislado de cada variable: HO, heterocigosidad observada; HE, heterocigosidad esperada; f, coeficiente de endogamia; T, valores de T; P, valores de P; C, coeficicente ± EE. (Los valores significativos se indican en negrita).

Ecological attribute Parameter

Adj–R = 0.008 2

Adj–R = 0.054

Adj–R = 0.139

IUCN category Intercept

0.614 ± 0.155 3.969 0.000

0.580 ± 0.152 3.817 0.000

–0.070 ± 0.128 –0.542 0.588

–0.014 ± 0.028 –0.512 0.609

0.070 ± 0.028 2.560 0.011

0.150 ± 0.023 6.462 0.000

–0.003 ± 0.043 –0.062 0.951

0.026 ± 0.043 0.612 0.541

0.058 ± 0.036 1.609 0.109

–0.042 ± 0.045 –0.925 0.356

0.017 ± 0.045 0.388 0.698

0.109 ± 0.038 2.879 0.004

–0.080 ± 0.043 –1.847 0.066 –0.008 ± 0.042 –0.196 0.844

0.126 ± 0.036 3.522 0.001

CR EW

– – – – – – – – –

–0.180 ± 0.091 –1.975 0.049

–0.203 ± 0.090 –2.262 0.024

Adj–R2 = 0.0028

–0.028 ± 0.076 –0.367 0.714

Adj–R2 = 0.004

T P

Adj–R2 = 0.0041

T P

Migratory habit Intercept

0.598 ± 0.154 3.893 0.000

0.622 ± 0.154 4.034 0.000

0.049 ± 0.136 0.365 0.715

Migratory – – – – – – – – – Non–migratory

–0.008 ± 0.019 –0.44 0.660

0.020 ± 0.019 1.022 0.308

Adj–R2 = 0.0208

0.025 ± 0.017 1.471 0.142

Adj–R2 = 0.029

T P

Adj–R2 = 0.0032

T P

Forest dependence Intercept

0.557 ± 0.154 3.618 0.000

High

0.596 ± 0.154 3.860 0.000

0.073 ± 0.138 0.533 0.594

– – – – – – – – –

Medium

0.010 ± 0.036 0.292 0.770 –0.001 ± 0.036 –0.015 0.988

–0.046 ± 0.032 –1.421 0.156

Low

–0.008 ± 0.035 –0.245 0.807 –0.007 ± 0.035 –0.193 0.847

0.002 ± 0.031 0.089 0.928

Not occurring in forests

0.062 ± 0.036 1.716 0.087

0.072 ± 0.036 1.976 0.049

Adj–R2 = 0.046

–0.001 ± 0.032 –0.049 0.960

Adj–R2 = 0.075

T P

Adj–R2 = 0.008

T P

Habitat type 0.685 ± 0.148 4.620 0.000

0.083 ± 0.135 0.611 0.541

Terrestrial –0.080 ± 0.020 –3.846 0.000 –0.097 ± 0.020 –4.889 0.000

–0.034 ± 0.018 –1.862 0.063

Intercept

0.632 ± 0.150 4.227 0.000

Not restricted to the terrestrial environment

– – – – – – – – –

328

in the population–level patterns of diversity. Neither did we find significant differences between the mean genetic diversity values of Bubo bubo; populations that are declining in many parts of its range in Europe (Tucker and Heath, 1994) despite being categorized as LC in the European Red List. In general, studies evaluating the genetic diversity of European raptors are being encouraged as a strategy for the conservation of these birds (e.g., Martinez–Cruz, 2011). Czech populations of Tetrao urogallus showed larger values of HO and HE than the Spanish populations. The distribution of this species extends across most of Europe, but its distribution in the Iberian Peninsula is restricted to northern Spain. Unlike other populations living in pure– or mixed–conifer forests, the local Tetrao urogallus cantabricus only inhabits purely deciduous forests, and this specificity has put the population at risk (Storch et al., 2006). A study examining the genetic differentiation between this and other European populations showed that the birds from Cantabria form a clade with low genetic variability that differs from all other populations (Rodríguez–Muñoz et al., 2007) We also found differences between the average Ho and He values obtained for a Spanish population and an Austrian population of Cyanistes caeruleus, with the largest heterozygosity found in the Spanish population. This species is widely distributed throughout Europe and is native to both of the surveyed countries, but studies by Kvist et al. (1999, 2004), who analysed mitochondrial DNA sequences, proposed that regions in Europe were recolonized by this species from two different Pleistocene refugia after the last ice age through a colonization route from the Balkans to central and northern Europe as well as a route from the Iberian peninsula to the north and east. The differences we found between the mean genetic diversity values for the species in Spain and Austria can be explained by the possible isolation of the populations in Pleistocene refugia, which would allow different degrees of change in the populations. For Nipponia nippon, we found similar mean genetic diversity values between a Japanese and a Chinese population. The historical distribution of this species included the Russian Far East, China, and Japan, but it is now extinct throughout most of its range (IUCN category EN). Drastic reductions in populations of Nipponia nippon were caused by deforestation of nesting habitat, over–hunting and loss of wetlands as well as use of agrochemicals in rice fields, especially during the 1950s, which reduced the abundance and diversity of its preys (Li et al., 2009; Changqing, 2010). Overall, the observed genetic diversity was low in both populations, and the inbreeding coefficients were positive and high, reflecting the degree of risk to the populations of this species. Genetic diversity, ecological attributes and conservaion status As expected, the non–independence of sister clades in ecological attributes and conservation status (i.e., related species tend to have more similar ecological attributes than expected and therefore also conser-

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vation status) add a significant effect to the patterns of genetic diversity observed for birds. When the phylogenetic signal was taken into account, the pGLS confirmed that conservation status is reflected in the levels of genetic diversity sampled in birds; the lower the heterozygosity values the higher the risk of extinction. Consistently, the lowest averages of HO and HE were observed in birds of the EW category. Birds classified as threatened, possibly occurring with small and fragmented populations, tended to have lower heterozygosity. The risk of extinction is reportedly higher for small populations (e.g., Mace et al., 2008; Frankham, 2015) because they are more susceptible to genetic drift with accumulation of deleterious recessive alleles due to inbreeding (Hedrick and Garcia–Dorado, 2016), and loss of locally adapted traits (Frankham, 1995). Non–migratory species showed the lowest levels of HO and HE,confirming our hypothesis. The dispersal ability through flight routes contributes significantly to the increase of gene flow in migratory species (Losos et al., 2013) and consequently decreases the population structure in this group of birds. Paradis et al. (1998) reported that migratory bird species disperse more than resident birds. Dispersal is a fundamental component of metapopulations, gene flow, and genetic structure (Neigel and Avise, 1993), and is dependent on phylogeny (Paradis et al., 1998). However, our results lead us to question whether the greater diversity observed in migratory species is not simply a sampling bias, since the dispersion of these species facilitates their sampling. Gilroy et al. (2016), however, observed that populations of migratory birds showed higher intra–population variability (migratory diversity) and considered that they tended to decline less because they are more resistant to environmental changes. When we analysed the isolated effect of each variable, we found the lowest significant mean values of HE occurred in bird species that naturally occur in forests and are therefore highly dependent on forest ecosystems. This result evidences the importance of forests as places of shelter and breeding for birds (e.g., LaManna and Martin, 2016; Selwood et al., 2017; Giubbina et al., 2018), so that species that have no access to food or reproductive resources in these environments may experience negative effects. These results are important at a time when many of the major forest ecosystems worldwide —especially in the neotropics— are experiencing severe disturbances (Hansen et al., 2013) which directly affect bird species (e.g., Pereira et al., 2014). Ram et al. (2017) showed that forest birds have more positive tendencies than non–forest birds in face of climatic changes, suggesting that these species are positively affected by factors other than climate. Alternatively, loss of diversity in non–forest birds can perhaps be explained by high exposure to predation and hunting in such environments. Features such as landscape composition influence predation patterns at finer scales (e.g., Thompson et al., 2002; Stephens et al., 2003; Chiavacci et al., 2018). For example, the abundance of some common nest predators (e.g., Procyon lotor) tends to be higher in more intensive agricultural landscapes (Chalfoun et al., 2002).

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1.2

329

Regression 95 % CI 95 % PI

1.0 0.8 0.6 0.4 0.2 0.0 1.2 1.0

y = –0.1291x + 0.7013 R2 = 0.0416

Regression 95 % CI 95 % PI

0.8 0.6 0.4 0.2 0.0 y = –0.1395x + 0.75 R2 = 0.0471

0.5

1.5 Locos (log10)

2.5

Fig. 3. Correlation between the genetic diversity estimates (HO and HE) and the number of loci evaluated in the articles: CI, confidence interval; PI, prediction interval. Fig. 3. Correlación entre las estimaciones de la diversidad genética (HO y HE) y el número de loci evaluados en los artículos: CI, intervalo de confianza; PI, intervalo de predicción.

The categories and criteria developed by the IUCN have been important in designing conservation plans and strategies (Miller et al., 2007), but current efforts have been focused on species and on the global conservation plan. However, conservation actions should be implemented at the population level as extinction rates are estimated to be three to eight times higher than extinction rates of species (Hughes et al., 1997), and it is at this level that substantial losses of genetic diversity occur (Garner et al., 2005). Many bird populations have been identified as threatened (e.g., Alves et al., 2010; Van De Pol et al., 2010; Fernandes–Ferreira et al., 2012; Dunham and Grand, 2017; Yong et al., 2018), which implies the risk to further decrease the heterozygosity values reported for species outside of the LC category and expose their populations to risk of local extinction (Garcia and Marini, 2006). However, since both population size and

geographical range are two of the key criteria used by the IUCN to assign threat status, and as both are directly linked to heterozygosity, several studies assessing conservation priorities have excluded species with small populations and/or populations with narrow geographical ranges (Fisher et al., 2003; Jones et al., 2003). Therefore, it is important to include genetic diversity of populations when determining global conservation actions. The lowest mean HO and HE values were observed in Otidiformes, Accipitriformes, Gaviiformes and Ciconiiformes, which supports the hypothesis that the solitary behaviour of species of Otidiformes (previously included in Gruiformes but currently considered a proper order) and aspects of their reproductive behaviour, including monogamy and nesting at ground level, might be affecting their genetic diversity as species with complex social systems are more vulnerable to

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the effects of low population densities (Brito, 2009). For example, breeding pairs show low reproductive success unless a minimum number of helpers are present to aid in nest defence against predators and food provisioning for chicks (Brito et al., 2004). Conversely, the genetic patterns of Accipitriformes (traditionally included in the order Falconiformes but currently elevated to a proper order; Hackett et al., 2008) have been strongly affected by interspecific hybridization and anthropogenic disturbances (e.g., Poulakakis et al., 2008; Nam and Lee, 2009; Väli et al., 2010). Falconiformes species are mostly birds of prey with naturally low abundances, so given the lack of mating partners these species tend to hybridize, promoting population declines that put them under threat (Randler, 2006). Additionally, species with a high degree of habitat specialization and small clutch sizes are strongly associated with small population sizes, limited geographical ranges and, thus, higher extinction risk (Krüger and Radford, 2008). Ferrer and Negro (2004) have demonstrated that large predators, such as eagles and lynxes, highly specialised in certain prey species, and with small populations, are permanently threatened with extinction. Furthermore, birds in the orders Gaviiformes and Ciconiiformes have been affected by changes in land use and degradation of freshwater systems because of their high dependence on aquatic habitats (Arzel et al., 2015), vulnerability to pathogens (Silva et al., 2010), and exposure to pollutants derived from aquatic contamination (Fontenelle, 2006). There are strong examples in the literature suggesting that populations of many Ciconiiformes have undergone genetic bottlenecks as evidenced by the loss of genetic diversity and an increase in deleterious mutations (Li et al., 2014) due to inbreeding, climate change, habitat loss, hunting and environmental pollution, especially by agrochemicals (e.g., Zhang et al., 2004; Miño et al., 2009). Significant correlations were observed between the number of loci used and the estimates of HO and HE. Although the relationships were negative, they nevertheless suggest that low microsatellite locus sample sizes may bias diversity estimates. Small numbers of loci can be used only when evaluating a large number of individuals and when the mean heterozygosity of the population is high (Nei, 1978). Thus, given the number of individuals sampled, the studies we evaluated apparently failed to follow the basic requirements for the estimation of genetic diversity (i.e., they used small numbers of loci despite the mean heterozygosity values being high). However, the number of alleles per locus seems to be a good indicator of accuracy when assessing genetic distances with microsatellite markers. Kalinowski (2002) showed that good results can be achieved using few loci with several alleles or many loci with few alleles. In recent centuries bird species have been deteriorating in status and becoming extinct at a rate that may be 2–3 orders of magnitude higher than in pre–human times (Brooke et al., 2008). Relating genetic diversity estimates with IUCN Red List categories represents an attempt to understand the circumstances under which a bird species becomes

extinct, since it is possible to link these figures to high rates of inbreeding or reduced effective population size and gene flow. Brooke et al. (2008) showed that conservation actions have benefited species on the verge of extinction, but are less directed or have less effect on moderately endangered species. We are aware that the IUCN has specific guidelines to address genetic issues in reintroductions and translocations of species (IUCN/SSC, 2013), but as the status of birds has worsened worldwide with populations declining faster than ever, especially those of the Pacific marine species (BirdLife, 2013), studies on genetic diversity of bird populations should be promoted to identify populations at risk. Conclusion Studies including bird genetic diversity data obtained using microsatellite markers increased significantly between 2013 and 2014, reflecting the popularization of this technique during this period. However, most of these studies were conducted on Passeriformes and/ or taxa belonging to the least concern (LC) IUCN category, suggesting that sampling effort is an obstacle to the application of molecular techniques to study less abundant and/or threatened species. Our findings show that ecological attributes of bird species such as migratory habit, forest dependence and habitat type have a significant effect on genetic diversity parameters. More importantly, we corroborate our hypothesis that bird species classified under the most threatened IUCN categories (i.e. EW) have lower values of genetic diversity especially for HO and HE, whereas species classified under LC have higher values. This indicates that populations with high genetic diversity have a larger effective population size and therefore a lower extinction risk. From the perspective of conservation genetics, we believe that genetic diversity data should be incorporated and support current criteria for the IUCN Red List to generate a more complex and realistic picture of the conservation status of avian species. Acknowledgements We would like to thank the Fundação de Amparo à Pesquisa do Estado de Goiás – FAPEG and Conselho Nacional de Desenvolvimento Científico e Tecnológico – CNPq for the doctoral scholarships. We also want to thank the three anonimous reviewers for their comments that helped us to improve the paper. References Abouheif, E., 1999. A method for testing the assumption of phylogenetic independence in comparative data. Evolutionary Ecology Research, 1: 895–909. Allendorf, F. W., Luikart, G., Aitken, S. N., 2012. Conservation and the genetics of populations., 2nd

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Authorship decisions in ecology, evolution, organismal biology and natural resource management: who, why, and how G. D. Grossman, D. R. DeVries

Grossman, G. D., DeVries, D. R., 2019. Authorship decisions in ecology, evolution, organismal biology and natural resource management: who, why, and how. Animal Biodiversity and Conservation, 42.2: 337–346, Doi: https://doi.org/10.32800/abc.2019.42.0337 Abstract Authorship decisions in ecology, evolution, organismal biology and natural resource management: who, why, and how. Publication in peer–reviewed journals is essential for scientific progress including: (1) advancement of knowledge, (2) societal benefits including scientifically–based decision–making, (3) evaluation of researcher productivity, and (4) obtaining and retaining a research or faculty position and facilitating future scientific contributions. As science becomes increasingly complex so do the results necessary for publication, which frequently necessitates collaboration among scientists from multiple and diverse fields. Nevertheless, collaborative publication always includes the possibility of misunderstandings and differences of opinion. Here we first review the published literature on authorship determination for scientific publications in ecology, evolution, organismal biology and natural resource management, including consideration of what constitutes authorship, consideration of author contributions, author order in a byline, and power relationships, after which we provide several examples of realistic authorship conflict scenarios for purposes of pedagogy and discussion with colleagues and students. Key words: Ghost authorship, Gift authorship, Power differentials in science, Multiple–authorship, Sole authorship, Scientific writing Resumen Decisiones sobre autoría en ecología, evolución, biología de organismos y gestión de recursos naturales; quién, cómo y por qué. La publicación en revistas con revisión crítica es fundamental para lograr progresos científicos como: 1) la mejora del conocimiento, 2) beneficios sociales como la toma de decisiones basada en datos científicos, 3) la evaluación de la productividad de los investigadores, y 4) la obtención y el mantenimiento de un puesto de docente universitario, y la facilitación de futuras contribuciones científicas. A medida que la ciencia se vuelve más compleja, también lo hacen los resultados necesarios para la publicación, que frecuentemente requieren la colaboración de científicos de distintos campos. No obstante, la publicación colaborativa siempre incluye la posibilidad de que se produzcan malentendidos y diferentes opiniones. En este estudio, primero examinamos los artículos publicados sobre la determinación de la autoría de publicaciones científicas en materia de ecología, evolución, biología de los organismos y gestión de recursos naturales, y luego se estudió en qué consiste la autoría y se analizaron las contribuciones del autor, el orden de los autores en la línea de firma y las relaciones de poder, tras lo cual proporcionamos varios ejemplos de situaciones realistas de conflictos de autoría con fines pedagógicos y para entablar un debate con compañeros y estudiantes. Palabras clave: Autoría fantasma, Autoría honorífica, Diferencias de poder en ciencia, Autoría colectiva, Autoría individual, Redacción científica Received: 18 XII 18; Conditional acceptance: 22 II 19; Final acceptance: 17 IV 19 Gary D. Grossman, Warnell School of Forestry and Natural Resources, University of Georgia, Athens, GA 30606 U.S.A.– Dennis R. DeVries, School of Fisheries, Aquaculture and Aquatic Sciences, Auburn University, Auburn, AL 36849 U.S.A. Corresponding author: G. Grossman. E–mail: gdgrossman@gmail.com ISSN: 1578–665 X eISSN: 2014–928 X

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Introduction Scientific progress is dependent upon publication of research results. Publication in peer–reviewed journals is essential for multiple reasons including: (1) advancement of accurate knowledge, (2) societal benefits including scientifically–based decision–making, (3) evaluation of scholarly productivity, (4) obtaining and retaining a research or faculty position, and (5) facilitating continued productivity. Given the benefits and potential rewards of scientific publication, it is not surprising that a variety of contentious issues may arise when determining who should and should not be an author on peer–reviewed publications. This problem has been recognized previously (see references) and valuable general resources on authorship determination are available on both the Committee on Publication Ethics website https://publicationethics. org/authorship and the CRediT website https://www. casrai.org/credit.html. Nonetheless, the potential problems involved in authorship determination have become more challenging as the numbers of authors per paper continues to increase (exponentially in ecological journals from 1.4 co–authors/paper in 1950 to 4.8 co–authors/paper in 2015: Logan, 2016; for examples from other fields, see Erlen et al., 1997; Cronin, 2001; Cronin et al., 2003; Papatheodorou et al., 2008; Dotson et al., 2011) and have become even more important when one considers that significant monetary rewards (i.e., grants and salary increases) may be tied to publications (Abritis and McCook, 2017; Quan et al., 2017). Here we review published work on authorship issues from diverse fields, and summarize conclusions and recommendations for researchers in ecology, evolution, organismal biology, and natural resource management (herefore EEONR), where these issues are not well studied. Finally, to provide a locus for increased discussion, strategy derivation and improved resolution of these issues, we elucidate several actual authorship conflict scenarios and describe their outcomes. These scenarios are certainly not exhaustive, but represent potentially common situations faced by researchers in EEONR. An informal poll of members of the listserve Ecolog in December 2018 indicated that conflicts over authorship likely are not uncommon, which suggests that existing publications on the topic (mostly in the biomedical field) may not be well known, or have not effectively dealt with the issues specific to EEONR. A review of issues encountered in determining authorship on scientific papers At the broadest level, there are two distinct groups of authorship issues: (1) inclusion of an author who has not earned authorship (commonly referred to as 'honorary authorship' or 'gift authorship'), and (2) exclusion of someone who has earned authorship (commonly referred to as 'ghost authorship'). Honorary or gift authorship may result from the belief that inclusion of a more senior author will improve chances of manuscript acceptance, especially in a prestigious

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journal, or may occur when a person in a superior power position forces a subordinate to include their name on a paper, even though they have not made a substantive contribution to the research. It also may result from a 'payback' scenario where authors exchange authorships to increase their publication rates (Feeser and Simon, 2008). Honorary or gift authorship is inappropriate, because gift authors receive publication credit that is undeserved, because it violates contemporary ethical standards of science. In addition, honorary/gift authors typically cannot explain the paper's contents nor address post–publication issues, particularly those arising from methodological or interpretational errors. In contrast, ghost authorship may result from: (1) exclusion of someone who has since left a laboratory, (2) professional or personal disagreements leading to author exclusion, or (3) omissions to obscure potential conflicts of interest. Less commonly, ghost authorship results from use of a professional writer to compose a paper (particularly where an author may have biases that they do not want to be apparent). The prevalence of honorary and ghost authorship may be surprisingly large, particularly in biomedical fields, with percentages ranging up to 19 % for honorary authorship in medical journals (Flannigan et al., 1998; Wislar et al., 2011) to between 8–75 % for ghost authorship (8–11 % in medical journals: Flannigan et al., 1998; Wislar et al., 2011; 64 % in hospital clinical research: Pignatelli et al., 2005; 75 % in industry–initiated trials, most often due to statisticians: Gotzsche et al., 2007). As one would likely expect, differences exist among disciplines in the percentage of ghost authorship, being similar in biology (56 %) versus all disciplines combined (55 %), but interestingly were lower for graduate (15 %) and undergraduate (9 %) students serving as ghost authors in biology versus other disciplines (22 % and 13 %, respectively; Jabbehdari and Walsh, 2017). Criteria and considerations for inclusion/ exclusion of authors Although there is little published work on authorship disputes in EEONR, a variety of criteria have been proposed to determine authorship in the biomedical/ health sciences, including the Vancouver Guidelines (International Committee of Medical Journal Editors, ICMJE, 2017). The Guidelines include four criteria, all of which are required for coauthorship. These criteria are (ICMJE, 2017): (1) substantial contributions to the conception or design of the work; or the acquisition, analysis, or interpretation of data for the work; (2) drafting the work or revising it critically for important intellectual content; (3) final approval of the version to be published; and (4) agreement to be accountable for all aspects of the work in ensuring that questions related to the accuracy or integrity of any part of the work are appropriately investigated and resolved. The use of these guidelines varies across disciplines and a current list of journals that follow ICMJE recommendations can be found at http://www.icmje. org/journals–following–the–icmje–recommendations/.

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Despite the widespread use of the ICMJE guidelines, they are not immune from problems and salient criticisms. Perhaps the most significant issue with the guidelines is the requirement for authors to meet all four authorship criteria; in particular the broadly–worded accountability criterion. Not surprisingly, as the complexity of our studies continues to increase, the number of researchers/coauthors required to generate a scientific paper also has increased (Cronin 2001; Cronin et al. 2003; Dotson et al., 2011; Gorham and Kelly, 2014; Sugrue and Carroll, 2015; Logan, 2016; Barlow et al., 2018; but see Papatheodorou et al., 2008 for an alternate view). This requisite rise in the number of authors makes it difficult for every author to be accountable for all aspects of the work, especially those portions outside of their own areas of expertise. The alternative to a more thorough and complete multi–authored paper is a series of short individual papers with fewer authors that certainly would stand a higher overall chance of rejection, as well as making it more difficult for readers to integrate and synthesize what is likely a very complex story. As examples, we surveyed criteria for authorship posed by several relevant scientific societies and journals relevant to EEONR, and not surprisingly, these criteria vary (see other examples in COPE, 2014). For example, the American Institute of Biological Sciences restricts authorship to individuals who have met three criteria: "(1) made a significant contribution to the conception and design or the analysis and interpretation of data or other scholarly effort, (2) participated in drafting the article or reviewing and/ or revising it for content, and (3) approved the final version of the manuscript" (https://www.aibs.org/ bioscience/authors_and_reviewers.html). In addition, the senior author is responsible for: (1) ensuring that all potential authors that meet the criteria are offered co–authorship, (2) preventing those who do not meet the criteria from obtaining authorship, and (3) obtaining approval of the final version of the manuscript from all co–authors. Co–authors assume full responsibility for all work submitted under their names. The American Fisheries Society has dealt with the authorship problem by requiring that authorship be restricted to individuals "making a significant contribution to the work such as: determining or developing study objectives, designing experimental, statistical, or analytical approaches, collecting data, analyzing data and interpreting outcomes, and preparing the paper (organizing, writing, revising, and proofreading the text)" and that "Each author should make two or more significant contributions that produce new information (https://fisheries.org/books–journals/writing–tools/ authorship–guidelines/)". Similarly, guidelines from the Ecological Society of America's Code of Ethics state that "researchers will claim authorship of a paper only if they have made a substantial contribution. Authorship may legitimately be claimed if researchers (1) conceived the ideas or experimental design, (2) participated actively in execution of the study, (3) analyzed and interpreted the data, or (4) wrote the manuscript" (note the use of the conjunction 'or'; https://www.esa.org/about/

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code–of–ethics/). These examples illustrate the cross–discipline variation that exists in authorship requirements even when explicitly detailed. Certainly part of this variation may be attributed to cultural differences among scientific societies with respect to authorship (Penders, 2016; Moffatt, 2018). To explore the effectiveness of the widely used ICMJE authorship guidelines, we consider published information on their performance. Multiple studies have identified issues associated with adherence to these guidelines including: (1) authors' general lack of awareness or familiarity with the guidelines (Bhopal et al., 1997; Hoen et al., 1998), and (2) failure to follow the guidelines even when publishing in journals that adhere to them (Bhopal et al., 1997; Marusic et al., 2011; Fong and Wilhite, 2017). Just determining whether an individual meets the requirements of the ICMJE Guidelines differs when criteria are evaluated on a 'percent participation', e.g, ordinal (i.e., using a scale that ranged from 0 = none to 4 = full participation in a category) versus a binary (i.e., a 'yes' or 'no' response) basis (Ivanis et al., 2008), making authorship determinations even more problematic (see Guallar (2007) for an alternative scaling system). Adding to this criterion bricolage is the aforementioned fact that guidelines differ among professional societies and their journals (Osborne and Holland, 2009; Bosch et al., 2012; Bosnjak and Marusic, 2012); in fact, da Silva and Dobranszki (2016) state that "scientists have the inherent right to determine who is an author of an article according to the ethical guidelines of their institutes, but these may differ from the guidelines indicated by publishers, while editors and publishers have the right to verify authorship". Given the diversity of methodological practices in the sciences, clearly it is difficult to erect an all–encompassing scaffolding upon which various levels of authorship can be hung. However, the one consistent criterion among societies and journals is that authors must make a 'significant' or 'substantial' contribution to earn authorship. The difficulty arises in defining a 'significant contribution' and whether that contribution must be made in a single or multiple areas. Most published examinations of authorship guidelines, rules, or scoring approaches refer to the ICMJE Guidelines, but as previously mentioned, these uidelines yield very different outcomes if the criteria are linked by an 'or' rather than an 'and' or require some arbitrary minimum percentage of contribution. In addition, there are other contributions that researchers typically consider significant, including: (1) conception of the ideas, (2) substantive input on experimental design, (3) data analysis, (4) interpretation of the data or results, (5) writing the entire or substantial portions of the manuscript, (6) revision of the manuscript in response to reviewer comments, and (7) review and approval of the final draft before submission (Schmidt, 1987; Ahmed et al.. 1997; Osborne and Holland, 2009; Clement, 2014). Additional considerations include obtaining funding and the need for input from external specialists (Hunt, 1991; Benson and Silver, 2013). The CRediT guidelines even specify 14 potential areas of contribution to a scientific work (https://www.

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casrai.org/credit.html; see also Brand et al. 2015) in an attempt at an official or approved approach to generate contributor statements in those journals that require them (see below). Another approach to clarify the diverse ways in which individuals may contribute to a paper and perhaps help to determine inclusion as an author included a list of 25 research activities used by the National Institutes of Health (Patience et al., 2019). In many scientific fields, funding may be primarily obtained via extensive grant proposals (NSF, USDA, NIH) that require elucidation of the conceptual background and structure of the research, experimental design, and statistical analyses of the proposed research. Although securing funding alone would not guarantee authorship, if the research upon which the paper was based included elements included in the grant proposal (e.g., conceptual ideas, experimental design, statistical techniques, etc.), then securing funding itself could constitute a 'significant contribution'; however, this would need to be discussed and concluded by all co–authors. Assessment and explicit descriptions of author's contributions Several authors (Schmidt, 1987; Ahmed et al., 1997; Guallar, 2007; Clement, 2014) advocate the use of assessment systems that presumably lead to more objective decisions regarding authorship. Assessment occurs via assignation of percentages or weights for the tasks associated with the completion and publication of a research study in combination with some minimum value established for authorship (e.g., Schmidt, 1987; Hunt, 1991; Galindo–Leal, 1996; Guallar, 2007). Typically, contributions for each potential author are assigned numerical scores (1–5 scale, Ahmed et al., 1997), and a matrix of tasks and contributions are then constructed for the project (e.g., Guallar, 2007; Clement, 2014; Roberts, 2017). Despite the desire for an objective approach to determining authorship, the instructions for authors or codes of ethics for few journals in organismal EEONR currently incorporate these approaches. In addition, the degree of objectivity of scoring approaches may be unclear; Ilakovac et al. (2007) found that survey responses regarding research contributions from corresponding authors differed through time, and also differed from their co–authors, documenting the presence of a high degree of subjectivity among coauthors. One example of this phenomenon is 'autobiographical memory', the subjectivity inherent in the memory and representation of what authors recalled about their own contributions. These phenomena might readily confound individual contribution scores. In an effort to increase author accountability, some journals now require that authors provide statements detailing their respective contributions to the publication (e.g., Lundberg and Flanagin, 1989), including: (1) performance of the research, (2) data analysis, (3) writing, (4) manuscript submission, and (5) revision of the paper (Weltzin et al., 2006; Feeser and Simon, 2008). As noted earlier, one group (CRediT;

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https://www.casrai.org/credit.html) has specified 14 potential areas of contribution that could be used in such contribution statements. To date, however, the inclusion of contribution statements has not been widely implemented, and still suffers from subjective memory issues both within and among individuals (Ilakovac et al., 2007) and from authors failing to draft contribution statements (Sauermann and Haeussler, 2017). In addition, it appears that responses may be influenced by the structure of the questions and form itself that is used to gather the information (Marusic et al., 2006). Consequently, this approach may not provide desired benefits for determining or justifying choice of authorship; however, enhanced communication among journals and/or publishers could lead to a more standardized approach to contribution statements that could be more broadly adopted and provide for more effective application of this approach (Sauermann and Haeussler 2017). The general approach of requiring some form of contribution statement at manuscript submission holds promise for the future, if for no other reason but that it requires research groups to discuss issues surrounding authorship. The order of authors and power–relationships Two additional issues associated with authorship determinations are: (1) decisions regarding the order of authorship, and (2) the role of differential power relationships (e.g., faculty, researchers, staff and students) in authorship decision–making. Given that the authorship inclusion or exclusion process is complex and full of subjectivity, one can only imagine the complexities of gaining author agreement on the order of all authors in a byline! Certainly some of the complexity in author order decisions arises because of different interpretations of the role and/or importance of the first author, the corresponding author, and the last author in a byline (Wren et al., 2007). Except for the case where a journal requires authors to be listed in alphabetical order (a practice that comes with its own set of consequences, such as reduced recognition of authors whose names occur later in the alphabet, Weber, 2018), there is general agreement that the first author of a paper is the person who contributed the most to a project, including production of the manuscript (Kempers, 2002; Tscharntke et al., 2007; Strange, 2008; Duffy, 2017; Logan et al., 2017; Tarkang et al., 2017). Nonetheless, first authors are not always corresponding authors (Duffy, 2017), especially in cases where a current or former student is the first author. As one might expect, the perceived importance of the corresponding author varies among individuals and disciplines. In addition, in many cases the position of last author is significant because it represents the person in whose lab the work has been completed. But sometimes the order of authors merely represents the declining order of contributions to the work (Feeser and Simon, 2008; Mulligan et al., 2014; Marusic et al., 2011). A survey of ecologists indicated some support for the last author being perceived as the most important author (Duffy, 2017), which likely

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reflects the view that this position typically represents the most senior author. Nonetheless, this response was not universal (Duffy, 2017) and is similarly not consistent across other disciplines (Kempers, 2002; Tscharntke et al., 2007; IWCSA, 2012; Logan et al., 2017; reviewed in Marusic et al., 2011). The ICMJE guidelines simply state that the order of authors should be a 'joint decision of the co–authors'. Clearly one effective mechanism for reducing authorship conflicts involves early discussions of authorship expectation among individuals with different levels of power such as students and their faculty advisors, or researchers and their staffs (Heffner, 1979; Guallar, 2007). Because students are typically relatively inexperienced in research and publication compared to their mentors, students may be at a disadvantage regarding decisions concerning author inclusion and order (Kwok, 2005; Maursic et al., 2011). This makes them potentially vulnerable to exploitation by superiors (Oberlander and Spencer, 2006). In addition, given that students are undergoing training, it may be unclear whether research tasks unrelated to their theses represent training or work warranting authorship. The field of psychology has published guidelines regarding authorship decisions involving dissertations (Costa and Gatz, 1992) which can be found in the current American Psychological Association's Ethical Principles of Psychologists and Code of Conduct. These guidelines state that a multi–authored paper based on a student's doctoral dissertation should list the student as senior author (http://www.apa.org/ethics/code/), which certainly is a good general rule. Nonetheless, this code does not address other student–supervisor situations, such as Master's theses or undergraduate students (Burks and Chumchal, 2009). Other fields and disciplines might consider following the APA's example in adopting policies that protect the rights of students in the publication process, as well as suggesting that individual institutions consider drafting their own guidelines for student co–authors. At a broad level, Fine and Kurdek (1993) provide three ethical principles for dealing with authorship in faculty–student projects: (1) beneficence (i.e., abstaining from injuring others, helping others further their important and legitimate interests by preventing or removing possible harms), (2) justice (the ethical duty to treat others fairly and to give them what they deserve), and (3) parentalism (treatment that restricts the liberty of individuals without their consent where the justification is either the prevention of some harm they might do to themselves or the production of some benefit they might not otherwise secure). These three principles should be considered by faculty/mentors (and perhaps more importantly by professional societies and universities) when working with student co–authors. There are many more authorship situations involving power imbalances, such as those between supervising administrators (e.g., Research Unit Heads, Deans or Department Heads) and more junior researchers/faculty, between untenured and tenured faculty, and between faculty and staff. To our knowledge, these cases have not been addressed in

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the published literature. Probably the most common of these power imbalance situations is between faculty and paid laboratory staff. Many researchers feel that laboratory staff perform research duties as part of their paid responsibilities, and hence, should not be considered coauthors. In our own labs, we use the principle that authorship is warranted when a staff member assumes a higher–level research role such as conceptualizing research projects, designing and performing experiments, and writing significant amounts of a manuscript. Nonetheless, clearly these are subjective decisions. Similarly, if a graduate student's stipend requires performance of research duties unrelated to their own graduate work, at what point do they earn authorship on papers resulting from their performance of these mandatory, yet non–thesis, duties? It seems logical that performance of higher– level duties to earn co–authorship, apply here as well. As with all creative ventures involving collaboration, discussions regarding research responsibilities and authorship should take place before the research project is started (Oberlander and Spencer, 2006; Guallar 2007). One approach is to generate a written plan complete with specification of who is responsible for each task (e.g., table 1), complete with temporal benchmarks to ensure that the project progresses as planned. Other examples of such systems are available on CRediT (https://www.casrai.org/credit. html) and Guallar (2007). The written plan should be evaluated by the research group quarterly or semiannually, and updated as needed. This should clarify research expectations and performance and facilitate early identification of performance problems. In addition to agreeing on satisfactory performance, participants should discuss potential consequences for unsatisfactory performance, including loss of authorship. If after multiple attempts to rectify and resolve a problem the responsible party(ies) decide to remove a coauthor, that collaborator should be notified in writing with confirmation of receipt of the message. Such steps will minimize future misunderstandings. Examples of authorship disputes and problematic situations Perhaps the most important point we would like to make is that members of a research group need to agree on standards both before and during a research project, and that these discussions need to incorporate or at least acknowledge differences in power. To illustrate potential problems that may be encountered as a research program progresses or even after fieldwork and analyses are completed, we provide three scenarios below, and discuss possible paths to their subsequent resolution or lack thereof. We describe these situations with two resolution perspectives in mind: 1) the strictly 'ethical' perspective in which a 'contract' is broken, and 2) the 'conflict resolutio' perspective in which the goal is not necessarily to employ a strictly ethical solution (about which there might be great differences of opinion) but rather to reach a mutually agreed upon solution to a 'broken

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contract' (Beer and Packard, 2012; Moore, 2014). The latter perspective allows greater latitude in dealing with mistakes, such as those that might be made by inexperienced researchers, in a non–punitive manner. These scenarios are real–world situations and their discussion has stimulated fruitful deliberations in GDG's graduate classes. Although we discuss resolution of these cases from both perspectives, our main purpose is to provide examples of authorship conflict scenarios for classroom and research group discussion. We recognize that these examples are presented from our perspective alone and that the circumstances of many authorship conflicts are too idiosyncratic and subjective to suggest that a single 'correct' resolution is possible. Scenario one A new PhD (first a Post–Doc, then an Assistant Professor) is conducting a poorly funded post–doctoral research program in community ecology when they are approached by a student who has just graduated with a bachelor's degree. The bachelor–level graduate (Volunteer) plans to remain in the locale and asks to volunteer on the project for 'experience' and as a potential springboard for graduate school. The Post–Doc can use the extra help and is also interested in maximizing the educational benefits for the Volunteer. Little is known about the organisms in this system so even basic biological information is likely publishable. The Post–Doc and the Volunteer hold several discussions regarding the project and reach an agreement that if the Volunteer fulfills a series of responsibilities that are defined and deemed reasonable by both individuals, they will be able to use a subset of specimens for a publishable study. The responsibilities are as follows: (1) the Volunteer will help on all collecting trips, (2) the Volunteer will learn how to perform the required lab analyses with the help of the Post–Doc, (3) the Volunteer will conduct the lab work, analyze the data, and perform statistical analyses with the aid of the Post–Doc, and (4) the Volunteer will write the first draft of a publishable manuscript on the species of interest, again with the help of the Post–Doc. The Post–Doc agrees to help with all aspects of the Volunteer's project, but the Volunteer must perform most of the work and therefore, will assume senior authorship on the manuscript. Without any discussion with the Post–Doc, the Volunteer abruptly goes overseas (for several years) before 50 % of the samples are collected. The Volunteer does no subsequent work on the project and the Post–Doc completes sampling with other team members. Two and a half years later, the Volunteer contacts the Post–Doc (now an Assistant Professor at a geographically distant university) and asks for 'their data'. However the Assistant Professor, having heard nothing from the Volunteer in the interim, has completed the research work with a different collaborator and a manuscript already has been submitted for publication. The Assistant Professor has included the Volunteer in the Acknowledgement section of the manuscript but given that the Volunteer did not satisfy

the terms of the research agreement, they were not included as a coauthor. Unfortunately, the Assistant Professor did not contact the Volunteer to discuss the decision, although the Volunteer had not provided any contact information and was overseas at the time. The Volunteer then contacted the Assistant Professor's former advisor claiming they have been cheated out of authorship on the project. After discussions with the Assistant Professor's former advisor (at this point the Volunteer refuses to communicate with the Assistant Professor), the following potential solutions are proposed by the various parties (certainly other possibilities exist and should be fodder for discussion by those using this paper in a teaching context): (1) the Volunteer should be given senior authorship on the submitted manuscript, (2) the Volunteer should be added as a co–author on the submitted manuscript, but then has to complete subsequent work to produce sufficient data for a second publishable manuscript; which they will write (with the help of the Assistant Professor) and be senior author, and (3) the situation will be left as it currently stands with the Volunteer's help acknowledged, and the Assistant Professor and their collaborator as senior and junior authors, respectively. From a strictly ethical perspective we believe that Solution 3 is the correct outcome, given that the Volunteer clearly failed to meet the responsibilities specified in the original agreement. However, from a conflict resolution perspective the Volunteer should interact with the Assistant Professor (possibly including the Assistant Professor's former advisor as a sort of 'mediator') and work out a mutually agreeable plan to receive some level of recognition/compensation. Solution 2 (and other potential alternatives) would satisfy this perspective, and provide an option to senior author a different paper. Nonetheless, the Volunteer refused to work to resolve the issue (or even communicate) with the Assistant Professor so the ultimate solution was Solution 3. Although in this case, 'ownership' of the data is relatively clear, in other situations it may not be (i.e., it could be owned by the PI, by the PI's institution, by the funding agency, or by combinations of individuals and organizations). Scenario two A Professor who is an editor for a prominent scientific journal would like to give a promising graduate student (henceforth Student) experience in reviewing scientific manuscripts. Over several months, the Professor gives the Student several manuscripts from the journal to review. For each review, the Professor assesses and edits the drafts of the review and then meets with the Student to discuss how the text should be modified to provide more appropriate and constructive criticisms for the authors. Each review assessment by the Professor involves several hours of their time not including the meetings with the Student. By the end of the 'tutorial' the Student has become a very good reviewer. Approximately six months later, and without telling the Professor, the Student submits and publishes a short article in a national journal on

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Table 1. A potential chart for assigning responsibilities and assessing authorship. Tabla 1. Una posible tabla para asignar responsabilidades y evaluar la autoría. Aspect

Author 1 % contribution

Author 2 % contribution

Author 3 % contribution

Author 4 % contribution

Research idea Design Data collection Data analysis interpretation Writing/revising Financial Other

'how to review a paper'. The article contains many pieces of advice regarding manuscript reviewing that the Student obtained from the Professor, but the Professor is not a coauthor nor even mentioned in the acknowledgements. From a strictly ethical perspective it is clear that the student should have coauthored the paper with the Professor, however, that is no longer an option. Consequently, the student should publish an addendum clarifying that much of the information in the article came from the Professor. From a conflict resolution standpoint, a number of possibilities exist, although none are completely satisfactory given that the article already is published and it is unlikely that a second article on the subject would be publishable. Nonetheless, the participants could agree to work on a different coauthored paper as a form of compensation as per Solution 2 in the previous scenario. Clearly this is an imperfect solution but as per current mediation practices it would at least represent a mutually agreed upon resolution. In addition, some organizations have advocated procedures where wronged authors might be added post–publication (Committee on Publication Ethics: https://publicationethics.org/ and https://www. casrai.org/credit.html). Scenario three A graduate student completes a strong PhD working on a problem that is part of a larger research program run for many years by his Professor. The Professor has provided some salary support (most support was from a fellowship won by the Student) for the former Student as well as providing expensive specialized equipment required for the study. The PhD leaves, and now five years later is employed in a professional position. Nevertheless, the PhD has not published any papers from their dissertation. Given the time that has passed, the manuscripts will require substantial editing for publication, including an updated literature

review. Although the Professor has repeatedly tried to contact the PhD, the former student has never responded. In discussion with his colleagues, the Professor suggests the following solutions: (1) the Professor edits and updates the manuscripts and is junior author. He then sends them to the PhD for review, with a time constraint stating that the manuscripts will be submitted if there is no further contact by a reasonable deadline, (2) the same constraints listed in Solution 1, but the Professor edits and updates the manuscripts and assumes the role of senior author; the PhD is moved to junior author, (3) the manuscripts are left unpublished, harming the both the PhD's and the Professor's career, and inhibiting the Professor's ability to obtain future funding. It is unclear whether there are any purely ethically–based solutions to this case; however, the most charitable would be for the Professor to employ Solution 1. Ideally, the PhD would be an active participant in the process and thus would satisfy the requirements of a mutually–agreed conflict–resolution solution as well. Depending on the amount of work required for revision, and the origin of the ideas in the manuscript, Solution 2 might be ethically appropriate as well. Both authors have experienced this situation. In one case, Solution 3 was what actually occurred, but only because the Professor was not interested in employing Solution 2 for concern regarding potential personal and professional conflicts. And in the other case, Solution 1 was what actually occurred, with minimal response from the former graduate student. Some faculty have tried to avoid these situations by requiring students to sign informal contracts stating that after a certain amount of time, say three years after graduation, the advisor has the right to update and submit the manuscript as first author. One of us (GDG) has used such contracts and found they do remove any ambiguity about the advisor's right to submit the manuscripts after a reasonable amount of

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time has passed. Nonetheless, GDG did not experience any increase in manuscript production by former students that could be attributed to use of a contract, nor increased resolution of who should update a manuscript and be senior author. Depending on the student, such an approach could appear unduly formal and legalistic, particularly given the power–structure difference between faculty and student, so such a practice must be approached carefully. Conclusions Peer–reviewed scientific publication is, and likely will continue to be, the standard mode for communicating important research results to research scientists, managers, teachers, and policy–makers, as well as an important metric by which scientific professionals are evaluated. Nonetheless, collaborative ventures are always open to misunderstandings and differences of opinion, and the increasing complexity of collaborative work only makes these issues more problematical. In our review of the literature, we described criteria for inclusion and ordering of individuals as authors for scientific publications, as well as techniques used by journals and publishers to better account for authorship in their published papers. In addition, for pedagogical purposes and discussion, we provided several realistic conflict scenarios. There are many valid opinions regarding how to decide authorship, some of which conflict, and praxis does not always follow the norms of a field. Some summary points and recommendations from our paper include: 1. Despite the importance of publication and authorship across all fields, the literature concerning authorship issues are contained in diverse and widely distributed publications, some of which may not be easily accessible. Consequently, there is a need for increased awareness of these issues among scientists and potential authors, as well as professional societies, journals, publishers, and employers. In fact, that is the purpose of our review of the literature. 2. Journals, publishers, and professional societies need to consider whether contribution statements may be used effectively, and if so, work to standardize their use to ensure they become a more effective tool for describing the roles of authors included in a byline (e.g., as suggested by CRediT). 3. Employers (e.g., agencies, universities) should also develop criteria for authorship decisions, perhaps involving standardization and use of contribution statements by students and employees. 4. Discussion of authorship issues needs to take place in both informal lab group settings and with professional societies, employers, etc. The scenarios we provide are intended to contribute to this pedagogical need. 5. Discussion among individuals should take place before a study is initiated and continued throughout the study so that all who should be considered as an author are considered, and no one is included as an author who has not satisfied the group's agreed–upon requirements for authorship.

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Spatial segregation between Iberian lynx and other carnivores G. Garrote, R. Pérez de Ayala

Garrote, G., Pérez de Ayala, R., 2019. Spatial segregation between Iberian lynx and other carnivores. Animal Biodiversity and Conservation, 42.2: 347–354, Doi: https://doi.org/10.32800/abc.2019.42.0347 Abstract Spatial segregation between Iberian lynx and other carnivores. The Iberian lynx (Lynx pardinus) is a specialist predator. Rabbits represent the bulk of its diet as for many other Iberian predators. This study addresses how the presence of the Iberian lynx affects the spatial distribution of the mesocarnivore community at landscape scale in the Sierra de Andújar. We studied mesocarnivore presence by sampling at 230 camera trapping stations, located in areas with and without lynx. We used a x2–test to compare the proportion of stations in which each species of carnivore were recorded in the zones with and without lynx. The proportion of camera trapping stations in which red fox (Vulpes vulpes), Egyptian mongoose (Herpestes ichneumon), beech marten (Martes foina), wildcat (Felis sylvestris) and common genet (Genetta genetta) were detected was significantly lower in the area where lynx were present than in the area where it was absent. No significant differences between the two types of areas were found for badgers (Meles meles). Our results highlight the role of the lynx as apex predators and the benefits that the recovery of Iberian lynx populations would entail in terms of trophic interactions and restored disrupted ecosystems processes. Key words: Intraguild competition, Carnivores, Phototrapping, Apex predator Resumen Segregación espacial entre el lince ibérico y otros carnívoros. El lince ibérico (Lynx pardinus) es un depredador especialista. El conejo constituye el grueso de su dieta, al igual que la de otros depredadores ibéricos. Este estudio analiza cómo la presencia del lince ibérico afecta a la distribución espacial de la comunidad de mesocarnívoros a escala de paisaje en la sierra de Andújar. Se estudió la presencia de mesocarnívoros mediante 230 cámaras de fototrampeo, instaladas en zonas con y sin presencia de lince. Se utilizó la prueba de la x2 para comparar la proporción de cámaras en las que se detectó cada una de las especies de carnívoros en las zonas con y sin lince. La proporción de cámaras que detectaron zorros (Vulpes vulpes), meloncillos (Herpestes ichneumon), garduñas (Martes foina), gatos monteses (Felis sylvestris) y ginetas (Genetta genetta) fue significativamente menor en las zonas con presencia de lince que en las zonas donde este estaba ausente. No se encontraron diferencias significativas en cuanto a la presencia de tejones (Meles meles) entre ambos tipos de zona. Nuestros resultados ponen de relieve la importancia del lince como depredador apical y los beneficios que podría reportar la recuperación de las poblaciones de lince ibérico en lo que concierne a las interacciones tróficas y el restablecimiento de los procesos ecosistémicos interrumpidos. Palabras clave: Competencia intragremial, Carnívoros, Fototrampeo, Depredador apical Received: 08 XI 18; Conditional acceptance: 05 II 19; Final acceptance: 16 V 19 Germán Garrote, Instituto de Biología de la Conservación (IBiCo), c/ Nebli 13, 28232, Madrid, España (Spain).– R. Pérez de Ayala, WWF/España, Gran Vía de San Francisco 8–D, 28005 Madrid, España (Spain). Corresponding author: G. Garrote. E–mail: gergarrote@gmail.com

ISSN: 1578–665 X eISSN: 2014–928 X

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Introduction Direct interactions between predators and other species can have indirect consequences further down the food web via trophic cascades (Ripple et al., 2016). Large carnivores play a key role in terrestrial ecosystems when they exert an influence on herbivores and so indirectly prevent overgrazing (McShea, 2005). They can also influence carnivore communities via intraguild interactions (Ritchie and Johnson, 2009) and indirectly prevent excessive predation on prey species by mesocarnivores (Elmhagen et al., 2010). This top–down cascade can influence ecosystem structures and biodiversity at both local and larger scales (Terborgh, 2001; Elmhagen et al., 2010). If healthy populations of top predators are to be maintained within ecosystems, these ecosystems should also contain healthy communities and populations of the many species that perform ecosystem services at lower trophic levels (Dobson et al., 2006; Haswell et al., 2017). However, the functional roles of top predators cannot be fully appreciated in isolation from bottom–up processes because the effects of nutrients, productivity (Pace et al., 1999) and anthropogenic habitat may bring about change (Litvaitis and Villafuerte, 1996; Estes, 1998; Elmhagen and Rushton, 2007). Competitive intraguild interactions have been proposed as highly important organizing mechanisms since, due to similarities in ecological niches, they limit the number of species that can be packed into an assemblage (Jaksic and Marone, 2007). Similar ecological preferences increase the risk of competition, whereas mechanisms such as resource partitioning, temporal or spatial avoidance strategies (Voigt and Earle, 1983; Johnson and Franklin, 1994; Kozlowski et al., 2008), or alternative foraging strategies (Husseman et al., 2003) facilitate coexistence. Interference interactions, harassment and injury caused by larger carnivores pose a risk to smaller mesopredators (Linnell and Strand, 2000; Haswell et al., 2018). Furthermore, as a result of interference competition, subordinate species are frequently restricted to suboptimal habitats (Tannerfeldt et al., 2002; Macdonald et al., 2004; Mitchell and Banks, 2005), which can have important implications for the demography and distribution of the species involved (Thompson, 1988; Holt and Polis, 1997; Atwood and Gese, 2008). The Iberian lynx (Lynx pardinus) is the top predator of the terrestrial vertebrate community in the Mediterranean ecosystem (Valverde, 1963). Listed as Endangered by the IUCN (Rodríguez and Calzada, 2015), the species reached its all–time minimum in the first years of the twenty–first century, when only 100 individuals in just two isolated populations –Andújar– Cardeña and Doñana– were known to exist (Guzmán et al., 2004; Simón et al., 2012). Since then, however, the Iberian lynx has undergone a significant increase in population size and range due to the measures implemented as part of conservation projects for the species (Simón et al., 2012), which include the creation of new populations through reintroduction. The Iberian lynx is a specialist predator. Rabbits represent the bulk of its diet in a similar manner to

that of many other Iberian predators (Cabezas–Díaz et al., 2011), possibly leading to interference or food competition. Previous studies of the relationships between Iberian lynx and other carnivores performed in Doñana have found that the Egyptian mongoose (Herpestes ichneumon) and genet (Genetta genetta) avoid lynx, while the Eurasian badger (Meles meles) is apparently indifferent to its presence. Although foxes (Vulpes vulpes) and lynx exhibit temporal segregation in their use of habitat (Fedriani et al., 1999), their spatial relationship remains unclear (Palomares et al., 1996). The relationship between wildcat (Felix sylvestris) and lynx has not been studied. This study addresses how the presence of the Iberian lynx affects the spatial distribution of the mesocarnivore community at a landscape scale in the Sierra de Andújar. We studied the spatial distribution of several species of mesocarnivores in areas where the lynx is absent and where it is present, taking into account the abundance of rabbits. Material and methods Study area The study area lies in the eastern Sierra Morena (SE Spain; fig. 1) and consists of a mountainous area with an altitudinal range of 200–1,500 m covered by well–preserved Mediterranean forests (Quercus ilex, Q. faginea and Q. suber) and scrublands (Quercus coccifera, Pistacia lentiscus, Arbutus unedo, Phillyrea angustifolia and Myrtus communis). The area is managed for big–game hunting and has high densities of red deer (Cervus elaphus) and wild boar (Sus scrofa). It is partially protected by the Parque Natural Sierra de Andújar. During the study period, the Andújar–Cardeña Iberian lynx population consisted of 60–110 individuals, distributed over an area of 15,000 ha (Guzmán et al., 2004). Camera trapping survey The spatial distribution of the carnivore community was estimated by sequential camera trapping surveys performed in December 1999–February 2000, November 2000–February 2001 and November 2001–February 2002. We used camera trapping data from the annual national Iberian lynx survey (Guzmán et al., 2004), which covers 85 % of the area potentially used by the Iberian lynx. We divided the study area into 12 survey blocks, each of which were surveyed by camera trapping for periods of two months. Once one block was finished, cameras were moved to the next survey block. We surveyed an almost continuous surface area of 7,800 ha using a total of 230 camera trapping stations (1999/2000: n = 28; 2000/2001: n = 168; 2001/2002: n = 39). In all, 115 out of 230 stations were located in areas in which the lynx are present, as defined by Guzman et al. (2004), and the other 115 stations were placed in areas without lynx (fig. 1).

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Spain

Andújar–Cardeña Doñana

Camera trap 2002 Iberian lynx distribution area

4 km

Felis sylvestris

Meles meles

Vulpes vulpes

Martes foina

Genetta genetta

Herpestes ichneumon

Fig. 1. Study area map. Camera trap stations located in areas with and without lynx, and stations in which each species of carnivore was recorded. Fig. 1. Mapa de la zona de estudio donde se representa la ubicación de las estaciones de fototrampeo en zonas con y sin lince, así como las estaciones donde se detectó cada una de las especies de carnívoros.

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We used 212 35–mm Canon Prima© classic photo film cameras with data registers and automatic flashes. The cameras were modified to allow activation via an external 25 × 25 cm pressure plate, positioned at a distance of 170 cm that was triggered when stepped on by an animal (Garrote et al., 2011). The cameras were placed in a small wooden box on pillars 30 cm above ground level. Urine from captive Iberian lynx, placed on an inert adjacent support, 50 cm above ground level and the pressure plate, was used as a lure. Lynx urine has been reported to be an excellent attractant for all carnivore species (Garrote et al., 2011; Monterroso et al., 2016). This attractant was replaced every 3–6 days. The distance between camera traps was 400–800 m. Camera–trap locations were located along suspected lynx travel routes (Garrote et al., 2012) such as roads or paths, chosen to maximize capture probabilities (Karanth and Nichols, 1998). Each camera was continuously active throughout the entire survey period for each block (two months). To describe the species distribution in the area, we calculated occupancy as the proportion of stations at which a species was detected in relation to the total number of stations (Sogbohossou et al., 2018). Rabbit abundance and habitat variables Rabbit abundances were estimated for each survey block by on–foot constant–speed itineraries lasting three hours. Rabbit latrines were counted every 15', and these counts were taken as the survey unit for the statistical analysis. Indirect surveys were carried out at the same time of the year (end of spring, when rabbit populations peak) under similar weather conditions. Every 15' we estimated, in a 25 m radius plot, the percentage of land surface covered by the following habitat categories: trees, scrubland lower than 50 cm in height, scrubland higher than 50 cm in height, pastureland and rocks. The percentage of covered land was divided into four categories scored as follows: 1 (0–25 %), 2 (> 25–50 %), 3 (> 50–75 %) and 4 (> 75). Statistical analysis We compared the mean values for rabbit abundance and for each habitat category obtained in the areas with and without lynx using a Mann–Whitney U–test. We used a x2–test to compare the proportion of stations in the zones with and without lynx in which each species of mesocarnivore was present. The carnivores with lower capture rates were grouped together to perform statistical analysis (minimum five expected records). Results The following carnivores were detected in this study: (Lynx pardinus, 9–15.9 kg), Eurasian badger (Meles meles), red fox (Vulpes vulpes), Egyptian mongoose (Herpestes ichneumon), beech marten (Martes foina), wildcat (Felis sylvestris), and common genet

(Genetta genetta). The proportion of camera trapping stations in which the fox and wildcat were detected was significantly lower in the area with lynx than in the area without lynx (table 1; fig. 1); no significant differences were found for the presence of the badger between both areas. Genet, beech marten and Egyptian mongoose were grouped together to perform the statistical analysis. The presence of this group of mesocarnivores was found to be significantly lower in the areas where lynx were present. No significant difference was found between zones with and without lynx for the habitat variables (table 2). As expected, rabbit abundance in areas with lynx was significantly higher than that in lynx–free areas since lynx distribution is dependent on rabbit abundance (table 2). Discussion With the exception of the badger, the presence of the Iberian lynx determines the distribution at the landscape scale of the mesocarnivores community in the study areas. No significant habitat differences were found between areas with and without lynx, while the highest rabbit abundances were detected in areas with lynx. As mentioned above, Iberian mesocarnivores preferably select rabbits as prey (Cabezas–Díaz et al., 2011). The most probable explanation for the observed distribution of mesocarnivores at a landscape scale is the interference competition between species in which the lynx is the dominant species. This is the first study to address a relationship between the Iberian lynx and wildcat, the only two sympatric wild felids present in the Iberian peninsula. Competition becomes greater as eco–morphological similarities or phylogenetic proximity between competing species increase (Cruz et al., 2018), and generally the larger dominant species exclude smaller or subordinate species from their territories by interference competition. Therefore, as expected, the larger Iberian lynx exerts strong interference competition on the smaller wildcat. This leads to fewer wildcats in those areas where lynx are present. Similar relationships of dominance have been described for other species of felines, such as the ocelot (Leopardus pardalis), which acts as a dominant carnivore over other smaller sympatric cats such as margay (Leopardus wiedii) and jaguarundi (Puma yagouaroundi) and so influences their ecological parameters (de Oliveira et al., 2010; Cruz et al., 2018). Previous studies have shown a high overlap in the diets, activity levels, habitat use and home range in radio–tracked foxes and lynx (Fedriani et al., 1999). Although it has been suggested that foxes mitigate lynx predation by modifying their spatial behaviour at home range level, no spatial segregation in these species has ever been found. Using a landscape approach, the present study demonstrates significant spatial segregation between foxes and lynx. These differences with previous work might be attributable to scale since certain studies have concluded that approaches at different scales can generate different

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Table 1. Total number of camera stations, positive stations for each species in zones with/without lynx, and positive stations per species. Genet, beech marten and Egyptian mongoose are grouped in 'others'. x2 results are shown. Tabla 1. Número total de estaciones de fototrampeo, número de estaciones positivas para cada especie en las zonas con y sin lince y estaciones totales positivas para cada especie. Las ginetas, las garduñas y los meloncillos están agrupados en la categoría "Others" (otras). Se muestran los resultados de las pruebas de la x2.

Total Badger Fox Wildcat Others

With lynx

115

Without lynx

115

Total

230 43 63 35 19

0.5

conclusions regarding interspecific interactions between species (e.g. (Tannerfeldt et al., 2002) for the Arctic red fox (Cruz et al., 2018). Previous studies (Palomares et al., 1996; Fedriani et al., 1999) have covered smaller areas than our study, which was performed at a much greater landscape scale. On the other hand, the relative densities of the mesocarnivores and their prey may also influence interactions (Creel, 2001; Berger and Gese, 2007). However, although no information is available for fox densities to compare these two study areas, the density of Matasgordas rabbit population (8 rabbits/ha; Villafuerte et al., 1997) is greater than that of Andújar (Simón et al., 2012). In areas or during periods of lower prey abundance, competition may play a more important role and interspecific interactions may change, resulting in increased interference competition (Creel, 2001). Lower prey densities can result in lower lynx tolerance toward foxes and, consequently, greater interference competition. Similar conclusions were reached by (Gese et al., 1996) in Yellowstone National Park, where coyotes tolerate red foxes during high prey years but not at other times. Although data regarding the presence of the smaller mesocarnivores (Mongoose, martens and genets) are scarce, our results concur with previously reports from Doñana, where mongoose and genets avoid areas where lynx are present. Iberian lynx and badgers seem to be particularly well predisposed to coexist (Palomares et al., 1996; Fedriani et al., 1999), and our results suggest that there is a complete spatial overlap between the species. Kleiman and Eisenberg (1973) suggest that this coexistence occurs as a result of a separation in their ecological niches, which is likely a consequence of evolution of different social systems. Similar interactions have been described between Eurasian lynx and wolves in Białowieza Forest (Schmidt, 2008) and between lynx and wolverine in northern Sweden (Schmidt, 2008). The Iberian lynx is a crepuscular species that preys mainly on rabbits (Fedriani et al., 1999), whereas badgers are much more nocturnal and are generalists with the capacity to survive on a

< 0.0001

0.0038

0.016

greater diversity of resources (Roper, 1994; Neal and Cheeseman, 1996; Revilla and Palomares, 2002). The food available for badgers in Mediterranean habitats varies greatly and badgers respond by shifting their diets accordingly between prey items (Virgós et al., 2004). However, niche differences alone cannot completely explain this coexistence. Foxes are even more adaptable than badgers and could potentially develop resource partitioning, temporal avoidance strategies (Voigt and Earle, 1983; Johnson and Franklin, 1994; Kozlowski et al., 2008), or different foraging strategies (Husseman et al., 2003) to facilitate coexistence. However, fox distribution is clearly influenced by the presence of lynx while badger distribution is not. The outcome of direct encounters between lynx and badgers is unknown but probably involves a risk of injury for both species. Therefore, the observed sympatry between Iberian lynx and badger is probably facilitated by a combination of both factors –the avoidance of injury and different foraging strategies.

Table 2. Mann–Whitney U–test results for the variables of habitat and rabbit abundance. Tabla 2. Resultados de las pruebas U de Mann–Whitney para las variables del hábitat y la abundancia de conejos. Pasture

p–level

–0.64 0.52

Scrub < 50cm

Scrub > 50 cm

0.96

0.33

Tree

–0.48 0.63

Rocks

1.28 0.2

Rabbit

2.08 0.03

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As a result of being a trophic specialist on rabbits, the abundance of its staple prey determines the lynx’s basic demographic parameters (Monterroso et al., 2016) and distribution (Guzmán et al., 2004), which thus implies that there is bottom–up control over Iberian lynx dynamics. Likewise, the presence or absence of the Iberian lynx, which is determined by rabbit abundance, affects the dynamics of subordinate carnivore species via a top–down control effect. The foraging theory suggests that animals adjust their behaviour accordingly to optimize foraging efficiency and overall fitness, and trade–off harvesting rates with fitness costs (Haswell et al., 2018). In the absence of Iberian lynx, sympatric mesocarnivores should ideally be distributed on the basis of habitat quality and preferred food availability (Van Der Meer and Ens, 1997; Roemer et al., 2009). The presence of the lynx forces smaller species to invest in antipredator behavioural strategies (Lima, 1998; Haswell et al., 2017) that can have negative consequences. For example, their access to high–quality foraging areas can be restricted (Ritchie and Johnson, 2009), which forces them to seek an alternative diet, adopt their life cycles to those of their new prey items, and adjust their feeding behaviour (Durant, 2000; Hayward and Slotow, 2009; Wikenros et al., 2014). This in turn can affect the size of the home range, increase travel costs or lead to shifts in habitat use (Caro and Stoner, 2003). The fitness costs of these antipredator responses could affect survival and reproduction, thereby ultimately having an impact on population dynamics (Creel and Christianson, 2008). On the other hand, a fall in lynx numbers is expected following rabbit declines, which will lead to a lessening of the top–down control on mesocarnivores numbers (Estes et al., 2011; Monterroso et al., 2016). Conservation implications Numerous studies have drawn attention to the importance of apex predators in suppressing populations of smaller predators (mesopredators) and thus their roles in moderating the impact of predation on smaller prey species (Crooks and Soulé, 1999; Johnson et al., 2007; Berger et al., 2008). The recovery and re–establishment of apex predator populations contribute not only to their conservation but also benefit biodiversity conservation via a relaxing of the impact of mesopredators on their prey (Ritchie and Johnson, 2009). This is positive for the restoration of disrupted ecosystem processes (Estes et al., 2011; Ritchie et al., 2012), particularly in terms of trophic interactions (Monterroso et al., 2016) but also for economic and social reasons (ecosystem services). Some areas in rural Spain have high rabbit densities and suitable habitat for the lynx. Most such areas are occupied by private, intensively managed, small–game hunting areas (rabbit and partridge; Delibes–Mateos et al., 2009). In these hunting estates strong predator control is traditional and still persists nowadays, both legally (leg–hold traps and snares when authorised under certain exceptional circumstances) and illegally (Villafuerte et al., 2000; Virgós and Travaini, 2005). Despite the possible negative effect on non–target

species, this practice requires important time and monetary expenditure, although the desired results are not always achieved (Harding et al., 2001). Lynx are viewed negatively by many hunters in the Iberian Peninsula since, as a trophic specialist that preys on rabbits, it competes for this highly important small– game species. Nevertheless, the Iberian lynx presence could be an effective, natural and inexpensive tool for predator control since it suppresses populations of smaller predators and thereby mitigates the impact that these mesopredators will have on game species (Palomares et al., 1995). This is a key argument for changing game managers’ opinions and for ensuring a favourable response to any lynx reintroduction project in its past range from where, ironically, it was eradicated by indiscriminate predator control (Gil–Sánchez and McCain, 2011). Acknowledgements This study was supported by DGCONA–MIMAM project 'Censo–Diagnóstico de las poblaciones de Lince Ibérico en España', and by IBiCO/WWF Spain/ Fundación Barcelona Zoo project 'Competencia interespecífica y coexistencia entre el lince ibérico (Lynx pardinus) y otros carnívoros'. We wish to express our gratitude to Nicolas Guzmán, Paco García, Aquilino Duque and Concha Iglesias who carried out fieldwork with us. We also thank the Organismo Autónomo de Parques Nacionales (Lugar nuevo), Parque Natural de la Sierra de Andújar, TRAGSA, Fundación CBD– Habitat, WWF/España, EGMASA, CMA Junta de Andalucía. We thank Jose Luis Tellería and Guillermo López for their constructive comments. References Atwood, T. C., Gese, E. M., 2008. Coyotes and recolonizing wolves: social rank mediates risk–conditional behaviour at ungulate carcasses. Animal Behaviour, 75: 753–762. Berger, K. M., Gese, E. M., 2007. Does interference competition with wolves limit the distribution and abundance of coyotes? Journal of Animal Ecology, 76: 1075–1085. Berger, K. M., Gese, E. M., Berger, J., 2008. Iindirect effects and traditional trophic cascades: a test involving wolves, coyotes, and pronghorn. Ecology, 89: 818–828. Cabezas–Díaz, S., Virgós, E., Mangas, J. G., Lozano, Jorge, J. G., 2011. The presence of a "competitor pit effect" compromises wild rabbit (Orcytolagus cuniculus) conservation. Animal Biology, 61: 319–334. Caro, T. M., Stoner, C. J., 2003. The potential for interspecific competition among African carnivores. Biological Conservation, 110: 67–75. Creel, S., 2001. Four factors modifying the effect of competition on carnivore population dynamics as illustrated by African wild dogs. Conservation Biology, 15: 271–274. Creel, S., Christianson, D., 2008. Relationships be-

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Accounting for uncertainty in assessing the impact of climate change on biodiversity hotspots in Spain D. Romero, J. Olivero, R. Real

Romero, D., Olivero, J., Real, R., 2019. Accounting for uncertainty in assessing the impact of climate change on biodiversity hotspots in Spain. Animal Biodiversity and Conservation, 42.2: 355–367, Doi: https://doi. org/10.32800/abc.2019.42.0355 Abstract Accounting for uncertainty in assessing the impact of climate change on biodiversity hotspots in Spain. Our limited understanding of the complexity of nature generates uncertainty in mathematical and cartographical models used to predict the effects of climate change on species’ distributions. We developed predictive models of distributional range shifts of threatened vertebrate species in mainland Spain, and in their accumulation in biodiversity hotspots due to climate change. We considered two relevant sources of climatological uncertainty that affect predictions of future climate: general circulation models and socio–economic scenarios. We also examined the relative importance of climate as a driver of species' distribution and taxonomic uncertainty as additional biogeographical causes of uncertainty. Uncertainty was detected in all the forecasts derived from models in which climate was a significant explanatory factor, and in the species with taxonomic uncertainty. Uncertainty in forecasts was mainly located in areas not occupied by the species, and increased with time difference from the present. Mapping this uncertainty allowed us to assess the consistency of predictions regarding future changes in the distribution of hotspots of threatened vertebrates in Spain. Key words: Climate change, Prediction accuracy, Taxonomic uncertainty, Threatened species, Uncertainty mapping Resumen Considerar la incertidumbre en la evaluación de los efectos del cambio climático en las zonas de gran diversidad de España. Nuestra comprensión incompleta de la complejidad de la naturaleza genera incertidumbre en los modelos matemáticos y cartográficos utilizados para predecir los efectos del cambio climático en la distribución de las especies. Se elaboraron modelos para predecir los cambios producidos por el cambio climático en la distribución de las especies de vertebrados amenazados en la España peninsular y en sus correspondientes zonas de alta biodiversidad. Se consideraron dos fuentes importantes de incertidumbre climática que afectan a las predicciones climáticas: los modelos de circulación general y el contexto socioeconómico. Asimismo, se analizó la importancia relativa del clima en cuanto factor determinante de la distribución de las especies y la incertidumbre taxonómica como causas biogeográficas añadidas de incertidumbre. Se detectó incertidumbre en todos los pronósticos realizados a partir de modelos en los que el clima era un factor explicativo significativo y en las especies con incertidumbre taxonómica. En los pronósticos, la incertidumbre se localizó principalmente en áreas no ocupadas por las especies y aumentó con el desfase temporal respecto al presente. La representación cartográfica de esta incertidumbre permitió evaluar la coherencia de las predicciones con respecto a los futuros cambios de la distribución de las zonas de alta biodiversidad de vertebrados amenazados de España. Palabras clave: Cambio climático, Exactitud de las predicciones, Incertidumbre taxonómica, Especies amenazadas, Mapas de incertidumbre Received: 15 X 18; Conditional acceptance: 11 III 19; Final acceptance: 17 VI 19 David Romero, Laboratorio de Desarrollo Sustentable y Gestión Ambiental del Territorio del Instituto de Ecología y Ciencias Ambientales de la Facultad de Ciencias de la Universidad de la República (Udelar), Iguá 4225, 11400 Montevideo, Uruguay (Uruguay).– Jesús Olivero, Raimundo Real, Grupo de Biogeografía, Diversidad y Conservación, Departamento de Biología Animal, Universidad de Málaga, Campus de Teatinos, 29071 Málaga, España (Spain). Corresponding author: David Romero. E–mail: davidrpbio@fcien.edu.uy ISSN: 1578–665 X eISSN: 2014–928 X

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Introduction Species distribution modelling (SDM) is useful to forecast the potential consequences of climate change on conservation of biodiversity (Dawson et al., 2011). A great deal of digital cartographic information has been produced related to observed and predicted climate, such as WorldClim (Hijmans et al., 2005; Fick and Hijmans, 2017), CliMond (Kriticos et al., 2012) and IPCC gas–emission scenarios (GESs) (Nakićenović et al., 2000). This large amount of information has significantly advanced predictive SDM. However, in SDMs and subsequent forecasts, uncertainties affect the reliability of predictions, leading to misconceptions and interpretation errors (Knutti, 2008), with critical consequences on the application of distribution forecasts to conservation policy (Real et al., 2010). Identifying the geographic distribution of uncertainty associated with a predictive model is, consequently, as important as the model mapping itself (Beale and Lennon, 2012; Kujala et al., 2013). Several sources of uncertainty have been analysed, such as the variety of SDM methods (Carvalho et al., 2011; Beale and Lennon, 2012), the inherent imperfection of atmospheric circulation models (i.e. GCMs, Knutti, 2008; Real et al., 2010), alternative proposals on future GESs (Real et al., 2010, Carvalho et al., 2011), the resolution of climate data (McInerny and Purves, 2011), and survey design (Tessarolo et al., 2014). However, other sources of uncertainty in mapping species distributions have seldom been studied (Rocchini et al., 2011). Few studies have assessed the effects of taxonomic uncertainty (Lozier et al., 2009; Romero et al., 2013; McInerny and Purves, 2011; Tessarolo et al., 2017), diversity of sources for climate data (Fernández et al., 2013; García–López and Real, 2014), behavioural plasticity of species in their response to climate change (Muñoz et al., 2015), correlations between climate, and other environmental factors (Real et al., 2013). These causes of uncertainty can affect model accuracy more than the availability of GCMs and GESs. The combination of models based on different species provides a dynamic measure of potential species’ richness, as it can fill gaps in distribution knowledge mainly due to sampling bias (Estrada and Real, 2018). This combination is particularly useful for predicting future changes in the distribution of biodiversity hotspots (Estrada et al., 2008; Real, et al., 2017). When focused on endangered species, it can enable forecasts with important applications for conservation. However, combining models requires the use of SDM outputs based on a commensurate index that provides comparable measures of the importance of different localities for different species. This index is provided by the favourability function (Real et al., 2006; Acevedo and Real, 2012) as it removes the effect of prevalence from predicted probabilities, and therefore more accurately describes the environmental conditions that facilitate species presence, regardless of the proportion of presences in the dataset (Barbosa and Real, 2012; Acevedo and Real, 2012). Favourability models can be combined

through the application of fuzzy–logic operations (Estrada et al., 2008; Barbosa and Real, 2012; Romero et al., 2014; Olivero et al., 2017). In this study, we took advantage of the properties of the favorability function regarding model combination to forecast how climate change may modify the location of biodiversity hotspots for threatened vertebrates in mainland Spain and to analyse the uncertainty associated with the resulting forecasts. We took into account the effect of alternative general circulation models, different gas–emission scenarios, the correlation between climate and other factors, and taxonomy on the forecasts. Specifically, we assessed the uncertainty associated with the identification of the areas where threatened species are most vulnerable to climate change. We also mapped the distribution of the degree of uncertainty, and quantified the reliability of forecasts across the study area. Material and methods Species and study area Spain comprises 84 % of the Iberian peninsula, a biogeographically relevant area for the conservation of biodiversity hotspots (Maiorano et al., 2013). We analyzed all threatened vertebrate species in mainland Spain, where more than sixty percent of those with European distributions (in terms of extent of occurrence) are found. Only 14 species had less than 90 % of their European distribution within Spain, namely Chioglossa lusitanica, Calotriton asper, Salamandra salamandra, Rana iberica, Iberolacerta bonnali, I. aranica, Mauremys leprosa, Cercotrichas galactotes, Emberiza shoeniclus, Tetrao urogallus, Tetrax tetrax, Arvicola sapidus, Microtus cabrere y Rhinolophos mehelyi, which present significant populations in Portugal or France. We modelled seven amphibian, seven reptile, twelve bird and six mammal species (table 1s). We selected the threatened species according to the IUCN criteria (vulnerable, endangered and critically endangered), adapted to Spain by national red books (Madroño et al., 2004; Pleguezuelos et al., 2004; Palomo et al., 2007). Exceptions were: Calotriton asper, selected because some authors consider it is threatened (Montori and Llorente, 2008); Salamandra salamandra, because its subspecies (S. s. longirostris), which is proposed to be a separate species (Dubois and Raffaëli, 2009), is vulnerable; and Tetrao urogallus and Emberiza shoeniclus, as all their populations in Spain are threatened subspecies. Presence in the 5,156 10 x 10–km UTM grid cells of mainland Spain were obtained from Pleguezuelos et al. (2004) for amphibians and reptiles, from Martí and del Moral (2003) for breeding birds, and from Palomo et al. (2007) for mammals. These data represent all or most of the distribution of the analysed species. Data of similar quality were partly unavailable from Portugal, which is why we restricted our analysis to Spain. In addition, the models were explicitly built to be helpful in conservation policy decisions, such as

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the design of nature reserve networks, which is decided at a national level. The cells without presence according to the respective distribution atlas were considered as absences. Although the number of presences was extremely low in a few cases (see table 1s), Proosdij et al. (2016) showed that the lower limit for presences needed to obtain good model performance depends on the species’ prevalence in the dataset, with absolute minimum sample sizes extremely low for narrow–ranged species. In our case, all cases with an extremely low number of presences corresponded to endemic and narrow–ranging species, and included the whole range of the species. Lobo and Tognelli (2011) showed that the model performance also depends on the number of absences used to calibrate the models, with better performance when absences are numerous and unbiased. In our case, the number of absences was always very high and unbiased, particularly when modelling species with highly restricted ranges. Predictors Factors other than climate should be considered in SDMs constructed to forecast changes induced by climate in species distribution (Aragón et al., 2010; Márquez et al., 2011). We assessed four factors that had a potential impact on species distributions (Márquez et al., 2011): climate, space, topography, and human influence (table 2s). Climate is the main driver of species distribution to be analysed when assessing the effect of climate change on species distribution. The use of the spatial factor in the models accounts for geographical trends that cannot be explained by climate (Legendre, 1993). These spatial trends may arise from population dynamics, dispersal capacities, and historical events that affected species distributions (Legendre, 1993; Real et al., 2003). The topographic factor allows us to assess whether there is any relationship between the topographic structure of the territory and the distribution of the species, independently of the relationship between topography and climate. Finally, human activity may have an effect on the availability and quality of habitats of many species, possibly interfering with the effects of climate (Delibes–Mateos et al., 2009). Table 2s shows the climatic, spatial, topographic and human variables and sources. The original resolution adopted for the variables was one km2 per pixel; we computed average values for each 10 × 10–km square using ArcGIS 10.0 zonal statistic tools (ESRI, 2011). With the longitude and latitude of the original spatial variables we built a single spatial variable to be used as a spatial predictor for every species. For this, we made a trend surface analysis by performing a backward stepwise logistic regression of each species’ presence/absence on nine spatial components that describe the spatial position of the data: X, Y, X2, Y2, X×Y, X3, Y3, X2×Y, Y2×X (X, latitude; Y, positive longitude). This produced a lineal combination of spatial components that we used as the single spatial predictor in the subsequent models. We used IBM SPSS statistics 21 (IBM, 2012) for this analysis.

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Selection of variables and model building Spearman correlation coefficients (r) were calculated to control the multicollinearity between independent variables. When a set of variables belonging to the same factor were correlated with r > 0.8, we selected just the variable with the most significant predictive power on the species presence; this was established by performing a logistic regression of presences/absences on the set of correlated variables separately and selecting the variable most significantly related with species presence/absence. This resulted in a reduced set of potential predictors that were tailor– suited for each species. We then performed logistic regressions of each species' presence/absence on the remaining variables separately. The false discovery rate (FDR) was used to control the increase in type I errors due to the number of remaining independent variables (Benjamini and Hochberg, 1995). The variables significantly related (p < 0.05) to the species distribution under a FDR < 0.05 were considered to be the further reduced subset of acceptable predictors. We next performed a multivariate forward–backward stepwise logistic regression of presences/ absences of each species on their corresponding reduced set of predictor variables to obtain probability values (P) of the species' presence in every square. Variables were included in the models according to the significance of their relationship with the species distribution while avoiding redundancy by checking at each step that the new variables added significant new information to the model. This was considered the current probability model for the species. Then, we modelled the present and future favourability for the presence of each species using the favourability function (Real et al., 2006, 2017). Favourability (F) was calculated from P using the following equation (Real et al., 2006): F = [P/(1 – P)] / [(n1/n0) + (P/[1 – P])] where n1 and n0 are the number of presences and absences, respectively. Climate in these models refers to the period 1961–1990; the models were later projected to the expected conditions of three future periods: 2011– 2040, 2041–2070 and 2071–2100 in order to obtain the different future forecasts. To this end, we applied the following equation, using climate–variable values referring to the corresponding period: F = ey / [(n1/n0) + ey] where e is the base of the natural logarithm, and y is the logit function of the probability model obtained by multivariate logistic regression. We used IBM SPSS statistics 21 (IBM, 2012) for this analysis. We evaluated the classification and discrimination capacity of the models using four indices: sensitivity, specificity, correct classification rate (CCR), and Cohen’s Kappa (Fielding and Bell, 1997). These indices assessed classification based on the 0.5–favou-

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rability threshold, which, in the favourability function, makes probability equal to overall prevalence. Discrimination capacity was evaluated using the area under the curve (AUC) of the receiver operating characteristic. Although discrimination cannot be considered an overall measure of model performance (Lobo et al., 2008), AUC provides a measure of the degree to which the modelled predictors allow separating presences from absences, which is informative when geographical extent, presence/absence dataset and modelling technique remain constant, as is the case here (Lobo et al., 2008). Forecasting the future distribution of biodiversity hotspots We considered four sources of uncertainty associated with future forecasts (fig. 1): (1) two alternative general circulation models (GCMs): CGCM2 (Canadian Climate Centre for Modeling and Analysis) and ECHAM4 (Max Planck Institut für Meteorologie) (IPCC, 2013), regionalized to Spain by the Spanish Meteorological Agency (AEMET) (Brunet et al., 2007); (2) two different Gas–Emission Scenarios or GESs for the 21st century from IPCC, (Nakićenović et al., 2000): A2 and B2, representing intermediate positions in the range of projected temperature changes, being medium–high and medium–low respectively (Brunet et al., 2007); (3) the degree to which climate affects distribution models, as a consequence of correlations between climate and other factors. We forecast distribution changes according to the spatial variation in the model that was exclusively explained by climate (pure effect); alternatively, we forecast changes according to the spatial variation potentially, although not exclusively, attributable to climate (apparent effect). We followed the method described in Real et al. (2013), which is based on variation partitioning (Legendre and Legendre, 1998); (4) taxonomic uncertainty, in those cases with taxonomic categorization under discussion. We constructed alternative models taking into account the taxonomic situations both before and after revision. This approach was used for: Salamandra salamandra longirostris, which is either considered a S. salamandra subspecies (García–París et al., 1998), or a species named S. longirostris (Dubois and Raffaëli, 2009); Calotriton arnoldi, recently separated from Calotriton asper (Carranza and Amat, 2005); Iberolacerta monticola (Pleguezuelos et al., 2004), recently categorised into I. monticola, I. cyreni, I. martinezricai, and I. galani (Arribas et al., 2006; Arribas and Carranza, 2004, 2015). Taxonomic alternatives resulted in two or more models for a single original species. On one hand, we considered a model for the species before the taxonomic revision; on the other hand, we considered a model performed by joining models for the species resulting from taxonomic revision using the fuzzy logic operator 'fuzzy union', which is equivalent to assigning the highest value observed in different models to each square (Zadeh, 1965). We calculated these fuzzy union values in each cell using the Max function from Microsoft Excel (version 2010). We produced favourability models for each species according to each source of uncertainty. This resulted

in eight favourability models for species with no taxonomic uncertainty and eight further more favourability models for each taxonomic alternative (subspecies) for species with taxonomic uncertainty. We applied the accumulated favourability to forecast the distribution of future diversity hotspots for threatened vertebrates in mainland Spain derived from these models. The accumulated favourability is a proxy for a diversity index (Estrada et al., 2008; Real et al., 2017), defined by the sum of favourability models of a group of species: n

AFj =

i=1

Fij

where Fij is the favourability value for species i in square j. This index was applied to the 32 species analysed for 1961–1990 and for three future periods (2011–2040, 2041–2070, and 2071–2100) according to the four sources of uncertainty. Uncertainty assessment We used fuzzy logic operators to summarize the effect of the different sources of uncertainty on the predictive models (Zadeh, 1965). Fuzzy union represented the highest favourability value predicted for a species in each cell according to any of the four sources of uncertainty. Fuzzy intersection represented the lowest favourability value predicted for a species in each cell according to any of the four sources of uncertainty. Fuzzy intersection indicates the minimum consensus among the models (Romero et al., 2016). Average favourability predicted for a species in each cell according to any of the four sources of uncertainty was also computed as an indicator of a balanced consensus among the models. We calculated these values in each cell using the Microsoft Excel functions (version 2010): Max, Min and Average, respectively. We obtained the accumulated favourability values resulting from the fuzzy union, the fuzzy intersection and the average favourability of the models produced according to the four uncertainty sources, which summarized the effect of the different sources of uncertainty on the forecasted biodiversity hotspots. We also calculated the fuzzy symmetric difference (Dubois and Prade, 1980) between all pairs of SDMs produced for the same species whose differences were based on a single source of uncertainty (i.e., a symmetric difference for each species, time eriod, and uncertainty source): FA▼B(j) = |FA(j) – FB(j)| where FA and FB represent favourability values for a given j square according to two alternative SDMs for the same species. The degree of uncertainty in each j square (Uj) was calculated as the fuzzy union (the maximum value) of all the symmetric differences applying to a species in a given time period. The mapping of Uj indicated the geographic distribution of

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Variability of nature in predictions and forecasts Future forecasts

Models from 1961 to 1990

Presence / Absence

General circulation models

Predictor variables

Gas emission scenarios

Degree of participation of climate

A2 CGCM2

ECHAM4

B2 A2 B2

Taxonomic uncertainty

Before revision

Apparent effect

Pure effect

After revision

Fig. 1. Schematic representation of the model diversification caused by the different sources of uncertainty: N, number of species resulting from taxonomic revision in species with taxonomic uncertainty; X2, duplication of forecasts due to the different consideration of climate as a driver of species distribution; and XN, multiplication of models due to different taxonomic alternatives for species with taxonomy uncertainty. Fig. 1. Esquema de la diversificación del modelo provocada por las distintas fuentes de incertidumbre: N, número de especies resultante de la revisión taxonómica en las especies de taxonomía incierta; X2, duplicación de los pronósticos debido a la diferente consideración del clima como factor determinante de la distribución de las especies; XN, multiplicación de modelos debido a las diferentes alternativas taxonómicas para las especies de taxonomía incierta.

the uncertainty associated to the predicted species distribution. Finally, we assessed the uncertainty associated to AFj values using the accumulated uncertainty index: n

AUj =

i=1

(Uij)

where Uij is the degree of uncertainty associated with the forecasts from different uncertainty sources obtained for species i in square j. Results Model assessment We obtained significant favourability models for the 32 species considered. Figure 2 shows an example for one species, and all models can be seen in supplementary material. Classification and discrimination assessments generally obtained high scores. On a scale ranging from 0 to 1, sensitivity was always higher than 0.6 (average 0.93); specificity was higher than 0.66 (average 0.88); CCR was higher than 0.69 (average 0.89); and Cohen's Kappa was higher than 0.1, with an average 0.42 or 'good' according to Fielding and Bell (1997). On a scale ranging from 0 to 1,

the AUC was always higher than 0.70 or 'acceptable' according to Hosmer and Lemeshow (2000), with an average of 0.95 or 'outstanding'. Relative importance of climate and other explanatory factors The pure effect of climate explained more than 40 % of the environmental favourability for 17.5 % of the species analysed: 27 % of reptiles, 25 % of birds, 9 % of amphibians, but 0 % of mammals (see Supplementary material, in fig. 1s, pictures 1–32). When the apparent effect of climate was considered, it explained more than 40 % of the environmental favourability in 72.5 % of the species: 83 % of mammals, 81 % of amphibians, 67 % of birds, and 64 % of reptiles. In contrast, non–climatic variables explained more than 40 % of the environmental favourability in 42.5 % of the species: 68 % of mammals, 64 % of amphibians, 36 % of reptiles, and 17 % of birds (see supplementary material, in fig. 1s, pictures 1–32). Effect of climate change on Spanish threatened vertebrates Eight different forecasts per species and future period were calculated, varying according to the sources of ambiguity considered in every case (see

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Distribution Aquila adalberti (Aves, Accipitriformes, Aquila) Spanish imperial eagle Forecast Predictions 1961–1990

Apparent effect of climate 2071–2100

Pure effect of climate 2071–2100

CGCM2–A2

CGCM2–B2

ECHAM4–A2

0 160 320 480 km Predicted favourability

Uncertainty in distribution forecast 2071–2100

Minimum favourability

Maximum favourability

Predicted favourability

Average favourability

Distribution of uncertainty

Predicted uncertainty

Fig. 2. Example of the methodology modelling for Aquila adalberti with the forecasts according to: two general circulation models, two gas–emissions scenarios and the different contribution of climate. Below, the consensual favourability values (minimum, maximum and average), and the uncertainty distribution in 2071–2100. Fig. 2. Ejemplo de la metodología de elaboración de modelos para Aquila adalberti con los pronósticos según: dos modelos generales de circulación, dos situaciones hipotéticas de emisiones de gases y la diferente contribución del clima. Abajo, los valores consensuados de favorabilidad (mínimo, máximo y promedio) y la distribución de la incertidumbre en el período 2071–2100.

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supplementary material, in fig. 1s, pictures 1–32), with eight more forecasts for each taxonomic alternative (subspecies) in the species with taxonomy uncertainty. When the apparent effect of climate was considered, the increment in favourability (Real et al., 2010) was positive for 48.8% of the species, negative for 39.6 % of the species, and nearly zero (< 1 %) for 12 % of the species. The average change between 1961–1990 and 2071–2100 for all models was 26 %. However, distributions forecasted according to the pure effect of climate were predicted to experience lower changes (average 7.3 %): favourability increased in 26.8 % of species, decreased in 27.4 %, and remained unchanged in 45.7 % (see supplementary material, in fig. 1s, pictures 1–32). A decrease in favourability over time was forecast for 60.4 % of the studied endemic mammals, 57.3 % of birds, 29.6 % of amphibians, and 27.1 % of reptiles. Opposite trends depending on the GCM considered were predicted only for five species (Alytes dickhilleni, Algyroides marchi, Chioglossa lusitanica, Lepus castroviejoi, and Pterocles alchata). Our SDMs forecast an exceptionally large decrease of favourability (> 50 % according to at least two methodological options), for another five species (Rana pyrenaica, Iberolacerta montícola, Chersophilus duponti, Pterocles alchata, and Tetrao urogallo). All the methodological options explored forecast a decrease in favourability in three species only (Iberolacerta bonnali, Iberolacerta montícola, and Chersophilus duponti). For seven more species (Chioglossa lusitanica, Triton pygmaeus, Mauremys leprosa, Arvicola sapidus, Galemys pyrenaicus, Microtus cabrerae, and Rhinolophus melei), a decrease in favourability was predicted according to all the forecasts based on the apparent effect of climate. Forecasted biodiversity hotspots and associated uncertainty Figure 3 shows the accumulated favourability (AF) of all species, representing the biodiversity hotspots. The accumulated favourability showed a general positive increase over time in the centre and south–west of Spain (fig. 3). In contrast, the lowest accumulated favourability values were detected in the north–east (see fig. 3, 4). The uncertainty associated to AF values can be seen in figure 4. Uncertainty was detected in all the forecasts involving climate variables and in the species subject to taxonomic uncertainty. Favourability forecasts showed low and highly localized uncertainty values for most species, with high uncertainty values generally aggregated at the edges of species distributions. In 87.5 % of the forecasts, the area affected by uncertainty increased as a function of the time elapsed from the present (fig. 4; and fig. 2s, pictures 1–32); a decrease was detected only in the case of Chersophilus duponti (fig. 2s, picture 19). Our models showed an aggregation of the highest uncertainty values in the eastern half of the Iberian peninsula, including the Pyrenees and some southern mountains (fig. 4). In the north–eastern Pyrenees, some species, such as Iberolacerta monticola,

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showed high discrepancies in future favourability values forecasted by alternative models (fig. 1s, picture 10.1), and therefore high uncertainty (fig 2s, picture 10.1). Discussion Effects of uncertainty in distribution models Several authors have indicated the importance of taking into account the different sources of uncertainty in species distribution models (Knutti, 2008; Real et al., 2010; Rocchini et al., 2011; Beale and Lennon, 2012). Our results also highlight that the implementation of the different uncertainty measures in SDMs is key for obtaining reliable results. Specifically, our analysis allowed us to identify the areas where the models from different uncertainty sources are consistent and the areas where the uncertainty was mainly located (Beale and Lennon, 2012; Kujala et al., 2013). The comparison between the consensus in the predictions between the different models and the associated uncertainty may thus be useful to locate the forecast important territories for conservation of species and the degree of reliability of these predictions (Beale and Lennon, 2012). Relative importance of climate as a driver of species distribution Climate was the most important factor influencing the distribution of birds and reptiles, whereas other factors were more influential for mammals and amphibians. This is probably a consequence of the greater dispersal ability of birds, which makes them less tied to regional influences such as historical events, geographic barriers, and local human influences (Cumming et al., 2012), and of the high dependence of reptiles on temperature (Adolph and Porter, 1993). However, much of the role of climate in explaining distributions fell on the intersection with other variables, in which the role of climate cannot be distinguished from the role of other factors (Real et al., 2013). In our models, climate not only had different degrees of influence depending on the species and the general circulation models (GCMs) and gas–emission scenarios (GES) analysed, but the relative contribution of climate was also responsible for the largest differences between forecasts. Thus, the analysis of this source of uncertainty is of great relevance for qualifying and geographically locating the reliability of distribution forecasts based on climate change. Climate change and predicted changes in favourability The concern about the negative impact of climate change on biodiversity was behind many of the studies on distribution forecasts in climate change scenarios (Bellard et al., 2012). Assessments of the positive impacts of climate change on species distributions are nevertheless accumulating (e.g. Araújo et al., 2006; Romero et al., 2013; Sorte et al., 2013; García–Valdés et al., 2015). We found positive, neutral, and negative effects depending on the species

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Accumulated favourability 2011–2040

CGCM2–A2

2041–2070

2071–2100

1961–1990

Apparent effect

Pure effect

CGCM2–B2 1961–1990

Apparent effect

Pure effect

ECHAM4–A2

2011–2040

2041–2070

2071–2100

1961–1990

Apparent effect

Pure effect ECHAM4–B2 1961–1990

Apparent effect 0 160 320 480 km

Pure effect Accumulated favourability 0

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Accumulated favourability 2011–2040

2041–2070

2071–2100

Accumulated minimun favourability

Accumulated maximun favourability

Accumulated average favourability

0 160 320 480 km Accumulated favourability

Accumulated uncertainty

0 160 320 480 km 0

Fig. 4. Consensual accumulated favourability values (minimum, maximum, and average) and accumulated uncertainty taking into account the 32 species analysed for the three time periods. Fig. 4. Valores consensuados de favorabilidad acumulada (mínimo, máximo y promedio) e incertidumbre acumulada teniendo en cuenta las 32 especies analizadas en los tres períodos de tiempo.

Fig. 3. Accumulated favourability for the 32 species analysed from 1961–1990 and for three future periods (2011–2040, 2041–2070, and 2071–2100) according to four sources of uncertainty: two general circulation models, two emissions scenarios, and the different contribution of climate to the species distribution (apparent or pure effect). Fig. 3. Favorabilidad acumulada de las 32 especies analizadas en el período 1961–1990 y en tres períodos futuros (2011–2040, 2041–2070 y 2071–2100) según las cuatro fuentes de incertidumbre: dos modelos de circulación general, dos situaciones hipotéticas de emisiones de gases y la diferente contribución del clima a la distribución de las especies (efecto aparente o efecto puro).

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considered. Consequently, our results suggest that climate change could harm many species by altering the local conditions they are adapted to, whereas it could contrarily move the environmental conditions closer to the optimal requirements of some other species. Furthermore, species can adapt to new climatic conditions by modifying their phenology (Parmesan, 2007) and physiology (Johansen and Jones, 2011), not only by changing their distribution ranges (Moreno–Rueda et al., 2011). Several authors have predicted that reptiles and amphibians have the highest susceptibility to climate change (Araujó et al., 2006; Maiorano et al., 2013). However, our results suggest that some bird species could experience the highest decrease in favourability values (around 30 %), whereas this decrease could be about half (around 15 %) in the case of amphibians, reptiles, and mammals. This discrepancy could be due to the fact that we forecast on the basis of different degrees of participation of climate in the models. For a large number of birds, the highest proportion of explanatory power in the models was due to climate, regardless of whether the pure effect or the apparent effect of this factor was considered (non–climatic factors had more explanatory power only in the models of 17 % bird species). In all cases, the proportion of change over time was higher when the apparent effect of climate was taken into account than when its pure effect was considered (average difference, 25 %). According to our forecasts, there are two groups of species for which conservation measures should be reinforced. The first group includes already endangered species (Iberolacerta bonnali, Iberolacerta montícola, and Chersophilus duponti), for which there were consensus on a predicted decrease in environmental favourability. We consider that these are good examples of reliable predictions that highlighted the suitable territories to incorporate in the decision–making framework for the conservation of these species (Beale and Lennon, 2012; Estrada and Real, 2018). Priority actions should therefore be implemented for these species in the territories with reliable forecasts of range contraction in order to prevent strong negative impacts due to climate change (Pleguezuelos et al., 2004; Madroño et al., 2004; Dawson et al., 2011). The second group includes species whose predicted trends are seriously affected by uncertainty, but for which a decrease in favourability is forecast by some of the models: Rana pyrenaica, Salamandra longirostris, Pterocles alchata, Tetrao urogallo, Arvicola sapidus, Galemys pyrenaicus, and Microtus cabrerae. For some of these species, we detected areas where the decrease in favourability was not affected by uncertainty: some squares at the southern limit of the Rana pyrenaica distribution; the core squares of the north–western populations of Tetrao urugallo, that is home to a large part of the genetic stock of populations at the southern limit of its global distribution, essential for the conservation of the genetic biodiversity of the species (Alda et al., 2013); the southern half of the Iberian peninsula, where the presence of Arvicola sapidus is scattered (Palomo et al., 2007); and the mid–western populations of

Microtus cabrerae, for which climate was not the main driver of its distribution or possible decline (Alagador and Cerdeira, 2018). Anyway, for this group of species our results highlighted territories to monitor the populations and evaluate the possibility of applying conservation measures (Dawson et al., 2011). Areas most vulnerable to climate change We found that a large part of western peninsular Spain has areas with favourable environmental conditions for a great majority of threatened vertebrates, an area that includes five Spanish National Parks (Estrada and Real, 2018). In contrast, the areas with the lowest favourability values and the highest uncertainty are located in some eastern areas. Wildlife managers will have to decide whether conservation priorities should focus on areas with the highest favourable environmental conditions for a greater number of species, or on areas with less favourable conditions but in which some species are present (Real et al., 2017). Management efforts could also be prioritized according to present or to future environmental conditions (Dawson et al., 2011; Beale and Lennon, 2012; Kujala et al., 2013). Our study contributes to this issue by identifying the areas in which environmental favourability is currently high, but in which a future decrease in favourability is predicted. Conservation measures should be reinforced, for example, by adapting the location and extension of protected spaces within the network of protected areas taking into account the dynamics of climate change (Estrada and Real, 2018), especially when the affected species are endemic or are narrowly distributed within the Iberian peninsula (e.g., Alytes dickhilleni, Iberolacerta monticola, Aquila adalberti, or Lynx pardinus). It is also important to monitor areas that could become favourable for some species in the future under the effect of climate change, especially areas in which the accumulated favourability could reach high values and uncertainty is low; these areas might act as refuges for species vulnerable to climate change. This is the situation of the following species: Salamandra salamandra in the northernmost ridge of the Iberian peninsula, the Duero and Guadalquivir valleys, and in southern Spain (Romero et al., 2012; Tejedo et al., 2003); Algyroides marchi in the southeast of the Iberian peninsula (Pleguezuelos et al., 2004); Emberiza shoeniclus in the Pyrenees and the northwest corner and eastern edge of the Iberian peninsula (Madroño et al., 2004); and Rhinolophus mehelyi in southwestern Spain (Palomo et al., 2007). Concluding remarks Based on SDMs, climate change could have both positive and negative impacts on biodiversity. These effects will probably affect the distribution of species to different extents, and predictions will be more or less accurate in different locations according to each individual case. Advances in SDM cannot claim to eliminate the uncertainty involved in predictions, because uncertainty is often a result of the complexity

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of nature and of our incomplete knowledge about how it works, and therefore about how to forecast its changes. The method used in this study takes into account a wide range of possible variations according to a range of sources of uncertainty, and also takes uncertainty into account to identify spatial overlap between alternative forecasts. Although climate was the most important factor influencing the distribution of threatened birds and reptiles, other factors were more influential than climate for threatened mammals and amphibians; m oreover, climate not only had different degrees of influence depending on the vertebrate group, but correlations between climate and other factors were responsible for the largest differences between alternative forecasts. Besides, imprecision in forecasts increased as predictions move forward in time. Acknowledging, identifying, and quantifying the degree of imprecision in equally probable models for the same species generate more accurate predictions and serve to assess reliability in forecasts based on climate change. Acknowledgements Dr. David Romero was supported by a grant from the ANII de Uruguay (2016–2018), and from the Graduate Academic Commission (CAP) of the Universidad de la República (2018–2020). We thank S. Coxon for his help in the English revision of the manuscript. References Acevedo, P., Real, R., 2012. Favourability: concept, distinctive characteristics and potential usefulness. Naturwissenschaften, 99: 515–522. Adolph, S. C., Porter, W., 1993. Temperature, activity and lizard life histories. Amer. Naturalist, 142: 273–295. Alda, F., González, M. A., Olea, P. P., Ena, V., Godinho, R., Drovetski, S. V., 2013. Genetic diversity, structure and conservation of the endangered Cantabrian Capercaillie in a unique peripheral habitat. European Journal of Wildlife Research, 59: 719–728. Alagador, D., Cerdeira, J. O., 2018. A quantitative analysis on the effects of critical factors limiting the effectiveness of species conservation in future time. Ecology and Evolution, 8: 3457–3467. Aragón, P., Rodríguez, M. A., Olalla–Tárraga, M. A., Lobo, J. M., 2010. Predicted impact of climate change on threatened terrestrial vertebrates in central Spain highlights differences between endotherms and ectotherms. Animal Conservation, 13: 363–373. Araújo, M. B., Thuiller, W., Pearson, G., 2006. Climate warming and the decline of amphibians and reptiles in Europe. Journal of Biogeography, 33: 1712–1728. Arribas, O., Carranza, S., 2004. Morphological and genetic evidence of the full species status of Iberolacerta cyreni martinezricai (Arribas, 1996). Zootaxa, 634: 1–24.

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Scat analysis as a preliminary assessment of moose (Alces alces andersoni) calf consumption by bears (Ursus spp.) in north–central British Columbia R. V. Rea, L. Ajala–Batista, D. A. Aitken, K. N. Child, N. Thompson, D. P. Hodder Rea, R. V., Ajala–Batista, L., Aitken, D. A., Child, K. N., Thompson, N., Hodder, D. P., 2019. Scat analysis as a preliminary assessment of moose (Alces alces andersoni) calf consumption by bears (Ursus spp.) in north– central British Columbia. Animal Biodiversity and Conservation, 42.2: 369–377, Doi: https://doi.org/10.32800/ abc.2019.42.0369 Abstract Scat analysis as a preliminary assessment of moose (Alces alces andersoni) calf consumption by bears (Ursus spp.) in north–central British Columbia. Moose (Alces alces andersoni) population numbers have decreased by 50–70 % in some parts of northern British Columbia (BC), Canada. Predation of moose calves by bears may be affecting moose populations in this area, but has gone undocumented. A total of 1,381 bear scats were collected during the spring and summer of 2014 and 2015. Hairs extracted from the scats were identified to species through hair scale imprints made in thermoplastic film, with the specific purpose of identifying the frequency of occurrence of moose calf hairs in scats. Only 27 scats (~2 %) contained moose calf hair. We discuss possible explanations for our findings. Key words: Alces alces andersoni, Moose neonate, Diet, Feces, Hair analysis, Predation Resumen Análisis de heces para la evaluación preliminar del consumo de becerros de alce (Alces alces andersoni) por osos (Ursus spp.) en la zona centroseptentrional de la Columbia Británica. La población de alce (Alces alces andersoni) ha disminuido entre el 50 % y el 70 % en algunos lugares del norte de la Columbia Británica, en Canadá. La depredación de becerros de alce por osos puede estar afectando a la población de alces en esta zona; sin embargo, no se ha documentado. Se recolectaron 1.381 heces de oso durante la primavera y el verano de 2014 y 2015. Con el propósito de determinar la frecuencia de presencia de pelo de becerro de alce en las heces, se identificaron las especies a las que pertenecían los pelos extraídos de las heces a través de las impresiones de las escamas de los mismos en películas termoplásticas. Solo 27 heces (~ 2 %) contenían pelos de becerros de alce. Analizamos las posibles explicaciones de los resultados. Palabras clave: Alces alces andersoni, Neonato de alce, Dieta, Heces, Análisis de pelo, Depredación Received: 30 X 18; Conditional acceptance: 15 IV 19; Final acceptance: 25 VI 19 Roy V. Rea, Ecosystem Science and Management Program, University of Northern British Columbia, 3333 University Way, Prince George, British Columbia, Canada V2N 4Z9.– Larissa Ajala–Batista, Zoology Departament, Federal University of Paraná, Coronel Francisco Heráclito dos Santos Avenue, Jardim das Americas, Curitiba, Paraná, Brazil 80050–540.– Daniel A. Aitken, College of New Caledonia, 3330 22nd Avenue, Prince George, British Columbia, Canada V2N 1P8.– Kenneth N. Child, 6372 Cornell Place, Prince George, British Columbia, Canada V2N 2N7.– Neil Thompson, Applied Forest Management, University of Maine at Fort Kent, 23 University Drive, Fort Kent, Maine 04743, U.S.A.– Dexter P. Hodder, John Prince Research Forest, P. O. Box 2378, Fort St. James, British Columbia, Canada V0J 1P0. Corresponding author: R. V. Rea. E–mail: reav@unbc.ca

ISSN: 1578–665 X eISSN: 2014–928 X

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Introduction Bears (Ursus spp.) prey on ungulates (Singer et al., 1997; De Barba et al., 2014), principally throughout the calving season (Zager and Beecham, 2006; Patterson et al., 2012). Moose (Alces alces) calves are an important food source for bears during late May and June (Swenson et al., 2007), and both black bears (Ursus americanus) and grizzly (brown) bears (Ursus arctos) eat moose calves. Up to 90 % of moose calf mortality by bears occurs before mid–July (Ballard et al., 1981) after which moose calf mortality by bears may decline as a result of increased mobility (Zager and Beecham, 2006) and decreased vulnerability of moose calves (Boertje et al. 1988), or shifts in bear diet to other food sources (Zager and Beecham, 2006). Adult grizzly bears account for a wide range of reported moose calf mortalities: 39 % in Alberta (Hauge and Keith 1981); 58 % in the Yukon Territory (Larsen et al., 1989); and 39 % to 79 % in Alaska (Ballard et al., 1981; Boertje et al., 1987, 1988). Black bears are known to be responsible for a similar range of moose calf mortalities: 45 % in eastern interior Alaska (Bertram and Vivion, 2002); and 60–70 % in south–western interior Alaska (Garneau et al., 2007). Some authors, however, suggest that bear predation on moose neonates in some parts of Alaska is insignificant: 11 % (LeResche, 1968), trace (Hatler, 1972). Black bears can have a significant impact on moose calf mortality when bear numbers are > 0.2/km2 (Ballard, 1992). Furthermore, bear predation can influence the survival and recruitment of calves where moose population numbers are in decline relative to predator abundance (Boertje et al., 1988; Ballard, 1992; Gasaway et al., 1992). The effects of bear predation on moose calves can be exacerbated where moose numbers are relatively low (0–0.65/km2) and where wolves are present (Messier, 1994; Crête and Courtois, 1997). It has been hypothesized that moose population declines in British Columbia (BC hereafter) may be due to a change in moose habitat due to a recent epidemic of mountain pine beetle (Dendroctonus ponderosae Hopkins, 1902) and subsequent salvage logging (Kuzyk, 2016). This salvage logging has led to dramatic increases in road densities and associated access for both hunters and predators which may have also contributed to the decline (Gorley, 2016). The role of predation in the declines of moose, however, has also been the topic of much conjecture. Most recently, a survival analysis from the interior of British Columbia reported high adult cow moose survival and suggested that current population declines may be more related to calf mortality and inferred that calves are likely being targeted and killed by bears (Mumma and Gillingham, 2019). The extent to which bears consume moose or other prey has been studied using a variety of techniques, including observational studies (Landers et al., 1979; Hamer and Herrero, 1987; MacHutchon and Wellwood, 2003), analysis of data from neck– mounted cameras (Brockman et al., 2017), isotope analysis (Costello et al., 2016), fatty acid signatures

(Thiemann, 2008), feeding/kill site investigations (Franzmann et al., 1980; Ballard et al., 1981; Hamer and Herrero, 1987; MacHutchon and Wellwood, 2003), analysis of stomach (Landers et al., 1979) and intestinal contents (Wilton et al., 1984), and analysis of scat contents (Landers et al., 1979; Hamer and Herrero, 1987; Mattson et al., 1991; Hewitt and Robbins, 1996; MacHutchon and Wellwood, 2003; Munro et al., 2006). Since bears are known to consume the entire carcass of a moose calf before leaving the kill site (Boertje et al., 1988), we assumed that evidence of calf consumption from both predation and scavenging would be revealed through the identification of hairs found in scats as described in studies of wolf and bear predation on moose calves in Ontario (Voigt et al., 1976; Popp et al., 2018) and polar bear predation on marine mammals in Norway (Iversen et al., 2013). We conjectured that rates of bear predation on moose calves in north–central BC would be similar to those reported in the bordering jurisdictions of Alaska, the Yukon Territory, and Alberta. We predicted that an analysis of bear scat contents collected during (May–June) and after (July) calving would indicate whether or not bears were preying on moose calves, reveal the frequency of occurrence of moose calves in scats, and help elucidate how bears might be impacting calf recruitment. Material and methods Study area Our study area is on the interior plateau of BC between the Rocky and Coastal Mountains. The study area is located within the sub–boreal ecotype (Eastman, 1983) where decades of forest harvesting have caused extensive modifications to the landscape (Kuzyk, 2016). The forests contain stands of all ages, from recently logged clear cuts to maturing plantations and uncut forests. The uncut forests are dominated by coniferous forests of hybrid white spruce (Picea engelmannii x Picea glauca) and subalpine fir (Abies lasiocarpa). Secondary successional sites are pioneered by lodgepole pine (Pinus contorta var. latifolia) and trembling aspen (Populus tremuloides) (Meidinger and Pojar, 1991). The area has a humid continental climate which is generally wet and cool, with precipitation evenly distributed throughout the year. Mean daily average temperatures are 4.3 ºC and range from a mean daily average of –7.9 ºC in January, to a mean daily average of 15.8 ºC in July. Mean annual precipitation is 595 mm, with 205 cm of it falling as snow (Government of Canada, 2016). Reported bear densities near our study area were between 0.10–0.27/km2 for black bears and 0.012–0.049/km2 for grizzlies (Mowat et al., 2002) prior to this study and appear to have remained relatively stable to the present (District Contact, Ministry of Forests, Lands and Natural Resource Operations and Rural Development MFLNRORD, unpublished data).

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Wolf densities in our study area are 10–44/1000 km2 (British Columbia MFLNRORD, 2014). Moose populations in north–central British Columbia (BC) around Prince George (PG) have declined from a density of 1.3 moose/km2 during the late 1990s (Heard et al., 1999) and early 2000s (Walker et al., 2006) to 0.45 moose/km2 in 2017 (Klaczek et al., 2017). Other large carnivores inhabiting our study area include wolves (Canis lupus), and cougars (Puma concolor) (Kuzyk et al., 2018). In addition to moose, large ungulates in the area include mule deer (Odocoileus hemionus), white–tailed deer (Odocoileus virginianus), elk (Cervus canadensis), and caribou (Rangifer tarandus) (Kuzyk et al., 2018). Field collections Bear scats were collected opportunistically along roads and trails in 2014 (16 May–27 October) and 2015 (12 May–30 July). We collected bear scat samples in the later summer and fall of 2014 because we wanted to know if bears were eating older calves. We did not collect scats in August, September or October of 2015 because samples collected during this period in 2014 contained no hairs from moose calves, which was consistent with the literature. Many of the roads and trails from which we collected were sampled in both years. In May and June of 2015, we added collections from areas where collared cow moose were calving in the John Prince Research Forest (JPRF). There, we collected 58 scats with an aim towards comparing differences in the frequency of occurrence of calf hairs in scats located inside of known calving areas (a biased sample) to the rest of the study area. Samples were collected and stored in Ziploc® freezer bags in the field. We collected what we considered to be fresh, recently deposited scats, but also collected some slightly older looking (e.g., covered in dust, dried out) samples. Our samples likely contained scats of both black and grizzly bears, but we did not attempt to distinguish between the two. A geo–reference and date of collection were recorded on each sample bag. All samples were stored in a freezer (–20ºC) until analysis. Scat analysis In the laboratory, scats were thawed, homogenized by hand mixing, and the piles divided into two equal portions. One portion was inspected for the presence of remains from moose calves, while the other portion was refrozen for use as reference material. The scat portions examined for calf remains were autoclaved for 60' at 121 ºC (Schwab et al., 2011). The autoclaved samples were then washed over a 1mm sieve screen for 10'' to 15'' until the rinse water ran clear. We tested the sieve screen with known moose calf hairs to ensure hairs were retained by the sieve. Washed scats were then put into paper bags and oven dried (70 ºC) to a constant weight. Finally, scat contents were disentangled manually and examined for the presence of hair, bones, hooves, and other body parts.

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Hairs and body parts were separated from the washed materials and sorted with tweezers. Species identification was difficult from bone fragments and hooves without DNA analyses or reference materials, so we focused our study on species identification using hair impressions as accomplished by others (Perrin, 1980; Cashman et al., 1992; Iversen et al., 2013; Popp et al., 2018). When hairs were present, we extracted all hairs from the scats and placed and stored them in petri dishes for imprinting and analysis. Our protocol for identifying hair was modified from the methods of Kennedy and Carbyn (1981). Hairs were initially separated into color and size types by visual inspection at 100x magnification with a compound microscope. We then made imprints of the cuticular scale patterns of all separate hairs (or one or two from tufts of hair) by placing hairs between two pieces (89 mm x 139 mm) of thermal laminating plastic (Swingline® GBC®, Lincolnshire, Illinois), clamping them between two microscope slides, and heating them at 120 ºC in a drying oven (Fisher Scientific Model V602G, Dubuque Iowa) for 160''. The same technique was used to make cuticular scale cast reference standards for hairs that we extracted from several dozen mammal study skins (including adult moose) previously collected in north–central BC and housed in a reference collection at the University of Northern BC. In addition, impressions were made of hairs from a calf moose collected in Jasper National Park, Alberta (The University of BC, Beaty Biodiversity Museum Catalogue # M000922 Collector: Ian McTaggart Cowan, 1944–MAY–27). All hair scale impressions were then examined under a compound microscope at 400x. Hairs were identified to species by comparing the impressions observed in thermoplastic film with known standards (Williamson, 1951). Adult and calf (neonate to three months old) moose hairs were distinguished using hair size, color, and cuticular patterns. Statistical analysis We applied statistical comparisons only to those samples (n = 1,319) collected in the May through July period that was consistent between the two years of collections. The x2–test (Gould and Gould, 2002) was used to compare the proportions of scats with and without hair from moose calves collected in 2014 and 2015. The 2014 and 2015 records were combined and the x2–test was used to compare the proportions of scats containing hairs from moose calves between the months of May, June, and July. For each of these three months, we reported the proportion of scats that contained hairs from moose calves. Dates of sample collections were recorded to day of the year where January 1 = day 1 and December 31 = day 365. The two–sample Wilcoxon rank–sum (Mann–Whitney) test was used to determine whether there was a difference between 2014 and 2015 in the days of the year when scats containing hairs from moose calves were collected. Significance of these statistical tests was assessed at α = 0.05 using STATA 12 (StataCorp, College Station, TX).

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124 ºW

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Fig. 1. Map of north–central British Columbia showing the locations of bear scats collected during May to October of 2014 and May to July of 2015: ×, scats without hairs from moose calves; Í, scats containing hairs from moose calves (n = 27) collected in 2014 (n = 11); Æ, scats containing hairs from moose calves collected in 2015 (n = 16). (x2 indicates symbols representing two nearby scats). Fig. 1. Mapa de la zona centroseptentrional de la Columbia Británica en el que se muestran las localizaciones de las heces de osos recolectadas entre mayo y octubre de 2014 y entre mayo y julio de 2015: ×, heces sin pelos de becerros de alce; Í. heces con pelos de becerros de alce (n = 27) recolectadas en 2014 (n = 11); Æ heces con pelos de becerros de alce recolectadas en 2015 (n = 16). ("x2" indica la presencia de dos heces cercanas).

Spatial relationships between the scats containing hairs from moose calves were described using the 'Generate Near Table' tool in ArcGIS (Version 10.4.1, ESRI 2016, Redlands, California) to calculate the distance between each point and its nearest neighbor. We described geographic clustering (grouping) of scats containing calf hairs using a circular buffer with a 6–km radius (to mimic 116 km2 home range; Young and Ruff, 1982). Results In addition to plant matter, 213 of our 1,319 scat samples contained hairs from various species of mammals, including bears, squirrels, hares, muskrat, various rodents, and ungulates (e.g., deer, elk, and moose; Reichert and Rea, unpublished data). Twenty–seven of the scats (2.05 %) collected in the May through July periods of 2014 and 2015 contained hair from moose calves. The frequency of occurrence of moose calf

hairs within bear scats increased from May to July of each year (fig. 1). Two scats collected on June 3, 2015 contained hairs from adult moose (0.15 %). In 2014, 11 of the 510 scats (2.16 %) collected from May to July contained hair from moose calves, while in 2015, 16 of the 809 scats (1.98 %) collected in the same months contained hair(s) from moose calves. There were no significant differences (x2 = 0.0500, df = 1, P = 0.823) in the frequency of occurrence of scats with and without hairs from moose calves collected between May (peak calving May 24, D. Aitken, unpublished data) through July 2014 and 2015. The proportion of scats containing moose calf hairs was significantly different (x2 = 7.1387, df = 2, P = 0.028) between the months of May, June, and July (fig. 2). The greatest proportion of scats with hairs from moose calves were collected in July (3.2 %) with fewer in June (2.3 %) and May (0.3 %). None of the 62 scats collected in the August to October 2014 period contained hairs from moose. Additionally, none of the 58 bear scat samples collected in June of

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4 3.5 3

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Fig. 2. Frequency of occurrence of scats containing moose calf hairs (out of every 100 scats collected) from May through July 2014 and 2015. Fig. 2. Frecuencia de heces con pelos de becerro de alce (por cada 100 heces recolectadas) entre mayo y julio de 2014 y 2015.

2015 from the six calving grounds in the John Prince Research Forest contained calf hairs. Calf hairs were observed in bear scats collected as early as 28 May (day 148) and as late as 30 July (day 211). There were no significant differences (z = 1.262, P = 0.2071) in the May, June, and July dates of collection of scats with hairs from moose calves between 2014 (n = 11) and 2015 (n = 16). Median date of collections of bear scats containing calf hairs for the two years was June 24 (day 175). Several of the scats with calf hair were geographically clustered (fig. 3). Five of these clusters were located approximately 90 km north of Prince George, while another two clusters were located approximately 50 km east of Prince George. Seven of the 27 6–km diameter buffers we placed around scats with hairs contained a single scat with calf hair. Eleven of the 27 buffers did not overlap each other (fig. 3). Six locations on the map to the north and east of Prince George (includes overlapping buffers) contained more than 1 scat with calf hair (n = 1–5 scats) within a 6–km radius area (fig. 3). Discussion Our examination of bear scats from north–central BC showed that bears consumed moose calves in both years of study. The frequency of occurrence of bear scats containing calf hair that we observed (~2 %) was similar to that reported in Algonquin Park where most moose calf mortalities were attributable to black bears (Patterson et al., 2012) and where 1.7 % of bear digestive contents examined in spring and summer contained moose calf hairs (Wilton et al., 1984). This is lower than what was reported in Alaska by Chatelain (1950) where 5.7 % of scats contained moose

calf hair, some of which LeResche (1968) speculated may have been from scavenging. Although some of the moose calves consumed by bears in our study may have been scavenged (possibly from previous bear kills; Boertje et al., 1988), both Franzmann et al. (1980) and Boertje et al. (1988) suggest that bears primarily kill, rather than scavenge calves. Ballard et al. (1981) reported 90% of calf mortality due to bears occurred before 19 July. Furthermore, Larsen et al. (1989), and Boertje et al. (1988) reported that predation by bears on moose calves was focused between May and late July/early August, after which predation rates declined due to the decreased availability and increased mobility of calves or alternate food sources becoming more abundant and available to bears (Adams et al., 1995; Boertje et al., 1988; Zager and Beecham, 2006). Unlike LeResche (1968) who reported bear predation occurrence to be highest in May and June with decreases in July, we found increasing evidence of calf hair in scats from May (during first estrous calving) to July. We do know the exact dates that our bear scats were collected. We do not know, however, exactly when each scat was deposited by bears and, therefore, cannot be certain of the exact date of calf consumption. Since we drove and collected scats on some roads every two weeks, we feel confident that most scats collected in our study would have been less than two weeks old. Backdating all scats that we collected containing calf hair by one to two weeks, better aligns our findings with expected predation events by bears that have been reported to eat calves of various ages also born in May and June in other jurisdictions (Ballard et al., 1981; Boertje et al., 1988; Swenson et al., 2007). This was true of both years, since we found no statistically significant difference between the dates on which scats with calf hairs were collected.

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123 ºW

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Fig. 3. Map of locations of bear scats containing moose calf hairs collected during the springs and summers of 2014 and 2015 in north–central British Columbia. Each scat location is buffered with a 6 km radius buffer that was used to estimate bear home range size (see Material and methods) and show overlap of possible home ranges relative to moose calf predation events. (x2 indicates symbols representing two nearby scats). Fig. 3. Mapa de las localizaciones de heces de osos con pelos de becerro de alce recolectadas durante la primavera y el verano de 2014 y de 2015 en la zona centroseptentrional de la Columbia Británica. Cada localización está rodeada por un radio de 6 km que se utilizó para estimar el tamaño del territorio del oso (véase el apartado "Material and methods") y mostrar posibles superposiciones de territorios relacionadas con episodios de depredación de becerros de alce. ("x2" indica la presencia de dos heces cercanas).

Some bears specialize on killing calves (Boertje et al., 1988), with some taking up to one calf daily (French and French, 1990). We, therefore, examined scat locations for possible clusters of scats containing calf hairs. Because bears defecate up to eleven times per day (Roth, 1980), clustering could be indicative of a single bear eating one calf but leaving behind several separate scats with remains of a single calf. Our nearest neighbor analysis suggested that about one quarter of the scats that had moose calf hairs were geographically clustered to the north and east of Prince George. Three quarters of the 6–km home range buffers we placed around the scats with calf hairs included more than one (and up to 5) scat with hair, suggesting that individual bears could theoretically be responsible for multiple predation events on calves. The exact number of bears consuming calves could possibly be determined through DNA analysis of intestinal mucosa extracted from the surface of scats at the time of collection (Lonsinger et al., 2015), but was not an explicit objective of our study and was not within the budget constraints of the project.

None of the 58 scats collected in our six calving grounds to the northwest of Prince George were found to contain calf hairs. Even with this form of biased sampling, where we expected calves would more likely be eaten by bears, we observed lower than expected frequency of occurrences (0 %) compared to the entire study area. This could have, however, been affected by the timing of our surveys since our data suggest bears were feeding on calves somewhat later in the season (after cows and calves had left the calving grounds) than anticipated. We sampled primarily along roads and trails and were not able to document whether bears were eating calves far from roads and trails. Bear scat samples collected in Alberta with the use of scat detection dogs were highly concentrated on industrial–use roads (Wasser et al., 2004), suggesting our collections along roads should be fairly representative of scats (albeit not necessarily prey) from across the landscape. If moose are calving far from roads and trails and bears are also consuming calves far from roads and trails, we may be underestimating predation on calves. We may

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also be underestimating calf predation if the reference proportion of the scat we did not analyze, by chance, contained evidence of calf predation not found in the portion we did assess. Several jurisdictions that are multi–prey and multi–predator systems surrounding our study area have reported significant impacts of bears on moose calves (Alberta: Hauge and Keith, 1981; Yukon Territory: Larsen et al., 1989; Alaska: Bertram and Vivion, 2002). Consequently, we have no reason to believe that these findings do not apply to north–central BC given the relatively high density of bears. Therefore, alternative explanations for our findings of a small number of scats with moose calf hairs may be simply that the ratio of calves to bears is low, that other food types (mostly vegetation) are much more common in spring and summer diets of bears, or that our collection procedures were unable to capture an unevenly distributed prey base. We determined that moose calves (n = 27) appear to comprise a much larger proportion of the spring and summer bear diet than adult moose (n = 2). If we scale our findings of 0.15 % of bear scats containing adult moose hairs to a recently published mortality study of adult cow–only moose where 6.6 % of cows were killed by bears (Mumma and Gillingham 2019), our 2 % of scats with calf hairs could suggest bears may be responsible for up to 88 % of calf mortalities in north–central BC. This is not unreasonable given the role that predators such as bears can have on moose calves (56–100 % of moose calf mortalities; Zager and Beecham, 2006), but is likely an overestimation given bears preferentially prey on adult cow moose that are pregnant or defending calves (seven cows to one bull moose killed by bears; Boertje et al., 1988). Developing a robust technique to age scats (e.g., travelling roads and trails and cleaning off all scats on a weekly basis) would help to pinpoint calf consumption dates better. Determining the species and individual identity of each bear from each scat and delineating the full range of diet items using genetic techniques for both could allow managers to determine the relative importance of moose calves in the diets of individual bears. However, DNA techniques would not distinguish between calf and adult moose, underscoring the importance of establishing baseline data from the present use of hair analysis. These parameters, combined with accurate estimates of: seasonal defecation rates of bears, moose and bear densities, spring and summer cow:calf ratios, moose to bear (both black and grizzly) ratios, and moose and bear home range sizes and overlap, could all help to establish more precisely the impacts of bears on moose. Finally, determining predation rates of bears on moose calves might help managers determine the overall impact of bears on a declining moose population, allowing them to weigh options about what, if anything, can be done to mitigate that impact. Acknowledgements We are grateful to Michael van Dijk, Lori Johnson, Caleb Sample, Corin MacPhail, Jonathan Shaw, Sandy

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van Dijk, Brittney Reichert, Brendan Carswell, Simon Harris, and Lena Richter for helping us process the scats and identify hairs and to John Orlowsky and Doug Thompson for space in the Enhanced Forestry Laboratory at UNBC. We thank the CNC Research Forest, the North–Central Guide Outfitters Association, Spruce City Wildlife Association, the John Prince Research Forest, the Vanderhoof Fish and Game Club, the Aleza Lake Research Forest and UNBC for funding the research. We are grateful to Shannon Crowley, Peter Forsythe, Gabrielle Aubertin, Warren Neuvonen, Hardy Griesbauer, Dan Ryan, Colin Chisholm, Samantha Pederson, Karl Domes, Dave Powe, Caroline Seip, Caslin Rea, Clara Temoin, Perry Temoin, Melissa Mjolsness, Leigh Anne Dutton, and Hardy Greisbauer for help with collections. Thanks to Chris Stinson at the Beaty Biodiversity Museum for providing moose calf hairs. We thank the Coordenação de Aperfeiçoamento de Pessoal de Nível Superior (Capes) for L. Ajala–Batista's scholarship in Canada. We are also grateful to Emygdio L. A. Monteiro Filho and Shelley Marshall and two anonymous referees for comments that greatly improved our manuscript. Our sincere thanks to Susana Aguilar for Spanish translations. References Adams, L. G., Singer F. J., Dale, B. W., 1995. Caribou calf in Denali National Park. Journal of Wildlife Management, 59(3): 584–594. Ballard, W. B., 1992. Bear predation on moose: a review of recent North American studies and their management implications. Alces Supplements, 1: 1–15. Ballard, W. B., Spraker, T. H., Kenton, P. T., 1981. Causes of Neonatal Moose Calf Mortality in South Central Alaska. The Journal of Wildlife Management, 45(2): 335–342. British Columbia MFLNRORD (British Columbia Ministry of Forests, Lands and Natural Resource Operations), 2014. Management Plan for the Grey Wolf (Canis lupus) in British Columbia. British Columbia Ministry of Forests, Lands and Natural Resource Operations, Victoria, BC. Bertram, M. R., Vivion, M. T., 2002. Moose mortality in eartern interior Alaska. The Journal of Wildlife Management, 66: 747–756. Boertje, R. D., Gasaway, W. C., Granaard, D. V., Kelleyhouse, D. G., 1988. Predation on moose and caribou by radio–collared grizzly bears in east central Alaska. Canadian Journal of Zoology, 66: 2492–2499. Boertje, R. D., Gasaway, W. C., Grangaard, D. V., Kelleyhouse, D. G., Stephenson, R. O., 1987. Factors limiting moose population growth in Game Management Unit 20E. Alaska Department of Fish and Game, Alaska. Brockman, C. J., Collins, W. B., Welker, J. M., Spalinger, D. E., Dale, B. W., 2017. Determining kill rates of ungulate calves by brown bears using neck–mounted cameras. Wildlife Society Bulletin, 41(1): 88–97.

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Cashman, J. L., Peirce, M., Krausman, P. R., 1992. Diets of mountain lions in southwestern Arizona. The Southwestern Naturalist, 37(3): 324–326. Chatelain, E. F., 1950. Bear–moose relationships on the Kenai Peninsula. Transactions of the North American Wildlife Conference, 15: 224–233. Costello, C. M., Cain, S. L., Pils, S., Frattaroli, L., Haroldson, M. A., Manen, F. T., 2016. Diet and macronutrient optimization in wild ursids: a comparison of grizzly bears with sympatric and allopatric black bears. Plos One, 11(5): e0153702, Doi:10.1371/ journal.pone.0153702 Crête, M., Courtois, R., 1997. Limiting factors might obscure population regulation of moose (Cervidae: Alces alces) in unproductive boreal forests. Journal of Zoology, 242(4): 765–781. De Barba, M., C. Boyer, M. F., Mercier, C., Rioux, D., Coissac, E., Taberlet, P., 2014. DNA metabarcoding multiplexing and validation of data accuracy for diet assessment: application to omnivorous diet. Molecular Ecology Resources, 14: 306–323. Eastman, D., 1983. Seasonal changes in crude protein and lignin of ten common forage species of moose in north–central British Columbia. Alces, 19: 36–70. Franzmann, A. W., Schwartz C. C., Peterson, R. O., 1980. Moose calf mortality in Summer on the Kenai Peninsula, Alaska. The Journal of Wildlife Management, 44(3): 764–768. French, S. P., French, M. G., 1990. Predatory behavior of grizzly bears feeding on elk calves in Yellowstone National Park, 1986–88. Bears: Their Biology and Management, 8: 335–341. Garneau, D. E., Boudreau, T., Keech, M., Post, E., 2007. Black bear movements and habitat use during a critical period for moose calves. Mammalian Biology, 73: 85–92. Gasaway, W. C., Boertje, R. D., Grangaard, D. V., Kelleyhouse, D. G., Stephenson, R. O., Larsen, D. G., 1992. The role of predation in limiting moose at low densities in Alaska and Yukon and implications for conservation. Wildlife monographs, 120: 3–59. Gould, J. L., Gould, G. F., 2002. Biostats Basics. W. H. Freeman and Company, New York. Gorley, A., 2016. A strategy to help restore moose populations in British Columbia. Ministry of Forests, Lands and Natural Resource Operations Fish and Wildlife Branch. Government of Canada, 2016. Canadian Climate Normals 1981–2010, http://climate.weather.gc.ca/ climate_normals/results_1981_2010_e.html?searchType=stnProv&lstProvince=BC&txtCentralLatMin=0&txtCentralLatSec=0&txtCentralLongMin=0&txtCentralLongSec=0&stnID=631&dispBack=0 [Accessed on March 2018]. Hamer, D., Herrero, S., 1987. Grizzly Bear Food and Habitat in the Front Ranges of Banff National Park, Alberta. Bears: Their Biology and Management, 7: 199–213. Hatler, D. F., 1972. Food habits of black bears in interior Alaska. Canadian Field–Naturalist, 86: 17–31. Hauge, T. M., Keith, L. B., 1981. Dynamics of moose populations in northeastern Alberta. The Journal of Wildlife Management, 45(3): 573–597.

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bear densities in interior British Columbia. Aurora Wildlife Research, Prince George. Mumma, M. A., Gillingham, M. P., 2019. Determining factors that affect survival of moose in central British Columbia. Technical Report to the Habitat Conservation Trust Foundation. Munro, R. H. M., Nielsen, S. E., Price, M. H., Stenhouse, G. B, Boyce, M. S., 2006. Seasonal and diel patterns of grizzly bear diet and activity in west–central Alberta. Journal of Mammalogy, 87(6): 1112–1121. Patterson, B. R., Benson, J. F., Middel, K. R., Silver, A., Obbard, M., 2012. Moose Calf Mortality in Central Ontario, Canada. The Journal of Wildlife Management, 77(4): 832–841. Perrin, B. S., 1980. Key to the mammals of the Andries Vosloo Kudu Reserve (Eastern Cape) based on their hair morphology, for use in predator scat analysis. South African Journal of Wildlife Research, 10: 1–14. Popp, J. N., Hamr, J., Larkin, L. J., Mallory, F. F., 2018. Black bear (Ursus americanus) and wolf (Canis spp.) summer diet composition and ungulate prey selectivity in Ontario, Canada. Mammal Research, 63: 433–441. Roth, H. U., 1980. Defecation Rates of Captive Brown Bears. Bears: Their Biology and Management, 4: 249–253. Schwab, C., Cristescu, B., Northrup, J. M., Stenhouse, G. B., Gänzle, M., 2011. Diet and environment shape fecal bacterial microbiota composition and enteric pathogen load of grizzly bears. Plos One, 6(12): e27905, Doi:10.1371/journal.pone.0027905. Singer, F. J., Harting, A., Symonds, K. K., Coughenour, M. B., 1997. Density dependence, compensation and environmental effects on elk calf mortality in

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Yellowstone National Park. The Journal of Wildlife Management, 61(1): 12–25. Swenson, J. E., Dahle, B., Busk, H., Opseth, O. L. E., Johansen, T., Söderberg, A., Wallin, K., Cederlund, G., 2007. Predation on moose calves by European brown bears. Journal of Wildlife Management, 71(6): 1993–1997. Thiemann, G. W., 2008. Using fatty acid signatures to study bear foraging: technical considerations and future applications. Ursus, 19(1): 59–72. Voigt, D. R., Kolenosky, G. B., Pimlott, D. H., 1976. Changes in Summer Foods of Wolves in Central Ontario. The Journal of Wildlife Management, 40(4): 663–668. Walker, A. B. D., Heard, D. C., Michelfelder, V., Watts, G. S., 2006. Moose density and composition around Prince George, British Columbia. Government of Canada, Prince George. Wasser, S. K., Davenport, B., Ramage, E. R., Hunt, K. E., Parker, M., Clarke, C., Stenhouse, G., 2004. Scat detection dogs in wildlife research and management: application to grizzly and black bears in the Yellowhead Ecosystem, Alberta, Canada. Canadian Journal of Zoology, 82: 475–492. Williamson, V. H. H., 1951. Determination of hairs by impressions. Journal of Mammalogy, 32(1): 80–84. Wilton, M. L., Carlson, D. M., McCall, C. I., 1984. Occurrence of neonatal cervids in the spring diet of black bear in south central Ontario. Alces, 20:95–105. Young, B. F., Ruff, R. L., 1982. Population Dynamics and Movements of Black Bears in East Central Alberta. The Journal of Wildlife Management, 46(4): 845–860. Zager, P., Beecham, J., 2006. The role of American black bears and brown bears as predators on ungulates in North America. Ursus, 17(2): 95–108.

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Evidence of character displacement in microhabitat use between two tropical sympatric Holcosus lizard species (Reptilia, Teiidae) F. Ruiz Real, J. Martín

Ruiz Real, F., Martín, J., 2019. Evidence of character displacement in microhabitat use between two tropical sympatric Holcosus lizard species (Reptilia, Teiidae). Animal Biodiversity and Conservation, 42.2: 379–388, Doi: https://doi.org/10.32800/abc.2019.42.0379 Abstract Evidence of character displacement in microhabitat use between two tropical sympatric Holcosus lizard species (Reptilia, Teiidae). Interspecific competition between sympatric related species leading to character displacement is critical for species coexistence, especially in tropical habitats. We examined microhabitat use of two sympatric species of tropical lizards of the genus Holcosus in relationship to the microhabitats available in two ecosystems. The species H. festivus lives exclusively in the forest and uses microhabitats in proportion to their availability; while the other, H. quadrilineatus, lives both in forest and on the beach and selects microhabitats with specific characteristics. In the ecosystem where these two lizards live in sympatry (forest), we observed a differential microhabitat use between the two species. However, these differences indicated changes in habitat use by H. quadrilineatus (the smaller species) concerning its patterns of habitat selection in the ecosystem (beach) where only this species occurs. The age of the lizards did not affect the patterns of selection of microhabitats of either species. Shifts in microhabitat use may allow coexistence in sympatry of both species, which might result from the competitive exclusion of the smaller species by the larger species. Key words: Interspecific competition, Holcosus, Lizards, Microhabitat use, Tropical habitats Resumen Evidencia del desplazamiento de caracteres en el uso de microhábitats entre dos especies simpátridas de lagartos tropicales del género Holcosus (Reptilia, Teiidae). La competencia interespecífica entre especies simpátricas relacionadas que conduce al desplazamiento de caracteres es crucial para la coexistencia de las especies, en especial en hábitats tropicales. Examinamos el uso de los microhábitats disponibles en dos ecosistemas por dos especies simpátricas de lagarto tropical del género Holcosus. La especie H. festivus vive exclusivamente en el bosque y utiliza microhábitats en proporción a su disponibilidad, mientras que la otra, H. quadrilineatus, vive tanto en el bosque como en la playa, y selecciona microhábitats con características específicas. En el ecosistema donde estos dos lagartos viven en simpatría (el bosque), observamos un uso diferente de microhábitats entre ambas especies. Sin embargo, estas diferencias indicaron cambios en el uso del hábitat de H. quadrilineatus (la especie más pequeña) con respecto a sus patrones de selección de hábitat en el ecosistema (la playa) donde solo está presente esta especie. La edad de los lagartos no afectó a los patrones de selección de microhábitat de ninguna especie. Los cambios en el uso de los microhábitats pueden permitir la coexistencia en simpatría de ambas especies, lo que podría ser consecuencia de la exclusión competitiva de la especie más pequeña por la más grande. Palabras clave: Competencia interespecífica, Holcosus, Lagartos, Uso del microhábitat, Hábitats tropicales Received: 04 III 19; Conditional acceptance: 17 IV 19; Final acceptance: 01 VII 19 F. Ruiz Real, J. Martín, Departamento de Ecología Evolutiva, Museo Nacional de Ciencias Naturales–CSIC, José Gutiérrez Abascal 2, 28006 Madrid, España (Spain). Corresponding author: Fátima Ruiz. E–mail: fatimaleal.3@gmail.com ISSN: 1578–665 X eISSN: 2014–928 X

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Introduction Interspecific competition has an important effect driving evolutionary and ecological diversification because when two or more species undergo intense competition for similar resources such as food or habitat (Begon et al., 1996; Dhont, 2011), they can diverge by ecological character displacement (Day and Young, 2004; Stuart and Losos, 2013). Interspecific competition may be especially important in tropical habitats due to the high biodiversity of species found in sympatry with similar ecological niches. For example, many studies examining the different interspecific interactions in the structure of the community of lizards in Neotropical biomes have found evidence of competition (Vitt et al., 2000a, 2000b, Hatano et al., 2001, Rocha et al., 2009). An assumption of the theories of competition is that the strength of between–species competition is inversely related to the amount of interspecific resource partitioning (Pacala and Roughgarden, 1982). Thus, when studying the interactions between pairs of similar species, it is necessary to quantify the degree of overlap by use of resources (Hurlbert, 1978). It has been shown that two species of sympatric lizards could coexist in the same place and time if there were a series of ecological parameters that would confer differentiating characteristics in, for example, the way of feeding, activity times, morphology, or physiology (Huey, 1979; Chase et al., 2002). However, the distinct spatial occupation on the available microhabitats within a given ecosystem is considered a critical factor determining the coexistence of sympatric species (Pianka, 1973; Schoener, 1974; Calsbeek, 20098), and it may also explain speciation processes (Losos, 2009). In addition, ontogenetic changes in size or any ecological aspect should be accounted for because these may differentially affect the intensity of potential competition, which may be age–specific (e.g., Smith 1981). Through these studies, it will be possible to determine factors that condition the life of the species, being able to approximate the impact that human activity may cause them (Böhm et al., 2013). In many tropical forests of Central America, two similar species of lizards of the genus Holcosus (formerly Ameiva) (Family Teiidae) co–occur (Savage, 2002; Abella et al., 2008). Holcosus festivus (also known as the Middle American ameiva or tiger ameiva) is a large lizard that reaches a total length of 34.5 cm. It inhabits mainly forest areas, with higher activity at midmorning on sunny days. Holcosus quadrilineatus (the four–lined ameiva) is smaller, reaching a total length of 28.3 cm, and it is found in open areas, forest margins and in clearings, where it is more active in the morning (Savage, 2002). Information on the ecology of these two species of lizards focuses on their temperature preferences (Hirth, 1965), thermoregulation and activity patterns (Vitt and Zani, 1996; Sebastián–González and Gómez, 2012), reproduction (Smith, 1968a, 1968b; Fitch, 1973), parasitology (Bursey et al., 2006), escape responses (Lattanzio, 2014), diet (Hirth, 1963; Whitfield and Donnelly, 2006), and other more general ecological aspects (Hirth, 1963;

Hillman, 1969; Fitch, 1973; Vitt and Zani, 1996). However, few studies are examining the possible competitive interactions between the two species (Sebastián–González and Gómez, 2012). Here, we investigate the partition of microhabitats between these two lizard species (H. festivus and H. quadrilineatus) relative to their availability and analyze whether ecological displacement to avoid interspecific competition may explain their coexistence in the same areas. We estimated the degree of overlap or ecological distribution in terms of microhabitat use between these two species that inhabit the same area (Pacuare Natural Reserve, Costa Rica). Holcosus quadrilineatus occupies two types of ecosystems (beach and forest) (fig. 1), whereas H. festivus occurs only in the forest area. We determined the structural characteristics of the microhabitats occupied by each species in the two types of ecosystems in relation to the microhabitats available in each area. Specifically, we aimed to determine: (1) whether there was a selection of specific microhabitats with respect to those available, or whether the habitat was simply used as a function of its availability; (2) whether there was a differential microhabitat use between species in those areas where they were sympatric; and (3) if there were such differences, whether this implied changes in habitat use with respect to the patterns of habitat selection in the areas where only one of the species was found. Finally, we examined (4) the effect of age on these patterns of habitat use and the possible competitive interactions between species. Material and methods Study area The study was performed in the Pacuare Natural Reserve, Costa Rica (10º 13' 50'' N / 83º 16' 72'' W to the north, 10º 12' 50'' N / 83º 3' 22'' W to the south). The Reserve comprises 800 ha of rainy secondary forest and swampy areas, surrounded by the Tortuguero channel system and the Caribbean Sea. More than 300 vertebrate species, including 52 reptile species, are found in the reserve. There are two research stations, a northern station and a southern station joined by a path system (Abella et al., 2008). The Reserve has two ecosystems: (1) a rainy secondary forest (fig. 1A); and (2) a beach area with a volcanic sand substrate (fig. 1B), located in the southern station. Therefore, the observations were divided according to the type of ecosystem (forest vs beach). In this reserve, only one study on thermoregulation and activity patterns of these Holcosus lizard species has been performed (Sebastián–González and Gómez, 2012), so knowing more about their ecology will allow improvement in the management and conservation of these species in the environment of the reserve. Data sampling The study was carried out at the beginning of the rainy season, during 15 days of May 2017. The reserve is divided by several trails: a main trail that connects the

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Fig. 1. Views of the two main ecosystems, a) forest and b) beach, used by Holcosus lizards in the study area. Fig. 1. Vista de los dos principales ecosistemas, a) el bosque y b) la playa, utilizados por los lagartos del género Holcosus en la zona de estudio.

southern station with the northern station, and several secondary trails. Every day, if the weather favored lizard activity, we slowly walked the trail system of the reserve looking for lizards during their hourly peaks of higher activity (morning: 06.30–09.00 h; midday: 10.00–12.30 h; afternoon: 15:00–17.00 h) (Savage, 2002). Surveys covered the beach area, and three sections of the forest, with a similar search effort in all the areas. The surveys followed trails that were repeated on different days, so there could be a potential repetition in the sighting of the same individual lizards. However, given that the abundance of lizards was high, and since we avoided repeating sampling in a specific area (a minimum distance of 4 m between observations was established), we are confident that the probability of repeated observations of the same individuals was low and did not affect the results. Every time that an individual lizard was sighted, we recorded the hour of the day and the species, and calculated the age of the individual. We distinguished adults from juveniles by their body size and morphological traits; H. quadrilineatus juveniles are less than 6 cm in length and H. festivus are less than 7–8 cm in length (Savage, 2002). Additionally, juveniles of H. quadrilineatus have two pairs of lateral and ventrolateral yellow stripes and blue tails (brown in adults), and juveniles of H. festivus have a vertebral bright blue stripe that extends from the tip of the snout almost to the tip of the tail; this changes to white or yellow with age (Savage, 2002). Then, we measured microhabitat use of the lizard by taking four transects of 2 m each, one at each of the four cardinal orientations (N, S, E, W) radiating from the point where the individual was first sighted. In each transect, the characteristics of the microhabitat were noted at four points at 50, 100, 150 and 200 cm from the central point (Martín and López, 1998). At each point, we noted the substrate type (sand, leaf litter, grass, bare soil or tree trunk). We also noted whether

there was tree cover above the point, and whether sun exposure wasshade or sunny. Furthermore, we assessed the moisture content of the substrate (dry or wet). Finally, we noted the contacts of a bar placed vertically with the vegetation at heights of 5, 10, 25 and 50 cm above each point, and the type of vegetation in each contact (Ipomoea sp. 1, Ipomoea sp. 2, Coccoloba sp., Hybiscus sp., Heliconia sp., or Fam. Rubiaceae, Piperaceae, Cyperaceae, Gramineae or Palmae). In this way, we obtained microhabitat data for 16 points around the observation point for each lizard, allowing us to calculate the percentages of the cover of each of the 23 microhabitat variables for each observation. We obtained 83 observations of lizards (50 H. quadrilineatus and 33 H. festivus). To study the availability of microhabitats, we used the same procedure to measure microhabitats at points (N = 35) determined at random while walking following straight line paths (a minimum distance of 4 m was established between points), in the same paths of the study area used to study lizards. Data analyses We performed a principal components analysis (PCA) to reduce the number of microhabitat variables (23) to a small number of principal components (PCs) that described the habitat. The original data (i.e. number of contacts with each microhabitat variable) were square–rooted to ensure that they were fitted to a normal distribution. We used analyses of variance (ANOVAs) to compare differences in the use of the different types of microhabitats described by the PCA (PC scores) between the points available in the habitat and those used by each of the species in each of the ecosystems (forest vs beach). Post–hoc tests (Tukey's tests) were used for pairwise comparisons. These analyses were carried out using the R Studio statistical software (R Core Team, 2017).

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Results Table 1 shows the average values for each of the microhabitat variables available and used by each species in each type of ecosystem (forest vs beach). The PCA performed with all microhabitat variables produced six principal components (PCs) with eigenvalues greater than 1, which together accounted for 64.4 % of the explained variance (table 2). The first component (PC1) represented a gradient from microhabitats with high cover of sand to areas with high cover of leaf litter. The second component (PC2) described a gradient from microhabitats with high grass cover to areas with high cover of tall (50 cm) bush vegetation. The third component (PC3) represented a gradient from dry to humid substrates. The fourth component (PC4) represented a gradient toward microhabitats with high cover of bush vegetation of 25 cm in height above the ground. The fifth component (PC5) described a gradient towards microhabitats with palms and the presence of tree trunks at the substrate level. Finally, the sixth component (PC6) represented a gradient from areas with high cover of Hybiscus sp. to areas with high cover of gramineous herbs. There were significant differences between groups (microhabitats available and used by each species in each type of ecosystem) in most PCs, except in the case of PC4 and PC5 (table 2). Regarding the PC1 (fig. 2A), H. quadrilineatus inhabiting beaches selected microhabitats with less leaf litter than expected given the average availability of leaf litter in the beach environment (Tukey's test, p = 0.0045). Also in the forest, although leaf litter was significantly more abundant here than on the beach (p < 0.0001), H. quadrilineatus selected substrates with significantly less leaf litter than what was available on average (p = 0.002). However, in the forest, H. quadrilineatus used microhabitats with significantly more leaf litter than those used by this species on the beach (p < 0.0001). In contrast, H. festivus used leaf litter in proportion to its availability in the forest (p = 0.92). This pattern resulted in significant differences between the two lizard species in the use of these substrates within the forest (p < 0.008), with H. festivus using microhabitats with more leaf litter than H. quadrilineatus. In relation to the PC2 (fig. 2B), H. quadrilineatus on the beach used microhabitats with grass or tall bushes according to what was expected given their availability (Tukey's test, p = 0.96). However, in the forest, although the availability of grass and tall bushes was similar to that on the beach (p = 0.28), H. quadrilineatus selected microhabitats with significantly much more grass cover and fewer bushes than what was available on average (p < 0.0001). Therefore, H. quadrilineatus used microhabitats with significantly more grass and fewer tall bushes in the forest than on the beach (p < 0.0001). In contrast, H. festivus used microhabitats with grass and bushes in the forest in proportion to their availability (p = 0.91). Between the two species, there were significant differences in the forest (p < 0.0001); H. quadrilineatus selected microhabitats with significantly more grass cover and fewer tall bushes than those used by H. festivus.

Regarding the PC3 (fig. 2C), on the beach H. quadrilineatus used substrates with humidity levels similar to those available on average in the environment (Tukey's test, p = 0.22). However, in the forest, this species used substrates that were significantly dryer than those available on average (p = 0.002). Therefore, H. quadrilineatus shifted its use of microhabitats depending on the ecosystem, using significantly more dry substrates in the forest than on the beach (p = 0.025) even though the average humidity of available substrates was significantly higher in the forest than on the beach (p = 0.04). H. festivus, on the other hand, used substrates with humidity levels similar to those available on average in the forest (p = 0.29). However, there were no significant interspecific differences in the use of substrates in relationship to their humidity within the forest (p = 0.19). For PC4 (fig. 2D) and PC5 (fig. 2E), there were no significant differences between groups. Thus, these variables (PC4, contact with vegetation at 25 cm; PC5, palms and tree trunks) did not seem to affect microhabitat selection and both species used them in proportion to availability in both ecosystems. In the case of PC6 (fig. 2F), both species used what was available in the environment (Tukey's tests, H. quadrilineatus: beach p = 0.13, forest p = 0.62, H. festivus: forest p = 0.18), which did not significantly differ between forest and beach (p = 0.97). Thus, there were no significant differences between species in the use of microhabitats with Hybiscus sp. or gramineous herbs within the forest (p = 0.94). Habitat use according to age classes After partitioning the observations according to the age of the lizards (juvenile vs adult), there were no significant differences between ages, within each species and habitat, in the use of microhabitats defined by any of the PCs (Tukey's tests, p > 0.40 in all between ages comparisons). That is, both juveniles and adults of the same species followed the same patterns of microhabitat use within each ecosystem. This result indicated that age did not affect the selection of microhabitats of each species, and therefore, age should not influence the possible competition between species. Discussion Considering the results of the availability and use of the microhabitats, we found that the two Holcosus lizard species differed in their general strategies of microhabitat use. While H. festivus, which was found only in the forest, used different microhabitats according to what was expected given their availability, H. quadrilineatus showed in many cases a distinct use of microhabitats relative to their availability. Thus, microhabitat variables such as leaf litter and grass cover, tall bushes and substrate humidity levels seem to affect the use of microhabitat by H. quadrilineatus.

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Table 1. Variables (mean ± SE) that characterize the microhabitats available and used by H. quadrileneatus (Hq) and H. festivus (Hf) lizards in two types of ecosystems (beach and forest). Tabla 1. Variables (media ± EE) que caracterizan los microhábitats disponibles y utilizados por H. quadrileneatus (Hq) y H. festivus (Hf) en dos tipos de ecosistema (la playa y el bosque).

Beach

Forest

Available Hq Available Hq

N = 15

N = 25

N = 20

N = 25

Hf N = 33

Substrate Sand (%)

84.2 ± 6.3

79.1 ± 5.9

–

Leaf litter (%)

12.9 ± 6.1

4.5 ± 2.1

88.4 ± 5.2

84.0 ± 3.7

88.3 ± 4.3

Grass (%)

–

10.0 ± 4.7

9.8 ± 4.8

9.2 ± 3.4

4.9 ± 3.1

Bare soil (%)

–

0.9 ± 0.7

6.0 ± 1.9

3.9 ± 1.6

Tree trunk (%)

2.9 ± 1.0

6.4 ± 2.2

0.9 ± 0.5

0.8 ± 0.5

2.9 ± 1.2

Dry (%)

47.1 ± 12.3

40.5 ± 9.7

10.0 ± 6.9

76.0 ± 8.7

27.3 ± 7.9

Humid (%)

52.9 ± 12.3

59.5 ± 9.7

90.0 ± 6.9

25.0 ± 8.7

72.7 ± 7.9

Sunny (%)

16.7 ± 8.1

27.8 ± 8.1

2.8 ± 2.8

25.8 ± 6.6

25.8 ± 4.2

Vegetation contacts 5 cm

3.8 ± 1.9

11.3 ± 3.1

2.5 ± 1.1

4.0 ± 1.1

2.8 ± 0.9

10 cm

9.1 ± 2.1

14.3 ± 3.4

10.9 ± 3.7

10.5 ± 2.5

4.7 ± 1.3

25 cm

12.1 ± 2.3

5.0 ± 1.1

16.5 ± 3.6

4.3 ± 1.2

8.9 ± 1.9

50 cm

5.8 ± 2.1

6.8 ± 1.9

10.3 ± 2.6

1.5 ± 0.7

10.8 ± 2.4

Tree cover (%)

56.3 ± 10.7

11.0 ± 4.4

91.3 ± 5.5

21.3 ± 8.1

99.2 ± 0.8

Vegetation type Ipomoea sp. 1

4.6 ± 2.0

16.5 ± 4.4

–

Ipomoea sp. 2

2.5 ± 1.8

5.3 ± 1.6

–

Coccoloba

19.2 ± 5.5

13.1 ± 2.4

–

Hybiscus

3.3 ± 2.2

4.2 ± 1.7

Heliconia

–

Rubiaceae

–

Piperaceae

–

Cyperaceae

–

3.4 ± 1.1

1.3 ± 0.5

7.0 ± 2.2

–

7.5 ± 2.4

–

7.0 ± 2.0

–

11.3 ± 4.0

–

4.7 ± 1.6

–

6.6 ± 3.0

16.8 ± 3.6

6.8 ± 2.2

Gramineae

–

6.3 ± 3.7

1.3 ± 0.8

0.7 ± 0.5

Palmae

–

0.2 ± 0.2

–

0.9 ± 0.6

Interestingly, some of these patterns of microhabitat selection of H. quadrilineatus were different in each ecosystem, but the differences in availability did not always seem to explain these shifts in microhabitat use. Thus, higher availability of leaf litter in the forest may clearly explain that H. quadrilineatus used microhabitats with more leaf litter here than on the beach. However, H. quadrilineatus used dryer substrates with more grass and fewer tall bushes in the forest than in the beach even though the availability of these microhabitats did not change (for bushes) or changed following the opposite pattern (e.g. hu-

midity) between ecosystems. If we consider that the patterns of microhabitat use on the beach ( where only H. quadrilineatus was found) were optimal for this species, the shifts observed in the forest populations may be explained by the interspecific competition with the sympatric H. festivus, which results in a competitive displacement to suboptimal microhabitats. With these shifts in habitat use, there was no ecological overlap in the forest between species in the use of microhabitats regarding grass or bush cover. Similar shifts in habitat use or displacement to sub–optimal microhabitats have been studied in detail in popu-

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Table 2. Principal components analysis for variables describing available microhabitats used by H. quadrileneatus and H. festivus lizards. Correlations in bold lettering correspond to variables significantly correlated with a PC (p < 0.0001). Results (F, p) of ANOVAs comparing PC scores between microhabitats available and used by each lizard species are shown. Tabla 2. Análisis de componentes principales para las variables que describen los microhábitats disponibles y utilizados por los lagartos H. quadrileneatus y H. festivus. Los valores en negrita corresponden a las correlaciones significativas de las variables con un componente principal (PC en su sigla en inglés; p < 0,0001). Se muestran los resultados (F, p) de los ANOVAS que comparan las puntuaciones de cada PC entre los microhábitats disponibles y los utilizados por cada especie de lagarto. Variable

PC–1 PC–2 PC–3 PC–4 PC–5 PC–6

Substrate Sand

–0.36 0.26 –0.05 0.14 –0.01 0.11

Leaf litter

0.40 –0.16 –0.02 –0.12 0.04 –0.07

Grass

–0.08

Bare soil

0.05 –0.29 –0.09 0.03 0.07 0.14

Tree trunk

–0.15 0.13 0.13 –0.12 0.45 0.15

Dry

–0.09

Humid Sunny

–0.37 0.30 0.14 –0.40 0.01

–0.18

–0.34 0.40 0.30 –0.21

0.08

0.22

0.33

–0.38

–0.30

0.21

0.02

–0.11

0.07

–0.17

–0.05

–0.19

Vegetation contacts 5 cm

–0.25

–0.21

0.34

0.01

–0.02

–0.16

10 cm

–0.14

–0.23

25 cm

0.14

0.37

0.20

0.04

0.10

0.25

0.44 –0.03 0.14

50 cm

0.20

0.34

0.17

0.29 –0.02 –0.17

Tree cover

0.35

–0.16

–0.07

–0.14

0.09

0.12

Vegetation type Ipomoea sp. 1

–0.30

0.01

0.27

–0.13

–0.06

–0.25

Ipomoea sp. 2 –0.26 0.05 0.17 –0.17 0.30 0.03 Coccoloba

–0.14 0.21 –0.13 0.28 –0.15 0.40

Hybiscus

–0.15 0.10 –0.04 0.15 –0.27 –0.47

Heliconia

0.24 0.07 0.16 0.17 –0.04 –0.03

Rubiaceae

0.27 0.20 0.20 0.01 0.09 –0.26

Piperaceae

0.24 0.16 0.22 0.09 0.23 0.06

Cyperaceae

0.06 –0.41 0.15 0.15 –0.07 0.04

Gramineae

0.11 –0.05 0.18 0.24 –0.07 0.36

Palmae

0.02 0.10 0.12 –0.02 0.60 –0.08

Eigenvalues

4.32 3.22 2.64 2.07 1.32 1.23

% variance explained

18.8

14.0

11.5

9.0

5.7

5.3

85.84

18.31

4.75

1.58

0.74

2.75

< 0.0001

ANOVAs: F4,113 p

0.0014 0.19 0.57 0.032

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A B Leaf litter

0 -1

-2 Sand

-3 -4

3 2

1 PC–2

PC–1

Contacts at 50 cm

Grass Avail. Hq Avail. Hq Hf Beach Forest

b,c

0 -1

-2 -3

Avail. Hq Avail. Hq Hf Beach Forest

C D 2 c

PC–3

Dry

Contacts at 2 25 cm

b,c

1 a,b,c

a,b

PC–4

Humid

-1

Avail. Hq Avail. Hq Hf Beach Forest E F Tree trunks 1 Palmae

Gramineae

0 -0.5 -1

-1

Hybiscus Avail. Hq Avail. Hq Hf Beach Forest

Avail. Hq Avail. Hq Hf Beach Forest

1 PC–6

PC–5

0.5

-1

c b,c a,b,c

a,b

Avail. Hq Avail. Hq Hf Beach Forest

Fig. 2. Mean (± SE) principal component (PC) scores from a PC analysis of all variables describing available microhabitats (open boxes) used by H. quadrilineatus (black boxes) and H. festivus (grey boxes) in each ecosystem (beach or forest). Means with the same letter above the bars did not differ significantly (p > 0.05) in post–hoc pairwise Tukey tests. Fig. 2. Media (± EE) de las puntuaciones de los PC resultantes de un análisis de componentes principales de todas las variables que describen los microhábitats disponibles (cajas blancas) y los utilizados por H. quadrilineatus (cajas negras) y H. festivus (cajas grises) en cada ecosistema (playa o bosque). Las medias con la misma letra encima no presentaron diferencias significativas (p > 0,05) en las pruebas de Tukey realizadas para comparar pares de medias a posteriori.

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lations of Anolis lizard species. For example, in the Bermudas, Anolis leachi is behaviorally dominant over A. grahami, so that A. grahami has to use perches of smaller diameter and lower height (Schoener, 1975; Losos, 1996). In this case both species can coexist because the subordinate, A. grahami, adapts to habitats unavailable to A. leachi. In Gran Cayman Island, the native A. conspersus has changed perch height and uses higher perches in open areas where it is sympatric with the introduced A. sagrei. In this case, competitive interactions are an improvement for ecological segregation among species (Losos et al., 1993). In another situation, native A. carolinenesis uses higher perches when its introduced congeneric competitor A. sagrei is present, suggesting that A. sagrei displaces the native A. carolinensis (Echternacht, 1999; Edwards and Lailvaux, 2012). As explained above, these differences in the use of microhabitats, which may allow the coexistence of two similar species that could otherwise be potential competitors, may be due to past or current competitive pressures (Schoener, 1975; Medel et al., 1988; Vanhooydonck et al., 2000; Vitt et al., 2000b; Kolbe et al., 2008). Nevertheless, the observed differential use of microhabitats may be a result of the different ecological requirements of each species (Huey et al., 1974: Barbault et al., 1985; Castilla and Bauwens, 1991; Daly et al., 2008). However, if the specific ecological requirements were the only explanation for the microhabitat use of H. quadrilineatus, then the same patterns or trends of habitat use would be observed in both types of ecosystems, but we found in contrasting opposite patterns between the forest and the beach in some cases. Other authors argue that the differential use of microhabitats may just be a reflection of differences in habitat availability (Johnson et al., 2006). However, our data showed that differences in availability between ecosystems alone could not explain all the shifts in microhabitat use observed in H. quadrilineatus. Within the forest, however, both species use microhabitats with similar humidity conditions, even though H. quadrilineatus used dryer microhabitats here than on the beach, which might reflect that this species was trying to avoid overlap in this factor too. Other microhabitat variables did not differ between species. Thus, there could be competence for some aspects of the microhabitat or these microhabitat factors might not be necessary for the ecological requirements of these lizards. Also, it is essential to consider environmental conditions when examining the existence of interspecific competition because when the availability of a required resource is high, competition between coexisting species may be weak (Paterson et al., 2018). Previous studies on the specificity of habitat use of both Holcosus lizard species indicated that H. quadrilineatus uses open areas that have less plant cover and greater sun exposure than H. festivus, with this species being found on the edges of forests, less exposed to the sun and at a lower temperature (Hillman, 1969). This finding reinforces our results, since in our study area, H. quadrilineatus was found both on the beach and in the forest, whereas H. festivus was found only in the forest. The studies carried out by

Sebastian–González and Gómez (2012) in the same Reserve indicated that H. festivus was also present in small numbers in the beach area, in contrast with our data. This difference may be due to the time of the year or the location (north or south research station) in which the different studies were conducted. Differences in size related to thermal biology between the two species can play a role in the differential use of habitats. Larger species have slower heating rates, but also a higher risk of overheating due to the slower cooling rate (Rubalcaba et al., 2019; Sebastian–González and Gómez, 2012). Also, larger species maintain higher body temperature for longer due to the lower convective heat dissipation and greater thermal inertia, which makes the loss of heat slower (Carrascal et al., 1992). Therefore, they will have a higher risk of overheating in warm regions (Rubalcaba et al., 2019), and will be forced to select more closed (shady) microhabitats than smaller species. Similarly, in some species of Sceloporus lizards, an effect of habitat structure and temperature of the environment has been observed in the selection of microhabitats (Adolph, 1990). Therefore, differences in habitat use might be only a result of differences in size–related thermal ecology. Body size differences can also be decisive in interspecific competition, as larger animals would be dominant (Miller, 1967; Schoener, 1983), displacing the smallest ones (Schoener, 1983; Tokarz, 1985; Losos, 1996; Sacchi et al., 2009). For example, Liolaemus lizard species show displacement in the use of microhabitats when they are in sympatry L. tenius restricts its microhabitat use (it was found only on trees and high perches) in the presence of L. pictus, which is large–bodied (Medel et al., 1988). Further studies should examine the use of microhabitats of both lizard species in other areas with similar habitats but where they do not coincide to determine whether they continue to use the same microhabitats or differ when they are in sympatry in the same locality. Such findings would reinforce our conclusion that these differences or similarities are due to the current competitive ecological interactions. It would also be of value to study the entire ecological niche (activity, diet, thermoregulation) of each species in situations of parapatry and sympatry, in order to obtain more information on their ecological and biological needs and to determine whether there is competition in any other of these ecological aspects. Finally, behavioral studies testing whether there are aggressive interactions or avoidance behavior would be important to discriminate the mechanisms involved in the competition between these species (Jenssen et al., 1984; Hess and Losos, 1991; Korner et al., 2000; Lailvaux et al., 2012). Acknowledgements We thank Sidney F. Gouveia and two anonymous reviewers for helpful comments, the Pacuare Reserve for allowing this study, and the teachers of the 'Master in Biodiversity in Tropical Areas and their Conservation' (UIMP–CSIC) for their advice and support.

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Medel, R. G., Marquet, P. A., Jaksic, F. M., 1988. Microhabitat shifts of lizards under different contexts of sympatry: a case study with South American Liolaemus. Oecologia, 76: 567–569. Miller, R. S., 1967. Pattern and process in competition. Advances in Ecological Research, 4: 1–74. Pacala, S., Roughgarden, J., 1982. Resource partitioning and interspecific competition in two two– species insular Anolis lizard communities. Science, 217: 444–446. Paterson, J. E., Weiss, S. L., Blouin–Demers, G., 2018. Experimental removal reveals only weak interspecific competition between two coexisting lizards. Canadian Journal of Zoology, 96: 888–896. Pianka, E. R., 1973. The structure of lizard communities. Annual Review of Ecology and Systematics, 4: 53–74. R Core Team, 2017. R: A Language and Environment for Statistical Computing. R Foundation for Statistical Computing, Vienna, Austria. Rocha, C. F. D., Sluys, M. Van., Vrcibradic, D., Kiefer, M. C., de Menezes, V. A., Siqueira, C. C., 2009. Comportamento de termorregulação em lagartos brasileiros. Oecologia Australis, 13: 115–131. Rubalcaba, J. G., Gouveia, S. F., Olalla–Tárraga, M. A., 2019. A mechanistic model to scale up biophysical processes into geographical size gradients in ectotherms. Global Ecology and Biogeography, 28(6): 793–803, doi:10.1111/geb.12893 Sacchi, R., Pupin, F., Gentilli, A., Rubolini, D., Scali, S., Fasola, M., Galeotti, P., 2009. Male–male combats in a polymorphic lizard: residency and size, but not color, affect fighting rules and contest outcome. Aggressive Behavior: Official Journal of the International Society for Research on Aggression, 35: 274–283. Savage, J. M., 2002. The Amphibians and Reptiles of Costa Rica. The University of Chicago Press, Chicago. Schoener, T. W., 1974. Resource partitioning in ecological communities. Science, 185: 27–39. – 1975. Presence and absence of habitat shift in some widespread lizard species. Ecological Monographs, 45: 233–258.

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– 1983. Field experiments on interspecific competition. The american naturalist, 122: 240–285. Sebastián–González, E., Gómez, R., 2012. Differences in habitat use of two sympatric species of Ameiva in East Costa Rica. Acta Herpetologica, 7: 347–353. Smith, D. C., 1981. Competitive interactions of the striped plateau lizard (Sceloporus virgatus) and the tree lizard (Urosaurus ornatus). Ecology, 62: 679–687. Smith, R. E., 1968a. Experimental evidence for a gonadal–fat body relationship in two teiid lizards (Ameiva festiva, Ameiva quadrilineata). The Biological Bulletin, 134: 325–331. – 1968b. Studies on reproduction in Costa Rican Ameiva festiva and Ameiva quadrilineata (Sauria: Teiidae). Copeia, 1968: 236–239. Stuart, Y. E., Losos, J. B., 2013. Ecological character displacement: glass half full or half empty? Trends in Ecology and Evolution, 28: 402–408. Tokarz, R. R., 1985. Body size as a factor determining dominance in staged encounters between male brown anoles (Anolis sagrei). Animal Behaviour, 33: 746–753. Vanhooydonck, B., Van Damme, R., Aerts, P., 2000. Ecomorphological correlates of habitat partitioning in Corsican lacertid lizards. Functional Ecology, 14: 358–368. Vitt, L. J., Sartorius, S. S., Avila–Pires, T. C. S., Esposito, M. C., Miles, D. B., 2000b. Niche segregation among sympatric Amazonian teiid lizards. Oecologia, 122: 410–420. Vitt, L. J., Souza, R. A., Sartorius, S. S., Avila–Pires, T. C. S., Espósito, M. C., 2000a. Comparative ecology of sympatric Gonatodes (Squamata: Gekkonidae) in the western Amazon of Brazil. Copeia, 2000: 83–95. Vitt, L. J., Zani, P. A., 1996. Ecology of the lizard Ameiva festiva (Teiidae) in southeastern Nicaragua. Journal of Herpetology, 30: 110–117. Whitfield, S. M., Donnelly, M. A., 2006. Ontogenetic and seasonal variation in the diets of a Costa Rican leaf–litter herpetofauna. Journal of Tropical Ecology, 22: 409–417.

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Diet to tissue discrimination factors for the blood and feathers of the monk parakeet (Myiopsitta monachus) and the ring–necked parakeet (Psittacula krameri) D. Mazzoni, N. A. Borray–Escalante, A. Ortega–Segalerva, L. Arroyo, J. González–Solís, J. C. Senar Mazzoni, D., Borray–Escalante, N. A., Ortega–Segalerva, A., Arroyo, L., González–Solís, J., Senar, J. C., 2019. Diet to tissue discrimination factors for the blood and feathers of the monk parakeet (Myiopsitta monachus) and the ring–necked parakeet (Psittacula krameri). Animal Biodiversity and Conservation, 42.2: 389–395, Doi: https://doi.org/10.32800/abc.2019.42.0389 Abstract Diet to tissue discrimination factors for the blood and feathers of the monk parakeet (Myiopsitta monachus) and the ring–necked parakeet (Psittacula krameri). Stable isotope analyses (SIAs) have been widely used in recent years to infer the diet of many species. This isotopic approach requires using diet to tissue discrimination factors (DTDFs) for each prey type and predator tissue, i.e., to determine the difference between the isotopic composition of the predator tissues and the different prey that conform its diet. Information on DTDF values in Psittaciformes is scarce. The aim of this study was to assess DTDF values for the carbon and nitrogen isotopes of the monk parakeet (Myiopsitta monachus) and the ring–necked parakeet (Psittacula krameri), two invasive alien species of concern. We fed captive birds of the two parakeet species on a single–species diet based on sunflower seeds to establish the DTDFs for the blood and feathers. In the monk parakeet (N = 9) DTDFs were Δδ13C 2.14 ‰ ± 0.90 and Δδ15N 3.21 ‰ ± 0.75 for the blood, and Δδ13C 3.97 ‰ ± 0.90 and Δδ15N 3.67 ‰ ± 0.74 for the feathers. In the ring–necked parakeet (N = 9), the DTDFs were Δδ13C (‰) 2.58 ± 0.90 and Δδ15N (‰) 2.35 ± 0.78 for the blood, and Δδ13C 3.64 ‰ ± 0.98 and Δδ15N 4.10 ‰ ± 1.84 for the feathers. DTDF values for the ring–necked parakeet blood were significantly higher than those for the monk parakeet blood. No difference was found between the two species in the DTDF for feathers. Our findings provide the first values of DTDFs for blood and feathers in these parakeets, factors that are key to infer the diet of these species based on SIA. Key words: Stable isotopes, Parrots, Captive birds, Control diet, Invasive species Resumen Factores de discriminación entre la dieta y los tejidos para la sangre y las plumas de la cotorra argentina (Myiopsitta monachus) y la cotorra de Kramer (Psittacula krameri). Los análisis de isótopos estables se han venido utilizando de forma generalizada en los últimos años para inferir la dieta de numerosas especies. El método isotópico consiste en utilizar factores de discriminación entre la dieta y los tejidos para cada tipo de presa y tejido de depredador, es decir, determinar la diferencia entre la composición isotópica de los tejidos del depredador y la de las distintas presas que integran su dieta. Hay muy poca información sobre los valores de los factores de discriminación entre la dieta y los tejidos relativa a los psitaciformes. La finalidad de este estudio fue evaluar los valores de los factores de discriminación de los isótopos de carbono y nitrógeno de la cotorra argentina (Myiopsitta monachus) y la cotorra de Kramer (Psittacula krameri), que son dos especies exóticas invasivas motivo de preocupación. Alimentamos a aves en cautividad de las dos especies de cotorra con una dieta exclusivamente a base de semillas de girasol para establecer los factores de discriminación relativos a la sangre y las plumas. En la cotorra argentina (N = 9), los factores de discriminación para la sangre fueron: Δδ13C 2,14 ‰ ± 0,90 y Δδ15N 3,21 ‰ ± 0,75 y para las plumas: Δδ13C 3,97 ‰ ± 0,90 y Δδ15N 3,67 ‰ ± 0,74. Los factores de discriminación de la cotorra de Kramer (N = 9) para la sangre fueron: Δδ13C (‰) 2,58 ± 0,90 y Δδ15N (‰) 2,35 ± 0,78 y para las plumas: Δδ13C 3,64 ‰ ± 0,98 y Δδ15N 4,10 ‰ ± 1,84. Los valores de los factores de discriminación entre la dieta y la sangre de la cotorra de Kramer fueron significativamente más elevados que los de la cotorra argentina. Con respecto a las plumas, no se observaron diferencias entre las dos especies. Nuestros ISSN: 1578–665 X eISSN: 2014–928 X

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resultados proporcionan los primeros valores de los factores de discriminación para la sangre y las plumas en cotorras, que son fundamentales para inferir la dieta de estas especies mediante el análisis de isótopos estables. Palabras clave: Isótopos estables, Cotorras, Aves en cautividad, Dieta de control, Especies invasoras Received: 02 I 19; Conditional acceptance: 08 IV 19; Final acceptane: 02 VII 19 D. Mazzoni, N. Borray, A. Ortega–Segalerva, L. Arroyo, J. C. Senar, Evolutionary and Behavioural Ecology Research Unit, Museu de Ciències Naturals de Barcelona, Psg. Picasso s/n., 08003 Barcelona, Espanya (Spain).– N. Borray, J. González–Solís, Institut de Recerca de la Biodiversitat (IRBio) and Department de Biologia Evolutiva, Ecologia i Ciències Ambientals, Facultat de Biologia, Universitat de Barcelona, Av. Diagonal 643, Barcelona 08028, Espanya (Spain). Corresponding author: D. Mazzoni. E–mail: daniele.mazzoni91@gmail.com

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Introduction Diet is a key aspect in ecological studies (Newton and Brockie, 1998). In the last few decades, analyses of stable isotopes of carbon and nitrogen have been widely used as ecological tracers to assess the diet of several species (Tieszen et al., 1983; Hobson and Clark, 1992b; Becker et al., 2007; Bearhop and Inger, 2008). Nevertheless, deriving quantitative estimates of dietary contributions needs precise estimates of the difference in isotopic composition between predator tissues and preys (diet to tissue discrimination factor, DTDF) (Bond and Diamond, 2011; Parnell et al., 2013). Recently, researchers have documented considerable variation in DTDF computations. Some reviews and meta–analyses summarized this variation as a function of a number of environmental and physiological factors that included environment (terrestrial, freshwater, marine), trophic level, taxon, tissue, metabolic rate (poikilotherm, homeotherm), nitrogenous excretion (ammonia, urea, uric acid), and sample treatment procedures (Dalerum and Angerbjörn, 2005; Caut et al., 2008; Phillips et al., 2014). As the DTDF is highlysensitive to these factors and is species–specific, it needs to be estimated for various taxa (McCutchan et al., 2003; Caut et al., 2009). However, there is a lack of information for the Psittaciformes order, to date computed only for the kea Nestor notabilis (Greer et al., 2015). The aim of this study was to estimate the DTDF of 13 C and 15N, the two most common stable isotopes used to estimate the diet of a species, for the blood and feathers of the monk parakeet (Myiopsitta monachus) and the ring–necked parakeet (Psittacula krameri). These two parakeets are currently widespread in Europe and other parts of the world and are considered to cause severe damage to agriculture, ornamental vegetation, and human facilities (Menchetti and Mori, 2014; Senar et al., 2016). DTDFs obtained from this study will allow the development of mixing models to investigate the diet of the species. Such information may be fundamental in controlling their populations. Material and methods The birds used in this study were sampled in the city of Barcelona. The two species have been present in the area since the early 1970s (Batllori and Nos, 1985). The population of the monk parakeet reached 5,000 individuals in 2015, while the population of the rose–ringed parakeet reached 340 individuals in 2014 (Senar et al., 2017a, 2017b). Avian samples We collected nine monk parakeets and nine rose–ringed parakeets of different ages and both sexes at the Ciutadella Park (Barcelona city) during autumn (September) 2017 and winter (January) 2018, after the end of their reproductive period. The birds were captured with a modified Yunnick trap located on a terrace of the Museu de Ciències Naturals de Barcelona. Monk parakeets were housed in captivity at the museum for

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90 days and rose–ringed parakeets for 75 days. As DTDFs differ among food sources, throughout the entire captivity period we fed parakeets with a single–species diet based on sunflower seeds. We provided this single food source for the entire study period to allow the isotope ratios in the parakeet tissues to equilibrate with their control diet (Kelly et al., 2012). Blood sampling was conducted biweekly by puncturing the brachial vein and collecting blood in an Eppendorf tube. This should allow the verification of whether the feathers grown after the acclimatization period reflect the isotope composition of sunflower seeds and whether they allow a reliable approximation of DTDF. We verified this by analyzing the stabilization of isotopes in the blood throughout the experiment. Breast feathers were removed at day 45 (monk parakeet) and 30 (rose–ringed parakeet) of the experiment, and hence were forced to grow under the same conditions for all individuals. We chose to remove the feathers after six weeks because complete blood turnover occurs generally in about 30 days (Pearson et al., 2003). New induced feathers were again removed after six weeks when they were fully grown and while they were still on the sunflower seed diet. To calculate the DTDF of the blood in both species, we used samples of the last biweekly puncturing, at the end of the experiment, for both species. Stable isotope analyses As the lipid component of the diet can affect δ13C values (Tarroux et al., 2010; Perkins et al., 2013), five samples of sunflower seeds were previously washed of lipids using successive rinses in a 2:1 chloroform: methanol solvent, and were later oven–dried at 50 ºC for 24 h. Feather samples were cleaned in a 0.1M NaOH solution to prevent contamination, oven–dried at 50 ºC for 24 hours and ground into a fine powder. Blood samples were lyophilized and ground to a fine powder, and then directly analyzed isotopically. We did not perform lipid extraction on whole–blood or feather samples because of the typically low proportion of lipids in bird blood and feathers (Bearhop et al., 2002) We did not apply any linear normalization equation (LNE) to the lipid content of the blood. In fact, when the ratio C:N is less or around 4, as we found in our samples, it is not necessary to take the lipid content of the samples for terrestrial animals into account (Post et al., 2007). The C:N ratios for the blood were 3.31 ± 0.03 and 3.40 ± 0.17 for the monk parakeet and the ringed–necked parakeet respectively. The stable isotope analyses were performed at the 'Serveis Científico–Tècnics' at the University of Barcelona. Stable carbon and nitrogen isotope assays for all samples were weighed using a micro–balance Mettler Toledo MX5 and placed in tin cups and crimped for combustion, using an elemental analysis–isotope ratio mass spectrometry (EA–IRMS). Stable isotope abundances are expressed in δ notation in parts per thousand (‰) according to: δX = [(Rsample/Rstandard) – 1] * 1,000 where X is

C or

N and R is the corresponding

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Table 1. Mean isotopic values (± SD) of δ13C (‰) and δ15N (‰) for the diet item (N = 5), and for blood samples (N = 9) and regrown feathers (N = 9) for monk and rose–ringed parakeets after being fed in captivity with a single–source diet based on sunflower seeds over 90 and 75 days, respectively. Tabla 1. Promedio de los valores isotópicos (± DE) de δ13C (‰) y δ15N (‰) relativos a la dieta (N = 5) y a las muestras de sangre (N = 9) y las plumas nuevas (N = 9) de las cotorras argentina y de Kramer tras haber sido alimentadas en cautividad con una fuente de alimentación exclusivamente a base de semillas de girasol durante 90 y 75 días, respectivamente.

Sunflower seeds

Monk parakeet Blood

Feathers

δ C (‰)

–27.04 ± 0.89

–24.90 ± 0.12

–23.07 ± 0.16

δ N (‰)

4.06 ± 0.73

7.27 ± 0.18

7.73 ± 0.13

13 15

C/12C or 15N/14N ratio. The standard values for 13C and 15N were Pee Dee Belemnite (PBD) and atmospheric nitrogen (AIR), respectively. Measurement errors (SD) were ± 0.15 ‰ for δ13C and ± 0.25 ‰ for δ15N. After allowing sufficient time for equilibration between the tissue and the control diet, the DTDF can be estimated by a calculation: 13

∆δXtissue = δXtissue – δXdiet i.e. the average isotopic value from the tissue minus the average isotopic value from the food. We estimate the variability of this calculation (standard deviation, SD) as the square root of summed variances for diet and tissues samples. (Giménez et al., 2016). We carried out Kruskal–Wallis tests in JMP (SAS Institute Inc., Cary, North Carolina) to assess whether there were any statistically significant differences in the DTDFs of the tissues between the two species as well as between tissues within the same species. To estimate the patterns of carbon and nitrogen turnovers, we applied the exponential model: y = a + beCT where y represents the isotope value of the tissue in question, a and b are parameters determined by initial and asymptotic conditions, C is the turnover rate of carbon in the tissue, and T is time (d) since the diet switch (Tieszen et al., 1983; Hobson and Clark, 1992a; Suzuki et al., 2005) in JMP (SAS Institute Inc., Cary, North Carolina). Results Table 1 shows the results of our isotopic analyses for the samples of sunflower seed, blood, and regrown feathers. The analysis of the stabilization of isotopes in the blood throughout the experiment provided a low fit to the exponential model (R2 < 0.20 in all cases). Half–life turnover rates for ringed–necked parakeets were estimated in 25.5 ± 20.1 days for δ13C and 21.3 ± 17.7 days for δ15N, whereas those

Rose–ringed parakeet Blood

Feathers

–24.46 ± 0.11 –23.40 ± 0.40 6.41 ± 0.28

8.16 ± 1.69

for monk parakeets were 52.2 ± 35.3 days for δ13C and 10.4 ± 4.31 days for δ15N. However, the curves approached an asymptote before the feathers were removed (fig. 1). Birds of both species maintained on the sunflower seed diet showed remarkably consistent isotope values for blood after this stabilization (fig. 1). Since we obtained suitable isotopic dietary shifts and stabilization was reached before inducing regrowth, for both the monk and the rose–ringed parakeets, we proceeded to estimate DTDF for each species (table 2). The two species differed in blood DTDF values for both isotopes (Kruskal–Wallis test: δ13C x2(1) = 12.79, p < 0.001; δ15N x2(1) = 12.79, p < 0.001). However, the DTDF values of the feathers did not differ significantly between the two species (Kruskal– Wallis test: δ13C x2(1) = 2.38, p = 0.12; δ15N x2(1) = 1.22, p = 0.26). The comparison of DTDF values between the two tissues within the same species resulted in statistically significant differences for both isotopes (monk parakeet blood vs. feathers: δ13C x2(1) = 12.79, p < 0.001; δ15N x2(1) = 12.79, p < 0.001; rose–ringed parakeet blood vs. feathers: δ13C x2(1) = 12.79, p < 0.001; δ15N x2(1) = 5.48, p = 0.02). Discussion The stabilization of both isotopes in monk parakeet blood seems to be reached earlier than in ring–necked parakeets. This suggests that the diet of the individuals in the wild was similar to the diet provided in captivity for monk parakeets but not so similar for rose–ringed parakeets. As a consequence, stabilization in isotope blood values followed a flat trend, and hence, the exponential model was not the most suitable model to explain the turnover patterns. This is not surprising since the 'exponential–fit' approach overemphasizes data at the extremes (Cerling et al., 2007). The DTDFs for δ15N of the feathers in M. monachus (3.67 ± 0.74 ‰) and in P. krameri (4.10 ± 1.84) are close to the frequently adopted DTDF value of 3.40 ‰ per trophic level (Post, 2002). Comparisons of δ13C DTDF values obtained in this work with that of other

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-24.3

-24.5

-24.5 δ13C (‰)

-24.3

-24.9

δ13C (‰)

-24.9 -25.2

-25.2

-25.5

-25.5 15

-25.8 45 60 75 90 0 15 Day

9.5

8.5

8.5 δ15N (‰)

δ15N (‰)

-25.8 0

7.5 6.5

45 Day

7.5 6.5 5.5

5.5 4.5

4.5 45 60 75 90 0 15 Day

45 Day

Fig. 1. Change in isotopic values for δ13C and δ15N in the blood of captive monk (Myiopsitta monachus) (left) and ring–necked parakeets (Psittacula krameri) (right) switched on day 1 from the wild diet to a sunflower seed diet. The yellow dots represent the day on which we removed the feathers. The red dots represent the day we took the last blood samples, and the regrown feathers. We present element– turnover patterns for carbon and nitrogen by fitting a single value to the means of the combined data for all individuals for the diet on each extraction. Error bars represent the standard deviation of the mean for all experimental birds. (N = 9 for each species). Fig. 1. Cambio en los valores isotópicos de δ13C y δ15N en la sangre de la cotorra argentina (Myiopsitta monachus) (izquierda) y la cotorra de Kramer (Psittacula krameri) (derecha) en cautividad en el primer día posterior al cambio de la dieta natural por la dieta a base de semillas de girasol. Los puntos amarillos representan el día en que se extrajeron las plumas. Los puntos rojos representan el día en que se tomaron las muestras de sangre y las plumas nuevas. Presentamos los patrones de recambio del carbono y el nitrógeno ajustando un único valor al promedio de los datos combinados de todos los individuos para la dieta en cada extracción. Las barras de error representan la desviación estándar de la media de todas las aves del experimento. (N = 9 para cada especie).

species is obscured by the fact that some studies do not present the δ13C values with prior lipid extraction (Hobson and Bairlein, 2003). The need to extract lipids from the food source derives from the fact that the carbon from the lipid fraction of the diet may not contribute to the carbon content of proteins in blood or keratin in feathers, and therefore the differential lipid contents of food can contribute to variation in DTDF as well as inflate its value, since isotopic values of the lipid fraction are typically isotopically depleted (Perkins et al., 2013). Regardless, we found that the DTDFs for δ13C that we obtained for both species

(monk parakeet: 3.97 ± 0.90 ‰ and the rose–ringed parakeet: 3.64 ± 0.98 ‰) are among the highest values reported for this isotope (Caut et al., 2009), considerably exceeding the commonly assumed 0–1 ‰ DTDF value per trophic level (Post, 2002). Nevertheless, the high DTDF values for δ13C that we found was similar to that found by Greer et al. (2015) in the Kea parrot. This may result from the fact that, parrots generally have higher basal metabolic rates (BMRs) than other bird species (McNab, 2012). DTDF values are positively correlated to the respiration rate (De Niro and Epstein, 1978). The higher blood DTDF

Mazzoni et al.

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Table 2. Diet to tissue discrimination factors (Δ13C and Δ15N) for feathers and blood for monk (N = 9) and rose–ringed parakeets (N = 9) derived from a single–source diet based on sunflower seeds over 90 and 75 days, respectively. Tabla 2. Factores de discriminación (Δ13C y Δ15N) entre la dieta y las plumas y la sangre de la cotorra argentina (N = 9) y la cotorra de Kramer (N = 9) obtenidos a partir de una sola fuente de alimentación a base de semillas de girasol durante 90 y 75 días, respectivamente.

Δδ13C (‰)

Δδ15N (‰)

Feathers Monk parakeet

3.97 ± 0.90

3.67 ± 0.74

Rose–ringed parakeet

3.64 ± 0.98

4.10 ± 1.84

Blood Monk parakeet

2.14 ± 0.90

3.21 ± 0.75

Rose–ringed parakeet

2.58 ± 0.90

2.35 ± 0.78

in the rose–ringed parakeet compared to that of the monk parakeet could, therefore, be a consequence of ring–necked parakeets displaying a higher BMR than monk parakeets. The Δ13C and Δ15N values from the feathers of the two species were higher than those from their blood (table 2). The differences between Δ13C values from blood and feathers are probably due to the fact that the two tissues differ in biochemical composition (Martínez del Rio et al., 2009). Feathers typically have more positive Δ13C and Δ15N values than blood or muscle (Caut et al., 2009). In conclusion, our study provides the first values of DTDFs for the monk and the rose–ringed parakeets and contributes to the scarce knowledge of DTDF values for the order Psittaciformes. Acknowledgements We thank Jéssica Borrego and Pilar Rubio for their help in the laboratory analyses. We thank Victor Peracho and Tomás Montalvo, from the Agencia de Salut Pública de Barcelona, for their support. Birds were captured and handled with special permission EPI 7/2015 (01529/1498/2015) from the Direcció General del Medi Natural I Biodiversitat, Generalitat de Catalunya. The present study was funded by CGL–2016–79568–C3–3–P research project to JCS from the Spanish Research Council (Ministry of Economics and Competitiveness). References Batllori, X., Nos, R., 1985. Presencia de la Cotorritagris (Myiopsitta monachus) y de la Cotorrita de collar (Psittacula krameri) en el área metropolitana de Barcelona. Misc. Zool., 9: 407–411. Bearhop, S., Inger, R., 2008. Applications of stable

isotope analyses to avian ecology. Ibis, 150(3): 447–461. Bearhop, S., Waldron, S., Votier, S. C., Furness, R. W., 2002. Factors that influence assimilation rates and fractionation of nitrogen and carbon stable isotopes in avian blood and feathers. Physiological and Biochemical Zoology, 75(5): 451–458. Becker, B. H., Newman, S. H., Inglis, S., Beissinger, S. R., 2007. Diet–feather stable isotope (D15N and D13C) fractionation in Common murres and other seabirds. The Condor, 109(2): 451–456. Bond, A. L., Diamond, A. W., 2011. Recent Bayesian stable–isotope mixing models are highly sensitive to variation in discrimination factors. Ecological Applications, 21(4): 1017–1023. Caut, S., Angulo, E., Courchamp, F., 2008. Discrimination factors (Δ15N and Δ13C) in an omnivorous consumer: effect of diet isotopic ratio. Functional Ecology, 22(2): 255–263. – 2009. Variation in discrimination factors (Δ15N and Δ13C): The effect of diet isotopic values and applications for diet reconstruction. Journal of Applied Ecology, 46(2): 443–453. Cerling, T. E., Ayliffe, L. K., Dearing, M. D., Ehleringer, J. R., Passey, B. H., Podlesak, D. W., Torregrossa, A.–M., West, A. G., 2007. Determining biological tissue turnover using stable isotopes: the reaction progress variable. Oecologia, 151(2): 175–189. Dalerum, F., Angerbjörn, A., 2005. Resolving temporal variation in vertebrate diets using naturally occurring stable isotopes. Oecologia, 144(4): 647–658. De Niro, M. J., Epstein, S., 1978. Influence of diet on the distribution of carbon isotopes in animals. Geochimica et Cosmochimica Acta, 42: 495–506. Giménez, J., Ramírez, F., Almunia, J., Forero, M. G., Stephanis, R. de, 2016. From the pool to the sea: Applicable isotope turnover rates and diet to skin discrimination factors for bottlenose dolphins (Tursiops truncatus). Journal of Experimental Marine

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Biology and Ecology, 475: 54–61. Greer, A. L., Horton, T. W., Nelson, X. J., Kurle, C., 2015. Simple ways to calculate stable isotope discrimination factors and convert between tissue types. Methods in Ecology and Evolution, 6(11): 1341–1348. Hobson, K. A., Bairlein, F., 2003. Isotopic fractionation and turnover in captive Garden Warblers (Sylvia borin): Implications for delineating dietary and migratory associations in wild passerines. Canadian Journal of Zoology, 81(9): 1630–1635. Hobson, K. A., Clark, R. G., 1992a. Assessing Avian Diets Using Stable Isotopes I: Turnover of 13C in Tissues. The Condor, 94(1): 181–188. – 1992b. Assessing Avian Diets Using Stable Isotopes II: Factors Influencing Diet–Tissue Fractionation. The Condor, 94(1): 189–197. Kelly, D. J., Robertson, A., Murphy, D., Fitzsimons, T., Costello, E., Gormley, E., Corner, L. A. L., Marples, N. M., 2012. Trophic enrichment factors for blood serum in the European badger (Meles meles). Plos One, 7(12): e53071. Martínez del Rio, C., Wolf, N., Carleton, S. A., Gannes, L. Z., 2009. Isotopic ecology ten years after a call for more laboratory experiments. Biological Reviews, 84(1): 91–111. McCutchan, J. H., Lewis, W. M., Kendall, C., McGrath, C. C., 2003. Variation in trophic shift for stable isotope ratios of carbon, nitrogen, and sulfur. Oikos, 102(2): 378–390. McNab, B. K., 2012. Extreme measures: The ecological energetics of birds and mammals. University of Chicago Press, Chicago, Illinois. Menchetti, M., Mori, E., 2014. Worldwide impact of alien parrots (Aves Psittaciformes) on native biodiversity and environment: a review. Ethology Ecology & Evolution, 26(2–3): 172–194. Newton, I., Brockie, K., 1998. Population Limitation in Birds. Elsevier Science. Parnell, A. C., Phillips, D. L., Bearhop, S., Semmens, B. X., Ward, E. J., Moore, J. W., Jackson, A. L., Grey, J., Kelly, D. J., Inger, R., 2013. Bayesian stable isotope mixing models. Environmetrics, 10(1–2): 387–399. Pearson, S. F., Levey, D. J., Greenberg, C. H., Martínez del Rio, C., 2003. Effects of elemental composition on the incorporation of dietary nitrogen and carbon isotopic signatures in an omnivorous songbird. Oecologia, 135(4): 516–523.

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Telahigue et al.

Animal Biodiversity and Conservation 42.2 (2019)

Animal Biodiversity and Conservation Animal Biodiversity and Conservation és (abans Miscel·lània Zoològica) és una revista interdisciplinària publicada, des de 1958, pel Museu de Ciències Naturals de Barcelona. Inclou articles d'investigació empírica i teòrica en totes les àrees de la zoologia (sistemàtica, taxonomia, morfologia, biogeografia, ecologia, etologia, fisiologia i genètica) procedents de totes les regions del món. La revista presta una atenció especial als estudis que plantegen un problema nou o que introdueixen un nou tema, amb unes hipòtesis i prediccions clares i als treballs que d'una manera o altre tinguin rellevància en la biologia de la conservació. No es publicaran articles purament descriptius o articles faunístics o corològics que descriguin la distribució en l'espai o en el temps dels organismes zoològics. Aquests treballs s'han de redirigir a la nostra revista germana Arxius de Miscel·lània Zoològica (www.amz. museucienciesjournals.cat). Els estudis realitzats amb espècies rares o protegides poden no ser acceptats tret que els autors disposin dels permisos corresponents. Cada volum anual consta de dos fascicles. Animal Biodiversity and Conservation es troba registrada en la majoria de les bases de dades més importants i està disponible gratuitament a internet a www.abc.museucienciesjournals.cat, de manera que permet una difusió mundial dels seus articles. Tots els manuscrits són revisats per l'editor executiu, un editor i dos revisors independents, triats d'una llista internacional, a fi de garantir–ne la qualitat. El procés de revisió és ràpid i constructiu. La publicació dels treballs acceptats es fa normalment dintre dels 12 mesos posteriors a la recepció. Una vegada hagin estat acceptats passaran a ser propietat de la revista. Aquesta es reserva els drets d’autor, i cap part dels treballs no podrà ser reproduïda sense citar–ne la procedència. Els drets d’autor queden reservats als autors, els qui autoritzen la revista a publicar l’article. Els articles es publiques amb una Llicència de Reconeixement 4.0 Internacional de Creative Commons: no es podrà reproduir ni reutilitzar cap part dels treballs publicats sense citar-ne la procedència.

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Animal Biodiversity and Conservation 42.2 (2019)

Animal Biodiversity and Conservation Animal Biodiversity and Conservation (antes Miscel·lània Zoològica) es una revista interdisciplinar, publicada desde 1958 por el Museu de Ciències Naturals de Barcelona. Incluye artículos de investigación empírica y teórica en todas las áreas de la zoología (sistemática, taxonomía, morfología, biogeografía, ecología, etología, fisiología y genética) procedentes de todas las regiones del mundo. La revista presta especial interés a los estudios que planteen un problema nuevo o introduzcan un tema nuevo, con hipòtesis y prediccions claras, y a los trabajos que de una manera u otra tengan relevancia en la biología de la conservación. No se publicaran artículos puramente descriptivos, o artículos faunísticos o corológicos en los que se describa la distribución en el espacio o en el tiempo de los organismes zoológicos. Esos trabajos deben redirigirse a nuestra revista hemana Arxius de Miscel·lània Zoològica (www.amz. museucienciesjournals.cat). 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 registrada 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 garantizar su calidad. El proceso de revisión es rápido y constructivo, y se realiza vía correo electrónico siempre que es posible. La publicación de los trabajos aceptados se realiza con la mayor rapidez posible, normalmente dentro de los 12 meses siguientes a la recepción del trabajo. Una vez aceptado, el trabajo pasará a ser propiedad de la revista. Ésta se reserva los derechos de autor, y ninguna parte del trabajo podrá ser reproducida sin citar su procedencia. Los derechos de autor quedan reservados a los autores, quienes autorizan a la revista a publicar el artículo. Los artículos se publican con una Licencia Creative Commons Atribución 4.0 Internacional: no se podrá reproducir ni reutilizar ninguna de sus partes sin citar la procedencia.

III

de animales, que los autores disponen de los permisos necesarios y que han cumplido la normativa de protección animal vigente. Los autores pueden enviar también sugerencias para asesores. Las pruebas de imprenta enviadas a los autores deberán remitirse corregidas al Consejo Editor en el plazo máximo de 10 días. Los gastos debidos a modificaciones sustanciales en las pruebas de im prenta, introducidas por los autores, irán a cargo de los mismos. El primer autor recibirá una copia electrónica del trabajo en formato PDF. Manuscritos Los trabajos se presentarán en formato DIN A–4 (30 líneas 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 ninguno, un servicio de corrección por parte de una persona especializada en revistas científicas. En cualquier caso debe presentarse siempre de forma correcta y con un lenguaje claro y conciso. Los caracteres en cursiva se utilizarán para los nombres científicos de géneros y especies y para los neologismos que no tengan traducción; las citas textuales, independientemente de la lengua en que estén, irán en letra redonda y entre comillas; el nombre del autor que sigue a un taxón se escribirá también en redonda. Se evitará el uso de términos extranjeros (p. ej.: latín, aleman,...). Al citar por primera vez una especie en el trabajo, deberá especificarse siempre que sea posible su nombre común. Los topónimos se escribirán bien en su forma original o bien en la lengua en que esté redactado el trabajo, siguiendo el mismo criterio a lo largo de todo el artículo. Los números del uno al nueve se escribirán con letras, a excepción de cuando precedan una unidad de medida. Los números mayores de nueve se escribirán con cifras excepto al empezar una frase. Las fechas se indicarán de la siguiente forma: 28 VI 99 (un único día); 28, 30 VI 99 (días 28 y 30); 28–30 VI 99 (días 28 al 30). Se evitarán siempre las notas a pie de página.

Normas de publicación

Formato de los artículos

Los trabajos se enviarán preferentemente de forma electrónica (abc@bcn.cat). El formato preferido es un documento Rich Text Format (RTF) o DOC, que incluya las figuras y las tablas. Las figuras deberán enviarse también en archivos separados en formato TIFF, EPS o JPEG. Debe incluirse, con el artículo, una carta donde conste que el trabajo versa sobre inves tigaciones originales no publicadas anteriormente y que se somete en exclusiva a Animal Biodiversity and Conservation. En dicha carta también debe constar, para trabajos donde sea necesaria la manipulación

Título. Será conciso pero suficientemente explicativo del contenido del trabajo. Los títulos con designaciones 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 esencia del manuscrito (introducción, material, métodos, resultados y discusión). Se evitarán las especulaciones y las citas bibliográficas. Irá encabezado por el título del trabajo en cursiva.

ISSN: 1578–665X eISSN: 2014–928X

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. Direccion postal del autor o autores, se publicarán tal como se indique en el manuscrito recibido. Identificadores de investigador (ORCID, ResearchID,…), al menos del investigador principal y de quien asuma la correspondencia posterior. (Título, Nombre de los autores, Abstract, Key words, Resumen, Palabras clave, Direcciones postalo e Identificadores de investigador conformarán la primera página.) Introducción. En ella se dará una idea de los antecedentes del tema tratado, así como de los objetivos del trabajo. Material y métodos. Incluirá la información referente a las especies estudiadas, aparatos utilizados, metodología de estudio y análisis de los datos y zona de estudio. Resultados. En esta sección se presentarán únicamente los datos obtenidos que no hayan sido publicados previamente. Discusión. Se discutirán los resultados y se compararán con otros trabajos relacionados. Las sugerencias sobre investigaciones futuras se podrán incluir al final de este apartado. Agradecimientos (optativo). Referencias. Cada trabajo irá acompañado de una bibliografía que incluirá únicamente las publicaciones citadas en el texto. Las referencias deben presentarse según los modelos siguientes (método Harvard): * Artículos de revista: Conroy, M. J., Noon, B. R., 1996. Mapping of species richness for conservation of biological diversity: conceptual and methodological issues. Ecological Applications, 6: 763–773. * Libros y otras publicaciones no periódicas: Seber, G. A. F., 1982. The estimation of animal abundance. C. Griffin & Company, London. * Trabajos de contribución en libros: Macdonald, D. W., Johnson, D. P., 2001. Dispersal in theory and practice: consequences for conservation biology. In: Dispersal: 358–372 (T. J. Clober, E. Danchin, A. A. Dhondt, J. D. Nichols, Eds.). Oxford University Press, Oxford. * Tesis doctorales: Merilä, J., 1996. Genetic and quantitative trait variation in natural bird populations. Tesis doctoral, Uppsala University. * Los trabajos en prensa sólo se citarán si han sido

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

Animal Biodiversity and Conservation 42.2 (2019)

Animal Biodiversity and Conservation Animal Biodiversity and Conservation (formerly Miscel·lània Zoològica) is an interdisciplinary journal published by the Museu de Ciències Naturals de Barcelona since 1958. It includes empirical 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 that expose a new problem or introduces a new topic, presenting clear hypotheses and predictions, and to studies related to Cconservation Biology. Papers purely descriptive or faunal or chorological describing the distribution in space or time of zoological organisms will not be published. These works should be redirected to our sister magazine Arxius de Miscel·lània Zoològica (www.amz.museucienciesjournals.cat). Studies concerning rare or protected species will not be accepted unless the authors have been granted the relevant permits or authorisation. Each annual volume consists of two issues. Animal Biodiversity and Conservation is registered in all principal data bases and is freely available online at www.abc.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 accepted, papers are published as soon as is practicable. This is usually within 12 months of initial submission. Upon acceptance, manuscripts become the property of the journal, which reserves copyright, and no published material may be reproduced or cited without acknowledging the source of information. All rights are reserved by the authors, who authorise the journal to publish the article. Papers are published under a Creative Commons Attribution 4.0 International License: no part of the published paper may be reproduced or reused unless the source is cited.

Information for authors Electronic submission of papers is encouraged (abc@ bcn.cat). The preferred format is DOC or RTF. All figures must be readable by Word, embedded at the end of the manuscript and submitted together in a separate attachment in a TIFF, EPS or JPEG file. Tables should be placed at the end of the document. A cover letter stating that the article reports original research that has not been published elsewhere and has been submitted exclusively for 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 authorisations. Authors may suggest referees for their papers. Proofs sent to the authors for correction should be returned to the Editorial Board within 10 days. Expenses due to any substantial alterations of the proofs will be charged to the authors. ISSN: 1578–665X eISSN: 2014–928X

The first author will receive electronic version of the article in PDF format. Manuscripts Manuscripts must be presented in DIN A–4 format, 30 lines, 70 keystrokes per page. Maintain double spacing throughout. Number all pages. Manuscripts should be complete with figures and tables. Do not send original figures until the paper has been accepted. The text may be written in English, Spanish or Catalan, though English is preferred. The journal provides linguistic revision by an author’s editor. Care must be taken to use correct wording and the text should be written concisely and clearly. Scientific names of genera and species as well as untranslatable neologisms must be in italics. 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. Foreing terms (e.g. Latin, German,...) should not be used. 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 consecutive days, e.g. 28th to 30th). Footnotes should not be used. 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. Author’s address will be published as they appear © 2019 Museu de Ciències Naturals de Barcelona Papers are published under a Creative Commons Attribution 4.0 International License

in the manuscript file. Researcher’s identifiers (ORCID, ResearchID,…), at least from the first and the corresponding authors. (Title, Name, Abstract, Key words, Resumen, Palabras clave and Author’s address and Researcher’s identifiers must constitute the first page) Introduction. Should include the historical background of the subject as well as the aims of the paper. Material and methods. This section should provide relevant information on the species studied, materials, methods for collecting and analysing data, and the study area. Results. Report only previously unpublished results from the present study. Discussion. The results and their comparison with related studies should be discussed. Suggestions for future research may be given at the end of this section. Acknowledgements (optional). References. All manuscripts must include a bibliography of the publications cited in the text. References should be presented as in the following examples (Harvard method): * Journal articles: Conroy, M. J., Noon, B. R., 1996. Mapping of species richness for conservation of biological diversity: conceptual and methodological issues. Ecological Applications, 6: 763–773. * Books or other non–periodical publications: Seber, G. A. F., 1982. The estimation of animal abundance. C. Griffin & Company, London. * Contributions or chapters of books: Macdonald, D. W., Johnson, D. P., 2001. Dispersal in theory and practice: consequences for conservation biology. In: Dispersal: 358–372 (T. J. Clober, E. Danchin, A. A. Dhondt, J. D. Nichols, Eds.). Oxford University Press, Oxford. * PhD thesis: Merilä, J., 1996. Genetic and quantitative trait variation in natural bird populations. PhD thesis, Uppsala University. * Works in press should only be cited if they have been accepted for publication: Ripoll, M. (in press). The relevance of population studies to conservation biology: a review. Animal Biodiversity and Conservation. References must be set out in alphabetical and chronological order for each author, adding the letters a, b, c,...

to papers of the same year. Bibliographic citations in the text must appear in the usual way: "...according to Wemmer (1998)...", "...has been defined by Robinson and Redford (1991)...", "...the prospections that have been carried out (Begon et al., 1999)..." Tables. Must be numbered in Arabic numerals with reference in the text. Large tables should be narrow (across the page) and long (down the page) rather than wide and short, so that they can be fitted into the column width of the journal. Figures. All illustrations (graphs, drawings, photographs) 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 tridimensional. 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 References) should not be numbered. Do not use more than three levels of headings. Manuscripts should not exceed 20 pages including figures and tables. If the article describes new taxa, type material must be deposited in a public institution. Authors are advised to consult recent issues of the journal and follow its conventions. Brief communications Brief communications should follow the same procedure as other articles and they will undergo the same review process. They should not exceed 2,300 words including title, abstract, figure and table legends, acknowledgements and references. The abstract should not exceed 100 words, and the number of references should be limited to 15. Section headings within the text are optional. Brief communications may have up to two figures and/or two tables but the whole paper should not exceed four published pages.

Animal Biodiversity and Conservation 42.2 (2019)

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Animal Biodiversity and Conservation 42.2 (2019)

317–335 Vitorino, L. C., Borges Souza, U. J., Jardim, T. P. F. A., Ballesteros–Mejia, L. Towards inclusion of genetic diversity measures into IUCN assessments: a case study on birds 337–346 Grossman, G. D., DeVries, D. R. Authorship decisions in ecology, evolution, organismal biology and natural resource management: who, why, and how 347–354 Garrote, G., Pérez de Ayala, R. Spatial segregation between Iberian lynx and other carnivores 355–367 Romero, D., Olivero, J., Real, R. Accounting for uncertainty in assessing the impact of climate change on biodiversity hotspots in Spain

369–377 Rea, R. V., Ajala–Batista, L., Aitken, D. A., Child, K. N., Thompson, N., Hodder, D. P. Scat analysis as a preliminary assessment of moose (Alces alces andersoni) calf consumption by bears (Ursus spp.) in north– central British Columbia 379–388 Ruiz Real, F., Martín, J. Evidence of character displacement in microhabitat use between two tropical sympatric Holcosus lizard species (Reptilia, Teiidae) 389–395 Mazzoni, D., Borray, N., Ortega–Segalerva, A., Arroyo, L., González–Solís, J., Senar, J. C. Diet to tissue discrimination factors for the blood and feathers of the monk parakeet (Myiopsitta monachus) and the ring–necked parakeet (Psittacula krameri)

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

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

Índex / Índice / Contents Animal Biodiversity and Conservation 42.2 (2019) ISSN 1578–665 X eISSN 2014–928 X 203–211 Ferreguetti, A. C., Duarte Rocha, C. F., Godoy Bergallo, H. Poaching in non–volant mammals in the Neotropical region: the importance of a metric to assess its impacts 213–226 Ortuño, V. M., Ledesma, E., Jiménez–Valverde, A., Pérez–Suárez, G. Studies of the mesovoid shallow substratum can change the accepted autecology of species: the case of ground beetles (Coleoptera, Carabidae) in the Sierra de Guadarrama National Park (Spain) 227–244 Dumont, L., Lauer, E., Zimmermann, S., Roche, P., Auliac, P., Sarasa, M. Monitoring black grouse Tetrao tetrix in Isère, northern French Alps: cofactors, population trends and potential biases 245–252 Saura–Mas, S., Benejam, L. Effects of the invasive crayfish Procambarus clarkii on growth and development of Pelophylax perezi tadpoles in field conditions 253–256 Sánchez–Tocino, L., de la Linde Rubio, A., López– Rodríguez, M. J., Tierno de Figueroa, J. M. Conservation status of Paramuricea clavata (Risso, 1826) (Anthozoa, Alcyonacea) in the Chafarinas Islands (Mediterranean Sea)

257–266 Martínez–Ortí, A., Vilavella, D., Bargues, M. D., Mas–Coma, S. Risk map of transmission of urogenital schistosomiasis by Bulinus truncatus Audouin, 1827) (Mollusca Gastropoda, Bulinidae) in Spain and Portugal 267–277 Polidori, C., Federici, M. Differential distribution of resources for females on a dioecious plant affects the small–scale distribution of male of an oligolectic bee 279–291 De la Hera, I. Seasonality affects avian species distribution but not diversity and nestedness patterns in the urban parks of Vitoria–Gasteiz (Spain) 293–300 Cordero–Rivera, A., Sanmartín–Villar, I., Sánchez Herrera, M., Rivas–Torres, A., Encalada, A. C. Survival and longevity in neotropical damselflies (Odonata, Polythoridae) 301–316 Oualid, J. A., Iazza, B., Tamsouri, N. M., El Aamri, F., Moukrim, A., López–González, P. J. Hidden diversity under morphology–based identifications of widespread invasive species: the case of the 'well–known' hydromedusa Craspedacusta sowerbii Lankester 1880

FUNDACIÓN ESPAÑOLA

Amb el suport de / Con el apoyo de / With the support of: FUNDACIÓN ESPAÑOLA PARA LA CIENCIA Y LA TECNOLOGÍA

Nº DE CERTIFICADO: FECYT-113/2019 FECHA DE CERTIFICACIÓN: 6 de octubre 2014 (4ª convocatoria) ESTA CERTIFICACIÓN ES VÁLIDA HASTA EL: 12 de julio de 2020