ABC 43-1 (2020) by Museu Ciències Naturals de Barcelona MCNB

en J. Hatchwell, Univ. of Sheffield, UK

Dibuix de la coberta / Dibujo de la portada / Drawing of the cover: Cotylorhiza tuberculata, borm groc o medusa ou ferrat, medusa del mediterraneo o medusa huevo frito, Mediterranean jelly or fried egg jellyfish (by Jordi Domènech) Localització / Localización / Locality: Platja del Far del Fangar, delta de l’Ebre; Playa 'Far del Fangar', delta del Ebro; 'Far del Fangar' Beach, Ebro Delta

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

Animal Biodiversity and Conservation 43.1, 2020 Autoedició: Montserrat Ferrer Fotomecànica i impressió: CEVAGRAF SCCL ISSN: 1578–665 X eISSN: 2014–928 X Dipòsit legal: B. 5357–2013

Animal Biodiversity and Conservation es publica amb el suport de / Animal Biodiversity and Conservation se publica con el apoyo de / Animal Biodiversity and Conservation is published with the support of: Asociación Española de Ecología Terrestre – AEET Sociedad Española de Etología y Ecología Evolutiva – SEEEE Sociedad Española de Biología Evolutiva – SESBE Disponible gratuitament a internet / Disponible gratuitamente en internet / Freely available online at: www.abc.museucienciesjournals.cat

Animal Biodiversity and Conservation 43.1 (2020)

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 Pelayo Acevedo Instituto de Investigación en Recursos Cinegéticos IREC–UCLM–CSIC–JCCM, Ciudad Real, 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 Pedro Aragón Universidad Complutense de Madrid, Madrid, Spain Miquel Arnedo Universitat de Barcelona, Barcelona, Spain Xavier Bellés Institut de Biología Evolutiva UPF–CSIC, Barcelona, Spain David Canal MTA Centre for Ecological Research, Vácrátót, Hungary 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 Manuel B. Morales CIBC–Universidad Autónoma de Madrid, 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

Animal Biodiversity and Conservation 43.1 (2020)

Potential effect of habitat disturbance on reproduction of the critically endangered harlequin frog Atelopus varius in Las Tablas, Costa Rica D. A. Gómez–Hoyos, R. Seisdedos–de–Vergara, J. Schipper, R. Allard, J. F. González–Maya Gómez–Hoyos, D. A., Seisdedos–de–Vergara, R., Schipper, J., Allard, R., González–Maya, J. F., 2020. Potential effect of habitat disturbance on reproduction of the critically endangered harlequin frog Atelopus varius in Las Tablas, Costa Rica. Animal Biodiversity and Conservation, 43.1: 1–7, Doi: https://doi.org/10.32800/ abc.2020.43.0001 Abstract Potential effect of habitat disturbance on reproduction of the critically endangered harlequin frog Atelopus varius in Las Tablas, Costa Rica. We studied a population of Atelopus varius in Las Tablas Protected Zone in southwest Costa Rica, where we estimated occupancy rates of tadpoles along the Cotón River. In addition, we report the first tadpoles observed in the wild in 20 years. Tadpole rate of occupancy was greater in habitat containing native forest than in disturbed areas bordering cattle pasture. This same pattern was also reflected in adult hotspots, where encounter rates were higher for adults in habitat surrounded by forest versus pasture. We present evidence for the potential effect of habitat modification on the presence and reproduction of A. varius and suspect that over time this modification impacts the species' demography. However, further study is necessary before we can confirm that habitat change alone was the key factor involved in patterns of decline for the species. Key words: Habitat disturbance, Hotspots, Occupancy, Tadpoles, Threats Resumen Posibles efectos de la perturbación del hábitat en la reproducción de la rana arlequín Atelopus varius críticamente amenazada en Las Tablas, Costa Rica. Estudiamos una población de Atelopus varius de la Zona Protegida Las Tablas, en el sureste de Costa Rica, donde estimamos la tasa de ocupación de renacuajos a lo largo del río Cotón. Además, reportamos los primeros renacuajos observados en el medio silvestre en 20 años. La tasa de ocupación de renacuajos fue mayor en los hábitats con bosque nativo que en las zonas perturbadas, aledañas a pastizales. Esta misma pauta se observa también en las zonas de alta concentración de adultos, donde la tasa de encuentro es más alta en los hábitats rodeados por bosque que en los pastizales. Presentamos algunos indicios de los efectos que la modificación del hábitat podría tener en la presencia y reproducción de A. varius y planteamos la posibilidad de que, con el tiempo, esta modificación pueda afectar también a su composición demográfica. Sin embargo, se necesitan más estudios antes de poder confirmar que el cambio de hábitat fue el único factor que determinó la tendencia decreciente de la especie. Palabras clave: Perturbación del hábitat, Zonas de alta concentración, Ocupación, Renacuajos, Amenazas Received: 23 IX 17; Conditional acceptance: 19 XII 17; Final acceptance: 03 VII 19 Diego A. Gómez–Hoyos, Rocío Seisdedos–de–Vergara, Jan Schipper, José F. González–Maya, ProCAT Internacional / Sierra to Sea Institute Costa Rica, Las Alturas, Puntarenas, Costa Rica.– Diego A. Gómez–Hoyos, Grupo de Investigación y Asesoría en Estadística / Grupo de Herpetología, Universidad del Quindío, Armenia, Colombia.– Jan Schipper, Ruth Allard, Arizona Center for Nature Conservation / Phoenix Zoo, Phoenix, Arizona, U.S.A. Corresponding author: Diego A. Gómez–Hoyos. E–mail: dgomez@procat–conservation.org

ISSN: 1578–665 X eISSN: 2014–928 X

Introduction The variable harlequin frog Atelopus varius is categorized as Critically Endangered on the IUCN Red List of Threatened Species following a drastic population decline since the 1980s (Pounds et al., 2010). Despite this decline, Atelopus varius is one of the few harlequin frog species remaining in Central America. Reproductive populations have recently been recorded in Costa Rica and Panama (González–Maya et al., 2013; Perez et al., 2014; Barrio–Amorós and Abarca, 2016), providing the opportunity to carry out research and conservation efforts to better understand the factors leading to the decline and how to address them. Although A. varius is the best–known species within the genus (Lötters, 1996), information about current threats and their effect is scarce. Ongoing factors impacting the species include chytrid fungus, invasive species, and habitat loss and degradation (Richards–Zawacki, 2009; Pounds et al., 2010; Perez et al., 2014; González–Maya et al., 2018). Chytrid fungus is the most well–known and persistent threat affecting the species (Brem and Lips, 2008), and the Atelopus genus, while other described threats both directly and indirectly impact the species (Pounds et al., 2010). Habitat disturbance is the main driver of amphibian population decline and extinction worldwide (Blaustein et al., 2011). However, the implications of habitat disturbance on population are unknown for many amphibians (Cushman, 2006), including the genus Atelopus. Understanding the effects of this threat factor is essential to develop and prioritize management actions to avoid the extinction of endangered species. Because the species is restricted to a very limited area along riverbanks, riparian habitat disturbance may have a negative effect on many aspects of the ecology of A. varius (Richards–Zawacki, 2009). Unfortunately, the effect at the population level is unknown. Here, we show evidence of the impact of habitat disturbance on the reproduction of A. varius along the Cotón River, Las Tablas Protected Zone, Costa Rica. We also present the first records of harlequin frog tadpoles in the wild in at least 20 years. Material and methods Our study site is located at Las Tablas Protected Zone (LTPZ), Puntarenas, Costa Rica. LTPZ is in southeastern Costa Rica, specifically along the Pacific flanks of the Talamanca mountain range (8.93º N and 82.82º W). LTPZ spans for over 19,000 ha and is part of La Amistad Biosphere Reserve. Our specific study site is situated in and along the Cotón River (fig. 1), at 1,300 m a.s.l., with a mean annual precipitation of 3,500 mm and a mean annual temperature of 19 ºC (González–Maya and Mata–Lorenzen, 2008). We conducted surveys along a 2.2 km section of the Cotón River, in LTPZ (fig. 1), along both river banks. Two experienced researchers (each with at least 1.5 years surveying the population) performed the surveys. One researcher focused on searching for adults and the other focused the search on tadpoles.

Gómez–Hoyos et al.

The same observers performed all the surveys during the study to minimize detection bias. Mean days surveyed per month were 7.4 (± 1.52) with a mean effort of 4.24 hours/day/person (± 1.61). The surveys were carried out monthly from October 2016 to April 2017, corresponding to the reproductive season of A. varius. The only break in surveys occurred following abnormally high water levels during November 2016 which prevented access to the study area due to hazardous conditions. All surveys consisted of visual encounter surveys (VES) along the river banks; for adults this entailed surveying at least 3 m from the water–line into the forest and for tadpoles we searched over and under rocks and in pools along the water's edge. Tadpoles of A. varius were identified by their exclusive ventral sucking disc and distinguishing chromatophores, which differ from larvae of other sympatric anuran species, as well as the depressed–flattened body and massive proximal caudal musculature (Lötters, 1996). Tadpole records include one or more individuals which were assigned to Gosner's stages (1960) per month through subaquatic photography. Each adult individual detected was captured and measured. We photographed each adult’s ventral and dorsal surfaces to 1) identify individuals and to create a capture–recapture histories database, 2) define movement of individuals and 3) prevent pseudoreplication. Only adult detections between October and December 2016 were used in the analysis, because this is time period when aggregation for reproduction occurs. Uncertainty about the identification between males and females did not allow discrimination by sex categories in all cases; only females with snout to vent length (SVL) > 33 mm were identified. Geographic location was recorded for each individual or tadpole and was georeferenced using QGIS software (QGIS Development Team, 2016). Each location was classified into one of two vegetation categories based on satellite imagery and verified using field observations, according to the composition and dominant vegetation elements: disturbed forest (perturbed) zone and late successional forest zone (fig. 1). The perturbed zone is dominated by pastures and scrublands with a few dispersed natural vegetation patches and crops. The late successional forest zone is dominated by natural vegetation that suffered selective wood extraction over 15 years ago and which is currently recovering. Capture–mark–recapture data were used to estimate the mean distance that A. varius individuals moved. We used the Points2One application (available at: https://plugins.qgis.org/plugins/points2one/) to calculate distance between capture and recapture locations for each individual. Mean distance covered by A. varius individuals was 76 m, so we established sub–transects of 100 m to divide the 2.2 km stretch we surveyed into 22 segments. The 100 m segments were used to identify aggregations of individuals of greater magnitude than that expected for random events. These aggregations were identified with the Getis–Ord Gi* statistic (Ord and Getis, 1995) return the z–score (GiZ) of each segment and 5 % significance, a GiZ greater than 1.96 indicates a hotspot. We used the

Animal Biodiversity and Conservation 43.1 (2020)

–82º50'

–82º49'

8º56'

Panama

8º57'

Costa Rica

8º57'

Nicaragua

0.25

–82º50'

0.25

0.50 km

–82º49'

Fig. 1. Study area in the Cotón River, Las Tablas Protected Zone, Costa Rica: location of the protected zone (white rectangle), perturbed zone (dotted line) and late successional forest zone (continuous line) of the studied stretch of the Cotón River (Satellite image: Google Inc.). Fig. 1. Zona de estudio en el río Cotón, en la Zona Protegida Las Tablas, Costa Rica: ubicación de la zona protegida (rectángulo blanco), de la zona perturbada (línea punteada) y de la zona de bosque en sucesión tardía (línea continua) del tramo estudiado del río Cotón (imagen satelital: Google Inc.).

fixed distance option of 200 m (two segments) to identify hotspots. This method works by evaluating each feature within the context of neighboring features, thus identifying the degree of spatial clustering. We used the Getis–Ord Gi* statistic because we assumed that the spatial distribution of harlequin frog (adult or tadpole) hotspots are spatially associated during the reproductive season since adults aggregate during these months (Lötters, 1996; Savage, 2002). We also estimated the occupancy rate of tadpoles along the river. The occupancy was modelled through Unmarked package (Fiske and Chandler, 2011) for the R language (R Core Team, 2016). We tested the null model (constant detection and occupancy) against detection (p) and occupancy (Psi) depending on vegetation cover zone (perturbed or late successional forest areas). The best–fitting model was selected based on Akaike's Information Criterion (AICc), which was adjusted for small sample size and derived measures (Burnham and Anderson, 2002). Model selection was performed with AICcmodavg package (Mazerolle, 2016). Our research permit was granted by the Sistema Nacional de Áreas de Conservación (SINAC) of Costa Rica (identification number: SINAC–ACLAP–GMRN– INV–105–2016).

Results We identified 87 adult individuals during the reproductive season (October–December 2016), where at least nine were females and three of them were in amplexus. Furthermore, 33 records of tadpoles with 195 individuals in total were taken (fig. 2). We detected between 1 and 55 individuals per record (mean: 5.9 ± 10.1). The tadpoles were between stages 26 in January 2016 and 39 in April 2016, Gosner's stages (1960). Despite equal survey effort, tadpoles were recorded almost exclusively in the late successional forest zone. We identified two hotspots for tadpole records using the defined segments of the Cotón River. The lengths of these hotspot segments were 200 and 400 m with GiZ value > 2.19 (fig. 3A). On the other hand, only 22 % of the adults were recorded in the perturbed zone. One adult hotspot was identified, and it was exclusively within the late successional forest zone. This hotspot was 400 m long with GiZ value > 2.16 (fig. 3B). Adults and tadpoles hotspots hadan overlap of 293 m (50.56 % and 75.85 % of tadpole and adults hotspots, respectively). The occupancy model supports the hypothesis that habitat disturbance has a negative effect on A. varius

Gómez–Hoyos et al.

Fig. 2. Tadpoles of Atelopus varius from Las Tablas Protected Zone, Costa Rica: A, ten individuals recorded; B, detail of a tadpole. Fig. 2. Renacuajos de Atelopus varius de la Zona Protegida Las Tablas, Costa Rica: A, diez individuos registrados; B, detalle de un renacuajo.

tadpole presence. The best–fitting model included both constant detection and occupancy rate depending on vegetation zone (table 1). Detection probability was 0.72 ± 0.07 (CI 95 %: 0.59–0.82). Occupancy in the perturbed zone was 0.095 ± 0.09 (CI 95 %: 0.02–0.36) and 0.98 ± 0.03 (CI 95 %: 0.84–0.99) in the late successional forest zone. Discussion Despite A. varius being the most studied species within the genus Atelopus (Lötters, 1996), ecological aspects are still almost entirely unknown. After the dramatic population decline and local extinctions in Costa Rica and Panama in the 1980s, remnant populations such as those studied here are an important opportunity for research, especially on demography and threat impacts (Muths et al., 2011). Herein, we present the first records of A. varius tadpoles in the wild in at least 20 years (Lötters, 1996; La Marca et al., 2005; Lötters pers. comm.). It is important to consider that almost all tadpoles were restricted to the forest zone of the Cotón River, where sun exposure is less due to continuous forest canopy cover. Hotspots of adult individuals were also restricted to the forest zone, and tadpole occupancy was higher in the forest than in the perturbed zone. This spatial pattern is the first evidence indicating that habitat disturbance affects reproduction and could have a strong impact on demographic parameters in A. varius. However, to avoid premature generalizations we recommend a multi–year study in order to confirm that the spatial pattern is explained by habitat disturbance as a driver of reproduction site selection in A. varius and not by some other factor. To date, habitat disturbance (specifically habitat loss) has not been a factor considered in the decline

and extinction of the Atelopus genus (La Marca et al., 2005). On the other hand, threat factors such as chytrid fungus have been well documented and have been associated with amphibian extinction and population decline processes (Ryan et al., 2008; Crawford et al., 2010), and more specifically for Atelopus species (Brem and Lips, 2008; La Marca et al., 2005). Although there is no doubt that habitat disturbance has been a factor involved in amphibian decline and extinction globally (Blaustein et al., 2011), including other Atelopus species (La Marca et al., 2005), we demonstrate a potential effect of habitat disturbance over A. varius reproduction and we suspect this disturbance also impacts population demography. However, we cannot confirm that habitat disturbance alone was the key factor involved in decline patterns for the species, although we suspect that it is a contributing factor. Species' disturbance tolerance is widely used as an indicator of ecosystem conservation status (Segurado et al., 2011) and it is also used during routine assessment of species' conservation status and categorization on the IUCN Red List of Threatened Species (IUCN, 2016). However, disturbance tolerance of A. varius is based on expert inference, with little empirical evidence (Segurado et al., 2011). Expert opinion is generally based on anecdotic and limited records, although in the absence of information it is an important resource when inferring impacts of known threats. However, it can be easily misapplied from one species to the next in the absence of specific data on occupancy during different life stages, specifically on breeding populations. This new information stands in contrast to inference sometimes made unknowingly by some experts. Therefore, we feel it is prudent and necessary to provide evidence that the presence of Atelopus varius individuals (especially adults) in a perturbed zone is not proof of a population' tolerance to disturbance,

Animal Biodiversity and Conservation 43.1 (2020)

–82º50'

–82º49'

0.1

0.1 0.2 km

0.1

0.1 0.2 km

8º56'

overlapping zone

8º56'

–82º50'

8º56'

–82º49'

Fig. 3. Hotspots for tadpoles (A) and adults (B) of Atelopus varius in the Cotón River, Las Tablas Protected Zone, Costa Rica. Records (white dots); hotspots with GiZ value > 1.96 (black segments); perturbed zone (dotted line). Hotspots overlapping zones are indicated with arrows and rectangles. Fig. 3. Zonas de alta concentración de renacuajos (A) y adultos (B) de Atelopus varius en el río Cotón, en la Zona Protegida Las Tablas, Costa Rica: registros (círculos blancos), zonas de concentración con GiZ > 1,96 (segmentos negros), y zona perturbada (línea punteada). Los puntos en los que se solapan varias zonas de alta concentración se indican con flechas y rectángulos.

Table 1. Model selection to explain detection and occupancy rate of Atelopus varius tadpoles in the Cotón River, Las Tablas Protected Zone, Costa Rica: K, number of parameter; AICc, Akaike's information criterion adjusted for small sample size. Tabla 1. Selección de modelos para explicar la detección y la tasa de ocupación de renacuajos de Atelopus varius en el río Cotón, en la Zona Protegida Las Tablas, Costa Rica: K, número de parámetros; AICc, criterio de información de Akaike ajustado para tamaños de muestra pequeños. Model K AICc ΔAICc

AICc Weight

Cumulative Weight

0.81

p(.) Psi(zone)

71.4

p(zone) Psi(zone)

74.41

0.18

0.99

p(.) Psi(.)

81.88

10.48

p(zone) Psi(.)

84.58

13.18

nor does it indicate that a population is breeding or thriving in disturbed habitat. Evidence from this study suggests that successful reproduction occurs almost exclusively in the forest zone (i.e. native habitat). We urge biology researchers and conservation scientists to avoid speculation about the effects of habitat disturbance on species, especially those on the brink of extinction. Instead we need to increase ecological research efforts for remnant populations of threatened species, including A. varius, with the goal of conducting appropriate conservation and management actions to avoid their extinction. Acknowledgements This project was partially funded by the Arizona Center for Nature Conservation/Phoenix Zoo, Disney Conservation Fund, Mikelberg Family Foundation, Finca Las Alturas del Bosque Verde and Rufford Foundation. We would also like to acknowledge the determination of Addison Fischer and Fernando Castañeda, whose conservation vision has preserved the upper Cotón River system and without whom this critically important remnant population of Atelopus varius would likely not have survived for us to study. References Barrio–Amorós, C. L., Abarca, J., 2016. Another surviving population of the Critically Endangered Atelopus varius (Anura: Bufonidae) in Costa Rica. Mesoamerican Herpetology, 3: 128–134. Blaustein, A. R., Han, B. A., Relyea, R. A., Johnson, P. T., Buck, J. C., Gervasi, S. S., Kats, L. B., 2011. The complexity of amphibian population declines: understanding the role of cofactors in driving amphibian losses. Annals of the New York Academy of Sciences, 1223: 108–119. Brem, F. M. R., Lips, K. R., 2008. Batrachochytrium dendrobatidis infection patterns among Panamanian amphibian species, habitats and elevations during epizootic and enzootic stages. Diseases of Aquatic Organisms, 81: 189–202. Burnham, K., Anderson, D., 2002. Model selection and multimodel inference: a practical information and theoretic approach. Springer, New York. Crawford, A. J., Lips, K. R., Bermingham, E., 2010. Epidemic disease decimates amphibian abundance, species diversity, and evolutionary history in the highlands of central Panama. Proceedings of the National Academy of Sciences, 107(31): 13777–13782. Cushman, S. A., 2006. Effects of habitat loss and fragmentation on amphibians: a review and prospectus. Biological Conservation, 128(2): 231–240. Fiske, I., Chandler, R. B., 2011. Unmarked: An R package for fitting hierarchical models of wildlife occurrence and abundance. Journal of Statistical Software, 43: 1–23. González–Maya, J. F., Belant, J. L., Wyatt, S. A., Schipper, J., Cardenal–Porras, J., Corrales, D.,

Gómez–Hoyos et al.

Cruz–Lizano, I., Hoepker, A., Escobedo–Galván, A. H., Castañeda, F., Fischer A., 2013. Renewing hope: the rediscovery of Atelopus varius in Costa Rica. Amphibia–Reptilia, 34(4): 573–578. González–Maya, J. F., Gómez–Hoyos, D. A., Cruz– Lizano, I., Schipper, J., 2018. From hope to alert: demography of a remnant population of the Critically Endangered Atelopus varius from Costa Rica. Studies on Neotropical Fauna and Environment, 53(3): 194–200. González–Maya, J. F., Mata–Lorenzen, J., 2008. Dung–beetles (Coleoptera: Scarabeidae) from the Zona Protectora Las Tablas, Costa Rica. Checklist, 4: 458–463. Gosner, K. L., 1960. A simplified table for staging anuran embryos and larvae with notes on identification. Herpetologica, 16(3): 183–190. IUCN, 2016. The IUCN Red List of Threatened Species. Version 2016–3, http://www.iucnredlist.org. [Accessed on 07 December 2016]. La Marca, E., Lips, K. R., Lötters, S., Puschendorf, R., Ibáñez, R., Rueda–Almonacid, J. V., Schulte, R., Marty, C., Castro, F., Manzanilla–Puppo, J., García–Pérez, J. E., Bolaños, F., Chaves, G., Pounds, J. A., Toral, E., Young, B. E., 2005. Catastrophic population declines and extinctions in neotropical harlequin frogs (Bufonidae: Atelopus). Biotropica, 37(2): 190–201. Lötters, S., 1996. The Neotropical Toad Genus Atelopus. Checklist – Biology – Distribution (M. Vences, F. Glaw, Eds.). Verlags GbR, Koln. Mazerolle, M. J., 2016. AICcmodavg: Model selection and multimodel inference based on (Q)AIC(c). R package version 2.1–0. https://cran.r–project.org/ package=AICcmodavg [Accessed on April 2018]. Muths, E., Scherer, R. D., Pilliod, D. S., 2011. Compensatory effects of recruitment and survival when amphibian populations are perturbed by disease. Journal of Applied Ecology, 48(4): 873–879. Ord, J. K., Getis, A., 1995. Local spatial autocorrelation statistics: distribution issues and an application. Geographical Analysis, 27: 286–306. Perez, R., Richards–Zawacki, C. L., Krohn, A. R., Robak, M., Griffith, E. J., Ross, H., Gratwicke, B., Ibáñez, R., Voyles, J., 2014. Field surveys in Western Panama indicate populations of Atelopus varius frogs are persisting in regions where Batrachochytrium dendrobatidis is now enzootic. Amphibian & Reptile Conservation, 8(2): 30–35. Pounds, J., Puschendorf, R., Bolaños, F., Chaves, G., Crump, M., Solís, F., Ibáñez, R., Savage, J., Jaramillo, C., Fuenmayor, Q., Lips, K., 2010. Atelopus varius. In: IUCN Red List of Threatened Species. Version 2017.1, 2010, http://www.iucnredlist.org [Accessed on April 2018] QGIS Development Team, 2016. QGIS Geographic Information System. Open Source Geospatial Foundation, http://qgis.osgeo.org [Accessed on April 2018] R Core Team, 2016. R: A language and environment for statistical computing. R Foundation for Statistical Computing, Vienna, Austria, https://www.R– project.org/ [Accessed on April 2018]

Animal Biodiversity and Conservation 43.1 (2020)

Richards–Zawacki, C. L., 2009. Effects of slope and riparian habitat connectivity on gene flow in an endangered Panamanian frog, Atelopus varius. Diversity and Distributions, 15: 796–806. Ryan, M. J., Lips, K. R., Eichholz, M. W., 2008. Decline and extirpation of an endangered Panamanian stream frog population (Craugastor punctariolus) due to an outbreak of chytridiomycosis. Biological Conservation, 141(6): 1636–1647.

Savage, J. M., 2002. The amphibians and reptiles of Costa Rica: a herpetofauna between two continents, between two seas. University of Chicago press, Chicago. Segurado, P., Santos, J. M., Pont, D., Melcher, A. H., Jalon, D. G., Hughes, R. M., Ferreira, M. T., 2011. Estimating species tolerance to human perturbation: expert judgment versus empirical approaches. Ecological Indicators, 11(6): 1623–1635.

Gómez–Hoyos et al.

Animal Biodiversity and Conservation 43.1 (2020)

Tree reproductive phenology determines the abundance of medium–sized and large mammalian assemblages in the Guyana shield of the Brazilian Amazonia A. R. Mendes Pontes, V. M. Guedes Layme, L. R. Rodrigues de Lucena, D. J. Chivers Mendes Pontes, A. R., Guedes Layme, V. M., Rodrigues de Lucena, L. R., Chivers, D. J., 2020. Tree reproductive phenology determines the abundance of medium–sized and large mammalian assemblages in the Guyana shield of the Brazilian Amazonia. Animal Biodiversity and Conservation, 43.1: 9-26, Doi: https://doi. org/10.32800/abc.2020.43.0009 Abstract Tree reproductive phenology determines the abundance of medium–sized and large mammalian assemblages in the Guyana shield of the Brazilian Amazonia. Assemblages of medium and large–sized mammals were studied in the Guyana shield of the Brazilian Amazonia. Diurnal and nocturnal line–transect samplings were carried out via the line–transect method in five different forest types along a 10–km transect, along which we also recorded habitat variables, such as tree species diversity, reproductive phenology, and residual fruit productivity. Group density was separately calculated for all mammalian species in the five forest types. Stepwise multiple regression analyses were performed to determine which habitat variables best predicted the mammalian species densities in the sampled forests. The sole determinants of mammalian densities in the forest types studied were basal area of each forest type, total number of tree species in each forest type, and tree reproductive phenology. Key words: Mammalian assemblages, Group density, Environmental determinants, Forest productivity, Tree reproductive phenology, Amazonia Resumen La fenología reproductiva de los árboles determina la abundancia de las comunidades de mamíferos de talla media y grande en el escudo de Guyana de la Amazonia brasileña. Se estudiaron varias comunidades de mamíferos de talla mediana y grande en el escudo de Guyana de la Amazonia brasileña. Se realizaron muestreos en transectos lineares diurnos y nocturnos de 10 km de longitud en cinco tipos distintos de bosque, a lo largo de los que también se registraron variables del hábitat, como la diversidad de especies arbóreas, la fenología reproductiva y la productividad residual de frutos. La densidad del grupo de calculó por separado para todas las especies de mamíferos en los cinco tipos de bosque. Se llevaron a cabo análisis escalonados de regresión múltiple para determinar qué variables del hábitat permitían predecir mejor la densidad de las especies de mamíferos en los bosques muestreados. Los únicos factores determinantes de la densidad de mamíferos en los tipos de bosques estudiados fueron el área basimétrica de cada tipo de bosque, el número total de especies arbóreas en cada tipo de bosque y la fenología reproductiva de los árboles. Palabras clave: Comunidades de mamíferos, Densidad de grupo, Factores determinantes ambientales, Productividad forestal, Fenología reproductiva de los árboles, Amazonia Received: 18 VI 18; Conditional acceptance: 22 X 18; Final acceptance: 16 VII 19 Antonio Rossano Mendes Pontes, David J. Chivers, Wildlife Research Group, Department of Anatomy, University of Cambridge, Downing Street, Cambridge CB2 3DY, England.– Viviane Maria Guedes Layme, Universidade Federal do Mato Grosso–UFMT, Instituto de Biociências, Avenida Fernando Correia s/n., Laboratório de Ecologia de Mamíferos, Cuiabá, Mato Grosso, CEP 78.060–900, Brazil.– Leandro Ricardo Rodrigues de Lucena, Universidade Federal Rural de Pernambuco, Campus de Serra Talhada, Departamento de Zootecnia, Av. Gregório Ferraz Nogueira, s/n., Serra Talhada, Pernambuco, CEP 66.909–535, Brazil.– David J. Chivers, Selwyn College, Grange Road, Cambridge CB3 9DQ, England. Corresponding author: A. R. Mendes Pontes. E–mail: antonio.rossano@inpa.gov.br ISSN: 1578–665 X eISSN: 2014–928 X

Introduction For many decades the relationship between body mass and abundance of animals has been a major subject of study (Elton, 1932, 1933). There is a general trend for abundance to decrease with increasing body mass, among different taxa, such as insects and birds (Juanes, 1986; Gaston and Lawton, 1988; Blackburn et al., 1990; Nee et al., 1991; Santini et al., 2017). Studies on mammals from different regions of the Neotropics has also showed that density of tropical mammals decreases with increasing biomass, biomass increases with increasing body weight, and biomass is positively related to home range size (Milton and May, 1976; Clutton–Brock and Harvey, 1977; Eisenberg, 1980; Damuth, 1981; Peters and Wassenberg, 1983; Peters and Raelson, 1984; Robinson and Redford, 1989; Eisenberg, 1990; Santini el al., 2017). Kinnaird and Eisenberg (1989) highlight, however, the importance of considering the dietary specialisation of the animals, which may cause variation in the relationship between density and biomass. They also show considerable scatter around the regression line, and explain that this is caused by diet and to a lesser extent by phylogeny. These abundance/body mass relationships depend on the availability and divisibility of food in the habitat (Hutchinson and MacArthur, 1959; Brown, 1981; Juanes, 1986; Maurer and Brown, 1988), which is also related to what Brown and Maurer (1986) envisioned, that larger species may monopolise the resources. This also agrees with Lawton's (1989) hypothesis that, through pre–emptive competition (Schoener, 1983), larger species have access to a greater share of the resources, to the detriment of the smaller. Furthermore, generalists should also occur at higher densities than specialists, for they have access to more food items (Brown, 1984; Gaston and Lawton, 1988). In northernmost Brazilian Amazonia, on the Guyana Shield, lies one of the poorest regions in the Amazon basin for mammalian species diversity. Within it, the poorest area overall is the central zone, where rainfall is low and the dry season protracted (Eisenberg and Redford, 1979; Emmons, 1984; Janson and Emmons, 1990; Mendes Pontes, 2004; Hoorn et al., 2010; Luna et al., 2017). Here the vegetation is a mosaic of campinas (open, scrub vegetation), campinaranas (forests with short thin trees that stand in irregular, floodable terrain), and seasonally–dry to dry forests (Mendes Pontes et al., 2012). In such habitat, terrestrial mammalian abundance may be higher than for arboreal species, and may fluctuate seasonally and yearly due to resource scarcity (Janzen, 1974; Eisenberg et al., 1979; August, 1983; Schaller, 1983; Mendes Pontes, 2004). Studies of medium– and large–sized mammals in northernmost Brazilian Amazonia are relatively recent and the focus has been mainly on the mammals of protected areas, such as Maracá Ecological Station (Nunes, 1992; Mendes Pontes, 1997, 1999, 2004; Fragoso, 1998; Mendes Pontes et al., 2007), and the Waimiri–Atroari (Mazurek et al., 2000), Yanomami and Macuxi Indian reserves (Fragoso, 2004), although most recently, Melo et al. (2015) and Luna et al. (2017)

Mendes Pontes et al.

studied the impacts on mammal assemblages of recent fishbone human settlements in Roraima state. It was revealed in these studies that the abundance of terrestrial mammal species can be greater than those of the arboreal ones, and that larger species can be more abundant than the smaller ones. A wide variety of variables may affect the abundance of mammalian assemblages not subjected to human interference, including climate, tectonics, fluvial history, topography (Tuomisto et al., 1995, 2003; Kristiansen et al., 2012; Pomara et al., 2012, 2014; Rossetti, 2014; Higgins et al., 2011, 2015), rainfall (Janson et al., 1981; Terborgh, 1983), site temperature (Peters and Raelson, 1984), plant anti– predator defences (Janzen, 1974; Montgomery and Sunquist, 1978), home range size (Milton and May, 1976), predators (Andrewartha, 1961), competition (Emmons, 1984), among many others. It has been shown in most studies, however, that mammalian abundance is, above all, a function of soil type, or, of geologically induced edaphic heterogeneity, drier climates, and longer dry seasons (Tuomisto et al., 1995, 2003, 2014; Hoorn et al., 2010; Kristiansen et al., 2012; Pomara et al., 2012, 2014; Rossetti, 2014; Higgins et al., 2011, 2015; Zuquim et al., 2014), and consequently, of forest structure and productivity (Eisenberg et al., 1979; Eisenberg, 1980; Emmons, 1984; Gentry and Emmons, 1987; Eisenberg, 1990; Kay et al., 1997; Hoorn et al., 2010; Marshall et al., 2014). This ultimately determines vegetation cover and complexity, both vertically and horizontally, as well as resource availability (Eisenberg and Thorington, 1973; August, 1983; Eisenberg, 1990; Glanz, 1990; Janson and Emmons, 1990; Malcolm, 1990; Chapman et al., 2002; Marshall and Leighton, 2006). In this context, the mammalian assemblages should be regulated by the unpredictable, yet recurrent, fluctuations of seasonality, which plays a crucial determining role, since it may cause the species to change diet (MacArthur, 1969; Janson et al., 1981; Terborgh, 1983), habitat use, home range (Peres, 1994; Zhang and Wang, 1995), migrate (Caldecott, 1988; Fragoso, 1998), scatter–hoard (Smythe, 1978), and intensity with which individuals compete for available resources, or ultimately die of starvation (Foster, 1982a, 1982b). In contrast, increased availability of food resources may result in increased species abundance (Eisenberg and Thorington, 1973; August, 1983; Malcolm, 1990; Chapman et al., 2002; Marshall and Leighton, 2006). Covariates of forest structure and productivity, such as stratification, floristic diversity, and phenology have been proposed as determinants of mammalian abundance (Malcolm, 1995; Gentile and Fernandez, 1999; Grelle, 2003; Marshall et al., 2014; Luna et al., 2017). Gadelha et al. (2017), studying a primate community in a habitat similar to that in the current study, found a significant association between mammalian abundance, and absolute dominance of trees, number of clearings, and vegetation height. No studies to date, however, had tested the effects of the forest structure and productivity covariates on whole assemblages of medium– and large–sized mammals in totally protected areas in the Brazilian Amazonia.

Animal Biodiversity and Conservation 43.1 (2020)

This is the first study for determining the drivers of the abundance of medium– and large–sized mammalian assemblages in highly–seasonal Amazonian forests. Our aim was to calculate the mammalian density in five forest types, and correlate them with the forest structure and the productivity covariates, such as basal area, of tree species diversity, phenology, and residual fruit productivity. Thus, we hypothesised that: (1) mammalian species diversity and densities are one of the poorest in the Amazonia due to the predicted hasher conditions of forest structure and productivity; (2) large–bodied mammals are more abundant than medium–sized ones, and terrestrial species, more abundant than arboreal ones due to the predicted more developed understorey and forest discontinuity of these highly–heterogeneous forest mosaics; and (3) mammalian densities are a function of tree reproductive phenology due to the scarcity and dramatic seasonal fluctuations that occur within, and between, forest types. Material and methods Study area Maracá Ecological Station is located at latitude 3º 15'–3º 35' N and longitude 61º 22'–61º 58' W, comprising 1,013 km2 of seasonally–dry forests (Ministério do Interior, 1977). Maracá is a fluvial island formed by the bifurcation of the Uraricoera River into the Maracá Channel (to the south), and Santa Rosa Channel (to the north) (fig. 1), although these channels do not comprise a barrier to most vertebrates, and many species can be seen regularly crossing the river (Mendes Pontes, unpubl. data). Though at least 95 % of the island is forested (Milliken and Ratter, 1990) (fig. 2, 3), it lies at the junction between Amazonian forests (Hylea) and the grassland–dominated Rupununi savannas, which lie to the south and east. Mean annual temperature during the study period, 1997/1998, was 31.6 ºC and rainfall was 1,577.3 mm, with a sharp decline in the dry season, from September to March (fig. 4). The trail for this study was cut in the same area used in the Maracá Rainforest Project (National Environment Office–SEMA, and The Royal Geographical Society–UK) (Thompson et al., 1992), where many of the trees studied were already identified and tagged and forest types identified (Milliken and Ratter, 1990). A 10–km trail, 1 m width, was cut from the easternmost point of the forested area towards the northwest in a straight line. The trail was alpha–numerically marked with tags and paint–marks, and debris were regularly removed to minimize the disturbance when walking. Forest types studied Five contiguous and sharply–differentiated forest types were identified along the 10–km study transect, according to Milliken and Ratter (1990): the two major types that dominate eastern Maracá, Terra Firme and Mixed forest; the two minor ones that occur in small patches inserted in, and completely surrounded by,

the major ones and have similar height, Pau Roxo and Buritizal; and the very short, shrubby, multi–trunked Carrasco forest. The main features of the five forest types described by Milliken and Ratter (1990) and complemented by this study (Supp Info 1) are: Terra Firme forest (fig. 2, 3) Normally about 25 to 35 m in height, emergent reaching 40 m. During the dry season some trees lose their leaves, including some emergent. Soils are Arenic Distrophic Plinthic Yellow Podzol (Carvalho et al., 1988; Thompson et al., 1992). During this study we increased the former species list of this forest type from 160 (Milliken and Ratter, 1990; Nunes, 1992) to 317 species of emergent, canopy and larger understory trees. The number of trees with animal dispersed fruit syndrome was 194 (61.2 %). The total basal area recorded by Milliken and Ratter (1990) was 182.7 m2/ha. The total area surveyed in this study was 45.5 ha. Mixed forest (fig. 2, 3) This has floristic composition and height similar to Terra Firme, but is distinguished by the presence of Peltogyne gracilipes Ducke (Leg. Caesalp.), a species absent from the Terra Firme forest, but which is particularly important due to its high density in some locations (Nascimento and Proctor, 1996; Nascimento et al., 1997), and by the large number of specimens of the palm Oenocarpus bacaba Mart and the giant herb Phenakospermum guyannense (A. Rich.) Endl. ex Miq. (Strelitziaceae) in the better developed ground layer. The soils were classified as Arenic Clayey Dystrophic Yellow and Red–Yellow Argisol (EMBRAPA, 2006). During this study we increased the former species list of this forest type from 201 (Milliken and Ratter, 1990) to 329 species (Mendes Pontes et al., 2013). The number of trees with animal dispersed fruit syndrome was 187 (56.8 %). The total basal area recorded by Milliken and Ratter (1990) was 109.7 m2/ ha. The total area surveyed in this study was 47.6 ha. Pau–Roxo forest (fig. 2, 3) Normally about 30 m in height. It is characterised by a predominance of Peltogyne gracilipes (Leg. Caesalp.), with very few other plant species, even being called Peltogyne–rich forest by Nascimento and Proctor (1996) and Nascimento et al. (1997). It is considered to be an almost–monodominant forest type, and is one of the tallest forest types in the study area, reaching 40 m (Milliken and Ratter, 1990). The soils were considered very sandy and acid with low concentrations of extractable nitrogen and phosphorous and exchangeable cations, especially for magnesium. It had less sand and more silt than the other forest types and varied from loamy sandy to silt clay (Nascimento et al., 1997). It was considered one of the poorest soils recorded in Amazonian lowland evergreen forests (Carvalho et al., 1988). It is possible to identify this forest from the air during the dry season when leaves are lost. Frequent canopy discontinuity means that this forest is usually fairly open. During this study we increased the former

Mendes Pontes et al.

62º 00'

61º 00'

Venezuela 4º 00'

Maracá Ecological Station

Uraricoera River

Guyana

Boa Vista

2º 00'

Roraima Roraima

Branco River

Brazil

Pará

Amazonas 100 km

Fig. 1. Maracá Ecological Station, located in the State of Roraima, on the Guyana Shield of Brazilian Amazonia. Fig. 1. Estación Ecológica de Maracá, situada en el Estado de Roraima, en el escudo de Guyana de la Amazonia brasileña.

species list of this forest type from 26 (Milliken and Ratter, 1990) to 47 species (Mendes Pontes et al., 2013). The number of trees with animal–dispersed fruit syndrome was 34 (72.3%). The total basal area recorded by Milliken and Ratter (1990) was 52.2 m2/ha. The total area surveyed in this study was 1.8 ha.

(Milliken and Ratter, 1990) to 48 species (Mendes Pontes et al., 2013). The number of trees with animal dispersed fruit syndrome was 29 (60.4 %). The total basal area recorded by Milliken and Ratter (1990) was 9 m2/ha. The total area surveyed in this study was 3.4 ha.

Buritizal forest (fig. 2, 3) Occurs exclusively along those streams which flood during the wet season. The commonest plant species is Mauritia flexuosa L. (Arecaceae). Trees may reach 23 m in height, forming a discontinuous canopy. Soils are Clayey Dystrophic Red–Yellow Latosol and Clayey Dystrophic Red–Yellow Argisol (EMBRAPA, 2006). During this study we increased the former species list of this forest type from 19

Carrasco forest (fig. 2, 3) Comprises very small trees that do not exceed 1.5 m in height, all with a very shrubby, multi–trunked form, which makes it almost impossible to penetrate this vegetation formation. Characteristic plant species are the spiny climbing palm Desmoncus polyacanthos Mart. and the also very spinore Bactris maraja Mart. (Arecaceae). No soil information is available for this forest type. During this study we increased the former

Animal Biodiversity and Conservation 43.1 (2020)

620000

390000

630000

640000

650000

Semi–deciduous closed canopy forests Intermediate forest types Evergreen closed canopy forests Open canopy forest types Buritizal forest

660000

670000

680000

Maracá Ecological Station Roraima, Brazilian Amazonia 390000

Vazante vegetation types Unflooded savanna Shallow water with emergent vegetation Trees overhanging water Deep water Cloud cover (high/low cloud) Cloud shadows

380000

Trail Station

370000

Satellite image was processed using ERDAS and GEMSTONE software at the Department of Geography, Univ. of Edinburgh, Scotland

360000

620000

630000

640000

650000

660000

670000

370000

360000

680000

Fig. 2. The forest types at Maracá Ecological Station (according to Milliken and Ratter, 1990) at the landscape scale (10 to 200 km; Peterson et al., 2011), showing the location of the study transect. Fig. 2. Los tipos de bosque en la Estación Ecológica de Maracá (según Milliken y Ratter, 1990) a escala de paisaje (de 10 a 200 km; Peterson et al., 2011), donde se indica la situación del transecto del estudio.

species list of this forest type from 34 (Milliken and Ratter, 1990) to 45 species (Mendes Pontes et al., 2013). The number of trees with animal dispersed fruit syndrome was 32 (71.1 %). The total basal area was 0.23 m2/ha. The total area surveyed in this study was 1.7 ha. Vegetation survey and forest productivity Tree species of each forest type were sampled along sections of the study transect by the point–centred quarter (PCQ) transect method (Muller–Dombois and Ellemberg, 1974), where, at each pre–established interval of 40 m, the four nearest trees with a DBH of at least 10 cm were marked. Using this method, we marked 504 trees, of which 493 were analyzed across the entire study period. Trees were marked with aluminium tags and vinyl flagging, and monitored monthly for leaves, flowers, and fruits (Mendes Pontes et al., 2013). Fruit quantities were assessed visually in the forest canopy by counting the number of fruits on an entire branch, or section of the crown, with the help of powerful binoculars, and extrapolating to the number of units of the same size in the whole tree crown. When fruits occurred in bunches, as in palms, we averaged the number of fruits obtained from fallen bunches, which could be counted more accurately. Flowers and fruits were deposited in the Herbarium and carpological collection of the Roraima State Museum, Boa Vista, Brazil (Mendes Pontes et al., 2013).

We also monitored residual fruit production via a raked–ground fruit survey (Sabatier, 1985; Guillotin et al., 1994; Zhang and Wang, 1995), which consisted of surveying all fallen fruits in the same sampling sections of the study transect at regular intervals and removing them all from the sampled areas after each line–transect sampling. Fruits were collected weekly for an entire year, identified, weighed, and samples of each new collection were preserved, and were later deposited in the first fruit collection for Roraima State, built by the authors and housed in the Roraima State Museum (Mendes Pontes et al., 2013). For each species fruit productivity was calculated by multiplying fruit number by the mean individual fruit weight of each species and total fruit production was obtained by summing up the production of all plant species sampled (Mendes Pontes et al., 2013). Tree–species identifications and fruit–dispersion syndromes followed Van Roosmalen (1985) and Lewis and Owen (1989), and, in all cases, were checked by the following experts: M. Van Roosmalen, L. Lohman, G. Lewis, and T. Pennington. During the study period, we also collected fruits ad libitum whenever they were found within the study area, so that we could sample species that were not registered by any of the previously given methods, although this information was used only to build the checklist of species of each forest type (Mendes Pontes et al., 2013).

Mendes Pontes et al.

10,000

8,000

7,000

(80 m) PRF.2

5,880 5,800

(100 m) PRF.1

4,830 4,730

6,000

5,000

4,000

3,000

2,000

TFF – Terra Firme forest MF – Mixed forest BF – Buritizal forest PRF – Pau Roxo forest CF – Carrasco forest

1,000

MF = 5,350 m

9,000

BF.7 (59 m) BF.6 (47 m)

6,937 6,883

BF.5 (54 m)

6,385 6,369

BF.3 (16 m)

4,897 4,830

BF.3 (67 m) Boundary of the two main forest types

TFF = 4,650 m

(171 m) CF

9,271 9,100

6,963 9,574 9,319 9,272

2,338 2,270

BF.2 (68 m)

1,155 1,125

BF.2 (30 m)

Fig. 3. Study transect with forest types (forest profiles are from Milliken and Ratter, 1990) at a micro–scale (site, or local scale, 10 m to 10 km; Peterson et al., 2011). Fig. 3. Transecto del estudio con los tipos de bosque (perfiles forestales de Milliken y Ratter, 1990) a microescala (lugar o escala local, de 10 m a 10 km; Peterson et al., 2011).

Vertebrates line transect sampling The trail was numerically marked at 50 m intervals, and had debris removed monthly to minimize disturbance. Diurnal line transect samplings were carried out for 12 sequential months, from the very beginning of the wet season (April 1997) to the end of the dry season (March

1998). Line–transect walks were carried out four times a week, from 06:00 to 12:00 h when sampling in Terra Firme forest (4,650 m), or from 07:00 to 13:00 h (much further from the station, in Mixed forest, 5,350 m). Average speed was 1 km per hour, with occasional stops of a few seconds to scan the habitat. No sightings were registered beyond 50 m in all forest types. For very large

45 40

Temperature (ºC)

35 30 25 20 15 10 5

9 10 11 12 1 Month

420 390 360 330 300 270 240 210 180 150 120 90 60 30 0

Rainfall (mm)

Animal Biodiversity and Conservation 43.1 (2020)

Fig. 4. Minimum, mean, and maximum temperature (lines) and rainfall (bars) at Maracá Ecological Station, Roraima, Brazil, during the study period (1997/1998): 1, January; 2, February; 3, March; 4, April; 5, May; 6, June; 7, July; 8, August; 9, September; 10, October; 11, November; 12, December. Fig. 4. Valores mínimo, medio y máximo de temperatura (líneas) y precipitación (barras) en la Estación Ecológica de Maracá, en Roraima (Brasil), durante el período del estudio (1997/1998). (Para las abreviaturas de los meses, véase arriba).

groups, counts were considered as minimum counts, and for animals sub–divided into smaller sub–groups, counts were considered as sub–groups (Mendes Pontes, 1999). Special care was taken to avoid double–counting the sighted animals, such as recording the direction the animal / group took after sighted (and even taking their coordinates), and following them for up to 15' to make sure they did not return to the trail. Nocturnal line–transect samplings were also carried out between April 1997 and March 1998, from 19:00 to 01:00 h, on the same basis as diurnal samplings, with the help of powerful hand spotlights. The data collected when an animal or group was encountered were: species, angle, perpendicular distance, location upon the trail and assumed direction. Date, time started and finished, observers’ names and total distance walked were also recorded every day (see Burnham et al., 1980; National Research Council, 1981; Brockelman and Ali, 1987; Buckland et al., 1993; Mendes Pontes, 1997, 1999). Total diurnal sampling transect length was the summation of all single diurnal walks; for nocturnal samplings it was the total length of the nocturnal walks. For those species active during the day and night, the total trail length was the summation of both diurnal and nocturnal samplings. To calculate group densities, we used the 5.0. version of the DISTANCE program (Buckland et al., 1993), which establishes the maximum effective strip width (ESW). The first step in calculating densities is, therefore, to check the

accuracy of the detection function g(x), which should be a monotonically–decreasing function, although in some cases there are not enough sightings to provide a smooth curve, and in some others, animals might be detected after they have moved a few metres away from the trail. The program DISTANCE considers these constraints, and chooses the most appropriate model to calculate densities (Buckland et al., 1993). Sightings were analyzed for the species in each forest type, during each season. For species with fewer than 25–30 sightings (the minimum number required to run the program DISTANCE) (Buckland et al., 1993), we combined the sightings of each species in all forest types, after having tested whether there were any statistically significant differences in the distribution of perpendicular distances across forest types, thus achieving enough sightings to run DISTANCE. ESW was calculated from DISTANCE, and densities were then calculated following Chiarello (1999). For species with very few sightings we used King's method (Robinette et al., 1974), which has provided satisfactory results (Nunes et al., 1988; Malcolm, 1990), in some cases, identical to those obtained via DISTANCE (Mendes Pontes, 1999). Ecological densities were calculated instead of crude densities, which are densities calculated separately for the five forest types. We calculated group densities for all the studied mammals, including the solitary ones, because

Mendes Pontes et al.

Table 1. Phenology of marked trees in the five studied forest types at Maracá Ecological Station, northernmost Brazilian Amazonia: T, total number of marked trees; Tlb, percentage (mean ± SD) of trees with leaf buds; Tfl, percentage (mean ± SD) of trees with flowers; Tfr, percentage (mean ± SD) of trees with fruits; kg, kg of fruit in the tree canopy. Tabla 1. Fenología de los árboles marcados en los cinco tipos de bosque estudiados en la Estación Ecológica de Maracá, en el extremo septentrional de la Amazonia brasileña: T, número total de árboles marcados; Tlb, porcentaje (media ± DE) de árboles con brotes de hojas; Tfl, porcentaje (media ± DE) de árboles con flores; Tfr, porcentaje (media ± DE) de árboles con frutos; kg, kilogramos de fruta en la copa del árbol. Forest Terra Firme forest

Season

Tlb (%)

203

Tfl (%)

Tfr (%)

Wet

9.9 (20.2 ± 17.3)

4.9 (10 ± 4.6)

13.4 (27.2 ± 7)

776.7

Dry

8.3 (16.8 ± 8.4)

6.0 (12.3 ± 5.3)

9.8 (20 ± 3.5)

532.8

Mixed forest

Wet

8.25 (16.5 ± 8.9)

3.1 (6.3 ± 4)

6.9(13.8 ± 4.7)

509.5

Dry

8.3 (16.5 ± 10.6)

2.6 (5.3 ± 2.3)

3.6 (7.3 ± 2.9)

634.5

Buritizal forest

13.2 (5.3 ± 1.4) 521.1

200

Wet

14.2 (5.7 ± 3)

5.7 (2.3 ± 2.2)

Dry

10.9 (4.3 ± 3.4)

7.5 (3 ± 1.1)

12 (4.8 ± 0.9)

353.7

Pau Roxo forest

Wet

9.5 (3.8 ± 2.6)

2.5 (1 ± 1.5)

8.2 (3.3 ± 1.4)

234.2

Dry

8.7 (3.5 ± 1)

4.2 (1.7 ± 1)

6.2 (2.5 ± 0.8)

129.7

Carrasco forest

Wet

22.7 (2.5 ± 2.2)

11.8 (1.3 ± 1.2)

9.0 (1 ± 0.6)

10.4

Dry

27.3 (3 ± 2.2)

9.0 (1 ± 0.9)

6.4 (0.7 ± 0.8)

40 11

in some cases, such as in red brocket deer Mazama americana Erxleben, we recorded up to three individuals together, the pair plus an infant, or in the case of the puma Puma concolor Linnaeus, we recorded 3 sub–adults in a cluster. Data analysis Principal components analysis (PCA) was performed with correlation matrices for the means of nine habitat variables calculated for each of the five studied forest types so that the entire set of variables could be reduced parsimoniously into a smaller number of descriptors, those sharing similar principal component scores, and so submitted to the same environmental constraints (Summerville et al., 2006). This allowed us to methodologically encompass as much variability as possible without losing information (Manly, 1986), and yielded uncorrelated composite measures that were used to fit stepwise multiple regressions to model the effects of these habitat variables on the densities of the studied mammalian species. The following variables were recorded per habitat: (1) tree basal area, (2) total number of tree species recorded during the study, (3) total number of animal dispersed fruit species, (4) mean number of fruit species recorded on the forest floor, (5) kg/ha/year of fruits on the forest floor, (6) mean number of marked trees with leaves, (7) mean number of marked trees with flower, (8) mean number of marked trees with fruits, and (9) kg of fruits in the canopy of marked trees.

As the quality measures of the trait variables were on very different scales, correlation matrices were preferred over covariance matrix, so avoiding biasing the eigenvalues (Jolliffe, 2002; Graham, 2003). PCA factors were tested using Horn´s parallel analysis (PA), since it is considered one of the most appropriate methods to select significant principal components (Zwick and Velicer, 1986), comparing PCA factor eigenvalues to threshold eigenvalues in order to find those factors with eigenvalues large enough to be retained. Factor variable loadings within significant factors were tested by Spearman rank correlation and only significant loadings were used in subsequent stepwise multiple regression analyses between the retained PCA axes and the ranked dependent variables (mammalian species densities) to determine which regression models that best predicted habitat preferences of the mammalian species. We adopted a significance level of 5 %. All analyses were performed using the R–program version 2.13.1 (R Core Team, 2012). Results Forest productivity: phenology of marked trees For the two major forest types that dominate eastern Maracá, Terra Firme and Mixed forest, Terra Firme had a higher percentage of trees with leaf buds, both in the wet season (9.9 %, n = 203) and in the

Animal Biodiversity and Conservation 43.1 (2020)

Table 2. Raked–ground fruit survey in the five studied forest types at Maracá Ecological Station, northernmost Brazilian Amazonia. Tabla 2. Estudio de los frutos recolectados en el suelo en los cinco tipos de bosque estudiados en la Estación Ecológica de Maracá, en el extremo septentrional de la Amazonia brasileña. Forest

Season

Fruit species (mean ± SD) on the forest floor

Kg/ha/year of fruit on the forest floor

Terra Firme forest

Wet

36 ± 7.7 (n = 216)

239

Dry

23.5 ± 4.3 (n = 141)

Mixed forest

Wet

16.7 ± 5.6 (n = 100)

176

Dry

10.5 ± 3.3 (n = 63)

53.5

Buritizal forest

Wet

6.8 ± 1.8 (n = 41)

2,062.5

Dry

5.3 ± 1.4 (n = 32)

3,212.2

Pau Roxo forest

Wet

2.5 ± 0.84 (n = 15)

750

Dry

1.5 ± 0.84 (n = 9)

79.9

Carrasco forest

Wet

2.2 ± 1.6 (n = 13)

4.16

Dry

0.7 ± 0.5 (n = 4)

0.4

dry season (8.3 %, n = 203), a higher percentage of trees with flowers (wet season: 4.9 %, dry season: 6 %, n = 203), and with fruits (wet: season 13.4 %, dry season: 9.8 %, n = 203). In the case of kilograms of fruits in the canopy of trees, while there were more in the wet season in Terra Firme forest (776.7 kg), there were more in the dry season in Mixed forest (634.5 kg) (table 1). For the two minor habitats, with similar forest height, Buritizal and Pau Roxo forest, Buritizal had a higher percentage of trees with leaf buds in both the wet (14.2 %, n = 40) and dry seasons (10.75 %, n = 40), as well as a higher percentage of flowers (wet: 5.7 %, dry: 7.5 %, n = 40), a higher percentage of fruits (wet: 13.2 %, dry: 12 %, n = 40), and more kilograms of fruits in the canopy of trees (wet: 521.1 kg, dry: 353.7 kg, n = 40) (table 1). In the shortest minor forest type, Carrasco, where only 11 trees with DBH ≥ 10 cm were recorded, more trees with leaf buds were recorded in the dry season (27.3 %, n = 11), whereas more trees with flowers (11.8 %, n = 11), with fruits (9 %, n = 11), and more kilograms of fruits in the canopy of trees (10.4 kg), were recorded in the wet season (table 1). Forest productivity: raked–ground fruit survey For the major forest types, the highest mean number of fruit species on the forest floor was recorded in Terra Firme forest, in both the wet (36 ± 7.7, n = 216), and dry seasons (23.5 ± 4.3, n = 141); in the case of the minor forest types of similar height, the highest mean number of fruit species on the forest floor was found in Buritizal forest, also in both the wet (6.8 ± 1.8,

n = 41), and dry seasons (5.3 ± 1.4, n = 32); in the lower minor habitat, Carrasco, the greatest number of fruit species was recorded in the wet season (2.2 ± 1.6, n = 13) (table 2). In the major forest types, the highest residual amounts of fruit on the forest floor in the wet season were recorded in Terra Firme (239 kg/ha/year), while in the dry season they were recorded in Mixed forest (53.5 kg/ha/year). For the minor forest types, Buritizal had by far the highest amounts of residual fruit on the forest floor in both the wet (2,062.5 kg/ha/year) and dry season (3,212.2 kg/ha/year); In Carrasco the highest residual amounts of fruit on the forest floor occurred in the wet season (4.16 kg/ha/year) (table 2). Mammalian species densities During this 12–month study, and across all five habitat types, 33 mammalian species were recorded over a total of 638 hours of line transect samplings, with 599 sightings recorded along the 1,180.2 km walked in a total of five habitat types (1,115.5 in diurnal, and 64.8 in nocturnal sampling). Density of terrestrial mammals (n = 14) was at least three times higher (1,636.85 groups/km2) than that of the arboreal and scansorial (n = 19) forms (484.05 groups/km2). The most abundant mammal species was Brazilian tapir, Tapirus terrestris with an overall density of 626.5 groups/km2, followed by red brocket deer, Mazama americana, with 449.4 groups/km2. The most abundant order was Artiodactyla, with 673 groups/km2, in which the largest contribution comes from M. americana, followed by Tayassu pecari, with 155 groups/km2, and Pecari tajacu, with

Mendes Pontes et al.

Table 3. Densities (groups/km2) in the different forest types at Maracá Ecological Station, Roraima, Brazilian Amazonia. (For abbreviations of forest types, see fig. 3). Tabla 3. Densidad (grupos/km2) en los diferentes tipos de bosque en la Estación Ecológica de Maracá, en Roraima, Amazonia brasileña. (Para las abreviaturas de los tipos de bosque, véase fig. 3). Species

Forest types

TFF Wet

Dry

Wet

BF Dry

Wet

Dry

PRF

Wet Dry

O. Artiodactyla F. Cervidae Mazama americana

11.4

13.3

9.7

183.3 85.7

133

F. Tayassuidae Tayassu pecari

2 4.1 0.7 1 26.8 119.4 0 1 0 0

Pecari tajacu

8.9

0.05

33.3

O. Carnivora F. Canidae Cerdocyon thous

7.3 3.3 0 0 0 0 0 0 0 0

Spheotos venaticus

0.6 0 0 0 0 0 0 0 0 0

F. Felidae Herpailurus yagouaroundi 3.3 0 0 0 0 0 0 0 0 0 Leopardus pardalis

0 0 0 0 0 0 0 0 0 0

L. tigrinus

0 0 0 0 0 0 0 0 0 0

L. wiedii

0 0 0 0 0 0 0 0 0 0

Panthera onca

0.6 1.1 0 0 2.9 2 0 0 0 0

Puma concolor

3.3 6.7 10 0 0 0 0 0 0 0

F. Mustelidae Eira barbara

0.8

0.2

0.5

Galictis vitatta

4 0 0 0 0 0 0 0 0 0

F. Procyonidae Bassaricyon beddardi 8.3 4.3 0 0 104.1 0 0 0 0 0 O. Perissodactyla F. Tapiridae Tapirus terrestris

4.5

150

300

O. Primates F. Cebidae Aotus trivirgatus

66.6 100 0 0 0 0 0 0 0 0

Ateles belzebuth

0.9 0.4 2.1 1.3 9.6 2.6 6 2.8 0 0

Alouatta macconnelli

0.7

0.3

0.4

0.5

3.7

Sapajus apella

0.4

0.15

0.3

Cebus olivaceus

1.6

2.5

1.4

1.3

1.7

5.5

1.6

Saimiri sciureus

0.6 1.2 0.3 0.2

0 3.3 2.4

O. Rodentia F. Cuniculidae Cuniculus paca

0 0 0 112.3 0 0 0 0 0 0

Animal Biodiversity and Conservation 43.1 (2020)

Table 3. (Cont.) Forest types Species

TFF Wet

Dry

Wet

Dry

Wet

PRF Dry

Wet

Dry

Wet Dry

F. Dasyproctidae Dasyprocta leporina

4.4

3.5

5.6

6.9

3.1

2.3

F. Erethizontidae Coendou prehensilis

F. Sciuridae 3

Sciurus igniventris

5.5

4.6

3.8

O. Cingulata F. Dasypodidae Cabassous unicinctus

0 0 0 0 0 0 0 0 0 0

Dasypus kappleri

0 0 0 0 0 0 0 0 0 0

D. novemcinctus

0 0 0 3.3 0 0 0 0 0 0

D. septemcinctus

0 0 0 3.3 0 0 0 0 0 0

Priodontes maximus

0 0 0 0 166.7 0 0 0 0 0

O. Pilosa F. Cyclopedidae Cyclopes didactylus

4.5

0.2

F. Myrmecophagidae 0

Myrmecophaga tridactyla

1.5

0.6

Tamandua tetradactyla 0.2 0.3 0.8 0 16.7 0 0 0 0 0 Total

180.2 166.1 48.4 148.8 691.8 402.5

Grand total

346.3

197.2

1,094.3

9.1

8.7 436.3 29.1 17.8

465.4

Table 4. Contribution of the components to the abundance of the mammalian species at Maracá Ecological Station, northernmost Brazilian Amazonia. Tabla 4. Contribución de los componentes a la abundancia de las especies de mamíferos en la Estación Ecológica de Maracá, en el extremo septentrional de la Amazonia brasileña.

Principal component Proportion of variance (%) PC1

Cumulative proportion (%)

0.8037 0.8037

PC2

0.1278

0.9315

PC3

0.04135

0.97285

PC4

0.01538

0.98823

PC5

0.00838

0.99661

PC6

0.00309 0.99970

PC7

0.00023 0.99993

PC8

0.00007 1

PC9

0.00000 1

Mendes Pontes et al.

Table 5. Importance of the variables in the components (those significant according to a Spearman rank correlation test are in bold) for the abundance of the mammalian species at Maracá Ecological Station, northernmost Brazilian Amazonia. Tabla 5. Importancia de las variables en los componentes (las que son significativas según una prueba de correlación por rangos de Spearman están en negrita) para la abundancia de las especies de mamíferos en la Estación Ecológica de Maracá, en el extremo septentrional de la Amazonia brasileña. Variables

PC1 –0.3581 –0.963

< 0.0001

2 Total number of tree species recorded during the study

–0.3536 –0.951

< 0.0001

3 Total number of animal–dispersed fruit species

–0.3583 –0.964

< 0.0001

4 Mean number of fruit species on the forest floor

–0.3510 –0.944

< 0.0001

5 Kg/ha/year of fruit on the forest floor

0.1234

6 Mean number of marked trees with leaves

–0.3661 –0.985

< 0.0001

Basal area

0.332

0.3484

–0.3520 –0.947

< 0.0001

8 Mean number of marked trees with fruits

–0.3513 –0.944

< 0.0001

9 Kg of fruit in the canopy of marked trees

–0.3139 –0.844

< 0.0001

Mean number of marked trees with flowers

69.2 groups/km2. The second most abundant order was Perissodactyla, with 626.5 groups/km2, with its only species, T. terrestris (table 3). Buritizal was the forest type with the highest overall mammal density, with 1,094.3 groups/km2, followed by Carrasco forest, with 465.4 groups/km2, and Terra Firme forest, with 346.3 groups/km2. The forest with the lowest density was Pau Roxo, with only 17.8 groups/km2 (table 3).

Selection of significant habitat variables via principal components analysis The principal components analysis results, subsequently tested via Horn´s parallel analysis, partitioned the 9 species traits into one factor with eigenvalues large enough to be retained (PC1), which explained 80.37 % of the total variation contained in the covariates (table 4). Spearman

Table 6. Stepwise multiple regression for the effect of the variables selected via PCA on the abundance of mammalian species at Maracá Ecological Station, northernmost Brazilian Amazonia: SE, standard error. Tabla 6. Regresión múltiple escalonada para determinar los efectos de las variables seleccionadas a través del análisis de componentes principales de la abundancia de especies de mamífero en la Estación Ecológica de Maracá, en el extremo septentrional de la Amazonia brasileña: SE, error estándard. Variables / parameters

Coefficient

p–value

Basal area

–6.21

2.52

0.0457

Total number of tree species recorded during the study

–6.25

2.20

0.0362

Mean number of marked trees with leaves

154.60

44.83

0.0183

Mean number of marked trees with flowers

109.04

42.95

0.0420

Mean number of marked trees with fruits

–33.23

12.91

0.0487

86.58

6.45

d.f.

5, 5

p–value 0.0308

800 700 600 500 400 300 200 100 0

200 150 100

m2/ha

Groups/km2

Animal Biodiversity and Conservation 43.1 (2020)

Density

50 Wet Dry Wet Dry Wet Dry Wet Dry Wet Dry TFF MF BF PRF CF Forest types

Basal area

Fig. 5. Relationship between mammalian density (groups/km2) and forest basal area at Maracá Ecological Station, Roraima, Brazil. (For abbreviations of forest types, see fig. 3).

800 700 600 500 400 300 200 100 0

350 300 250 200 150 100 50 0

m2/ha

Groups/km2

Fig. 5. Relación entre la densidad de mamíferos (grupos/km2) y área basimétrica forestal en la Estación Ecológica de Maracá, en Roraima (Brasil). (Para las abreviaturas de los tipos de bosque, véase fig. 3).

Wet Dry Wet Dry Wet Dry Wet Dry Wet Dry TFF MF BF PRF CF Forest types

Density Number of tree species

Fig. 6. Relationship between mammalian density (groups/km2) and number of tree species per forest type at Maracá Ecological Station, Roraima, Brazil. (For abbreviations of forest types, see fig. 3). Fig. 6. Relación entre la densidad de mamíferos (grupos/km2) y el número de especies arbóreas por tipo de bosque en la Estación Ecológica de Maracá, en Roraima (Brasil). (Para las abreviaturas de los tipos de bosque, véase fig. 3).

Rank Correlation tests of loading coefficients of trait variables identified eight significant factor variable loadings (P < 0.05) for the first factor (PC1) (table 5). Stepwise multiple regressions for the effects of environmental variables on mammalian species density Stepwise multiple regressions identified five of the eight environmental variables selected via PCA as predicting a significantly high amount of the variability in the mammalian species density in the forest types studied (F = 6.45 , d.f. = 5,5, P = 0.0308, R2 = 86.6) (table 6). Density was negatively associated with tree

basal area, total number of tree species registered during the study, and mean number of marked trees with fruits (fig. 5–7, table 6), and positively related to mean number of marked trees with leaves, and mean number of marked trees with flowers (fig. 7, table 6). Discussion Maracá Ecological Station is located in the central zone of the Guyana Shield, one of the driest regions in Brazilian Amazonia, with one of the longest dry seasons, most frequent canopy discontinuities, and consequently, a very well–developed understorey (Mendes Pontes, 2004; Mendes Pontes et al., 2013).

Mendes Pontes et al.

30 28 26 24 22 20 18 16 14 12 10 8 6 4 2 0

700 650 600 550 500 450 400 350 300 250 200 150 100 50 0

Groups/km2

Tree reproductive phenology

Wet Dry Wet Dry Wet Dry Wet Dry Wet Dry TFF MF BF PRF CF Forest types

Mean number of threes with leaf Mean number of threes with flower Mean number of threes with fruit Density

Fig. 7. Relationship between mammalian density (groups/km2) and tree reproductive phenology at Maracá Ecological Station, Roraima, Brazil. (For abbreviations of forest types, see fig. 3). Fig. 7. Relación entre la densidad de mamíferos (grupos/km2) y la fenología reproductiva de los árboles en la Estación Ecológica de Maracá, en Roraima (Brasil). (Para las abreviaturas de los tipos de bosque,, véase fig. 3).

One of the most remarkable features of these seasonally–dry forests was the dramatic seasonal fluctuation on the availability of edible plant parts, both in the canopy and on the floor of the five forest types studied. As is typical of these regions (Janzen, 1974; Eisenberg and Redford, 1979; Eisenberg et al., 1979; August, 1983; Schaller, 1983; Emmons, 1984), the abundance of the terrestrial mammalian fauna was higher than the arboreal fauna (more than three times higher: see Mendes Pontes, 2004). As the most abundant orders, browsers and grazers (Artiodactyla and Perissodactyla) could benefit most from a well–developed understory. In this scenario, tree reproductive phenology was the sole determinant of the overall abundance in the mammalian assemblages, and in turn it translated into more food resources for the vegetarians/primarily vegetarian species, and indirectly, also for meat eaters. Thus, we confirmed for this site what had already been suggested in the literature on a much broader scale, that forest productivity is the main determinant of abundance of mammalian assemblages (Eisenberg et al., 1979; Eisenberg, 1980; Gentry and Emmons, 1987; Eisenberg, 1990; Kay et al., 1997; Hoorn et al., 2010; Marshall et al., 2014) more specifically, this being productivity as mediated by tree reproductive phenology (a relationship reported by Marshall et al., 2014).

Indirectly, our findings also support the assumption that soil type is the main determinant of mammalian abundance (Tuomisto et al., 1995, 2003, 2014; Hoorn et al., 2010; Kristiansen et al., 2012; Pomara et al., 2012, 2014; Rossetti, 2014; Higgins et al., 2011, 2015; Zuquim et al., 2014), since it will ultimately determine the structure and productivity of the forests (Eisenberg et al., 1979; Eisenberg, 1980; Emmons, 1984; Gentry and Emmons, 1987; Eisenberg, 1990; Kay et al., 1997; Hoorn et al., 2000; Marshall et al., 2014) via resource availability and seasonality, and hence the intensity of food competition between species (Connell, 1980; Chesson and Huntley, 1989; Chesson and Rosenzweig, 1991). It has been shown in our previous studies (Gadelha et al., 2017; Luna et al., 2017) that at the regional scale (Peterson et al., 2011), the abundance of the mammalian assemblages can be determined by the physical structure of forests, whether it be in highly–heterogeneous vegetation mosaics, such as in the northern part of the Guyana Shield, or the ombrophilous forests of the southern part, or a transition between them. Now, we have also shown that at the local scale (Peterson et al., 2011), forest structure and productivity, and more specifically food availability and its fluctuations, determined the abundance of the mammalian assemblages.

Animal Biodiversity and Conservation 43.1 (2020)

Basal area, total number of tree species, and total number of tree species with fruits also influenced the recorded mammalian densities, but because in these more open, discontinuous forests of the Guyana Shield the bulk of the mammalian abundance is made up of large, terrestrial, mostly browser and grazer species (at Maracá three times higher than smaller species), the basal area and total number of tree species related negatively to mammalian densities. As a result, Artiodactyla and Perissodactyla, the most abundant orders, favoured more open–canopy forests, which had more abundant undergrowth on which they could feed. The total number of tree species with fruits also showed a negative relationship with mammalian densities, possibly because the most abundant orders/ species were primarily browsers and grazers, despite also having a diet containing a considerable proportion of fruits (Emmons and Feer, 1997). Finally, Maracá Ecological Station is located in the centre of the Guyana Shield and is covered by a mosaic of contiguous forest types that are highly seasonal and have a long dry season. It harbors mammalian assemblages that are one of the poorest in species diversity in the northernmost Brazilian Amazonia, with large–bodied mammals being more abundant than medium–sized mammals, and terrestrial species more abundant than arboreal species. In these assemblages, the mammalian species favoured forest types that were more open and had more zoochorous fruit species. In this scenario, the sole determinant of their densities was the tree assemblage reproductive phenology, confirming what has been previously postulated in the literature, that mammalian abundances should be a function of forest productivity. Acknowledgements We thank Dr T. Pennington, Dr G. Lewis, Dr L. Lohman, and Dr M. Van Roosmalen for crosschecking plant/fruit identifications. This study was funded by the Brazilian Ministry of Education–CAPES. We thank the following institutions for sponsoring field work: Alliance for the Preservation of Species–Germany; Fauna and Flora International–UK; The Isaac Newton Trust, Trinity College and Satusoma Trust, University of Cambridge–UK; British Airways Assisting Conservation–UK; The Margaret Mee Amazon Trust–UK. The Brazilian Environmental Office kindly granted permission to carry out this study in Maracá. Adrian Barnett corrected the English. References Andrewartha, H. G., 1961. An introduction to the study of animal populations. University of Chicago Press, Chicago. August, P. V., 1983. The role of habitat complexity and heterogeneity in structuring tropical mammal communities. Ecology, 64: 1495–1507. Blackburn, T. H., Harvey, P. H., Pavel, M. D., 1990. Species number, population density and body

size relationship in natural communities. Journal of Animal Ecology, 59: 335–345. Brockelman, W. Y., Ali, R., 1987. Methods of surveying and sampling forest primate populations. 2. In: Primate conservation in the tropical rainforest: 23–62 (C. W. Marsh, R. A. Mittermeier, Eds.). John Wiley and Sons, New York. Brown, J. H., 1981. Two decades of homage to Santa Rosalia: toward a general theory of diversity. American Zoologist, 21: 255–279. Brown, J. W., 1984. On the relationship between abundance and distribution of species. American Naturalist, 124: 255–279. Brown, J. W., Maurer, B. A., 1986. Body size, ecological dominance, and Cope's rule. Nature, 324: 248–250. Buckland, S. T., Anderson, D. R., Burnham, K. P., Laake, J. L., 1993. Distance sampling. Estimating abundance of biological populations. Chapman and Hall, London. Burnham, K. P., Anderson, D. R., Laake, J. L., 1980. Estimation of density from line transect sampling of biological populations. Wildlife Monographs, 72: 1–202. Caldecott, J., 1988. Hunting and wildlife management in Sarawak. IUCN, Gland, Switzerland. Carvalho, A. P., Larach, J. O. I., Camargo, M. N., 1988. Critérios para distinção de classes de solos e de fases de unidade de mapeamento. Empresa Brasileira de Pesquisa Agropecuária. Serviço Nacional de Conservação e Levantamento de Solos, Rio de Janeiro, Brazil. Chapman, C. A., Chapman, L. J., Bjorndal, K. A., Onderdonk, D. A., 2002. Application of protein– to–fiber ratios to predict colobine abundance on different spatial scales. International Journal of Primatology, 23: 283–310. Chiarello, A. G., 1999. Effects of fragmentation of the Atlantic forest on mammal communities in south–eastern Brazil. Biological Conservation, 89(1): 71–82. Chesson, P. L., Huntley, N., 1988. Community consequences of life history traits in a variable environment. Annales Zoologici Fennici, 25: 5–16. Chesson, P. L., Rosenzweig, M., 1991. Behaviour, heterogeneity and the dynamics of interacting species. Ecology, 72: 1187–1195. Clutton–Brock, T. H., Harvey, P. H., 1977. Species differences in feeding and ranging behaviour in primates. In: Primate ecology: studies of feeding and ranging behaviour in lemur, monkeys and apes: 557–582 (T. H. Clutton–Brock, Ed.). Academic Press, London. Connell, J. H., 1980. Diversity and the evolution of competitors, or the ghost of competition past? Oikos, 35: 131–138. Damuth, J., 1981. Population density and body size in mammals. Nature, 290: 699–670. Eisenberg, J. F., 1980. The density and biomass of tropical mammals. In: Conservation biology: 35–55 (M. Soulé, B. Wilcox, Eds.). Sinauer Associates, Sunderlands, Mass. – 1990. Neotropical mammal communities. In: Four

neotropical rainforests: 35–55 (A. H. Gentry, Ed.). Yale University Press, New Haven, CT. Eisenberg, J. F., Redford, K., 1979. A biogeographical analysis of a mammalian fauna of Venezuela. In: Vertebrate ecology in the northern neotropics: 31–36 (J. F. Eisenberg, Ed.). Smithsonian Institution Press, Washington D.C. Eisenberg, J. F., Thorington, Jr., R. W., 1973. A preliminary analysis of a Neotropical mammal fauna. Biotropica, 5(3): 150–161. Eisenberg, J. F., O'connell, M. A., August, P., 1979. Density, productivity, and distribution of mammals in two Venezuelan habitats. In: Vertebrate ecology in the northern neotropics: 187–207 (J. F. Eisenberg, Ed.). Smithsonian Institution Press, Washington D.C. Elton, C., 1932. Territory among wood ants (Formica rufa L.) at Picket hill. Journal of Animal Ecology, 1: 69–76. – 1933. The ecology of animals. Methuen, London. EMBRAPA, 2006. Sistema brasileiro de classificação de solos. Centro Nacional de Pesquisa de Solos, Rio de Janeiro. Emmons, L. H., 1984. Geographic variation in densities and diversities of non–flying mammals in Amazonia. Biotropica, 16: 210–222. Emmons, L. H., Feer, F., 1997. Neotropical rainforest mammals: a field guide. University of Chicago Press, Chicago. Foster, R. B., 1982a. Famine on Barro Colorado island. In: The ecology of a tropical forest: Seasonal rythms and long–term changes: 201–212 (E. G. Leigh Jr., A. S. Rand, D. M. Windsor, Eds.). Smithsonian Institution Press, Washington, D.C. – 1982b. The seasonal rhythm of fruit fall on Barro Colorado island. In: The ecology of a tropical forest. Seasonal rhythms and long–term changes: 151– 172 (E. G. Leigh Jr., A. S. Rand, D. M. Windsor, Eds.). Smithsonian Institution Press, Washington, D.C. Fragoso, J., 1998. Home range and movement patterns of white–lipped peccary, Tayassu pecari herds in the northern Brazilian Amazon. Biotropica, 30: 458–469. – 2004. A long–term study of white–lipped peccary, Tayassu pecari population fluctuation in northern Amazonia: Anthropogenic vs natural causes. In: People in nature: wildlife conservation in south and central America: 286–296 (K. Silvius, R. E. Bodmer, J. M. Fragoso, Eds). Columbia University Press, New York. Gadelha, J. R., Melo, E. R. A., Silva, M. N. D., Silva Júnior, A. P., Filgueiras, B. K. C., Mendes Pontes, A. R., 2017. Habitat heterogeneity as a local and regional–scale driver of primate assemblage structure in northernmost Brazilian Amazonia. Animal Biology, 67: 263–286. Gaston, K. J., Lawton, J. H., 1988. Patterns in body size, population dynamics, and regional distribution of bracken herbivores. American Naturalist, 132: 662–680. Gentile, R., Fernandez, F. A. S., 1999. Influence of habitat structure on a streamside small mammal

Mendes Pontes et al.

community in a Brazilian rural area. Mammalia, 63: 29–40. Gentry, A. H., Emmons, L. H., 1987. Geographic variation in fertility, phenology, and composition of Neotropical forests. Biotropica, 19: 216–227. Glanz, W. E., 1990. Neotropical mammal densities: how unusual is the community in Barro Colorado island, Panama? In: Four neotropical rainforests: 287–313 (A. H. Gentry, Ed.). Yale University Press, New Haven, CT. Graham, M. H., 2003. Confronting multicollinearity in ecological multiple regression. Ecology, 84: 2809–2815. Grelle, C. E. V., 2003. Forest structure and vertical stratification of small mammals in a secondary Atlantic forest, southeastern Brazil. Studies on Neotropical Fauna and Environment, 38: 81–85. Guillotin, M., Dubost, G., Sabatier, D., 1994. Food choice and food competition among the three major primate species of French Guiana. Journal of Zoology London, 233: 551–579. Higgins, M. A., Asner, G. P., Anderson, C. B., Martin, R. E., Knapp, D. E., Tupayachi, R., Perez, E., Elespuru, N., Alonso, A., 2015. Regional–scale drivers of forest structure and function in north western Amazonia. Plos One, 10(3): e0119887, Doi:10.1371/journal.pone.0119887 Higgins, M. A., Ruokolainen, K., Tuomisto, H., Llerena, N., Cardenas, G., Phillips, O. L., Vásquez, R., Räsänen, M.,, 2011. Geological control of floristic composition in Amazonian forests. Journal of Biogeography, 38: 2136–2149. Hoorn, C. F. P., Wesselingh, H., Ter Steege, M. A., Bermudez, M. A., Mora, A., Sevink, J. et al., 2010. Amazonia through time: andean uplift, climate change, landscape evolution and biodiversity. Science, 330: 927–931. Hutchinson, G. E., MacArthur, R. H., 1959. A theoretical ecological model of size distributions among species of animals. American Naturalist, 93: 117–125. Janson, C. H., Emmons, L. H., 1990. Ecological structure of the non–flying mammal community at Cocha Cashu Biological Station, Manu National Park, Peru. In: Four neotropical rainforests: 314–338 (A. H. Gentry, Ed.). Yale University Press, New Haven. Janson, C. H., Terborgh, J., Emmons, L., 1981. Non–flying mammals as pollinating agents in the Amazonian forest. Biotropica (Supplement: Reproductive Botany), 13(2): 1–6. Juanes, F., 1986. Population density and body size in birds. American Naturalist, 128: 921–929. Janzen, D. H., 1974. Tropical black water rivers, animals, and mast fruiting by the Dipterocarpaceae. Biotropica, 6: 69–103. Jolliffe, I. T., 2002. Principal component analysis. Springer–Verlag, New York. Kay, R. F., Madden, R. H., Van Schaik, C., Higdon, D., 1997. Primate species richness is determined by plant productivity: Implications for conservation. Proceedings of the National Academy of Sciences of the United States, 94: 13023–13027. Kinnaird, M., Eisenberg, J. F., 1989. A consideration

Animal Biodiversity and Conservation 43.1 (2020)

of body size, diet, and population biomass for Neotropical mammals. In: Advances in neotropical mammalogy: 595–603 (K. H. Redford, J. F. Eisenberg, Eds.). Sandhill Crane Press, Gainesville. Kristiansen, T., Svenning, J. C., Eiserhardt, W. L., Pedersen, D., Brix, H., Kristiansen, S. M., Knadel, M., Grández, C., Balslev, H., 2012. Environment versus dispersal in the assembly of western Amazonian palm communities. Journal of Biogeography, 39: 1318–1332. Lawton, J. H., 1989. What is the relationship between population density and body size in animals? Oikos, 55: 429–434. Lewis, G. P., Owen, P. E., 1989. Legumes of the ilha de Maracá. Royal Botanic Gardens, London. Luna, R. B., Reyes, A. F. A., Lucena, L. R. R., Mendes Pontes, A. R., 2017. Terrestrial mammal assemblages in protected and human impacted areas in Northern Brazilian Amazonia. Nature Conservation, 22: 147–167. MacArthur, R. H., 1969. Species packing and what competition minimizes. Proceedings of the National Academy of Sciences of the United States, 1: 1–11. Malcolm, J. R., 1990. Estimation of mammalian densities in continuous forest north of Manaus. In: Four neotropical rainforests: 339–357 (A. H. Gentry, Ed.). Yale University Press, New Haven, CT. – 1995. Forest structure and the abundance and diversity of Neotropical small mammals. In: Forest canopies: 179–197 (M. D. Lowman, N. M. Nadkarni, Eds.). Academic Press, San Diego. Manly, B. F. J., 1986. Multivariate statistical methods: a primer. Chapman and Hall, London. Marshall, A. J., Beaudrot, L., Wittmer, H. U., 2014. Responses of primates and other frugivorous vertebrates to plant resource variability over space and time at Gunung Palung National Park. International Journal of Primatology, 35: 1178–1201. Marshall, A. J., Leighton, M., 2006. How does food availability limit the population density of white bearded gibbons? In: Feeding ecology of the apes and other primates: 311–333 (G. Hohmann, M. Robbins, C. Boesch, Eds.). Cambridge University Press, Cambridge. Maurer, B. A., Brown, J. H., 1988. Distribution of energy use and biomass among species of North– American terrestrial birds. Ecology, 69: 1923–1932. Mazurek, R. R. S., Temehe, P., Xinymy, F., Waraié, H., Sanapyty, G., Ewepe, M., 2000. Subsistence hunting among the Waimiri Atroari Indians in central Amazonia, Brazil. Biodiversity and Conservation, 9(5): 579–596. Melo, E. R., Gadelha, J. R., Silva, M. N. D., Silva Jr., A. P., Mendes Pontes, A. R., 2015. Diversity, abundance and the impact of hunting on large mammals in two contrasting forest sites in northern Amazon. Wildlife Biology, 21(5): 234–245. Mendes Pontes, A. R., 1997. Habitat partitioning among primates in Maracá Island, Roraima, northern Brazilian Amazonia. International Journal of Primatology, 18: 131–157. – 1999. Environmental determinants of primate abundance in Maracá Island, Roraima, Brazil-

ian Amazonia. Journal of Zoology London, 247: 189–199. – 2004. Ecology of a community of mammals in a seasonally dry forest in Roraima, Brazilian Amazonia. Mammalian Biology, 69: 319–336. Mendes Pontes, A. R., Chivers, D. J., Lee, P., 2007. Effect of biomass on assemblages of large mammals in a seasonally dry forest in the Brazilian Amazonia. Journal of Zoology London, 271: 278–287. Mendes Pontes, A. R., Holderbaum, J. M., Layme, V. M. G, Chivers, D. J., 2013. The forests of Maracá, Northern Brazilian Amazonia: structure, phenology, fruit assessment and an illustrative field guide. Editora Universitária da UFPE, Recife, Brazil. Mendes Pontes, A. R., Paula, M. D., Magnusson, W. E., 2012. Low primate diversity and abundance in northern Amazonia and its implications for conservation. Biotropica, 44: 834–839. Milliken, W., Ratter, J., 1990. The Vegetation of the Ilha de Maracá. First report of the vegetation survey of the Maracá Rainforest Project. Royal Botanic Garden, Edinburgh. Milton, K., May, M. L., 1976. Body weight, diet and home range area in Primates. Nature, 259: 459–463. Ministério do Interior, 1977. Programa das Estações Ecológicas. Secretaria Especial do Meio Ambiente. Série Meio Ambiente, 2: 1–39. Montgomery, G. G., Sunquist, M. E., 1978. Habitat selection and use by two–toed and three–toed sloths. In: The ecology of arboreal folivores: 324–360 (G. G. Montgomery, Ed.). Smithsonian Institution Press, Washington, D.C. Muller–Dombois, D., Ellemberg, H., 1974. Aims and methods of vegetation ecology. John Wiley and Sons, New York. Nascimento, M. T., Proctor J., 1996. Seed attack by beetles and leaf–cutter ants on Peltogyne gracilipes Ducke (Caesalpiniaceae) on Maracá island, Brazilian Amazonia. Journal of Tropical Ecology, 12: 723–727. Nascimento, M. T., Proctor, J., Vilella, D. M., 1997. Forest structure, floristic composition and soils of an Amazonian monodominant forest on Maracá island, Roraima, Brazil. Edinburgh Journal of Botany, 54: 1–38. Nee, S., Read, A. F., Greenwood, J. J. D., Harvey, P. H., 1991. The relationship between abundance and body length in British birds. Nature, 351: 312–313. NRC (National Research Council), 1981. Techniques for the study of primate population ecology. National Academic Press, Washington, D.C. Nunes, A. P., 1992. Uso do habitat, comportamento alimentar e organização social de Ateles belzebuth belzebuth (Primates: Cebidae). PhD thesis, Universidade Federal do Pará and Museu Paraense Emílio Goeldi. Nunes, A. P., Ayres, J. M., Martins, E. S., Silva, J. S., 1988. Primates of Roraima (Brazil). I. Northeastern part of the Territory. Boletim do Museu Paraensi Emilio Goeldi. Serie Zoologia, 4(1): 1–33. Peres, C. A., 1994. Primate response to phenological changes in an Amazonia Terra Firme forest.

Biotropica, 26(1): 98–112. Peters, R. H., Raelson, J. V., 1984. Relations between individual size and mammalian population density. American Naturalist, 124(4): 498–517. Peters, R. H., Wassenberg, K., 1983. The effects of body size on animal abundance. Oecologia, 60: 89–96. Peterson, A. T., Soberón, J., Pearson, R. G., Anderson, R. P., Martínez–Meyer, E., Nakamura, M. et al., 2011. Ecological niches and geographic distributions. Princeton University Press, Princeton. Pomara, L. Y., Ruokolainen, K., Tuomisto, H., Young, K. R., 2012. Avian composition co–varies with floristic composition and soil nutrient concentration in Amazonian upland forests. Biotropica, 44: 545–553. Pomara, L. Y., Ruokolainen, K., Young, K. R., 2014. Avian species composition across the Amazon River: the roles of dispersal limitation and environmental heterogeneity. Journal of Biogeography, 41: 784–796. R Core Team, 2012. R: A language and environment for statistical computing. R Foundation for Statistical Computing, Vienna. http://www.R–project.org/ Robinette, W. L., Loveloss, C. M., Jones, D. A., 1974. Field test of strip census methods. Journal of Wildlife Management, 38(1): 81–96. Robinson, J. G., Redford, K. H., 1989. Body size, diet, and population variation in Neotropical forest mammal species: predictors of local extinction? In: Advances in neotropical mammalogy: 567–594 (K. H. Redford, J. F. Eisenbergh, Eds.). Sundhill Crane Press, Gainesville. Rossetti, D. F., 2014. The role of tectonics in the late quaternary evolution of Brazil’s Amazonian landscape. Earth–Science Reviews, 139: 362–389. Sabatier, D., 1985. Saisonnalite et determinisme du pic de frutification en forest guyanaise. Revue D’Ecologie–La Terre et la Vie, 40: 289–320. Santini, L., Isaac, N. J. B., Maiorano, L., Ficetola, G. F., Huijbregts, M. A. J., Carbone, C., Thuiller, W., 2017. Global drivers of population density in terrestrial vertebrates. Global Ecology and Biogeography, 27: 968–979 Schaller, G. B., 1983. Mammals and their biomass on a Brazilian ranch. Arquivos de Zoologia, 31: 1–36.

Mendes Pontes et al.

Schoener, T. W., 1983. Field experiments in interspecific competition. American Naturalist, 122: 240–283. Smythe, N., 1978. The natural history of the Central American Agouti (Dasyprocta punctata). Smithsonian Contributions to Zoology, 257: 1–52. Summerville, K. S., Conoan, C. J., Steichen, R. M., 2006. Species traits as predictors of the frequency of occurrence of Lepidoptera in restored and remnant tallgrass prairies. Ecological Applications, 16: 891–900. Terborgh, J., 1983. Five new world primates: a study in comparative ecology. Princeton University Press, Chicago. Thompson, J., Proctor, J., Viana, V., Milliken, W., Ratter, J. A., Scott, D. A., 1992. Ecological studies on a lowland evergreen rain forest on Maraca Island, Roraima, Brazil. I. Physical environment, forest structure and leaf chemistry. Journal of Ecology, 80: 689–703. Tuomisto, H., Ruokolainen, K., Kalliola, R., Linna, A., Danjoy, W., Rodriguez, Z., 1995. Dissecting Amazonian biodiversity. Science, 269: 63–66. Tuomisto, H., Ruokolainen, K., Yli–Halla, M., 2003. Dispersal, environment, and floristic variation of western Amazonian forests. Science, 299: 241–244. Tuomisto, H., Zuquim, G., Cárdenas, G., 2014. Species richness and diversity along edaphic and climatic gradients in Amazonia. Ecography, 37: 1034–1046. Van Roosmalen, M. G. M., 1985. Fruits of the Guianan flora. Institute of Systematic Botany. Utrecht University, Utrecht. Zhang, S., Wang, L., 1995. Comparison of three fruit census methods in French Guiana. Journal of Tropical Ecology, 11: 281–295. Zuquim, G., Tuomisto, H., Jones, M. M., Prado, J., Figueiredo, F. O. G., Moulatlet, G. M., Flavia, R. C., Costa, F. R. C., Quesada, C. A., Emilio, T., 2014. Predicting environmental gradients with fern species composition in Brazilian Amazonia. Journal of Vegetation Science, 25: 1195–1207. Zwick, W. R., Velicer, W. F., 1986. Comparison of five rules for determining the number of components to retain. Psychological Bulletin, 99: 432–442.

Animal Biodiversity and Conservation 43.1 (2020)

A difference between sexes: temporal variation in the diet of Carollia perspicillata (Chiroptera, Phyllostomidae) at the Macaregua cave, Santander (Colombia) Á. Alviz, J. Pérez–Torres Alviz, A., Pérez–Torres, J., 2020. A difference between sexes: temporal variation in the diet of Carollia perspicillata (Chiroptera, Phyllostomidae) at the Macaregua cave, Santander (Colombia). Animal Biodiversity and Conservation, 43.1: 27–35, Doi: https://doi.org/10.32800/abc.2020.43.0027 Abstract A difference between sexes: temporal variation in the diet of Carollia perspicillata (Chiroptera, Phyllostomidae) at the Macaregua cave, Santander (Colombia). Organisms adjust their foraging strategies to optimize the energetic costs during foraging with respect to benefits gained. These strategies are usually different in males and females due to their specific requirements during reproduction. Knowing the temporal dietary composition and variation may help us understand how intrinsic factors can influence diet during the breeding season. Seba's short–tailed fruit bat (Carollia perspicillata) plays an important role in seed dispersal throughout the Neotropics. Seasonal dietary changes related to resource availability have been documented but dietary differences between males and females have not been analyzed. We tested the hypothesis that dietary breadth increases and varies between males and females of Carollia perspicillata during the breeding season. We collected 295 fecal samples (from 236 males and 182 females) between June 2012 and April 2013 at the Macaregua cave (Santander, Colombia). Sex, diet and overlap were recorded. Time series analysis of dietary variation were estimated and related to food (fruits and flowers) availability. Males were found to include 18 seed morphospecies within their diet, while females included 16 seed morphospecies. Ficus, Vismia and Acacia were the most commonly consumed plant genera within the diet of both males and females. The time series analysis throughout the year indicated that males had greater dietary diversity than females. Dietary richness for males peaked multiple times, while dietary richness for females peaked only once during the transition period between pregnancy and lactation. We recorded significant sex differences in the value of importance of plants in the diet, evenness, and dominance of plant species consumed, as well as differential consumption over the seasons. Knowing the variations in the diet allows us to address the differences between the foraging strategies that females and males use in response to energy demands, movement patterns and habitat use. This is essential to understand all those processes that organisms must carry out for their survival and maintenance. Key words: Bats, Conservation, Feeding ecology, Foraging, Nutrition Resumen Una diferencia entre sexos: variación temporal en la dieta de Carollia perspicillata (Chiroptera, Phyllostomidae) en la cueva de Macaregua, en Santander (Colombia). Los organismos ajustan sus estrategias de alimentación para optimizar la energía invertida en la búsqueda de alimento respecto a la energía obtenida. Estas estrategias suelen ser distintas entre machos y hembras debido a los requerimientos específicos durante la reproducción. Conocer la composición de la dieta y su variación en el tiempo puede ayudarnos a entender la forme en que los factores intrísecos pueden influir en la dieta durante la época reproductiva. Carollia perspicillata es una de las especies más importantes en el proceso de dispersión de semillas del Neotrópico. Se han documentado cambios estacionales en la dieta relacionados con disponibilidad de recursos, pero no se han descrito las diferencias en la alimentación entre ambos sexos. Evaluamos la hipótesis que en la época reproductiva, la dieta se vuelva más variada y se diferencia entre machos y hembras de Carollia perspicillata. Entre junio de 2012 y abril de 2013 se obtuvieron recolectaron 295 muestras de materia fecal (de 236 machos y 182 hembras) en la cueva de Macaregua, en Santander (Colombia). Se registraron el sexo, la dieta y el solapamiento. Se realizaron análisis de series cronológicas de la variación de la dieta y se relacionaron con la disponibilidad de alimento (frutos y flores) a lo largo del tiempo. ISSN: 1578–665 X eISSN: 2014–928 X

Alviz and Pérez–Torres

En su dieta se identificaron 18 morfoespecies de semillas para machos y 16 para hembras. Las especies más consumidas fueron de los géneros Ficus, Vismia y Acacia, tanto en machos como en hembras. El análisis de series cronológicas realizadas durante el año mostró que la diversidad de la dieta de los machos fue mayor que la de las hembras. Mientras que los machos presentaron más de un pico de riqueza de la dieta, las hembras mostraron solo uno que coincidió con el período de transición entre el embarazo y la lactancia. Se observaron diferencias significativas entre sexos en cuanto al valor de la importancia de las plantas en la dieta, a la riqueza y dominancia de las especies vegetales consumidas, además de un consumo diferenciado según la época del año. Conocer las variaciones que ocurren en la dieta nos permite abordar las diferencias entre las estrategias de alimentación que las hembras y los machos utilizan en respuesta a las demandas de energía, los patrones de movimiento y el uso del hábitat. Esto es esencial para comprender todos los procesos que los organismos deben llevar a cabo para su supervivencia y mantenimiento. Palabras claves: Murciélagos, Conservación, Ecología de la alimentación, Alimentación, Nutrición Received: 11 VII 18; Conditional acceptance: 21 I 19; Final acceptance: 05 IX 19 Ángela Alviz, Jairo Pérez–Torres, Unidad de Ecología y Sistemática (UNESIS), Laboratorio de Ecología Funcional, Departamento de Biología, Facultad de Ciencias, Pontificia Universidad Javeriana, 110231 Bogotá, Colombia. Corresponding author: Ángela Alviz. E–mail: aalviz86@gmail.com Orcid ID Á. Alviz: 0000–0002–7844–1418; J. Pérez–Torres: 0000–0001–7121–6210

Animal Biodiversity and Conservation 43.1 (2020)

Introduction Carollia perspicillata is the most abundant and widely distributed member of the genus Carollia (Fleming, 1988) and it is found in most evergreen forests in the neotropics (Cloutier and Thomas, 1992). Its diet is composed of a variety of fruits, flowers and insects. It is one of the main seed dispersers, helping to maintain plant heterogeneity, principally through the dispersion of pioneer plant species such as Piper and Cecropia (Barboza–Márquez and Aguirre, 2010; Kunz et al., 2011). Species of the genera Piper and Solanum comprise most of its diet, followed by a large variety of secondary plant species of the genera Cecropia and Vismia (Barboza–Marquez and Aguirre, 2010; Sánchez et al., 2012). During the dry season, when fruit production is low, its diet is mainly composed of nectar and pollen (Charles–Dominique, 1991; Cloutier and Thomas, 1992). Diet analyses have shown that C. perspicillata also consumes insects, possibly as a dietary supplement due to the low protein content of fruits (Mello et al., 2004a). The dietary variation of C. perspicillata has been described in Costa Rica (Fleming and Heithaus, 1986), Rio de Janeiro and southeastern Brazil (Marinho–Filho, 1991; Mello et al., 2004a), but diet variation between males and females has not been reported. Dietary differences between males and females could be the result of seasonal changes of resources or energetic demands during different reproductive stages. According to reports, females’ energetic demands increase continuously from fertilization to time of weaning (Sánchez et al., 2012). This is reflected in a 45 % increase in food consumption from pregnancy to lactation (Dietz and Kalko, 2006). Males only increase their energetic demands during gonad development and spermatogenesis (Cryan and Wolf, 2003; Pfeiffer and Mayer, 2013). Therefore, there likely exists a difference between male and female dietary items as sexes exhibit different energetic demands during their reproductive seasons. Understanding temporal dietary variation of this species might therefore allow us to explain behavioral traits, the relationship with the plants they feed on, their foraging patterns, and variations between males and females. We analyzed the temporal variations of dietary preferences for male and female C. perspicillata in a cave in eastern Colombia, and evaluated the relationship of these changes with reproductive stages and resource availability. We evaluated whether the richness of the diet was greater during the breeding season, both for females (pregnant or lactating) and for males (scrotal testes), and also whether the amplitude of the diet of females was greater than that of the males, regardless of the season. We hypothesized that, due to the differences in energetic demands between males and females in relation to their reproductive seasons and foraging activities, a temporal variation in the components of the diet between sexes would be evident. Since the females present an increase in nutritional requirements during pregnancy and lactation, there would

be marked variations throughout the year due to the increase in the consumption of food items during these reproductive seasons. Material and methods Macaregua cave (6º 39’ 36.2’’ N; 73º 6’ 32.3’’ W) is located in Las Vueltas, Municipality of Curití, Santander, Colombia at an elevation of 1.566 m, embedded in a remnant of tropical dry forest (fig. 1). The average annual precipitation is 1.499 mm, with the highest peak of rainfall (285 mm) occurring in April– May and a second lower peak in September– October (GELT, 2013). The entrance to the cave is surrounded by tropical dry forest vegetation. Inside the cave there is a dry room of about 80 m long, and a moist room of about 610 m long with a running stream (Perez– Torres et al., 2015). Sampling was conducted every two months between June 2012 and June 2013 to include both the dry and wet seasons. Bats were captured using cone traps and a mist net (6 x 3 m) located at the entrance to the cave, at between 02:00 and 05:00 h for four days each sampling month. The bats were mainly trapped when returning to the cave after foraging (Fleming and Heithaus, 1986; Charles–Dominique, 1991). Each individual was marked with forearm bands. Bats captured with traps were placed in individual cloth bags for 30' to 1 hour (Galindo–González, 1998; Heer et al., 2010; Mello et al., 2004b) to obtain fecal samples. Data collected for each individual were sex (male, female), reproductive status (males: testicles impalpable, inguinal or scrotal; females: active, inactive, pregnant, lactating) (Burnett and Kunz, 1982; Kunz and Anthony, 1982; Anthony, 1996), weight, and morphological measurements (Schmieder et al., 2015). Fecal samples were stored individually in microcentrifuge tubes with 70 % ethanol. The seeds found in the samples were separated according to morphology (size, color and shape) and initially classified by morphotype. Seeds were identified with the help of specialists, by comparing the morphotypes with reference collections at the Museum and Herbarium of Pontificia Universidad Javeriana and with specialized literature (Lobova et al., 2009). To characterize the resource availability, vegetation sampling was conducted using five 50 m x 2 m transects (Dinerstein, 1986) located in the vicinity of the cave. The farthest transect was located at 4 km away from the cave and the closest was at 500 m. In addition, once the seeds retrieved from fecal samples were identified, phenological observations were made during each field trip in all established transects. Every day during the field trips the identified species of plants were examined in order to estimate the relative proportion of flowers and fruits during each sampling month. To check the reproductive status, the species of plants were characterized into four categories: 1) inactive (no flowers or fruits), 2) flowering (more flowers than fruits), 3) intermediate (same amount of flowers and fruits), and 4) fructification (more fruits than flowers) (Mello et al., 2004a, 2004b).

Alviz and Pérez–Torres

CepitaM

Aratoca

av i

Las Vueltas

Bolivar

Mogotes

Cesar Norte de Santander

San Gil Antioquia

20 km

Boyacá

Fig. 1. Study area location in Vereda Las Vueltas, Municipality of Curití, Santander (Colombia). General information on the location of the country, the department and the village where the sampling was carried out is included. Fig. 1. Ubicación de la zona de estudio en Vereda Las Vueltas, en el municipio de Curití, Santander (Colombia). Información general sobre la ubicación del país, el departamento y la aldea en los que se llevó a cabo el muestreo.

We analyzed the species composition and abundance of each dietary item consumed by both females and males. The number of samples in which a food item was found was taken into account in relation to the total number of samples analyzed. Diet breadth was evaluated using Levin's standardized index (B) following the assumptions proposed by Krebs (1998), where B = Y2 / (total number of sampled individuals), R represents the total resources; s, total species; and Nj, number of individuals found using the j resource. Diet overlap was estimated using a simplified Morisita index (CH) (Krebs, 1998), where CH = / ; , , represents the proportion of the resource i from the total resources used by males and females, j y k (i = 1, 2, 3…., n). A confidence level of 95 % was taken into account in all analyses. Temporal dietary behavior was analyzed using a time series analysis (Box et al., 1994; Ziebarth et al., 2010), defined as a succession of observations of a variable taken in several instants of time. This test was used to model the mechanism that gives rise to the series observed through time. To run the test, we took time (independent) versus items consumed by males and females (dependent), and several variables of the time series analysis such as

the media, medium, trend, coefficients of symmetry and kurtosis. The normality of the test was analyzed usingthe Jarque–Bera estimate. Finally, consumption ratios between males and females were compared with an odds ratio test and we assessed temporal covariation between sexes using a Spearman’s rank correlation test (Zar, 2010). Results and discussion During the sampling period, 236 males (56.46 %) and 182 females (43.54 %) were captured. In total, 160 fecal samples were collected from males and 135 from females. From these samples, 18 seed morphospecies were found in the fecal samples of males, and 16 seed morphospecies were found in fecal samples of females (table 1). Ficus, Vismia and Acacia were the most common plant genera, including more than 50 % of the diet for both sexes. Males most frequently consumed Ficus gigantosyce (29.82 %), followed by Vismia glaziovii (16.67 %) and Acacia farnesiana (14.04 %). Females consumed high proportions of Vismia glaziovii (23.40 %) followed by

Animal Biodiversity and Conservation 43.1 (2020)

Table 1. Seed morphotypes found in the diet of males and females of Carollia perspicillata between June 2012 and July 2013 at the Macaregua cave. The abbreviation of the biological collection, the encounter frequency in the faeces (n), and the proportion of each item (%) within the total sample is included.. Tabla 1. Morfotipos de semilla encontrados en la dieta de los machos y las hembras de Carollia perspicillata entre junio de 2012 y junio de 2013 en la cueva de Macaregua. Se incluyen la abreviación de la colección biológica, la frecuencia de encuentro en las heces (n) y la proporción de cada elemento (%) en la muestra total. Morphotypes

Males

Females

Abbreviation

(%)

Ficus gigantosyce

29.82

17.02

Vismia glaziovii

19 16.67

22 23.40

Acacia farnesiana

16 14.04

18 19.15

Piper aduncum

9 7.89

4 4.26

Myrcia popayanensis

8 7.02

4 4.26

Myrcia sp. 1

6.14

9.57

Ficus sp. 1

3.51

3.19

Acacia sp. 1

2.63

5.32

Undetermined 1

Indv.1

1.75

3.19

Undetermined 2

Indv.2

1.75

1.06

Solanum mauritianum

2 1.75

1 1.06

Piper nigrum

2 1.75

0 0

Piper sp. 1

Vismia sp. 2

0.88

Indv.3

0.88

1.06

Vismia sp. 3

0.88

1.06

Asteraceae

Ast.

1 0.88

0 0

Piper sp. 2

0.88

Acacia sp. 2

1.06

Solanum sp. 1

1.06

Indv.4

4.26

Undetermined 3

Undetermined 4

Acacia farnesiana (19.15 %) and Ficus gigantosyce (17.02 %). The remaining morphotypes showed similar frequencies. In addition, males were more likely to consume more Ficus, Piper and Myrcia, while females had higher probabilities of consuming Vismia, Acacia and Myrcia. The remaining identified morphotypes had similar probabilities of being consumed by both sexes (fig. 2). Structure and composition indexes were calculated to compare the assembly of plants species (table 2). Overall, males and females had similar dietary compositions. There was, however, variation across the sampling months. The major differences occurred during the wet season when rainfall and fructification were at their highest. During the wet season (June and August), dietary richness between males and females differed significantly (June: α males = 0.79, α

females = 14.12, n = 26, d.f. = 25, p = 0.02; August: α males = 8.86, α females = 0, n = 54, d.f. = 53, p = 0.03). Males consumed Ficus gigantosyce more frequently, while females consumed more Vismia glaziovii. Females ate a wide variety of items in June (14 spp.) but consumed only a few items in August (2 spp.) despite the high resource availability. This finding may indicate a level of resource selection by females. In contrast, males consumed one food item in June, but expanded their diet considerably in August (1 to 4). No significant difference in dietary richness was observed in the dry season (December) (α males= 2.63, 1.78 ± 2.63; α females = 2.39, 1.05 ± 2.39; n = 53, d.f. = 52, p = 0.3). However, males presented a wider dietary breadth than females during the other sampling months. Females had a wider diet breadth during April and October (wet season). The diets of

Alviz and Pérez–Torres

2.5

Males

Females

Odds ratio

2 1.5 1 0.5 0

P1 M1 Indv.2 S1 M2 F1 V3 Indv.3 A3 V1 Indv.1 M3 A2 Morphotypes

Fig. 2. Odds ratio of the items consumed by male and female Carollia perspicillata at the Macaregua cave. Table 1 shows the morphotypes. The proportion is the number of times an item is present in the total samples per bat species. In this case, females and males are taken as two different species. (For abbreviations of morphotypes, see table 1). Fig. 2. Oportunidad relativa de los elementos consumidos por los machos y las hembras de Carollia perspicillata en la cueva de Macaregua. En la tabla 1 se muestran los morfotipos. La proporción es el número de veces que un elemento está presente en todas las muestras por especie de murciélago. En este caso, las hembras y los machos se consideran como si fueran dos especies diferentes. (Para las abreviaturas de los morfotipos, véase tabla 1).

Table 2. Structure and composition indices calculated for the ensemble of plants consumed by male and female Carollia perspicillata at the Macaregua cave. Tabla 2. Índices de estructura y composición calculados para el conjunto de plantas consumidas por los machos y las hembras de Carollia perspicillata en la cueva de Macaregua.

Sampling months

Indices

Sex

Jun

Aug

Oct

Dec

Feb

Apr

Total

Equitability (J’)

Male

0.00

0.89

0.95

0.68

0.84

0.57

0.78

Female

0.97

1.00

0.92

0.69

0.86

0.55

0.81

0.02 0.22 0.37 0.84 0.78 0.82 0.2

Male

1.00

0.21

0.23

0.52

0.24

0.52

0.15

Female

0.22

0.25

0.39

0.47

0.23

0.62

0.14

0.02 0.66 0.11 0.97 0.92 0.28 0.4

Male

0,00

0.79

0.77

0.48

0.77

0.48

0.85

Female

0.78

0.75

0.61

0.53

0.77

0.38

0.86

0.02 0.62 0.05 0.97 0.92 0.28 0.4

Male

0.80

8.86

2.63

2.47

3.09

2.03

6.48

Female

14.12

0.00

2.39

1.55

3.01

1.59

6.1

0.02 0.03 1.00 0.3 0.99 0.72 0.99

Male

1.00

2.33

Female

4.17

M and F

0.16

Dominance (λ)

Diversity (D)

Richness (α)

Breadth (B) Overlap (C_H)

4.46

5.78

2.5

4.46

7.58

2.34

3.2

2.63

2.52

5.14

8.19

0.91

0.89

0.91

0.46

0.95

0.92

Animal Biodiversity and Conservation 43.1 (2020)

Males Females

Lineal (males) Lineal (females)

Items

Jun

Aug

Oct Dec Sampling month

Feb

Apr

Fig. 3. Temporal series of the items consumed by male and female Carollia perspicillata in each sampling month at the Macaregua cave. The series shows important peaks of consumption of items in October and April for males, and a peak in April for females. Fig. 3. Serie cronológica de los elementos consumidos por los machos y las hembras de Carollia perspicillata en cada mes del muestreo en la cueva de Macaregua. En la serie se muestran picos importantes de consumo de elementos en octubre y abril para los machos y un pico en abril para las hembras.

males and females showed a high degree of similarity for much of the year (CH: Aug = 0.91, Oct = 0.89, Dec = 0.91, Apr = 0.95), with the exception of June (wet season; CH = 0.16) and February (dry season; CH = 0.46), when they exhibited a differential intake of resources (fig. 3). In general, the plant communities that constitute the diets of male and female Seba's short–tailed fruit bat were highly rich and diverse because there was no clear dominance among the most frequently consumed species and their distributions were highly equitable. Even so, the diet of the males was richer, with a slightly more pronounced dominance of some species, contrasting with the diet of females that was more equitable and more diverse. This indicates that the males consume more items, but that the diet of the females is more varied. This variation may be a response to high energy demands by females during reproductive stages, especially during pregnancy and lactation (Angell et al., 2013; Pfeiffer and Mayer, 2013). Unlike males, females must invest more time foraging to obtain food with a different nutritional composition to satisfy energy demands and pre– and post–natal growth (Angell et al., 2013). For this reason, variation can be related to the nutritional quality of plants (Barclay and Jacobs, 2011). Females exhibited a high consumption of fruits of Ficus during August and October, when lactation occurs. The fruits of Ficus contain low levels of nitrogen, which is an essential component during lactation (Herbst, 1986; Shanahan et al., 2001). However, females can meet their nitrogen demands by consuming fruits of Acacia (Hackett et al., 2013).

Fruits of Vismia and Myrcia also contain some calcium content (Charles–Dominique, 1991; López and Vaughan, 2007), so calcium–rich Ficus fruits can be supplementary (Shanahan et al., 2001; Angell et al., 2013). Barclay and Jacobs (2011) reported that calcium is a limiting resource for pregnant bats because their diets are relatively low in calcium. Bat may thus supplement their diet with Ficus fruits due to their high calcium content (Shanahan et al., 2001). This is consistent with other studies that have evaluated the nutritional content of the fruits of Ficus because of their high frequency in the diet of some frugivorous bats (Herbst, 1986; Charles–Dominique, 1991; Korine et al., 2000; Angell et al., 2013; Saldaña–Vásquez et al., 2013). Ficus fruits may be key resources for replenishing energy requirements of females, which may explain their high consumption during reproductive season. Males have a high richness in consumed items so are not as limited as females because their energetic costs are less than those of pregnant and lactating females (Cryan and Wolf, 2003; Encarnação and Dietz, 2006; Almenar et al., 2011). In other geographical regions, males have richer diets than females (Cryan and Wolf, 2003; Angell et al., 2013). However, males of C. perspicillata at the Macaregua cave appeared to have additional energetic demands due to the harem maintenance (McCracken and Wilkinson, 2000; Ortega et al., 2008). If energetic costs increase by maintaining harems, males may need to increase consumption of items and adapt their diet to meet their nutritional demands. In addition, males showed increased consumption of items between August and October, which coincided with reproductive inactivity (inguinal testicles),

and again between February and April when males were reproductively active (scrotal testicles). In general, there was a temporary variation in the diet components between males and females, taking into account the differences that have been reported in terms of energy cost related to reproductive times and resources availability. Unlike what was predicted, there was an increase in the consumption of food items among males during reproductive times, and higher dietary differences with respect to females, possibly due to the fact that they are social males and harem defenders. On the other hand, females were expected to exhibit higher consumption rates of food items during breastfeeding and during periods of higher availability. This means that the consumption that is being presented in the cave differs in time and space due to the additional variations in foraging strategies and social behavior. Knowing the variations that occur in the diet of males and females allows us to more specifically address the differences in the foraging strategies applied regarding movement patterns, differential habitat use, nutritional requirements and energy demands. Diet knowledge is essential to understand all the processes that organisms must carry out for their survival and maintenance. This information is extremely important in order to develop conservation plans for threatened species and ecosystems based on the potential in the resources availability and how organisms adapt and respond to environmental disturbances. Conclusion The diet of Carollia perspicillata was composed of 26 morphotypes, the most representative items being species of the genus Ficus, Vismia, Acacia and Myrcia. This is the first record of Myrcia popayanensis in the diet of Carollia perspicillata. The bats' diet was rich and diverse throughout the year, suggesting that Carollia perspicillata is a generalist species. Males exhibited high consumption of Ficus gigantosyce, while females consumed more Vismia glaziovii fruits. The consumption of resources differed between sexes, as reflected in both foraging strategies and reproductive times of the year. Acknowledgements The authors are grateful to the Laboratorio de Ecología Funcional (LEF–PUJ) for logistic and financial support. We thank Flor Daza and Elías Gómez, owners of Finca La Palma, for their hospitality. This work is part of the project 'Ecología de murciélagos de sistemas cavernícolas del Departamento de Santander' supported by Pontificia Universidad Javeriana (ID 5696). References Almenar, D., Aihartza, J., Goiti, U., Salsamendi, E. G. I., 2011. Reproductive and age classes do not

Alviz and Pérez–Torres

change spatial dynamics of foraging long–fingered bats (Myotis capaccinii). European Journal of Wildlife Research, 57: 929–937. Angell, R. L., Butlin, R. K., Altringham, J. D., 2013. Sexual segregation and flexible mating patterns in temperate bats. Plos One, 8: 1–7. Anthony, E., 1996. Age determination in bats. In: Ecological and behavioral methods for the study of bats: 47–58 (T. Kunz, Eds.). Smithsonian Institution Press. Washington, USA. Barboza–Márquez, K., Aguirre, L. F., 2010. Patrones reproductivos del murciélago frugívoro de cola corta (Carollia perspicillata) relacionados con la fenología de Piper en un bosque montano de Bolivia. Revista Boliviana de Ecología y Conservación Ambiental, 27: 43–52. Barclay, R. M. R., Jacobs, D. S., 2011. Differences in the foraging behaviour of male and female Egyptian fruit bats (Rousettus aegyptiacus). Canadian Journal of Zoology, 89: 466–473. Box, G. E. P., Jenkins, G. M., Reinsel, G. C., 1994. Time Series Analysis: Forecasting and Control. Prentice Hall, Englewood Cliffs, NJ. Burnett, C. D., Kunz, T. H., 1982. Growth Rates and Age Estimation in Eptesicus fuscus and Comparison with Myotis lucifugus. Journal of Mammalogy, 63(1): 33–41. Charles–Dominique, P., 1991. Feeding strategy and activity budget of the frugivorous bat Carollia perspicillata (Chiroptera: Phyllostomidae) in French Guiana. Journal of Tropical Ecology, 7: 243–256. Cloutier, D., Thomas, D. W., 1992. Carollia perspicillata. Mammalian Species, 417: 1–9. Cryan, P. M., Wolf, B. O., 2003. Sex differences in the thermoregulation and evaporative water loss of a heterothermic bat, Lasiurus cinereus, during its spring migration. The Journal of Experimental Biology, 206: 3381–3390. Dietz, M., Kalko, E. K. V., 2006 Seasonal changes in daily torpor patterns of free–ranging female and male Daubenton's bats (Myotis daubentonii). Journal of Comparative Physiology, 176: 223–231. Dinerstein, E., 1986. Reproductive ecology of fruit bats and the seasonality of fruit production in a costa rican cloud forest. Biotropica, 18: 307–318. Encarnação, J. A., Dietz, M., 2006. Estimation of food intake and ingested energy in Daubenton’s bats (Myotis daubentonii) during pregnancy and spermatogenesis. European Journal of Wildlife Research, 52: 221–227. Fleming, T. H., 1988. The Short–Tailed Fruit Bat: A Study in Plant–Animal Interactions. The University of Chicago Press, United States. Fleming, T. H., Heithaus, E. R., 1986. Seasonal Foraging Behavior of the Frugivorous Bat Carollia perspicillata. Journal of Mammalogy, 67: 660–671. Galindo–González, J., 1998. Dispersión de semillas por murciélagos: su importancia en la conservación y regeneración del bosque tropical. Acta Zoológica Mexicana, 73: 57–74. GELT (Gobierno en línea del orden territorial), 2013. Sitio oficial del Municipio de Curití en Santander, Colombia. Online: http://curiti–santander.gov.co/

Animal Biodiversity and Conservation 43.1 (2020)

Hackett, T. D., Korine, C., Holderied, M. W., 2013. The Importance of Acacia Trees for Insectivorous Bats and Arthropods in the Arava Desert. Plos One, 8(2): e52999, https://doi.org/10.1371/journal. pone.0052999 Heer, K., Albrecht, L., Kalko, E. K. V., 2010. Effects of ingestion by neotropical bats on germination parameters of native free–standing and strangler Figs (Ficus sp., Moraceae). Oecologia, 163: 425–435. Herbst, L. H., 1986. The Role of Nitrogen from Fruit Pulp in the Nutrition of the Frugivorous Bat Carollia perspicillata. Biotropica, 18(1): 39–44. Korine, C., Kalko, E. K. V., Herre, E. A., 2000. Fruit characteristics and factors affecting fruit removal in a Panamanian community of strangler figs. Oecologia, 123: 560–568. Krebs, C. J., 1998. Ecological Methodology, 2nd Ed. Benjamin Cummings, Menlo Park, California. Kunz, T. H., Anthony, E. L. P., 1982. Age Estimation and Post–Natal Growth in the Bat Myotis lucifugus. Journal of Mammalogy, 63(1): 23–32. Kunz, T. H., Braun de Torres, E., Bauer, D., Lobova, T., Fleming, T. H., 2011. Ecosystem services provided by bats. The Year in Ecology and Conservation Biology, 1223(2011): 1–38. Lobova, T., Geiselman, C. K., Mori, S. A., 2009. Seed dispersal by bats in the Neotropics. The New York Botanical Garden, Bronx, New York, United States. Lopez, J. E., Vaughan, C., 2007. Food niche overlap among neotropical frugivorous bats in Costa Rica. Revista de Biologia Tropical, 55(1): 301–313. https://doi.org/10.15517/rbt.v55i1.6082 Marinho–Filho, J. S., 1991. The coexistence of two frugivorous bat species and the phenology of their food plants in Brazil. Journal of Tropical Ecology, 7(1): 59–67. McCracken, G. F., Wilkinson, G. S., 2000. Bat Mating Systems. In: Reproductive Biology of Bats: 321–357 (P. H. Krutzsch, Ed.). Academic Press, A Harcourt Science and Technology Company. San Diego, United States. Mello, M. A., Schittini, G. M., Selig, P., Bergallo, H. G., 2004a. A test of the effects of climate and fruiting

of Piper species (Piperaceae) on reproductive patterns of the bat Carollia perspicillata (Phyllostomidae). Acta Chiropterologica, 6: 309–318. – 2004b. Seasonal variation in the diet of the bat Carollia perspicillata (Chiroptera: Phyllostomidae) in an Atlantic Forest area in southeastern Brazil. Mammalia, 68: 49–55. Pérez–Torres, J., Martínez–Medina, D., Ríos–Blanco, C., Peñuela–Salgado, M., Brito–Hoyos D., Martínez–Luque, L., 2015. Macaregua: the cave with the highest bat richness in Colombia. Check List, 11(2): 1616. Ortega, J., Guerrero, J. A., Maldonado, J. E., 2008. Aggression and tolerance by dominant males of Artibeus jamaicensis: strategies to maximize fitness in harem groups. Journal of Mammalogy, 89(6): 1372–1378. Pfeiffer, B., Mayer, F., 2013. Spermatogenesis, sperm storage and reproductive timing in bats. Journal of Zoology, 289: 77–85. Saldaña–Vásquez, R. A., Sosa, V. J., Iñiguez–Dávalos, L. I., Schondube, J. E., 2013. The role of extrinsic and intrinsic factors in Neotropical fruit bat–plant interactions. Journal of Mammalogy, 94(3) : 632–639. Sánchez, M. S., Giannini, N. P., Barquez, R. M., 2012. Bat frugivory in two subtropical rain forests of Northern Argentina: Testing hypotheses of fruit selection in the Neotropics. Mammalian Biology, 77: 22–31. Schmieder, D. A., Benítez, H. A., Borissov, I. M., Fruciano, C., 2015. Bat Species Comparisons Based on External Morphology: A Test of Traditional versus Geometric Morphometric Approaches. Plos One, 10(5): e0127043. Doi: 10.1371/journal. pone.0127043. Shanahan, M., Samson, S., Compton, S. G., Corlett, R., 2001. Fig–eating by vertebrate frugivores: a global review. Biological Reviews, 76: 529–572. Zar, J., 2010. Biostatistical Analysis. Prentice Hall, Inc. Ney Jersey, United States. Ziebarth, N. L., Abbott, K. C., Ives, A. R., 2010. Weak population regulation in ecological time series. Ecology Letters, 13: 21–31.

Alviz and Pérez–Torres

Brief communication 37

Animal Biodiversity and Conservation 43.1 (2020)

European free–tailed bat fatalities at wind farms in southern Spain A. R. Muñoz, M. Á. Farfán

Muñoz, A. R., Farfán, M. Á., 2020. European free–tailed bat fatalities at wind farms in southern Spain. Animal Biodiversity and Conservation, 43.1: 37–41, Doi: https://doi.org/10.32800/abc.2020.43.0037 Abstract European free–tailed bat fatalities at wind farms in southern Spain. Wind is increasingly used as a renewable energy all around the world. Although wind turbines help reduce greenhouse gas emissions, the costs to wildlife cannot be overlooked. To date, monitoring programs and research have mainly focused on the impact of wind farms on birds but negative effects on bats have also reported. Here we compile information related to European free–tailed bat deaths at wind farms in southern Spain. In a world where the demand for renewable energy is rising we highlight the need to better understand and prevent bat fatalities. Key words: Bat collision, Monitoring program, Tadarida teniotis, Wind farm Resumen Mortalidad del murciélago rabudo en los parques eólicos del sur de España. El viento se está utilizando cada vez más como fuente de energía renovable en todo el mundo. Aunque las turbinas eólicas ayudan a reducir la emisión de gases de efecto invernadero, no se pueden pasar por alto los costos que conllevan para la fauna salvaje. Hasta la fecha, los programas de seguimiento e investigación se han centrado principalmente en el impacto de los parques eólicos en las aves, aunque se ha constatado que también afectan a los murciélagos. En este trabajo se compila información sobre la mortalidad del murciélago rabudo en los parques eólicos del sur de España y destacamos la necesidad de conocer mejor y evitar la mortalidad de murciélagos en un mundo donde la demanda de energías renovables no para de aumentar. Palabras clave: Colisión de murciélagos, Programa de seguimiento, Tadarida teniotis, Parque eólico Received: 16 IV 18; Conditional acceptance: 09 VII 18; Final acceptance: 05 IX 19 Antonio–Román Muñoz, Miguel Ángel Farfán, Biogeography, Diversity, and Conservation Research Team, Depto. de Biología Animal, Facultad de Ciencias, Universidad de Malaga, 29071 Malaga, España (Spain).– Miguel Ángel Farfán, Biogea Consultores, c/ Navarro Ledesma 243 portal 4, 3º C, 29010 Málaga, España (Spain). Corresponding author: Antonio–Román Muñoz. E–mail: roman@uma.es ORCID–ID: A.–R. Muñoz: 0000–0002–0253–7632; M. Á. Farfán: 0000–0002–4617–6517

ISSN: 1578–665 X eISSN: 2014–928 X

Muñoz and Farfán

Introduction Human activities cause environmental modifications such as fragmentation, destruction and degradation of habitats, exposing many animal populations to novel perturbations and consequent declines worldwide (Scheffer et al., 2001). Reducing greenhouse gases emission to prevent anthropogenic climate change has enhanced the innovation, development and application of renewable energy sources, but unfortunately, renewable energies, and in particular wind power, may come at some risk for wildlife. Wind turbines, for example, can cause large numbers of fatalities among flying animals (Ferrer et al., 2012; Zimmerling and Francis, 2016). Most research to date has focused on the impact of wind farms on birds, especially those of conservation concern and particularly raptors or other large species, but information related to bats is increasing and giving rise to serious concerns (e.g. Lintott et al., 2016). Here we compile the information available regarding fatalities of the European free–tailed bat Tadarida teniotis at wind farms in southern Spain. We report new distributional data for the species, discuss its probable underestimated distribution area, and point out some important aspects that should be taken into account to facilitate compatibility between the development of renewable energies and bat conservation. Material and methods The European free–tailed bat is the only representative of the family Molossidae in Europe. It is mostly a Palaeartic species whose range extends into the Indomalayan region (Benda and Piraccini, 2016). It is well known in the Mediterranean basin, in Portugal, Spain, Morocco and Algeria, and eastwards to the Middle East, Saudi Arabia, Iran, Iraq, Azerbaijan, Turkmenistan, Afghanistan, India, and possibly Yunnan (China). It is widely–dispersed throughout Spain, including the Canary and Balearic archipelagos, but little information is available at the population level (Balmorí, 2007). A brief glance at the most recently available distribution map, in the Atlas of Terrestrial Spanish Mammals and Red Book (Palomo et al., 2007), shows that existing information is limited (fig. 1). Furthermore, the distribution data are biased towards some areas that seem to be surveyed systematically, such as the Canary Islands, Extremadura and Navarra, while information seems to be partial in areas such as Andalusia. The study was performed from 2006 and 2016. We focused our research on southern Spain, particularly the provinces of Cadiz and Malaga (fig. 1). Most information comes from the Andalusia Environmental Agency, which provided all the data on bat fatalities recorded for Cadiz at 56 wind farms containing a total of 801 wind turbines. The information from Malaga is derived from two wind farms, one with 13 wind turbines and the other with four turbines. Monitoring at the former was performed regularly over two years but irregularly at the latter. In Cadiz, the Environ-

mental Agency requires all wind farms to develop a surveillance program to document all accidents caused to flying animals. This program operates on a daily basis from dawn to dusk (between eight and 14 daylight hours in winter and summer, respectively) and is carried out by trained observers who work in a coordinated manner. In Malaga, searches for animal carcasses were made on a weekly basis at one of the studied wind farms, as required by the same Environmental Agency, and six times a year at the second wind farm. Results During the study period, a total of 15 European free–tailed bats were found dead at 11 wind farms, all having collided with wind turbines. At two wind farms, one in Cadiz and other in Malaga, three fatalities were found (each site indicated by a star in fig. 1). All records collected for this study are new distributional data for the species, considering the specific information provided in the Atlas of Terrestrial Spanish Mammals and Red Book (Balmorí, 2007) (represented by grey squares in fig. 1). The new records correspond to the following 10 x 10 UTM squares: UF38, TF92 y TF93 in Malaga province, and TE69, TF60, TF53, TF44, TF40, TF35, QA64 and QA55 in Cadiz province. Fatalities were not found every year. Pooling all data, we found that mortality was highest in 2012 and 2014 (fig. 2). During the study period, mortality peaked in October, although cases were reported during other months, including January (fig. 2). Discussion The European free–tailed bat is a relatively unknown species, and aspects such as population size and movements in the Iberian Peninsula are open questions. The fact that all fatalities in this study correspond to new distribution records for the species is a clear indicator of the lack of knowledge related to this bat. In Spain, only some wind farms are required to develop surveillance programs and these focus mainly on bird mortality (De Lucas et al., 2012). The effects of wind power on bats seem to take second place for the environmental authorities, even though consequences can be serious, as demonstrated in other regions of the world (e.g. Lehnert et al., 2014). The absence of a standardised method to monitor bat fatalities and the current limitations of the existing monitoring protocols (such as the lack of training of observers in bat identification) suggest that the impact of wind farms on the European free–tailed bat is likely greater than that reflected in our figures. The fatalities reported here occurred practically throughout the year with a clear peak in October. Although Arlettaz et al. (2000) demonstrated that the European free–tailed bat can be considered a hibernating species in the Swiss Alps —the highest

Animal Biodiversity and Conservation 43.1 (2020)

1:8.000.000

Spain

1:1.125.000

Fig. 1. A, species distribution for the UTM 10 x 10 km squares in Spain (dark squares indicate presences). The study area is indicated in solid lines (provinces of Cadiz and Malaga); B, study area and locations where fatalities were found (black dots); stars indicate the two wind farms that each had three fatalities. UTM 10 x 10 km squares are shown; those in grey are squares where the species was previously reported according to the Atlas of Terrestrial Spanish Mammals and Red Book. Fig. 1. A, distribución de la especie en las cuadrículas UTM 10 x 10 km en España (la presencia se indica con cuadrículas oscuras). La zona del estudio se indica con líneas continuas de color negro (provincias de Cádiz y Málaga); B, zona del estudio y localización de los ejemplares fallecidos encontrados (puntos negros); las estrellas indican la ubicación de los dos parques eólicos en los que se produjeron tres fallecimientos. Se muestran las cuadrículas UTM 10 x 10 km; en gris se indican las cuadrículas donde ya se tenía constancia de la presencia del murciélago rabudo según el Atlas y Libro Rojo de los Mamíferos Terrestres de España.

northern limit for this species— it can be active all year round in the south of the Iberian Peninsula (Marques et al., 2004). This peak in October could be related to flocking and courtship behaviour during autumn, but the proximity of the study area to the African continent, 14 km. at the shortest distance, could also play a role. The species is found on both sides of the Strait of Gibraltar and its capacity for movement is well known, with single bouts of activity averaging more than 6 hours per night (Marques et al., 2004). Being a fast flying species with low manoeuvrability compared to others bat species, the effects of turbines could have far reaching consequences. As with other flying animals (Schuster et al., 2015), it is likely that the consequences of wind farms are not the only factor responsible for the death of the

bats. However, disturbance, habitat loss, and barrier effects may also play a role. In the absence of population estimates for the European free–tailed bat in most of its distribution range, greater concern among environmental authorities is called for. Deaths of other species of the same family, particularly those in tropical regions, are commonly reported, such as for the Egyptian free–tailed Bat Tadarida aegyptiaca in South Africa (Doty and Martin, 2013), and the Brazilian free–tailed Bat Tadarida brasiliensis in South America (Barros et al., 2015). The impact on this latter species is severe; 245 of the 336 bat fatalities reported from Brazil corresponded to Brazilian free–tailed bats. Our results may be biased. Nevertheless, as information about the impact of wind farms on bats in southern Spain is limited, in light of the high number

Muñoz and Farfán

Number of bat fatalities

5 4 3 2 1 0

2006 2007 2008 2009 2010 2011 2012 2013 2014 2015 2016 Year

Number of bat fatalities

8 7 6 5 4 3 2 1 0

Jan Feb Mar Apr My

Jun Jul Aug Sep Oct Nov Dec Month

Fig. 2. A, distribution of European free–tailed bats found dead in wind farms per year; B, monthly distribution of dead European free–tailed bats, grouping data from 2006 to 2016. Fig. 2. A, distribución anual de los ejemplares de murciélago rabudo fallecidos en parques eólicos; B, distribución mensual de ejemplares de murciélago rabudo fallecidos; agrupando la información del período 2006–2016.

of wind power facilities established over the last two decades, greater knowledge and understanding of bat fatalities is necessary (see Rodrigues et al., 2015). This is especially important if we consider that the demand for this type of renewable energy has been increasing rapidly and will likely continue to do so. Acknowledgements ARM was supported by the Juan de la Cierva Program (FJCI–2014–20653) (Ministerio de Ciencia e Innovación, Spain). The study was partially funded by the Spanish Ministry of Agriculture, Food and Environment, and the National Park Network of Spain (Project 1098/2014). References Arlettaz, R., Ruchet, C., Aeschimann, J., Brun, E., Genoud, M., Vogel, P., 2000. Physiological traits

affecting the distribution and wintering strategy of the bat Tadarida teniotis. Ecology, 81: 1004–1014. Balmorí, A., 2007. Tadarida teniotis Rafinesque, 1814. In: Atlas y Libro Rojo de los Mamíferos Terrestres de España: 230–233 (L. J. Palomo, J. Gisbert, J. C. Blanco, Eds.). Dirección General para la Biodiversidad–SECEM–SECEMU, Madrid. Barros, M. A. S., De Magalhães, R. G., Rui, A. M., 2015. Species composition and mortality of bats at the Osório Wind Farm, southern Brazil. Studies on Neotropical Fauna and Environment, 50: 31–39. Benda, P., Piraccini, R., 2016. Tadarida teniotis. The IUCN Red List of Threatened Species 2016: e.T21311A22114995, http://dx.doi.org/10.2305/ IUCN.UK.2016–2.RLTS.T21311A22114995.en [Downloaded on 03 December 2019]. De Lucas, M., Ferrer, M., Bechard M. J., Muñoz A. R., 2012. Griffon vulture mortality at wind farms in southern Spain: distribution of fatalities and active mitigation measures. Biological Conservation, 147: 184–189.

Animal Biodiversity and Conservation 43.1 (2020)

Doty, A. C., Martin, A. P., 2013. Assessment of bat and avian mortality at a pilot wind turbine at Coega, Port Elizabeth, Eastern Cape, South Africa. New Zealand Journal of Zoology, 40: 75–80. Ferrer, M., De Lucas, M., Janss, G. F. E., Casado, E., Muñoz, A. R., Bechard, M. J., Calabuig, C. P., 2012. Weak relationship between risk assessment studies and recorded mortality in wind farms. Journal of Applied Ecology, 49: 38–46. Lehnert, L. S., Kramer–Schadt, S., Schönborn, S., Lindecke, O., Niermann, I., Voigt C. C., 2014. Wind farm facilities in Germany kill noctule bats from near and far. Plos One, 9(8): e103106, https://doi. org/10.1371/journal.pone.0103106 Lintott, P. R., Richardson, S. M., Hosken, D. J., Fensome, S. A., Mathews, F., 2016. Ecological impact assessments fail to reduce risk of bat casualties at wind farms. Current Biology, 26: 1135–1136. Marques, J. T., Rainho, A., Carapuço, M., Oliveira, P., Palmeirim, J. M., 2004. Foraging behaviour and habitat use by the European free–tailed bat Tadarida teniotis. Acta Chiropterologica, 6: 99–110.

Palomo, L. J., Gisbert, J., Blanco, J. C. (Eds.), 2007. Atlas y Libro Rojo de los Mamíferos Terrestres de España. Dirección General para la Biodiversidad– SECEM–SECEMU, Madrid. Rodrigues, L., Bach, L., Dubourg–Savage, M.J., Karapandza, B., Kovac, D., Kervyn, T., Dekker, J., Kepel, A., Bach, P., Collins, J., Harbusch, C., Park, K., Micevski, B., Miderman, J., 2015. Guidelines for consideration of bats in wind farm projects. Revision 2014. EUROBATS Publication Series N°6 (English version). UNEP/EUROBATS Secretariat, Bonn, Germany. Schuster, E., Bulling, L., Köppel, J., 2015. Consolidating the state of knowledge: a synoptical review of wind energy's wildlife effects. Environmental Management, 56: 300–331. Scheffer, M., Carpenter, S., Foley, J., Folke, A. C., Walker, B., 2001. Catastrophic shifts in ecosystems. Nature, 413: 591–596. Zimmerling, J. R., Francis, C. M., 2016. Bat mortality due to wind turbines in Canada. The Journal of Wildlife Management, 80: 1360–1369.

Muñoz and Farfán

Animal Biodiversity and Conservation 43.1 (2020)

Taxonomic nestedness based on guilds? Bird assemblages of the Jardines de la Reina National Park, Cuba, as study case A. García–Quintas, D. Fundora Caballero, A. Parada Isada García–Quintas, A., Fundora Caballero, D., Parada Isada, A., 2020. Taxonomic nestedness based on guilds? Bird assemblages of the Jardines de la Reina National Park, Cuba, as study case. Animal Biodivesity and Conservation, 43.1: 43–54, Doi: https://doi.org/10.32800/abc.2020.43.0043 Abstract Taxonomic nestedness based on guilds? Bird assemblages of the Jardines de la Reina National Park, Cuba, as study case. Nestedness is a widely known structuring model in insular and fragmented biotas that has often been assessed, but most studies to date have used a taxonomic approach. However, the relevance of an approach using functional groups has become increasingly highlighted in community ecology research. In this study, we evaluated the occurrence of nested structure in the Jardines de la Reina National Park bird assemblages as a whole, and its trophic guilds by following three different grouping criteria. We constructed species presence–absence matrices for each guild and estimated the degree of nestedness with the metric based on the overlap and decreasing fill, assessing its significance by means of two null models. Overall bird assemblage was significantly nested (NODF = 76.99; p = 0.01) whereas terrestrial insectivores (NODF = 81.32) and insectivores (NODF = 80.04) were the only trophic guilds (out of 19) that showed significant nestedness (p ≤ 0.01). These results could provide evidence of the structural and functional cohesion of avifauna at the study site, especially among its insect–eating taxa. Taxonomic nestedness based on a guilds approach may help identify suitable conservation strategies for avian communities inhabiting naturally fragmented areas such as the Jardines de la Reina National Park. Key words: Functional grouping, Low–lying islands, Nested community, Null model, Protected area, Trophic guild Resumen ¿Anidamiento taxonómico basado en gremios? Los ensamblajes de aves del Parque Nacional Jardines de la Reina, en Cuba, como caso de estudio. El anidamiento es un modelo de estructuración bien conocido en biotas insulares y fragmentadas que se ha estudiado a menudo, aunque en la mayoría de los estudios realizados hasta la fecha se ha hecho desde un enfoque taxonómico. No obstante, en los estudios sobre ecología de comunidades se resalta cada vez más la importancia de adoptar un enfoque que utilice grupos funcionales. En este estudio se evaluó el grado de estructura de anidamiento en los ensamblajes de aves del Parque Nacional Jardines de la Reina en general y en sus gremios tróficos, siguiendo tres criterios de clasificación. Se confeccionaron matrices de presencia-ausencia de especies para cada gremio, se calculó el grado de anidamiento a partir del relleno superpuesto y decreciente y se analizó su significación mediante dos modelos nulos. La comunidad general de aves estuvo significativamente anidada (NODF = 76,99; p = 0,01) mientras que los insectívoros terrestres (NODF = 81,32) y los insectívoros (NODF = 80,04) fueron los únicos gremios tróficos que presentaron anidamiento significativo (p ≤ 0,01). Estos resultados podrían poner de manifiesto la cohesión estructural y funcional de la avifauna en la zona del estudio, especialmente en los taxones insectívoros. Así, el enfoque del anidamiento taxonómico basado en gremios puede ayudar a determinar las estrategias de conservación adecuadas para las comunidades de avifauna que habitan en zonas naturalmente fragmentadas, como el Parque Nacional Jardines de la Reina. Palabras clave: Agrupamiento funcional, Islas bajas, Comunidad anidada, Modelo nulo, Zona protegida, Gremio trófico Received: 03 VI 19; Conditional acceptance: 05 IX 19; Final acceptance: 10 IX 19 Antonio García–Quintas, Centro de Investigaciones de Ecosistemas Costeros (CIEC), Cayo Coco, Ciego de Ávila, 69400 Cuba.– Daylon Fundora Caballero, Unidad de Medio Ambiente, Delegación Territorial del CITMA Ciego de Ávila, 69400 Cuba.– Alain Parada Isada, Environmental and Life Sciences Graduate Program, Trent University, 1600 West Bank Drive, Peterborough, ON K9J 7B8 Canada. Corresponding author: Antonio García Quintas. E–mail: agquintas86@gmail.com ISSN: 1578–665 X eISSN: 2014–928 X

Introduction The complexity of natural communities has long proved a fruitful theoretical framework for the inception of numerous hypotheses related to and assembly rules at explaining species coexistence patterns. The most prevalent of these patterns are aggregation, segregation, turnover, and nestedness (Ulrich and Gotelli, 2013). Although these patterns and their underlying mechanisms have been the subject of considerable research efforts, tackling these theoretical principles separately is still a pervasive approach across many such studies. However, different structural patterns of species assemblages may well coexist in any given matrix (Ulrich and Gotelli, 2012). Ulrich et al. (2017) stated that species interactions, distribution of functional traits and stochastic events of colonization and extinction can be readily inferred from the analysis of presene-absence . Among these patterns, biotic nestedness has become a recurring topic of ecological research in the past few decades (e.g., Patterson and Atmar, 1986; Patterson, 1987; Wright et al., 1998; Almeida–Neto et al., 2008; Ulrich et al., 2009). Interestingly, prevalence of this particular phenomenon across natural communities has shifted from commonplace to scarce through continued improvements of statistical toolkits (Matthews et al., 2015). While the definition of nestedness itself has been targeted for extensive criticism and revision (Dormann et al., 2009), its essence has remained nearly unchanged in that 'communities exhibit nested structures if poor species assemblages are non–random subgroups of those with greater species richness' (Patterson, 1987). More recently, Matthews et al. (2015) noted that ecological similarities among species may be overlooked by the traditional taxonomic approach (species composition), which could result in functional redundancy of the ecosystems. This latter subject stresses the need to include functional criteria in nestedness studies by providing a more accurate ecological context. The concept of functional diversity, i.e. the breadth of ecological functions within any species assemblage (Petchey and Gaston, 2006; Bender et al., 2017), has been applied to the nestedness approach (e.g. Matthews et al., 2015; Bender et al., 2017; Aspin et al., 2018; Peláez and Pavanelli, 2018). Thus, functional nestedness was defined as the degree to which the set of functions present in a species–poor site are also present in richer sites, with a greater number of species, revealing a gradient in functional redundancy (Matthews et al., 2015; Bender et al., 2017). The core of this nestedness type is the 'species eco–functional traits' such as physiological, morphological, biochemical and behavioral characteristics of the individuals, related to the functions of the ecosystems (Gómez–Ortiz and Moreno, 2017). Similar to its taxonomic homologous, functional nestedness can be represented in a presence–absence matrix with the difference that the eco–functional traits substitute the species'names (see Bender et al., 2017). To date, this is the most widely accepted approach to assess the functional nestedness within the natural species assemblages.

García–Quintas et al.

Nevertheless, since Bender et al. (2017) and Gómez–Ortiz and Moreno (2017) state that functional diversity may be too quantified through functional groups or guilds, we suggest that another new approach can be used to assess the species functional roles in the nestedness context. Guilds constitute 'groups of species that exploit the same kind of environmental resources in a similar manner' (Simberloff and Dayan, 1991; Heino, 2009), and are defined based on a set of eco–functional traits shared among the species within the guild in question. Therefore, nestedness can be potentially assessed for any given species assemblages by using the concept of guilds/functional groups. A hypothetical presence–absence matrix (species vs. sites) may or may not be taxonomically nested, but different degrees of nestedness between some of its underlying submatrices can still occur. If such submatrices are built based on a given guild classification system, we would be using a similar procedure to that of the functional nestedness within the overall species assemblage (fig. 1). Taxonomic nestedness based on guilds could be an ecologically meaningful approach since it enables researches for several reasons: first, to detect the true nested species subgroups according to its functional role within entire species assemblage; second, to identify important ecological functions (determined by the nested guilds) that contribute to the communities' stability and structural cohesion in fragmented landscapes; and third, to prioritize conservation efforts towards species subgroups relevant in maintaining the assemblages' natural cohesion/organization over individual species. Further studies focused on the underpinnings of functional nestedness could shed light on how to assess the effects of biological conservation threads at varying spatio–temporal scales from long–term population datasets. Trophic guilds are particularly useful as they inform us about trophic web structures and the ecosystems' energy flow. In Cuba, bird assemblages of the Jardines de la Reina archipelago (JRA) show a consistent year–round nested structure throughout the annual cycle, unaltered by migration–driven species turnovers (García–Quintas and Parada, 2014). The Jardines de la Reina National Park (JRNP), the largest marine protected area with the highest number of cays in the Caribbean, is located within this insular region. Species relationships with its habitats and critical food resources are well known to shape birds' distribution patterns at varying spatial scales. Thereby, marine bird assemblages tend to display extensive and homogeneous distributional areas due to the spatial representativeness and interconnectivity of foraging sites across insular regions. On the contrary, trophic guilds whose species rely heavily on terrestrial food resources would naturally show comparatively more patchy and reduced distribution ranges due to differences on food items (availability and quality) among habitats and locations.. Therefore, we would expect that the nestedness degree of bird assemblages in the JRNP increases from the trophic guilds that include species dependent on the marine resources to the guilds composed by species consuming exclusively terrestrial items. Thus,

Animal Biodiversity and Conservation 43.1 (2020)

Species

1 2 3 4 5 6 7 8 9 10 11 12 13 14

Sites A B C D E F G H I J

2 3 5 7 Species guildα 9 10

1 4 6 8 11 12 13 14

A B C D E F G H I J

NODF = 65.04 A B C D E F G H I J

Species guildβ

NODF = 61.99

NODF = 54.52

Fig. 1. A theoretical example of the taxonomic nestedness based on guilds. The example on the left represents a general matrix (species vs. sites) sorted according to the classical taxonomic nestedness approach. On the right we showed two submatrices of the general matrix where its species compositions constitute distinct functional groups (α and β guilds): gray squares, presences; white squares, absences. Differences between NODF (nestedness metric based on the overlap and decreasing fill) values show that several nested statuses, corresponding to distinct guilds, can coexist into either matrix. Fig. 1. Ejemplo teórico del anidamiento taxonómico basado en gremios. A la izquierda está representada una matriz general (especies frente a sitios) ordenada según el enfoque del anidamiento taxonómico clásico. A la derecha se presentan dos submatrices de la matriz general cuyas composiciones de especies constituyen grupos funcionales distintos (gremios α y β): cuadros grises, presencias; cuadros blancos, ausencias. Las diferencias entre los valores de NODF (índice de anidamiento basado en el relleno superpuesto y decreciente) muestran que en cualquier matriz pueden coexistir varios estados anidados correspondientes a gremios distintos.

the aims of this study are: 1) to test out the occurrence of varying degrees of nestedness between different trophic guilds underliyng the overall taxonomic nestedness in the JRNP; 2) to evaluate and compare the nested pattern between trophic guilds based on distinct grouping schemes; and 3) to identify functional groups of birds which significantly contribute to the overall nested structure in the archipelago as priority conservation targets. Material and methods Study area The JRA stretches along Cuba's south–eastern coast from the Ancón Peninsula (Sancti Spíritus province) to Cabo Cruz (Granma province) and it comprises about 661 cays and islets. Among the three main insular sub–regions in this archipelago, the Doce Leguas cays, which encompass the JRNP, are the most extensive. This protected area extends across 87 km south of Ciego de Ávila and Camagüey provinces and includes relatively large cays such as Caguama (7.7 km2), Grande (24.3 km2) and Caballones (33.5 km2) (fig. 2).

This insular region encompasses relatively small and low–lying islands of recent geological history and oceanic origin that have arisen from storm movement of offshore sediments (García–Quintas and Parada, 2017). Sandy shores, dunes and shallow coastal lagoons constitute the most remarkable landscape features in these cays. In the JRNP, there are three main vegetation types, namely, mangrove forests (mostly dominated by Rhizophora mangle), sandy coastal scrubs (typical plant species are Metopium toxiferum, Coccothrinax littoralis, Erithalis fruticosa, Chamaecrista lineata, Salvia bahamensis and Crossopetalum rhacoma), and sandy and rocky vegetation complexes. Mangrove forests are the most widely distributed vegetation type, featuring floristic and physiognomic variants throughout the archipelago (Parada and García–Quintas, 2012). In general, these coastal vegetation complexes harbor low plant species richness and some of their floristic features can be found intermixed with those of adjacent coastal scrubs. Vascular flora of the JRA is represented by 40 families, 97 genera and 113 infra–generic taxa with 4.5 % taxa being endemic to Cuba (Acevedo, 2013). Recent contributions to the study of the avifaunal distribution within the JRA have improved knowledge

García–Quintas et al.

21º30'0'' W

Sancti Spiritus Ciego de Ávila

N S

Camagüey

Gulf of Ana Maria

21º0'0'' W 20º30'0'' N

Caribbean Sea 1 cm = 5 km 0 10 20 40 60

79º30'0'' W

Caribbean Sea

80 km

78º45'0'' W

78º0'0'' W

Jardines de la Reina National Park Contours Cuba Fig. 2. Geographical location of the Jardines de la Reina National Park south of Ciego de Ávila and Camagüey provinces, southern Cuba. Studied cays: 1, Bretón; 2, Cinco Balas; 3, Alcatracito; 4, Alcatraz; 5, Grande; 6, Caballones; 7, Anclitas; 8, Boca Piedra Chiquita; 9, Boca de la Piedra de Piloto; 10, Piedra Grande; 11, Las Cruces; 12, Cachiboca; 13, Largo; 14, Boca Rica; 15, Juan Grin; 16, Boca Seca; 17, Camposanto; 18, Caguama; 19, Cabeza del Este. Fig. 2. Ubicación geográfica del Parque Nacional Jardines de la Reina, al sur de las provincias Ciego de Ávila y Camagüey, en el sur de Cuba. (Para las abreviaturas de los cayos estudiados, véase arriba).

of the JRA biodiversity, a previously relatively poorly studied area compared to other Cuban insular and coastal regions (Parada and García–Quintas, 2012). At present, 92.6 % of the avifauna documented in JRA (121 species, Parada et al., 2015) is represented in JRNP. Three taxa are regarded as Near Threatened: reddish egret (Egretta rufescens), Cuban black hawk (Buteogallus gundlachii) and white crowned pigeon (Patagioenas leucocephala) (Birdlife International, 2018). Data source and processing We obtained the presence–absence data of bird species occurring in 19 of the Doce Leguas cays from earlier works (Parada and García–Quintas, 2012; García–Quintas and Parada, 2017), along with our latest documented occurrences in the study site. General physical attributes of these cays and the sampling effort employed by each one are in García–Quintas and Parada (2017). We created a presence–absence (1–0) matrix whose rows and columns represented species and cays (taxonomic approach), respectively. Data were entered in the matrix in increasing order; starting with most widely distributed species on the top row and the cay with the highest species richness on the left column.

We excluded species with no explicit reference to their locality name when first reported in the study area (e.g., solitary sandpiper Tringa solitaria, chuck–will's–widow Antrostomus carolinensis). Species were grouped into trophic guilds according to three distinct classification schemes: most detailed and comprehensive criteria used for the Cuban avifauna (Kirkconnell et al., 1992), a parsimonious simplification of previous criteria (Andraca, 2010: Variant III) and the simplest and most inclusive criteria (Pizarro et al., 2012) (table 1s in supplementary material). Presence–absence matrices were created limited to the species in a trophic guild, for each guild of each of the three grouping schemes (based on guilds approach), separately, and arranged the same way as the combined matrix. Trophic guilds made up by fewer than five species were excluded from analyses to avoid small matrices. To estimate nestedness, we used the nestedness metric based on the overlap and decreasing fill NODF (Almeida–Neto et al., 2008) featured in NODF 2.0 (Almeida–Neto and Ulrich, 2010). NODF values ranged from 0 (no nestedness) to 100 (perfect nestedness), and were compared with those obtained from 1,000 simulations of two null models. We used null models

Animal Biodiversity and Conservation 43.1 (2020)

Table 1. Trophic guilds structure (guilds ≥ 5 species) of bird assemblages in 19 cays of the Jardines de la Reina National Park, southern Cuba: FGI, foliage–gleaner insectivore; ASI, air–sallier insectivore; PGIF, pecker– gleaner insectivore–frugivore; FGIF, foliage–gleaner insectivore–frugivore; GGG, ground–gleaner granivore; WAC, water–ambusher carnivore; WPP, water–plunger piscivore; SGC, shoreline–gleaner carnivore; SPC, shoreline–prober carnivore; TI, terrestrial insectivore; AI, aerial insectivore; GA, granivore (Andraca, 2010); GCIF, ground–dwelling insectivore–frugivore; FIF, foliage insectivore–frugivore; P, predator; CARA, carnivore (Andraca, 2010); I, insectivore; GP, granivore (Pizarro et al., 2012); CARP, carnivore (Pizarro et al., 2012). Tabla 1. Estructura en gremios tróficos (gremios ≥ 5 especies) de los ensamblajes de aves en 19 cayos del Parque Nacional Jardines de la Reina, en el sur de Cuba: FGI, insectívoro de follaje con espigueo; ASI, insectívoro aéreo; PGIF, insectívoro–frugívoro con picoteo y espigueo; FGIF, insectívoro–frugívoro de follaje con espigueo; GGG, granívoro de suelo con espigueo; WAC, carnívoro acuático con vuelo en picada; WPP, piscívoro buceador; SGC, carnívoro de orilla con espigueo; SPC, carnívoro de orilla-prober; TI, insectívoro terrestre; AI, insectívoro aéreo; GA, granívoro (Andraca, 2010); GCIF, insectívoro–frugívoro de suelo; FIF, insectívoro–frugívoro de follaje; P, depredador; CARA, carnívoro (Andraca, 2010); I, insectívoro; GP, granívoro (Pizarro et al., 2012); CARP, carnívoro (Pizarro et al., 2012). Species

Kirkconnell et al. (1992)

Andraca (2010) Variant III

Pizarro et al. (2012)

Setophaga petechia FGI TI I Setophaga tigrina

FGI TI I

Setophaga americana FGI TI I Setophaga palmarum TI I Setophaga ruticilla AI I Seiurus aurocapilla TI I Parkesia noveboracensis

Bubulcus ibis TI I Charadrius vociferus TI I Oreothlypis peregrina FGI TI I Vireo olivaceus

FGI

Vireo altiloquus

FGI TI I

Vireo griseus

FGI

I I

Sphyrapicus varius TI I Xiphidiopicus percussus

TI I

Mniotilta varia TI I Geothlypis trichas

Setophaga caerulescens

Setophaga dominica TI I Setophaga discolor TI I Chlorostilbon ricordii I Contopus caribaeus

ASI

Contopus virens

ASI

Myiarchus sagrae

ASI

Tyrannus dominicensis

ASI

Tyrannus caudifasciatus

ASI

Chordeiles gundlachii AI I Hirundo rustica AI I Progne cryptoleuca AI I

García–Quintas et al.

Table 1. (Cont.)

Species

Kirkconnell et al. (1992)

Andraca (2010) Variant III

Pizarro et al. (2012)

Petrochelidon fulva AI I Chordeiles minor AI I Zenaida asiatica

GGG GA GP

Zenaida macroura

GGG GA GP

Zenaida aurita

GGG GA GP

Columbina passerina GGG GA GP Geotrygon montana GGG GA GP Tiaris olivaceus GA GP Passerina caerulea GA GP Passerina cyanea GA GP Agelaius humeralis GA GP Coccyzus americanus PGIF GCIF I Coccyzus minor

PGIF GCIF I

Crotophaga ani

PGIF GCIF I

Dumetella carolinensis PGIF GCIF I Mimus polyglottos

PGIF GCIF I

Quiscalus niger

PGIF GCIF I

Turdus plumbeus GCIF I Piranga rubra

FGIF FIF I

Piranga olivacea

FGIF FIF I

Pheucticus ludovicianus FGIF

FIF

Icterus galbula

FGIF FIF I

Icterus melanopsis

FGIF FIF I

Falco columbarius

D CARP

Falco peregrinus

D CARP

Buteo jamaicensis

D CARP

Tyto alba

D CARP

Asio dominguensis

D CARP

Spatula discors CARP Nyctanassa violacea CARP Buteogallus gundlachii CARP Ardea alba

WAC CARA CARP

Ardea herodias

WAC CARA CARP

Butorides virescens

WAC CARA CARP

Egretta tricolor

WAC CARA CARP

Egretta rufescens

WAC CARA CARP

Egretta caerulea

WAC CARA CARP

Egretta thula

WAC CARA CARP

Eudocimus albus CARA CARP Platalea ajaja CARP

Animal Biodiversity and Conservation 43.1 (2020)

Table 1. (Cont.)

Species

Kirkconnell et al. (1992)

Andraca (2010) Variant III

Pizarro et al. (2012)

Pelecanus occidentalis WPP CARA CARP Hydroprogne caspia

WPP CARA CARP

Thalasseus maximus WPP CARA CARP Sternula antillarum

WPP CARA CARP

Sula leucogaster

WPP CARA CARP

Sula dactylatra

WPP CARA CARP

Pandion haliaetus

WPP CARA CARP

Leucophaeus atricilla CARA CARP Charadrius wilsonia

SGC

CARA CARP

Charadrius semipalmatus

SGC

CARA CARP

Actitis macularius

SGC

CARA CARP

Pluvialis squatarola

SGC

CARA CARP

Calidris minutilla

SGC

CARA CARP

Calidris mauri

SGC

CARA CARP

Calidris alba

SGC

CARA CARP

Arenaria interpres

SGC

CARA CARP

Tringa melanoleuca

SPC

CARA CARP

Tringa semipalmata

SPC

CARA CARP

Himantopus mexicanus

SPC

CARA CARP

Limnodromus griseus

SPC

CARA CARP

Rallus crepitans

SPC

CARA CARP

Phalacrocorax auritus CARA Megaceryle alcyon CARA

Fixed–Fixed (FF) (Connor and Simberloff, 1979; Gotelli, 2000) and Proportional–Proportional (PP) (Ulrich and Gotelli, 2012) to estimate the significance level of the nestedness degree at p ≤ 0.01. Results from FF null model were prioritized over those obtained from the PP null model when matrix filling ranged between 35 and 45 % according to Ulrich and Gotelli's (2012, 2013) recommendations. The nestedness degree was compared between trophic guilds using the Z–transformed score (NODFobs – NODFexp / SDexp). Results We analysed a total of 115 bird species based on revised compilation works (table 2s in supplementary material). These included the latest additions to Caballones cay: semipalmated plover (Charadrius semipalmatus), zenaida dove (Zenaida aurita), gray catbird (Dumetella carolinensis), common yellowthroat

(Geothlypis trichas) and yellow–throated warbler (Setophaga dominica). Highest species numbers were documented in Anclitas (88), Caguama (76) and Grande (74). Species such as white–winged pigeon (Zenaida asiatica) and yellow warbler (Setophaga petechia) showed the widest distributional ranges across JRNP (18 cays). Avian guild structure yielded nine–, seven– and three– trophic groupings in accordance to classification criteria of Kirkconnell et al. (1992), Andraca (2010) (Variant III) and Pizarro et al. (2012), respectively (table 1). Guilds comprising the highest species numbers were Insectivore (43) and Carnivore (38) according to this latter classification scheme, whereas the best represented ones by the first two criteria (i.e., Kirkconnell's and Andraca's) were shoreline pecker Carnivore (eight) and Carnivore (31), respectively. Overall, avian assemblages in the JRNP showed a significant nested structure based on the difference between the observed (76.99) and simulated NOFD

García–Quintas et al.

Table 2. Mean ± SD of nestedness degree (CLlower – CLupper 95 %) of avian trophic guilds (≥ 5 species) in 19 cays of the Jardines de la Reina National Park, southern Cuba. Grouping classifications were based on criteria of Kirkconnell et al. (1992), Andraca (2010) and Pizarro et al. (2012): NODF, nestedness metric based on overlap and decreasing fill; NODFPP, null model 'proportional–proportional'; p, probability. Tabla 2. Media ± DE (LCinferior – LCsuperior 95 %) del grado de anidamiento de los gremios tróficos de aves (≥ 5 especies) en 19 cayos del Parque Nacional Jardines de la Reina, en el sur de Cuba. Las clasificaciones en grupos se basaron en los criterios de Kirkconnell et al. (1992), Andraca (2010) y Pizarro et al. (2012): NODF, índice de anidamiento basado en el relleno superpuesto y decreciente; NODFPP, modelo nulo "proporcional–proporcional"; p, probabilidad. Matrix fill (%) / Trophic guilds size (columns x rows) NOFDobs

NOFDPP N = 1000

77.10 ± 3.40

0.34

Kirkconnell et al. (1992) Water–ambusher

58.6 / 19 x 7

78.79

Carnivore Water–plunger

41.4 / 19 x 7

69.18

(69.44–82.86) 67.61 ± 5.78

Piscivore

(55.10–76.94)

Shoreline–gleaner

40.00 ± 3.12

28.3 / 19 x 8

42.12

Carnivore

(34.20–46.44)

Shoreline–prober

12.96 ± 1.66

16.8 / 19 x 5

13.44

Carnivore Ground–gleaner

31.6 / 19 x 5

54.42

Granivore Foliage–gleaner

33.8 / 19 x 7

69.79

54.28 ± 2.60 64.43 ± 4.43

Air–sallier

49.05 ± 5.50

Insectivore Pecker–gleaner

28.9 / 19 x 6

48.43

Insectivore–frugivore Foliage–gleaner

6.3 / 19 x 5

1.93

Insectivore–frugivore

0.40 0.43

(48.62–58.29) (54.94–72.41)

49.62

0.24

(9.71–16.35)

Insectivore 52.6 / 19 x 5

0.44

0.09 0.48

(39.17–60.09) 49.11 ± 3.45

0.43

(41.91–55.73) 1.68 ± 0.39

0.50

(1.10–2.21)

Andraca (2010) (Variant III) Carnivore

39.0 / 19 x 31

81.72

Predator

9.5 / 19 x 5

2.76

77.58 ± 2.42 (72.76–81.99) 3.80 ± 0.80

(2.21–5.25)

Granivore

59.19 ± 4.51

28.7 / 19 x 9

61.07

(49.66–67.98)

Aerial Insectivore

61.46 ± 3.87

40.2 / 19 x 11

61.06

Foliage

6.3 / 19 x 5

1.93

Insectivore–frugivore Ground–dwelling

27.1 / 19 x 7

49.52

1.66 ± 0.41 50.18 ± 3.68

Terrestrial Insectivore

70.50 ± 3.43

0.33 0.48 0.50

(1.10–2.21) (43.39–57.36)

81.32

0.14

(54.54–69.37)

Insectivore–frugivore 29.1 / 19 x 19

0.03

(63.56–76.93)

0.44 0.00

Animal Biodiversity and Conservation 43.1 (2020)

Table 2. (Cont.) Matrix fill (%) / Trophic guilds size (columns x rows) NOFDobs

NOFDPP N = 1000

74.17 ± 2.47

0.12

Pizarro et al. (2012) Carnivore 33.2 / 19 x 38

77.08

Granivore

28.7 / 19 x 9

(69.18 – 78.72)

61.07

Insectivore

29.9 / 19 x 43

80.04

values according to the PP null model (NODF = 2.54; p = 0.01). Results from FF null model (NODF = 76.68; p = 0.28) were not considered for analyses because its matrix filling just reached the 30.3 %. In contrast, most of the trophic guilds that we assessed did not exhibit nested structure, except for Terrestrial Insectivore (TI) and Insectivore (I) by Andraca's (2010) and Pizarro et al. (2012) classifications, respectively (table 2), with the highest nestedness degree reported in TI (fig. 3). On these matrices, no FF null model results were considered for the same reason of the matrix fill (table 3s in supplementary material).

0.35

73.27 ± 2.40

0.00

(68.43 – 78.06)

avian metacommunities throughout the archipelago (García–Quintas and Parada, 2017). Up to 42 species appear to be prone to local extirpation according to observed avian nestedness, owing to their spatially patchy occurrence (table 2s in supplementary material). The inclusion of these uncommon species among the main conservation targets of the protected area should not be entirely based on their scarcity in the JRNP. These species encompass waterfowl (e.g., blue–winged teal (Spatula discors), red–breasted merganser (Mergus serrator), greater

Discussion

Z = 3.15

Z = 2.82

80 NODF

Firstly, our results suggest that the overall avian assemblage in the JRNP is taxonomically nested. Such structuring corresponds to the interspecific differences of abundance and distribution patterns since the most abundant species at a local scale tend to influence neighbouring species assemblages to a greater extent than less common species (Patterson and Atmar, 1986). Thus, at our study site, avifauna inhabiting cays with relatively lower habitat diversity (reduced resources/ niches) may become impoverished through extinction, colonization avoidance by dispersing birds that perceive no suitable habitat, and competitive exclusion, as opposed to avifauna in cays with higher landscape complexity or more resources for species to coexist. Avian nestedness within the JRNP is consistent with findings from similar works focused at a broader spatial scale, namely, the Jardines de la Reina Archipelago as a whole (García–Quintas and Parada, 2014, 2017). This lends further support to the idea of nestedness persistence at varying spatial scales; an aspect which has remained poorly studied (Méndez, 2004). It is plausible to assert that nestedness within the cays of Laberinto de las Doce Leguas (where JRNP is located) accounts for much of the observed nested structures in the entire archipelago, because these cays exhibit the highest number of habitats for birds. Indeed, this factor is thought to be pivotal in the unfolding and persistence of nestedness in

59.39 ± 4.47 (50.55 – 68.21)

75 70 65 60

Observed Simulated TI I Trophic guilds

Fig. 3. Comparisons of nestedness degree between insectivore bird species groupings (Z– transformed score) in 19 cays of the Jardines de la Reina National Park, southern Cuba. Trophic guilds: TI, terrestrial insectivore (Andraca, 2010); I, insectivore (Pizarro et al., 2012). Fig. 3. Comparación del grado de anidamiento entre grupos de especies de aves insectívoras (puntuación Z–transformada) en 19 cayos del Parque Nacional Jardines de la Reina, en el sur de Cuba. Gremios tróficos: TI, insectívoro terrestre (Andraca, 2010); I, insectívoro (Pizarro et al., 2012).

García–Quintas et al.

yellowlegs (Tringa melanoleuca), sanderling (Calidris alba), short–billed dowitcher (Limnodromus griseus)) and raptors (e.g., red–tailed hawk (Buteo jamaicensis), peregrine falcon (Falco peregrinus), barn owl (Tyto alba)), that are usually regarded as generalist species that exploit food resources across vast areas of broadly represented habitats. Rare occurrence of avian species in the study site is likely explained by the absence of critical habitats/ecological niches, which in turn, highlights the drawbacks of nestedness studies to facilitate fact–based decisions in prioritizing conservation goals when used as the sole set of criteria (Cutler, 1994). However, the fact that significant nestedness was only reported in a small number of the trophic guilds in the JRNP avifauna may be an indication of a differential contribution of these functional groups to the overall nested structure. Thus, the occurrence of nestedness of functional groups of species brings to light ecological and biogeographical features frequently overlooked by studies focused on taxonomic nestedness only. For this reason Ulrich et al. (2017) proposed that robust identification of nested structures in natural communities demands the collection of environmental data and functional traits related to its component species. In this study, nested trophic guilds grouped the insectivore species, more specifically, the TI guild which was exclusively made up of terrestrial species. Therefore, it appears that nestedness of terrestrial insect-eating bird species could be driven by differences in food availability (insects and spiders) throughout the JRNP which, in turn, may be linked to ecological and landscape features of these cays (vegetation types, substrates, seasonal bodies of water). Species clustered in the remaining trophic guilds either relied upon different food sources (e.g., seeds, fruits, larger living preys), or foraged mostly along coastal wetlands such as shorelines and shallow coastal waters (e.g., shorebirds, piscivores, wading birds), or exploited vast foraging areas (e.g., raptors). In general, weak reliance of these species on terrestrial habitats through frequent dispersal/foraging movements between fragments (cays) and the matrix (sea water) may have diluted the effects of nestedness-generating factors. Besides, the assumption of a total isolation between islands and the neighbouring sea biotas may not be very realistic (Herrera, 2011). Although the overall species assemblage in the JRNP is organized in a nested structure, insectivore species were the only significantly nested trophic guild. This indicates that taxonomic nestedness based on guilds may shed light on finer–grain structuring patterns in metacommunities that are otherwise unnoticed if analyzed through the taxonomic approach alone. Additionally, this also highlights the importance of the insect–eating avifauna in shaping avian terrestrial communities in JRNP as a whole. As a result, our analyses provided a list of 13 locally extinction–prone species (those occurring in one or two cays in JRNP) when terrestrial insectivore (Andraca, 2010) and insectivore (Pizarro et al., 2012) guilds were combined (table 4s in supplementary material). Pizarro et al.'s (2012) classification scheme was the broadest of all three and included few trophic guilds,

with the highest species number per grouping. On the contrary, numerous guilds made up of much fewer species were identified by the criteria of Kirkconnell et al. (1992), with a small portion of those included in the analyses (> 5 species). This is largely due to the high degree of specificity (type of food, foraging substrate and behavior) by which any given species is assigned to a trophic guild. The variant III from Andraca (2010) rendered a trade–off between classification specificity and the resulting number of groupings, and was thus being the most parsimonious approach we assessed. This classification scheme may be suitable to characterize avifaunas at both regional (e.g., nation–wide, large mountain ranges) and local scales (e.g., sub– archipelagos). Nonetheless, no classification system should be deemed more adequate per se over others unless the study spatial scale, research objectives and species assemblages’ main features are taken into consideration. Our findings lend further support to the plausibility of using several guild classification criteria in the same hypothesis–testing framework as the recommended approach by Milesi et al. (2002), since detection of nested trophic guilds may be partly influenced by the chosen classification scheme. Therefore, we ratify the usage of various grouping criteria, or recommend using those classification systems that are neither too broad nor too detailed. The conservation status of avian functional groups and the ecosystem implications of bird declines have been widely addressed (Sekercioglu et al., 2004), as has the importance of tropical insectivorous birds to various types of landscapes and habitats (Gradwohl and Greenberg, 1982; Greenberg et al., 2000; Van Bael et al., 2003; Perfecto et al., 2004). Accordingly, insectivore species may require the implementation of specific management and conservation strategies given their high sensitivity to orderly loss of nested guilds based on the fragments' surface area (Matthews et al., 2015). Moreover, its functional extinction/deficiency may likely cause trophic cascades (Sekercioglu et al., 2004), especially when in absence of other functionally equivalent taxa that can replace these birds' ecosystem services. While none of the insectivore species in JRNP is currently classified as endangered, their numerical importance has been highlighted by greater species number per family (e.g., Parulidae and Tyrannidae) and individual species abundance (e.g., yellow warbler, prairie warbler Setophaga discolor) (Parada et al., 2015). Therefore, we recommend the inclusion of the bird species of this trophic guild among the conservation goals of the next JRNP Management Plan since this will target its functional roles in the entire ecosystem for protection as opposed to the typical taxonomic–based conservation. Acknowledgements We thank all those technicians at Coastal Ecosystems Research Center who timely contributed to this study. Moreover, we are grateful to David Wright, Yandy Rodríguez and Yunier Olivera for their valuable suggestions and contributions.

Animal Biodiversity and Conservation 43.1 (2020)

References Acevedo, C. J., 2013. Lista florística del archipiélago de los Jardines de la Reina, Cuba. Revista Cubana de Ciencias Biológicas, 2(3): 62–72. Almeida–Neto, M., Guimarães, P., Guimarães, P. R. Jr., Loyola, R. D., Ulrich, W., 2008. A consistent metric for nestedness analysis in ecological systems: reconciling concept and measurement. Oikos, 117: 1227–1239. Almeida–Neto, M., Ulrich, W., 2010. A straightforward computational approach for quantifying nestedness using abundance data. Environmental Modelling & Software, 26: 173–178. Andraca, L., 2010. Reglas de ensamblaje y modelos de coexistencia de especies en las comunidades de aves cubanas. Tesis de licenciatura, Universidad de La Habana. Aspin, T. W. H., Matthews, T. J., Khamis, K., Milner, A. M., Wang, Z., O'callaghan, M. J., Ledger, M. E., 2018. Drought intensification drives turnover of structure and function in stream invertebrate communities. Ecography, 41(12): 1992–2004. Bender, M. G., Leprieur, F., Mouillot, D., Kulbicki, M., Parravicini, V., Pie, M. R., Barneche, D. R., Oliveira–Santos, L. G. R., Floeter, S. R., 2017. Isolation drives taxonomic and functional nestedness in tropical reef fish faunas. Ecography, 40(3): 425–435. Connor, E. F., Simberloff, D., 1979. The assembly of species communities: chance or competition? Ecology, 60(6): 1132–1140. Cutler, A. H., 1994. Nested biotas and biological conservation: metrics, mechanisms, and meaning of nestedness. Landscape and Urban Planning, 28: 73–82. Dormann, C. F., Fründ, J., Blüthgen, N., Gruber, B., 2009. Indices, graphs and null models: analyzing bipartite ecological networks. The Open Ecology Journal, 2: 7–24. García–Quintas, A., Parada, A., 2014. Effects of migrations on the nestedness structure of bird assemblages in cays of the Jardines de la Reina archipelago, Cuba. Animal Biodiversity and Conservation, 37(2): 127–139. – 2017. Underlying factors promoting nestedness of bird assemblages in cays of the Jardines de la Reina archipelago, Cuba. Animal Biodiversity and Conservation, 40(1): 7–16, https://doi. org/10.32800/abc.2017.40.0007 Gómez–Ortiz, Y., Moreno, C. E., 2017. La diversidad funcional en comunidades animales: una revisión que hace énfasis en los vertebrados. Animal Biodiversity and Conservation, 40(2): 165–174, https:// doi.org/10.32800/abc.2017.40.0165 Gotelli, N. J., 2000. Null model analysis of species co–ocurrence patterns. Ecology, 81(9): 2606–2621. Gradwohl, J., Greenberg, R., 1982. The effect of a single species of avian predator on the arthropods of aerial leaf litter. Ecology, 63: 581–583. Greenberg, R., Bichier, P., Cruz, A., MacVean, C., Perez, R., Cano, E., 2000. The impact of avian insectivory on arthropods and leaf damage in some Guatemalan coffee plantations. Ecology,

81(6): 1750–1755. Heino, J., 2009. Species co–occurrence, nestedness and guild–environment relationships in stream macroinvertebrates. Freshwater Biology, 54: 1947–1959. Herrera, J. M., 2011. El papel de la matriz en el mantenimiento de la biodiversidad en hábitats fragmentados. De la teoría ecológica al desarrollo de estrategias de conservación. Ecosistemas, 20(2): 21–34. Kirkconnell, A., Garrido, O. H., Posada, R. M., Cubillas, S. O., 1992. Los grupos tróficos en la avifauna cubana. Poeyana, 415: 1–21. Matthews, T. J., Sheard, C., Cottee–Jones, H. E. W., Bregman, T. P., Tobias, J. A., Whittaker, R. J., 2015. Ecological traits reveal functional nestedness of bird communities in habitat islands: a global survey. Oikos, 124: 817–826. Méndez, M., 2004. La composición de especies de aves en islas y paisajes fragmentados: un análogo ecológico de las muñecas rusas. El Draque, 5: 199–212. Milesi, F. A., Marone, L., Lopez de Casenave, J., Cueto, V. R., Mezquida, E. T., 2002. Gremios de manejo como indicadores de las condiciones del ambiente: un estudio de caso con aves y perturbaciones del hábitat en el Monte central, Argentina. Ecologia Austral, 12: 149–161. Parada, A., García–Quintas, A., 2012. Avifauna de los archipiélagos del sur de Ciego de Ávila y Camagüey, Cuba: una revisión taxo–ecológica actualizada. Mesoamericana, 16(1): 35–55. Parada, A., García–Quintas, A., Hernández, D., 2015. The avifauna of sandy coastal vegetation during migration in the Laberinto de las Doce Leguas sub– archipelago, southern Cuba. Cotinga, 37: 48–58. Patterson, B. D., 1987. The principle of nested subsets and its implications for biological conservation. Conservation Biology, 1(4): 323–334. Patterson, B. D., Atmar, W., 1986. Nested subsets and the structure of insular mammalian faunas and archipelagos. Biological Journal of the Linnean Society, 28: 65–82. Peláez, O., Pavanelli, C. S., 2018. Environmental heterogeneity and dispersal limitation explain different aspects of β diversity in Neotropical fish assemblages. Freshwater Biology, 64(3): 497–505, https://doi.org/10.1111/fwb.13237 Perfecto, I., Vandermeer, J. H., López, G., Ibarra, G., Greenberg, Bichier, P., Langridge, S., 2004. Greater predation in shaded coffee farms: the role of resident neotropical birds. Ecology, 85(10): 2677–2681. Petchey, O. L., Gaston, K. J., 2006. Functional diversity: back to basics and looking forward. Ecology Letters, 9: 741–758. Pizarro, J. C., Anderson, C. B., Rozzi, R., 2012. Birds as marine–terrestrial linkages in sub–polar archipelagic systems: avian community composition, function and seasonal dynamics in the Cape Horn Biosphere Reserve (54–55ºS), Chile. Polar Biology, 35: 39–51. Sekercioglu, C. H., Daily, G. C., Ehrlich, P. R., 2004.

Ecosystem consequences of bird declines. Proceedings of the National Academy of Sciences, USA, 101: 18042–18047. Simberloff, D., Dayan, T., 1991. The guild concept and the structure of ecological communities. Annual Review of Ecology and Systematics, 22: 115–143. Ulrich, W., Almeida–Neto, M., Gotelli, N. J., 2009. A consumer’s guide to nestedness analysis. Oikos, 118: 3–17. Ulrich, W., Gotelli, N. J., 2012. A null model algorithm for presence–absence matrices based on proportional resampling. Ecological Modelling, 244: 20–27. – 2013. Pattern detection in null model analysis.

García–Quintas et al.

Oikos, 122: 2–18. Ulrich, W., Kryszewski, W., Sewerniak, P., Puchalka, R., Strona, G., Gotelli, N. J., 2017. A comprehensive framework for the study of species co–occurrences, nestedness and turnover. Oikos, doi: 10.1111/oik.04166. Van Bael, S. A., Brawn, J. D., Robinson, S. K., 2003. Birds defend trees from herbivores in a Neotropical forest canopy. Proceedings of the National Academy of Sciences, USA, 100(14): 8304–8307. Wright, D. H., Patterson, B. D., Mikkelson, G. M., Cutler, A., Atmar, W., 1998. A comparative analysis of nested subset patterns of species composition. Oecología, 113: 1–20.

Animal Biodiversity and Conservation 43.1 (2020)

Supplementary material

Table 1s. Trophic guilds of the avian assemblages in the Jardines de la Reina National Park, Cuba, according to three classification systems. Tabla 1s. Gremios tróficos de los ensamblajes de aves en el Parque Nacional Jardines de la Reina, en Cuba, según tres sistemas de clasificación. Kirkconnell et al. (1992) Bark–gleaner insectivore BGI Bark–excavator insectivore BEI Foliage–gleaner insectivore FGI Bark–and–foliage gleaner insectivore BFGI Ground–ambusher insectivore GAI Ground–pecker insectivore GPI Ground–and–water pecker insectivore GWPI Ground–and–bark–excavator insectivore GBEI Ground–and–foliage gleaner–pecker GFGPI Air–sallier lower–canopy–gleaner ASLCGI Air–sallier insectivore ASI Aerial insectivore AI Aerial nocturnal insectivore ANI Air–sallier insectivore–frugivore ASIF Air–sallier–and–forager insectivore–frugivore ASFIF Foliage–gleaner insectivore–frugivore FGIF Pecker–gleaner insectivore–frugivore PGIF Ground–forager insectivore–frugivore GFIF Floral–hover–gleaner nectarivore–insectivore FHGNI Frugivore–nectarivore FN Ground–gleaner granivore GGG Ground–and–foliage granivore GFG Andraca (2010) (Variant III) Terrestrial Insectivore IT Aerial–insectivore IAA Foliage insectivore–frugivore FIF Ground–dwelling insectivore–frugivore GIF Nectarivore–insectivore NI Frugivore FA Granivore G Predator D Insectivore GFGPI Insectivore ASLCGI Omnivore OA Scavenger SCAVA Carnivore–phytophage C–P Water–dabbler WDA Carnivore CAR

Frugivore F Frugivore–granivore FG Ground–and foliage pecker–and–gleaner Granivore–insectivore GFPGGI Aerial Predator AP Aerial nocturnal predator ANP Air–sallier nocturnal predator ASNP Insectivore and small–vertebrate predator ISIP Crustaceovore C Molluscovore M Omnivore O Scavenger Scav Water–ambusher carnivore WAC Shoreline–prober carnivore SPC Shoreline–gleaner carnivore SGC Water–plunger piscivore WPP Water–prober carnivore–phytophage WPCP Pecker phytophage–carnivore PPC Diver carnivore–phytophage DCP Mud strainer MS Water–dabbler omnivore WDO Water–dabbler–and–diver piscivore WDDP Air–sallier–and water–diver piscivore ASWDP Water–diver piscivore

WDP

Pizarro et al. (2012) Insectivore I Frugivore FP Granivore GP Carnivore CARP Carnivore–scavenger Scavenger

CAR–S SCAV

Omnivore

Piscivore

PIS

Crustaceovore CA Molluscovore

García–Quintas et al.

Table 2s. Incidence matrix of 115 bird species occurring in 19 cays of the Jardines de la Reina National Park, southern Cuba: * presence; blank, not recorded; Oo, overall occurrence. Tabla 2s. Matriz de incidencia de 115 especies de aves presentes en 19 cayos del Parque Nacional Jardines de la Reina, en el sur de Cuba: * presencia; vacío, no registrada; Oo, presencia generalizada.

Species/Cay

Anc

Cag

Gra Cab Bre

Zenaida asiatica

White–winged pigeon

* * * * *

Setophaga petechia

Yellow warbler

* * * * *

Thalasseus maximus

Royal tern

Patagioenas leucocephala

White crowned pigeon

Phalacrocorax auritus

Double–crested cormorant

Ardea alba

Great egret

Quiscalus niger

Greater Antillean grackle

Pelecanus occidentalis

Brown pelican

Pandion haliaetus

Osprey

Tyrannus dominicensis

Gray kingbird

Ardea herodias

Great blue heron

Fregata magnificens

Magnificent frigatebird

Agelaius humeralis

Tawny–shouldered blackbird

* * * *

Vireo altiloquus

Black whiskered vireo

* * * * *

Chlorostilbon ricordii

Cuban emerald

* * * * *

Charadrius wilsonia

Wilson's plover

* * * *

Tyrannus caudifasciatus

Loggerhead kingbird

* * * * *

Butorides virescens

Green heron

Cathartes aura

Turkey vulture

Anhinga anhinga

Anhinga

* * * * *

Buteogallus gundlachii

Cuban black hawk

Egretta tricolor

Tricolored heron

Egretta rufescens

Reddish egret

* * * * *

Contopus caribaeus

Cuban pewee

* * * * *

Myiarchus sagrae

La Sagra's flycatcher

Eudocimus albus

White ibis

* * * *

Setophaga discolor

Prairie warbler

* * * * *

Arenaria interpres

Ruddy turnstone

Chordeiles gundlachii

Antillean nighthawk

Hirundo rustica

Barn swallow

* * * *

Setophaga ruticilla

American redstart

* * * * *

Xiphidiopicus percussus Cuban green woodpecker

* * * *

Parkesia noveboracensis

Northern waterthrush

Thalasseus sandvicensis

Sandwich tern

Egretta thula

Snowy egret

Mniotilta varia

Black–and–white warbler

Geothlypis trichas

Common yellowthroat

Animal Biodiversity and Conservation 43.1 (2020)

iii

Table 2s. (Cont.)

CaE

Cin

Cru Cac BPP

BSe

BPC

BGr

Alc

JGr

Alz

BRi

Cam Lar

* * * * * * * * * * * * * 18 * * * * * * * * * * * * * 18 *

17 17

* *

* * 15

* * * * * * * * * * 14

* * * * * * * * 13

* * * * * * * * 13 * * * * * * * * 12 * * * * * * * 12 *

12 12

* * * * * * * 12 *

* * * * * * 11 * * * * * * 11 *

* * * * * * * 11 * * * * * 10

* * * * 8 * * * 8 * * * * 8 *

García–Quintas et al.

Table 2s. (Cont.)

Species/Cay

Anc

Gra Cab Bre

Egretta caerulea

Little blue heron

Zenaida macroura

Mourning dove

Calidris minutilla

Least sandpiper

Platalea ajaja

Roseate spoonbill

Rallus crepitans

Clapper rail

Setophaga americana

Northern parula

* * * * *

Actitis macularius

Spotted sandpiper

Setophaga caerulescens Black–throated blue warbler Setophaga palmarum

Palm warbler

Cag

* *

* * * * *

* * * *

Setophaga dominica

Yellow–throated warbler * * * *

Megaceryle alcyon

Belted kingfisher

Leucophaeus atricilla

Laughing gull

* * *

Dumetella carolinensis

Gray catbird

* * * * *

Petrochelidon fulva

Cave swallow

* * *

Coccyzus americanus

Yellow–billed cuckoo

* * *

Pluvialis squatarola

Black–bellied plover

* * * *

Seiurus aurocapilla

Ovenbird

Tringa semipalmata

Willet

* * *

Sternula antillarum

Least tern

* *

Charadrius semipalmatus

Semipalmated plover

Progne cryptoleuca

Cuban martin

* * * *

Setophaga tigrina

Cape may warbler

* * *

Turdus plumbeus

Red–legged thrush

* * *

Falco columbarius

Merlin

* * *

Mimus polyglottos

Northern mockingbird

* * *

Vireo olivaceus

Red–eyed vireo

* * *

Himantopus mexicanus

Black–necked stilt

* * *

Crotophaga ani

Smooth–billed ani

Passerina cyanea

Indigo bunting

* * *

Nyctanassa violacea

Yellow–crowned night–heron

* *

Zenaida aurita

Zenaida dove

* * *

Buteo jamaicensis

Red–tailed hawk

* *

Tringa melanoleuca

Greater yellowlegs

* *

Columbina passerina

Common ground–dove * *

Falco peregrinus

Peregrine falcon

* *

Bubulcus ibis

Cattle egret

* *

Vireo griseus

White–eyed vireo

* *

Helmitheros vermivorum Worm–eating warbler

* *

Patagioenas squamosa

Scaly–naped pigeon

Piranga olivacea

Scarlet tanager

Animal Biodiversity and Conservation 43.1 (2020)

Table 2s. (Cont.)

CaE

Cin Cru Cac BPP

BSe

BPC BGr

Alc

JGr

Alz

BRi

Cam Lar

* *

5 *

* 5 * 5

* * 5

* 4

* 4 * * 4

3 * 3

García–Quintas et al.

Table 2s. (Cont.)

Species/Cay Spatula discors

Blue–winged teal

Anc

Cag

Gra Cab Bre

* *

Calidris mauri

Western sandpiper

* *

Setophaga citrina

Hooded warbler

* *

Coccyzus minor

Mangrove cuckoo

Protonotaria citrea

Prothonotary warbler *

Oreothlypis peregrina

Tennessee warbler *

Geotrygon montana

Ruddy quail–dove

Tyto alba

Barn owl

Chordeiles minor

Common nighthawk

Mergus serrator

Red–breasted merganser

Limnodromus griseus

Short–billed dowitcher

Setophaga castanea

Bay–breasted warbler *

Setophaga fusca

Blackburnian warbler *

Numenius phaeopus

Whimbrel

Catharus minimus

Gray–cheeked thrush

Catharus fuscescens

Veery

Sphyrapicus varius

Yellow–bellied sapsucker *

Sula leucogaster

Brown booby *

Icteria virens

Yellow–breasted chat *

Tiaris olivaceus

Yellow–faced grassquit

Piranga rubra

Summer tanager

Charadrius vociferus

Killdeer *

Pheucticus ludovicianus Rose–breasted grosbeak * Passerina caerulea

Blue grosbeak

Contopus virens

Eastern wood–pewee *

Dolichonyx oryzivorus

Bobolink

Hydroprogne caspia

Caspian tern

Calidris alba

Sanderling

Icterus galbula

Baltimore oriole *

Asio dominguensis

Short–eared owl

Sula dactylatra

Masked booby *

Icterus melanopsis

Cuban oriole *

Catharus ustulatus

Swainson's thrush

Species's total

* 73

Animal Biodiversity and Conservation 43.1 (2020)

vii

Table 2s. (Cont.)

CaE

Cin Cru Cac BPP

BSe

BPC BGr

Alc

JGr

Alz

BRi

Cam Lar

* 1

García–Quintas et al.

viii

Table 3s. Mean ± SD of nestedness degree (CLlower – CLupper 95 %) of avian trophic guilds (≥ 5 species) in 19 cays of the Jardines de la Reina National Park, southern Cuba. Grouping classifications were based on criteria of Kirkconnell et al. (1992), Andraca (2010) and Pizarro et al. (2012): NODF, nestedness metric based on overlap and decreasing fill; NODFFF, null model 'Fixed–Fixed'; p, probability.

Trophic guilds

Matrix fill (%) / size (columns x rows) NOFDobs

NOFDFF N = 1000

79.58 ± 0.97

0.18

Kirkconnell et al. (1992) Water–ambusher

58.6 / 19 x 7

78.79

Carnivore Water–plunger

41.4 / 19 x 7

69.18

(77.06–80.74) 69.33 ± 1.38

Piscivore

(65.39–70.74)

Shoreline–gleaner

41.96 ± 0.65

28.3 / 19 x 8

42.12

Carnivore Shoreline–prober

16.8 / 19 x 5

13.44

13.55 ± 0.53 (12.66–14.23)

Ground–gleaner

55.97 ± 0.36

54.42

Granivore Foliage–gleaner

33.8 / 19 x 7

69.79

69.79 ± 0.15 (69.79–69.79)

Air–sallier

49.68 ± 0.54

49.62

Insectivore Pecker–gleaner

28.9 / 19 x 6

48.43

Insectivore–frugivore Foliage–gleaner

6.3 / 19 x 5

1.93

Insectivore–frugivore

0.38 0.01

(55.06–56.54)

Insectivore 52.6 / 19 x 5

0.50

(40.82–42.62)

Carnivore 31.6 / 19 x 5

0.28

0.50 0.50

(48.47–50.13) 48.49 ± 0.23

0.45

(48.12–48.70) 1.72 ± 0.30 (1.10–1.93)

0.50

Animal Biodiversity and Conservation 43.1 (2020)

Tabla 3s. Media ± DE (LCinferior – LCsuperior 95 %) del grado de anidamiento de los gremios tróficos de aves (≥ 5 especies) en 19 cayos del Parque Nacional Jardines de la Reina, en el sur de Cuba. Las clasificaciones en grupos se basaron en los criterios de Kirkconnell et al. (1992), Andraca (2010) y Pizarro et al. (2012): NODF, índice de anidamiento basado en el relleno superpuesto y decreciente; NODFFF, modelo nulo FF "Fijo–Fijo"; p, probabilidad.

Trophic guilds

Matrix fill (%) / size (columns x rows) NOFDobs

NOFDFF N = 1000

81.94 ± 0.75

0.32

Andraca (2010) (Variant III) Carnivore

39.0 / 19 x 31

81.72

Predator

9.5 / 19 x 5

2.76

(80.13–82.99) 3.81 ± 0.68

(2.21–4.97)

Granivore

62.10 ± 0.86

28.7 / 19 x 9

61.07

(60.06–63.33)

Aerial

61.40 ± 1.07

40.2 / 19 x 11

61.06

Insectivore Foliage

6.3 / 19 x 5

1.93

1.72 ± 0.30 (1.10–1.93)

Ground–dwelling

49.65 ± 0.17

49.52

Insectivore–frugivore

(49.22–49.96)

Terrestrial

81.21 ± 0.49

29.1 / 19 x 19

81.32

Insectivore

0.12 0.31

(58.76–62.86)

Insectivore–frugivore 27.1 / 19 x 7

0.11

0.50 0.19 0.48

(80.15–81.87)

Pizarro et al. (2012) Carnivore

33.2 / 19 x 38

77.08

78.33 ± 0.80

(76.48–79.50)

Granivore

62.08 ± 0.85

28.7 / 19 x 9

61.07

Insectivore

29.9 / 19 x 43

80.04

0.08 0.12

(60.14–63.33) 80.05 ± 0.65 (78.58–80.93)

0.41

GarcĂaâ&#x20AC;&#x201C;Quintas et al.

Table 4s. Incidence matrix of 43 insectivore bird species found in 19 cays of the Jardines de la Reina National Park, southern Cuba: * presence; blank, not recorded; Oo, overall occurrence. Matrix contains species grouped in the trophic guilds Terrestrial Insectivore (Andraca, 2010) and Insectivore (Pizarro et al., 2012).

Species / Cay Setophaga petechia Quiscalus niger Tyrannus dominicensis Chlorostilbon ricordii Vireo altiloquus Tyrannus caudifasciatus Myiarchus sagrae Contopus caribaeus Setophaga discolor Chordeiles gundlachii Xiphidiopicus percussus Hirundo rustica Setophaga ruticilla Parkesia noveboracensis Mniotilta varia Geothlypis trichas Setophaga caerulescens Setophaga palmarum Setophaga dominica Setophaga americana Dumetella carolinensis Seiurus aurocapilla Petrochelidon fulva Coccyzus americanus Progne cryptoleuca Mimus polyglottos Vireo olivaceus Setophaga tigrina Turdus plumbeus Crotophaga ani Vireo griseus Bubulcus ibis Piranga olivacea Oreothlypis peregrina Charadrius vociferus Sphyrapicus varius Chordeiles minor Coccyzus minor Icterus galbula Icterus melanopsis Contopus virens Piranga rubra Pheucticus ludovicianus Species's total

Gra Anc Cag Cab Bre CaE Cac BPC * * * * * * * * * * * * * * * * * * * * * * * * * * * * * * * * * * * * * * * * * * * * * * * * * * * * * * * * * * * * * * * * * * * * * * * * * * * * * * * * * * * * * * * * * * * * * * * * * * * * * * * * * * * * * * * * * * * * * * * * * * * * * * * * * * * * * * * * * * * * * * * * * * * * * * * * * * * * * * * * * * * * * * * * * * * * * * * * * * * * * * * * 34 33 33 24 20 13 10 10

Animal Biodiversity and Conservation 43.1 (2020)

Tabla 4s. Matriz de incidencia de 43 especies de aves insectívoras encontradas en 19 cayos del Parque Nacional Jardines de la Reina, en el sur de Cuba: * presencia; vacío, no registrada; Oo, presencia generalizada. La matriz contiene las especies agrupadas en los gremios tróficos Insectívoro Terrestre (Andraca, 2010) e Insectívoro (Pizarro et al., 2012).

Cru BPP Alc Cin BSe JGr BGr Alz Cam BRi Lar * * * * * * * * * 18 * * * * * * * * * 17 * * * * * * 15 * * * * * * * * * * * * * 13 * * * * 12 * * * 11 * * * * * * * 10 * * * 9 * * * 8 * * * 4 10 9 9 8 7 6 6 6 3 3 0

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

Muñoz and Farfán

Animal Biodiversity and Conservation 43.1 (2020)

Coexistence of jaguars (Panthera onca) and pumas (Puma concolor) in a tropical forest in south–eastern Mexico D. M. Ávila–Nájera, C. Chávez, S. Pérez–Elizalde, J. Palacios–Pérez, B. Tigar Ávila–Nájera, D. M., Chávez, C., Pérez–Elizalde, S., Palacios–Pérez, J., Tigar, B., 2020. Coexistence of jaguars (Panthera onca) and pumas (Puma concolor) in a tropical forest in south–eastern Mexico. Animal Biodiversity and Conservation, 43.1: 55–66, Doi: https://doi.org/10.32800/abc.2020.43.0055 Abstract Coexistence of jaguars (Panthera onca) and pumas (Puma concolor) in a tropical forest in south–eastern Mexico. The biological ranges of the jaguar (Panthera onca) and puma (Puma concolor) overlap in the Yucatan Peninsula, corresponding to the most important population of jaguars in Mexico. The goal of this study in the El Eden Ecological Reserve (EER) was to investigate the factors that permit these two predators to coexist in the dense vegetation of medium–stature tropical forest and secondary forest in the north–eastern Yucatan Peninsula. We assessed their spatial and temporal overlap using Pianka’s index, and evaluated their habitat use by applying occupancy models. A total sampling effort of 7,159 trap–nights over 4 years produced 142 independent photographic records of jaguars, and 134 of pumas. The felids showed high to very high overlap in their use of different vegetation (0.68–0.99) and trail types (0.63–0.97) and in their activity patterns (0.81–0.90). However, their peak activity patterns showed some temporal separation. Time of day, particularly for peak activity time, was the best predictor to explain the coexistence of the felids in this habitat. While occupancy models showed that the presence of potential prey species and vegetation type could predict the presence of felids in the study area. Natural disturbances during 2010 (hurricane) and 2011 (fire) drastically changed habitat use and activity patterns, resulting in pumas and jaguars adjusting their resource–use and activity pattern through a strategy of mutual evasion. Keys words: Big cats, Activity pattern, Habitat use, Prey, Occupancy models Resumen Coexistencia del jaguar (Panthera onca) y el puma (Puma concolor) en un bosque tropical del sureste de México. La distribución del jaguar (Panthera onca) y el puma (Puma concolor) se superponen en la Península de Yucatán, donde se encuentra la población más importante de jaguares en México. El objetivo de este estudio, realizado en la Reserva Ecológica El Eden, fue estudiar los factores que permiten que estos dos depredadores coexistan en la densa vegetación de la selva mediana tropical y los bosques secundarios del noreste de la península de Yucatán. En el estudio se evaluó la superposición en el tiempo y el espacio utilizando el índice de Pianka y se analizó el uso que hacen del hábitat estas dos especies mediante modelos de ocupación. Un esfuerzo de muestreo total de 7.159 noches/trampa durante cuatro años produjo 142 registros fotográficos independientes de jaguares y 134 de pumas. Los félidos mostraron una superposición alta o muy alta en el uso de vegetación (0,68–0,99) y los tipos de senderos (0,63–0,97) y en sus patrones de actividad (0,81–0,90). Sin embargo, sus picos de actividad muestran una cierta separación temporal. El momento del día, en particular para los picos de actividad, fue el factor que mejor explicaba la coexistencia de los félidos en este hábitat. Los modelos de ocupación mostraron que la presencia de presas potenciales y el tipo de vegetación podrían predecir la presencia de félidos en la zona del estudio. Las perturbaciones naturales acaecidas durante 2010 (huracán) y 2011 (incendio) cambiaron drásticamente el uso del hábitat y los patrones de actividad de forma que los pumas y los jaguares adaptaron el uso de los recursos y sus patrones de actividad mediante una estrategia de evasión mutua. Palabras clave: Grandes felinos, Patrón de actividad, Uso del hábitat, Presas, Modelos de ocupación ISSN: 1578–665 X eISSN: 2014–928 X

Ávila–Nájera et al.

Received: 14 II 19; Conditional acceptance: 04 VI 19; Final acceptance: 20 IX 19 Dulce María Ávila–Nájera, Departamento de Investigación, Universidad Intercultural del Estado de México, Libramiento Francisco Villa, s/n., San Felipe del Progreso, Estado de México, C.P. 50640, México.– Cuauhtémoc Chávez, Departamento de Ciencias Ambientales, CBS Universidad Autónoma Metropolitana, Unidad Lerma, Hidalgo Pte. 46, Col. La Estación Lerma, Estado de México, C.P. 52006, México.– Sergio Pérez–Elizalde, Colegio de Postgraduados, Campus Montecillo, Carretera México–Texcoco km 36.5, Montecillo, Texcoco, C.P. 56230, México.– Jaime Palacios–Pérez, Wildlife Conservation Society, Ecuador. Avenida de los Granados N40–53 y París, Quito, Ecuador.– Barbara Tigar, School of Forensic and Applied Sciences, University of Central Lancashire, Preston, PR1 2HE UK. Corresponding author: Cuauhtémoc Chávez. E–mail: j.chavez@correo.ler.uam.mx

Animal Biodiversity and Conservation 43.1 (2020)

Introduction Jaguars (Panthera onca) and pumas (Puma concolor) occur sympatrically in their neotropical ranges, with both species experiencing continued range contractions resulting from habitat loss and fragmentation, and anthropogenic activites including direct persecution (Sanderson et al., 2002; Scognamillo et al., 2003). Most jaguar studies focus on their central and southerly populations, and tropical biomes, while their more northerly populations in the Yucatan Peninsula in south–eastern Mexico are poorly known (Faller et al., 2007; Chávez, 2010). Pumas have been widely studied throughout the most northerly parts of their range, particularly temperate and continental parts of the USA and Canada, but little is known of their tropical (Foster et al., 2010a) and Mexican populations (Monroy–Vilchis and Soria–Díaz, 2013). Furthermore, the majority of coexistence studies on these felids are from humid tropical and sub–tropical forests (Nuñez et al., 2002; Scognamillo et al., 2003; Foster et al., 2010a; Faller et al., 2007) and semiarid regions (Astete et al., 2017; Gutierrez–González and López–González, 2017). The coexistence of two similar–sized carnivores has stimulated research into the mechanisms that allow them to partition resources, including specialization in their temporal and spatial use of prey or habitats (Carothers and Jaksic, 1984; Linnel and Strand, 2000; Donadio and Buskirk, 2006; Foster et al., 2013). Complex interactions between coexisting jaguars and pumas are related to their habitat and prey use (Woodroffe, 2001; Scognamillo et al., 2003; Foster et al., 2010a, 2010b; Sollman et al., 2012). Evidence to support this includes their differential use of vegetation, particular densely vegetated habitats (Hanski, 1994; Creel and Creel, 1996; Durant, 1998; Fedriani et al., 1999; Maffei et al., 2004; Chávez, 2010; Di Bitetti et al., 2010; Foster et al., 2013) and temporal differences that facilitate evasion (Aranda and Sánchez–Cordero, 1996; Romero–Muñoz et al., 2010) such as different activity regimes to help avoid conflict (Paviolo et al., 2009; Di Bitetti et al., 2010; Foster et al., 2013; Hérnandez–Saint Martín et al., 2013; Àvila–Nàjera et al., 2016). Examples of dietary specialization include the dominant species —usually considered to be the jaguar (Sollman et al., 2012)— selecting larger prey, and changes to niche breadth seen from differential prey selection by size, age and taxa (Gittleman, 1985; Aranda, 1994; Karanth and Sunquist, 1995; Aranda and Sánchez–Cordero, 1996; Taber et al., 1997; Karanth and Nichols, 1998; Núñez et al., 2000; Scognamillo et al., 2003; Chávez, 2010; Di Bitetti et al., 2010; Foster et al., 2013). The many small reserves and protected areas in the Yucatan Peninsula, Mexico, form patches of interconnected natural habitat (Pozo et al., 2011) where sympatric jaguar and puma populations occur in a landscape mosaic dominated by semi–natural environments and human activity (Zarza et al., 2007). Camera traps are increasingly used in ecological and behavioural studies of large nocturnal predators that roam widely over their extensive home ranges (Núñez

et al., 2002; Chávez, 2010; Foster et al., 2013). In the present study we used camera trap evidence to investigate which factors permit jaguar and puma to coexist in the tropical forest of the El Eden Ecological Reserve (EER) in the north–eastern Yucatan Peninsula. We assessed the degree of overlap in their resource–use (spatial and temporal) and applied occupancy models (MacKenzie et al., 2006) to evaluate how differences in their habitat use and activity regimes allow them to coexist. We tested the hypotheses that different habitat components directly affect the temporal and spatial occurrence of jaguars and pumas, and that flexibility in their daily activity patterns and habitat use allow them to minimize their interactions with each other, and avoid direct competition. Material and methods Study area The El Eden Ecological Reserve covers an area of 3,077 ha of the northernmost tropical forests of North America, and is congruent with the larger Yum Balam Protected Area (Navarro et al., 2007) (fig. 1E). It consists mainly of medium stature tropical forest (MSTF) with secondary forest (acahual) being the dominant tree species described (Schultz, 2003). Fieldwork Camera traps operating 24 h/d were deployed in July– September 2008, October–December 2010, May–July 2011, and August–November 2012 (Cuddeback expert, Capture, Capture IR, Moultrie and Wildview), and images were downloaded every 15 days. Traps were sited using the Mexican National Census of the jaguar and its prey design (CENJAGUAR) (Chávez et al., 2007), with up to three cameras placed 1.5–3 km apart in 9 km2 plots. At least one site per plot had paired cameras to capture images of both sides of any animal that triggered the trap. Cameras were placed along forest paths, firebreaks and minor roads, and were re–positioned each year across the two dominant vegetation types (MSTF and seconday forest), as shown in figure 1A–1D. Camera trap analysis We identified individual jaguars by their coat patterns and markings, and pumas by their scars, coloration patterns, and body shape (Kelly et al., 2008). Photographs were grouped by trap site to perform the analyses. Photographs were considered as independent events (1) when the same individual was photographed again more than 30' later, (2) when different individuals could be distinguished in consecutive photos, (3) when several individuals were clearly identifiable in a single photo and (4) when individuals could not be identified in consecutive photos, in which case a new event was recorded after 3 h (Ávila–Nájera et al., 2016). All records were placed into one of three time classes: nocturnal (20:00–06:00 h), diurnal

Ávila–Nájera et al.

(08:00–18:00 h) or crepuscular (06:00–08:00 h and 18:00–20:00 h) (Gómez et al., 2005). Any cameras > 1.5 km apart were considered as independent sampling units, and assumed to be equally accessible to all felids. The cameras were grouped by vegetation type (MSTF and secondary forest) and site characteristics (forest path, firebreak or road). Statistical analyses The overlap in activity pattern and habitat use was estimated via Pianka's Index (Pianka, 1973), where 0 indicates no overlap and 1 is complete overlap in resource use. All tests and graphics were calculated using the R statistical package (version 3.1.0). To understand changes or differences in the proportion of sites occupied by jaguars and pumas, the imperfect detection of these species had to be taken into account as this can result in some occupied sites appearing to be unoccupied. We estimated the probability of occupancy (ψ) and detection (p) of jaguar and puma based on their detection rates from the 70 sampling days in 2008 and 2012. These data were used to run occupancy models for each species and each sampling period or year (MacKenzie et al., 2006). The models that we considered assume that occupancy was either constant across sites ψ or varied by site according to the variables ψ (type of prey, prey interactions, co–predators, vegetation type or trail type). Detectability was either constant across both years and sites or varied according to features of the camera trap site (path type, vegetation type or co–predators). Final model selection used Akaike's Information Criteria for small sample size (AICc) to identify the most parsimonious model, balancing model fit and parameter precision, where models with lower AICc are considered best. It should be mentioned that the reserve was affected by a hurricane in September 2010 and by a fire that occurred outside EER in May 2011. The results reported therefore take these changes into account in the environment. Results A total sampling effort of 7,159 trap nights over the four years produced 142 independent photographic records of jaguars and 134 of pumas. Habitat use by felids During the study most jaguars were recorded in secondary forest (80 %) as seen in 2008, 2010 and 2012 (95 %, 62 % and 75 % respectively), although the majority of sightings in 2011 were from MSTF (71 %). The only significant differences in jaguar habitat use were seen in 2008 (x2 = 21.88, p < 0.05) and 2011 (x2 = 159.98, p < 0.05). Pumas occurred in both forest types, and used secondary forest and MSTF roughly equally in 2008 (54 % and 46 % respectively), but were seen more often in MSTF in 2010 and 2011 (75 % and 53 % respectively) and more often in secondary

forest in 2012 (78.6 %). The only significant differences in puma habitat use occurred in 2008 (x2 = 8.22, p = 0.02) and 2011 (x2 = 159.9, p < 0.05). More jaguars were seen on roads (67 %–88 %) than on forest paths (20 %) or firebreaks (13 %), and difference in the type of trail used by jaguars in 2008 was significant (x2 =36.88, p < 0.05) and 2011 (x2 = 228.76, p < 0.05). However, pumas were mainly seen on forest paths (46 %–75 %) and roads (44 %), with only 10 % recorded on firebreaks. In 2012, only 21% of puma records were from forest paths. There was a significant difference in the type of trail used by pumas in 2011 (x2 = 228.76, p < 0.005) and 2012 (x2 = 9.10, p = 0.03). Activity patterns of felids Both felids were predominantly crepuscular–nocturnal (jaguar 69 % and pumas 64 %) although about a third of all sightings were diurnal (fig. 2). Their activity patterns differed between the years, with less nocturnal and more diurnal activity in 2010, and predominantly nocturnal jaguar activity (86 %) with no crepuscular sightings in 2011, compared with 48 % nocturnal and 17 % crepuscular records for puma. There were significant differences in in the activity patterns of jaguars in 2011 (x2 = 176.67, p ˃ 0.00) and 2012 (x2 = 1.32, p ˃ 0.01) and of pumas in 2008 (x2 = 1053.77, p ˃ 0.00), 2011 (x2 = 176.67, p ˃ 0.00) and 2012 (x2 ˃ 1053.77, p ˃ 0.00). Spatial and temporal overlap of felids The felids showed a high to very high overlap in the use of vegetation and path types (0.63–0.99), particularly in 2011 and 2012. There was also a very high overlap in the their activity patterns (0.81–0.90) in most years (table 1). Occupancy models The best occupancy model for pumas was in 2008 (0.68 with 0.30–0.88 CIs, and an AIC of 234.76 and AIC Wgt of 0.96), and was mainly explained by the presence of collared pecaries and by vegetation type (table 2). However, pumas were also affected by the presence of jaguars, which when included as a variable, produced a model for 2012 which had the lowest AIC value (184.17). However, for both years and all variables selected, none of the models or variables (AIC, ΔAIC and AIC Wgt) predicted the presence of jaguar in the El Eden. The models with the lowest AIC are shown in table 2. However, the ΔAIC values showed no difference between the models. Discussion Habitat use by felids Jaguars were seen on all trail types but were more frequently seen on roads, while pumas used paths,

Animal Biodiversity and Conservation 43.1 (2020)

Acahual (secondary forest)

Mexico Pacific Ocean

Gulf of Mexico Study Area Yucatan Peninsula

Mixed farming Savanna Medium forest Camera station Sink hole 0

10 km

500 km

Fig. 1. Camera trapping stations (black circles) at the El Eden Ecological Reserve, Quintana Roo, Mexico plotted by year of study with the major vegetation types: A, 2008; B, 2010; C, 2011; D, 2012. Fig. 1. Estaciones de trampeo con cámaras (círculos negros) en la Reserva Ecológica El Edén, en Quintana Roo, México, por año de estudio con los principales tipos de vegetación: A, 2008; B, 2010; C, 2011; D, 2012.

firebreaks and roads according to availability, although this changed following the fire. In other studies, jaguars frequently use roads because they facilitate movement and scent marking (Maffei et al., 2004), although male jaguars are more likely to use roads than females (Conde et al., 2010; Maffei et al., 2011). Pumas also take advantage of roads to move around their home ranges, but the proximity to fire damage

combined with the high presence of co–predators like jaguars may have made them less favorable to pumas in 2011. Normally the roads in EER have little human traffic, but the fire in 2011 resulted in firefighters and people hired to put out the fires frequently travelling on the roads and around the reserve, creating high levels of disturbance. The secondary forest surrounding many roads in EER is extremely dense, so

Ávila–Nájera et al.

Table 1. Overlap in resource use (Pianka Index) between jaguars (Panthera onca) and pumas (Puma concolor) in the El Eden Ecological Reserve, Quintana Roo, Mexico: CI, confidence interval. Tabla 1. Superposición en el uso de recursos (índice de Pianka) entre el jaguar (Panthera onca) y el puma (Puma concolor) en la Reserva Ecológica El Edén, en Quintana Roo, México: CI, intérvalo de confianza.

Median overlap (Pianka's Index) SD

CI 2.5%

CI 97.5%

Vegetation type

2008

0.68

0.11

0.45

0.87

2010

0.75

0.18

0.35

1.00

2011

0.90

0.10

0.64

1.00

2012

0.99

0.01

0.95

1.00

Path type

2008

0.63

0.11

0.40

0.83

2010

0.76

0.17

0.35

0.99

2011

0.90

0.10

0.64

1.00

2012

0.97

0.03

0.89

1.00

Activity pattern

2008

0.90

0.07

0.73

0.99

2010

0.88

0.11

0.59

1.00

2011

0.81

0.11

0.56

0.96

2012

0.88

0.08

0.71

0.99

Two–hour time periods

2008

0.67

0.10

0.46

0.85

2010

0.28

0.14

0.04

0.60

2011

0.28

0.15

0.04

0.61

2012

0.64

0.10

0.43

0.83

roads are highly likely to be used by both felids, as is common in other parts of their range (Dickson and Beier, 2002; Harmsen et al., 2010; Rodríguez–Soto et al., 2013), although pumas generally prefer small paths with high tree cover elsewhere in Mexico (Lira and Naranjo, 2003). Although we observed changes in the jaguars use of resources, secondary forests were used consistently to a greater or lesser extent over the study period. In more humid tropical forests, dense horizontal and vertical vegetation cover is thought to be essential for their permanency (Scognamillo et al., 2003; Conde et al., 2010), but while both forms of vegetation are present in EER they were not significant factors in the occupancy models. After the fire in 2011, the area immediately around the perimeter was damaged and cameras recorded less activity; both species are known to be sensitive to changes in the level of human activity, and prey and co–predator abundance (Carrillo, 2000; Novack et al., 2005; Haines, 2006; McLoughlin et al., 2010; Foster et al., 2013). Environmental changes following natural or anthropogenic disturbances are therefore likely to effect the interactions between puma and jaguar, and may result in changes in behaviour and resource use.

Activity patterns of felids Jaguars in the El Eden were active 24 h/d although they were mainly crepuscular and nocturnal, with most activity occurring early in the morning, as in the southern Yucatan Peninsula (Chávez et al., 2007). We observed two nocturnal activity peaks, and most crepuscular activity occurred around dusk. Jaguars’ activity patterns vary across their range and are thought to be influenced by the activity patterns of their prey (Carrillo, 2000; Scognamillo et al., 2003). For example, some jaguars are predominatly diurnal (Rabinowitz and Nottingham, 1986; Álvarez–Castañeda and Patton, 2000; Maffei et al., 2004; Harmsen et al., 2009; Maffei et al., 2011; Foster et al., 2013) and least active at midnight (Maffey et al., 2004) as are their most important prey such as Mazama sp. and Tayassu sp. (Barrientos and Maffei, 2000). Pumas at EER were also active 24 h/d, and they were mainly cathemeral. There was a strong positive association between peak puma activity and that of their main prey (nine–banded armadillos, collared peccaries and red brocket deer) and a negative temporal association with jaguars, suggesting a possible copredator evasion strategy (Ávila–Nájera et al., 2018b). Similar activity patterns have been

Animal Biodiversity and Conservation 43.1 (2020)

Table 2. Occupany models for jaguar (Panthera onca) and puma (Puma concolor) in the El Eden Ecological Reserve, Quintana Roo, Mexico in 2008 and 2012: AIC, Akaike's information criteria for small sample sizes; ψ, occupancy probability; p, detection probability. Tabla 2. Modelos de ocupación del jaguar (Panthera onca) y el puma (Puma concolor) en la Reserva Ecológica El Edén, en Quintana Roo, México en 2008 y 2012: AIC, criterio de información de Akaike para muestras pequeñas; ψ, probabilidad de ocupación; p, probabilidad de detección.

Year

Predictor variables

AIC

∆AIC AIC Wgt

2008

ψ(white tailed deer), p(vegetation)

233.7

0.00 0.24

ψ(puma), p(vegetation)

233.7

0.07 0.23

ψ(pecari), p(vegetation)

234.63 0.93 0.15

Ψ(red brocket deer), p(vegetation)

234.91 1.21 0.13

2012

Ψ(coati * white tailed deer * red brocket deer * pecari), p(vegetation)

Jaguar

300.35 0.00 0.37

Ψ(.), p(vegetation)

301.41 1.06 0.22

Ψ(.), p(trail)

301.41 1.06 0.22

Ψ(.), p(.)

302.92 2.57 0.10

2008

Ψ(pecari), p(vegetation)

234.76 0.00 0.96

Ψ(.), p(type_trail)

241.52 6.76 0.03

Ψ(.), p(vegetation)

241.52 6.76 0.03

Ψ(coati + armadillo + opossum + white tailed deer + red brocket deer + pecari), p(vegetion)

Puma

259.25 24.49 0.00

Ψ(.), p(jaguar)

261.79 27.03 0.00

Ψ(.), p(.)

265.22 30.46 0.00

2012

Ψ(.), p(jaguar)

184.17 0.00 0.45

Ψ(vegetation), p(jaguar)

186.17 2.00 0.16

Ψ(coati * white tailed deer * red brocket deer * pecari), p(jaguar)

186.17 2.00 0.16

Ψ(red brocket deer), p(jaguar)

186.17 2.00 0.16

Ψ(pecari + red brocket deer + white tailed deer), p(jaguar)

190.74 6.00 0.02

Ψ(.), p(.)

reported in other Mexican studies (Hernández–Saint Martin et al., 2013). However, in other parts of their range pumas are predominantly crepuscular with peak activity between 02:00 and 10:00 h (Hernández–Saint Martin et al., 2013) and in a tropical forest in the south of Mexico, pumas are more diurnal and jaguars are nocturnal (De la Torre et al., 2017). Other studies show a negative influence of human activity on puma activity (Chávez, 2010; Foster et al., 2010; Rodríguez–Soto et al., 2013). Pumas have become more nocturnal in order to avoid human contact and as a result of human impact are absent or considered to be endangered in parts of Mexico where they used to be abundant (Chávez, 2010). To ensure their sustainable and long–term survival. We therefore need to understand how best to conserve them in protected areas like EER in the south of Mexico.

Spatial and temporal overlap of felids This camera trapping survey of sympatric jaguar and puma populations in the El Eden found evidence of their coexistence in a relatively small reserve consisting mainly of MSTF in North Eastern Yucatan. Despite the high degree of overlap in both habitat and resource–use, there were some differences in peak activity times and association with other species, including a jaguar evasion strategy by pumas. This has also been seen in their sympatric populations in tropical areas (Scognamillo et al., 2003; Di Bitetti et al., 2010) and between other coexisting large felids (Ramesh et al., 2012). The differences between habitat use over the four years in EER also suggests flexibility in how they use shared resources, and includes changes in habitat–use following disturbances

Ávila–Nájera et al.

Jaguar

Puma 2008

2010

2011

2012

Fig. 2. Activity patterns of jaguars (Panthera onca) and pumas (Puma concolor) based on camera trap records from the El Eden Ecological Reserve, Quintana Roo, Mexico, plotted by study year (2008, 2010, 2011 and 2012). Fig. 2. Patrones de actividad del jaguar (Panthera onca) y el puma (Puma concolor) basada en registros de cámaras en la Reserva Ecológica El Edén, en Quintana Roo, México por año de estudio (2008, 2010, 2011 y 2012).

Animal Biodiversity and Conservation 43.1 (2020)

such as Hurricane Karl in September 2010 and a fire around the reserve in May 2011. Occupancy model The environmental variables that influence the presence of the jaguar and its resource–use patterns vary across its range. In Belize, Davis et al. (2010) found that jaguars were positively associated with canopy height, length of trails and the presence of small birds and large mammals. However, in the El Eden, the year of study had a significant impact on all the models tested. Sightings of all species in the El Eden decreased following the disturbances in 2010 and 2011, and may in part explain this effect. However in The Reserve, distance to water bodies did not affect jaguar or puma occurrence because water is available throughout the year and is not a limiting factor, unlike in other parts of their range (Davis et al., 2010). In addition, secondary forest was positively associated with jaguar sightings. This is probably because many of the roads are in the secondary forest, where jaguars can move more easily and hence find prey, as seen by Davis et al. (2010) in Belize. Another strong predictor of puma and jaguar presence in the study was the activity pattern and spatial overlap of prey species (Ávila–Nájera et al., 2016). Similarly important prey occur throughout the jaguar’s geographic range (Ceballos et al., 2005; Chávez et al., 2007; Davis et al., 2010; Harmsen et al., 2010; Romero–Muñoz et al., 2010). In the El Eden, the most frequently recorded species in the camera traps during the study, for example in 2011, were humans (unpublished data), and human activity was a significant negative factor in predicting jaguar and puma sightings, agreeing with similar findings by Davis et al. (2010). Many researchers have noted the negative impact of humans on the presence of jaguars, and there is evidence of significant hunting and poaching in the Calakmul region of the Yucatan (Ceballos et al., 2005). In addition, subsistence hunting depletes prey availablity because hunters favour the same species as the felids (Chávez, et al 2007; Ávila–Nájera et al., 2011; Foster et al., 2014). Predators select habitats as a result of the complex interactions between factors including their population density and the presence of other predators, as well as the abundance and diversity of prey and the level of human activity (Hojnowski et al., 2012; Foster et al., 2014). In EER, slight differences in peak activity time, and dietary preferences (Ávila–Nájera et al., 2018a, 2018b) facilitate the coexistence of two large predators despite the high overlap in their resource use. Studies in other ecosystems have also reported differential habitat use by coexisiting predators (Jonson et al., 1980; Davis et al., 2010; Sollmann et al., 2012) along with a positive dependence upon each other (Gutierrez–González and López–González, 2017). However, this has not been previously reported for these felids in Mexico, which generally show complete overlap in their habitat–use and diet in dry deciduous tropical forest in western Mexico (Núñez et al., 2002).

None of the variables used in the occupancy models were able to predict the presence of jaguar in the reserve. This may suggest that they use it as a corridor to travel between larger reserves, since they require extensive areas of home range, and the protected natural areas and surrounding areas serve as important biological corridors that encourage biodiversity conservation (Domínguez, 2009). In contrast, in 2008, puma occupancy was dependent on one of its main prey, the collared peccary, although the 2012 model depended on both its copredator (which it avoided) and prey abundance. All the occupancy models tested suggest that pumas are occasional residents in the EER, and that their presence is associated with that of their prey, such as peccaries. However, this remains untested. In conclusion, the factors that allow jaguars and pumas to coexist in EER are the differences in their activity patterns, especially their peak activity times, as well as their diets (the latter tested in a previous investigation within the reservation), as reported in previous studies from similar habitats in this region. However, natural perturbations like hurricanes and fire triggered changes in the habitat use and activity pattern of both felids. This showed that they were able to modify their behaviour and the level of interaction in order to avoid contact with each other. However, several aspects require deeper analysis, such as individual interactions between males and females of the same species or co–predators. Acknowledgements We thank Marco Antonio Lazcano for allowing us access to the facilities at the El Eden, Kathy Cabrero of the Center of Tropical Research, at the Universidad Veracruzana, Erik Torres, Juan Castillo, Alejandro Pacheco, Brady Hollinsgworth and other staff and volunteers at Global Vision International for support with fieldwork at the El Eden Ecological Reserve. The study was funded by the Mexican National Council of Science and Technology (CONACYT) doctoral grant number 211454 awarded to Dulce Maria Avila Najera, and Project Financing Promep grant 54310009 to Cuauhtémoc Chávez (UAM–PTC–333). References Álvarez–Castañeda, S. T., Patton, J. L., 2000. Mamíferos del Noroeste de México II. Centro de Investigaciones Biológicas del Noroeste. La Paz, Baja Californi Sur. Aranda, M., 1994. Importancia de los pecaríes (Tayassu sp.) en la alimentación del jaguar (Panthera onca). Acta Zoologica Mexicana, 62: 11–22. Aranda, M., Sánchez–Cordero, V., 1996. Prey spectra of sympatric jaguar Panthera onca and puma Puma concolor at the Calakmul Biosphere Reserve, Campeche, Mexico. Studies on Neotropical Fauna Environment, 31: 43–45. Astete, S., Marinho–Filho, J., Kajin, M., Penido, G.,

Zimbres, B., Sollmann, R., Jácomo, A. T. A., Torres, N. M., Silveira, L., 2017. Forced neighbours: Coexistence between jaguars and pumas in a harsh environment. Journal of Arid Environments, 146: 27–34. Ávila–Nájera, D. M., Chávez, C., Barreto, M. A., Mendoza, G. D., Pérez–Elizalde, S,. 2016. Patrones de actividad y traslape entre grandes felinos y sus principales presas en el norte de Quintana Roo, México. THERYA 7(3): 439–448. Ávila–Nájera, D. M., Chávez, C., Pérez–Elizalde, S., Guzmán–Plazola, R. A., Mendoza, G. D., Lazcano– Barrero, M. A., 2018b. Ecology of Puma concolor (Carnivora: Felidae) in a Mexican tropical forest: adaptation to environmental disturbances. Journal of Tropical Ecology, 66(1): 78–90. Ávila–Nájera, D. M., Palomares, F., Chávez, C., Tigar, B., Mendoza, G. D., 2018a. Jaguars (Panthera onca) and pumas (Puma concolor) diet in Quintana Roo, Mexico. Animal Biodiversity and Conservation, 41(2): 257–266, https://doi.org/10.32800/abc.2018.41.0257 Ávila–Nájera, D. M., Rosas–Rosas, O., Tarango– Arámbula, L., Martínez–Montoya, J. F., Santoyo– Brito, E., 2011. Conocimiento, uso y valor cultural de seis presas del jaguar (Panthera onca) y su relación con éste, en San Nicolás de los Montes, San Luis Potosí, México. Revista Mexicana de Biodiversidad, 82: 1020–1028. Barrientos, S., Maffei, L., 2000. Radio–telemetría de la hurina Mazama gouazoubira en el campamento Cerro Cortado, Izozog, Santa Cruz, Bolivia. In: Manejo de fauna silvestre en Amazonia y Latinoamerica. Asunción: CITES Paraguay: 369–372 (E. Cabrera, C. Mercolli, R. Resquin, Eds.). Fundación Moises Bertoni, University of Florida. Florida. Carothers, J. H., Jaksic, F. M., 1984. Time as a niche difference: the role of interference competition, Oikos, 42: 403–406. Carrillo, E., 2000. Ecology and conservation of white– lipped peccaries and jaguars in Corcovado National Park, Costa Rica. PhD dissertation, University of Massachusetts, Massachusetts. Ceballos, G., Chávez, C., Zarza, H., Manterola, C., 2005. Ecología y conservación del jaguar en la región de Calakmul. CONABIO. Biodiversitas, 62: 1–7. Chávez, C., 2010. Ecología y conservación del jaguar (Panthera onca) y puma (Puma concolor) en la región de Calakmul y sus implicaciones para la conservación de la Península de Yucatán. PhD thesis, Universidad de Granada, España. Chávez, C., Ceballos, G., Amín, M., 2007. Ecología poblacional del jaguar y sus implicaciones para su conservación en la Península de Yucatán. In: Conservación y Manejo del Jaguar en México, estudios de caso y perspectivas: 91–100 (G. Ceballos, C. Chávez, R. List, H. Zarza, Eds.). CONABIO, UNAM, Alianza WWF, Telcel, Distrito Federal. Conde, D. A., Colchero, F., Zarza, H., Christensen, N. L., Sexton, J. O., Manterola, C., Chávez, C., Rivera, A., Azuara, D., Ceballos, G., 2010. Sex matters: modeling male and female habitat differences for jaguar conservation. Biological Conservation, 143:

Ávila–Nájera et al.

1980–1988. Creel, S., Creel, N. M., 1996. Limitation of African wild dogs by competition with large carnivores. Conservation Biology, 10: 526–538. Davis, L., Kelly, M., Stauffer, D. F., 2010. Carnivore co–existence and habitat use in the Mountain Pine Ridge Forest Reserve, Belize. Animal Conservation, 4:56–65. De la Torre, A., Núñez, J. M., Medellín, R., 2017. Spatial requeriments of jaguar and pumas in Southern Mexico. Mammalian Biology, 84: 52–60. Di Bitetti, M., De Angelo, C., Di blanco, Y., Paviolo, A., 2010. Niche partitioning and species coexistence in Neotropical felid assemblage. Acta Oecologica, 36: 403–412. Dickson, B., Beier, P., 2002. Home–range and habitat selection by adult cougars in Southern California. The Journal of Wildlife Management, 66(4):1235–1245. Domínguez, C. E., 2009. Conectividad biológica y Social, Zonas de Influencia de las Áreas Naturales Protegidas. Serie Conocimiento/Número 5. CONABIO, México. Donadio, E., Buskirk, S. W., 2006. Flight behavior of guanacos and vicunas in areas of western Argentina with and without poaching. Biological Conservation, 127: 139–145. Durant, S., 1998. Competition refuges and coexistence: an example from Serengeti carnivores. Journal of Animal Ecology, 67: 370–386. Faller, J. C., Chávez, C., Johnson, S., Ceballos, G., 2007. Densidad de la población de jaguares en el Noreste de la Península de Yucatán. In: Conservación y manejo del Jaguar en México estudios de caso y perspectivas. México: 111–121 (G. Ceballos, C. Chávez, R. List, H. Zarza, Eds.). Alianza WWF, Telcel, CONABIO, CONANP, EcoCiencias S.C., Ciudad de México. Fedriani, J., Palomares, F., Delibes, M., 1999. Niche relations among three sympatric Mediterranean carnivores. Oecologia, 121: 138–148. Foster, R., Harmsen, B., Doncaster, P., 2010a. Habitat use by sympatric jaguar and puma acroos a gradient of human disturbance in Belize. Biotropica, doi:10.1111/j.1744–7429.2010.00641.x – 2010b. The food habits of jaguars and pumas across a gradient of human disturbance. Journal of Zoology, 280: 309–318. Foster, R., Harmsen, B. J, MacDonald, D., Collins, J., Urbina, Y., Garcia, R., Doncaster, P., 2014. Wild meat: a shared resource amongst people and predators. Oryx, 50(1): 63–75. Foster, V., Sarmento, P., Sollman, R., Torres, N., Jácomo, A., Negroes, N., Fonseca, C., Silveira, L., 2013. Jaguar and puma activity patterns and predator–prey interactions in four Brazilian biomes. Biotropica, 45: 373–379. Gittleman, J. L., 1985. Carnivore body size: ecological and taxonomical correlates. Oecologia (Berl), 67: 540–554. Gómez, H., Wallace, B., Ayala, G., Tejada, R., 2005. Dry season activity periods of some Amazonian mammals. Studies on Neotropical Fauna and

Animal Biodiversity and Conservation 43.1 (2020)

Environment, 40: 91–95. Gutiérrez–González, C., López–González, C., 2017. Jaguar interactions with pumas and prey at the northern edge of jaguars´range. Peer J., 5: e2886, doi: 10.7717/peerj.2886 Haines, A., 2006. Is there competition between sympatric jaguar Panthera onca and puma Puma conclor? Acta Zoologica Sinica, 52 (6): 1142–1147. Hanski, I., 1994. Spatial scale, patchiness and population dynamics on land. Philosophical Transactions Royal Society of London. Biological Science, 343: 19–25. Harmsen, B., Foster, R., Silver, S., Ostro, L., Doncaster, P., 2009. Spatial and temporal interactions of sympatric jaguar (Panthera onca) and puma (Puma concolor). Journal of Mammalogy, 90: 612–620. – 2010. Differential Use of Trails by Forest Mammals and the Implications for Camera–Trap Studies: A case study from Belize. Biotropica, 42: 126–133. Hernández–SaintMartin, A., Rosas–Rosas, O., Palacios–Núñez, J., Tarango–Arambula, L., Clemente–Sánchez, F., Hoogesteijn, A., 2013. Activity patterns of jaguar, puma and their potential prey in San Luis Potosi, México. Acta Zoológica Mexicana (n.s.), 29(3): 520–533. Hojnowski, E., Miquelle, D. G., Myslenkov, A. I., Strindberg, S., Smirnov, E. N., Goodrich, J. M., 2012. Why do Amur tigers maintain exclusive home ranges? Relating ungulate seasonal movements to tiger spatial organization in the Russian Far East. Journal of Zoology, 287: 276–282. Johnson, H. D., 1980. The comparison of usage and availability measurements for evaluating resource preference. Ecology, 61: 65–71. Karanth, K. U., Nichols, J. D., 1998. Estimation of Tiger densities in India using photographic captures and recaptures. Ecology, 79: 2852–2862. Karanth, K. U., Sunquist, M. E., 1995. Prey selection by tigers, leopard, and dhole in tropical forest. Journal of Animal Ecology, 64: 439–450. Kelly, M. J., Noss, A. J., Di Bitetti, M. S., Maffei, L., Arispe, L., Paviolo, A., de Angelo, C., Di Blanco, Y., 2008. Estimating puma densities from camera trapping across three study sites: Bolivia, Argentina, and Belize. Journal of Mammalogy, 89: 408–418. Linnell, J. C., Strand, O., 2000. Interference interactions, coexistence and conservation of mammalian carnivores. Diversity and Distribution, 6: 169–176. Lira, T. I., Naranjo, E., 2003. Abundancia, preferencia de hábitat e impacto del ecoturismo sobre el puma y dos de sus presas en la Reserva de la Biósfera El Triunfo, Chiapas, México. Revista Mexicana de Mastozoología, 7: 20–39. Mackenzie, D. I., Nichols, J., Royle, J., Pollock, K., Bailey, L., Hines, J. E., 2006. Occupancy estimation and modeling: Inferring patterns and dynamics of species occurrence. Elsevier Publishing, California. Maffei, L., Cuellar, E., Noss, A., 2004. One thousand jaguars (Panthera onca) in Bolivia's chaco? Camera trapping in the Kaa–Iya National Park. Journal of Zoology London, 262: 295–304. Maffei, L., Noss, A., Silver, S., Kelly, J., 2011. Abundance Density Case Study: Jaguars in the Ameri-

cas. In: Camera Traps in Animal Ecology. Methods and Analyse:119–144 (A. O'Connell, J. Nichols, U. Karanth, Eds.). Springer, Japan. Mcloughlin, D. P., Morris, D. W., Fortin, D., Wal, E., Contasti, A., 2010. Considering ecological dynamics in resource selection functions. Journal of Animal Ecology, 79: 4–12. Monroy–Vilchis, O., Soria–Díaz, L., 2013. Ecología de Puma concolor en la Sierra anchititla. Estado de México, México. Universidad Autónoma del Estado de México. Estado de México. Navarro, C. J., Ramolina, J. F., Pérez, J. J., 2007. El jaguar en Yum Balam y el Norte de Quintana Roo. In: Conservación y manejo del jaguar en México: estudio de caso y perspectivas: 123–132 (G. Ceballos, C. Chávez, R. List, H. Zarza, Eds.). CONABIO, Alianza WWF, Telcel, Universidad Nacional Autónoma de México, Ciudad de México. Novack, A. J., Main, M. B., Sunquist, M. E., Labisky, R. F., 2005. Foraging ecology of jaguar Panthera onca and puma Puma concolor in hunted and non– hunted sites within the Maya Biosphere Reserve, Guatemala. Journal of Zoology, 267(2): 167–178. Núñez, R., Miller, B., Lindzey, F., 2000. Food habits of jaguar and pumas in Jalisco, México. Journal of Zoology. London, 252: 373–379. – 2002. Ecología del jaguar en la reserva de la biosfera Chamela–Cuixmala, Jalisco, México.In: El jaguar en el nuevo milenio: 107–126 (R. A. Medellin, C. Equihua, C. Chetkiewizc, P. G. Craeshaw, A. Rabinowitz, K. H. Redford, J. G. Robinson, E. W. Sanderson, A. B. Taber, Eds.). Universidad Nacional Autónoma de México and Wildlife Conservation Society, Ciudad de México, Mexico. Paviolo, A., Di Blanco, Y., De Angelo, C., Di Bitetti, M., 2009. Protection affects the abundance and activity patterns of puma in the Atlantic forest. Journal of Mammalogy, 90: 963–964. Pianka, E. R., 1973. The structure of lizard communities. Annual Review of Ecology and Systematics. 4: 53–74. Pozo, C., Armijo–Canto, N., Calmé, S. (Eds.), 2011. Riqueza Biológica de Quintana Roo. Un análisis para su conservación, Tomo I. El cColegio de la Frontera Sur (Ecosur), Comisión Nacional para el Conocimiento y Uso de la Biodiversidad (Conabio), Gobierno del Estado de Quintana Roo y Programa de Pequeñas Donaciones. Ciudadd e México. Rabinowitz, A. R., Nottingham, B. G., 1986. Ecology and behavior of the jaguar (Panthera onca) in Belize, Central America. Journal of Zoology, 210: 149–159. Ramesh, T., Kalle, R., Sankar, K., Qureshi, Q., 2012. Spatio–temporal partitioning among large carnivores in relation to major prey species in Western Ghats. Journal of Zoology, 287: 268–275. Rodríguez–Soto, C., Hernández–Téllez, M., Montoy– Vilchis, O., 2013. Distribución y uso del hábitat de Puma concolor en la Sierra Nanchititla, México. In: Ecología de Puma concolor en la Sierra Nanchititla, México: 47–63 (O. Monrroy–Vilcis, L. Soria–Díaz, Eds.). Universidad Autónoma del Estado de México, Estado de México.

Romero–Muñoz, A., Maffei, L., Cuéllar, E., Noss, J. A., 2010. Temporal separation between jaguar and puma in the dry forests of southern Bolivia. Journal of Tropical Ecology, 26: 303–311. Sanderson, E. W., Chetkiewicz, C. L. B., Medellin, R. A., Rabinowitz, A. Redford., K. H. Robinson, J. G., Tober, A. B., 2002. Un análisis del estado de conservación y distribución de los jaguares a través de sus áreas de distribución. In: El jaguar en el nuevo milenio: 551–600 (R. A. Medellin, C. Equihua, C. Chetkiewizc, P. G. Craeshaw, A. Rabinowitz, K. H. Redford, J. G. Robinson, E. W. Sanderson, A. B. Taber, Eds.). Universidad Nacional Autónoma de México and Wildlife Conservation Society, Ciudad de México, Mexico. Schultz, G., 2003. Structure and diversity of the forests at the El Eden Ecological Reserve. In: The lowland Maya area. Three millennia at the human–wildland interface: 91–114 (A. Gómez–Pompa, M. F. Allen, S. Fedick, J. J. Jiménez–Osorio, Eds.). The Haworth Press, Binghamton, New York. Scognamillo, D. M., Maxit, I. E., Sunquist, M., Polisar, J., 2003. Coexistence of Jaguar (Panthera onca) and puma (Puma concolor) in a mosaic landsca-

Ávila–Nájera et al.

pe in the Venezuelan llanos. Journal of Zoology London, 259: 269–279. Sollmann, R., Malzoni–Furtado, M., Hofer, H., Jácomo, A. T. A., Mundim, N., Silveira, L., 2012. Using occupancy models to investigate space partitioning between two sympatric large predators, the jaguar and puma in central Brazil. Mammalian Biology, 77: 41–46. Taber, A. B., Novaro, A. J., Neris, N., Colman, F. H., 1997. The food habits of sympatric jaguar and puma in the Paraguayan Chaco. Biotropica, 29: 204–213. Woodroffe, R., 2001. Strategies for carnivore conservation: Lessons from contemporary extinctions. In: Carnivore Conservation (J. L. Gittleman, R. K. Wayne, D. W. Macdonald, S. M. Funk, Eds.). Cambridge University Press, Cambridge. Zarza, H., Chávez, C., Ceballos, G., 2007. Uso de hábitat del jaguar a escala regional en un paisaje dominado por actividades humanas en el Sur de la Península de Yucatán. In: Conservación y manejo del jaguar en México: estudios de caso y perspectivas: 101–110 (G. Ceballos, C. Chávez, R. List, H. Zarza, Eds.). CONABIO, Alianza WWF, Telcel, UNAM, Ciudad de México.

Animal Biodiversity and Conservation 43.1 (2020)

Comparison of 12 DNA extraction kits for vertebrate samples I. Martincová, T. Aghová

Martincová, I., Aghová, T., 2020. Comparison of 12 DNA extraction kits for vertebrate samples. Animal Biodiversity and Conservation, 43.1: 67–77, Doi: https://doi.org/10.32800/abc.2020.43.0067 Abstract Comparison of 12 DNA extraction kits for vertebrate samples. Obtaining high quality DNA extractions is a crucial step for molecular biology research projects. At present, numerous protocols are available for vertebrate tissue extractions. In the present study we compared eleven column–based protocols and one HotSHOT protocol using similar conditions (i.e., type of sample, weight of starting material). We evaluated time of extraction, quality and quantity of DNA yield, and price of extraction for a single sample. Based on our analysis, the most successful kits for producing DNA with the highest concentration and purity are the JetQuick® Genomic DNA Purification Kit (Genomed) and the NucleoSpin® Tissue (Macherey–Nagel). Nevertheless, it is highly recommended to test various extraction kits with specific samples to find the optimal kit in all aspects of time, quality and cost for a particular project. Key words: DNA isolation, DNA concentration, DNA purity, Price of extraction kit Resumen Comparación de 12 kits de extracción de ADN de muestras de vertebrados. Obtener extracciones de ADN de buena calidad es un paso crucial para los proyectos de investigación del ámbito de la biología molecular. En la actualidad, existen numerosos protocolos para la extracción en tejidos de vertebrados. En el presente estudio comparamos 11 protocolos de extracción por columnas y un protocolo HotSHOT utilizando condiciones parecidas (como el tipo de muestra o el peso del material inicial). Evaluamos el tiempo de extracción, la calidad y la cantidad de ADN obtenido y el precio de la extracción de una única muestra. Según nuestro análisis, los kits que dieron mejores resultados para producir ADN con la mayor concentración y pureza son JetQuick® Genomic DNA Purification Kit (Genomed) y el NucleoSpin® Tissue (Macherey–Nagel). No obstante, se recomienda encarecidamente probar varios kits de extracción con muestras específicas para encontrar el que sea mejor en cuanto al tiempo, la calidad y el costo en relación con un proyecto concreto. Palabras clave: Aislamiento de ADN, Concentración de ADN, Pureza del ADN, Precio del kit de extracción Received: 01 III 19; Conditional acceptance: 11 VI 19; Final acceptance: 01 X 19 Iva Martincová, Tatiana Aghová, Institute of Vertebrate Biology, Czech Academy of Sciences, Květná 8, 603 65 Brno, Czech Republic.– Iva Martincová, Department of Botany and Zoology, Faculty of Science, Masaryk University, Kotlářska 2, 611 37 Brno, Czech Republic.– Tatiana Aghová, Department of Zoology, National Museum, Václavské náměstí 68, 115 79 Prague, Czech Republic. Corresponding author: T. Aghová. Current address: Center of Oncocytogenomics, Institute of Medical Biochemistry and Laboratory Diagnostics, General University Hospital and First Faculty of Medicine, Charles University in Prague, U Nemocnice 499/2, 128 08, Prague, Czech Republic. E–mail: tatiana.aghova@gmail.com

ISSN: 1578–665 X eISSN: 2014–928 X

Martincová and Aghová

Introduction Recent expansion of new approaches in molecular genetics allows researchers to obtain genetic information from a variety of sources: e.g., hair (Grisedale et al., 2018), faeces (De Barba et al., 2017), urine (Hausknecht et al., 2007), shed feathers (Valsecchi, 1998), eggshells (Lee and Prys–Jones, 2008), and owl pellets (Taberlet and Fumagalli, 1996). Additionally, improved protocols for non–invasive samples enable DNA isolation from environmental samples such as water and soil (Taberlet et al., 1999; Thomsen et al., 2012; Pilliod et al., 2013; Català et al., 2015; Jerde et al., 2019; Stat et al., 2019) and from museum samples (Payne and Sorenson, 2002; Wandeler et al., 2003; Flagstad et al., 2003; Wisely et al., 2004; Hedmark and Ellegren, 2005; Horváth et al., 2005; Harper et al., 2006; Stuart et al., 2006; Ciborowski et al., 2007; Wandeler et al., 2007). Regardless of the source or eventual use of the DNA, the goals of all extraction methods are the same: 1) to release genetic material from its source (fluid, tissue or microbe); 2) to stabilize nucleic acids against degradation; 3) to remove amplification inhibitors; 4) to concentrate the nucleic acid material into an appropriate volume of an aqueous solution compatible with downstream application; and 5) to standardize the methods to support accurate, sensitive and reproducible laboratory assays (Attia et al., 1996; Fox et al., 2007; Hill, 2011; Boesenberg–Smith et al., 2012). The few previous studies that compared DNA extraction methods include: bacterial and fungal communities (e.g. Queipo–Ortuño et al., 2008; Tomaso et al., 2010; Vesty et al., 2017; Rodrigues et al., 2018); mollusc (Der Sarkissian et al., 2017); invertebrate (Kranzfelder et al., 2016; Schiebelhut et al., 2017); and ancient samples and formalin–fixed tissue samples (Rohland and Hofreiter, 2007; Janecka et al., 2015; Gamba et al., 2016). Furthermore, these studies are not comparable, as each experiment employed a different source of genetic material (e.g. different tissue, age of material, storage conditions) and different extraction kits. The extraction of DNA from vertebrate samples is usually less challenging than in the previously mentioned groups. However, the comparison of extraction kits for vertebrate samples is still lacking. Advances in Next Generation Sequencing demand the use of high quality DNA. Therefore, we decided to compare 12 DNA extraction protocols under similar conditions, including: initial weight of isolated sample, equipment, laboratory technician, and measurements of quality and quantity. The goal of this study was to compare the concentration and purity of isolated DNA, and the time and price of single extraction from vertebrate tissue. Material and methods For the purpose of this experiment we used 12 different extraction kits from eight manufacturers. We tried to cover the diversity of extraction kits, including those kits commonly cited in the literature (based on Google Scholar Search, 25th of June 2019). The kits were: High Pure PCR Template Preparation Kit (Roche)

72,200 Google Scholar hits; Dneasy® Blood & Tissue Kit (Qiagen) 46,700 hits; NucleoSpin® Tissue (Macherey–Nagel) 26,800 hits followed by less–known and/or less cited in literature; E.Z.N.A.® MicroElute Genomic DNA Kit (Omega) 128 hits; Quick–gDNATM MiniPrep (Zymo Research) 219 hits; Invisorb® Spin Tissue Mini Kit (Stratec) 381 hits. Beside the DNA concentration and purity, we also assessed the price of kits and hand–on time demanded for single extraction. The eleven extraction kits are based on silica membrane columns and the last extraction is HotSHOT (Truett et al., 2000). The list of used extraction kits: ‹1› E.Z.N.A.® Tissue DNA Kit (Omega); ‹2› High Pure PCR Template Preparation Kit (Roche); ‹3› E.Z.N.A.® MicroElute Genomic DNA Kit (Omega); ‹4› JetQuick® Genomic DNA Purification Kits (Genomed); ‹5› Dneasy® Blood & Tissue Kit (Qiagen); ‹6› E.Z.N.A.® Forensic DNA Kit (Omega); ‹7› NucleoSpin® Tissue (Macherey–Nagel); ‹8› Quick–gDNATM MiniPrep (Zymo Research); ‹9› UltraClean® Tissue & Cells DNA Isolation Kit (MoBio); ‹10› UltraClean® Tissue & Cells DNA Isolation Kit (MoBio) with Proteinase K; ‹11› Invisorb® Spin Tissue Mini Kit (Stratec); and ‹12› HotSHOT (Truett et al., 2000). Starting material We applied extraction protocols to four types of vertebrate samples: finger (phalange), spleen and tail from the domestic house mouse (Mus musculus) and blood samples obtained from grey partridge (Perdix perdix). Quality bias of samples (due origin, age, storage conditions, etc.) was eliminated by using only fresh samples. The tissue was weighed using Kern ABT analytical balances (Kern, ABT 120–5 DNM, resolution 0.1 mg). The average weight of finger was 1.78 mg (median 1.77 mg, 1.00–2.33 mg), spleen: 5.70 mg (median 5.95 mg, 2.08–7.70 mg), tail: 7.75 mg (median 8.00 mg, 2.54–10.30 mg). For the blood samples we used a comparable volume of clot (cca 5 ul). Details are available in table 1s in supplementary material. To confirm the consistency of results obtained with the same protocol, three extraction kits were used: ‹4› JetQuick, ‹2› Roche, and ‹5› Qiagen to extract several samples of the same tissue type (N = 16, 12, 12 respectively, for samples of comparable weight for finger, spleen and tail). Protocol We followed the manufacturers' recommendations for lysis and purification of DNA (table 1). For three protocols ‹8, 9, 10› we used a homogenization step using the MagNA Lyser Instrument (Roche) and bead tubes (provided in Mobio kit). The tissue was mixed using a thermo–shaker until complete lysis. The extent of disruption (in %) was recorded after two hours (more details in table 2s in supplementary material). For protocols ‹1–11› we used the same amount of elution buffer, 100 µl for finger, and 200 µl for all other tissues. The elution step was repeated twice with the same amount of elution buffer (100 µl and 200 µl, respectively).

Animal Biodiversity and Conservation 43.1 (2020)

Finger

Spleen

Tail

Blood

Fig. 1. DNA quality visualisation using 5 µl of DNA on 1 % agarose gel for 30 min with 1 kb ladder. Fig. 1. Visualización de la calidad del ADN utilizando 5 µl de ADN en gel de agarosa al 1 % durante 30 min con 1 kb Ladder.

Extraction protocol ‹12› does not employ silica– based columns. Tissues were incubated in 75 µl of extraction solution (25 mM NaOH, 0.2 mM EDTA, pH12) 95 ºC for 50 min and then the same amount of 40 mM Tris–HCl (pH 5) was added. The final solution was incubated for one hour at 4 ºC (Truett et al., 2000; Reichard et al., 2008). Repeatability of DNA extraction The effect of tissue weight on the final DNA concentration was tested by multiple extractions of samples of a comparable size (ranging from 0.72 mg to 3 mg in weight) using three kits: JetQuick® Genomic DNA Purification Kits (Genomed) ‹4› was used for extraction of 16 samples, ‹2› High Pure PCR Template Preparation Kit (Roche) and ‹5› Dneasy® Blood & Tissue Kit (Qiagen) for extraction of 12 samples each. We reported the relative concentration as the ratio of concentration per 1 mg of weight (c/W). Measuring Primary verification of DNA quality was tested using gel agarose electrophoresis under the following

conditions: 5 µl of DNA in 1 % agarose gel, running for 40 min. The gel was visualized by GenoPlex documentation system and GenoCapture software (fig. 1). The isolates were subsequently assessed for quantity and quality using Quibit® fluormeter and DS–11 Spectrophotometer. We measured the 1 st and 2 nd elution (5µl of DNA for each) using Quibit® fluorometer with Qubit dsDNA BR Assay kit. DNA purity was evaluated using A 260/A280 ratio via DS–11 Spectrophotometer (DeNovix). Following the manufacturers' recommendation, the blank (i.e., water or elution buffer) for each extraction kit was used together with 1 µl of DNA. The A260/ A280 ratio measured with fluorometry should be ~1.8 (Santos et al., 2009). Values higher than 2.0 indicate basic contamination, while values lower than 1.7 relate to acidic contamination of phenol or proteins. Cost The cost per single extraction was calculated for the commercial kits by dividing the cost by the number of extractions. The prices are valid for 2019 in the Czech Republic.

Martincová and Aghová

Table 1. Overview of 12 extraction kits, information about lysis, purification of DNA from manufacturers' protocol, time extraction, concentration of DNA measured at Qubit ® fluorometer, purity ration measured at spectrophotometer and costs in Czech Republic (2019). Lysis: SM, starting material (mg); LB, Lysis buffer (µl, nls, no lysis step); PK, proteinase K (µl); T, temperature (ºC). Purification: BB, binding buffer (µl); I, incubation (ºC for 10'); A, alcohol (µl; E, ethanol; I, isopropanol); EB, extra buffer (µl; HBC, HBC Buffer; IRB, Inhibitor Removal buffer); WB, wash buffer (µl); E, elution (µl); TEB (ºC). Time: L, lysis (h); Ho, Hand–on (min). Concentration and Purity ratio (ng/µl: F, finger; S, spleen; T, tail; B, blood). C, cost for 1 reaction (€). ◊ in protocol was used Proteinase K for sample lysis; * Because of low concentration of DNA (< 10 ng/µl) spectrophotometer can't measuere precise purity ratio; † price from 2016, after this date MoBio stop producing UltraClean® Tissue & Cells DNA Isolation Kit. (For abbreviations of extraction kits, see Material and methods and table 2). Extraction kits Lysis

Purification

SM LB PK T BB I A EB WB1 1 E.Z.N.A.® Tissue DNA Kit – Omega 30 200 25 55 220 70 220 (E) 500 HBC 700 2 High Pure PCR Template Preparation Kit – Roche 25–50 200 40 55 200 70 100 (I) 500 IRB 500 3 E.Z.N.A.® MicroElute Genomic DNA Kit – Omega < 10 200 20 55 220 70 220 (E) 500 HBC 700 4 JetQuick® Genomic DNA Purification Kit – Genomed 10–20 200 20 55 200 70 200 (E) no 500 5 Dneasy® Blood & Tissue Kit – Qiagen < 25 180 20 56 200 56 200 (E) no 500 6 E.Z.N.A.® Forensic DNA Kit -– Omega 30 200 25 55–60 225 60 200 (I) 500 HBC 700 7 NucleoSpin® Tissue – Machery–Nagel < 25 180 25 56 100 70 210 (E) no 500 8 Quick-gDNA TM MiniPrep – Zymo Research < 25 500 nls no no no no 200 9 UltraClean ® Tissue & Cells DNA Isolation Kit – MoBio 1–25 700 nls no no no no no 10 UltraClean ® Tissue & Cells DNA Isolation Kit – MoBio ◊ 1–25 700 20 56 no no no no no 11 Invisorb® Spin Tissue Mini Kit – Stratec 5–40 400 40 52 200 no no no 550 12 HotSHOT NA no no no no no no no no

Visualisation Results were visualized using packages ggplot2 and ggthemes in the R statistical environment (R Team, 2015). The extraction kits are coded and numbered ‹1–12›. A full list of names is available in table 1. The relative DNA concentration was calculated as the ratio of concentration per 1 mg of weight (c/W; table 1; fig. 2). For each extraction kit, four sample types were visualized (finger, spleen, tail and blood). First and second elution of each sample was shown. To provide a lucid comparison of concentrations

WB2

TEB

700 100–200 70 500

200

700 10–50

70 70

500 25–200 65–70 500

200

700

100

600

100

500

> 50

400

550 200 52 no

between various tissues and elution, a final list was created and provided in table 1s in supplementary material. Results Table 1 summarizes results. It contains information about lysis conditions and DNA purification, time of lysis and hands–on time, DNA yield, DNA purity, and the cost per sample obtained from the 12 extraction protocols.

Animal Biodiversity and Conservation 43.1 (2020)

Tabla 1. Resumen de los 12 kits de extracción, información sobre lisis, purificación del ADN del protocolo de los fabricantes, tiempo de extracción, concentración de ADN medida en el fluorímetro Qubit®, índice de pureza medido en un espectrofotómetro en costos en la República Checa (2019). (Para las abreviaturas, véase la cabecera de la tabla en inglés).

Time

Concentration

Purity ratio A260/A280

5.26

21.10

86.40

11.90

2.16*

1.91

1.9

2.01

1.82

.44

11.50

49.60

2.63

2.33*

1.74 1.87

1.76*

3.05

5.23

8.16

93.50

13.30

2.07* 1.61* 1.87

1.68

3.10

3.68

12.00

43.00

10.80

2.07*

2.01 1.89

1.83

2.46

12.40 10.90

49.90

8.72

1.95

1.78 1.81

2.1

4.42

7.02

1.44

25.10

1.70

1.64

1.85* 1.74

1.33*

2.37

13.40 17.90

83.80

8.10

2.02

2.03 1.86

1.95

2.79

0.50

1.15

0.50

–0.99* 0.56* 0.36* 0.62*

2.19

0.50

1.04* 1.06* 0.97*

1.2*

11.46†

0.5

0.50

0.96* 1.01* 1.88* 1.38*

11.46†

6 35

25.00 28.70 87.00 16.40

2.09 2.12 1.87 1.96

3.37

1.19

1.7*

< 1.0

120

4.08

1.41

1.46

1.94* 1.66* 2.02*

Time

Concentration and purity of DNA

The extraction protocols differ in time of lysis and hands–on time. Three of the protocols did not require a lysis step ‹8› Quick–gDNATM MiniPrep (Zymo Research), ‹9› UltraClean® Tissue & Cells DNA Isolation Kit (MoBio) and ‹12› HotSHOT; therefore, the time of extraction was less than 2 hours. The lysis time was an interval from 3 hours (‹1›) to overnight lysis (‹2›, ‹3›, ‹4›). When we compared hands–on time after lysis, the time variability was minimal, ranging from 35 min (‹1›) to 50 min (‹1›, ‹6›). More details are given in table 1.

The absolute and relative DNA concentration from the first and second elution for each tissue is shown in table 1 and figure 2. Overall, the second elution had similar or lower concentration than the first one. The only exception was blood sample isolation using extraction kits: ‹1, ‹2›, ‹4›, ‹5›. The highest DNA yields were obtained using ‹11›, which provided constantly high DNA concentration for all used tissues. Contrary to this, kits ‹8›, ‹9› and ‹10› provided only low–concentration DNA in all cases.

Martincová and Aghová

Spleen

4 3 2 1 0

1 2 3 4 5 6 7 8 9 10 11 12 Tail

Finger

1 2 3 4 5 6 7 8 9 10 11 12 Blood

10 5 0

DNA concentration (ng/µl)

1 2 3 4 5 6 7 8 9 10 11 12 Extraction kit First elution

5 0

1 2 3 4 5 6 7 8 9 10 11 12 Extraction kit Second elution

Fig. 2. Comparison of obtained DNA concentrations between 12 extraction kits. To visualize finger, spleen and tail the relative concentrations were used (c/W; absolute concentration available in table 1). The absolute concentration (ng/µl) was visualized for blood samples. In the case that only the first elution is visualised, the second elution was not provided ‹12›, or the second elution was not measurable on fluorometer (e.g. ‹8›, ‹9›, ‹10›). Fig. 2. Comparación entre las concentraciones del ADN obtenido con 12 kits de extracción. Para visualizar las muestras de dedo, bazo y cola se utilizaron las concentraciones relativas (c/W; concentración absoluta disponible en la tabla 1). La concentración absoluta (ng/µl) se visualizó para muestras de sangre. Cuando solo se visualizó la primera elución, la segunda ‹12› no se proporcionó o no fue mensurable en el fluorímetro (p. ej. ‹8›, ‹9›, ‹10›).

The DNA quality was assessed using agarose gel electrophoresis. Several cases of very low DNA concentration (less than 0.5 ng/µl; ‹8›, ‹9›, ‹10›, ‹12›) produced no visible band. Additionally, the quality/rate of fragmentation in several of the extraction kits was dependent on tissue type. For example, the highest DNA concentration was obtained from tail and this DNA also displayed extensive degradation/fragmentation visible on gel. The results of DNA extraction repeatability are shown in figure 3. ‹4› JetQuick® Genomic DNA Purification Kits (Genomed) and ‹2› High Pure PCR Template Preparation Kit (Roche) showed similar results (fig. 3) with concentrations ranging between 3.73 ng/µl to 3.92 ng/µl and 12.6 ng/µl and 12.7 ng/µl, respectively. Results from ‹5› Dneasy® Blood & Tissue Kit (Qiagen) displayed much higher variance. The minimal concentration of reference extraction (using the same amount of tissue) was 1.74 ng/ul, while the maximal concentration was 31.4 ng/µl.

The results of purity measurements are available in table 1. None of the tested kits showed A260/A280 ratios within interval 1.8–2.0. Therefore, we defined pure DNA to be in interval 1.6–2.1. The best results according to these parameters were achieved using kits ‹4›, ‹5› and ‹7›. DNA concentration from other kits was lower than10 ng/µl, which may have resulted in an inaccurate A260/A280 ratio. Price Table 1 states the cost of DNA extraction. The price for ‹9› and ‹10› is from 2016; after this date MoBio Laboratory no longer produced the UltraClean® Tissue & Cells DNA Isolation Kit. We provide only an approximate price for ‹12› HotSHOT, because this approach uses basic chemicals, which are routinely available in molecular laboratories. The average price for a commercial kit per extraction is 4.12 € (median

Animal Biodiversity and Conservation 43.1 (2020)

DNA concentration (ng/µl)

0 ‹4› Genomed

‹2› Roche Extraction kit

‹5› Qiagen

Fig. 3. Repeatability of DNA extraction using three reference kits: ‹4› JetQuick® Genomic DNA Purification Kits (Genomed), ‹2› High Pure PCR Template Preparation Kit (Roche), and ‹5› Dneasy® Blood & Tissue Kit (Qiagen). Boxplot represent relative DNA concentration accomplished by multiple isolation of DNA (16, 12, 12 samples respectively; more information in table 7s in supplementary material). Extraction kits ‹4› and ‹2› showed similar results, while ‹5› displayed much greater variance. Fig. 3. Repetibilidad de la extracción de ADN utilizando tres kits de referencia: ‹4› JetQuick® Genomic DNA Purification Kits (Genomed), ‹2› High Pure PCR Template Preparation Kit (Roche) y ‹5› Dneasy® Blood & Tissue Kit (Qiagen). El diagrama de caja representa la concentración relativa de ADN alcanzada mediante el aislamiento múltiple de ADN (16, 12 y 12 muestras respectivamente; más información en la tabla 7s del material suplementario). Los kits de extracción ‹4› y ‹2› dieron resultados parecidos, mientras que con el ‹5› la varianza fue mucho mayor.

2.92 €). The price varied from less than 1 € ‹12› to 11.46 € ‹9›,10› per extraction (see table 1). Discussion The DNA quality, quantity and purity have crucial effect for downstream molecular analysis, therefore the methods of DNA extraction should be thoughtfully selected (Sagi et al., 2009). The main goal of this study was to compare DNA extraction (comprising yield, purity, cost and hands–on time) between twelve kits designed for DNA isolation. To make the comparison transparent, we defined three categories (best, average, worst) evaluating pros and cons of each kit in relation to hands–on time, DNA concentration, DNA purity and costs (table 2). Time One of the factors of DNA extraction is time (both the total extraction time and the time of lysis). Presented extraction kits differ in time of lysis, because of the different proteinase effectiveness. In general, longer lysis resulted in extracting higher amounts of DNA from the same amount of starting material (Janecka et al.,

2015). Nevertheless, there are extraction kits ‹1›, ‹3›, which provided 80–100 % disrupted tissue 2 hours after lysis, while proteinases from other kits worked slower and took from 6 hours to overnight to provide results. To improve the yield of DNA and reduce lysis time, it is useful to disrupt tissue with pestles or beads and use extraction kits with effective Proteinase K (e.g. kits ‹1›, ‹3›). There are extraction protocols without silica columns (e.g. ‹12›; Truett et al., 2000; Wang and Storm, 2006; Meeker et al., 2007; Montero–Pau, 2008; Jiang et al., 2014) that are not only fast (2 hours) and cheap but also allow extraction of 96 samples on a plate. Moreover, there are kits that provide super–fast DNA extraction (in less than 10'), such as QuickExtractTM DNA Extraction Solution (Lucigen). This protocol was used in recent studies: e.g., NGS barcoding (Kutty et al., 2018), gene expression (Fernández et al., 2012), and non–invasive seed extraction (Al–Amery et al., 2016). Al–Amery et al. (2016) compared this QuickExtractTM with the CTAB protocol and obtained similar results regarding quality and quantity of DNA. Alternatively, there are extraction kits which use a bead homogenization instead of proteinase lysis (e.g. ‹8›, ‹9›; Guimaraes et al., 2011; Liakopoulos et al., 2014; Mavroidi et al., 2014; Delherbe, 2015; Minogue

Martincová and Aghová

Table 2. Comparison of the time, DNA concentrations, purity and costs between 12 extraction kits: *** short time, high concentration, good purity and good price; ** medium time, concentration, purity and price; * long time, low concentration of DNA, low purity and high price: ◊ in protocol was used Proteinase K for sample lysis. (For details see tables 3s–6s in supplementary material). Tabla 2. Comparación del tiempo, las concentraciones de ADN, índice y los costos de los 12 kits de extracción. *** tiempo breve, concentración elevada, buena pureza y buen precio, ** tiempo, concentración, pureza y precio medios, * tiempo largo, concentración de ADN baja, baja pureza y precio elevado; ◊ en el protocolo se utilizó Proteinasa K para la lisis de la muestra). (Para obtener información más detallada, véanse las tablas 3s–6s del material suplementario). Extraction kit

Time

Concentration of DNA

Purity of DNA

Cost

1 E.Z.N.A.® Tissue DNA Kit 2 High Pure PCR Template Preparation Kit 3 E.Z.N.A.® MicroElute Genomic DNA Kit 4 JetQuick® Genomic DNA Purification Kit 5 Dneasy® Blood & Tissue Kit 6 E.Z.N.A.® Forensic DNA Kit 7 NucleoSpin® Tissue 8 Quick–gDNATM MiniPrep 9 UltraClean® Tissue & Cells DNA Isolation Kit 10 UltraClean® Tissue & Cells DNA Isolation Kit ◊ 11 Invisorb® Spin Tissue Mini Kit 12 HotSHOT

** * * * ** ** * *** *** *** ** ***

*** ** ** *** ** * *** * * * *** *

** ** ** *** *** ** *** * * * ** *

*** ** ** ** * *** ** *** * * * ***

et al., 2015). These kits provide the possibility to extract DNA in less than one hour. Tomaso et al., (2010) compared five kits (QIAamp™ DNA Mini Kit (Qiagen), peqGold™ Tissue DNA Mini Kit (PeqLab), ‹9› UltraClean™ Tissue and Cells DNA Isolation Kit (MoBio), ‹2› DNA Isolation Kit for Cells and Tissues (Roche), and NucleoSpin™ Tissue (Macherey–Nagel) and with exception of ‹9› all of them yielded enough DNA for real–time PCR assay. Measurement (Qubit, NanoDrop) Determining DNA concentration and purity is an important step for downstream applications, such as polymerase chain reaction. Two types of measurements are preferable: fluorometry (Qubit) and spectrophotometry (e.g. NanoDrop, DeNovix). Qubit fluorometer calculates the total amount of DNA in one sample (O’Neill et al., 2011). The spectrophotometric instrument detects all particles that absorb light at 260 nm (DNA, RNA, single or double stranded, proteins, contaminants; O’Neill et al., 2011). Fluorometer and spectrophotometer results are not always correlated. Spectrophotometry measurements are usually higher than Qubit results, indicating that the DNA sample may contain a mixture of double– and single–stranded DNA, contaminants which scatter light, or UV–absorbing materials that are not nucleic acids (O’Neill et al., 2011). To take advantage we used both approaches for DNA quantification, Qubit values to determine concentration of double strand DNA and

spectrophotometer to obtain information about DNA purity. The difference in concentration measurements between Qubit and spectrophotometer could indicate the presence of single strand DNA, RNA, proteins and/ or contaminants (in our case the biggest difference in extraction kit ‹12›). The combination of both approaches provides the most complete and correct information about DNA sample quality (Simbolo et al., 2013) the quality of DNA can vary depending on the source or extraction method applied. Thus a standardized and cost–effective workflow for the qualification of DNA preparations is essential to guarantee interlaboratory reproducible results. The qualification process consists of the quantification of double strand DNA (dsDNA, Qubit measurements as the DNA quantity indication (DNA concentration), and spectrophotometry as the information about the DNA quality (purity ratio A260/A280). Relationship of time, quality and price When selecting an extraction method, many aspects are to be considered. The purity of the nucleic acid in obtained sample, the cost–effectiveness of the procedure, the duration of exposure to dangerous chemicals, the amount of hands–on time, and the related number of required steps must be taken into account (Boesenberg–Smith et al., 2012). If the short–time extraction is needed, kits ‹8–10› or ‹12› should be used. If the high yield of DNA is the main concern, kits ‹1›, ‹4›, ‹7› or ‹11› are suitable. In case of

Animal Biodiversity and Conservation 43.1 (2020)

molecular methods requesting high quality pure DNA, kits ‹4›, ‹5› or ‹7› are preferable. Some laboratories have limited budgets, in which case we recommend kits ‹1›, ‹6›, ‹8›, ‹12›. In case of rare samples that demand good quality and quantity, it is advantageous to use ‹4› or ‹7›. To sort this puzzle out, testing an extraction kit should belong to standard practice in any laboratory. The testing should include at least three different protocols. It is appropriate to optimize the selected extraction kit, e.g. amount of starting material, time of lysis and/or usefulness of performing 2nd elution. When the amount of starting material is larger than recommended, it is suggested in ‹11› protocol to double the amount of lysis buffer and proteinase. If extracting limited and rare samples, then the second elution is recommended because up to 42 % of DNA is still bound to the membrane surface after the first elution (Janecka et al., 2015). Conclusion and perspectives Selection of the best DNA extraction kit depends on many factors: starting material, sample size, number of samples (single column extraction or 96–plate extraction), extraction time, expected concentration and purity (e.g., for microsatellites required DNA quality is lower compared to whole genome sequencing requirements) and price per one extraction. The kit choice can be adjusted according to special conditions of any samples (forensic, micro–samples, blood), price (big difference between kits), or extraction time. If many samples need to be processed, some extraction kits (Qiagen, Stratec, HotSHOT) provide the possibility to extract 96 samples in plates and thus significantly decrease the extraction time. Some kits, such as NucleoSpin (Macherey–Nagel), are forensic quality, treated to prevent DNA contamination. In conclusion, there is no best kit for any particular project, only the best solution. For this reason we strongly recommend trials be performed with 3–5 extraction kits in advance to check which kit provides the best results for specific samples. Acknowledgements This study was fully supported by the Ministry of Culture of the Czech Republic: project NAKI II (DG16P02B038). We would like to thank T. Kralová and L. Ďureje for providing the samples, J. Moravec for discussion about experiment design and J. Šmíd, J. Bryja and A. Ribas about valuable comments to the manuscript. We thank M. M. McDonough for language correction and comments on the earlier version of the manuscript. Montserrat Ferrer, David Buckley and two anonymous referees provided very useful comments on a previous version of the manuscript. References AL–Amery, M., Fukushige, H., Serson, W., Hildebrand, D., 2016. Nondestructive DNA extraction techni-

ques for soybean (Glycine Max) Seeds. Journal of Crop Improvement, 30: 165–175. Attia, M. A. M., Zekri, A. R. N., Goudsmit, J., Boom, R., Khaled, H. M., Mansour, M. T., De Wolf, F., Alam El–Din, H. M., Sol, C. J. A., 1996. Diverse patterns of recognition of hepatitis C virus core and nonstructural antigens by antibodies present in Egyptian cancer patients and blood donors. Journal of Clinical Microbiology, 34: 2665–2669. Boesenberg–Smith, K. A., Pessarakli, M. M., Wolk, D. M., 2012. Assessment of DNA yield and purity: an overlooked detail of PCR troubleshooting. Clinical Microbiology Newsletter, 34: 1–6. Català, S., Pérez–Sierra, A., Abad–Campos, P., 2015. The use of genus–specific amplicon pyrosequencing to assess Phytophthora species diversity using eDNA from soil and water in northern Spain. Plos One, 10: e0119311. Ciborowski, K. L., Consuegra, S., García De Leániz, C., Wang, J., Beaumont, M. A., Jordan, W. C., 2007. Stocking may increase mitochondrial DNA diversity but fails to halt the decline of endangered Atlantic salmon populations. Conservation Genetics, 8: 1355–1367. De Barba, M., Miquel, C., Lobréaux, S., Quenette, P. Y., Swenson, J. E., Taberlet, P., 2017. High‐throughput microsatellite genotyping in ecology: Improved accuracy, efficiency, standardization and success with low–quantity and degraded DNA. Molecular Ecology Resources, 17: 492–507. Delherbe, N., 2015. Hemolymph bacterial community of Pacific Oyster (Crassostrea gigas) in response to long–term hypercapnia. Capstone Pap. Retrieved from https://escholarship.org/uc/item/0m Der Sarkissian, C., Pichereau, V., Dupont, C., Ilsøe, P. C., Perrigault, M., Butler, P., Chauvaud, L., Eiríksson, J., Scourse, J., Paillard, Ch., Orlando, L., 2017. Ancient DNA analysis identifies marine mollusc shells as new metagenomic archives of the past. Molecular Ecology Resources, 17: 835–853. Fernández, J. J., Martínez, R., Andújar, E., Costa, A., 2012. Gene expression profiles in the cerebellum of transgenic mice over expressing the human FMR1 gene with CGG repeats in the normal range. Genetics and Molecular Research, 11: 467–483. Flagstad, Walker, C. W., Vilà, C., Sundqvist, A. K., Fernholm, B., Hufthammer, A. K., Wiig, Koyola, I., Ellegren, H., 2003. Two centuries of the Scandinavian wolf population: patterns of genetic variability and migration during an era of dramatic decline. Molecular Ecology, 12: 869–880. Fox, A. J., Taha, M. K., Vogel, U., 2007. Standardized nonculture techniques recommended for European reference laboratories. FEMS Microbiology Reviews, 31: 84–88. Gamba, C., Hanghøj, K., Gaunitz, C., Alfarhan, A. H., Alquraishi, S. A., Al‐Rasheid, K. A., Bradley, D. G., Orlando, L., 2016. Comparing the performance of three ancient DNA extraction methods for high‐throughput sequencing. Molecular Ecology Resources, 16: 459–469. Grisedale, K. S., Murphy, G. M., Brown, H., Wilson, M. R., Sinha, S. K., 2018. Successful nuclear

DNA profiling of rootless hair shafts: a novel approach. International Journal of Legal Medicine, 132: 107–115. Guimaraes, A. M. S., Vieira, R. F. C., Poletto, R., Vemulapalli, R., Santos, A. P., Moraes, W. De, 2011. A quantitative TaqMan PCR assay for the detection of Mycoplasma suis. Journal of Applied Microbiology, 111: 417–425. Harper, G. L., Maclean, N., Goulson, D., 2006. Analysis of museum specimens suggests extreme genetic drift in the adonis blue butterfly (Polyommatus bellargus). Biological Journal of the Linnean Society, 88: 447–452. Hausknecht, R., Gula, R., Pirga, B., Kuehn, R., 2007. Urine – A source for noninvasive genetic monitoring in wildlife. Molecular Ecology Notes, 7: 208–212. Hedmark, E., Ellegren, H., 2005. Microsatellite genotyping of DNA isolated from claws left on tanned carnivore hides. International Journal of Legal Medicine, 119: 370–373. Hill, C. E., 2011. Nucleic acid isolation: overview of sample preparation methods. In: Molecular microbiology: diagnostic principles and practices, 2nd ed: 119–125 (D. H. Persing, Ed.). ASM Press, Washington, DC. Horváth, M. B., Martínez–Cruz, B., Negro, J. J., Kalmár, L., 2005. An overlooked DNA source for non‐invasive genetic analysis in birds. Journal of Avian Biology, 36: 84–88. Janecka, A., Adamczyk, A., Gasin, A., 2015. Comparison of eight commercially available kits for DNA extraction from formalin–fixed paraffin–embedded tissues. Analytical Biochemistry, 476: 8–10. Jiang, L., Mancuso, M., Lu, Z., Akar, G., Cesarman, E., Erickson, D., 2014. Solar thermal polymerase chain reaction for smartphone–assisted molecular diagnostics. Scientific Reports, 4: 4137. Jerde, C. L., Wilson, E. A., Dressler, T. L., 2019. Measuring global fish species richness with eDNA metabarcoding. Molecular Ecology Resources, 19: 19–22. Kranzfelder, P., Ekrem, T., Stur, E., 2016. Trace DNA from insect skins: a comparison of five extraction protocols and direct PCR on chironomid pupal exuviae. Molecular Ecology Rresources, 16: 353–363. Kutty, S. N., Wang, W., Ang, Y., Tay, Y. C., Ho, J. K. I., Meier, R., 2018. Next–Generation identification tools for Nee Soon freshwater swamp forest, Singapore. The Gardens' Bulletin, Singapore, 70: 155–173. Lee, P. L. M., Prys–Jones, R. P., 2008. Extracting DNA from museum bird eggs, and whole genome amplification of archive DNA. Molecular Ecology Resources, 8: 551–560. Liakopoulos, A., Mavroidi, A., Panopoulou, M., Zachariadou, L., Chatzipanagiotou, S., Spiliopoulou, I., Zerva, L., Petinaki, E., 2014. Molecular characterization of Streptococcus agalactiae from vaginal colonization and neonatal infections: a 4–year multicenter study in Greece. Diagnostic Microbiology and Infectious Disease, 78: 487–490. Mavroidi, A., Liakopoulos, A., Sarrou, S., Miriagou, V., Petinaki, E., 2014. Identification and characterization of genetic structures coding for carbape-

Martincová and Aghová

nemases in enterobacteria from Central Greece. Diagnostic Microbiology and Infectious Disease, 81: 47–49. Meeker, N. D., Hutchinson, S. A., Ho, L., Trede, N. S., 2007. Method for isolation of PCR–ready genomic DNA from zebrafish tissues. Biotechniques, 43: 610–614. Minogue, E., Tuite, N. L., Smith, C. J., Reddington, K., Barry, T., 2015. A rapid culture independent methodology to quantitatively detect and identify common human bacterial pathogens associated with contaminated high purity water. BMC Biotechnology, 15: 6. Montero‐Pau, J., Gómez, A., Muñoz, J., 2008. Application of an inexpensive and high‐throughput genomic DNA extraction method for the molecular ecology of zooplanktonic diapausing eggs. Limnology and Oceanography: Methods, 6: 218–222. O'Neill, M., McPartlin, J., Arthure, K., Riedel, S., McMillan, N. D., 2011. Comparison of the TLDA with the Nanodrop and the reference Qubit system. Journal of Physics: Conference Serie, 307(1): 012047. IOP Publishing. Payne, R. B., Sorenson, M. D., 2002. Museum collections as sources of genetic data. Bonner Zoologische Beitrage, 51: 97–104. Pilliod, D. S., Goldberg, C. S., Arkle, R. S., Waits, L. P., Richardson, J., 2013. Estimating occupancy and abundance of stream amphibians using environmental DNA from filtered water samples. Canadian Journal of Fisheries and Aquatic Sciences, 70: 1123–1130. Queipo–Ortuño, M. I., Tena, F., Colmenero, J. D., Morata, P., 2008. Comparison of seven commercial DNA extraction kits for the recovery of Brucella DNA from spiked human serum samples using real–time PCR. European Journal of Clinical Microbiology and Infectious Diseases, 27: 109–114. R Team Core, 2015. R: A language and environment for statistical computing. R Foundation for Statistical Computing, Vienna, Austria. Reichard, M., Smith, C., Bryja, J., 2008. Seasonal change in the opportunity for sexual selection. Molecular Ecology, 17: 642–651. Rodrigues, P., Venâncio, A., Lima, N., 2018. Toxic reagents and expensive equipment: are they really necessary for the extraction of good quality fungal DNA?. Letters in Applied Microbiology, 66: 32–37. Rohland, N., Hofreiter, M., 2007. Comparison and optimization of ancient DNA extraction. Biotechniques, 42: 343–352. Sagi, N., Monma, K., Ibe, A., Kamata, K., 2009. Comparative evaluation of three different extraction methods for rice (Oryza sativa L.) genomic DNA. Journal of Agricultural and Food Chemistry, 57: 2745–2753. Santos, S., Sá, D., Bastos, E., Guedes–Pinto, H., Gut, I., Gärtner, F., Chaves, R., 2009. An efficient protocol for genomic DNA extraction from formalin–fixed paraffin–embedded tissues. Research in Veterinary Science, 86: 421–426. Schiebelhut, L. M., Abboud, S. S., Gómez Daglio, L. E., Swift, H. F., Dawson, M. N., 2017. A comparison

Animal Biodiversity and Conservation 43.1 (2020)

of DNA extraction methods for high‐throughput DNA analyses. Molecular Ecology Resources, 17: 721–729. Simbolo, M., Gottardi, M., Corbo, V., Fassan, M., Mafficini, A., Malpeli, G., Lawlor, R. T., Scarpa, A., 2013. DNA qualification workflow for next generation sequencing of histopathological samples. Plos One, 8: e62692. Stat, M., John, J., DiBattista, J. D., Newman, S. J., Bunce, M., Harvey, E. S., 2019. Combined use of eDNA metabarcoding and video surveillance for the assessment of fish biodiversity. Conservation Biology, 33: 196–205. Stuart, B. L., Dugan, K. A., Allard, M. W., Kearney, M., 2006. Extraction of nuclear DNA from bone of skeletonized and fluid‐preserved museum specimens. Systematics and Biodiversity, 4: 133–136. Taberlet, P., Fumagalli, L., 1996. Owl pellets as a source of DNA for genetic studies of small mammals. Molecular Ecology, 5: 301–305. Taberlet, P., Waits, L. P., Luikart, G., 1999. Noninvasive genetic sampling: look before you leap. Trends in Ecology and Evolution, 14: 323–327. Thomsen, P. F., Kielgast, J., Iversen, L. L., Wiuf, C., Rasmussen, M., Gilbert, M. T. P., Orlando, L., Willerslev, E., 2012. Monitoring endangered freshwater biodiversity using environmental DNA. Molecular Ecology, 21: 2565–2573. Tomaso, H., Kattar, M., Eickhoff, M., Wernery, U., Al Dahouk, S., Straube, E., 2010. Comparison of

commercial DNA preparation kits for the detection of Brucellae in tissue using quantitative real–time PCR. BMC Infectious Diseases, 10: 100. Truett, G. E., Heeger, P., Mynatt, R. L., Truett, A. A., Walker, J. A., Warman, M. L., 2000. Preparation of PCR–quality mouse genomic DNA with hot sodium hydroxide and tris (HotSHOT). Biotechniques, 29: 52–54. Valsecchi, E., 1998. Tissue boiling: a short‐cut in DNA extraction for large‐scale population screenings. Molecular Ecology, 7: 1243–1245. Vesty, A., Biswas, K., Taylor, M. W., Gear, K., Douglas, R. G., 2017. Evaluating the impact of DNA extraction method on the representation of human oral bacterial and fungal communities. Plos One, 12: e0169877. Wandeler, P., Hoeck, P. E., Keller, L. F., 2007. Back to the future: museum specimens in population genetics. Trends in Ecology and Evolution, 22: 634–642. Wandeler, P., Smith, S., Morin, P. A., Pettifor, R. A., Funk, S. M., 2003. Patterns of nuclear DNA degeneration over time—a case study in historic teeth samples. Molecular Ecology, 12: 1087–1093. Wang, Z., Storm, D. R., 2006. Extraction of DNA from mouse tails. Biotechniques, 41: 410–412. Wisely, S. M., Maldonado, J. E., Fleische, R. C., 2004. A technique for sampling ancient DNA that minimizes damage to museum specimens. Conservation Genetics, 5: 105–107.

Martincová and Aghová

Supplementary material

Table 1s. Details about the DNA extraction. For each extraction kit we note: W, weight of sample (in g); c1, concentration of the first elution (ng/ul); c2, concentration of the second elution (ng/ul); c/W, relative concentration. (◊ in protocol was used Proteinase K for sample lysis)

Finger

Extraction kit

c1/W

c2/W

E.Z.N.A.® Tissue DNA Kit

1.00

5.26

1.63

5.26

1.63

High Pure PCR Template Preparation Kit

1.70

4.44

2.91

2.61

1.71

E.Z.N.A.® MicroElute Genomic DNA Kit

1.70

5.23

2.08

3.08

1.22

JetQuick® Genomic DNA Purification Kit

2.10

3.68

2.48

1.75

1.18

Dneasy® Blood & Tissue Kit

1.55

12.40

12.00

8.00

7.74

E.Z.N.A.® Forensic DNA Kit

2.07

7.02

3.61

3.39

1.74

7 NucleoSpin® Tissue

1.75 13.40 13.00 7.66 7.43

8 Quick–gDNA

1.78

0.50

0.10

0.28

0.06

1.63

0.50

0.31

MiniPrep

UltraClean® Tissue & Cells DNA Isolation Kit

10 UltraClean® Tissue & Cells DNA Isolation Kit ◊

2.33 0.50 NA 0.21 NA

11 Invisorb® Spin Tissue Mini Kit

1.78

12 HotSHOT

1.97 1.19 NA 0.60 NA

25.00

10.30

14.04

5.79

Table 2s. Information about the lysis progress after 2 hours. Percentage represent ratio of lysis tissue (100%. fully dissolved tissue). (◊ in protocol was used Proteinase K for sample lysis) Tabla 2s. Información sobre el progreso de la lisis a las 2 horas. El porcentaje representa el índice de lisis del tejido (100 %. tejido totalmente disuelto). (◊ en el protocolo se utilizó Proteinasa K para la lisis de la muestra).

Lysis after 2 hours

Extraction kit

Finger Spleen Tail Blood

E.Z.N.A.® Tissue DNA Kit

100 %

95 %

100 %

High Pure PCR Template Preparation Kit

85 %

E.Z.N.A.® MicroElute Genomic DNA Kit

100 %

50 %

30 %

100 %

JetQuick® Genomic DNA Purification Kit

90 %

60 %

100 %

Dneasy® Blood & Tissue Kit

100 %

70 %

100 %

E.Z.N.A.® Forensic DNA Kit

100 %

80 %

90 %

100 %

NucleoSpin® Tissue

80 %

60 %

80 %

40 %

100 %

8 Quick–gDNA

MiniPrep

UltraClean® Tissue & Cells DNA Isolation Kit

10 UltraClean® Tissue & Cells DNA Isolation Kit ◊ 11 Invisorb® Spin Tissue Mini Kit

100 %

12 HotSHOT

no lysis step no lysis step short lysis step 100 % no lysis step

Animal Biodiversity and Conservation 43.1 (2020)

iii

Tabla 1s. Información detallada sobre la extracción de ADN. Para cada kit de extracción tomamos nota de: W, peso de la muestra (en g); c1, concentración de la primera elución; c2, concentración de la segunda elución; c/W, concentración relativa. (◊ en el protocolo se utilizó Proteinasa K para la lisis de la muestra).

Spleen W

c1/W c2/W

Tail W (g) L (mm)

Blood

c2 c1/W c2/W

4.70

21.10 6.68 4.49

1.42

6.40

9.00

86.40

4.70

11.50 7.25 2.45

1.54

8.80

7.00

49.60 14.30 5.64 1.63

2.63

3.95

6.30

8.16

4.04 1.30

0.64

4.10

7.00

93.50

4.27 22.80 1.04

13.30

8.88

5.90

12.00 2.15 2.03

0.36

9.70

7.00

43.00

9.98

10.80 17.30

6.60

10.90 9.65 1.65

1.46

7.50

7.00

49.90 35.50 6.65 4.73

8.72

10.80

7.70

1.44

0.06

10.30 7.00

25.10

1.70

1.27

0.50 0.19

8.04 13.50 1.26

6.51

4.43 1.03 2.44 0.63

11.90 14.50

6.50 17.90 17.60 2.75 2.71

11.00 7.00 83.80 26.30 7.62 2.39 8.10 2.99

5.55

1.15

0.09

9.60

7.00

0.50

0.05

0.50

7.58

0.50

8.50

3.00

0.50

0.06

0.50

0.50 0.21 NA

0.07

4.82 0.50 NA 0.10 NA

7.42 7.00 0.50 NA 0.07 NA 0.50 NA

6.00

7.10

28.70 6.82 4.78

1.14

2.08 4.08 NA 1.96 NA

7.00

87.00 32.40 12.25 4.56

16.40

7.54

2.54 7.00 1.41 NA 0.56 NA 1.46 NA

Table 3s. Order of extraction kits based on sum of relative concentration, including the first and the second elution for all four kind of tissue (finger, spleen, tail, blood). (◊ in protocol was used Proteinase K for sample lysis). Tabla 3s. Kits de extracción ordenados en función de la suma de la concentración relativa, incluidas las eluciones primera y segunda para los cuatro tipos de tejido (dedo, bazo, cola y sangre). (◊ en el protocolo se utilizó Proteinasa K para la lisis de la muestra).

Sum of reative concentration (1st + 2nd for finger, spleen, tail and blood)

Extraction kit 11

Invisorb® Spin Tissue Mini Kit

66.50836208

Good

E.Z.N.A.® Tissue DNA Kit

52.7006383

Good

NucleoSpin® Tissue

41.64634366

Good

JetQuick® Genomic DNA Purification Kit

38.89349409

Good

E.Z.N.A.® MicroElute Genomic DNA Kit

28.41650794

Medium

Dneasy® Blood & Tissue Kit

28.1037449

Medium

High Pure PCR Template Preparation Kit

20.52925475

Medium

E.Z.N.A.® Forensic DNA Kit

11.42614579

Medium

12 HotSHOT

4.580717485

Low

8 Quick–gDNATM MiniPrep

1.134375949

Low

0.872711527

Low

UltraClean® Tissue & Cells DNA Isolation Kit

UltraClean® Tissue & Cells DNA Isolation Kit 0.818326715 Low ◊

Martincová and Aghová

Table 4s. Order of extraction kits based on sum of the time: L, lysis (h); HT, hand–on time (min.); TT, total time (min.). (◊ in protocol was used Proteinase K for sample lysis). Tabla 4s. Kits de extracción ordenados en función de la suma del tiempo: L, lisis (h); HT, tiempo de manipulación (min.); TT, tiempo toal (min.). (◊ en el protocolo se utilizó Proteinasa K para la lisis de la muestra). Extraction kit

8 Quick–gDNA

MiniPrep

UltraClean® Tissue & Cells DNA Isolation Kit

10 UltraClean® Tissue & Cells DNA Isolation Kit ◊ 12 HotSHOT

Short

0,5 50 80 Short 0 120 120 Short

E.Z.N.A.® Tissue DNA Kit

230

Medium

E.Z.N.A.® Forensic DNA Kit

290

Medium

Dneasy® Blood & Tissue Kit

340

Medium

395

Medium

11 Invisorb® Spin Tissue Mini Kit 7

NucleoSpin® Tissue

405

Long

High Pure PCR Template Preparation Kit

645

Long

E.Z.N.A.® MicroElute Genomic DNA Kit

650

Long

JetQuick® Genomic DNA Purification Kit

655

Long

Table 5s. List of extraction kits with result of purity measurements obtained by spectrophotometer. Purity ratio A260/A280 for finger, spleen, tail blood. * mean out of reccomended values. (◊ in protocol was used Proteinase K for sample lysis). Tabla 5s. Lista de kits de extracción con los resultados de las mediciones de la pureza obtenidas en el espectrofotómetro: Relación de pureza A260 / A280 para el dedo, el bazo y sangre de la cola: * media fuera de los valores recomendados, (◊ en el protocolo se utilizó Proteinasa K para la lisis de la muestra). Extraction kit

Purity ration

E.Z.N.A.® Tissue DNA Kit

2,16*

1,91

1,9

2,01

Medium

High Pure PCR Template Preparation Kit

2,33*

1,74

1,87

1,76*

Medium

E.Z.N.A.® MicroElute Genomic DNA Kit

2,07*

1,61*

1,87

1,68

Medium

JetQuick® Genomic DNA Purification Kit

2,07*

2,01

1,89

1,83

Good

Dneasy® Blood & Tissue Kit

1,95

1,78

1,81

2,1

Good

E.Z.N.A.® Forensic DNA Kit

1,64

1,85*

1,74

1,33*

Medium

7 NucleoSpin ® Tissue

2,02 2,03 1,86 1,95 Good

8 Quick–gDNATM MiniPrep

–0,99*

0,56*

0,36*

0,62*

Low

1,04*

1,06*

0,97*

1,2*

Low

10 UltraClean® Tissue & Cells DNA Isolation Kit

0,96*

1,01*

1,88*

1,38*

Low

11 Invisorb® Spin Tissue Mini Kit

2,09

2,12

1,87

1,96

Medium

12 HotSHOT

1,7*

1,94*

1,66*

2,02*

Low

UltraClean® Tissue & Cells DNA Isolation Kit ◊

Animal Biodiversity and Conservation 43.1 (2020)

Table 6s. Order of extraction kits based on cost per reaction (in €). (◊ in protocol was used Proteinase K for sample lysis). Tabla 6s. Kits de extracción ordenados en función del costo por reacción (euros). (◊ en el protocolo se utilizó Proteinasa K para la lisis de la muestra.) Extraction kit

Cost for 1 reaction

12 HotSHOT

< 1

Low

1.82

Low

E.Z.N.A.® Tissue DNA Kit

8 Quick–gDNA

2.19

Low

E.Z.N.A.® Forensic DNA Kit

2.37

Low

JetQuick® Genomic DNA Purification Kit

2.46

Medium

NucleoSpin® Tissue

2.79

Medium

High Pure PCR Template Preparation Kit

3.05

Medium

E.Z.N.A.® MicroElute Genomic DNA Kit

3.10

Medium

Invisorb® Spin Tissue Mini Kit

3.37

High

Dneasy® Blood & Tissue Kit

4.42

High

UltraClean® Tissue & Cells DNA Isolation Kit

11.46

High

UltraClean® Tissue & Cells DNA Isolation Kit ◊

11.46 High

MiniPrep

Table 7s. List of samples used for repeatability analysis: W, weight (mg); C, concentration (ng/ul). Tabla 7s. Lista de las muestras utilizadas para los análisis de repetibilidad: W, peso (mg); C, concentración (ng/ul). ID

Sample

Roche

1.58 4.31

2 JetQuick 3.18 5

Roche

2.12 7.1

JetQuick 2.09 11.8

Roche

1.69 5.4

JetQuick 2.84 12.6

Roche

1.63 7.61

JetQuick 2.17 4.23

Roche

1.5 6.59

JetQuick 2.44 5.53

Roche

2.57 3.92

JetQuick 2.01 3.82

Roche

1.9 6.28

JetQuick 1.97 5.32

Roche

1.56 12.7

JetQuick 1.74 6.65

Qiagen

1.9 3.01

10 JetQuick 2.42 7.47

Qiagen

2.22 5.79

JetQuick 2.07 10.9

Qiagen

1.53 11.1

12 JetQuick 3 7.34

Qiagen

1.92 1.74

13 JetQuick 2.47 6.53

Qiagen

1.74 7.61

14 JetQuick 2.5 4.49

Qiagen

0.79 14.9

15 JetQuick 2.48 7.23

Qiagen

0.9 24.1

16 JetQuick 2.69 4.2

Qiagen

0.97 25.9

Roche

2.02 7.29

Qiagen

0.97 30.3

Roche

1.95 8.13

Qiagen

0.98 31.4

Qiagen

0.72 28.4

Qiagen

1.32 2.78

ID 1

Sample

JetQuick 1.98 3.73

Roche

2.09 7.36

Roche

2.11 7.56

Martincovรก and Aghovรก

Animal Biodiversity and Conservation 43.1 (2020)

Does biogeography need species? Ş. Procheş

Procheş, Ş., 2020. Does biogeography need species? Animal Biodiversity and Conservation, 43.1: 79–87, Doi: https://doi.org/10.32800/abc.2020.43.0079 Abstract Does biogeography need species? The non–equivalence of species defined using different species concepts has recently been highlighted as a serious impediment for conservation efforts. The question arises then, to what extent biogeographical studies, and especially macroecological studies, might also be hampered by the numerous problems pertaining to multi–species datasets. An examination of what is meant by species across spatial scales reveals an important discontinuity. Over and above the much–debated species concepts the word 'species' describes, in fact, two distinct ideas. One, applicable at the local scale, is critical in a community ecology context. The second refers to non–equivalent units in the global inventory of biodiversity, useful for reference purpose, but problematic where analysis is concerned. The majority of biogeographical studies are in fact relevant to those intermediate spatial scales where neither meaning truly applies. Multi–species lineages that are comparable in one or another respect (such as equal–age lineages and similar–range lineages) are probably more accurate units for testing biogeographical hypotheses. Key words: Equal–age lineages, Global scale, Higher taxa, Local scale, Range–defined lineages, Species concepts Resumen ¿La biogeografía necesita especies? Recientemente se ha señalado que la falta de equivalencia de las especies definidas utilizando diferentes conceptos de especie es un grave obstáculo para las iniciativas de conservación. La pregunta que se plantea es hasta qué punto los estudios biogeográficos y, en especial los macroecológicos, podrían verse también perjudicados por los numerosos problemas relacionados con las bases de datos de múltiples especies. Si se analiza lo que se entiende por especie en distintas escalas espaciales se observa una importante falta de uniformidad. Al margen de los conceptos de especie ampliamente debatidos, el término “especie” describe, de hecho, dos ideas diferentes. La primera, aplicable a escala local, es fundamental en el contexto de la ecología comunitaria. La segunda hace referencia a unidades no equivalentes del inventario mundial de biodiversidad, útil con fines de consulta, pero problemática en lo que respecta al análisis. En realidad, la mayor parte de los estudios biogeográficos se realizan en escalas espaciales intermedias, en las que ninguna de las dos ideas es verdaderamente válida. Los linajes de múltiples especies que son comparables en un sentido y otro (como los linajes de la misma edad y los linajes de rango parecido) son unidades probablemente más precisas para comprobar las hipótesis biogeográficas. Palabras clave: Linajes de la misma edad, Escala mundial, Taxones superiores, Escala local, Linajes definidos por el rango, Conceptos de especie Received: 08 II 19; Conditional acceptance: 28 III 19; Final acceptance: 31 X 19 Şerban Procheş, Centre for Functional Biodiversity and Discipline of Geography, Westville Campus, University of KwaZulu–Natal, PB X54001, Durban 4000, South Africa. Corresponding author: Şerban Procheş. E–mail: setapion@gmail.com ORCID: http://orcid.org/0000–0002–3415–6930

ISSN: 1578–665 X eISSN: 2014–928 X

Procheş

Introduction In the language of ecologists, biogeographers and conservationists alike, 'species' is one of the most commonly used words, topping key word rankings (Kissling et al., 2015; Chytrý et al., 2017), or even being disqualified from such rankings on this very account (Franklin, 2013). However, not everyone using the word 'species' means the same thing, and it is often unclear exactly who means what. A variety of species concepts are employed by taxonomists when defining species and the results of classification efforts based on different concepts can differ widely (Wheeler and Meier, 2000). This is of particular concern to conservation biologists. Regarding the prioritisation of species, an alarm bell has recently been rung (Garnett and Christidis, 2017) calling for some form of standardisation in terms of the species concepts employed. This call is in fact echoing older concerns (Avise and Mitchell, 2007). When prioritising areas rather than species, concern over the use of species–based measures is shown by increasing supplementation with data on phylogenetic relatedness and functional diversity (e.g. Pollock et al., 2017). Biogeography feeds both data and analysis methods to conservation science, and biogeographers need to stay alert to any concerns raised by conservation scientists as to the validity or reliability of such inputs. Beyond conservation, the theoretical understanding of biogeography depends heavily on what species are understood to be, and on how species concepts are applied. While it has been acknowledged that species need geography (de Queiroz, 2007), it has been considered a given that (bio) geography needs species, and perhaps such a statement needs to be qualified. The literature on species concepts is substantial (Wheeler and Meier, 2000; Coyne and Orr, 2004; de Queiroz, 2007; Sangster, 2014). The question remains though whether the differences between such concepts actually encapsulate the deepest divide in the discussion over what species (should) mean. It may be useful to take a step back and revisit where the idea of species originates, to what extent contemporary usage is compatible with those initial intentions, and whether those intentions in turn are compatible with the questions contemporary biogeography is aiming to answer. I will attempt this here, before considering whether there are any alternatives to species in biogeographical studies. A need for order There is little doubt that humans have a basic and profound need to classify living beings, as with all other things —and that the need to describe life forms as species has its roots in our cognitive processes (Kunz, 2012). Even details such as taxonomic typification (Witteveen, 2015) have clear psychological justifications. If this is the case, one can of course question how real such entities are. The widespread explanation that species have a real existence whe-

reas higher taxa are artificial human constructs is presently being eroded at both ends, with suggestions that neither are species particularly natural (Mischler, 2010; Kunz, 2012; Slater, 2016), nor are e.g. genera substantially more artificial (Humphreys and Linder, 2009; Barraclough and Humphreys, 2015). A recent contribution (Barrowclough et al., 2016) suggests that in fact, at the global scale, subspecies may hold more predictive power than species. Nevertheless, in order for species–based approaches to continue existing for centuries in both their folk and scientific forms, species must presumably have some predictive value (Andersson, 1990). Species across scales To understand the origin of the species idea, one needs to refer to the local scale. Among traditional societies, taxonomic knowledge is often detailed, and there is a good match between what scientists and hunter–gatherers perceive as different species within local ecological communities —in the case of both higher plants and tetrapod vertebrates (Berlin, 1992). Those studying biotic assemblages first hand, whether scientifically or otherwise, primarily apply their species concepts at the local scale, where species are represented by populations. Even at this level, morphological and behavioural variation within species exists, but key characteristics of what makes species different (look different, occupy different niches, behave differently, do not interbreed; Chambers, 2012) are seldom contradicted across fine spatial and temporal scales. Difficulties start to arise when geographic variation kicks in. How broad an area needs to be for variation to be described as 'geographic' is a function of species’ ability to disperse and thus keep the gene flow going. In snails or fruit flies, it may be a couple of kilometres (Cowie and Holland, 2008). In some waders or birds of prey, the world is hardly enough (Procheş and Ramdhani, 2013). In these well–dispersed groups, there can be substantial gene flow between populations across several or all continents, and sometimes only a couple of isolated islands have populations showing significant morphological discrepancies. But in most groups, over scales broad enough to result in reduced gene flow, new traits and combinations thereof appear. It is at this point that the issue of delimiting species arises, and with it the matter of employing one or another species concept. Without molecular data, one commonly used guideline is to look at the level of variation that allows two species to co–exist without or with minimal hybridisation, and assume that non–co–occurring populations showing comparable differences are also distinct species (Sangster, 2014). Where molecular methods are employed, decision processes can be similar. Quite often, an important criterion here is reciprocal monophyly, even though a large proportion of all accepted species are in fact paraphyletic (Ross, 2014). Some molecular studies show patterns scarcely

Animal Biodiversity and Conservation 43.1 (2020)

Table 1. Do species concepts make sense when employed at the local and global scale? An evaluation using criteria from Coyne and Orr (2004). Tabla 1. ¿Tienen sentido los conceptos de especie cuando se utilizan a escala local y mundial? Una evaluación utilizando los criterios de Coyne y Orr (2004).

Criteria (species concepts) Local scale

Global scale

Interbreeding (biological species) Different species are isolated,

with occasional hybridisation.

Often untestable in nature; many closely related species pairs never meet. Some isolating

Yes (almost always)

barriers may not function in the same way

across geographic space.

Mostly unknown

Genetic or phenotypic cohesion (genotypic / cohesion / recognition species) Cohesion is maintained within

Levels of cohesion vary with distance;

species by interbreeding and same

the relevant variables vary in ways which can be

environmental pressures.

continuous or multimodal.

Mostly yes

Variable

Evolutionary cohesion (ecological / evolutionary species) Maintained through ecological niche,

Different populations have different evolutionary

defined at ecosystem level; essentially

trajectories, depending on chance and different

species are represented by populations

local environments.

which have clearly described evolutionary

Mostly not

trajectories. Yes (almost always) Evolutionary history (phylogenetic species) A species has a common history,

In most species there is a common evolutionary

except for occasional in situ

history to the exclusion of all other organisms,

speciation events.

albeit a large proportion of all species are

Yes (almost always)

paraphyletic.

matching what one sees based on morphology, and the matter of revising species delimitation can then be deferred. Often though, molecular studies result in the description of new cryptic species, that are hard to recognise based on morphology alone. Such descriptions have sometimes been hailed as great victories for conservation, only to be re–assessed soon thereafter (Morrison et al., 2009). The global species inventory and the northern bias in taxonomy The species thus described based on morphology and, increasingly, on molecular studies then take their place in a global list of species, which is widely perceived to represent the most reliable measure of

Mostly yes, but often not

global biodiversity, and the estimation of which has become an important topic in itself (Costello et al., 2013). It can be argued, however, that given the lack of uniformity in the methods used, the global number of species represents a better measure of the effort we have put in subdividing biodiversity than a measure of the biodiversity itself. According to most, if not all, major species concepts, it is easier to speak of species at the local scale than globally (table 1). To explain the current global–scale use of species, it may be useful to review some potential biases in research relevant to the use of species. One such bias is the group of organisms under study. It is certainly easier to speak of species in the case of vertebrates than in the case of plants. It is even more complicated in the case of bacteria, algae or fungi, where reproductive isolation is broken more

Procheş

Table 2. The logical validity of adding up species to produce values for Whittaker's (1977) levels of biodiversity measurement, with one level added here (global). Tabla 2. Validez lógica de sumar especies para producir los valores que permitan calcular los niveles de biodiversidad de Whittaker (1977), con un nivel añadido (el mundial).

Diversity level 'Point' diversity Fully valid while noting rare events such as Alpha diversity (local) hybridisation or the accidental occurrence of individuals from elsewhere within or even outside region. Beta diversity (local turnover) Largely valid, except for events such as those listed 'Pattern' diversity (within beta patterns) above, and replacement by closely related species Gamma diversity (intermediate) in different habitat types. These species pairs do not typically co–occur and their acceptance as different species is down to the species concepts employed. Delta diversity (within region turnover) Potentially valid if counting allopatric species pairs/ Epsilon diversity (regional) swarms as single species. 'Global' diversity Possibly valid in special cases such as taxa with low numbers of species and high species distinctiveness.

often via hybridisation and horizontal transfer of genes (Yakimowski and Rieseberg, 2014; Dudgeon et al., 2017). In most countries and ages, zoological studies have edged botanical ones, not to mention microbiology or mycology, even though animals only represent less than 1 % of the total biomass in most terrestrial ecosystems. This has been attributed to anthropomorphism (Wilson et al., 2007), but being able to delimit species more easily may have also had something to do with it. Second, geographic intraspecific variation needs time to arise. In places where most animals and plants are recent arrivals, variation is minimal (Hewitt, 2004), and organisms can be easily classified over broader areas. Is it then coincidental that the effort of extrapolating the idea of species from local to global took flight in Linnaeus' post–glacial landscapes of Sweden, and most of world’s taxonomists are, to the present day, based in places that used to be under ice at the last glacial maximum? (See patterns in the distribution of taxonomists in Gaston and May, 1992.) Where do species fit in biogeography? Biogeography interacts with the idea of species at multiple levels. Most often the very studies defining species include distribution data and as such qualify as biogeography. Analysing these distributions does in fact lend taxonomic and phylogenetic studies a cutting edge, and thus the biogeographical component often takes precedence. In its macroecological incarnation, biogeography studies multi–species datasets, and more specifically

the geographic distributions of multiple species. These are analysed in relation to other variables, either measured for individuals belonging to those putative species (traits such as body size, trophic group, metabolic rate, etc.) or are, as is the case with the geographic range itself, estimated for the species as a whole (age, origin, total abundance). Often, species richness for a given group becomes a variable onto itself, and can be measured at various scales. The broader the scale, the more likely is an operational geographic unit to harbour several closely related allopatric species, meaning the number of species in such a unit depends on taxonomic treatment. A proposed assessment of how valid the use of species is as relevant to the different levels of biodiversity measurement (Whittacker, 1977) is presented in table 2. Among the variables mentioned above, species richness and geographic range are arguably the most sensitive to potential lumping and splitting exercises. However, functional traits also vary geographically, and the inclusion or exclusion of certain populations may change a species’ standing. A conceptual model of how species usage may have expanded from local to global scale, including some of the biases involved and some potential difficulties, is presented in fig. 1. Alternatives Many of the key patterns in biodiversity and biogeography can already be confirmed after discarding species. Barcoding, even though in some ways a black box, is increasingly used as a species–free

Animal Biodiversity and Conservation 43.1 (2020)

Human need for classifying nearby world objects

Intuitive concept

Long history of usage Extensive data available

Local scale origin Lack of consistent definition /usage at larger scales Instability

Potential for inaccuracies

Key unit for biodiversity measurement

Key unit for conservation

Key unit for biodiversity mapping

Fig. 1. A conceptual model tracing the species idea from its inception (human need to classify) to the potential inaccuracies plaguing it today (both in bold boxes). The key uses of species–biodiversity measurement and mapping, and conservation prioritisation, in light grey boxes. Dark grey boxes depict the points where problems intervene. Fig. 1. Modelo conceptual para hacer un seguimiento de la idea de especie desde su concepción (necesidad de los humanos de clasificar) hasta la actualidad, con las posibles inexactitudes que plantea (ambas en cuadros negros). Los principales usos de las especies: medición y cartografía de la biodiversidad y establecimiento de prioridades de conservación se indican en los cuadros gris claro. En los cuadros gris oscuro se indican los puntos problemáticos.

form of biodiversity measurement (Waterton et al., 2013). Species–area curves are paralleled by phylogenetic diversity–area curves —not particularly different except for very fine spatial scales (Procheş et al., 2009; Helmus and Ives, 2012). Species colonisation of islands or new habitats is probably better described as lineage colonisation, and this assists with appreciating the distinction between the number of colonisation events and subsequent diversification. The latitudinal gradient in species richness can be illustrated equally well with multi–species lineages (Davies and Buckley, 2012; Procheş et al., 2015) and with measures of genetic diversity within species (Adams and Hadly, 2012; Araújo and Costa–Pereira, 2013). Simulated genetic and species diversity measures are strongly matched under a variety of modelled scenarios (Vellend, 2005), and this has been tested with real–world data across sites. For example, Cleary et al. (2006) found a 96% match in Bornean butterflies. At broader scales, genetic diversity is also starting to be mapped (Miraldo et al., 2016), and the species–based hotspots of plant diversity have been confirmed using plant phylogenetic diversity, with an 80 % match (Daru et al., 2019). In the continuum between allele diversity and the diversity of broad multi–species lineages,

at any spatial scale other than the local community scale, the need for employing species diversity is decreasing. Insofar as species are road markers to indicate where about in the living world a study is positioned, they are very much still necessary, and the most stringent philosophical approaches questioning the use of species (Stamos, 2004; Kunz, 2012; Slater, 2016) will concede this. Where species become units of analysis, and as such assumed to be somehow equivalent to one another, caution is required. Nevertheless, metrics based on species, and even higher taxa, have been shown to correlate well with arguably more objective measures such as phylogenetic diversity across a variety of studies (e.g. Forest et al., 2007; Pollock et al., 2017). Insofar as species–based metrics remain the easiest to assemble for any given study, one should cautiously employ them —while keeping in mind at all times their shortcomings. A lineage–based future? Most patterns and processes relevant to biogeography can be re–examined using sets of multi–species lineages carefully designed for the key questions

Procheş

Table 3. An outline for using age–based and range–based broad lineages as biodiversity units. Tabla 3. Resumen para utilizar linajes amplios basados en la edad y el rango como unidades de biodiversidad.

Age–based broad lineages

Range–based broad lineages

Methodology

Divide phylogenetic tree into all lineages

Select from phylogenetic tree lineages

present at a given time

at any depth that fit a certain

in the geological past.

geographic pattern.

Possible

Can include extinct lineages or not.

Can be defined as a % match for a

variations.

region, biome, realm or whole world.

Biogeographical Map the diversity and endemism

Map the diversity of such lineages to

application

of such lineages to highlight areas with

highlight most representative areas for

maximal refugial value since age used

a region, biome etc., or, map cosmopolitan

to define them (Davies and Buckley, 2012;

lineages globally to indicate regions

Procheş et al., 2015).

of maximum historical connectivity

(Procheş and Ramdhani, 2013).

Macroecological Compare lineages of the same age to see

Compare lineages with similar

application

how traits affect their survival, distribution,

distribution to understand how traits

diversity (Procheş et al., 2019).

affect the period of time necessary to

achieve that distribution breadth

(Procheş and Ramdhani, 2013).

Other

Suggested for usage in place of taxonomic Can be of ecological interest too,

potential uses

ranks (older lineages, inclusive of

younger ones, in place of higher and lower certain range likely subdivide niche

taxa respectively, Avise and Johns, 1999).

of each study. Two of the several possible ways of defining such lineages are detailed in table 3. One of these is equal–age lineages. Their use as taxa was suggested by Henning (as long ago as 1936), resurfaced with the work of Avise and Johns (1999), and is presently taken quite seriously in fungal systematics (Zhao et al., 2016). The idea of mapping them may be rooted in basal versus derived lineage comparisons (Hawkins et al., 2006), and true equal–age lineages were first mapped by Davies and Buckley (2012; 'local lineages through time'). Such maps can illustrate lineage survival from specific age intervals, and provide a picture of refugial value (Procheş et al., 2015; Padayachee and Procheş, 2016) that cannot be encapsulated by phylogenetic diversity, where recent diversifications can add up to match the values derived from ancient lineages (Forest et al., 2007). For other types of questions, it may be more appropriate to look at equal–range (or at least comparable–range) lineages, from narrow endemic lineages to cosmopolitan ones. Lineage range dynamics has been employed in the understanding of differential survival in refugia (Waldron, 2010), and in the mapping

as lineages occurring throughout a space in a manner that is of regional or global relevance.

of endemism (Rosauer et al., 2015). Age comparisons for lineages defined based on their range was also used to illustrate the effects of body size and dispersal ability (Procheş and Ramdhani, 2013). One should keep in mind that lineages are not entirely free from some of the problems that plague species–based approaches. For example, hybridisation and horizontal transfer of genes that happened long ago mean that any phylogeny from which lineages are derived is not a perfect model. Additionally, the same phylogenetic correction methods used for species will be needed to identify or eliminate patterns derived from lineages being more or less closely related to each other. Nevertheless, lineages present at least two major advantages: they can be made approximately equivalent for the purpose required in a given study, and, if looking at lineages old enough, they come in manageable numbers, meaning one can actually interpret the results with one's own knowledge of biodiversity. Even though the main benefit of using lineages is their equivalence at a scale where species are anything but equivalent, lineages can equally well be used at the local scale (where species do not raise the same

Animal Biodiversity and Conservation 43.1 (2020)

Table 4. A semi–quantitative summary of the data availability, usefulness, feasibility and costs of species in comparison to different lineage types. While current data is certainly best for species, broad lineages are more manageable in terms of number, can be defined quite accurately, and have minimal costs in terms to putting together data sets, as available phylogenies will be in most cases sufficient, especially in the case of age–based lineages. Narrow lineages are numerous, expensive to list (e.g. barcoding) and their definition can vary substantially, but may nevertheless be useful in understanding patterns of plant or animal survival in the geologically recent past, which may be important from a conservation perspective. Tabla 4. Resumen semicuantitativo de la disponibilidad, la utilidad, la viabilidad y los costos de las especies en comparación con diferentes tipos de linajes. Si bien los datos actuales son sin duda más adecuados para las especies, los linajes amplios son más manejables en cuanto al número, se pueden definir con bastante precisión y conllevan un costo mínimo en lo que se refiere a la compilación de conjuntos de datos, puesto que las filogenias disponibles serán suficientes en la mayoría de los casos, en especial en los linajes basados en la edad. Los linajes reducidos son numerosos, caros de identificar (por ejemplo, mediante el código de barras) y su definición puede variar de forma sustancial, aunque pueden resultar útiles para entender los patrones de supervivencia de plantas y animales en el pasado geológico reciente, lo que puede ser importante desde el punto de vista de la conservación.

Broad lineages (age–based)

Numbers Definition accuracy

(range–based)

Narrow leneages Species (species level or below)

* **** *****

*****

****

***

Current data availability

*****

***

Costs of future data acquisition

****

*****

problems) if cross–scale analyses are envisaged (Procheş et al., 2019). Zooming in to a finer phylogenetic scale, within– species lineages have been studied in a variety of ways, and are already contributing a lot to biogeography. Nevertheless, a lot more can be done in terms of finding commonality of patterns. This call, voiced by Hickerson et al. in their 2010 review of phylogeographic research, remains very much relevant today. Searching for such common patterns across lineages may in fact have been more hampered than helped by referring everything to species level. However, one cannot speak of phylogeography and species in one breath, without mentioning speciation. The onset of reproductive isolation in previously cohesive lineages is key to the global accumulation of biodiversity, and this may be the strongest indication that we do need species, and not only to have names for things. This reproductive isolation is highly sensitive to geography. Distance can both increase and decrease the chances of reproductive isolation (Coyne and Orr, 2004; Turelli et al., 2013), and so, at least in that context, biogeography may yet need species after all. Outlook The contemporary understanding of global biodiversity cannot be separated from the Darwinian 'tree of life' model. That Darwin chose to call his most influential work 'The origin of species' may therefore seem ironic

from a present–day perspective. Where branches of the tree of life co–exist locally, they can certainly be called species in that particular local context. When they do not, their specific status remains a moot point. For conservation purposes, one can call them species, while being aware that, in the global context, a different term should be used. Contemporary biogeography incorporates numerous types of studies, and as such is not bound to a particular range of spatial scales. However, in its most stringent definition it should perhaps refer to that precise interval between the local and the global, where neither meaning of the word 'species' truly applies. I started this piece by highlighting conservation–related concerns, and it is probably appropriate to end with conservation too. Species distributions alone are not an ideal predictor of an area's conservation value (Araújo and Williams, 2000). Species occurrence in a given place does not mean that the species will be able to sustain itself if conserved there but extirpated in all other places. On the contrary, high within–species genetic diversity is a strong indication that the species indeed has a long history of occurrence there. It is true that environmental change does not always follow the same routine, and anthropogenic change may indeed act very differently from past changes that resulted in the current spatial pattern in genetic variation. Nevertheless, more often than not, documenting high genetic diversity in an area makes it likely that a lineage will also survive there in the future. Similarly, across broader spatial and phylogenetic scales, the occurrence of multiple ancient

lineages in a region is an indication that the region has served as a refugium through geological time. Often this regional–level survival is decoupled from the specifics of the environments where lineages live, with lineages from diverse environments sharing the same regional or global survival patterns. This allows painting a picture of truly global, multi–environment biodiversity hotspots (Forest et al., 2007; Daru et al., 2019; Igea and Tanentzap, 2019). A lineage focus, whether one refers to narrow genetic lineages or broad multi–species ones, or ideally a combination of the two, is more likely to capture what is needed in conservation than the current species focus, and assembling lineage data sets for this purpose is at least partly achievable (table 4). The word 'lineage' is perhaps ambiguous, and this may have hindered its widespread use in comparative studies, but, as stated here, there is enough ambiguity with 'species' too, even though most conservation–minded people may feel they have an intuitive understanding of what it means. If conservation is to a great extent driven by popular buy–in, and an acceptance of the keywords used is critical, perhaps biogeography can lead an effort to detach itself from a species–centred approach, whether the replacement be lineages, or another. References Adams, R. I., Hadly, E. A., 2012. Genetic diversity within vertebrate species is greater at lower latitudes. Evolutionary Ecology, 27: 133–143. Andersson, L., 1990. The driving force: species concepts and ecology. Taxon, 39: 375–382. Araújo, M. S., Costa–Pereira, R., 2013. Latitudinal gradients in intraspecific ecological diversity. Biology Letters, 9: 20130778. Araújo, M. B., Williams, P. H., 2000. Selecting areas for species persistence using occurrence data. Biological Conservation, 96: 331–345. Avise, J. C., Johns, G. C., 1999. Proposal for a standardized temporal scheme of biological classification for extant species. Proceeding of the National Academy of Sciences of the USA, 96: 7358–7363. Avise, J. C., Mitchell, D., 2007. Time to standardize taxonomies. Systematic Biology, 56, 130–133. Barrowclough, G. F., Cracraft, J., Klicka, J., Zink, R. M., 2016. How many kinds of birds are there and why does it matter? Plos One, 11: e0166307. Barraclough, T. G., Humphreys, A. M., 2015. The evolutionary reality of species and higher taxa in plants: a survey of post–modern opinion and evidence New Phytologist, 207: 291–296. Berlin, B., 1992. Ethnobiological classification: principles of categorization of plants and animals in traditional societies. Princeton: Princeton University Press. Chambers, G., 2012. The species problem: seeking new solutions for philosophers and biologists. Biology, Philosophy, 27: 755–765. Chytrý, M., Chiarucci, A., Pillar, V. D., Pärtel, M., 2017. Applied Vegetation Science enters its 20th year. Applied Vegetation Science, 20: 1–4.

Procheş

Cleary, D. F. R., Fauvelot, C., Genner, M. J., Menken, S. B. J., Mooers, A. Ø., 2006. Parallel responses of species and genetic diversity to El Niño Southern Oscillation–induced environmental destruction. Ecology Letters, 9: 304–310. Costello, M. J., May, R. M., Stork, N. E., 2013. Can we name earth’s species before they go extinct? Science, 339: 413–416. Cowie, R. H., Holland, B. S., 2008. Molecular biogeography and diversification of the endemic terrestrial fauna of the Hawaiian Islands. Philosophical Transactions of the Royal Society B, 363: 3363–3376. Coyne, J. A., Orr, H. A., 2004. Speciation. Sunderland: Sinauer Associates. Daru, B. H., le Roux, P. C., Gopalraj, J., Park, D. S., Holt, B. G., Greve, M., 2019. Spatial overlaps between the global protected areas network and terrestrial hotspots of evolutionary diversity. Global Ecology and Biogeography, 28: 757–766. Davies, T. J., Buckley L. B., 2012. Exploring the phylogenetic history of mammal species richness. Global Ecology and Biogeography, 21: 1096–1105. de Queiroz, K., 2007. Species concepts and species delimitation. Systematic Biology, 56: 879–886. Dudgeon, S., Kübler, J. E., West, J. A., Kamiya, M., Krueger–Hadfield, S. A., 2017. Asexuality and the cryptic species problem. Perspectives in Phycology, 4: 47–59. Forest, F., Greyner, R., Rouget, M., Davies, J. T., Cowling, R. M., Faith, D. P., Balmford, A., Manning, J. C., Procheş, Ş., van der Bank, M., Reeves, G., Hedderson, T. A. J., Savolainen, V., 2007. Preserving the evolutionary potential of floras in the biodiversity hotspots. Nature, 445: 757–760. Franklin, J., 2013. Species distribution models inconservation biogeography: developments and challenges. Diversity and distributions, 19: 1217–1223. Garnett, S. T., Christidis, L., 2017. Taxonomy anarchy hampers conservation. Nature, 546: 25–27. Gaston, K. J., May, R. M., 1992. The taxonomy of taxonomists. Nature, 356: 281–282. Hawkins, B. A., Diniz–Filho, J. A. F., Jaramillo, C. A., Soeller, S. A., 2006. Post–Eocene climate change, niche conservatism, and the latitudinal diversity gradient of New World birds. Journal of Biogeography, 33: 770–780. Igea, J., Tanentzap, A. J., 2019. Multiple macroevolutionary routes to becoming a biodiversity hotspot. Science Advances, 5: eaau8067. Helmus, M. R., Ives, A. R., 2012. Phylogenetic diversity–area curves. Ecology, 93: S31–S43. Hennig, W., 1936. Beziehungen zwischen geographischer Verbreitung und systematischer Gliederung bei einigen Dipterenfamilien: ein Beitrag zum Problem der Gliederung systematischer Kategorien höherer Ordnung. Zoologischer Anzeiger, 116: 161–175. Hewitt, G. M., 2004. Genetic consequences of climatic oscillations in the Quaternary. Philosophical Transactions of the Royal Society of London B, 359: 183–195. Hickerson, M. J., Carstens, B. C., Cavender–Bares,

Animal Biodiversity and Conservation 43.1 (2020)

J., Crandall, K. A., Graham, C. H., Johnson, J. B., Rissler, L., Victoriano, P. F., Yoder, A. D., 2010. Phylogeography's past, present, and future: 10 years after Avise, 2000. Molecular Phylogenetics and Evolution, 54: 291–301. Humphreys, A. M., Linder, H. P., 2009. Concept versus data in delimitation of plant genera. Taxon, 58: 1054–1074. Kissling, W. D., Hardisty, A., García, E. A., Santamaria, M., De Leo, F., Pesole, G., Freyhof, J., Manset, D., Wissel, S., Konijn, J., Los, W., 2015. Towards global interoperability for supporting biodiversity research on essential biodiversity variables. Biodiversity, 16: 99–107. Kunz, W., 2012. Do species exist? Principles of taxonomic classification. Wiley, Weinheim. Miraldo, A., Li, S., Borregaard, M. K., Flórez–Rodríguez, A., Gopalakrishnan, S., Rizvanovic, M., Wang, Z.–H., Rahbek, C., Marske, K. A., Nogués– Bravo, D., 2016. An Anthropocene map of genetic diversity. Science, 6307: 1532–1535. Mishler, B. D., 2010. Species are not uniquely real biological entities. In: Contemporary debates in philosophy of biology: 110–122 (F. J. Ayala, R. Arp Eds.). Blackwell, Oxford. Morrison W. R. III, Lohr, J. L., Duchen, P., Wilches, R., Trujillo, D., Mair, M., Renner, S. S., 2009. The impact of taxonomic change on conservation: Does it kill, can it save, or is it just irrelevant? Biological Conservation, 142: 3201–320. Padayachee, A. L., Procheş, Ş., 2016. Patterns in the diversity and endemism of extant Eocene age lineages across southern Africa. Biological Journal of the Linnean Society, 117: 482–491. Pollock, L. J., Thuiller, W., Jetz, W., 2017. Large conservation gains possible for global biodiversity facets. Nature, 546: 141–144. Procheş, Ş., Forest, F., Veldtman, R., Chown, S. L., Cowling, R. M., Johnson, S. D., Richardson, D. M., Savolainen, V., 2009. Dissecting the plant–insect diversity relationship in the Cape. Molecular Phylogenetics and Evolution, 51: 94–99. Procheş, Ş., Ramdhani, S., 2013. Eighty–three lineages that took over the world: a first review of terrestrial cosmopolitan tetrapods. Journal of Biogeography, 40: 1819–1831. Procheş, Ş., Ramdhani, S., Ali, J. R., Perera, S. J., Gairola, S., 2015. Global hotspots in the present–day distribution of ancient animal and plant lineages. Scientific Reports, 5: 15457. Procheş Ş., Sukri, R. S., Jaafar, S. M., Sieben, E. J. J., Zaini, N. H., Abas, N., Suhaini, S. N., Juhairah Manjul, N. M., Julaihi, M. A. H., Marshall, D. J., Slik, J. W. F., Moodley, D., 2019. Soil niche of rain forest plant lineages: Implications for dominance on a global scale. Journal of Biogeography, 46:

2378–2387. Rosauer, D. F., Catullo, R. A., VanDerWal, J., Moussalli, A., Moritz, C., 2015. Lineage range estimation method reveals fine–scale endemism linked to Pleistocene stability in Australian rainforest herpetofauna. Plos One, 10: e0126274. Ross, H. A., 2014. The incidence of species–level paraphyly in animals: A re–assessment. Molecular Phylogenetics and Evolution, 76: 10–17. Sangster, G., 2014. The application of species criteria in avian taxonomy and its implications for the debate over species concepts. Biological Reviews, 89: 199–214. Slater, M., 2016. Are species real?: an essay on the metaphysics of species. London: Palgrave Macmillan. Stamos, D. N., 2004. The species problem: biological species, ontology, and the metaphysics of biology. Lexington Books, Lanham. Turelli, M., Lipkowitz, J. R., Brandvain, Y., 2013. On the Coyne and Orr–igin of species: effects of intrinsic postzygotic isolation, ecological differentiation, X chromosome size, and sympatry on Drosophila speciation. Evolution, 68: 1176–1187. Vellend, M., 2005. Species diversity and genetic diversity: parallel processes and correlated patterns. The American Naturalist, 166: 199–215. Waldron, A., 2010. Lineages that cheat death: surviving the squeeze on range size. Evolution, 64: 2278–2292. Waterton, C., Ellis, R., Wynne, B., 2013. Barcoding nature: shifting cultures of taxonomy in an age of biodiversity loss. Abingdon: Routledge. Wheeler, Q. D., Meier, R., 2000. Species concepts and phylogenetic theory: a debate. Columbia University Press, New York. Whittaker, R. H., 1977. Evolution of species diversity in land communities: In: Evolutionary Biology, 10: 1–67 (M. K. Hecht, W. C. Steere, B. Wallace, Eds.). Plenum, New York. Wilson, J. R. U., Procheş, Ş., Braschler, B., Dixon, E. S., Richardson, D. M., 2007. The (bio)diversity of science reflects the interests of society. Frontiers in Ecology and the Environment, 5: 409–414. Witteveen, J., 2015. Naming and contingency: the type method of biological taxonomy. Biology, Philosophy, 30: 569–586. Yakimowski, S. B., Rieseberg, L. H., 2014. The role of homoploid hybridization in evolution: a century of studies synthesizing genetics and ecology. American Journal of Botany, 101: 1247–1258. Zhao, R.–L., Zhou, J.–L., Chen, J., Mărgăritescu, S., Sánchez–Ramírez, S., Hyde, K. D. Callac, P., Parra, L. A., Li, G.–J., Moncalvo, J.–M., 2016. Towards standardizing taxonomic ranks using divergence times – a case study for reconstruction of the Agaricus taxonomic system. Fungal Diversity, 78: 239–292.

ProcheĹ&#x;

Animal Biodiversity and Conservation 43.1 (2020)

The evolution of crypsis when pigmentation is physiologically costly G. Moreno–Rueda

Moreno–Rueda, G., 2020. The evolution of crypsis when pigmentation is physiologically costly. Animal Biodiversity and Conservation, 43.1: 89–96, Doi: https://doi.org/10.32800/abc.2020.43.0089 Abstract The evolution of crypsis when pigmentation is physiologically costly. Predation is one of the main selective forces in nature, frequently selecting for crypsis in prey. Visual crypsis usually implies the deposition of pigments in the integument. However, acquisition, synthesis, mobilisation and maintenance of pigments may be physiologically costly. Here, I develop an optimisation model to analyse how pigmentation costs may affect the evolution of crypsis. The model provides a number of predictions that are easy to test empirically. It predicts that imperfect crypsis should be common in the wild, but in such a way that pigmentation is less than what is required to maximise crypsis. Moreover, optimal crypsis should be closer to “maximal” crypsis as predation risk increases and/or pigmentation costs decrease. The model predicts for intraspecific variation in optimal crypsis, depending on the difference in the predation risk or the costs of pigmentation experienced by different individuals. Key words: Predation, Pigmentation, Coloration Resumen La evolución de la cripsis cuando la pigmentación es fisiológicamente costosa. La depredación es una de las principales fuerzas de selección de la naturaleza y a menudo favorece la cripsis en las presas. Por lo general, la cripsis visual implica el depósito de pigmentos en el tegumento. Sin embargo, adquirir, sintetizar, movilizar y mantener los pigmentos puede ser fisiológicamente costoso. En este estudio he elaborado un modelo de optimización para analizar cómo pueden afectar los costes de la pigmentación a la evolución de la cripsis. El modelo proporciona una serie de predicciones que son fáciles de probar empíricamente. Predice que la cripsis imperfecta debería ser común en la naturaleza, pero de manera que la pigmentación fuera inferior a la necesaria para que la cripsis sea máxima. Además, la cripsis óptima debería estar más cerca de la cripsis “máxima” a medida que aumenta el riesgo de depredación o disminuye el coste de la pigmentación. El modelo también predice la existencia de variación intraespecífica en la cripsis óptima, que depende de la diferencia en el riesgo de depredación o de los costes de la pigmentación que soportan los diferentes individuos. Palabras clave: Depredación, Pigmentación, Coloración Received: 26 III 18; Conditional acceptance: 19 VI 18; Final acceptance: 18 XI 19 Gregorio Moreno–Rueda, Depto. de Zoología, Fac. de Ciencias, Univ. de Granada, 18071 Granada, Spain. Corresponding author: G. Moreno–Rueda. E–mail: gmr@ugr.es

ISSN: 1578–665 X eISSN: 2014–928 X

Moreno–Rueda

Introduction Predation is one of the main selective forces in nature, so select for potential prey to have a number of adaptations to avoid being devoured (Ruxton et al., 2004; Caro, 2005; Stevens and Merilaita 2011; Cooper and Blumstein, 2015). One of the main means for potential prey to avoid predation is to evade predator detection, frequently by developing cryptic coloration (Endler, 1978; Ruxton et al., 2004; Caro, 2005; Merilaita and Stevens, 2011; Cuthill, 2019). Crypsis by colour matching involves prey presenting coloration similar to the background where they are most often exposed to predator attacks to avoid being detected by visual predators (Endler, 1990; Stevens and Merilaita, 2009; Merilaita et al., 2017). Indeed, experimental studies have shown that less colour difference between the animal and the background, from the predator’s perspective (i.e. the greater the degree of visual crypsis), equates to less risk of being depredated (Cooper and Allen, 1994; Bond and Kamil, 2002; Stuart–Fox et al., 2003; Cuthill et al., 2005; Cook et al., 2012; Dimitrova and Merilaita, 2014; Merilaita and Dimitrova, 2014; Troscianko et al., 2016; Michalis et al., 2017; Walton and Stevens, 2018). Colour is a complex trait that can be defined according to the HSL model, in which H is the hue, the 'colour' in common parlance, S is the saturation, the purity of the hue, and L is lightness, the quantity of light reflected by the surface (Montgomerie, 2006). To attain visual crypsis, animals generally need to deposit pigments in teguments exposed to the visual system of potential predators in order to develop a tegument coloration resembling the background. To this end, animals must acquire or synthesise one or more pigments that provide a hue of coloration similar to the background colour. Once the proper pigment has been produced, its saturation and lightness will provide the correct concealment against the background. For example, given the proper pigment, maximal crypsis may vary according to the degree of saturation, with crypsis declining when saturation deviates from the optimal level (fig. 1). The level of saturation of a colour is typically determined by the concentration of the pigment providing that colour (Ito and Wakamatsu, 2003; McGraw and Gregory, 2004; McGraw and Wakamatsu, 2004; McGraw et al., 2005; Fargallo et al., 2007; McGraw and Toomey, 2010; Roulin et al., 2013). Nevertheless, the acquisition, synthesis, mobilisation and maintenance of pigmentation may be costly (review in Hill and McGraw, 2006). These costs, though extensively studied in the context of social signals, have received little or no attention in the context of visual crypsis (but see Rodgers et al., 2013). For example, one of the pigments most used in crypsis is melanin (Hubbard et al., 2010; also see Galván et al., 2017), which is also involved in social signals (Jawor and Breitwisch, 2003). Melanin is synthesised from the amino acid tyrosine (with cysteine intervening in the synthesis of pheomelanin) in complex pathways occurring in melanocytes (McGraw, 2006). It is

then stored in organelles called melanosomes and transferred to keratinocytes in the tegument. Melanin–based tegument coloration therefore depends on the concentration of melanin in melanosomes or the concentration of melanosomes in the tegument (Grether et al., 2004). Melanin synthesis may be energetically costly and constrained by a low dietary intake of amino acids (McGraw, 2006). Moreover, the biochemical pathways in which melanin metabolism intervenes and its hormonal regulation may impact on the immune system, oxidative balance, and other physiological processes (Ducrest et al., 2008). But not only pigmented colours may be expensive. For example, white colours, involving unpigmented structural colorations (Prum, 2006), require certain production or maintenance costs (Poston et al., 2005; Moreno–Rueda, 2010; Vágási et al., 2010). Therefore, maintaining a certain coloration in the integument could prove costly, even in the absence of social selection, and these costs may have an effect on the evolution of crypsis, presumably reducing the optimal level of crypsis. In this paper, I hypothesise that the evolution of crypsis is constrained by the costs of pigmentation. Furthermore, if cryptic pigmentation is costly and these costs differ among individuals, then we can expect intra– and interspecific variation in cryptic coloration. Here, I formally present this hypothesis by examining the evolution of crypsis based on a scenario of costly pigmentation and by developing a simple optimisation model. The model Let us consider an animal that, when not preyed upon, reaches its maximal residual fitness, W = 1. Its residual fitness would descent to zero if depredated and the probability of being preyed upon is directly related to the probability of being detected by the predator, i.e. the inverse of the animal's degree of crypsis. Assuming that visual crypsis is achieved solely through one pigment, crypsis may be considered a function of the amount of the pigment deposited in the tegument. Therefore, the animal’s fitness will be modified by the function P(x), where x is the quantity of pigment deposited and P is the probability of avoiding predator detection through crypsis. Obviously, even with maximal crypsis, there is still a chance of being depredated, but this probability is constant with respect to pigmentation and so it is not considered here. Let us assume that the animal is typically present on only one type of background and has only one type of predator, thus avoiding the heterogeneity due to different backgrounds and predators with different visual systems, which is beyond the scope of this paper. Hence, a specific amount of pigment provides a maximum degree of crypsis (xcrypsis). Therefore, when x = xcrypsis, P(xcrypsis) ~ 1; that is, the animal has the lowest probability of being preyed upon. However, if x ≠ xcrypsis, P(x) < 1, the animal would have a higher probability of being depredated, thereby reducing its fitness. As such, fitness as a consequence of predation is a function of the level of pigment the animal

Animal Biodiversity and Conservation 43.1 (2020)

Fig. 1. Example of how the level of saturation may affect visual crypsis. The circle has the same values of hue and lightness as the background (H = 150, L = 150), but differs in saturation; saturation is 100 for the background and 25 (A), 125 (B) and 250 (C) for the circle. It is evident that the circle is most cryptic in (B), where the overall difference in saturation is the smallest. Fig. 1. Ejemplo de cómo el grado de saturación puede afectar a la cripsis visual. El círculo tiene los mismos valores de matiz y luminosidad que el fondo (H = 150, L = 150), pero difiere en saturación; la saturación es 100 para el fondo y 25 (A), 125 (B) y 250 (C) para el círculo. Es evidente que el círculo más críptico es el (B), donde se produce la menor diferencia general en la saturación.

uses for crypsis. At the same time, if the acquisition, synthesis, deposition or maintenance of the pigment imposes a physiological cost on the animal in other fitness parameters (energy, immunity, oxidative balance, etc.), then fitness would also be reduced according to a function that depend directly on the quantity of pigment used for crypsis: C(x). Therefore, the fitness of an animal adopting a strategy of visual crypsis (everything being equal) may be defined by the equation: W = [1 – C(x)] · P(x)

[1]

In this equation fitness ranges from 0 to 1, it is reduced by the function C(x) whose values lies between 0 and 1 and modified by a probability function of escaping predator detection, P(x). Note that this includes the extreme possibilities in which the physiological costs are so high that the animal dies, when C(x) = 1, or when there is no cost for the case of C(x) = 0. The effect of quantity of pigment deposited on the probability of being depredated can be considered a quadratic function such that: P(x) = 1 – b(xcrypsis – x)2, where b is a parameter related to the strength of the selective pressure from predators; that is, b indicates how much fitness decreases as a consequence of increased predation risk as the degree of crypsis deviates from maximal crypsis (xcrypsis) (fig. 2). Higher values of b indicate a steeper decrease in fitness as the level of crypsis declines (fig. 2). Meanwhile, C can be approached as a linear function such as C(x) = cx, where c is the degree to which fitness is affected by the quantity of pigment (x) deposited in the tegument (fig. 2). Given these equations, fitness can be expressed as: W = [1 – cx] · [1 – b(xcrypsis – x)2]

[2]

Results To examine how the optimal pigmentation values (x*) vary depending on the three parameters defined in the model (c, b and xcrypsis), I ran computer simulations where the values of a specific parameter were allowed to vary along a continuum (0 to 1) while the other two parameters remained fixed, and the value of x* (which maximises fitness) was estimated. The results show that higher values of both b and xcrypsis increase the optimal value of x. On the other hand, the optimal value of x decreases monotonically with an increase in the costs c of depositing pigment in the tegument (fig. 3). In other words, the optimal pigmentation level increases when greater pigmentation is necessary for maximal crypsis and also when fitness decreases steeply if crypsis diminishes; however, the optimal pigmentation level decreases when the cost of pigmentation increases. The model also shows that x* is typically smaller than xcrypsis except when c = 0. That is, the costs associated with pigmentation mean that optimal pigmentation is below the level that produces maximum visual crypsis. Moreover, the function of x* according to b depended on the costs; the function was smoother for low values of c, while x* was relatively higher for low values of b (fig. 4). Discussion The model indicates that when greater quantities of pigment are needed for maximal crypsis (xcrypsis), the optimal quantity of pigment (x*) the animal should deposit in its tegument is also higher. This finding implies that in the case of crypsis resulting from eumelanin, for example, darker substrates require the incorporation of more eumelanin in the teguments.

Moreno–Rueda

Relative impact on fitness

1.0 0.8 Low b Mean b High b

0.6

Low c Mean c High c

0.4 0.2 0.0 0.0

0.2

0.4 0.6 Values of b and c

0.8

1.0

Fig. 2. Fitness associated with different values of b and c according to the level of pigmentation (x). The data consider a maximal crypsis at a pigmentation level of 0.5 (xcrypsis = 0.5). The higher the risk of predation when non–cryptic (b), the steeper the function of fitness related to pigmentation. The higher the cost of pigmentation, the greater the fitness lost. Fig. 2. Eficacia biológica asociada a diferentes valores de b y c según el grado de pigmentación (x). Los datos apuntan que la cripsis es máxima cuando el grado de pigmentación es de 0,5 (xcrypsis = 0,5). Cuanto mayor es el riesgo de depredación en ausencia de cripsis (b), más pronunciada es la función de la eficacia biológica relacionada con la pigmentación. Cuanto mayor sea el coste de la pigmentación, mayor será la pérdida de eficacia biológica.

On the other hand, the optimal quantity of eumelanin would be lower for light substrates. Thus, one prediction from the model is a match between the background colour and the quantity of pigment used for crypsis, which is supported by a number of studies that have reported a relationship between substrate hue and animal coloration (Nachman et al., 2003; Laurent et al., 2016). However, the model also predicts that x* will generally be smaller than xcrypsis, so crypsis is not usually maximal. Typically, imperfect crypsis has been attributed to background heterogeneity (Hughes et al., 2019), gene flow (Rosenblum, 2006), constrictions for crypsis (Cutchill, 2019), or a conflict with other selective pressures on coloration such as sexual selection (Martin and Badyaev, 1996). As such, the present model suggests that imperfect crypsis may be widespread in the wild. According to the model, maximal crypsis is only reached when the cost of pigmentation is null, so x* = xcrypsis. Pigmentation costs could be zero if animals compensate for it, such as in the case of an excess of amino–acids required in pigment synthesis. Nevertheless, for very high predation–risk values (b) and/or reduced pigmentation costs (c), the optimal level of pigmentation may be very close to the level needed for maximal crypsis. Mathematically, if b >> c, then x* ~ xcrypsis. Notice that

the model not only predicts imperfect crypsis, but also that animals will deposit less pigment than the amount required to maximise crypsis. This means that animals should have less saturated coloration than necessary for maximal crypsis rather than an excess of pigment. As far as I am aware, this prediction from the model has not been studied empirically. Parameter b is an indicator of how the risk of being depredated varies as crypsis deviates from maximum, i.e. the crypsis–dependent risk of predation. The higher the b, the quicker fitness is lost because the tegument coloration differs from the background (fig. 2). The model predicts higher values of x* for a greater crypsis–dependent risk of predation. Endler (1978) stated that if selection by predation is weak, crypsis may not be very accurate. My model conceptualises why the level of crypsis depends on the risk of predation, the key factor being the cost associated with pigmentation. Although b is determined by the predator's perception, if pigmentation costs were null, crypsis would be maximal even for a very low predation risk. In other words, in the long evolutionary run, and with no other selective pressures intervening, a weak predation risk would only allow reduced visual crypsis if pigmentation were costly. However, if the risk is very high, the optimal level of pigmentation will approach that of maximal crypsis even for high pigmentation costs (fig. 4).

Animal Biodiversity and Conservation 43.1 (2020)

Optimal level of pigmentation (x*)

0.6 xcrypsis b c

0.5 0.4 0.3 0.2 0.1 0.0 0.2

0.3

0.4 0.5 0.6 Values for xcrypsis (b or c)

0.7

0.8

Fig. 3. Optimal level of pigmentation (x*) as a function of pigmentation level that maximises crypsis (xcrypsis), predation risk when non–cryptic (b) and costs associated with pigmentation (c). Fig. 3. Grado óptimo de pigmentación (x*) en función del grado de pigmentación en el que la cripsis es máxima (xcrypsis), el riesgo de depredación en ausencia de cripsis (b) y los costes asociados a la pigmentación (c).

The costs of pigmentation, on the other hand, affect the way in which x* varies with predation risk, with lower values and steeper functions as the costs become higher (fig. 4). The model therefore predicts that crypsis should be higher as predation risk increases; this conjecture is supported by some studies (e.g. Endler, 1980). Nevertheless, the model also stresses that, at the intraspecific level, individuals with a higher predation risk (e.g. those with a lower capacity for escape) are expected to be more cryptic (provided all else is equal, i.e. no other antipredator strategies are involved). As expected, the optimal level of pigmentation, which maximises fitness, should decline as the cost of pigmentation increases. As explained earlier, acquisition, mobilisation, synthesis and maintenance of pigmentation may be costly, at least under certain circumstances. An important implication of the model is that if pigmentation costs vary between individuals, then the optimal level of crypsis should also vary between individuals. To the best of my knowledge, the literature is bereft of any studies on this aspect. So the model predicts that, all else being equal, individuals for whom pigmentation is more costly should be less cryptic than those with less costly pigmentation. For example, given that grooming is very costly in terms of time and energy (Walther and Clayton, 2005), unwell individuals might reduce time spent grooming (Yorinks and Atkinson, 2000) and hence plumage coloration fades (Zampiga et al., 2004). These less cryptic individuals may therefore favour alternative

antipredator strategies, such as aposematism (Merilaita and Tullberg, 2005), or their lower crypsis (and hence higher mortality prospects) may affect their life history, for example, favouring fast life styles (Roff, 2002). What is more, costs might not be limited to pigmentation but also affect pigmentation pattering, which is very important for camouflage (Merilaita and Lind, 2005). In conclusion, if pigmentation for crypsis is costly, it involves several ecological and evolutionary implications that require further investigation. This model, however, is a first approach and has a number of limitations. Firstly, it assumes that only one pigment is involved in tegument coloration. This situation is probably very rare in nature, where coloration typically results in a mix of pigments and structural traits (Grether et al., 2004). Moreover, changes in tegument coloration can be achieved not only by varying the concentration of a pigment, but also by changing the pigment (e.g. replacing eumelanin with pheomelanin) or by pigment abrasion (Negro et al., 2019). In such cases, if each pigment is associated with a different cost, increasing crypsis could result in a lower physiological cost than remaining less cryptic (Grether et al., 2004). Furthermore, camouflage frequently results from a pattern of several colours providing disruptive crypsis (Cuthill et al., 2005), a complication that is not captured by the model. The model only considers what happens when the prey moves across one type of substrate and has one type of predator. Heterogeneity regarding the microhabitats the prey uses is known to constrain

Moreno–Rueda

Optimal level of pigmentation (x*)

0.5

0.4

c = 0.1

0.3

c = 0.2 c = 0.3

0.2

0.1

0.0 0.0

0.1

0.2

0.3 0.4 0.5 0.6 0.7 Risk of predation (b)

0.8

0.9

1.0

Fig. 4. Optimal level of pigmentation (x*) for maximal crypsis at a pigmentation level of 0.5 (xcrypsis = 0.5), as a function of the predation risk when non–cryptic (b) for different pigmentation costs (c). The optimal level of pigmentation increases monotonically with the risk of predation, approaching an asymptote at xcrypsis = 0.5. The higher the pigmentation cost, the lower the optimal pigmentation for a given risk predation and the asymptote is reached more slowly. Fig. 4. Grado óptimo de pigmentación (x*) para una cripsis máxima a un grado de pigmentación de 0,5 (xcrypsis = 0,5), en función del riesgo de depredación (b), para diferentes costes de pigmentación (c). El grado óptimo de pigmentación aumenta con el riesgo de depredación y se acerca a una asíntota en xcrypsis = 0,5. Cuanto mayor es el coste de pigmentación, menor es la pigmentación óptima para una depredación de riesgo dada y más lentamente se alcanza la asíntota.

the evolution of crypsis (Hughes et al., 2019). Prey found living in several habitats are predicted to evolve a compromise coloration or customise their crypsis for a specific microhabitat, usually the most frequent or that in which they are at most risk (Merilaita et al., 1999; Houston et al., 2007; Nilsson and Ripa, 2010). Incorporating pigmentation costs in these models might modify their predictions as prey could specialise their crypsis for microhabitats where it is less costly. Regarding the effect of multiple predators, theory predicts the evolution of monomorphic prey (cryptic against the most danger predator) or a polymorphism (Endler, 1988). The consideration of pigmentation costs would probably affect these predictions, favouring the less expensive crypsis under certain circumstances. The movement of prey may also affect the effectiveness of crypsis (Hall et al., 2013). Future extensions of this model should explore all these limitations. In conclusion, this model suggests that crypsis evolution may be strongly affected by pigmentation costs. The model presented here provides a number of new predictions that are simple to test empirically: (1) imperfect crypsis should be common in the wild; (2) the level of pigmentation should typically be lower (not higher) than the level of maximum crypsis; (3) crypsis should be closer to maximum when risk of

predation is higher; (4) crypsis should be closer to maximum when the cost of pigmentation is lower; and (5) crypsis should present intraspecific variation: individuals that are generally more exposed to predators and those for which pigments are less costly should show more cryptic coloration. Acknowledgements Comments from Mar Comas, the Associate Editor, two anonymous referees and Sami Merilaita have helped to greatly improve the manuscript. References Bond, A. B., Kamil, A. C., 2002. Visual predators select for crypticity and polymorphism in virtual prey. Nature, 415: 609–613. Caro, T., 2005. Antipredator defenses in birds and mammals. Chicago University Press, Chicago. Cook, L. M., Grant, B. S., Saccheri, I. J., Mallet, J., 2012. Selective bird predation on the peppered moth: the last experiment of Michael Majerus. Biology Letters, 8: 609–612. Cooper, J. M., Allen, J. A., 1994. Selection by wild

Animal Biodiversity and Conservation 43.1 (2020)

birds on artificial dimorphic prey on varied backgrounds. Biological Journal of the Linnean Society, 51: 433–446. Cooper, W. E. J., Blumstein, D. T., 2015. Escaping from predators: an integrative view of escape decisions. Cambridge University Press, Cambridge. Cuthill, I. C., 2019. Camouflage. Journal of Zoology, 308: 75–92. Cuthill, I. C., Stevens, M., Sheppard, J., Maddocks, T., Párraga, C. A., Troscianko, T. S., 2005. Disruptive coloration and background pattern matching. Nature, 434: 72–74. Dimitrova, M., Merilaita, S., 2014. Hide and seek: properties of prey and background patterns affect prey detection by blue tits. Behavioral Ecology, 25: 402–408. Ducrest, A.–L., Keller, L. F., Roulin, A., 2008. Pleitropy in the melanocortin system, coloration and behavioural syndromes. Trends in Ecology & Evolution, 23: 502–510. Endler, J. A., 1978. A predator’s view of animal color patterns. Evolutionary Biology, 11: 319–364. – 1980. Natural selection on colour patterns in Poecilia reticulata. Evolution, 34: 76–91. – 1988. Frequency–dependent predation, crypsis and aposematic coloration. Philosophical Transactions of the Royal Society B: Biological Sciences, 319: 505–523. – 1990. On the measurement and classification of colour in studies of animal colour patterns. Biological Journal of the Linnean Society, 41: 315–352. Fargallo, J. A., Laaksonen, T., Korpimäki, E., Wakamatsu, K., 2007. A melanin–based trait reflects environmental growth conditions of nestling male Eurasian kestrels. Evolutionary Ecology, 21: 157–171. Galván, I., García–Campa, J., Negro, J. J., 2017. Complex plumage patterns can be produced only with the contribution of melanins. Physiological and Biochemical Zoology, 90: 600–604. Grether, G. F., Kolluru, G. R., Nersissian, K., 2004. Individual colour patches as multicomponent signals. Biological Reviews, 79: 583–610. Hall, J. R., Cuthill, I. C., Baddeley, R., Shohet, A. J., Scott–Samuel, N. E., 2013. Camouflage, detection and identification of moving targets. Proceedings of the Royal Society B: Biological Sciences, 280: 20130064. Hill, G. E., McGraw, K. J., 2006. Bird coloration. Vol. I: Mechanisms and measurements. Harvard University Press, Cambridge. Houston, A. I., Stevens, M., Cuthill, I. C., 2007. Animal camouflage: compromise or specialize in a 2 patch–type environment? Behavioral Ecology, 18: 769–775. Hubbard, J. K., Uy, J. A. C., Hauber, M. E., Hoekstra, H. E., Safran, R. J., 2010. Vertebrate pigmentation: from underlying genes to adaptive function. Trends in Genetics, 26: 231–239. Hughes, A., Liggins, E., Stevens, M., 2019. Imperfect camouflage: how to hide in a variable world? Proceedings of the Royal Society B: Biological Sciences, 286: 20190646.

Ito, S., Wakamatsu, K., 2003. Quantitative analysis of eumelanin and pheomelanin in humans, mice, and other animals: a comparative review. Pigment Cell Research, 16: 523–531. Jawor, J. M., Breitwisch, R., 2003. Melanin ornaments, honesty, and sexual selection. The Auk, 120: 249–265. Laurent, S., Pfeifer, S. P., Settles, M. L., Hunter, S. S., Hardwick, K. M., Ormond, L., Sousa, V. C., Jensen, J. D., Rosenblum, E. B., 2016. The population genomics of rapid adaptation. Molecular Ecology, 25: 306–323. Martin, T. E., Badyaev, A. V., 1996. Sexual dichromatism in birds: importance of nest predation and nest location for females versus males. Evolution, 50: 2454–2460. McGraw, K. J., 2006. Mechanisms of melanin–based coloration. In: Bird coloration. Vol. I: Mechanisms and measurements: 243–294 (G. E. Hill, K. J. McGraw, Eds.). Harvard University Press, Cambridge. McGraw, K. J., Gregory, A. J., 2004. Carotenoid pigments in male American goldfinches. Biological Journal of the Linnean Society, 83: 273–280. McGraw, K. J., Safran, R. J., Wakamatsu, K., 2005. How feather colour reflects its melanin content. Functional Ecology, 19: 816–821. McGraw, K. J., Toomey, M. B., 2010. Carotenoid accumulation in the tissues of zebra finches: predictors of integumentary pigmentation and implications for carotenoid allocation strategies. Physiological and Biochemical Zoology, 83: 97–109. McGraw, K. J., Wakamatsu, K., 2004. Melanin basis of ornamental feather colors in male Zebra Finches. Condor, 106: 686–690. Merilaita, S., Dimitrova, M., 2014. Accuracy of background matching and prey detection: predation by blue tits indicates intense selection for highly matching prey colour pattern. Functional Ecology, 28: 1208–1215. Merilaita, S., Lind, J., 2005. Background–matching and disruptive coloration, and the evolution of cryptic coloration. Proceedings of the Royal Society B: Biological Sciences, 272: 665–670. Merilaita, S., Scott–Samuel, N. E., Cuthill, I. C., 2017. How camouflage works. Philosophical Transactions of the Royal Society B: Biological Sciences, 372. Merilaita, S., Stevens, M., 2011. Crypsis through background matching. In: Animal camouflage: Mechanisms and function: 17–33 (M. Stevens, S. Merilaita, Eds.). Cambridge University Press, Cambridge. Merilaita, S., Tullberg, B. S., 2005. Constrained camouflage facilitates the evolution of conspicuous warming coloration. Evolution, 59: 38. Merilaita, S., Tuomi, J., Jormalainen, V., 1999. Optimization of cryptic coloration in heterogeneous habitats. Biological Journal of the Linnean Society, 67: 151–161. Michalis, C., Scott–Samuel, N. E., Gibson, D. P., Cuthill, I. C., 2017. Optimal background matching camouflage. Proceedings of the Royal Society B: Biological Sciences, 284: 20170709.

Montgomerie, R., 2006. Analazing colors. In: Bird coloration. Vol. I: Mechanisms and measurements: 90–147 (G. E. Hill, K. J. McGraw, Eds.). Harvard University Press, Cambridge. Moreno–Rueda, G., 2010. Uropygial gland size correlates with feather holes, body condition and wingbar size in the house sparrow Passer domesticus. Journal of Avian Biology, 41: 229–236. Nachman, M. W., Hoekstra, H. E., D'Agostino, S. L., 2003. The genetic basis of adaptive melanism in pocket mice. Proceedings of the National Academy of Sciences, 100: 5268–5273. Negro, J. J., Galván, I., Potti, J., 2019. Adaptive plumage wear for increased crypsis in the plumage of Palearctic larks (Alaudidae). Ecology, 100: e02771. Nilsson, J., Ripa, J., 2010. The origin of polymorphic crypsis in a heterogeneous environment. Evolution, 64: 1386–1394. Poston, J. P., Hasselquist, D., Stewart, I. R. K., Westneat, D. F., 2005. Dietary amino acids influence plumage traits and immune responses of male house sparrows, Passer domesticus, but not as expected. Animal Behaviour, 70: 1171–1181. Prum, R. O., 2006. Anatomy, physics, and evolution of structural colors. In: Bird coloration, vol. I: Mechanisms and measurements: 295–353 (G. E. Hill, K. J. McGraw, Eds.). Harvard University Press, Cambridge. Rodgers, G. M., Gladman, N. W., Corless, H. F., Morrell, L. J., 2013. Costs of colour change in fish. Journal of Experimental Biology, 216: 2760–2767. Roff, D. A., 2002. Life history evolution. Sinauer Associates, Sunderland. Rosenblum, E. B., 2006. Convergent evolution and divergent selection: lizards at the White Sands ecotone. American Naturalist, 167: 1–15. Roulin, A., Mangels, J., Wakamatsu, K., Bachmann, T., 2013. Sexually dimorphic melanin–based colour polymorphism, feather melanin content, and wing

Moreno–Rueda

feather structure in the barn owl (Tyto alba). Biological Journal of the Linnean Society, 109: 562–573. Ruxton, G. D., Allen, W. L., Sherratt, T. N., Speed, M. P., 2004. Avoiding attack: The evolutionary ecology of crypsis, aposematism, and mimicry. Oxford University Press, Oxford. Stevens, M., Merilaita, S., 2009. Animal camouflage: current issues and new perspectives. Philosophical Transactions of the Royal Society B: Biological Sciences, 364: 423–427. – 2011. Animal camouflage: Mechanisms and function. Cambridge University Press, Cambridge. Stuart–Fox, D. M., Moussalli, A., Marshall, N. J., Owens, I. P. F., 2003. Conspicuous males suffer higher predation risk. Animal Behaviour, 66: 541–550. Troscianko, J., Wilson–Aggarwal, J., Stevens, M., Spottiswoode, C. N., 2016. Camouflage predicts survival in ground–nesting birds. Scientific Reports, 6: 19966. Vágási, C. I., Pap, P. L., Barta, Z., 2010. Haste makes waste: Accelerated molt adversely affects the expression of melanin–based and depigmented plumage ornaments in house sparrows. Plos One, 5: e14215. Walther, B. A., Clayton, D. H., 2005. Elaborate ornaments are costly to maintain. Behavioral Ecology, 16: 89–95. Walton, O. C., Stevens, M., 2018. Avian vision models and field experiments determine the survival value of peppered moth camouflage. Communications Biology, 1: 118. Yorinks, N., Atkinson, C. T., 2000. Effects of malaria on activity budgets of experimentally infected juvenile Apapane (Himantione sanguinea). The Auk, 117: 731–738. Zampiga, E., Hoi, H., Pilastro, A., 2004. Preening, plumage reflectance and female choice in budgerigars. Ethology Ecology & Evolution, 16: 339–349.

Animal Biodiversity and Conservation 43.1 (2020)

Terrestrial mammal community richness and temporal overlap between tigers and other carnivores in Bukit Barisan Selatan National Park, Sumatra M. L. Allen, M. C. Sibarani, L. Utoyo, M. Krofel

Allen, M. L., Sibarani, M. C., Utoyo, L., Krofel, M., 2020. Terrestrial mammal community richness and temporal overlap between tigers and other carnivores in Bukit Barisan Selatan National Park, Sumatra. Animal Biodiversity and Conservation, 43.1: 97–107, DOI: https://doi.org/10.32800/abc.2020.43.0097 Abstract Terrestrial mammal community richness and temporal overlap between tigers and other carnivores in Bukit Barisan Selatan National Park, Sumatra. Rapid and widespread biodiversity losses around the world make it important to survey and monitor endangered species, especially in biodiversity hotspots. Bukit Barisan Selatan National Park (BBSNP) is one of the largest conserved areas on the island of Sumatra, and is important for the conservation of many threatened species. Sumatran tigers (Panthera tigris sumatrae) are critically endangered and serve as an umbrella species for conservation, but may also affect the activity and distribution of other carnivores. We deployed camera traps for 8 years in an area of Bukit Barisan Selatan National Park (BBSNP) with little human activity to document the local terrestrial mammal community and investigate tiger spatial and temporal overlap with other carnivore species. We detected 39 mammal species including Sumatran tiger and several other threatened mammals. Annual species richness averaged 21.5 (range 19–24) mammals, and remained stable over time. The mammal order significantly affected annual detection of species and the number of cameras where a species was detected, while species conservation status did not. Tigers exhibited a diurnal activity pattern, and had the highest temporal overlap with marbled cats (Pardofelis marmorata), dholes (Cuon alpinus), and Malayan sun bears (Helarctos malayanus), but little overlap with other carnivores. These findings suggest that some smaller carnivores might be adjusting temporal activity to avoid tigers or mesocarnivores. The stable trends in richness of terrestrial mammal species show that BBSNP remains an important hotspot for the conservation of biodiversity. Key words: Activity patterns, Carnivores, Conservation, Interspecific interactions, Panthera tigris sumatrae Resumen Riqueza de la comunidad de mamíferos terrestres y solapamiento temporal entre el tigre y otros carnívoros en el Parque Nacional Bukit Barisan Selatan, Sumatra. Debido a la pérdida rápida y generalizada de biodiversidad en todo el mundo, es importante estudiar las especies en peligro de extinción, en especial en zonas de gran biodiversidad, y de hacer un seguimiento de dichas especies. El Parque Nacional Bukit Barisan Selatan (BBSNP en sus siglas en inglés) es una de las mayores zonas de conservación de la isla de Sumatra y es importante para la conservación de muchas especies amenazadas. El tigre de Sumatra (Panthera tigris sumatrae) se encuentra en peligro crítico de extinción y sirve de especie paraguas para la conservación, pero también puede afectar a la actividad y la distribución de otros carnívoros. Utilizamos cámaras de trampeo durante 8 años en una zona del Parque Nacional BBSNP con escasa actividad humana, a fin de documentar la comunidad local de mamíferos terrestres y estudiar el solapamiento espacial y temporal del tigre con otras especies de carnívoros. Detectamos 39 especies de mamíferos como el tigre de Sumatra y otros varios mamíferos amenazados. La riqueza anual de especies se situó de media en 21,5 mamíferos (intervalo 19–24) y se mantuvo estable a lo largo del tiempo. A diferencia de la situación de conservación de la especie, el orden de los mamíferos tuvo un efecto significativo en la detección anual de especies y el número de cámaras en las que se detectó una especie. El tigre mostró una pauta de actividad diurna y el mayor solapamiento temporal con el gato jaspeado (Pardofelis marmorata), el cuón (Cuon alpinus) y el oso malayo (Helarctos malayanus), pero poco solapamiento con otros carnívoros. Estos resultados sugieren que algunos carnívoros de menor talla podrían estar ajustando la actividad temporal para ISSN: 1578–665 X eISSN: 2014–928 X

Allen et al.

evitar a los tigres o a mesocarnívoros. La tendencia estable de la riqueza de especies de mamíferos terrestres pone de manifiesto que el BBSNP sigue siendo una zona importante para la conservación de la biodiversidad. Palabras clave: Pautas de actividad, Carnívoros, Conservación, Interacciones interespecíficas, Panthera tigris sumatrae Received: 31 VII 19; Conditional acceptance: 20 IX 19; Final acceptance: 18 XI 19 Maximilian L. Allen, Illinois Natural History Survey, University of Illinois, 1816 S. Oak Street, Champaign, IL 61820, U.S.A.– Marsya C. Sibarani, Laji Utoyo, Wildlife Conservation Society–Indonesia Program, Jalan Tampomas Ujung No. 35, Bogor, West Java 16151, Indonesia.– Miha Krofel, Department of Forestry, Biotechnical Faculty, University of Ljubljana, Večna pot 83, SI–1000 Ljubljana, Slovenia. Corresponding author: Maximilian L. Allen. E–mail: maxallen@illinois.edu

Animal Biodiversity and Conservation 43.1 (2020)

Introduction The effects of humans are widespread and in the past centuries have led to historically high rates of extinction around the world (Pimm et al., 1995; Chapin et al., 1998). In order to guide effective conservation efforts to address biodiversity loss it is important to survey and monitor endangered species and the biodiversity in crucial areas for conservation, such as large protected areas in biodiversity hotspots (Johnson et al., 2009; Gopal et al., 2010). The effects of humans are often strongest on carnivores (Ripple et al., 2014), which occupy high trophic levels and structure ecosystems and community composition through predation and other interspecific interactions (Estes and Palmisano, 1974; McLaren and Peterson, 1994). Carnivores also act as umbrella species as they require large areas for viable populations and through their preservation protect the habitat of many other co–occurring species (Sibarani et al., 2019). Camera trapping is one of the many emerging technologies that is increasingly being used to monitor wildlife (O'Brien, 2008; Karanth and Nichols, 2010), and can be a critical non–invasive tool in documenting cryptic and endangered wildlife (Linkie et al., 2007; Tobler et al., 2008). Surveys of carnivores are important, particularly when they are threatened or endangered, and surveys via camera trap also allow for surveying a diversity of other species. The Indonesian island of Sumatra is located in one of the global hotspots of biodiversity and represents a conservation priority (Myers et al., 2000). Bukit Barisan Selatan National Park (BBSNP) is one of the largest conserved areas on the island of Sumatra, making it important for the conservation of several critically endangered species (e.g., Sumatran rhinoceros, Dicerorhinus sumatrensis, and Sunda Pangolin, Manis javanica) and subspecies (e.g., Sumatran tigers, Panthera tigris sumatrae; and Sumatran elephant, Elephas maximus sumatranus) (O’Brien and Kinnaird, 1996; Pusparini et al., 2018; Allen et al., 2019). Tigers are an endangered apex carnivore throughout their range (Goodrich et al., 2015), but four subspecies are likely now extinct in the wild (Seidensticker et al., 1973; Goodrich et al., 2015). The Sumatran tiger subspecies is one of the most critically endangered carnivores in the world (Linkie et al., 2008b), and Sumatran tigers serve as an umbrella species for scientific studies and conservation in many areas of their range. As a national park, BBSNP acts as a preserve and stronghold for biodiversity, but there is no buffer between the park and adjacent agriculture, resulting in frequent illegal encroachment into the park (O’Brien and Kinnaird, 1996; Pusparini et al., 2018). Repeated surveys are needed to understand the biodiversity of the park, as well as trends in threatened and endangered populations over time. The ecology of most carnivore species occurring on Sumatra, including their activity patterns, is poorly studied (Hunter, 2015), but it is important to understand in order to develop effective means for their conservation. The interactions between carnivores are important as the conservation of one species

can have detrimental effects on other species, and management plans need to account for interspecific interactions to mitigate these side–effects (Krofel and Jerina, 2016). Camera trapping is frequently used to monitor wildlife, providing a wealth of information on the spatial and temporal activity of species in the local community (Swanson et al., 2015; Rich et al., 2016; Allen et al., 2019). Temporal patterns are important aspects of niche partitioning among sympatric carnivores (Romero–Munoz et al., 2010; Karanth et al. 2017; Herrera et al., 2018), with subordinate carnivores often adjusting their temporal activity to avoid overlap with dominant carnivores (Foster et al., 2013; Lynam et al., 2013; Wang et al., 2015), but not always (Balme et al., 2017; Allen et al., 2018). Determining temporal patterns and overlap among species is a way to inform our understanding of cryptic species and their interactions (e.g., Van Schaik and Griffiths, 1996; Linkie and Ridout, 2011; O'Brien et al., 2003). We deployed camera traps in BBSNP in Sumatra over eight years to document trends in the local terrestrial mammalian community, as well as temporal overlap between tigers and other carnivores to inform conservation efforts. Our objectives were: (1) Determine the trends in annual mammal species richness, and relative abundance of species in the study area. We hypothesized that trends in richness would be stable due to the relatively short time period of the surveys. We also hypothesized that camera traps would detect higher relative abundances for Artiodactyla than Carnivora which occur at lower densities and primate species due to their arboreal nature. (2) Define factors affecting annual detection for species. We hypothesized that mammals species from lower trophic levels and of lower conservation concern would be detected in more years due to their greater abundance. (3) Compare mammal detections from our camera trapping with previous surveys using track surveys and interviews with local experts from BBNSP by O'Brien and Kinnaird (1996). (4) Determine factors affecting species occupancy. We hypothesized that camera traps would detect higher occupancies for Artiodactyla than Carnivora which occur at lower densities and primate species due to their arboreal nature. We also hypothesized that the conservation status of species would be related to occupancy, with endangered species being detected at fewer cameras then species of less concern. (5) Analyze the temporal overlap of tigers with four other felids and six other carnivores, hypothesizing that subordinate competitor carnivores would have low temporal overlap with the apex predator, the Sumatran tiger (e.g., Lynam et al., 2013; Wang et al., 2015). Material and methods Study area Our study site is located in BBSNP in the South Barisan Range ecosystem on the Indonesian island of Sumatra (fig. 1). BBSNP is the third largest protected area (3,560 km²) on Sumatra (O'Brien and Kinnaird,

Allen et al.

100

1996), spanning two provinces: Lampung and Bengkulu. Topography ranges from coastal plains and lowland rainforest at sea level in the southern peninsula of the park to mountains up to 1,964 m in the middle to northern parts of the park (Pusparini et al., 2018). The park contains montane forest, lowland tropical forest, coastal forest and mangrove forest. Rainfall is most abundant in the monsoon season from November to May, with approximately 3,000–4,000 mm of rainfall (O'Brien et al., 2003); and annual temperatures are between 22 ºC to 35 ºC (O'Brien et al., 2003). BBSNP contains a high diversity of wildlife, with tigers and 76 other species listed in CITES with Endangered to Critical IUCN status. Field methods We set camera traps to monitor biodiversity as part of the Tropical Ecology and Assessment Monitoring (TEAM) network for Bukit Barisan (teamnetwork.org). We set two arrays of 30 camera traps placed at a density of 1 camera trap per 2 km2 (fig. 1) covering a total of 128.43 km2. We attempted to set each camera trap annually from 2010–2017, and we deployed the camera arrays sequentially rather than simultaneously within the same dry season from April to July (Array 1 from April to May and Array 2 from June to July) to complete at least 30 days of sampling for each point. We placed camera traps in strategic locations on game trails. We set all camera traps in lowland forests, with an elevation range of 16 to 320 m. Statistical analyses We defined a detection event as any series of photos triggered by a human or wildlife species. To avoid pseudo–replication, we considered consecutive photo captures of the same species within 30' to be the same event (Rovero and Zimmermann, 2016; Allen et al., 2018). We calculated the number of independent events for each species and relative abundance (RAB) as: RAB = events / trap nights x 1,000 and then report the mean annual RAB (0 RAB) for each species. We calculated annual species richness by totaling the number of unique mammal species detected each year. We also calculated the naïve annual occupancy and mean naïve annual occupancy for each species (Nichols et al., 2007; O'Connell and Bailey, 2011). We used generalized linear mixed models (GLMMs) to determine if the RAB of species was affected by either their order or conservation status, using the annual RAB of a species as our dependent variable, their order or conservation status as the independent variable, and species as a random effect. We also used GLMs to determine if the number of years a species was detected was affected by either their order or conservation status, using the number of years a species was detected as our dependent variable, their order or conservation status as the independent

variable, and species as a random effect. We also compared species richness between our surveys and the previous survey by O'Brien and Kinnaird (1996), using a t–test to compare our annual values to the previous value. We then used GLMMS to determine if the occupancy of species varied annually by either their order or conservation status. We used the annual occupancy of a species as our dependent variable, their order or conservation status as the independent variable, and the species as a random effect. We used kernel density estimation to determine activity patterns and quantify overlap among species (Ridout and Linkie, 2009). We considered interactions with other carnivores for which we obtained > 3 detections. Other felids included Asiatic golden cat (Catopuma temminckii), leopard cat (Prionailurus bengalensis), marbled cat (Pardofelis marmorata), and Sunda clouded leopard (Neofelis diardi). Other carnivores included banded linsang (Prionodon linsang), banded palm civet (Hemigalus derbyanus), binturong (Arctictis binturong), dhole (Cuon alpinus), masked palm civet (Paguma larvata), and sun bear (Helarctos malayanus). We changed the time of each event to radians for each species, and then used the overlap package (Meredith and Ridout, 2017) in program R version 3.3.1 (R Core Team, 2016) to fit the data to a circular kernel density and estimated the activity level at each time period from the distribution of the kernel density. We then used the overlapEst function to test for the degree of overlap in activity patterns between tigers and the other species using their Δ1 scores (where a higher score indicates more overlap). We calculated 95 % confidence intervals by bootstrapping 10,000 estimates of activity for each species, and then using the bootEst and bootCI functions to estimate overlap between each species pair based on the boot0 score. Results Sixty camera traps functioned from 2010 to 2017 for a total of 11,896 trap nights, registering 53,120 photos, representing 3,245 independent detection events of 49 species. We detected one critically endangered species, Sunda pangolin (0 RAB = 1.31), and two critically endangered subspecies, Sumatran tiger (0 RAB = 2.41) and Sumatran elephant (0 RAB = 1.42), as well as with seven endangered species and seven vulnerable species (table 1). We found that the mammal order had a significant effect on the relative abundance of species, with Artiodactyla species having higher RAB (0 = 28.81) than Carnivora species (0 = 0.78, F = –3.55, p = 0.0004), but not other orders (p > 0.12). The conservation status of species, however, did not have a significant effect on the relative abundance of species (p > 0.51). Our observed annual species richness averaged 21.5 (range 19–24) mammals, with a relatively stable trend that did not vary significantly across time (df = 7, F = 0.91, p = 0.37). We documented eight species in all eight years, three species in seven years; but nine species were detected in only one year. We found

Animal Biodiversity and Conservation 43.1 (2020)

101

BBSNP

at um

Camera trap locations Array 1 0 Array 2

10 km

Fig. 1. Study site within Bukit Barisan National Park on the island of Sumatra (BBSNP), and camera trap arrays. We did not display the coordinate reference grids because of conservation concerns and some of the recorded species are hunted for illegal wildlife trading. Fig. 1. Zona del estudio dentro del Parque Nacional Bukit Barisan en la isla de Sumatra (BBSNP) y disposición de las cámaras de trampeo. No se muestran las cuadrículas de coordenadas de referencia por motivos de conservación y porque algunas de las especies son objeto de caza para el comercio ilegal.

that the mammal order had a significant effect on the number of years detected, with Artiodactyla species found in more years (0 = 6.50) than Carnivora species (0 = 3.33, F = –2.81, p = 0.005), but not other orders (p > 0.08) including Eulipotyphla (0 = 6.00), Perissodactyla (0 = 8.00), Pholidota (0 = 7.00), Primates (0 = 4.00), Proboscidea (0 = 6.00), Rodentia (0 = 4.75), or Scandentia (0 = 4.00). The conservation status of species, however, did not have a significant effect on how many years a species was detected (p > 0.53). The number of species we observed in any given year did not vary significantly from the previous surveys by O'Brien and Kinnaird (1996) (df = 7, p = 0.77). We detected 39 mammal species, 26 (excluding humans and domestic dogs) of which had not been documented in previous surveys by O'Brien and Kinnaird (1996); but we did not detect eight mammal species which had previously been detected (table 1). We found that the species' order had a significant effect on their occupancy, with Artiodactyla species having significantly higher mean annual occupancy (0 = 0.28) than Carnivora (0 = 0.02, T = –3.62, p = 0.0003), but not other orders (p > 0.08). The conservation status of species, however, did not have a significant effect on how many cameras they were documented at annually (p > 0.59). Tigers exhibited a diurnal activity pattern (fig. 2, 3). We documented four other felid species, all with lower relative abundance than tigers. Marbled cats

had a peak of activity in the morning and were active during the day, leading to the highest overlap with tigers (fig. 2). Asian golden cats were crepuscular with their highest activity at dawn, leading to some overlap with tigers, while leopard cats and Sunda clouded leopards were primarily nocturnal and had little overlap with tigers (fig. 2). We documented six other carnivore species, with banded palm civets and sun bears having higher RAB than tigers. Dholes were diurnal and had the highest temporal overlap with tigers, while Malayan sun bears were cathemeral and had less overlap with tigers. Banded linsangs, banded palm civets, binturongs, and masked palm civets were primarily nocturnal and exhibited little overlap with tigers (fig. 2). Discussion BBSNP and other protected areas in Sumatra contain many threatened and endangered species whose populations are imperiled primarily by encroachment and habitat destruction. Effective conservation for species or ecological communities is dependent on international teamwork among government agencies, local communities, and scientific organizations. Using camera trap surveys, we were able to monitor numerous mammal species, including critically endangered Sunda pangolins, Sumatran tigers and Sumatran elephants, along with over a dozen other threatened

Allen et al.

102

Table 1. Mammal species, IUCN status (IUCN: EN, endangered; NT, near threatened; VU, vulnerable;; LC, least concern; CE, critically endangered; NE, not evaluated), number of years documented (N), mean annual relative abundance (0 RAB), mean annual percent area occupied (0 PAO), and whether the species was documented in previous surveys O’Brien and Kinnaird (1996), last column: * subspecies critically endangered Tabla 1. Especie de mamífero, situación de la UICN, número de años documentado (N), abundancia media anual relativa (0 RAB), superficie media ocupada anual (0 PAO) y si la especie se había documentado en estudios anteriores O'Brien y Kinnaird (1996), en la última columna: * subespecies en peligro crítico. (Para las abreviaturas de la situación de la IUCN, véase arriba).

Common name

Scientific name

IUCN

Tiger

Panthera tigris

EN* 6 2.41 0.05 Yes

Asian golden cat

Catopuma temminckii

NT 6 0.88 0.03 No

Marbled cat

Pardofelis marmorata

NT 4 0.6 0.02 No

Sunda clouded leopard

Neofelis diardi

VU 3 0.39 0.01 Yes

Leopard cat

Prionailurus bengalensis

LC 7 1.07 0.03 No

Oriental small–clawed otter

Aonyx cinerea

Binturong

Arctictis binturong

VU 2 0.34 0.01 No

Small–toothed palm civet

Arctogalidia trivirgata

0.07

< 0.01

Domestic dog

Canis familiaris

0.15

< 0.01

Dhole

Cuon alpinus

EN 2 0.33 0.01 No

Otter civet

Cynogale bennettii

Malayan sun bear

Helarctos malayanus

VU 7 2.61 0.07 Yes

Banded palm civet

Hemigalus derbyanus

NT 6 2.68 0.07 No

Short–tailed mongoose

Herpestes brachyurus

NT 1 0.11 0.00 No

Eurasian otter

Lutra lutra

Hairy–nosed otter

Lutra sumatmna

EN 0 0.00 0.00 Yes

Yellow–throated marten

Martes flavigula

LC 1 0.07 0.00 No

Asian palm civet

Paradoxurus hermaphroditus

Masked palm civet

Paguma larvata

LC 6 0.83 0.03 No

Banded linsang

Prionodon linsang

LC 4 0.43 0.01 No

Domestic water buffalo

Bubalus bubalis

NE 0 0.00 0.00 Yes

Plantain squirrel

Callosciurus notatus

LC 0 0.00 0.00 Yes

Sumatran serow

Capricornis sumatraensis

VU 3 0.32 0.01 No

Sumatran rhinoceros

Dicerorhinus sumatrensis

CE 0 0.00 0.00 Yes

Moonrat

Echinosorex gymnura

LC 6 2.34 0.05 No

Sumatran elephant

Elephas maximus

EN* 6 1.42 0.03 Yes

Human

Homo sapiens

NE 5 1.24 0.03 No

Dark–handed gibbon

Hylobates agilis

EN 0 0.00 0.00 Yes

Common porcupine

Hystrix brachyura

LC 8 30.34 0.33 No

Three–striped ground squirrel Lariscus insignis

LC 8 9.95 0.10 No

Long–tailed macaque

Macaca fascicularis

LC 3 0.24 0.01 Yes

Pigtail Macaque

Macaca nemestrina

VU 8 43.76 0.63 No

Sunda pangolin

Manis javanica

CE 7 1.31 0.04 No

Red muntjac

Muntiacus muntjak

LC 8 71.99 0.71 Yes

0 RAB

0.10

0.07

0.08

0.11

0 PAO

< 0.01

Animal Biodiversity and Conservation 43.1 (2020)

103

Table 1. (Cont.)

0 PAO

Scientific name Presbytis melalophos

EN 4 0.48 0.02 Yes

Flying fox

Pteropus vampyrus

NT 0 0.00 0.00 Yes

Black giant squirrel

Ratufa bicolor

NT 0 0.00 0.00 Yes

Sambar deer

Rusa unicolor

VU 8 9.69 0.18 Yes

Horse–tailed squirrel

Sundasciurus hippurus

Wild boar

Sus scrofa

LC 8 29.4 0.39 Yes

Siamang

Symphalangus syndactylus

Malayan tapir

Tapirus indicus

EN 8 8.07 0.15 Yes

Silvery lutung

Trachypithecus cristatus

NT 0 0.00 0.00 Yes

Lesser mouse deer

Tragulus kanchil

LC 4 21.99 0.23 No

Greater mouse deer

Tragulus napu

LC 8 12.55 0.17 No

Long–tailed porcupine

Trichys fasciculata

LC 2 0.63 0.01 No

Large treeshrew

Tupaia tana

LC 4 1.37 0.03 No

or endangered mammals. Tigers are threatened and declining worldwide (Seidensticker, 2010; Walston et al., 2010), and protected areas, including BBSNP, are important for their populations (Kawanshi and Sunquist, 2004, Wibisono et al., 2009) our surveys showed the importance of BBSNP to the critically endangered tiger population in Sumatra. High species richness of terrestrial mammal species, which remained similar across the eight years of our survey, also confirms the conservation value of BBSNP and other protected areas for other threatened and endangered mammal species (Linkie et al., 2008a). Camera trapping is an informative way to gather ecological data, especially for cryptic or rare species, but is best used in conjunction with other surveys. The project documented 26 mammal species with camera trap surveys that had not been detected by O'Brien and Kinnaird (1996) in a previous survey using transects and interviews with local citizens, but our survey missed eight that had been previously documented. The additional species we identified were primarily terrestrial species, while the ones not documented were primarily arboreal species. Both survey methods detected a similar number of species in any given annual survey, though not the same species. Each method has costs and benefits that should be considered in future studies. Camera trapping is effective for documenting terrestrial cryptic mammal species, and can be used 24 hours a day over weeks or months, while transect surveys are more effective for documenting birds and amphibians. There is the possibility of misidentification of photographs or signs with either method, although misidentification should be less frequent for photographs. Using both methods in conjunction is a good approach to document the

IUCN

0 RAB

Common name Sumatran surili

1 1

0.07 0.14

< 0.01 < 0.01

No Yes

mammal community, especially in such a critically important area for conservation. Tigers exhibited diurnal activity patterns and had moderate temporal overlap with marbled cats, dholes, and sun bears, but most other carnivores had little temporal overlap with tigers. Sun bears are known for their arboreal habits and insectivorous–frugivorous diet, and therefore have less overlap in dietary and spatial use with tigers, but are also generally diurnal throughout their range (Fitzgerald and Krausman, 2002). Common leopards (Panthera pardus) were extirpated from Sumatra, and now dholes and Sunda clouded leopards are the other carnivores nearest to tigers in size, and may be their closest competitors. Sympatric carnivores that are smaller than their competitors use adaptive strategies, including temporal avoidance, to exploit the same resources and avoid intra–guild predation (Lesmeister et al., 2015; Wang et al., 2015). Contrary to our hypothesis, dholes were also diurnal and exhibited greater temporal overlap with tigers than we expected. This may be due to lack of fear of tigers on the part of dholes (e.g., Burton, 2019). Among felids, Sunda clouded leopards and leopard cats were nocturnal, while Asian golden cats and marbled cats were crepuscular, which is generally in accordance with previous studies (Van Schaik and Griffiths, 2009; Grassman et al., 2005). Temporal patterns and overlap can be complex in ecosystems with many carnivores, as competitive suppression of mesocarnivores by apex carnivores can release subordinate small carnivores from competitive pressure (e.g., Levi and Wilmers, 2012; Wang et al., 2015; Allen et al., 2017), but species are most likely to avoid the species that they perceive as the greatest threat. For example, marbled cats had a high

Allen et al.

104

A Δ1 = 0.41 (0.23–0.60)

Tiger As. golden cat

0.20

Δ1 = 0.29 (0.11–0.48)

0.10

Tiger Leopard cat

0.10 0.0

0.0 00:00 06:00 12:00 18:00 24:00

00:00 06:00 12:00 18:00 24:00

Δ1 = 0.20 (0.41–0.93) C

Δ1 = 0.29 (0.09–0.54)

Tiger Marbled cat

0.10

0.0

0.0 00:00 06:00 12:00 18:00 24:00

00:00 06:00 12:00 18:00 24:00

E Δ1 = 0.21 (0.05–0.41)

Δ1 = 0.25 (0.11–0.41)

Tiger Banded linsang

0.10

Tiger S. clouded leopard

Tiger Banded palm civet

0.10

0.0

00:00 06:00 12:00 18:00 24:00 G Δ1 = 0.18 (0.03–0.35) 0.20

Tiger Binturong

Δ1 = 0.60 (0.42–0.77) Tiger Dhole

0.20 0.10

0.10

0.0 00:00 06:00 12:00 18:00 24:00

0.0 00:00 06:00 12:00 18:00 24:00 Δ1 = 0.26 (0.09–0.46) I 0.10

00:00 06:00 12:00 18:00 24:00

Tiger Masked palm civet

Δ1 = 0.60 (0.42–0.77)

J 0.12

Tiger Sun bear

0.06 0.0

0.0 00:00 06:00 12:00 18:00 24:00 00:00 06:00 12:00 18:00 24:00 Time of day

Fig. 2. Temporal activity and overlap (kernel density), of tigers and other carnivores: A, Asian golden cat; B, leopard cat; C, marbled cat; D, Sunda clouded leopard; E, banded linsang; F, banded palm civet; G, binturong; H, dhole; I, masked palm civet; J, sun bear. Tiger activity is represented as a solid line and the other carnivore activity as a dotted line, and temporal overlap as the shaded area. Fig. 2. Actividad temporal y solapamiento (densidad de kernel) del tigre y otros carnívoros: A, gato dorado asiático; B, gato de Bengala; C, gato jaspeado; D, pantera nebulosa de Borneo; E, linsang rayado; F, civeta de las palmeras rayada; G, binturong; H, cuón; I, civeta de las palmeras enmascarada; J, oso malayo. La actividad del tigre se representa con una línea continua y la de los otros carnívoros con una línea discontinua; el solapamiento temporal es la superficie sombreada.

Animal Biodiversity and Conservation 43.1 (2020)

degree of overlap with tigers, and this may be due to temporally avoiding Sunda clouded leopards. Sunda clouded leopards are also arboreal, which increase the probability of encounters with marbled cats and may thus be perceived as a more direct threat. The temporal patterns and overlap among the carnivore community suggest that tigers may be structuring the carnivore guild, but smaller carnivores may also use different resources (prey and habitat) as a means of limiting competition and overlap with tigers (Karanth et al., 2017). Camera trapping surveys using TEAM protocols appear effective for monitoring the richness and relative abundance of the terrestrial mammal community, but may best be used in conjunction with other survey methods. Our surveys focused on terrestrial mammals, but camera trapping can also be effective for arboreal species with appropriate adjustments (Gregory et al., 2014). The TEAM protocol is set for short bursts (one month) of camera trapping, and to be effective for monitoring threatened and endangered terrestrial species. Surprisingly, the conservation status of species did not predict the number of years they were detected, relative abundance or occupancy. This may be due to the inherent differences in abundance among species of different trophic levels, or because species can be locally abundant but of conservation concern globally. Our surveys also highlight the importance of BBSNP and other parks for biodiversity and many endangered species, and there is much potential to use BBSNP for future species–specific surveys, including for critically endangered Sunda pangolins, Sumatran tigers or Sumatran elephants. The development of new analyses, such as kernel density overlap (Ridout and Linkie, 2009), help us understand the ecological interactions of species, and development of new techniques in the future should be used for further understanding cryptic species. Acknowledgements All data in this publication are available through the Tropical Ecology Assessment and Monitoring (TEAM) Network, a collaboration between Conservation International, the Missouri Botanical Garden, the Smithsonian Institution, and the Wildlife Conservation Society. The work was partially funded by these institutions, the Gordon and Betty Moore Foundation, and the Illinois Natural History Survey. Monitoring activities were managed by the Wildlife Conservation Society in collaboration with the Bukit Barisan Selatan National Park and the Ministry of Environment and Forestry, Republic of Indonesia. We thank all the field staff and forest rangers involved in the camera trap deployment, and W. Marthy and F. R. Affandi for field coordination. References Allen, M. L., Gunther, M. S., Wilmers, C. C., 2017. The scent of your enemy is my friend? The acquisition of large carnivore scent by a smaller carnivore.

105

Journal of Ethology, 35: 13–19. Allen, M. L., Peterson, B., Krofel, M., 2018. No respect for apex carnivores: distribution and activity patterns of honey badgers in the Serengeti. Mammalian Biology, 89: 90–94. Allen, M. L., Sibarani, M. C., Krofel, M., 2019. Predicting preferred prey of Sumatran tigers (Panthera tigris sumatrae) via spatiotemporal overlap. Oryx, 53: 54. Balme, G. A., Pitman, R. T., Robinson, H. S., Miller, J. R. B., Funston, P. J., Hunter, L. T. B., 2017. Leopard distribution and abundance is unaffected by interference competition with lions. Behavioral Ecology, 28: 1348–1358. Burton, A., 2019. Do dholes kill tigers? Frontiers in Ecology and the Environment, 17: 300. Chapin, F. S., Sala. O. E., Burke, I. C., Grime, J. P., Hooper, D. U., Lauenroth, W. K., Lombard, A., Mooney, H. A., Mosier, A. R., Naeem, S., Pacala, S. W., Roy, J., Steffen, W. L., Tilman, D., 1998. Ecosystem consequences of changing biodiversity. Experimental evidence and a research agenda for the future. Bioscience, 48: 45–52. Estes, J. A., Palmisano, J. F., 1974. Sea otters: their role in structuring nearshore communities. Science, 185: 1058–1060. Fitzgerald, C. S., Krausman, P. R., 2002. Helarctos malayanus. Mammalian Species, 696: 1–5. Foster, V. C., Sarmento, P., Sollmann, R., Tôrres, N., Jacomo, A. T. A., Negrões, N., Fonseca, C., Silveira, L., 2013. Jaguar and puma activity patterns and predator–prey interactions in four Brazilian biomes. Biotropica, 45: 373–379. Goodrich, J.M., Lyam, A., Miquelle, D. G., Wibisono, H. T., Kawanishi, K., Pattanavibool, A., Htun, S., Tempa, T., Karki, J., Jhala, Y. V., Karanth, U. K., 2015. Panthera tigris. The IUCN Red List of Threatened Species, 8235: e.T15955A50659951. Gopal, R. Q., Quershi, Bhardwaj M., Singh, R. K. J., Jhala, Y. V., 2010. Evaluating the status of the endangered tiger panthera tigris and its prey in Panna Tiger Reserve, Madhya Pradesh, India. Oryx, 44: 383–389. Grassman, Jr. LI., Tewes, M. E., Silvy, N. J., Kreetiyutanont, K., 2005. Ecology of three sympatric felids in a mixed evergreen forest in north–central thailand. Journal of Mammalogy, 86: 29–38. Gregory, T., Rueda, F. C., Deichmann, J., Kolowski, J., Alonso, A., 2014. Arboreal camera trapping: taking a proven method to new heights. Methods in Ecology and Evolution, 5: 443–451. Herrera, H., Chávez, E. J, Alfaro, L. D., Fuller, T., Montalvo, V., Rodrigues, F., Carrillo, E., 2018. Time partitioning between jaguar panthera onca, puma puma concolor and ocelot Leopardus pardalis (Carnivora: Felidae) in costa Rica's dry and rainforests. Revista De Biología Tropical, 66: 1559–1568. Hunter, L., 2015. Wild cats of the world, First Edition. Bloomsbury Natural History, New York, USA, Johnson, A. C., Vongkhamheng, Saithongdam T., 2009. The diversity, status and conservation of small carnivores in a montane tropical forest in northern laos. Oryx, 43: 626–633.

106

Karanth, K. U., Nichols, J. D., 2010. Non–invasive survey methods for assessing tiger populations.; tigers of the world: The science, politics, and conservation of Panthera tigris. In: Tigers of the World, 2nd edition: 241–261 (R. Tilson, P. J. Nyhus, Eds.). Elsevier, New York. Karanth, K. U., Srivathsa, A., Vasudev, D., Puri, M., Parameshwaran, R., Kumar, N. S., 2017. Spatio–temporal interactions facilitate large carnivore sympatry across a resource gradient. Proceedings of the Royal Society B, 284: 20161860. Kawanishi, K., Sunquist, M. E., 2004. Conservation status of tigers in a primary rainforest of peninsular malaysia. Biological Conservation, 120: 329–344. Krofel, M., Jerina, K., 2016. Mind the cat: Conservation management of a protected dominant scavenger indirectly affects an endangered apex predator. Biological Conservation, 197: 40–46. Lesmeister, D. B., Nielsen, C. K., Schauber, E. M., Hellgren, E. C., 2015. Spatial and temporal structure of a mesocarnivore guild in midwestern north America. Wildlife Monographs, 191: 1–61. Levi, T., Wilmers, C. C., 2012. Wolves–coyotes–foxes: A cascade among carnivores. Ecology, 93: 921–929. Linkie, M., Dinata, Y., Nugroho, A., Haidir, I. A., 2007. Estimating occupancy of a data deficient mammalian species living in tropical rainforests: Sun bears in the Kerinci Seblat region, Sumatra. Biological Conservation, 137: 20–27. Linkie, M., Ridout, M. S., 2011. Assessing tiger–prey interactions in Sumatran rainforests. Journal of Zoology, 284: 224–229. Linkie, M. R., Smith, J., Zhu, Y., Martyr, D. J., Suedmeyer, B., Pramono, J., Leader–Williams, N., 2008a. Evaluating biodiversity conservation around a large Sumatran protected area. Conservation Biology, 22: 683–690. Linkie, M., Wibisono, H. T., Martyr, D. J., Sunarto, S., 2008b. Panthera tigris ssp. sumatrae. The IUCN Red List of Threatened Species, 8235: e.T15966A5334836. Lynam, A. J., Jenks, K. E., Tantipisanuh, N., Chutipong, W., Ngoprasert, D., Gale, G. A., Steinmetz, R., Sukmasuang, R., Bhumpakphan, N., Grassman, LI., Cutter, P., Kitamura, S., Reed, D. H., Baker, M. C., Mcshea, W., Songsasen, N., Leimgruber, P., 2013. Terrestrial activity patterns of wild cats from camera–trapping. Raffles Bulletin of Zoology, 61: 407–415. McLaren, B. E., Peterson, R. O., 1994. Wolves, moose, and tree rings on Isle Royale. Science, 266: 1555–1558. Meredith, M., Ridout, M. S., 2017. Overview of the overlap package, R package version 3.3.1, http:// cran.r-project.org/web/packages/overlap/index.html [Accessed on July 2019] Myers, N., Mittermeier, R. A., Mittermeier, C. G., da Fonseca, G. A. B., Kent, J., 2000. Biodiversity hotspots for conservation priorities. Nature, 403: 853–858. Nichols, J. D., Hines, J. E., Mackenzie, D. I., Seamans, M. E., Gutierrez, R. J., 2007. Occupancy

Allen et al.

estimation and modeling with multiple states and state uncertainty. Ecology, 88:1395–400. O'Brien, T. G., 2008. On the use of automated cameras to estimate species richness for large– and medium–sized rainforest mammals. Animal Conservation, 11: 179–181. O'Brien, T. G., Kinnaird, M. F., 1996. Birds and mammals of the Bukit Barisan Selatan National Park, Sumatra, Indonesia. Oryx, 30: 207–217. O'Brien, T. G., Kinnaird, M. F., Wibisono, H. T., 2003. Crouching tigers, hidden prey: Sumatran tiger and prey populations in a tropical forest landscape. Animal Conservation, 6: 131–139. O'Connell, A. F., Bailey, L. L., 2011. Inference for occupancy and occupany dynamics. In: Camera traps in animal ecology: Methods and analyses: 191–205 (A. F. O'Connell, J. D. Nichols, K. U. Karanth, Eds.). Springer, Tokyo. Pimm, S. L., Russell, G. J., Gittleman, J.L., Brooks, T. M., 1995. The future of biodiversity. Science, 269: 347–350. Pusparini, W., Batubara, T., Surahmat, F., Ardiantiono, Sugiharti T., Muslich, M., Amama, F., Marthy, W., Andayani, N., 2018. A pathway to recovery: the critically endangered Sumatran tiger Panthera tigris sumatrae in an 'in danger' UNESCO world heritage site. Oryx, 52: 25–34. R Core Team, 2016. R, A language and environment for statistical computing. R Foundation for Statistical Computing, Vienna, Austria. Rich, L. N., Miller, D. A. W., Robinson, H. S., McNutt, J. W., Kelly, M. J., 2016. Using camera trapping and hierarchical occupancy modelling to evaluate the spatial ecology of an African mammal community. Journal of Applied Ecology, 53: 1225–1235. Ridout, M. S., Linkie, M., 2009. Estimating overlap of daily activity patterns from camera trap data. Journal of Agricultural, Biological, and Environmental Statistics, 14: 322–337. Ripple, W. J., Estes, J. A., Beschta, R. L., Wilmers, C. C., Ritchie, E. G., Hebblewhite, M., Berger, J., Elmhagen, B., Letnic, M., Nelson, M. P., Schmitz, O. J., Smith, D. W., Wallach, A. D., Wirsing, A. J., 2014. Status and ecological effects of the world’s largest carnivores. Science, 343(6167): 1241484, Doi: 10.1126/science.1241484. Romero–Muñoz, A., Maffei, L., Cuéllar, E., Noss, A. J., 2010. Temporal separation between jaguar and puma in the dry forests of southern Bolivia. Journal of Tropical Ecology, 26: 303–311. Rovero, F., Zimmermann, F., 2016. Camera trapping for wildlife research, 1st Edition. Pelagic Publishing, Exeter. Seidensticker, J., 2010. Saving wild tigers: A case study in biodiversity loss and challenges to be met for recovery beyond 2010. Integrative Zoology, 5: 285–299. Seidensticker, J. C., Hornocker, M. G., Wiles, W. V., Messick, J. P., 1973. Mountain lion social organization in the Idaho Primitive Area. Wildlife Monographs, 35: 3–60. Sibarani, M. C., Di Marco, M., Rondinini, C., Kark, S., 2019. Measuring the surrogacy potential of

Animal Biodiversity and Conservation 43.1 (2020)

charismatic megafauna species across taxonomic, phylogenetic and functional diversity on a megadiverse island. Journal of Applied Ecology, 2019: 1–12. Swanson, A., Kosmala, M., Lintott, C., Simpson, R., Smith, A., Packer, C., 2015. Snapshot Serengeti, high–frequency annotated camera trap images of 40 mammalian species in an African savanna. Scientific Data, 2: 150026. Tobler, M. W., Carrillo–Percastegui, S. E., Pitman, R. L., Mares, R., Powell, G., 2008. Further notes on the analysis of mammal inventory data collected with camera traps. Animal Conservation, 11: 187–189. Van Schaik, C. P., Griffiths, M., 1996. Activity periods of indonesian rain forest mammals. Biotropica, 28: 105–12. Walston, J., Robinson, J. G., Bennett, E. L., Breit-

107

enmoser, U., da Fonseca, G. A. B., Goodrich, J., Gumal, M., Hunter, L., Johnson, A., Ullas Karanth, K., Leader–Williams, N., MacKinnon, K., Miquelle, D., Pattanavibool, A., Poole, C., Rabinowitz, A., Smith, J. L. D., Stokes, E. J., Stuart, S. N., Vongkhamheng, C., Wibisono, H., 2010. Bringing the tiger back from the brink–the six percent solution. Plos Biology, 8: e1000485. Wang, Y., Allen, M. L., Wilmers, C. C., 2015. Mesopredator spatial and temporal responses to large predators and human development in the Santa Cruz Mountains of California. Biological Conservation, 190: 23–33. Wibisono, H. T., Figel, J. J., Arif, S. M., Ario, A., Lubis, A. H., 2009. Assessing the Sumatran tiger Panthera tigris sumatrae population in Batang Gadis national park, a new protected area in Indonesia. Oryx, 43: 634–638.

Muñoz and Farfán

Animal Biodiversity and Conservation 43.1 (2020)

Terrestrial mammal community richness and temporal overlap between tigers and other carnivores in Bukit Barisan Selatan National Park, Sumatra M. L. Allen, M. C. Sibarani, L. Utoyo, M. Krofel

Allen et al.

Animal Biodiversity and Conservation 43.1 (2020)

Allen et al.

100

Animal Biodiversity and Conservation 43.1 (2020)

101

BBSNP

at um

Camera trap locations Array 1 0 Array 2

10 km

Allen et al.

102

Common name

Scientific name

IUCN

Tiger

Panthera tigris

EN* 6 2.41 0.05 Yes

Asian golden cat

Catopuma temminckii

NT 6 0.88 0.03 No

Marbled cat

Pardofelis marmorata

NT 4 0.6 0.02 No

Sunda clouded leopard

Neofelis diardi

VU 3 0.39 0.01 Yes

Leopard cat

Prionailurus bengalensis

LC 7 1.07 0.03 No

Oriental small–clawed otter

Aonyx cinerea

Binturong

Arctictis binturong

VU 2 0.34 0.01 No

Small–toothed palm civet

Arctogalidia trivirgata

0.07

< 0.01

Domestic dog

Canis familiaris

0.15

< 0.01

Dhole

Cuon alpinus

EN 2 0.33 0.01 No

Otter civet

Cynogale bennettii

Malayan sun bear

Helarctos malayanus

VU 7 2.61 0.07 Yes

Banded palm civet

Hemigalus derbyanus

NT 6 2.68 0.07 No

Short–tailed mongoose

Herpestes brachyurus

NT 1 0.11 0.00 No

Eurasian otter

Lutra lutra

Hairy–nosed otter

Lutra sumatmna

EN 0 0.00 0.00 Yes

Yellow–throated marten

Martes flavigula

LC 1 0.07 0.00 No

Asian palm civet

Paradoxurus hermaphroditus

Masked palm civet

Paguma larvata

LC 6 0.83 0.03 No

Banded linsang

Prionodon linsang

LC 4 0.43 0.01 No

Domestic water buffalo

Bubalus bubalis

NE 0 0.00 0.00 Yes

Plantain squirrel

Callosciurus notatus

LC 0 0.00 0.00 Yes

Sumatran serow

Capricornis sumatraensis

VU 3 0.32 0.01 No

Sumatran rhinoceros

Dicerorhinus sumatrensis

CE 0 0.00 0.00 Yes

Moonrat

Echinosorex gymnura

LC 6 2.34 0.05 No

Sumatran elephant

Elephas maximus

EN* 6 1.42 0.03 Yes

Human

Homo sapiens

NE 5 1.24 0.03 No

Dark–handed gibbon

Hylobates agilis

EN 0 0.00 0.00 Yes

Common porcupine

Hystrix brachyura

LC 8 30.34 0.33 No

Three–striped ground squirrel Lariscus insignis

LC 8 9.95 0.10 No

Long–tailed macaque

Macaca fascicularis

LC 3 0.24 0.01 Yes

Pigtail Macaque

Macaca nemestrina

VU 8 43.76 0.63 No

Sunda pangolin

Manis javanica

CE 7 1.31 0.04 No

Red muntjac

Muntiacus muntjak

LC 8 71.99 0.71 Yes

0 RAB

0.10

0.07

0.08

0.11

0 PAO

< 0.01

Animal Biodiversity and Conservation 43.1 (2020)

103

Table 1. (Cont.)

0 PAO

Scientific name Presbytis melalophos

EN 4 0.48 0.02 Yes

Flying fox

Pteropus vampyrus

NT 0 0.00 0.00 Yes

Black giant squirrel

Ratufa bicolor

NT 0 0.00 0.00 Yes

Sambar deer

Rusa unicolor

VU 8 9.69 0.18 Yes

Horse–tailed squirrel

Sundasciurus hippurus

Wild boar

Sus scrofa

LC 8 29.4 0.39 Yes

Siamang

Symphalangus syndactylus

Malayan tapir

Tapirus indicus

EN 8 8.07 0.15 Yes

Silvery lutung

Trachypithecus cristatus

NT 0 0.00 0.00 Yes

Lesser mouse deer

Tragulus kanchil

LC 4 21.99 0.23 No

Greater mouse deer

Tragulus napu

LC 8 12.55 0.17 No

Long–tailed porcupine

Trichys fasciculata

LC 2 0.63 0.01 No

Large treeshrew

Tupaia tana

LC 4 1.37 0.03 No

IUCN

0 RAB

Common name Sumatran surili

1 1

0.07 0.14

< 0.01 < 0.01

No Yes

Allen et al.

104

A Δ1 = 0.41 (0.23–0.60)

Tiger As. golden cat

0.20

Δ1 = 0.29 (0.11–0.48)

0.10

Tiger Leopard cat

0.10 0.0

0.0 00:00 06:00 12:00 18:00 24:00

00:00 06:00 12:00 18:00 24:00

Δ1 = 0.20 (0.41–0.93) C

Δ1 = 0.29 (0.09–0.54)

Tiger Marbled cat

0.10

0.0

0.0 00:00 06:00 12:00 18:00 24:00

00:00 06:00 12:00 18:00 24:00

E Δ1 = 0.21 (0.05–0.41)

Δ1 = 0.25 (0.11–0.41)

Tiger Banded linsang

0.10

Tiger S. clouded leopard

Tiger Banded palm civet

0.10

0.0

00:00 06:00 12:00 18:00 24:00 G Δ1 = 0.18 (0.03–0.35) 0.20

Tiger Binturong

Δ1 = 0.60 (0.42–0.77) Tiger Dhole

0.20 0.10

0.10

0.0 00:00 06:00 12:00 18:00 24:00

0.0 00:00 06:00 12:00 18:00 24:00 Δ1 = 0.26 (0.09–0.46) I 0.10

00:00 06:00 12:00 18:00 24:00

Tiger Masked palm civet

Δ1 = 0.60 (0.42–0.77)

J 0.12

Tiger Sun bear

0.06 0.0

0.0 00:00 06:00 12:00 18:00 24:00 00:00 06:00 12:00 18:00 24:00 Time of day

Animal Biodiversity and Conservation 43.1 (2020)

105

106

Allen et al.

Animal Biodiversity and Conservation 43.1 (2020)

107

Muñoz and Farfán

Animal Biodiversity and Conservation 43.1 (2020)

109

Effectiveness of the aposematic Eumaeus childrenae caterpillars against invertebrate predators under field conditions N. Ruiz–García

Ruiz–García, N., 2020. Effectiveness of the aposematic Eumaeus childrenae caterpillars against invertebrate predators under field conditions. Animal Biodiversity and Conservation, 43.1: 109–114, Doi: https://doi. org/10.32800/abc.2020.43.0109 Abstract Effectiveness of the aposematic Eumaeus childrenae caterpillars against invertebrate predators under field conditions. Eumaeus childrenae (Lepidoptera, Lycaenidae) caterpillars are specialist herbivores that feed on Dioon holmgrenii (Cycadacea). They are a well–documented case of chemical protection by sequestering cycasin and related compounds from their host. In this study we evaluated the effectiveness of aposematic defenses against chemical and visual invertebrate predators in wild populations of E. childrenae reared on D. holmgrenii. The results from field experiments indicated that the estimated survival and the intrinsic rate of increase in cohorts with predator exclusion were twice those in cohorts without predator exclusion. The visually oriented predators observed were Mischocyttarus wasps and assassin bugs, and the chemically oriented predators were Wasmannia, Crematogaster and Ectatomma ants. Other mortality factors observed were egg cannibalism, nuclear polyhedrosis virus and, reported for the first time, larval parasitism by fly larvae and a fungus. Key words: Cycasin, Chemical defense, Chemical ecology, Tritrophic interaction, Demography, Cycads Resumen Eficacia de las orugas aposemáticas de Eumaeus childrenae contra depredadores invertebrados en el medio natural. Las orugas de Eumaeus childrenae (Lepidoptera, Lycaenidae) son herbívoros especialistas que se alimentan de Dioon holmgrenii (Cycadacea). Constituyen un caso bien documentado de protección química mediante el secuestro de cicasina y de compuestos relacionados de su hospedante. En este estudio evaluamos la eficacia del aposematismo contra depredadores invertebrados con mecanismos químicos y visuales de caza en una población silvestre de E. childrenae criada en D. holmgrenii. Los resultados de los experimentos de campo indican que la supervivencia estimada y la tasa de cambio intrínseca fueron el doble en las cohortes con exclusión de depredadores que en aquellas donde no se excluyó a los depredadores. Los depredadores con mecanismos de caza visual observados fueron avispas del género Mischocyttarus sp. y chinches, mientras que los depredadores con mecanismos químicos de caza fueron varias especies de hormigas de los géneros Wasmannia, Crematogaster y Ectatomma. Otros factores de mortalidad observados fueron el canibalismo de huevos, el virus de la poliedrosis nuclear y el parasitismo de las larvas por larvas de moscas y un hongo, que se observan por primera vez. Palabras clave: Cicasina, Defensa química, Ecología química, Interacción tritrófica, Demografía, Cícadas Received: 10 VI 19; Conditional acceptance: 11 X 19; Final acceptance: 22 XI 19 Noe Ruiz–García, Instituto de Ecología, Universidad del Mar, km 3.5 carretera Puerto Escondido–Oaxaca, San Pedro Mixtepec, Oaxaca 71980, México. ORCID: https://orcid.org/0000–0001–7319–9311 E–mail: nruizg@zicatela.umar.mx

ISSN: 1578–665 X eISSN: 2014–928 X

Ruiz–García

110

Introduction Chemically defended animals often deter predators by stimulating their visual and chemical sensory channels (Rojas et al., 2018). The aposematic coloration of prey is a defense mechanism against visual predators (Dell'Aglio et al., 2016). Aposematic prey are under positive frequency–dependent selection, whereby the efficiency of a defense signal increases in relation to local abundance and background (Seymoure et al., 2018). Chemically protected or unpalatable prey have chem2 s ical compounds with a noxious taste or odor that deters predators (Brower, 1984). Once predators have tasted their toxins, they will usually avoid them (Leimar et al., 1986). Plant–derived chemical defense has been documented in many lepidopteran species (Opitz and Müller, 2009), such as the Eumaeus species. This species sequesters cycasin and related compounds from its cycad plant hosts (Rothschild et al., 1986; Bowers and Larin, 1989; Nash et al., 1992). The butterfly Eumaeus childrenae Gray is a specialist herbivore that feeds on Dioon holmgrenii De Luca, Sabato and Vázquez Torres (1981) an endemic cycad species of the physiographic province Sierra Madre del Sur and the Pacific Coastal Plain of the state of Oaxaca, Mexico (Cervantes–Zamora et al., 1990). The cycad D. holmgrenii is considered threatened and endangered due to its low genetic diversity (González–Astorga et al., 2008) and disturbance of its habitat (Velasco–García et al., 2016). Previous research confirms that this species of the genus Eumaeus sequesters and accumulates cycacin when feeding on cycad leaves and strobili (Rothschild et al., 1986; Bowers and Larin, 1989). Observational and feeding experiments suggest that cycasin in eggs and larvae produces a deterrent effect against ants (Bowers and Larin, 1989; Castillo–Guevara and Rico–Gray, 2002) and makes E. atala adults unpalatable to birds (Bowers and Farley, 1990). It thus provides chemical protection against predators throughout all stages of the E. atala life cycle (Rothschild et al., 1986). This chemical protection is signaled to its predators through the red coloration of the larvae (Schneider et al., 2002) that is magnified by their gregarious behavior (Bowers and Farley, 1990). While these laboratory studies demonstrate that the sequestration of cycasin and related compounds are potentially important determinants of predation rates, no field experiments have yet been performed to test the deterrent effect of cycasin and aposematism against other predators such as mites and spiders (Beltrán–Valdez and Torres–Hernández, 1995) or against unreported predators, parasites and diseases in wild populations. We investigated the effectiveness of specific defenses against multiple species (Camara, 1997) and report the results of field experiments that measured the efficacy of aposematic defenses against chemical and visual predators in wild populations of E. childrenae reared on its natural host plant, D. holmgrenii. We used the survival rate from oviposition until adult emergence and the intrinsic rate of reproduction to measure deterrent efficiency against all invertebrate preda-

tors and parasites in the life cycle of E. childrenae in two conditions: with and without predator exclusion. An additional aim was to investigate the possible presence of new predators and parasites in the Neotropical region. Material and methods Study area The demographic parameters of E. childrenae were determined in the locality of D. holmgrenii located at kilometer 244 of the Oaxaca to Puerto Escondido highway, at 16º 1' 47.1'' N and 97º 3' 59.9'' W. It has an area of approximately 4.9 hectares, at 650 m to 850 m a.s.l.. The climate is warm and subhumid with rains in summer, the most humid of the subhumid, with percentage of winter rain less than 5 [Aw2(w)]. The average annual temperature is 26 ºC and the average annual rainfall is 1,500 mm (INEGI et al., 2008). Vegetation corresponds to a transition zone between oak forest and sub–deciduous forest. Cohorts From June 2015 to February 2017 we recorded 29 cohorts with 1404 eggs on accessible host plants. Survival rates until adult emergence were estimated in 21 cohorts and 1,215 eggs; eight cohorts that were completely destroyed or that moved to inaccessible plants were excluded. The cohorts were grouped according to two rearing conditions: without predator exclusion (WPE) (14 cohorts with 819 eggs on leaves and 5 cohorts with 283 eggs on strobili), and with predator exclusion (PE) (2 cohorts with 113 eggs on leaves). The PE cohorts were placed in exclusion using a cylindrical cage (1 m long per 0.5 m diameter). The cage was wired and covered with a fine mesh that excluded predators such as ants, wasps, bugs and birds. The caged larvae were fed host cycad leaves cut from the most recent flush of leaves. The pupae of three cohorts that grew on WPE leaves were kept in net screen cages (1.25 × 1.25 × 1.25 m) until adult emergence. In both cages, the sex ratio and the number of oviposited eggs were recorded in order to calculate the fertility table. Each cohort was visited every day to record the time of molting. The behavioral response of the larvae when attacked by the predators was recorded. The behavior of each predator species observed attacking larvae was also recorded, including predation on other prey. Digital photographs were taken to document the predators observed in each stage. Data analysis The survival profile was estimated using the Kaplan– Meier nonparametric estimator. A pairwise comparison of survival profiles considering reared conditions was used with a Mantel–Haenszel test. The observed significance level was adjusted for the number of comparisons using the Bonferroni procedure (Lee and Wang, 2003). The fertility table was built using a simple

Animal Biodiversity and Conservation 43.1 (2020)

1.0

0.0

20 30 40 Days after oviposition

Adult

Pupae

Pre–pupae

0.2

Instar III

Instar II

Instar I

Survival proportion

0.4

Instar IV

Egg

PE leaves WPE leaves WPE strobili

0.8

0.6

111

Fig. 1. Survival rate of Eumaeus childrenae from egg to adult emergence in leaf and stroboli of Dioon holmgrenii under field conditions in the physiographic province of Sierra Sur, Oaxaca, Mexico: PE, with predation exclusion; WPE, without predation exclusion. Fig. 1. Índice de supervivencia de Eumaeus childrenae desde el huevo hasta la aparición del adulto en hojas y estróbilos de Dioon holmgrenii en el medio natural en la provincia fisiográfica de Sierra Sur, Oaxaca, en México: PE, con exclusión de depredadores; WPE, sin exclusión de depredadores.

decrement abridged life table grouped by stages, the number of eggs, and the sex ratio observed in cages. The net reproduction rate (R0), mean generation time (TC) and intrinsic rate of increase (r . ln R0 · TC–1) were estimated following Carey (1993). Results The survival rate up until to adult emergence differed between cohorts on WPE leaves, WPE strobili, and PE leaves (x2 = 16.4, p < 0.000272). The survival rate (0.230 ± 0.039) was highest on PE leaves, followed by WPE leaves (0.15 ± 0.039) and WPE strobili (0.046 ± 0.012). Figure 1 shows the survival profile for each rearing condition. We merged the survival rates on WPE leaves and WPE strobili because they were similar (x2 = 0.8, p < 0.374). The WPE survival rate (0.123 ± 0.029) was lower (x2 = 12.5, p < 0.001) than that in the PE cohorts. In the PE cohorts, the net reproduction rate was 0.436 individuals per female, with a generation time of 53.2 days; the estimated intrinsic rate of increase was –4.8. In the WPE cohorts, the net reproduction rate was 0.244 individuals per female, with a generation time of 54 days, and the estimated intrinsic rate of increase was –5.4. In both cages, the observed sex ratio was 1:1 (exact binomial test: 0.437, p = 0.261). The average number of oviposited eggs on leaves was 54.6 ± 6.7, and 69.7 ± 6.7 on strobili.

Mortality factors observed on eggs were lack of hatching in complete cohorts of up to 65 eggs, and the sucking of deutoplasm through the micropyle by recruiting and mass attack by ants of the genera Wasmannia sp. and Crematogaster sp. (Hymenoptera, Formicidae) (fig. 1s–A, 1s–B in supplementary material). The attacks by these two species of ants destroyed complete cohorts of up to 93 eggs during the rainy season (June to September). These ants carry the whole dead caterpillar to their nest (fig. 1s–C in supplementary material). They were not observed attacking live caterpillars. Complete destruction of cohorts also occurred as a result of egg cannibalism by first instar larval stages (fig. 3s in supplementary material). Also, a beetle of the genus Dasydactylus sp. (Coleoptera, Languriidae) occasionally ate the eggs, biting and killing larvae and pupae (fig. 2s in supplementary material). In the first and second instar larval stages we observed greater natural mortality in both conditions, that is, both with and without predator exclusion cages. One possible explanation for this observation is the antiherbivory defense of cycasyn and related compounds, but this antiherbivory effect has not been documented previously. First and second instar larvae were predated by solitary ants of the genus Ectatomma tuberculatum (Hymenoptera, Formicidae) (fig. 1s–D in supplementary material). At the beginning of the dry season (October), assassin bug nymphs sucked the haemolymph of third and

112

fourth instar larvae, prepupae and pupae, leaving behind dry exoskeletons (fig. 5s in supplementary material). A species of wasp Mischocyttarus sp. and Mischocyttarus mexicanus ssp. mexicanus (Hymenoptera, Vespidae) were also observed killing third and fourth instar larvae and prepupae by biting caterpillars, which were then chewed and swallowed (fig. 4s in supplementary material). These wasps killed up to 27 % of one cohort. This wasp partitioned the prey into chunks for manageable mastication and consumption. They did not sting the caterpillars. Mortality of third and fourth instar larvae was observed during the rainy season, showing typical symptoms of polyhedrosis virus (fig. 6s–A in supplementary material). The dead larvae were transported by ants of the genus Crematogaster sp. to their nest. Prepupal mortality was observed at the beginning of the dry season, showing typical symptoms of polyhedrosis virus, which caused 40 % mortality in one cohort (fig. 6s–B in supplementary material). Parasitism of pupae by fly larvae was also observed during the rainy season (fig. 8A–8B in supplementary material). A tachinid adult fly was observed visiting the pupae, but parasitized pupae were not caged in order to obtain adult parasitoids. Parasitism by a fungus in prepupae and pupae was observed during the rainy season (fig. 7s–C, 7s–8D in supplementary material). The parasitized pupae were not collected to identify the fungus. Furthermore, adult individuals were observed trapped in spider webs. Discussion The bright red colouring of Eumaeus species, which is magnified by the gregarious habits of the larvae, is produced by cycasin and protects against predators (Bowers and Larin, 1989; Nash et al., 1992; Schneider et al., 2002). However, the relevance of the deterrent effect of chemical protection against predators on the survival and reproduction rate of Eumaeus has not yet been determined. The cluster of white eggs, which are quite visible on the leaves and strobili of D. holmgrenii, are similar to those reported on E. atala reared in Zamia integrifolia lf, a case of egg protection by aposematism (Rothschild et al., 1986; Schneider et al., 2002). This hypothesis could be correct in the case of visual predators, because only chemically–oriented predators such as ants were observed consuming and transporting eggs. Even though the deterrent effect of cycasin against ants has been verified in laboratory assays (Bowers and Larin, 1989; Castillo–Guevara and Rico–Gray, 2002), the continuous foraging activity of individuals of Wasmannia sp. and Crematogaster sp. means that when the first individuals stop feeding due to the deterrent effect, other individuals arrive and continue the foraging until the entire cohort is destroyed. It has been documented that ants use chemical detection mechanisms to hunt (Greeney et al., 2012). Thus, cycasin content in eggs would be expected to have a deterrent effect. However, ants have been found to be the main

Ruiz–García

predators of Eumaeus species (Castillo–Guevara and Rico–Gray, 2002; Smith, 2002), not only consuming deutoplasm, but also transporting E. childrenae eggs and larvae (both dead and alive) to their nests. The social ant predators were not observed attacking live caterpillars, but transporting dead caterpillar to their nest, suggesting that cycasin was not an important defense against social ant predators. One possibility is that the larvae size deterrent social ants from attacking. The social ants have been observed attacking smaller live larvae (fig. 1s–E in supplementary material), such as Rhopalotria sp. larvae, which also contain cycasin inside intact idioblast cells in their gut (Vovides, 1991). Conversily, the aposematism and cycasin do not deter freely foraging ants from attacking first and second instar larvae that are gregarious. E. tuberculatum ants are common pretators on Dioon plants and male strobili. They capture live larvae and adult Rhopalotria sp., Pharaxonotha sp. and bees that using strobili (fig. 1s–F in supplementary material). Furthermore, in another species cosidered protedted by aposematism, it has been documented that the larvae have others morphological defenses that deter ant (de la Fuente et al., 1994/1995; Dyer, 1997; Massuda and Trigo, 2009). Like other plant–feeding Lepidoptera (Lefèvre et al., 2012), newly hatched caterpillars often feed initially on their chorion, but when there are two or more clutches of eggs, the caterpillars that hatched first fed on the unhatched eggs of other clutches (cannibalism), devouring a complete cohort of 113 eggs. This behavior has been reported in other species of Lepidoptera (Barros–Bellanda and Zucoloto, 2001). The bright red colouring and its magnification by agglomeration constitute a defense mechanism against predators such as birds and mantids (Greeney et al., 2012), wasps and assassin bugs (Johnson, 1983; Dyer, 1997) that rely on visual predation mechanisms (Ruxton and Sherratt, 2006). In the present field study, we did not observe bugs and wasps attacking the gregarious first and second instar larvae. Both wasps and bugs were observed attacking the third and fourth instar larvae, which are larger and live solitarily. However, upon contact with wasps or bugs, caterpillars showed active responses, including escape and lifting their posterior segments. In contrast, Junonia coenia Hübner, another caterpillar considered protected by its aposematism, shows no behavioral defenses when attacked by the wasp Polistes fuscatus Fabricius (Stamp, 1992). In another study, wasps (Polistes instabilis Saussure) were deterred by brightly colored prey (Dyer, 1997). Both previously cited studies support general theories about aposematism. However, our field observations indicate that both species of Mischocyttarus wasp are not dettered by aposematism because the feed on larvae and pre–prepupae during moulting process, when they are aggregated and immobile (fig. 4s–C.1, 4s–C.2 in supplementary material). This observation has not been reported previously and suggests that cycasin and aposematic colouring are not useful defense against these wasps. Even

Animal Biodiversity and Conservation 43.1 (2020)

higher predation was reported in another aposematic caterpillar, Hemileuca lucina (Saturnidae), with the wasp Polistes sp., which killed up to 99 % of one cohort (Stamp and Bowers, 1988). The only larval disease reported is the polyhedrosis virus (Koi and Daniels, 2015). No other diseases or parasites have been reported to affect the Eumaeus species. Pupae were parasitized by fly larvae during the rainy season. Flies are generally considered generalist parasitoids that are little affected by plant allelochemicals sequestered by their larval host (Mallampalli et al., 1996). Fungi and the polyhedrosis virus were also observed parasitizing pupae during the rainy season. There are no previous experiments about unpalatable effects of deterrent contents in the pupae against predators, parasites and diseases. A possible explanation for the parasitism observed in the present study is the increased susceptibility of pupae due to the decrease in their cycasin content. This decrease amounted to 75 % compared to the content of cycasin in fourth instar larvae (Schneider et al., 2002). The survival rate until adult emergence in the WPE cohorts was half that observed in the PE cohorts. In addition, the estimated intrinsic rate of increase was greater in PE cohorts. Thus, we suggest that chemical protection was insufficient to protect E. childrenae and that the observed mortality factors are important top–down forces against this herbivore in wild conditions. This could also be true for the other species of the genus Eumaeus, since the average mortality rate observed in the present study was similar to the values reported for eggs (31.25 %) and larvae (64.9 %) of Eumaeus atala Godart reared on the cycad Zamia loddigesii Miq. (Castillo–Guevara and Rico–Gray, 2002), and to E. atala survival rate (17.1 %) from egg to pupae when reintroduced to its host Z. pumila (Smith, 2002). Although aposematic coloration and chemical defenses are well–known defenses against predators in the Eumaeus species, our field data suggest that some invertebrate predators are not deterred by warning coloration or chemical protection. Also, our results suggest that the ecological importance of aposematic coloration and chemical defenses as protective mechanisms are overestimated and other strategies against enemies are undervalued, because caterpillars also show strategies such as behavioral responses against predators. Upon contact with wasps or bugs, caterpillars show behavioral responses against enemies, mainly evasive responses, including escape and lifting their posterior segments. Another defense observed in the field is the hiding position of feeding larvae (e.g. under leaflets and sporophylls) to avoid encounters with predators. Acknowledgements We thank the two anonymous reviewers who provided helpful comments to improve the final version of this paper.

113

References Barros–Bellanda, H. C. H., Zucoloto, F. S., 2001. Inflence of chorion ingestion on the performance of Ascia monuste and its association with cannibalism. Ecological Entomology, 26: 557–561. Beltrán–Valdez, J. M., Torres–Hernández, L., 1995. Aspectos cualitativos de la herbivoría de Eumaeus debora Hubner. (Lepidoptera: Lycaenidae) sobre Ceratozamia mexicana Bröng. (Cycadales: Zamiaceae). La ciencia y el Hombre, 19: 91–104. Bowers, M. D., Farley, S., 1990. The behaviour of grey jays, Perisoreus canadensis, towards palatable and unpalatable Lepidoptera. Animal Behaviour, 39: 699–705. Bowers, M. D., Larin, Z., 1989. Acquired chemical defense in the lycaenid butterfly, Eumaeus atala. Journal of Chemical Ecology, 15: 1133–1146. Brower, L. P., 1984. Chemical defense in butterflies. In: The biology of butterflies: 109–132 (R. I. Vane–Wright, P. R. Ackery, Eds.). Academic Press, London. Camara, M. D., 1997. Predator responses to sequestered plant toxins in buckeye caterpillars: are tritrophic interactions locally variable? Journal of Chemical Ecology, 23: 2093–2106. Carey, J. R., 1993. Applied demography for biologists with special emphasis on insects. Oxford University Press, New York. Castillo–Guevara, C., Rico–Gray, V., 2002. Is cycasin Eumaeus minyas (Lepidoptera: Lycaenidae) a predator deterrent? Interciencia, 27: 465–470. Cervantes–Zamora, Y., Cornejo–Olgín, S. L., Lucero– Márquez, R., Espinoza–Rodríguez, J. M., Miranda– Viquez, E., Pineda–Velázquez, A., 1990. Provincias fisiográficas de México, escala 1:4000000. In: Clasificación de regiones naturales de México 1. Tomo II, Sección IV, 10.1. Atlas nacional de México (1990–1992). Instituto de Geografía, UNAM, México. de la Fuente, M., Dyer, L. A., Bowers, M. D., 1994/1995. The iridoid glycoside, catalpol, as a deterrent to the predator Camponotus floridanus (Formicidae). Chemoecology, 5/6: 13–18. Dyer, L. A., 1997. Effectiveness of caterpillar defenses against three species of invertebrate predators. Journal of Research on the Lepidoptera, 34: 48–68. Dell'Aglio, D. D., Stevens, M., Jiggins, C. D., 2016. Avoidance of an aposematically coloured butterfly by wild birds in a tropical forest. Ecological Entomology, 41: 627–632. González–Astorga, J., Vergara–Silva, F., Vovides, A. P., Nicolalde–Morejón, F., Cabrera–Toledo, D., Pérez–Farrera, M. A., 2008. Diversity and genetic structure of three species of Dioon Lindl. (Zamiaceae, Cycadales) from the Pacific seaboard of Mexico. Biological Journal of the Linnean Society, 94: 765–776. Greeney, H. F., Dyer, L. A., Smilanich, A. M., 2012. Feeding by lepidopteran larvae is dangerous: A review of caterpillars' chemical, physiological, morphological, and behavioral defenses against natural enemies. Invertebrate Survival Journal, 9: 7–34.

114

INEGI, CONABIO, INE, 2008. Ecorregiones terrestres de México. Escala 1:1000000. México. Johnson, L. K., 1983. Apiomerus pictipes. In: Costa Rican natural history: 684–687 (D. H. Janzen, Ed.). University of Chicago Press, Chicago. Koi, S., Daniels, J., 2015. New and revised life history of the Florida hairstreak Eumaeus atala (Lepidoptera: Lycaenidae) with notes on its current conservation status. Florida Entomologist, 98: 1134–1147. Lee, E. T., Wang, J. W., 2003. Statistical methods for survival data analysis. John Wiley & Sons, Inc., New Jersey. Lefèvre, T., Chiang, A., Kelavkar, M., Li, H., Li, J., Lopez–Fernandez de Castillejo, C., Oliver, L., Potini, Y., Hunter, M. D., de Roode, J. C., 2012. Behavioural resistance against a protozoan parasite in the monarch butterfly. Journal of Animal Ecology, 81(1): 70–79. Leimar, O., Enquist, M., Sillen–Tullberg, B., 1986. Evolutionary stability of aposematic coloration and prey unprofitability: a theoretical analysis. The American Naturalist, 128: 469–490. Mallampalli, N., Barbosa, P., Weinges, K., 1996. Effects of condensed tannins and catalpol on growth and development of Compsilura concinnata (Diptera: Tachinidae) reared in gypsy moth (Lepidoptera: Lymantriidae). Journal of Entomological Science, 31: 289–300. Massuda, K. F., Trigo, J. R., 2009. Chemical defence of the warningly coloured caterpillars of Methona themisto (Lepidoptera: Nymphalidae: Ithomiinae). European Journal of Entomology, 106: 253–259. Nash, R. J., Bell, A. A., Ackery, P. R., 1992. The protective role of cycasin in cycad–feeding lepidoptera. Phytochemistry, 31: 1955–1957. Opitz, S. E. W., Müller, C., 2009. Plant chemistry and insect sequestration. Chemoecology, 19: 117–154. Rojas, B., Burdfield–Steel, E., De Pasqual, C., Gordon, S., Hernández, L., Mappes, J., Nokelainen, O., Rönkä, K., Lindstedt, C., 2018. Multimodal

Ruiz–García

aposematic signals and their emerging role in mate attraction. Frontiers in Ecology and Evolution, 6: 93. Rothschild, M., Nash, R. J., Arthur, B. E., 1986. Cycasin in the endangered butterfly Eumaeus atala. Florida. Phytochemistry, 25: 1853–1854. Ruxton, G. D., Sherratt, T. N., 2006. Aggregation, defence and warning signals: the evolutionary relationship. Proceedings of the Royal Society B, 273: 2417–2424. Schneider, D., Wink, M., Sporer, F., Lounibos, P., 2002. Cycads: their evolution, toxins, herbivores and insect pollinators. Naturwissenschaften, 89: 281–294. Seymoure, B. M., Raymundo, A., McGraw, K. J., Owen–McMillan, W., Rutowski, R. L., 2018. Environment–dependent attack rates of cryptic and aposematic butterflies. Current Zoology, 64: 663–669. Smith, E. M., 2002. The effects of season, host plant protection, and ant predators on the survival of Eumaeus atala (Lycaenidae) in re–establishments. Joumal of the Lepidopterists' Socciety, 56: 272–276. Stamp, N. E., 1992. Relative susceptibility to predation of two species of caterpillar on plantain. Oecologia, 92: 124–129. Stamp, N. E, Bowers, M. D., 1988. Direct and indirect effects of predatory wasps (Polistes sp.: Vespidae) on gregarious caterpillars (Hemileuca lucina: Saturniidae). Oecologia, 75: 619–624. Velasco–García, M. V., Valdez–Hernández, J. I., Ramírez–Herrera, C., Hernández–Hernández, M. L., López–Upton, J., López–Mata, L., López– Sánchez, H., 2016. Estructura, heterogeneidad de estadios y patrón de dispersión espacial de Dioon holmgrenii (Zamiaceae). Botanical Sciences, 94: 75–87. Vovides, A. P., 1991. Cone ideoblasts of eleven cycad genera: morphology distribution and significance. Botanical Gazette, 152: 91–99.

Animal Biodiversity and Conservation 43.1 (2020)

Supplementary material

Fig. 1s. Depredation of egg and caterpillar of Eumaeus childrenae by ants: A, the Wasmannia sp. ants sucking deutoplasm through the micropyle; B, the Crematogaster sp. predating eggs; and C, transport dead caterpillar; D, Ectatomma tuberculatum ants transport live caterpillar; E, social ants feed on live larvae of Rhopalotria sp. and other unidentified larvae; F, E. tuberculatum ants feed on live caterpillar and adult of Rhopalotria sp. and bees that visit male strobili. Fig. 1s. Depredación de los huevos y las orugas de Eumaeus childrenae por hormigas: A, hormigas del género Wasmannia succionando deutoplasma a través del micrópilo; B, hormigas Crematogaster sp. depredando huevos; C, transporte de un oruga muerta; D, una hormiga de Ectatomma tuberculatum transportando una oruga viva; E, hormigas sociales alimentándose de larvas vivas de Rhopalotria sp. y otras larvas sin identificar; F, hormigas de E. tuberculatum alimentándose de orugas y adultos vivos de Rhopalotria sp. y abejas que visitan los estróbilos macho.

Ruiz–García

Fig. 2s. A, depredation of eggs by beetle Dasydactylus sp.; B, killing and ate fourth instar caterpillar; C, Ate on exuviae of second instar larval. Fig. 2s. A, escarabajo Dasydactylus sp. alimentándose de huevos; B, matando y alimentándose de una oruga de cuarto instar; C, alimentándose de la exuvia de una larva de segundo instar.

Fig. 3s. Cannibalism of egg by first instar larval. Fig. 3s. Canibalismo de huevos por larvas de primer instar.

Animal Biodiversity and Conservation 43.1 (2020)

Fig. 4s. Predation of caterpillar by wasp: A, Mischocyttarus sp. bitter third instar larval; B, consuming caterpillar of second instar; C.1, Mischocyttarus mexicanus ssp. mexicanus; and C.2, Mischocyttarus sp. biting and killing third instar caterpillar during moulting time. Fig. 4s. Una avispa alimentĂĄndose de una oruga. A, Mischocyttarus sp. mordiendo una larva de tercer instar; B, consumiendo una oruga de segundo instar; C.1, Mischocyttarus mexicanus ssp. mexicanus; y C.2, Mischocyttarus sp. mordiendo y matando orugas de tercer instar durante el perĂodo de muda.

iii

Muñoz and Farfán

Animal Biodiversity and Conservation 43.1 (2020)

115

Nevado de Toluca: habitat for Romerolagus diazi? O. Monroy–Vilchis, A. A. Luna–Gil, A. R. Endara–Agramont, M. M. Zarco–González, G. A. González–Desales

Monroy–Vilchis, O., Luna–Gil, A. A., Endara–Agramont, A. R., Zarco–González, M. M., González–Desales, G. A., 2020. Nevado de Toluca: habitat for Romerolagus diazi? Animal Biodiversity and Conservation, 43.1: 115–121, Doi: https://doi.org/10.32800/abc.2020.43.0115 Abstract Nevado de Toluca: habitat for Romerolagus diazi? The volcano rabbit (Romerolagus diazi), also known as teporingo or zacatuche, is a small rabbit that is endemic to Mexico. In this study we characterized its potential habitat in the Area of Protection of Flora, and Fauna Nevado de Toluca, Mexico. Between April 2016 and November 2017, we sampled 1,807 units to determine the presence of this species using indirect evidence. We found dung pellets that could be attributed to R. diazi in 41 (2.27 %) of the sampled units. In 10 % of these units, we set up camera traps to confirm the presence of the species. Sites with presumed R. diazi pellets were characterised by rocky terrain, with Pinus hartwegii as the dominant tree species, and Festuca tolucensis as the dominant grass. Overall herbaceous cover was over 70 %. Sites observed to have a negative effect on the presence of the pellets were areas with livestock grazing and induced burning. The results of camera trapping did not reveal the presence of R. diazi in Nevado de Toluca. Key words: Teporingo, Conservation, High mountain forest Resumen El Nevado de Toluca: ¿un hábitat para Romerolagus diazi? El conejo de los volcanes (Romerolagus diazi), también conocido como teporingo o zacatuche, es un pequeño conejo endémico de México. En este estudio se caracterizó su hábitat potencial en el Área de Protección de Flora y Fauna Nevado de Toluca, en México. Entre abril de 2016 y noviembre de 2017, se analizaron 1.807 unidades de muestreo para determinar la presencia de esta especie de forma indirecta. Encontramos excrementos atribuibles a R. diazi en 41 unidades de muestreo (el 2,27 %). En el 10 % de estas unidades se colocaron cámaras de trampeo para confirmar la presencia de la especie. Los sitios con presencia de excrementos supuestamente pertenecientes a R. diazi están en terrenos rocosos, donde la especie dominante arbórea es Pinus hartgewii y la herbácea, Festuca tolucensis. El porcentaje de cobertura herbácea fue superior al 70 %. Se observó que el pastoreo y los incendios inducidos tienen un efecto negativo en la presencia de los excrementos. Los resultados del muestreo con cámara no revelaron la presencia de R. diazi en el Nevado de Toluca. Palabras clave: Teporingo, Conservación, Bosque de alta montaña Received: 31 V 19; Conditional acceptance: 22 VII 19; Final acceptance: 02 XII 19 O. Monroy–Vilchis, A. A. Luna–Gil, M. M. Zarco–González, G. A. González–Desales, Centro de investigación en Ciencias Biológicas Aplicadas, Universidad Autónoma del Estado de México, Instituto Literario 100, C.P. 50170, Toluca, México.– A. R. Endara–Agramont, Instituto de Ciencias Agropecuarias y Rurales, Universidad Autónoma del Estado de México, Instituto Literario 100, C.P. 50170, Toluca, México. Corresponding author: M. M. Zarco–González. E–mail: martha.zarco.g@gmail.com

ISSN: 1578–665 X eISSN: 2014–928 X

Monroy–Vilchis et al.

116

Introduction

Material and methods

The diversity of hares and rabbits (Order Lagomorpha, Family Leporidae) is high in Mexico, with 15 species. Seven of these are endemic (Farías, 2011). The volcano rabbit, (Romerolagus diazi Ferrari–Pérez 1893), also known as zacatuche or teporingo, is a small, monospecific rabbit found only in the mountains around the southern part of Mexico City. It lives in small groups in runways among grass tussocks, typical of its distribution area. The volcano rabbit is considered endangered (SEMARNAT, 2010; IUCN, 2017). It is threatened by the loss and fragmentation of its habitat (Hoth et al., 1987; Velázquez et al., 2011), introduced species (dogs and cats), hunting, and more recently, climate change (López et al., 1996; Anderson et al., 2009). The main causes for the loss of habitat are anthropogenic fires, livestock grazing, and logging (López et al., 1996). The volcano rabbit is known to occur along the central Trans–Mexican Volcanic System (TMVS) in discontinuous patches in four volcanoes (Popocatépetl, Iztaccíhuatl, El Pelado and Tláloc) that cover approximately 386 km2 (Velázquez, 1994). Although there is no historical evidence of sightings of the species, some researchers (Velázquez et al., 1996a) have argued that the rabbit could potentially be found in areas of similar habitat in other volcanoes and mountains in the TMVS such as the Nevado de Toluca, Nevado de Colima, Volcán Tancítaro, Cofre de Perote and Pico de Orizaba. However, credible evidence of their presence in these sites is lacking (Velázquez et al., 1996a; Olascoaga et al., 2015). Volcano rabbits are strongly associated with pine–grassland habitat (Cervantes et al., 1990; Fa et al., 1992; Velázquez, 2012). They are generally found where tussock grasses, genera Muhlenbergia, Festuca and Jarava (Velázquez, 1994), are denser in cover and taller (Fa et al., 1992; Rizo et al., 2015), providing the species not only with a source of food but also shelter (Trigo et al., 2003). They may also use trunks and stony mounds as refuge (Cervantes and Martínez, 1996). Volcano rabbits have been found in areas with no or moderate slopes (Martínez–Calderas et al., 2016) where dominant vegetation is Festuca tolucensis (70 % cover)/Pinus spp. (10 %), and Muhlenbergia macroura (70 %)/Pinus spp. (20 %, Velázquez et al., 1996b). In 1975, a volcano rabbit was found in the Nevado de Toluca, in Central Mexico, about 80 km west of Mexico City and near the city of Toluca. It is now deposited in the Instituto Politécnico Nacional Museum (Galindo, C., pers. comm.). There is also anecdotal information regarding sightings in the 1980s, but no evidence was found during a field study in 1987 (Hoth et al., 1987). In 1998, Ceballos et al. (1998) reported finding dung pellets attributable to the species but no further details were given. On the basis of these previous uncertain findings, we attempted to confirm the presence of the volcano rabbit in the Protected Area of Flora and Fauna Nevado de Toluca, focussing on sites with environmental characteristics that are similar to those in its area of distribution.

Study area The Protected Area of Flora and Fauna Nevado de Toluca National Park (PAFFNT) is located within the TMVS between 18º 51' 3'' and 19º 19' 03'' N and 99º 38' 54'' and 100º 09' 58'' W (Brunett et al., 2010). The PAFFNT, covering 53,590 ha (CONANP, 2016), is the fourth highest mountain (4,660 m) in the country (Candeau and Franco, 2007). The area is dominated by coniferous forests [29.5 % pine (Pinus spp.), 35.7 % fir (Abies religiosa)] with some oak Quercus (0.3 %) and alder Alnus jorullensis (3.9 %) and agriculture 17.4 % (CONANP, 2016). In addition, 4.2 % consists of alpine grassland communities (Candeau and Franco, 2007). Pine–grassland habitat occupies 12,924 ha. Sampling Between 22 April 2016 and 05 July 2017 we searched for the presence of the volcano rabbit in pine–grassland habitat in the PAFFNT (fig. 1). We established 1807 sampling units (SU) above 3,000 m (Dauber, 1995) situated along every 100 m (3,100, 3,200, etc). Each SU consisted of a circular plot of 0.1 ha (17.86 m radius) to guarantee a minimum sampling intensity (recommended 0.44 % for 12,924 ha (Dauber, 1995). In each SU we noted the following qualitative variables: presence/absence of rocky outcrops, evidence of recent fire (< 1 year), livestock grazing, logging, and presence of ravines, trails or reforestation. We also recorded the dominant arboreal species (DAS), dominant herbaceous species (DHS), dominant shrub species (DSS) and density of the forest, classified as: dense (80–100 %), semi–dense (50–79 %) and open (< 50 %; Regíl, 2005). Quantitative variables analysed were altitude (m a.s.l.), slope and aspect (measured in degrees), percentage of herbaceous cover (PHC), percentage of shrub cover (PSC), number of seedlings, and number of tree stumps. Two axes were drawn in the north–south and east–west directions crossing the central point of the SU. Within each SU, we searched for rabbit pellets. Two other lagomorph species are found in the area (Sylvilagus floridanus and S. cunicularius). Rabbit pellets found were photographed to determine whether they corresponded to R. diazi (Aranda, 2012). Using this indirect approach, we analysed each photographic record digitally (with ImageJ software) to verify the dorso–ventrally compressed spherical shape and maximum diameter of 10 mm that is typical for the species (fig. 2). Statistical analysis We analysed the data by dividing the SUs into two groups: one with presence of pellets attributable to R. diazi (41 sites) and another without pellets (1,766 sites). To identify the variables that correctly classified the SU with and without pellets, we performed a stepwise discriminant analysis, with the quantitati-

Animal Biodiversity and Conservation 43.1 (2020)

400000

410000

117

420000

430000

Gulf of Mexico

ex i

Pa c

ifi

2120000 2110000

2110000

2120000

2130000

United States of America

Sampling units with pellets Dense pine forest Semi–dense pine forest Open pine forest Study area

400000

410000

420000

8 km

430000

Fig. 1. Sampling units with pellets attributable to Romerolagus diazi in the Protected Area of Flora and Fauna Nevado de Toluca. Fig. 1. Unidades de muestreo con excrementos atribuibles a Romerolagus diazi en el Área de Protección de Flora y Fauna Nevado de Toluca.

ve variables. We tested the independence of each qualitative variable and the presence of pellets, with contingency tables and x2–tests. Camera trapping In 10 % of the sites with pellets attributable to R. diazi, a camera trap was placed in front of latrines to obtain photographic records of the species using the latrine. The cameras were operated day and night for 6 months (July to December, 2017). The presence of different vertebrate species and their relative abundance was obtained using the following indices: IAR = (C/SE) * 100 trap–days where: C is the number of photographic events and SE, the sampling effort (number of cameras per day of monitoring; Maffei et al., 2002; Jenks et al., 2011; Monroy–Vilchis et al., 2011). Results We compared the environmental variables with those from other studies with confirmed presence of R. diazi: Pelado volcano (Velázquez et al., 1996a), Tláloc volcano (Velázquez and Heil, 1994), Popocatépetl (Velázquez et al., 1996b) and Iztaccíhuatl (Hunter

and Cresswell, 2015; table 1). We found pellets traditionally attributable to R. diazi in 41 SUs (2.27 %), 36 of which were in Pinus spp. forest and 29 in zones with varying density of pine forest (fig. 1). The 41 SUs with pellets attributable to R. diazi were found between 3,305 and 3,874 m, with a higher proportion between 3,500 and 3,700 m (fig. 2). Seven dominant tree species were recorded in these areas: Pinus hartwegii, Alnus jorullensis, Cupressus lusitanica, Abies religiosa, Pinus montezumae, Pinus patula and Pinus pseudostrobus. However, the presence of pellets attributable to the volcano rabbit was recorded in only four Sus (table 2), with P. hartwegii being the most frequent, in 36 SUs (87.80 % of SUs with scat). Three of the seven dominant genera in the herbaceous layer were present in SUs with pellets: Festuca (56.10 %), Muhlenbergia (39.02 %) and Calamagrostis (4.88 %). Within the shrub cover, 16 species were identified, but the pellets atributable to R. diazi were associated with six of these (table 2), in order of frequency: Senecio cinerarioides (26.83 %), Lupinus montanus (24.39 %), Barkleyantus salicifolius (21.95 %), Symphoricarpos microphyllus (2.44 %), Acaena elongata (2.44 %) and Baccharis conferta (2.44 %). Eight SUs did not have any shrub cover (19.51 %). Results of the stepwise discriminant analysis showed that only the percentage of herbaceous cover was a significant predictor for the presence of pellets attributable to the volcano rabbit (Wilk's

Monroy–Vilchis et al.

118

40 35

No SU

30 25 20 15 10 5 2

10 mm

Fig. 2. Diameter of rabbit pellets found in 41 sampling units in the Protected Area of Flora and Fauna Nevado de Toluca. Fig. 2. Diámetro de los excrementos de conejo encontrados en 41 unidades de muestreo en el Área de Protección de Flora y Fauna Nevado de Toluca.

Lambda = 0.99, F = 8.31, p = 0.003). The mean percentage of herbaceous vegetation in SU with pellets (74.85%) was higher than that in SUs without pellets (61.77%). Of the qualitative variables, only two had an influence on the presence of pellets: rocky outcrop (x2 = 8.38, df = 1, p = 0.003) and livestock grazing (x2 = 2.85, df = 1, p = 0.001). Rocky outcrops that favored the presence of pellets traditionally attributable to the volcano rabbit were

recorded in 70.73 % of the SU with pellets. Grazing was recorded in only one SU with pellets, showing the negative influence of livestock herding on pellets attributable to the volcano rabbit. We obtained a total of 23 independent photographic records (IR) out of 8,601 records from a sampling effort of 165 camera trap days. The specific richness was five species of three orders (Passeriformes, Lagomorpha and Rodentia; table 3).

Table 1. Variables associated with the presence / absence of Romerolagus diazi. Tabla 1. Variables asociadas a la presencia y ausencia de Romerolagus diazi. Variable

Other studies

Present study

Presence of R. diazi

3,400–4,000

3,500–3,700

DAS

Pinus hartwegii

DHS

Festuca tolucensis

Festuca spp.

DSS

–

Senecio cinerarioides

Recent fire

Presence of R. diazi

Absence of R. diazi

Livestock grazing

Rocky outcrop Altitude (m a.s.l.)

Presence of R. diazi

Absence of R. diazi

Aspect

–

Northeast hillside

Slope (º)

–

19 (5–38)

PSC (%)

–

16 (0–70)

PHC (%)

> 70

Animal Biodiversity and Conservation 43.1 (2020)

119

Discussion The presence of the volcano rabbit in the Nevado de Toluca has not been confirmed in any scientific study although pellets considered attributable to the species have been reported (Ceballos et al., 1998). In our study, pellets attributable to the volcano rabbit were mostly associated with rocky outcrops, where P. hartwegii was the dominant arboreal species, F. tolucensis was the dominant herbaceous species and the percentage of herbaceous cover was > 70 %. These characteristics are similar to those reported in the Pelado volcano, Mountain Tlaloc and Iztaccíhuatl, where the species is known to exist (Velázquez et al., 1996a, 1996b; Hunter and Cresswell, 2015). The PAFFNT is a relatively small area. It is subject to anthropogenic pressure such as agriculture, livestock grazing, logging and feral species (dogs and cats; CONANP, 2016). However, in our study, SUs with pellets were found in habitat considered optimal for the species since they show environmental characteristics similar to other sites of confirmed distribution (Velázquez, 1996). The presence of the volcano rabbit is related to several factors. The most positive factor is the herbaceous stratum (Fa et al., 1992; Velázquez et al., 1996b) that not only provides protection and refuge from predators (Trigo et al., 2003) but also contributes to diet. Several negative factors have also been described, particularly livestock grazing, controlled burning, and feral dogs (Weber, 2010; García–Aguilar, 2012). In a study conducted in the Iztaccíhuatl volcano, the species was most abundant in the habitat with the highest percentage of grassland and least abundant in areas with more hunting and grazing (Hunter and Cresswell, 2015). This coincides with our results, since the sites with pellets attributed to R. diazi were found in areas with over 70 % of herbaceous cover, mainly Festuca. The incidence of forest fires within the area is another threat to their ecosystems. In the APFFNT 80 % of fires are intentional (PROBOSQUE, 2012). The main type of fire during the last ten years was the superficial or creeping type (CONANP, 2016), whose most severe damage is reflected on the herbaceous

Table 2. Frequencies in sampling units (F) of dominant species/genus for the arboreal, herbaceous and shrub cover. Tabla 2. Frecuencia de las especies o géneros dominantes (F) de la cobertura arbórea, herbácea y arbustiva en las unidades de muestreo. Cover Dominant species / genus

Arboreal P. hartwegii 36 A. jorullensis 3 C. lusitanica 1 P. pseudostrobus 1 Herbaceous Festuca 23 Muhlenbergia 16 Calamagrostis 2 Shrub S. cinerarioides 11 Lupinus montanus 10 B. salicifolius 9 S. microphyllus 1 A. elongata 1 B. conferta

Not shrub cover

cover. In other distribution sites of R. diazi, however, it has been reported that controlled burning has a significant positive effect on the appearance of the rabbit (Hunter and Cresswell, 2015). In the PAFFNT, however, the incidence of recent fires is not related to the presence of pellets.

Table 3. Species registered by camera traps in the sampling units with pellets attributable to R. diazi: IR, independent records; IAR, relative abundance index. Tabla 3. Especies registradas con cámaras en las unidades de muestreo que presentaban excrementos atribuibles a R. diazi: IR, registros independientes; IAR, índice de abundancia relativa. Order

Family

Passeriformes Troglodytidae

Species

Troglodytes aedon

3 0.36

IAR

Emberizidae Oriturus superciliosus

8 0.97

Lagomorpha Leporidae

Sylvilagus floridanus

1 0.12

Sylvilagus cunicularius

7 0.85

Rodentia Cricetidae

Neotoma sp.

4 0.48

120

Pinus hartwegii grows between 3,000 and 4,000 m on high mountain forests in Mexico (Rzedowski, 2006; Endara et al., 2013). Forests of P. hartwegii are closely related to the presence of pellets attributable to R. diazi. Endara et al. (2013) state that the areas with the greatest deterioration dynamics are Pinus spp. High mountain ecosystems are among the most threatened and impacted ecosystems worldwide due to climate change, which is transforming their landscape, substantially changing environmental conditions, and affecting the viability of biodiversity in these areas (Astudillo–Sánchez et al., 2017). Camera trapping is a common technique to study medium and large mammals. In rabbits, it has been used to record their presence and calculate their relative abundance (Monroy–Vilchis et al., 2011; Hernández–Hernández et al., 2018). Although we were unable to confirm the presence of R. diazi in the Nevado de Toluca, pellet data suggest it is present the in PAFFNT. Because the evidence is indirect we consider the occurrence of the species should be verified using direct methods such as photographic records. Further photo–trapping should also be carried out in the other sites with ideal habitat characteristics for the species. Ecological and social studies are needed in these sites to gain further knowledge on how to extend the distribution of the species to the Nevado de Toluca. Acknowledgements We wish to thank the National Forestry Commission for financing the study through the CONAFOR–UAEM 3668 2014E Project and UAEM with the project 4343/2017/CI, all the people of the communities of Nevado de Toluca for sharing their knowledge, and the high mountain research group and students who participated in the field work. We also thank John Fa and two anonymous reviewers for their comments on the manuscript. References Anderson, B., Akçakaya, H., Araújo, M., Fordham, D., Martinez, M., Thuiller, B., Brook, B., 2009. Dynamics of range margins for metapopulations under climate change. Proceedings of the Royal Society of London, Series B, Biological Sciences, 276: 1415–1420. Aranda, S., 2012. Manual para el rastreo de mamíferos silvestres de México. CONABIO, México. Astudillo–Sánchez, C., Villanueva–Díaz, J., Endara– Agramont, A., Nava–Bernal, G., Gómez–Albores, M., 2017. Influencia climática en el reclutamiento de Pinus hartwegii lindl. del ecotono bosque–pastizal alpino en monte Tláloc, México. Agrociencia, 51: 105–118. Brunett, E., Baró, J., Cadena, E., Esteller, M., 2010. Pago por servicios ambientales hidrológicos: caso de estudio Parque Nacional del Nevado de Toluca, México. Ciencia Ergo Sum, 17(3): 286–294.

Monroy–Vilchis et al.

Candeau, D., Franco, S., 2007. Dinámica y condiciones de vida de la población del Parque Nacional Nevado de Toluca (PNNT) en la generación de presión a los ecosistemas circundantes y de impactos ambientales a través de un sistema de información geográfica. Investigaciones Geográficas, 62: 44–68. Ceballos, G., Vieyra, B, Ramírez–Pulido, J., 1998. A recent record of the volcano rabbit (Romerolagus diazi) from the Nevado de Tolcua, state of Mexico. Revista Mexicana de Mastozoología, 2: 149–150. Cervantes, F., Lorenzo, C., Hoffmann, R., 1990. Romerolagus diazi. Mammalian Species, 360: 1−7. Cervantes, F., Martínez, J., 1996. Historia natural del conejo zacatuche o teporingo (Romerolagus diazi). In: Ecología y conservación del conejo zacatuche y su hábitat: 29–40 (A. Velázquez, F. Romero, P. López, Eds.). FCE–UNAM, México D.F. CONANP (Comisión Nacional de Áreas Naturales Protegidas), 2016. Programa de Manejo del Área de Protección de Flora y Fauna Nevado de Toluca. Dauber, E., 1995. Guía práctica y teórica para el diseño de un inventario forestal de reconocimiento. BOLFOR, Santa Cruz, Bolivia. Endara, A., Calderón, R., Nava, G., Franco, S., 2013. Analysis of fragmentation processes in high–mountain forests of the centre of Mexico. American Journal of Plant Sciences, 4: 697–704. Fa, J. E., Romero, F. J., López–Paniagua, J., 1992. Habitat use by parapatric rabbits in a Mexican high–altitude grassland system. Journal of Applied Ecology, 29: 357–370. Farías, G., 2011. Conceptos ecológicos, métodos y técnicas para la conservación de conejos y liebres. In: Temas sobre conservación de vertebrados silvestres en México: 229–248 (O. Sánchez, P. Zamorano, P. Recagno, H. Moya, Eds.). INECC, México. García–Aguilar, M. C., 2012. Monitoreo de la población de perros ferales en la Isla de Cedros, Baja California, y las amenazas a la mastofauna nativa. Acta Zoológica Mexicana, 28(1): 37–48. Hernández–Hernández, J. C., Ruiz–Gutiérrez, F., Vázquez–Arroyo, E., Chávez, C., 2018. Registros notables del grisón (Galictis vittata), nutria de río neotropical (Lontra longicaudis) y conejo (Sylvilagus sp) en la Reserva de la Biosfera La Encrucijada, Chiapas, México. Revista Mexicana de Mastozoología, 1: 8–12. Hoth, J., Velázquez, A., Romero, F., Leon, L., Aranda, M., Bell, D., 1987. The volcano rabbit: a shrinking distribution and a threatened habitat. Oryx, 21: 85–91. Hunter, H., Cresswell, W., 2015. Factors affecting the distribution and abundance of the endangered volcano rabbit Romerolagus diazi on the Iztaccíhuatl volcano, México. Oryx, 49: 366–375. IUCN, 2017. The IUCN Red List of Threatened Species, versión 2017-1, http://www.iucnredlist. org [Accessed on 20 August 2017]. Jenks, K., Chanteap, P., Damrongchainarong, K., Cutter, P., Redford, T., Lynam, A., Howard, J., Leimgruber, P., 2011. Using relative abundance

Animal Biodiversity and Conservation 43.1 (2020)

indices from camera–trapping to test wildlife conservation hypotheses: an example from Khao Yai National Park, Thailand. Tropical Conservation Science, 4: 113–131. López, P., Romero, J., Velázquez, A., 1996. Las actividades humanas y su impacto en el hábitat del conejo zacatuche. In: Ecología y conservación del conejo zacatuche y su hábitat: 119–131 (A. Velázquez, F. J. Romero, P. López, Eds.). FCE– UNAM, México D.F. Maffei, L., Cuéllar, E., Noss, A., 2002. Uso de trampas–cámara para la evaluación de mamíferos en el ecotono Chaco–Chiquitanía. Revista Boliviana de Ecología y Conservación Ambiental, 11: 55–62. Martínez–Calderas, J., Palacio–Núñez, J., Martínez–Montoya, J., Clemente–Sánchez, F., Sánchez–Rojas, G., Olmos–Oropeza, G., 2016. Abundancia relativa y distribución de lagomorfos en el Centro–Norte de México. Agroproductividad, 9(9): 67–72. Monroy–Vilchis, O., Zarco–González, M., Rodríguez– Soto, C., Soria–Díaz, L., Urios, V., 2011. Fototrampeo en la sierra Nanchititla, México: abundancia relativa y patrón de actividad. Revista de Biología Tropical, 59(1): 373–383. Olascoaga, E. A., Córdova, I. A., Rivera, R. A., Soto, M. S., Olivares, P. J., 2015. Rabbit zacatuche (Romerolagus diazi): an endemic species of the valley of Mexico. International Journal of Information Research and Review, 2(3): 504–507. PROBOSQUE (Protectora de Bosques del Estado de México), 2012. Reporte de incendios. Estado de México, México. Regíl, H., 2005. Análisis del cambio de uso de suelo y vegetación para la obtención de la dinámica de perturbación–recuperación de las zonas forestales en el Parque Nacional Nevado de Toluca 1972– 2000. Tesis de licenciatura, Facultad de Geografía, Universidad Autónoma del Estado de México. Rizo, A., Guerrero, J., Mihart, M., Romero, A., 2015. Relationship between the abundance of the endangered volcano rabbit (Romerolagus diazi) and vegetation structure in the Sierra Chichinautzin mountain range, Mexico. Oryx, 49: 360–365. Rzedowski, J., 2006. Vegetación de México, 1ra Edición digital. Comisión Nacional para el Conocimiento y Uso de la Biodiversidad, México. SEMARNAT (Secretaria del Medio Ambiente y Re-

121

cursos Naturales), 2010. Norma Oficial Mexicana NOM–059–SEMARNAT–2010. Ley General de Vida Silvestre. Diario Oficial de la Federación (DOF): 38–39. Trigo, B. N., Chimal, H. A., Heil, G. W., Bobbink, R., Verduyn, B., 2003. Classification and mapping of the vegetation using field observations and remote sensing. In: Ecology and man in Mexico's Central Volcanoes area: 19–48 (G. W. Heil, R. Bobbink, B. N. Trigo, Eds.). Kluwer Academic Publishers, Netherlands. Velázquez, A., 1994. Distribution and population size of Romerolagus diazi on El Pelado Volcano, Mexico. Journal of Mammalogy, 75: 743–749. – 1996. Geo–ecología del volcán Pelado, México: estudio integral enfocado hacia la conservación del conejo zacatuche. In: Ecología y conservación del conejo zacatuche y su hábitat: 102–118 (A. Velázquez, F. J. Romero, P. López, Eds.). FCE–UNAM, México D.F. – 2012. El contexto geográfico de los lagomorfos de México. Therya, 3(2): 223–238. Velázquez, A., Heil, G., 1994. Habitat suitability study for the conservation of the volcano rabbit (Romerolagus diazi). Journal of Applied Ecology, 33(3): 543−554. Velázquez, A., Larrazábal, A., Romero, J., 2011. Del conocimiento específico a la conservación de todos los niveles de organización biológica. El caso del zacatuche y los paisajes que denotan su hábitat. Investigación Ambiental, 3(2): 59–62. Velázquez, A., Romero, F., León, L., 1996a. Fragmentación del hábitat del conejo zacatuche. In: Ecología y conservación del conejo zacatuche y su hábitat: 73–86 (A. Velázquez, F. J. Romero, P. López, Eds.). FCE–UNAM, México D.F. Velázquez, A., Romero, F., López–Paniagua, J., 1996b. Amplitud y utilización del hábitat del conejo zacatuche. In: Ecología y conservación del conejo zacatuche y su hábitat: 89–101 (A. Velázquez, F. J. Romero, P. López, Eds.). FCE– UNAM, Méxic, D.F. Weber, M., 2010. Perros (Canis lupus familiaris) y gatos (Felis catus) ferales en la Reserva de la Biosfera Los Petenes, Campeche, México: Diagnóstico, efectos en la fauna nativa y perspectivas de control. Reporte Técnico. Proyecto PNUD–CONANP SDP–18–2008: 9–57.

Muñoz and Farfán

Animal Biodiversity and Conservation 43.1 (2020)

123

Wild boar diet and its implications on agriculture and biodiversity in Brazilian forest–grassland ecoregions I. B. Cervo, D. L. Guadagnin

Cervo, I. B., Guadagnin, D. L., 2020. Wild boar diet and its implications on agriculture and biodiversity in Brazilian forest–grassland ecoregions. Animal Biodiversity and Conservation, 43.1: 123–136, https://doi.org/10.32800/ abc.2020.43.0123 Abstract Wild boar diet and its implications on agriculture and biodiversity in Brazilian forest–grassland ecoregions. We aimed to describe the composition of Sus scrofa diet in three Brazilian ecoregions characterized by a mosaic of forests and grasslands: Pampa, Araucaria Forest and Pantanal. We evaluated the possible risks that the species may represent for agriculture and conservation of biodiversity by analyzing the stomach content of 118 boars. We examined dietary patterns in each ecoregion using PCA (principal component analysis) and verified how diet varies according to individual attributes through redundancy analysis. We visualized the composition of macronutrients in a multidimensional space by means of RMT (right–angled mixture triangle). The wild boars presented a diverse diet, influenced by season, time of day, and local availability of resources. Cultivated grains and herbs were the most commonly consumed items, leading to a high carbohydrate intake. Damage to agriculture is potentially high given the large consumption of cultivated grains. Population growth and expansion may be limited by the low availability of protein in the ecoregions. Key words: Invasive species, Sus scrofa, Biodiversity, Agriculture Resumen La dieta del jabalí y sus implicaciones en la agricultura y la biodiversidad de las ecorregiones de bosques y pastizales del Brasil. En este artículo tratamos de describir la composición de la dieta de Sus scrofa en tres ecorregiones brasileñas caracterizadas por un mosaico de bosques y pastizales: la pampa, el bosque de araucaria y el pantanal. Evaluamos los riesgos que la especie puede representar para la agricultura y la conservación de la biodiversidad analizando el contenido estomacal de 118 jabalíes. Analizamos sus hábitos alimentarios en cada ecorregión utilizando el análisis de componentes principales y comprobamos que la dieta varía en función de las características de cada individuo mediante un análisis de la redundancia. Representamos la composición de macronutrientes en un espacio multidimensional mediante un diagrama triángular rectangular. El jabalí presentó una dieta diversa, influida por la estación, el momento del día y la disponibilidad local de recursos. Los cereales y hierbas cultivados fueron los productos consumidos más habitualmente, lo que apunta una ingesta elevada de carbohidratos. Los daños provocados a la agricultura podrían ser elevados dado el gran consumo de cereales cultivados. El crecimiento y la expansión de la población pueden verse limitados por la escasa disponibilidad de proteína en las ecorregiones. Palabras clave: Especie invasiva, Sus scrofa, Biodiversidad, Agricultura Received: 21 X 19; Conditional acceptance: 02 XII 19; Final acceptance: 10 I 20 Isadora Bisognin Cervo, Demetrio Luis Guadagnin, Laboratório de Conservação e Manejo da Vida Silvestre, Instituto de Biociências, Programa de Pós–Graduação em Ecologia, Universidade Federal do Rio Grande do Sul, 9500 Avenida Bento Gonçalves, Porto Alegre, RS, Brazil, CEP 91501–970. Corresponding author: I. B. Cervo: isacervo@hotmail.com; dlguadagnin@gmail.com ISSN: 1578–665 X eISSN: 2014–928 X

Cervo and Guadagnin

124

Introduction The introduction of organisms by humans in the last 200 years, either accidentally or intentionally, overcame the dispersion by natural forces in previous periods of Earth's history (Mack et al., 2000; Lockwood et al., 2013). The wild boar Sus scrofa L. is one of worst invasive species at a global level (Lowe et al., 2000). Wild boar damage agricultural crops (Nunley, 1999; Schley and Roper, 2003; Deberdt and Scherer, 2007), attack domestic animals (Nunley, 1999; Deberdt and Scherer, 2007), serve as reservoirs of diseases (Nunley, 1999; Deberdt and Scherer, 2007), threaten native species by predation or competition (Wood and Roark, 1980), alter ecosystem processes (Wilcox and Van Vuren, 2009), and favor other exotic species. Wild boar of different lineages have been found in the wild in Brazil since they were accidentally introduced in the late nineteenth century. Later, as of 1960, they were being deliberately introduced for hunting and commercial purposes. Populations also expanded as they invaded over borders from nearby countries (Deberdt and Scherer, 2007; García et al., 2011; Pedrosa et al., 2015). The dietary niche and the feeding habits are fundamental to our understanding of how species and individuals behave adaptively and what damage they can cause once introduced (Senior et al., 2016). The feeding habits of wild boar explain much of the unwanted ecological and economic effects (Zeman et al., 2018); the active rooting and foraging (Jones et al., 1994; Crooks, 2002) and their generalist diet (Hahn and Eisfeld, 1998; Schley and Roper, 2003; Morelle and Lejeune, 2015) explain their ability to exploit a wide range of resources and to maintain large populations. Agricultural crops are not only damaged in their search for food but also contribute to the establishment and expansion of wild boars (Keiter and Beasley, 2017). Moreover, wild boar may endanger wild species by predation, competition or habitat alteration (Bevins et al., 2014). In South Brazil, for example, preliminary assessments suggest that wild boar consume large amounts of Araucaria angustifolia (Deberdt and Scherer, 2007), a key resource for several vertebrate species in winter (Deberdt and Scherer, 2007; Zanin Hegel and Ángelo Marini, 2013). They also predate toads, eggs and nestlings of ground nesting birds, and lambs (Chimera et al., 1995; Deberdt and Scherer, 2007). Their diet is influenced by a number of factors, including habitat, season, circadian activity, and individual traits (Keuling et al., 2008). It has been reported, for example, that roots of plants appear to be consumed more often in the winter and green parts in the spring (Ballari and Barrios–García, 2014), and that diet and movements are influenced by seasonal availability of agricultural crops (Hahn and Eisfeld, 1998; Schley and Roper, 2003; Morelle and Lejeune, 2015) and hunting pressure (Keuling et al., 2008). The moment and methods of capture may introduce bias in dietary analysis (Scillitani et al., 2010), since behavior, including diet, changes throughout the day and some control techniques include baiting. Diet is determined by the breadth of the dietary niche of a species, or its degree of generalism, which

involves at least three interrelated dimensions (Westoby, 1978; Machovsky–Capuska et al., 2016). These are the range of physical attributes of food resources (food composition niche), the nutritional compositions of these resources (food exploitation niche), and the range of macronutrient diet composition (macronutrient niche). The degree of generalism of a species with regard to macronutrients and food composition may help reveal a population's suitability to a new environment, allowing the assessment of its colonization potential and geographical expansion (Hutchinson, 1957). The wild boar is known as a generalist regarding food exploitation and composition and has a broad macronutrient niche (Senior et al., 2016). In this work, we characterize and compare the wild boar dietary niche in three neotropical ecoregions in Brazil where the species is established and currently expanding over wild and agricultural land (Pedrosa et al., 2015). Specifically, according to patterns in other regions invaded by wild boars, we expect that: 1) the diet varies between sexes, with females ingesting more protein than males, but not between ages, due to the gregarious feeding behavior, nor concerning the time of the day, due to behavioral plasticity; 2) since the wild boar is a generalist in food composition and exploitation, we expect the range of food resources to be broad and to vary regionally and seasonally; and 3) cultivated grains, when available, will play an important role in the diet, due to their high nutritional value and ease of access. We also expect the dietary niche of wild boar in Brazil to fall within the macronutrient niche space established by Senior et al. (2016). Material and methods Study area We collected samples from three ecoregions (fig. 1) corresponding to three biogeographical provinces (Morrone, 2014): Pantanal Ecoregion (Pantanal Province), Uruguayan Savannas (Pampa Province) and Brazilian Araucaria Forest (Araucaria angustifolia Forest Province). In Brazil, wild boar invasion is particularly widespread in these three ecoregions (Pedrosa et al., 2015). All three ecoregions consist of mosaics of forests, shrublands and grasslands. The climate in the Araucaria Forest is temperate and humid, with an average annual temperature of 17 ºC. The total annual precipitation is 1,500–2,000 mm (Cfb) (Overbeck et al., 2009; Suertegray and da Silva, 2009). The climate in the Pampa is subtropical, with a mean annual temperature of 18 ºC and total annual rainfall of 1,500 mm (Cfa) (Moreno, 1961; IBGE, 2012). In these two ecoregions the critical season for vegetation growth is the winter, but droughts can occur in summer months. The climate in the Pantanal is tropical, with an average annual temperature of 26 ºC and total annual rainfall of 1,600 mm, alternating a wet season from November to April and a dry season from May to October (Aw) (Cadavid Garcia, 1984). The Pantanal is a lowland Savanna subjected to annual flooding in the wet season, extending from

Animal Biodiversity and Conservation 43.1 (2020)

125

67º30' W

62º30' W

17º30' S

Equator

57º30' W

52º30' W

47º30' W

42º30' W

Santa Lucia Farm

Bolivia Brazil

N S

22º30' S

Paraguay Bolivia

Tropic of Capricorn

Brazil

Paraguay Tropic of Capricorn Argentina

Uruguay

27º30' S

Argentina

André da Rocha Campestre da Serra Alegrete

32º30' S

Santana do Livramento Dom Pedrito

200 km

Uruguay

Cambarà do Sul Sao Francisco de Paula

Encruzilhada do Sul Araucaria Forest Pampa Pantanal

Fig. 1. Map of the three ecoregions studied: Araucaria Forest, Pampa and Pantanal, with the respective sampling sites of stomach contents of wild boar in 2015 and 2016. Fig. 1. Mapa de las tres ecorregiones estudiadas: el bosque de araucaria, la pampa y el pantanal, con los respectivos sitios donde se recogieron las muestras de contenido estomacal de jabalíes entre 2015 y 2016.

four to nine months and covering up to 90 % of the area (Cordeiro, 2004; Pott and Pott, 2004). About 33 % to 50 % of the Pantanal area is flooded annually (Apollonio et al., 1988; Silva et al., 2000). Diet analysis We obtained the stomach content of wild boar from individual hunters officially authorized to conduct wild boar control under Brazilian regulations. We obtained 45 samples from the Pampa between June 2015 and October 2016, 15 samples from the Araucaria Forest from September to October 2015, and 58 samples from the Pantanal, 31 of which were collected in September and October 2015 (flooding season) and 27 in June 2016 (dry season). All animals were culled with firearms in active boar beats (Araucaria and Pantanal) or attracted to stands of blinds with baiting. Each animal was obtained from a separate beat.Immediately after culling, we incised the stomach and collected samples of 500 ml from the centre of its content, stored them in vials containing a solution of 90 % alcohol 70 %, 5 % formalin and 5 % acetic acid (Skewes et al., 2007). These procedures were carried out in the field with assistance from the hunters. We were unable to weigh the carcasses. Each sample was washed in a 1.7 mm sieve (Wood and Roark, 1980). We determined the food items under stereoscopic lens, classifying them into eight categories: herbs and leaves, cultivated grains, wild seeds, fruits, roots, wood (parts of tree trunks and bark), invertebrates,

and vertebrates. Baiting with corn used by some hunters was not included in the food items of the animals captured using this technique. We calculated and recorded the percentage by volume of each food item by displacing the water in a volumetric beaker, according to Skewes et al. (2007). For each animal we recorded the sex (male or female), age (juvenile or adult, based on body size and tusk development), the method of capture (hunting with dogs, trapping, nocturnal hunting with searchlights) and the time of capture (morning, afternoon or night). We looked carefully at the stomach contents for remains of species considered particularly vulnerable to predation by wild boar (ground nesting birds, toads, bulbous perennial herbs, and seeds of Araucaria angustifolia). Data analysis We checked sample sufficiency through rarefaction. We used principal component analysis (PCA) to explore the dietary patterns in each ecoregion and redundancy analysis (RDA) to explore how the diet varies according to ecoregions, sex, age and the capture method, time of the day and season. In these ordinations, cultivated grains and wild seeds were pooled in order to evaluate the importance and variation of grain consumption. The data were previously standardized by Hellinger’s transformation. We analyzed the significance of the RDA through the permutation test. These multidimensional techniques are suitable to obtain an overview of the data and to

Cervo and Guadagnin

Results

7 3

equivalent to 17 Kj/g and lipids to 37 Kj/g (Raubenheimer and Rothman, 2013). The primary axes X and Y represent carbohydrates and proteins, respectively, while the implicit Z axis represents lipids. Each point represents the percentage participation in Kj of proteins (P), lipids (L) and carbohydrates (C) found in a food item. Segments of lines joining the points of the graph in a minimum convex polygon demarcate the nutrient space potentially accessible to wild boars in each ecoregion or season (Coogan et al., 2014). This nutrient space defines the fundamental niche of macronutrients, the space of the RMT in which the population may persist (Machovsky–Capuska, 2016). We calculated and designed the RMT in the LibreOffice Calc spreadsheet (LibreOffice, 2018).

126

Pantanal

Food items 4 5

Araucaria Forest

Pampa

20 30 40 Individual number

Fig. 2. Rarefaction curve using samples of stomach contents of wild boar from the Araucaria Forest, Pampa and Pantanal. Fig. 2. Curva de rarefacción utilizando muestras de contenido estomacal de jabalíes del bosque de araucaria, la pampa y el pantanal.

orient further hypothesis testing. We employed the vegan package (Oksanen et al., 2017) of the software R (R Core Team, 2017) to perform these analyses. We calculated the standardized Levin's index (Feinsinger et al., 1981) to estimate the breadth of the food composition niche within ecoregions and the Pianka index to estimate the food composition niche overlap between ecoregions (Colwell and Futuyma, 1971). For both indexes we used the proportion of each food item in each ecoregion, pooling the data of all stomach contents. These indexes vary from a minimum breath or overlap of 0 to a maximum of 1. We calculated the index in the LibreOffice Calc spreadsheet (LibreOffice, 2018). We analyzed the food and macronutrient composition through the RMT (right–angled mixture triangles) technique used to visualize the distribution of macronutrients in a multidimensional space (Machovsky–Capuska et al., 2016). In order to evaluate the percentage of each macronutrient in the diet, we first estimated the percentage of contribution of each food in grams to the total diet of each ecoregion. We then estimated the percentage of carbohydrates, lipids and proteins of each food from published data (appendix 1). Finally, we transformed them into energy content, taking carbohydrates and proteins as

Rarefaction curves indicated that the number of samples was sufficient to characterize the diet in the Pantanal and Pampa, but insufficient to characterize those in the Araucaria Forest (see fig. 2). Consumption of herbs and leaves, cultivated grains, wild seeds, fruits, roots and vertebrates varied greatly across ecoregions and seasons, as is evident in the PCA. The first two axes of the PCA cumulatively explained 68.4 % of the variations in the proportion of food items among the individuals in the three ecoregions (see fig. 3). The diet showed a common pattern across ecoregions (as shown in table 1). In summary, herbs/leaves and roots were the items consumed in highest volume (respectively 11.4 to 44.1 % and 19.3 to 28.7 %) and frequency (33.3 to 65.5 % and 46.7 to 53.4 %) in all three ecoregions; cultivated grains were an important part of the diet in the two ecoregions characterized by agricultural matrix (Pampa and Araucaria Forest; above 40 % in frequency and volume); vertebrates, invertebrates and fruits were consumed in lower volume and frequency. Crops found in the stomachs were oats, sorghum, ryegrass, corn, rice and soybean. The consumption of vertebrates was frequent in the Pantanal (24.1 %, mainly amphibians) and in the Pampa (30.4 %, mainly sheep and armadillos Dasypus sp.). Fruits were frequent (15.5 %) in the Pantanal, with bocaiúva Acrocomia aculeata (Jacq.) Lodd. Mart. being the most consumed item in this category. We found baiting with corn in three samples, two from the Araucaria Forest (13 %) and one from the Pampa (2 %). Corn baiting was identified by the pink color of seeds treated with fungicides. The first two axes of the PCA in dry and flooding seasons in the Pantanal explained 80 % of the variation (see fig. 4). The wild boar diet in the Pantanal was composed mainly of roots in the flooding season (77.6 % in volume), and herbs/leaves in the dry season (56.7 % in volume), also including fruits and roots, as shown in table 2. The major items in the diet of males and females were the same across ecoregions: cultivated grains (25.2 % and 26.5 %, respectively), herbs/leaves (24.3 % and 38.1 %) and roots (26.1 % and 22.7 %), although males ate almost four times more wild seeds

Animal Biodiversity and Conservation 43.1 (2020)

–5

10 10

–10

Herbs and leaves

Roofs

–5

Wild seeds Woods Fruits Invertebrates Vertebrates

–10

Cultivated grains

–0.10 Pantanal Pampa Araucaria Forest –0.15

–0.10

–15

–0.15

PC2 –0.05 0.00

0.05

0.10

–15

127

–0.05 0.00 PC1

0.05

0.10

Fig. 3. Principal component analysis (PCA) depicting variations in the proportions of food items (cultivated grains, woods, invertebrates, vertebrates, fruits, wild seeds, roots and herbs/leaves) among individual wild boar in three Brazilian ecoregions: Pantanal, Pampa and Araucaria Forest. Fig. 3. Análisis de componentes principales en el que se muestra la variación de la proporción de productos alimenticios (cereales cultivados, maderas, invertebrados, vertebrados, frutas, semillas silvestres, raíces e hierbas u hojas) en los individuos de jabalí en las tres ecorregiones del Brasil: el pantanal, la pampa y el bosque de araucaria.

Table 1. Percentage of volume and frequency of occurrence of food items in the stomachs of boar in three Brazilian ecoregions in 2015 and 2016. Tabla 1. Porcentaje del volumen y la frecuencia de la presencia de productos alimenticios en el estómago de los jabalíes en tres ecorregiones del Brasil en 2015 y 2016.

Pantanal (58 samples)

Pampa (45 samples) Volume

Frequency

Araucaria Forest (15 samples)

Volume

Frequency

Herbs and leaves

44.1

65.5

11.4

34.8

20.9

33.3

Cultivated grains

0.0

48.1

43.5

44.4

40.0

Fruits

7.0 15.5

0.2 6.5

0.0 0.0

Roots

28.7

53.4

28.5 47.8

19.3 46.7

Wild seeds

15.9

56.9

0.0

12.6

Vertebrates

0.4

24.1

9.4 30.4

1.7 13.3

Invertebrates

3.8

39.6

2.3 50.0

0.1 13.3

Woods

0.1

5.2

0.1 4.3

0.8 13.3

0.0

Volume

Frequency

6.7

Cervo and Guadagnin

128

Dry season

PC2 0.1

0.2

0.3

Flooding season

–5

5 0

Cultivated grains Woods Vertebrates Invertebrates

Roots

Herbs and leaves

–0.1

–5

–0.1

0.0

Fruits Wils seeds

0.0

0.1 PC1

0.2

0.3

Fig. 4. Principal component analysis (PCA) depicting variations in the proportions of food items (cultivated grains, woods, invertebrates, vertebrates, fruits, wild seeds, roots and herbs/leaves) among individual wild boar in the dry and flooding seasons in the Pantanal ecoregion. Fig. 4. Análisis de componentes principales en el que se muestra la proporción de productos alimenticios (cereales cultivados, maderas, invertebrados, vertebrados, frutas, semillas silvestres, raíces e hierbas u hojas) en los individuos de jabalí en la estación seca y la estación de las inundaciones en la ecorregión del pantanal.

Table 2. Percentage of volume and frequency of occurrence of food items in stomachs of boar for the flooding and dry seasons in the Pantanal ecoregion in 2015 and 2016. Tabla 2. Porcentaje del volumen y la frecuencia de la presencia de productos alimenticios en el estómago de los jabalíes en la estación de las inundaciones y la estación seca en la ecorregión del pantanal en 2015 y 2016. Pantanal (15 samples) Flooding season

Dry season

Volume

Frequency

Volume

Frequency

Herbs and leaves

6.4

40.7

56.7

90.3

Cultivated grains

0.0

Fruits

4.2

7.4

20.6 80.6

Roots

77.6 96.3

0.8 16.1

Wild seeds

1.9

20.6

Vertebrates

29.6

80.6

0.0 0.0

0.6 45.2

Invertebrates 9.8 70.4

0.6 12.9

Woods

0.1 9.7

0.0 0.0

Animal Biodiversity and Conservation 43.1 (2020)

129

Table 3. Percentage of volume of food items in stomachs of male and female boar in the three Brazilian ecoregions in 2015 and 2016. Tabla 3. Porcentaje del volumen de los productos alimenticios encontrados en el estómago de machos y hembras de jabalíes en las tres ecorregiones del Brasil en 2015 y 2016..

Pantanal (58 samples)

Pampa (45 samples)

Araucaria Forest (15 samples)

Total

Females Males

Females

Males

Herbs and leaves

62.6

26.6

12.4

8.2

37.9

8.5

38.1

24.3

Cultivated grains

0.0

53.3

69.8

31.7

54.6

26.5

25.2

Fruits

7.0 20.1 0.5 0.0 0.0 0.1

Roots

20.0 30.0 23.9 14.5 30.4 23.8 22.7 26.1

Wild seeds

4.8

Vertebrates

0.3 0.4 6.4 4.8 0.0 10.6 3.0 4.0

Invertebrates

5.1 2.5 3.1 0.2 0.0 2.2 3.7 2.8

Woods

0.1 0.1 0.3 2.5 0.0 0.0 0.2 0.1

20.3

0.0

than females (8.6 % in volume for males, 2.2 % for females), as shown in table 3. The RDA showed that dietary differences were highly variable among individuals, being related to the spatial and temporal distribution of the samples (F = 2.325, p < 0.001), as shown in table 4. The first two axes of the RDA cumulatively explained 15.7 % of dietary differences between individuals in relation to the time of year, the ecoregion, and the method and period of the hunting. Cultivated grains were consumed in a greater proportion in the Pampa, during the night, and when hunting with dogs. Roots and invertebrates were consumed more often during times of less abundance of resources, in the Pantanal, and in the morning or afternoon. Herbs/leaves were more consumed in the Araucaria Forest or in trapping (fig. 5, table 4). The niche breadth measured by the standardized Levins index was 0.34 in the Araucaria Forest, 0.28 in the Pampa and 0.32 in the Pantanal. In the Pantanal, this index was 0.09 in the dry season and 0.15 in the flooding season. The niche overlap was low when the three studied ecoregions were compared (Pianka index = 0.15), being higher among the Pampa and Araucaria Forest and smaller between these and the Pantanal, as shown in table 5. The overlap was also low between flooding and dry seasons (Pianka Index = 0.10). The analysis of the composition of macronutrients through RMT (fig. 6) indicated that the wild boar diet in the three ecoregions is on the margin of the macronutrient niche space established by Senior et al. (2016), especially due to the low intake of proteins. The diet was mainly composed of carbohydrates in all ecoregions. The proportion of carbohydrates in the diet was lower in Pantanal. In all cases the energy from proteins was below 20 %, close to the lowest values found in other countries (Machovsky–Capuska

0.0

3.5 8.9 2.2

8.6

et al., 2016). Protein intake was relatively higher in the Pampa, mainly due to the contribution of vertebrates. Lipid intake in Pantanal (48 %) was high, compared to that in other ecoregions studied. The lipid intake in this region was mainly due to the contribution of palm fruits. The diet in the flooding and dry seasons in the Pantanal was predominantly composed of carbohydrates (fig. 7). However, in the flooding season, carbohydrates corresponded to almost 90 % of the diet due to the higher intake of grasses and roots, while in the dry season they corresponded to about 50 %. Discussion We found that wild boars introduced in the Araucaria Forest, Pampa and Pantanal, three Brazilian ecoregions characterized by mosaics of grasslands and forests, have a diverse diet that is influenced by season, capture method and differences in local availability of resources, but not by individual traits. The consumption of cultivated grains, when available, was found in most samples, demonstrating the importance of these food items. The availability of cultivated grains (in our case oats, sorghum, rye, corn, rice and soybean), either used as supplementary feeding or for baiting, is related to the wild boar population and impact increase in Europe (Ballari et al., 2015; Miloš et al., 2016). In all ecoregions, the diet was located at the fringe of the ideal target that maximizes fitness, as proposed by Senior et al. (2016), and the supply of proteins was generally critical. It is known that the wild boar is an opportunistic animal, feeding on any available food resource, although there are preferences (Schley and Roper, 2003). Seasonal differences in habitat use may also be related to changes in food availability (Oja et al.,

Cervo and Guadagnin

130

Hunting with traps

Araucaria Forest

Axis II

Herbs and leaves

Vertebrates

Afternoon captures

Dry season Night captures Panama ecoregion Hunting with dogs

Fruits

Pantanal ecoregion Flooding season

Invertebrates

Cultivated grains

Roots

Axis I Fig. 5. Redundancy analysis (RDA) depicting the relationship of differences in the diet of wild boar according to the spatial and temporal distribution of the samples in the Pantanal, Pampa and Araucaria Forest. Fig. 5. Análisis de la redundancia en el que se muestra la relación de las diferencias en la dieta del jabalí según la distribución espacial y temporal de las muestras en el pantanal, la pampa y el bosque de araucaria.

Table 4. Significance of the association of attributes of the wild boars and the spatial and temporal distribution of the samples (redundancy analysis axes) with the differences in the proportion of food items in the stomach contents of wild boar in three Brazilian ecoregions in 2015 and 2016. Tabla 4. Significación de la asociación de las características de los jabalíes en la distribución espacial y temporal de las muestras (ejes del análisis de la redundancia) con las diferencias en la proporción de los productos alimenticios encontrados en el contenido estomacal del jabalí en tres ecorregiones del Brasil en 2015 y 2016. Factor Variance F P

2015). Herbs/leaves and roots formed the basis of the diet in the three ecoregions, perhaps because they are the most abundant and most widely available items everywhere (Chimera et al., 1995; Ballari and Barrios–García, 2014). Moreover, since pigs have difficulty extracting energy from fresh herbage (Wie-

Table 5. Food niche overlap (Pianka index, PI) of wild boar captured in the Pantanal, Pampa and Araucaria Forest ecoregions in 2015 and 2016. Tabla 5. Solapamiento del nicho alimentario (índice de Pianka, PI) de los jabalíes capturados en las ecorregiones del pantanal, la pampa y el bosque de araucaria en 2015 y 2016.

Capture shift

0.043

4.161 0.001

Capture method

0.024

2.343 0.017

Period of the year

0.015

2.827 0.037

Pantanal, Pampa and Araucaria Forest 0.15

Ecoregion

0.023 2.261 0.025

Pantanal and Araucaria Forest

0.41

Age strata

0.005

Pantanal and Pampa

0.55

Sex

0.002 0 0.747

Pampa and Araucaria Forest

0.94

Flooding and dry season

0.1

0.996

0.37

Residual 0.504

131

1.0

Animal Biodiversity and Conservation 43.1 (2020)

% Carbohydrates 0.4 0.6 0.8

Pantanal Pampa Araucaria Forest

0.0

0.2

ds pi Li

0.0

0.2

0.4 0.6 % Proteins

0.8

1.0

Fig. 6. Right–angled mixture triangle (RMT) of the three ecoregions studied in 2015 and 2016: Pantanal, Pampa and Araucaria Forest. The smaller sympols are the composition of macronutrients (carbohydrates, proteins and lipids) as a percentage of each food eaten by boar in each ecoregion. The larger symbols (diets) are the proportion of energy carbohydrates, proteins and lipids that the food set found in the diet provides the boar in each ecoregion. The gray area represents an estimate of the fundamental macronutrient niche found in the work of Senior et al. (2016), corresponding to the convex polygon formed by all diets. The areas surrounded by pink, green and black lines represent, respectively, the estimation of the fundamental macronutrient niche of the Pantanal, Pampa and Araucaria Forest. Fig. 6. Diagrama triangular rectangular de la composición de la dieeta de las tres ecorregiones estudiadas en 2015 y 2016: el pantanal, la pampa y el bosque de araucaria. Los símbolos pequeños representan la composición de macronutrientes (hidratos de carbono, proteínas y lípidos), expresada en porcentaje, de cada alimento consumido por el jabalí en cada ecorregión. Los símbolos grandes (dietas) representan la proporción de la energía que procede de los carbohidratos, las proteínas y los lípidos respecto de toda la energía que el conjunto de alimentos encontrados en la dieta proporciona a los jabalíes en cada región. El área gris representa una estimación del nicho fundamental relativo a los macronutrientes encontrado en el trabajo de Senior et al. (2016), que corresponde al polígono convexo formado por todas las dietas. Las áreas delimitadas por líneas rosas, verdes y negras representan, respectivamente, la estimación del nicho de macronutrientes fundamentales del pantanal, la pampa y el bosque de araucaria.

ren, 2000; Edwards, 2003; Massei and Genov, 2004), they need to feed themselves abundantly with herbs and leaves in order to extract enough energy for their survival. Rooting is also considered a signal of scarcity of preferred above ground resources (Zeman et al., 2018). The intake of cultivated grains in the Pampa and Araucaria Forest and fruits and wild seeds in the Pantanal was high in the seasons where these items were available, suggesting that these are preferred items, known to have high energy value (Caley, 1993). The lipid intake in Pantanal in the dry season was among the highest reported globally (Senior et al., 2016). Roots and invertebrates are important foods in the Pantanal, especially in the dry season. The niche overlap between peccaries and wild boars increases in the flooding season, when both species increase the consumption of fresh plants and fruits and reduce rooting (Sicuro and Oliveira, 2002).

The differences in stomach content as a function of hunting time and method probably reflect a habitat and feeding shift for refuges rather than biases due to collection method. Hunting influences wild boar behavior (Scillitani et al., 2010) and feeding patterns. Each mode of hunting occurs in a specific place and period of the day, and the stomach contents reflect the most recent feeding activity. Our results from stomach content analysis suggest that wild boar tend to move either to cultivated or open areas more frequently at night. Sex and age group had negligible influences on the variation between individuals regarding stomach content. Although wild boar nutritional needs were reported to vary with age (Dardaillon, 1986), sex, and reproduction (Wilcox and Van Vuren, 2009), not all studies have found significant differences (Wood and Roark, 1980; Loggins et al., 2002, Adkins and

Cervo and Guadagnin

1.0

132

ds pi Li

0.0

% Carbohydrates 0.2 0.4 0.6 0.8

Dry season Flooding season

0.0

0.2

0.4 0.6 % Proteins

0.8

1.0

Fig. 7. Right–angled mixture triangle (RMT) of the Pantanal ecoregion in the flooding and dry seasons in 2015 and 2016. The smaller symbols represent the composition of macronutrients (carbohydrates, proteins and lipids) in percentage of each food ingested by the boars in each period. The larger symbols (diets) are the proportion of energy carbohydrates, proteins and lipids that the food set found in the diet provides to the boar in each period. The areas surrounded by the blue and red lines represent, respectively, the estimation of the fundamental macronutrient niche of the Pantanal in the flooding season and in the dry season. Fig. 7. Diagrama triangular rectangular de la composición de la dieeta de la ecorregión del pantanal en la estación seca y la estación de las inundaciones en 2015 y 2016. Los símbolos pequeños representan la composición de macronutrientes (hidratos de carbono, proteínas y lípidos), expresada en porcentaje, de cada alimento consumido por los jabalíes en cada período. Los símbolos grandes (dietas) representan la proporción de la energía que procede de los carbohidratos, las proteínas y los lípidos respecto de toda la energía que el conjunto de alimentos encontrados en la dieta proporciona a los jabalíes en cada período. Las áreas delimitadas por líneas azules y rojas representan, respectivamente, la estimación del nicho de macronutrientes fundamentales del pantanal en estación de las inundaciones y la estación seca.

Harveson, 2006). We interpret this as a consequence of the difficulty in adjusting the diet under the limiting conditions of wild environments. It is also possible that differences are undetected because hunted animals do not represent the complete age structure of the population. The gregarious habit could also minimize differences in the food content of animals of the same group, but the individuals we collected were always independent sampling units. In our data, the consumption of native fauna was sporadic. We recorded a few samples of amphibians in the Pantanal and armadillos Dasypus sp. in the Pampa and bird feather in the Araucaria Forest. We did not record the egg intake found in the Pantanal by Desbiez et al. (2009), nor did we record the intake of fauna and flora of special conservation concern. The main economic effect of the presence of wild boar, based on the stomach contents, is the consumption of cultivated grains, whereas in the livestock sector, the damage appears to be sporadic or localized. Cultivated grains were found in most of the sample from the Pampa and Araucaria Forest. We recorded

a few stomachs with sheep meat and wool. Most of the vertebrate samples contained fly larvae, indicating that at least part of this consumption was of carcasses. This finding agrees with the opportunistic scavenger habit of wild boars in the Pampa ecoregion (Herrero and Fernández de Luco, 2003; Desbiez et al., 2009), as well as in its native range (Espadas et al., 2010). Alternatively, boars could be actively searching for alternative protein rich resources, which are scarce in the available plant material (Ballari and Barrios– García, 2014). In the three ecoregions, the diet we recorded is at the margin of the ideal target that maximizes the fitness of wild boar according to Senior et al. (2016), mainly due to the low protein intake, suggesting an unbalanced diet. Potential consequences on reproduction and caring capacity are expected (Senior et al., 2016) and deserve further research. The domestic pig, under optimum conditions for fattening and reproduction, requires about 14.5 % of protein in periods of growth and lactation (Zardo and Lima, 1999). Crude protein content in diet of feral pigs is optimal above

Animal Biodiversity and Conservation 43.1 (2020)

12 % (Coblentz and Baber, 1987). These targets are higher than the amount we found in Pantanal (9.3 %) and in the Araucaria Forest (10.3 %), but not in the Pampa (15.2 %). In the studied ecoregions, this may apply particularly in the Pampa and in the dry season in the Pantanal. Comparing our RMT results with Senior et al. (2016), correspondence of the food niche with climate becomes evident. The food niche in the Pampa and Araucaria Forest resembles that found in Australia, in Italian Piedmont (Cfa) and in France (Cfb). None of the regions revised by Senior et al. (2016) have a tropical seasonal climate similar to that of the Pantanal (Tropical Savanna, Aw). Our data highlight the plasticity of the wild boar diet. We recognize some limitations of our study. We obtained an unbalanced sample between the regions and seasons of the year, and the sampling in the Araucaria Forest ecoregion was insufficient to recover the whole set of potential foods. However, we were still able to characterize regional patterns and provide an overview of the wild boar food and macronutrient niches for these three regions where the species is already established and expanding. We segregated the items into general categories, which allowed us to provide information about regional differences in major items, the macronutrient share in the diets and relative niche breath based on major food categories. We were unable to discriminate the intake at the species level, which precludes the identification of predation over species of special conservation concern, a central question about effects of invasive species.However, we adopted a standard and precise volumetric analysis (Zeman et al., 2016) and are confident that vertebrates and insects were predated only passively or opportunistically, at low levels. Moreover, we did not find remains of major plant families, including endangered species particularly vulnerable to wild boar feeding habits, such as Cactaceae, Amarillydaceae, Liliaceae and Amaranthaceae. Molecular techniques are best suited for recovering the presence of items consumed in low quantity (Robeson II et al., 2017) but can only provide frequency measures. Acknowledgements This study was partially supported by Grant no. 400713/2013–6 from the Conselho Nacional de Desenvolvimento Científico e Tecnológico. The authors would also like to thank Coordenação de Aperfeiçoamento de Pessoal de Nível Superior (CAPES) for the Programa de Excelência Acadêmica (PROEX) and the Portal de Periódicos. References Adkins, R. N., Harveson, L. A., 2006. Summer diets of feral hogs in the Davis Mountains, Texas. The Southwestern Naturalist, 51: 578–580. Apollonio, M., Randi, E., Toso, S., 1988. The systematics of the wild boar (Sus scrofa L.) in Italy.

133

Italian Journal of Zoology, 55: 213–221. Ballari, S. A., Barrios–García, M. N., 2014. A review of wild boar Sus scrofa diet and factors affecting food selection in native and introduced ranges. Mammal Review, 44(2): 124–134. Ballari, S. A., Cuevas, M. F., Ojeda, R. A., Navarro, J. L., 2015. Diet of wild boar (Sus scrofa) in a protected area of Argentina: the importance of baiting. Mammal Research, 60(1): 81–87, https:// doi.org/10.1007/s13364–014–0202–0 Bevins, S. N., Pedersen, K., Lutman, M. W., Gidlewski, T., Deliberto, T. J., 2014. Consequences Associated with the Recent Range Expansion od Nonnative Feral Swine. BioScience, 64(4): 291–299. Bukkens, S. G. F., 1997. The nutritional value of edible insects. Ecol. Food. Nutr., 36(2–4): 287–319. Cadavid Garcia, E. A., 1984. O clima do Pantanal Mato–Grossense. EMBRAPA/UEPAE de Corumbá: 14–42. Caley, P., 1993. Population dynamics of feral pigs (Sus scrofa) in a tropical riverine habitat complex. Wildl. Res., 20: 625–636. Chimera, C., Coleman, M., Parkes, J. P., 1995. Diet of Feral Goats and Feral Pigs on Auckland Island. New Zealand.Journal of Ecology, 19: 203–207. Coblentz, B. E., Baber, D. W., 1987. Biology and control of feral pigs on Isla Santiago, Galapagos, Ecuador. Journal of Applied Ecology, 24: 403–418. Coimbra, M. C., Jorge, N., 2011. Proximate composition of guabiroba (Syagrus oleracea), jerivá (Syagrus romanzoffiana) and macaúba (Acrocomia aculeata) palm fruits. Food Research International, 44: 2139–2142. Colwell, R. K., Futuyma, D. J., 1971. On the measurement of niche breadth and overlap. Ecology, 52(4): 567–576. Coogan, S. C. P., Raubenheimer, D., Stenhouse, G. B., Nielsen, S. E., 2014. Macronutrient Optimization and Seasonal Diet Mixing in a Large Omnivore, the Grizzly Bear: A Geometric Analysis. Plos One, 9(8): e105719, https://doi.org/10.1371/journal. pone.0105719 Cordeiro, J. L. P., 2004. Estrutura e heterogeneidade da paisagem de uma unidade de conservação no nordeste do Pantanal (RPPN SESC Pantanal), Mato Grosso, Brasil: efeitos sobre a distribuição e densidade de antas (Tapirus terrestris) e de cervos–do–pantanal (Blastocerus dichotomus). PhD thesis, UFRGS, Porto Alegre. Crooks, J. A., 2002. Characterizing ecosystem–level consequences of biological invasions: the role of ecosystem engineers. Oikos, 97: 156–166. Dardaillon, M., 1986. Seasonal variations in habitat selection and spatial distribution of wild boar (Sus scrofa) in the Camargue, Southern France. Behavioural Processes, 13: 251–268. Deberdt, A. J., Scherer, S. B., 2007. O javali asselvajado: Ocorrência e manejo da espécie no Brasil. Natureza, Conservação, 5(2): 31–44. Desbiez, A. L. J., Santos, S. A., Keuroghlian, A., Bodmer, R. E., 2009. Niche Partitioning Among White–Lipped Peccaries (Tayassu pecari), Collared Peccaries (Pecari tajacu), and Feral Pigs (Sus

134

scrofa). Journal of Mammalogy, 90(1): 119–128. Edwards, S. A., 2003. Intake of nutrients from pasture by pigs. Proceedings of the Nutrition Society, 62; 257–265. Espadas, I. J., Sánchez–Zapata, J. A., Botella, F., Eguía, S., 2010. Los Carroñeros facultativos y la caza mayor: estrategia de aprovechamiento de carroña. Galemys, 22: 595–609. Feinsinger, P., Spears, E. E., Poole, R. W., 1981. A simple measure of niche breath. Ecology, 62: 27–32. García, G., Vergara, J., Lombardi, R., 2011. Genetic characterization and phylogeography of the wild boar Sus scrofa introduced into Uruguay. Genetics and Molecular Biology, 337: 329–337. Hahn, N., Eisfeld, D., 1998. Diet and habitat use of wild boar (Sus scrofa) in SW–Germany. Faune Sauvage, 15: 595–606. Henry–Silva, G. G., Camargo, A. F. M., 2002. Valor nutritivo de macrófitas aquáticas flutuantes (Eichhornia crassipes, Pistia stratiotes e Salvinia molesta) utilizadas no tratamento de efluentes de aqüicultura. Acta Scientiarum, Biological Sciences: 519–526, https://doi.org/10.4025/actascibiolsci. v24i0.2353 Herrero, J., Fernández de Luco, D., 2003. Wild boars (Sus scrofa) in Uruguay: scavengers or predators? Mammalia, 67: 485–491. Hutchinson, G. E., 1957. Concluding remarks. Cold Spring Harbor Symposia on Quantitative Biology, 22: 415–427. IBGE, 2012. Manual Técnico da Vegetação Brasileira, Rio de Janeiro. Instituto Brasileiro de Geografia e Estatística, Rio de Janeiro, RJ. Jones, C. G., Lawton, J. H., Shachak, M., 1994. Organisms as ecosystem engineers. Oikos, 69: 373–386. Keiter, D. A., Beasley, J. C., 2017. Hog Heaven? Challenges of Managing Introduced Wild Pigs in Natural Areas. Natural Areas Journal, 37(1): 6–16, https://doi.org/10.3375/043.037.0117 Keuling, O., Stier, N., Roth, M., 2008. How does hunting influence activity and spatial usage in wild boar Sus scrofa L.? European Journal of Wildlife Research, 54(4): 729–737. LibreOffice, 2018. The Document Fundation. Available at: https://pt–br.libreoffice.org/sobre–nos/credits/ Lockwood, J. L., Hoopes, M. F., Marchetti, M. P., 2013. Invasion ecology, 2nd Ed., Wiley Blackwell, Malden, MA. Loggins, R. E., Wilcox, J. T., Van Vuren, D. H., 2002. Seasonal diets of wild pigs in oak woodlands of the central coast region of California. California Fish and Game, 88: 28–34. Lowe S., Browne M., Boudjelas S., D. P. M., 2000. 100 of the World's Worst Invasive Alien Species. Published by The Invasive Species Specialist Group ISSG a specialist group of the Species Survival Commission SSC of the World Conservation Union IUCN First published as special liftout in Aliens, 12(3): 1–12. Available online at: http://interface.creative.auckland.ac.nz/database/ species/reference_files/100English.pdf [Accessed

Cervo and Guadagnin

on December 2019]. Machovsky–Capuska, G. E., Senior, A. M., Simpson, S. J., Raubenheimer, D., 2016. The multidimensional nutritional niche. Trends in Ecology & Evolution, 31: 355–365, https://doi.org/10.1016/j. tree.2016.02.009 Mack, R. N., Simberloff, D., Lonsdale, W. M., Evans, H., Clout, M., Bazzaz, F. A., 2000. Biotic invasions: causes, epidemiology, global consequences, and control. Ecological Applications, Ecological Society of America, 10(3): 689–710, Massei, G., Genov, P., 2004. The environmental impact of wild boar. Galemys, 16: 135–145. Miloš, J., Michaela, H., Tomáš, K., Jaroslav, Č., 2016. Creeping into a wild boar stomach to fi nd traces of supplementary feeding. Wildlife Research, 43(7): 590–598, https://doi.org/10.1071/WR16065 Morelle, K., Lejeune, P., 2015. Seasonal variations of wild boar Sus scrofa distribution in agricultural landscapes: a species distribution modelling approach. European Journal of Wildlife Research, 61(1): 45–56. Moreno, J. A., 1961. Clima do Rio Grande do Sul. Secretaria da Agricultura do Rio Grande do Sul, Porto Alegre, 42. Morrone, J. J., 2014. Biogeographical regionalisation of the Neotropical region. Zootaxa, 3782(1): 1–110, http://dx.doi.org/10.11646/zootaxa.3782.1.1 NEPA, 2011. Tabela Brasileira de Composição de Alimentos. UNICAMP, Campinas. Nunley, G. L., 1999. The cooperative Texas wildlife damage management program and feral swine damage management. In: Proceedings of the Feral Swine Symposium, June 2–3, Texas: 27–30. Animal Helath Comission, Fort Worth. Oja, R., Zilmer, K., Valdmann, H., 2015. Spatiotemporal Effects of Supplementary Feeding of Wild Boar (Sus scrofa) on Artificial Ground Nest Depredation. Plos One: 10(8): e0135254, https://doi.org/10.1371/ journal.pone.0135254 Oksanen, J., Blanchet, F. G., Kindt, R., Legendre, P., Minchin, P., O'Hara, B., Simpson, G., Solymos, P., Stevens, H., Wagner, H., 2017. Vegan: Community Ecology Package, version 2.5–6, Overbeck, G. E., Müller, S. C., Fidelis, A., Pfadenhauer, J., De Patta Pillar, V., Casagrande Blanco, C., Boldrini, I. I., Both, R., Dias Forneck, E., 2009. Os Campos Sulinos: um bioma negligenciado. In: Campos Sulinos, conservação e uso sustentável da biodiversidade: 26–41 (V. de Patta Pillar, S. C. Müller, Z. M. de Souza Castilhos, A. V. Ávila Jacques, Eds.). Ministerio de Meio Ambiente, Brasilia. Pedrosa, F., Salerno, R., Fabio Vinicius Borges, Padilha Mauro, G., 2015. Current Distribution of Invasive Feral Pigs in Brazil: Economic Impacts and Ecological Uncertainty. Natureza, Conservação, 13(1): 84–87. Pinheiro, R. S. B., Garcia da Silva Sobrinho, A., Yamamoto, S. M., Barbosa, J. C., 2007. Composição tecidual dos cortes da carcaça de ovinos jovens e adult. Pesquisa Agropecuaria Brasileira, Brasília, 42(4): 565–571. Pott, A., Pott V. J., 2004. Features and conservation of

Animal Biodiversity and Conservation 43.1 (2020)

the Brazilian Pantanal wetland. Wetlands Ecology and Management, 12: 547–552. R Core Team, 2017. R: A language and environment for statistical computing. Foundation for Statistical Computing, Vienna, Austria. Raubenheimer, D., Rothman, J. M., 2013. Nutritional Ecology of Entomophagy in Humans and Other Primates. Annu. Rev. Entomol., 58: 141–160. Robeson II, M. S., Khanipov, K., Golovko, G., Wisely, S. M., White, M. D., Bodenchuck, M., Smyser, T. J., Fofanov, Y., Fierer, N., Piaggio, A. J., 2017. Assessing the utility of metabarcoding for diet analyses of the omnivorous wild pig (Sus scrofa). Ecology and Evolution, 8(1): 185–196. Schley, L., Roper, T. J., 2003. Diet of wild boar Sus scrofa in Western Europe, with particular reference to consumption of agricultural crops. Mammal Review, 33(1): 43–56. Scillitani, L., Monaco, A., Toso, S., 2010. Do intensive drive hunts affect wild boar (Sus scrofa) spatial behaviour in Italy? Some evidences and management implications. European Journal of Wildlife Research, 56(3): 307–318. Senior, A. M., Grueber, C. E., Machovsky–Capuska, G., Simpson, S. J., Raubenheimer, D., 2016. Macronutritional consequences of food generalism in an invasive mammal, the wild boar. Mammalian Biology – Zeitschrift für Säugetierkunde, 81(5): 523–526, https://doi.org/10.1016/j.mambio.2016.07.001 Sicuro, F. L., Oliveira, L. F. B., 2002. Coexistence of Peccaries and Feral Hogs in the Brazilian Pantanal Wetland: an Ecomorphological View. Journal of Mammalogy, 83(1): 207–217. Silva, M. P., Mauro, R., Mourao, G., Coutinho, M., 2000. Distribuição e quantificação de classes de vegetação do Pantanal através de levantamento aéreo. Revista Braileira de Botânica, 23(2): 143–152, http://dx.doi.org/10.1590/S0100– 84042000000200004 Skewes, Ó., Rodríguez, R., Jaksic, F. M., 2007. Ecología trófica del jabalí europeo (Sus seroja) silvestre en Chile. Revista Chilena de Historia Natural, 80(3): 295–307. Suertegray, D. M. A., da Silva, L. A. B., 2009. Tchê Pampa. Campos Sulinos – conservação e uso

135

sustentável da biodiversidade, Brasília/DF. Tacon, A. G. J., Stafford, E. A., Edwards, C. A., 1983. A preliminary investigation of the nutritive value of three terrestrial lumbricid worms for rainbow trout. Aquaculture, 35: 187–199. USDA/APHIS/WS, 2010. Feral Hog Biology, Impacts and Eradication Techniques. Wildlife Services, New Mexico. Waibel, P., Liu, J., Noll, S., 1987. Utilization of feather meal, blood meal, and meat and bone meal in market turkey diets. In: Proceedings of the Minnesota Conference on Turkey Research, nov. 5–6, Miscellaneous Publication, 43: 101–110. Minnesota Agricultural Experiment Station, University of Minnesota. Westoby, M., 1978. What are the Biological Bases of Varied Diet? The Amercian Naturalist, 985: 627–631. Wieren, S. E. van, 2000. Digestibility and voluntary intake of roughages by wild boar and Meishan pigs. Animal Science, 71(1): 149–156. Wilcox, J. T., Van Vuren, D. H., 2009. Wild pigs as predators in oak woodlands of California. Journal of Mammalogy, 90: 114–118. Wood. G. W., Roark, D. N., 1980. Food habits of feral hogs in coastal South Carolina. Journal of Wildlife Management, 44: 506–511. Zanin Hegel C. G., Ângelo Marini, M., 2013. Impacto do javali Europeu, Sus scrofa, em um fragmento da Mata Atlântica Brasileira. Neotropical Biology and Conservation, 8(1): 17–24, doi: 10.4013/ nbc.2013.81.03 Zardo, A. O., Lima, G. J. M.., 1999. Alimentos para suínos, BIPERS, 71 p. (Boletim Informativo. Pesquisa, Extensão). Publicação conjunta do CNPSA e da EMATER/RS. Zeman, J., Hrbek, J., Drimaj, J., Kudláček, T., Kamler, J., Plhal, R., Heroldova, M., 2016. Comparison of three methods to evaluate wild boar diet. Folia Zoologica, 65(3): 221–224, https://doi.org/10.25225/ fozo.v65.i3.a7.2016 Zeman, J., Hrbek, J., Drimaj, J., Kudláček, T., Heroldova, M., 2018. Habitat and management influence on a seasonal diet composition of wild boar. Biologia, 73: 259–265, https://doi.org/10.2478/ s11756–018–0027–4

Cervo and Guadagnin

136

Appendix 1. Values and sources of information regarding the percentage of carbohydrates, lipids and proteins for each food item according to macro–nutritional analyses. Apéndice 1. Valores y fuentes de información relativa al porcentaje de carbohidratos, lípidos y proteínas de cada producto alimenticio según los análisis macronutricionales.

Item

Protein Carbohydrate Lipid

Herbs and leaves

0.25

Bromeliads

0.04 0.95 0.01 USDA/APHIS/WS (2010)

Corn and oats

0.15

Legumes

0.69 0.78

0.04

Source

0.07

USDA/APHIS/WS (2010) USDA/APHIS/WS (2010)

0.0028 0.997 0.0002 USDA/APHIS/WS (2010)

Corn

0.11 0.83 0.06 USDA/APHIS/WS (2010)

Corn and sorghum

0.12

0.82

0.05

USDA/APHIS/WS (2010)

Salvinia

0.34

0.51

0.15

Henry–Silva and Camargo (2002)

Oats

0.19 0.73 0.08 USDA/APHIS/WS (2010)

Sorghum

0.14 0.81 0.05 USDA/APHIS/WS (2010)

Rice

0.14 0.82 0.04 USDA/APHIS/WS (2010)

Rice and soybeans

0.28

Soybeans

0.42 0.36 0.22 USDA/APHIS/WS (2010)

Rosaceae

0.11 0.82 0.07 USDA/APHIS/WS (2010)

Grass

0.22 0.74 0.04 USDA/APHIS/WS (2010)

Amphibians

0.98

0.00

0.02

Waibel et al. (1987)

Mammals

0.9

0.00

0.1

Waibel et al. (1987)

Sheep

0.58

0.00

0.42

Pinheiro et al. 2007

Armadillo

0.9

0.00

0.1

Waibel et al. (1987)

Vertebrates

0.91

0.00

0.09

Waibel et al. (1987)

Birds

0.75

0.00

0.25

Waibel et al. (1987)

Roots

0.07 0.92 0.01 USDA/APHIS/WS (2010)

Fruits

0.17

0.27

0.56

Coimbra and Jorge (2011)

Seeds

0.25

0.69

0.06

Coimbra and Jorge (2011)

Araucaria seed

0.06

0.92

0.02

NEPA (2011)

Bivalve mollusks

0.69

0.22

0.09

USDA/APHIS/WS (2010)

Ants and beetles

0.45

0.00

0.55

Bukkens (1997)

Crustaceans

0.89 0.05 0.06 USDA/APHIS/WS (2010)

Invertebrates

0.63

0.18

0.2

Bukkens (1997)

Earthworms

0.79

0.18

0.03

Tacon et al. (1983)

Fly larva

0.75

0.18

0.07

Bukkens (1997)

Beetle larva

0.62

0.00

0.38

Bukkens (1997)

Wood

0.07 0.78 0.16 USDA/APHIS/WS (2010)

0.59

0.13

USDA/APHIS/WS (2010)

Animal Biodiversity and Conservation 43.1 (2020)

137

Recommendations for the management of sarcoptic mange in free–ranging Iberian ibex populations J. Espinosa, J. M. Pérez, A. Raéz–Bravo, P. Fandos, F. J. Cano–Manuel, R. C. Soriguer, J. R. López–Olvera, J. E. Granados

Espinosa, J., Pérez, J. M., Baéz–Bravo, A., Fandos, P., Cano–Manuel, F. J., Soriguer, R. C., López–Olvera, J. R., Granados, J. E., 2020. Recommendations for the management of sarcoptic mange in free–ranging Iberian ibex populations. Animal Biodiversity and Conservation, 43: 137–149, https://doi.org/10.32800/abc.2020.43.0137 Abstract Recommendations for the management of sarcoptic mange in free–ranging Iberian ibex populations. In recent decades, sarcoptic mange has become the main driver of demographic changes in Iberian ibex (Capra pyrenaica) populations in the Iberian Peninsula. Given this species' economic and ecological importance, priority must be given to management measures aimed at limiting the effects of this disease. However, despite the wealth of research on sarcoptic mange in ibex, no common patterns of action are yet available to manage this disease under field conditions. The lack of national and international protocols aimed at controlling sarcoptic mange has favoured the spontaneous emergence of various disease management initiatives in Spain. However, very little information is available concerning this trend and what there is tends to be available only as 'grey literature' or is consigned to the memory of local observers. Traditional strategies designed to combat this disease include the administration of medicated feed and the non–selective culling of mangy ibex. Here, we propose a management approach that takes into account aspects relating to the ecology and conservation of ibex populations, as well as public–health–related factors. Our recommendations are based on knowledge of the disease and host–parasite interaction, and aim to promote long–term advances in its control. Moreover, we discuss the efficacy of the measures traditionally used in mange management. The overall aim is to encourage debate between wildlife managers and motivate the development of alternative management strategies. Key words: Capra pyrenaica, Conservation, Management strategies, Sarcoptes scabiei, Wild populations Resumen Recomendaciones para el manejo de la sarna sarcóptica en poblaciones silvestres de cabra montés. En las últimas décadas, la sarna sarcóptica se ha convertido en la principal causa de los cambios demográficos en las poblaciones silvestres de cabra montés (Capra pyrenaica) de la península ibérica. Dada la importancia ecológica y económica de esta especie, se debe dar prioridad a las medidas de gestión destinadas a limitar los efectos de esta enfermedad. Sin embargo, a pesar del gran número de estudios que existen sobre la sarna sarcóptica en la cabra montés, actualmente no hay ningún protocolo de actuación común para el manejo de esta enfermedad sobre el terreno. La ausencia de protocolos nacionales e internacionales destinados a controlar la sarna sarcóptica ha favorecido la aparición espontánea de diversas iniciativas de gestión en España. Sin embargo, existe muy poca información sobre esta tendencia y la que hay solo suele estar disponible en la literatura gris o en la memoria de los observadores locales. Algunas de las estrategias tradicionales diseñadas para combatir esta enfermedad son la administración de piensos medicados y el sacrificio generalizado de los animales afectados. En este trabajo, proponemos un enfoque de gestión que tenga en cuenta aspectos relacionados con la ecología y la conservación de la cabra montés, además de factores relacionados con la salud pública. Nuestras recomendaciones se basan en el conocimiento de la enfermedad y la interacción entre el parásito y el hospedador y tienen por objeto impulsar progresos a largo plazo en su control. Además, analizamos la eficacia de las medidas utilizadas tradicionalmente en el manejo de la enfermedad. El objetivo general es fomentar el debate entre los gestores de fauna silvestre y motivar la elaboración de estrategias de gestión alternativas. Palabras claves: Capra pyrenaica, Conservación, Estrategias de gestión, Sarcoptes scabiei, Poblaciones silvestres ISSN: 1578–665 X eISSN: 2014–928 X

138

Espinosa et al.

Received: 3 IX 19; Conditional acceptance: 10 XII 19; Final acceptance: 10 I 20 José Espinosa, Department of Animal Health, Instituto de Ganadería de Montaña (IGM), CSIC–ULe, Facultad de Veterinaria, Campus de Vegazana s/n., 24071 León, Spain.– Jesús María Pérez, Departamento de Biología Animal, Biología Vegetal y Ecología, Universidad de Jaén, Campus Las Lagunillas s/n., 23071 Jaén, Spain. –Arián Ráez–Bravo, Jorge Ramón López–Olvera, Wildlife Ecology & Health group (Wild&EH) and Servei d’Ecopatologia de Fauna Salvatge (SEFaS), Departament de Medicina i Cirurgia Animals, Facultat de Veterinària, Universitat Autònoma de Barcelona (UAB), 08193, Bellaterra, Barcelona, Spain.– Paulino Fandos, Agencia de Medio Ambiente y Agua, Isla de la Cartuja, E–41092 Sevilla, Spain.– Francisco Javier Cano–Manuel, Departamento Actuaciones Forestales, Delegación Territorial, Consejería de Agricultura, Pesca, Ganadería y Desarrollo Sostenible, Granada, Spain.– Ramón Casimiro Soriguer, Estación Biológica de Doñana (CSIC), Av. Américo Vespucio, s/n., E–41092 Sevilla, Spain.– José Enrique Granados, Espacio Natural Sierra Nevada, Carretera Antigua de Sierra Nevada km 7, E–18071 Pinos Genil, Granada, Spain. Corresponding author: José Espinosa. E–mail: jespic@unileon.es ORCID ID: J. Espinosa: 0000-0002-9036-1402; J.M. Pérez: 0000-0001-9159-0365; A. Ráez-Bravo: 00000002-3190-1659; P. Fandos: 0000-0002-9607-8931; R. C. Soriguer: 0000-0002-9165-7766; J. López-Olvera: 0000-0002-2999-3451; J. E. Granados: 0000-0002-9787-9896

Animal Biodiversity and Conservation 43.1 (2020)

Introduction Awareness of the importance of actively managing infectious diseases in wild animals is a relatively novel phenomenon. Until recently, the general attitude was that 'nature can manage on its own'. However, the presence of humans and the enormous pressure they exert on the environment as a means of satisfying their requirements distorts this natural balance and artificial control measures are needed (Lyles and Dobson, 1993). Two good examples of such distortions are the elimination of large predators and the loss of biodiversity (Packer et al., 2003; Keesing et al., 2006). In many zones, this has led to an unsustainable overabundance of wild animals in their chosen habitats, which creates ideal conditions for the flare–up of disease (Rossi et al., 2005; Gortázar et al., 2006; Vicente et al., 2007). If these diseases are zoonotic in character or imply a threat to human economic activities, human action becomes inevitable as, for example, in the cases of rabies in wild carnivores (Pastoret and Brochier, 1999), classical swine fever in wild boar (Kaden et al., 2000) and tuberculosis in badgers (Woodroffe et al., 2005). Furthermore, the emergence of virulent forms of infectious agents or highly susceptible hosts may also jeopardize the structure of wild populations (Woodroffe, 1999), as occurred in the case of sarcoptic mange caused by the Sarcoptes scabiei mite. In the Iberian ibex (Capra pyrenaica), sarcoptic mange is at the root of the most serious demographic changes affecting this mountain ungulate in the Iberian Peninsula (Fandos, 1991; Pérez et al., 1997), to the extent that representative populations have become severely depleted and others are currently threatened (e.g. ibex populations in the Ports of Tortosa and Beceite and in the Game Reserve of Muela de Cortes) (unpublished data). Although sarcoptic mange is a constant threat to all populations of this mountain ungulate, currently this disease has not curbed the demographic trend of this species. However, the ecological, economic and social importance of the ibex (Granados, 2001), together with the sanitary risks that sarcoptic mange poses, obliges authorities and wildlife managers to instigate management control measures. One of the great inherent difficulties is that, like other infectious diseases (Kock et al., 2018), mange epidemics break out unexpectedly, often for obscure reasons (Pence and Ueckermann, 2002). As well, the pathogenesis of sarcoptic mange (in terms of morbidity, mortality and population effects) generally varies greatly from one affected area to another (Fandos, 1991; Mörner, 1992; Pérez et al., 1997; Fernández–Morán et al., 1997) and its effects are difficult to predict in affected populations for the first time. Given that the eradication of this disease is practically impossible (Wobeser, 2002), attempts are made to reduce its impact to 'tolerable' levels using a variety of control strategies areas; as such, management tasks designed to combat this disease are extremely complex. Despite the amount of research that has been carried out on sarcoptic mange in the ibex, no common evidence–based management strategy designed to combat this disease under field conditions exists. This absence of any national or international protocol for

139

sarcoptic mange has led to the emergence in Spain of spontaneous disease management initiatives that lack any consensus regarding which strategies are the most appropriate. Traditional strategies include the administration of medicated feed or the non–selective culling of mangy ibex. However, many management techniques generate serious social conflicts between animal rights activists, hunters and the local environmental agencies in charge of hunting activities. Furthermore, very little information is available on this question and what is available is generally either 'grey literature', that is, unpublished reports and conference proceedings, or it simply resides in the memories of local observers (Sánchez–Isarria et al., 2007a, 2007b). For this reason, we believe that it is essential to draw up a series of proposals for improved management and control of the spread of mange in the Iberian ibex. We believe that the selection of the most appropriate management techniques requires a clear understanding of the cause and ecology of this disease, as well as full knowledge of the course of the disease in individual ibex and the population biology of the parasite–host interaction. In light of the four categories used for the management of wildlife diseases (prevention, control, eradication and doing nothing i.e. laissez–faire) (Wobeser, 1994, 2002; Artois et al., 2001, Artois, 2003), here we propose action that lies halfway between laissez–faire and control, given that prevention and eradication under field conditions is an extremely complex task. We use published scientific evidence on sarcoptic mange in the ibex to develop a more 'ecological' approach to the management of this disease, which we believe is the best strategy for both the conservation of the species and future prevention. Additionally, we discuss the effectiveness of the measures traditionally used in mange management. We hope that this work will stimulate a debate among wildlife managers and motivate the development of alternative management strategies. Our aim is to promote a consensus regarding the best measures to adopt when confronted with sarcoptic mange, while ensuring optimal conservation of ibex. Preliminary considerations Initially, it is important to highlight certain aspects of the disease that will serve as premises in control strategies: (1) As a parasitic disease whose main mode of transmission is direct contact, sarcoptic mange can be categorized as a density–dependent process (Pence and Windberg, 1994; León–Vizcaino et al., 1999). (2) The clinical course in affected individuals (and therefore its effect on populations) is variable (Fandos, 1991; Górtazar et al., 1998; González–Candela et al., 2004). Effects will be conditioned by intrinsic factors relating to each individual and/or population (sex, age, genetics, health status, previous contact with the mite, etc.) and/or extrinsic factors (time of year, population density, infective dose, availability of trophic resources, etc.) (Rossi et al., 2007; Sarasa et al., 2010; López– Olvera et al., 2015; Pérez et al., 2017). (3) Except in residual or highly fragmented populations, in which the

Espinosa et al.

140

capacity to respond to changes is low (Skerratt, 2005), although mortality may at first be high (in many cases favoured by elevated densities), initial appearances can be deceptive and have no significant long–term effects on population dynamics (Pérez et al., 1997; Little et al., 1998). Given these considerations, the management measures taken to limit the effects of sarcoptic mange can be either active or pro–active. Active management refers to actions applied to the affected ibex population and/or the environment if sarcoptic mange is present in the ibex population and there is a desire to reduce its impact. Given its speed and effectiveness, host density reduction by culling is the most commonly used method (Sánchez–Isarria et al., 2007a). In the case of directly transmitted infections, population reduction is based on the epidemiological theory that –regardless of the severity of lesions– the per capita rate of disease transmission and the prevalence of the disease will grow as the population increases (Wobeser, 2002; Maxwell et al., 2013). The use of medicated feed containing Ivermectin or other macrocyclic lactones placed at different points in the wild has also been used as a mange management strategy in free–ranging ibex populations (Sánchez– Isarria et al., 2007b). Pro–active management measures include actions that aim to prevent future outbreaks of the disease and/or reverse the effects of mange epidemics. This group of techniques includes research, habitat improvement and the setting up of infrastructure. Furthering knowledge of pathological, immunological and epidemiological aspects of this disease improves our understanding of its development and helps determine in a coherent fashion the best measures for reducing its impact (Espinosa, 2018). Other important preventive measures in mange management in ibex include the restoration of pastures, the sowing of forage in time of scarcity or the modification of drinking fountains, aimed at strengthening the health status of the population; the veterinary control of livestock and safe translocations of ibex, to reduce the risk of mange transmission between domestic herbivores and ibex and/or disease–free ibex populations (Fandos 1991; Pérez et al., 1996); or the establishment of a stock reservoir of ibex in order to combat massive mortality outbreaks or for the strengthening some of ibex populations (Espinosa et al., 2017b). Management proposals against sarcoptic mange and discussion Here we suggest a series of alternative strategies for managing sarcoptic mange in free–ranging ibex populations. The selection of the most appropriate management technique requires a clear understanding of the ecology of the disease (including how the disease affects the individual), as well as intimate knowledge of the population biology of the parasite–host interaction. In the case of mange in ibex, we propose selective, less invasive action based on scientific evidence that, compared to traditional management measures, will

provide more long–term benefits and better protection against future mange outbreaks. In the interaction between sarcoptic mange and the Iberian ibex, four pathological phases or periods have been used to characterize the severity of infestations, according to the percentage of skin surface affected: 0, ibexes without skin lesions; 1, skin surface affected < 25 %; 2, skin surface affected > 25 and < 50 %; 3, skin surface affected > 50 and < 75 %; and 4, skin surface affected > 75 %) (León–Vizcaíno et al., 1999; Pérez et al., 2011). Given this, our management proposals focus exclusively on the selective culling of ibex in the chronic or final phases of the disease (phases 3 and 4). We rule out massive culling in mangy ibex populations and attempts to control the disease in the wild using pharmacological treatments. As we argue below, we believe that selective culling of ibexes in the final phases of the disease is the most appropriate and most reasonable measure for tackling sarcoptic mange in the ibex. Selective removal of infested ibexes From a clinical point of view, under natural conditions the multi–systemic clinical picture is severe and entails a very marked reduction in body condition, disorders of haematological and biochemical parameters (Pérez et al., 2015), septicaemic processes (Espinosa et al., 2017d), oxidative stress phenomena (Espinosa et al., 2017c), and an increase in inflammation biomarkers causing tissue damage in dermal and non–dermal tissues (Raéz–Bravo et al., 2015; Espinosa et al., 2017d), all of which greatly reduce survival possibilities and/ or hamper the recovery of ibexes in chronic phases of disease. For ethical and humanitarian reasons, the ending of the suffering of infected animals is necessary. In addition, ibex that have reached these stages of the disease can be a direct or indirect potential source of infestation for the rest of the population (Arlian et al., 1984; Pérez et al., 2011). Thus, unlike a non–intervention (laissez–faire) strategy (Wobeser, 2002), our low–level intervention approach will help reduce the risks of mange transmission within a population. Sarcoptic mange has side–effects that, in final phases of the disease, negatively affect the reproductive physiology of both male and female ibex (Sarasa et al., 2011; Espinosa et al., 2017a) and hinders their reproductive success. This makes them ineffective in prolonging the species and therefore unable to transmit to their offspring any type of response developed against the disease. In addition, given that mange is transmitted mainly by direct contact, recently born young are likely to be infected and to increase the affected population. This assumption is based on the observation of very young ibex with sarcoptic mange in herds with mangy adult specimens (Espinosa et al., 2017a), as well as the finding of mangy carcasses of juvenile ibex (J. E. Granados, pers. comm.).Thus, severely ill ibexes, with a low reproductive capacity and with a high risk of spreading the infestation to the rest of ibex population are determining factors to selectively remove these individuals and thus contribute to the control of the disease.

Animal Biodiversity and Conservation 43.1 (2020)

141

Fig. 1. Iberian ibex with sarcoptic mange: A, female ibex in early phases of the disease; B, male ibex with severe sarcoptic mange. Unlike the previous animal, according to our management proposals, this ibex should be removed from the ibex population. Fig. 1. Cabras montesas afectadas por sarna sarcóptica. A, hembra de cabra montés en las primeras fases de la enfermedad; B, macho de cabra montés con sarna sarcóptica intensa. A diferencia del animal anterior, de acuerdo con nuestras propuestas de manejo, este ejemplar debería ser eliminado de la población.

In all cases, the selective culling of these ibex must be carried out whenever possible by specialized staff using firearms. Capture using anaesthetic darts or systems of physical restraint that involves the pharmacological administration of a lethal drug –and therefore further manipulation– is a far more laborious and costly process (e.g. additional staff, the transfer of carcasses for incineration, etc.) with greater chances of failure on the welfare ground (due to additional capture–related stress; López–Olvera et al., 2009). The thorough disposal of the mangy skin of culled ibex reduces the possibilities of disease contamination and transmission and, unlike laissez–faire strategies (Wobeser, 2002) that advocate leaving dead ibex in the wild, eliminates a risk factor for the rest of the ibex population and for sympatric species. Effects of massive lethal control of sarcoptic mange Attempts to reduce or eliminate sarcoptic mange from the population by culling all mangy ibex– regardless of the severity of lesions– may also have unintended consequences on the population. No effective results for this type of management measure have ever been reported. For example, the spread of mange in Northern chamois (Rupicapra rupicapra) in the Eastern Alps (Italy, Austria and Slovenia) and Southern chamois in the Cantabrian Mountains (Spain) could not be controlled by culling visibly infected individuals (Meneguz et al., 1996; Fernández–Morán et al., 1997). One of the problems of this technique is that, in the event of epizootic outbreaks of disease, most of the large–scale culls take place before epidemiological studies of the response of the population to the disease are performed and so do not discriminate between

ibex in regression stages and those that are resistant to the disease. We believe that ibex in the initial and intermediate stages of the disease (phases 1 and 2) (León–Vizcaíno et al., 1999; Pérez et al., 2011) should never be removed from the population (fig.1). In the initial stages of the disease (Phases 1 and 2), the identification of sarcoptic mange with binoculars or telescopes can be confused in ibexes that, although without being parasitized by S. scabiei, show alterations of the coat due the seasonally heavy shedder (Valldeperes et al., 2019). Experimental infestations carried out on ibex from the Sierra Nevada Natural Space showed a wide variety of clinical responses varying from animals that progressed to severe or chronic phases to others that developed self–limiting clinic processes with mange lesions covering less than 50 % of the body surface, spontaneous regression lesions (move from phases 3–4 to 2–1) and even full recovery (Espinosa et al., 2017c) (fig. 2). Furthermore, scientific evidence show that a considerable proportion of mangy ibex recover naturally (Alasaad et al., 2013). Similar results were obtained in experimentally infested Northern chamois (R. rupicapra) in the Alps (Menzano et al., 2002) As well, the loss of genetic diversity from the population through the removal of resistant ibex or those in a recovery stage can have long–term negative consequences and even give rise to future, more severe epidemics due to a loss of herd immunity (Ebinger et al., 2011).The culling of diseased animals can even select for increased virulence as it means that there will be a greater number of relatively more susceptible hosts available for pathogens; this, in turn, stimulates pathogens to transmit more quickly to susceptible hosts to avoid being culled along with their hosts (Choo et

142

al., 2003). Another counterproductive consequence is that a reduction in population size by culling is often offset by the effects of compensatory reproduction and immigration (Caughley and Sinclair, 1994). Social structure disruption with increased movement and therefore increased contact rate (Donnelly et al., 2006) may lead to depopulation. This has been reported in the case of bovine tuberculosis in cattle and European badgers (Meles meles) (Woodroffe et al., 2006). Non–lethal control: drug treatment of infested animals The clinical management of wildlife is becoming increasingly frequent but is usually conducted only at the population level since individual treatment is largely impractical. Nonetheless, in tiny populations facing high extinction risks, vaccination and individual treatment may help manage a variety of infections (Wodroffe, 1999; Wobeser, 2002). For example, successful treatment against sarcoptic mange in a small population of Arctic foxes (Alopex lagopus) and individual treatment with anthelmintics in red grouse (Lagopus lagopus scoticus) have been reported (Dobson and Hudson 1992; Bornstein et al., 2001). Usually, it is impossible to capture and treat an entire population and so not all management programs have obtained conclusive results. Despite showing positive effects at an individual level in some cases, data on long–term population effects or re–infection rates of treatment against mange in Mednyi arctic foxes (Alopex lagopus semenovi), cheetahs (Acinonyx jubatus) and mountain gorillas (Gorilla beringei beringei) are inconclusive (Mwanzia et al., 1995; Goltsman et al., 1996; Kalema–Zikusoka et al., 2002). In the same way, following field trials the use of Ivermectin ‘bio–bullets’ to treat bighorn sheep (Ovis canadensis nelsoni) for psoroptic mange was rejected as a realistic management tactic (Jessup et al., 1991). If, in tiny populations, pharmacological treatment as a management strategy proves not to be totally efficient, in larger populations or over larger areas, such intervention may be inappropriate, unsustainable or simply impractical. Part of the problem in this regard is that the drugs that are used have often not been tested extensively in free–ranging wild animals, so that their actual efficacy is unknown. In addition, there is no consensus between the dose of drug administered and the severity of the infestation of the treated animals. Most of the studies consulted on attempts to control mange in free–ranging populations based their outcomes upon a few recaptured or remotely observed individuals and the long–term effects at the population level were in most cases inconclusive (see table 1). In the Iberian Peninsula, attempts to control sarcoptic mange outbreaks in ibex and chamois (Rupicapra pyrenaica parva) (Fernández–Morán et al., 1997) by dispersing feed treated with Ivermectin is a commonly used management strategy. Taking into account results in other species (Rowe et al., 2019), we believe that this strategy is impractical as to date there is no proof as to its effectiveness. In fact, treatment with Ivermectin is likely not warranted until designed studies have demonstrated its efficacy in free–ranging Iberian ibex populations.

Espinosa et al.

Captive mangy ibex treated with Ivermectin via subcutaneous injection needed four weeks of treatment before complete parasitological and clinical recovery. However, no positive results were obtained in chronically ill ibex (León–Vizcaíno et al., 2001), similarly as in other species (Skerratt, 2003; Kido et al., 2014). We consider that in ibex with severe sarcoptic mange the multi–systemic complications derived from the action of the mites (Espinosa et al., 2017d, 2017c) require additional treatment if the animal is to completely recover, which include the provision of intravenous fluids, antimicrobials and high–caloric nutrition. In free–ranging ibex, parenteral antiparasitic treatment is considered impractical for economic reasons, while the capture and handling of severely affected ibex for treatment purposes may often result in short–term mortality (López–Olvera et al., 2009). When attempting to control sarcoptic mange epizootics in ibex by administrating drugs orally, it must be taken into account that many areas of the natural habitat of the ibex are large and inaccessible (Acevedo and Cassinello, 2009). Therefore, it may never be possible to successfully implement long–term sarcoptic mange control in free–ranging ibex populations since dispensing drugs in this fashion implies no control of doses, no guarantee of repeated individual treatment, and very little acquired knowledge of effective therapeutic doses. It is also important to know whether other sympatric species act as reservoirs of the disease and contribute to its maintenance in the target host population, as shown in other multi–species models (Gakuya et al., 2012). We also believe that other ecological, ethical and public health issues must be addressed when contemplating the free dispensing of drugs in the natural environment (Artois et al., 2011; Crozier and Schulte–Hostedde, 2014). For example, the distribution of medicated feed at specific points can involve the aggregation of animals and therefore greater contact and an increase in infestation rates (Anderson and May, 1978). Environmental pollution via excreta with antiparasitic residues can cause a loss of biological diversity amongst invertebrate, vertebrate and microbial fauna, which in turn can be competitors in exogenous phases of other parasites (Verdú et al., 2018). Another important aspect is the development of acquired resistance by parasites to the chemicals present in the medication, which, due to the selective pressure in resistant organisms, is a clear risk in any program dependent upon the continued and widespread use of chemotherapy (Wobeser, 2002). This is significant not only for the treatment of mange in the ibex but also for the management of disease in domestic livestock and even in humans. Other less well–known effects, such as the teratogenic effects of Ivermectin intake during pregnancy or lactation, must also be taken into account (Bialek and Knobloch, 1999). The abuse of antiparasitic drugs can also affect the nature of the parasite–host interaction. For example, it is well known that previous contact with the mite induces a more intense and effective immune response in re–infestations (Sarasa et al., 2010). Thus, host– parasite relationships can be modified by continuous

Animal Biodiversity and Conservation 43.1 (2020)

143

Fig. 2. Iberian ibex experimentally infested with Sarcoptes scabiei showing a self–limiting process. A, development of the first lesions in the inoculation area (inter–scapular region); B, extension of the lesions over 50 % of the body surface; C, regression of the lesions, leaving small lesion centres on the rump; D, complete recovery of the animal with the disappearance of mangy lesions. To date this animal has not developed any further lesions or clinical signs of disease (Espinosa et al., 2017c). Fig. 2. Cabra montés infectada experimentalmente con Sarcoptes scabiei que muestra un proceso de infección autolimitante. A, aparición de las primeras lesiones en la zona de inoculación (región interescapular); B, extensión de las lesiones en el 50 % de la superficie corporal; C, regresión de las lesiones, que dejan pequeños centros de lesión en la grupa; D, recuperación completa del animal con la desaparición de las lesiones sarnosas. En la actualidad, este animal no ha vuelto a manifestar lesiones ni signos clínicos de la enfermedad (Espinosa et al., 2017c).

interference to the immune system as a result of the lack of sufficient parasitic antigens able to produce efficient and protective stimulation (Pedersen and Fenton, 2015). Therefore, the development of a correct immune response and the appearance of resilient ibex may be delayed or interrupted. Even if treatment does manage to eliminate mites from mangy ibex, recovered animals will not acquire long–lasting immunity and re–infestation may occur (Pederson, 1984; Wobeser, 2002). Another unintended consequence of the non–selective use of antiparasitic drugs in the wild is the possible modification of the balance with other parasites in both healthy and mangy ibex (Pérez et al., 2003; Thomas et al., 2005). Based on the above, we believe that in free–ranging species the application of uncontrolled antiparasitic treatment is inappropriate since evidence of proven efficacy is lacking and there is no solid scientific base to support its use as a

management measure. Until such scientific support is obtained, this type of management measure should be limited exclusively to the control of sarcoptic mange in captive wildlife (Rowe et al., 2019) or domestic livestock sharing territory with the ibex, thereby ensuring the safe and efficient control of a significant risk factor for the ibex population (Granados, 2001). On the other hand, it will be interesting to evaluate the efficacy of the new generation of isoxazoline parasiticides as an alternative to the use of macrocyclic lactones in the treatment of sarcoptic mange in ibex (Van Wick and Hashem, 2019). For more information on the success of such treatment in wild species, treated animals will have to be fitted with radio–collars to guarantee better individual monitoring. Finally, we believe that, bearing in mind the considerations outlined above, the culling of mangy ibex in chronic phases is justified. The public can be

Espinosa et al.

144

Table 1. Summary of attempts to treat mange in free–ranging wildlife and reported results. Treatments in captive wildlife are excluded: I, infestation (OM, Otodectic mange; PM, Psoroptic mange; SM, sarcoptic mange).

Host species Mednyi artic fox

I OM

Severity

Treatment and doses

Unreported

Alugan spray and ivermectin

(Alopex lagopus semenovi)

Unknow doses

Cheetah

Ivermectin: one doses

Mild to severe

(Acinonyx jubatus) and others

(200 μg/kg SC)

Bighorn sheep

Ivermectin 'bio–bullet'.

Mild or severe

(Ovis canadensis)

Doses unknown strategy

Red fox

Ivermectin;

Unreported

(Vulpes vulpes)

one doses (300 μg/kg SC)

Southern hairy nosed wombat SM

Ivermectin:

Mild or severe

(Lasiorhinus latrifrons)

one doses (300 μg/kg SC)

Bare–nosed wombat

Ivermectin: two doses

Mild to moderate

(Vombatus ursinus)

(400–800 μg/kg SC) +

Amitraz: one doses

(0.02 5% topical wash)

Mountain gorilla

Ivermectin: one doses

Mild to severe

(Gorilla beringei beringei)

(170–670 μg/kg IM) +

Antibiotic and

vitamin supplements

Hanuman langur

Tebrub: 30 doses (250 mg PO)

Moderate

(Semnophitecus entellus)

Mebhydrolin: 30 doses (25 mg PO)

Ivermectin: one doses (1 mg/kg SC)

Chlorpheniramine maleate:

one dose (10 mg SC)

Vicuña

Ivermectin (unknown doses

Mild to severe

Moderate to severe Unreported

(Vicugna vicugna) Giraffe

(Giraffa camelopardis) Agile wallaby

Moderate

(Macropus agilis) Gorilla

Severe

Ivermectin: two doses (300 μg/kg SC) Ivermectín: two doses

(Gorilla beringei beringei)

(200 mg/Kg SC) +

Antibiotic

Iberian ibex

Foxim: two doses

Mild to severe

(Capra pyrenaica)

(500 mg/l topical wash) +

Ivermectin: two doses

(02 mg/kg SC)

Koala

Amitraz: two doses

Unreported

(Phascolarcos cinereus)

(0.025 % topical wash)

Animal Biodiversity and Conservation 43.1 (2020)

145

Tabla 1. Resumen de los intentos de tratar la sarna en especies de fauna silvestre y resultados obtenidos. Quedan excluidos los tratamientos en cautividad: I, infestación (OM, sarna otodéctica; PM, sarna psoróptica; SM, sarna sarcóptica).

Effects on population

Reference

Non–significant increase in cub survival; effect

Goltsman et al. (1996)

on population viability unknown Recovery of some treated individual; inconclusive

Gayuka et al. (2012)

long–term population effects Not reported, but dismissed as a management

Jessup et al. (1991)

Ineffective results both in the individual

Newman et al. (2002)

and in the population Effective treatment only in animals with mild

Ruykys et al. (2013)

mange; population effects unknown Recovery of some treated individual; population

Skerratt et al. (2004)

effects unknown Recovery of all affected animals and mange

Kalema–Zikusoka et al. (2002)

control in the gorilla population

Recovery only of animals with parenteral

Chhangani et al. (2001)

treatment; population effects untested

Inconclusive treatment results

Gómez–Puerta et al. (2013)

Treated animals recovered. Certain success

Alasaad et al. (2012)

at the population level Successful on an individual level; population

McLelland and Youl (2005)

effects not evaluated Successful on an individual level and mange control

Graczyk et al. (2001)

in the gorilla population Recovery of some treated mild mangy individual.

León–Vizcaíno et al. (2001)

Dismissed as a management strategy at the population level Complete success at the individual and population level

Brown et al. (1982)

146

convinced of the need to remove some specimens without culling all affected ibex. Using solid arguments and without applying extreme management measures the aversion generated by human intervention in the control of the disease in wildlife can be minimized. Conclusions Parasites are natural parts of ecosystems and for this reason the presence of parasites in hosts does not necessarily imply that these wild populations are in danger of disappearing. Wildlife is a societal resource and provides ecological services that are vital for sustaining economies and human health. However, sometimes (as in the case of sarcoptic mange in Spanish ibex) a parasitic disease causes the death of a part of the population, thereby endangering the economic activities (e.g. ECO–tourism, hunting) that are derived from it. When attempting to control the disease and limit its effects, a lack of information and the need to make decisions quickly in a 'crisis' situation lead to the application of strategies that are not appropriate for the management of the disease. We believe that the management of diseases in wild animals generally requires solid scientific evidence grounded on corrective measures with potentially irremediable consequences for the future of the wild population. Short–term specific measures including pharmacological treatment and mass culls are too costly, too limited in duration and have little effect on overall population health. Given that it is difficult to predict where and when the next outbreak of sarcoptic mange will occur, further research is needed on the real effectiveness of the different management strategies of sarcoptic mange in field conditions, with proven results that may help wildlife managers in the decision–making process. Acknowledgements The authors acknowledge the support during this study from the Sierra Nevada Natural Space staff and the Consejería de Medio Ambiente (Junta de Andalucia). The scientific results on which this work is based were financed by the projects CGL2012–40043–C02–01, CGL2012–40043–CO2–02 and CGL2016–80543–P) (MEC, Spanish Government) and by the PAIDI Research Group RNM–118 from the Junta de Andalucía. We would also like to thank Michael Lockwood for the English revision. The authors would like to thank two anonymous expert reviewers and journal editors for their valuable comments on the manuscript. References Acevedo, P., Cassinello, J., 2009. What we know about Capra pyrenaica biology and ecology: A critical review and proposal for forthcoming agenda. Mammal Review, 39: 17–32. Alasaad, S., Ndeereh, D., Rossi, L., Bornstein, S., Permunian, R., Soriguer, R. C., Gakuya, F., 2012.

Espinosa et al.

The opportunistic Sarcoptes scabiei: a new episode from giraffe in the drought–suffering Kenya. Veterinary Parasitology, 185: 359–363. Alasaad, S., Granados, J. E., Fandos, P., Cano–Manuel, F. J., Soriguer, R. C., Pérez, J. M., 2013. The use of radio–collars for monitoring wildlife diseases: a case study from Iberian ibex affected by Sarcoptes scabiei in Sierra Nevada, Spain. Parasites & Vectors, 6(1): 242. Anderson, R. M., May, R. M., 1978. Regulation and stability of host parasite population interactions: I. Regulatory processes. Journal of Animal Ecology, 47: 219–247. Arlian, L. G., Runyan, R. A., Achar, S., Estes, S. A., 1984. Survival and infestivity of Sarcoptes scabiei var. canis and var. hominis. Journal of the American Academy of Dermatology, 11: 210–215. Artois, M., 2003. Wildlife infectious disease control in Europe. Journal of Mountain Ecology, 7: 89–97. Artois, M., Blancou, J., Dupeyroux, O., Gilot–Fromont, E., 2011. Sustainable control of zoonotic pathogens in wildlife: how to be fair to wild animals? Revue Scientifique et Technique–OIE, 30: 733. Artois, M., Delahay, R., Guberty, V., Cheeseman, C., 2001. Control of infectious diseases of wildlife in Europe. Veterinary Journal, 162: 141–152. Bialek, R., Knobloch, J., 1999. Parasitic infections in pregnancy and congenital parasitoses. II. Helminth infections. Zeitschrift für Geburtshilfe und Neonatologie, 203: 128–133. Bornstein, S., Mörner, T., Samuel, W. M., 2001. Sarcoptes scabiei and sarcoptic mange. In: Parasitic diseases of wild mammals, 2º edition: 107–120 (W. M. Samuel, M. J. Pybus, A. A. Kocan, Eds.). Iowa State University Press, Ames, Iowa, U.S.A. Brown, A. S., Seawright, A. A., Wilkinson, G.T., 1982. The use of amitraz in the control of an outbreak of sarcoptic mange in a colony of koalas. Australian Veterinary Journal, 58: 8–10. Caughley, G., Sinclair, A. R. E., 1994. Wildlife Ecology and Management. In: Blackwells Science: 334 (Editors). Editorial, Cambridge, Massachusetts, U.S.A. Chhangani, A. K., Mathur, B. R. J., Mohnot, S. M., 2001. First record of mange (sarcoptic mange) in free ranging Hanuman langur (Semnopithecus entellus) and its treatment around Jodhpur (Rajasthan), India. Intas Polivet, 2: 261–265. Choo, K., Williams, P. D., Day, T., 2003. Host mortality, predation and the evolution of parasite virulence. Ecology Letters, 6: 310–315. Crozier, G. K. D., Schulte–Hostedde, A. I., 2014. The ethical dimensions of wildlife disease management in an evolutionary context. Evolutionary Applications, 7: 788–798. Dobson, A. P., Hudson, P. J., 1992. Regulation and stability of a free–living host–parasite system: Trichostrongylus tenuis in red grouse. II. Population models. Journal of Animal Ecology, 61: 487–498. Donnelly, C. A., Woodroffe, A. R., Cox, D. R., Bourne, F. J., Cheeseman, C. L., Clifton–Hadley, R. S., Wei, G., Gettinby, G., Gilks, P., Jenkins, H., Johnston, W. T., Le Fevre, A. M., McInerney, J. P., Morrison,

Animal Biodiversity and Conservation 43.1 (2020)

W. I., 2006. Positive and negative effects of widespread badger culling on tuberculosis in cattle. Nature, 439: 843–846. Ebinger, M., Cross, P., Wallen, R., White, P. J., Treanor, J., 2011. Simulating sterilization, vaccination, and test–and–remove as brucellosis control measures in bison. Ecological Applications, 21: 2944–2959. Espinosa, J., 2018. Caracterización de la patología asociada a diferentes grados de infestación por Sarcoptes scabiei en cabra montés (Capra pyrenaica). Implicaciones en la gestión de la sarcoptidosis. PhD thesis, University of Murcia, Spain. Espinosa, J., Granados, J. E., Cano–Manuel, F. J., López–Olvera., J. R., Ráez–Bravo, A., Romero, D., Soriguer, R. C., Pérez, J. M., Fandos, P., 2017a. Sarcoptes scabiei alters follicular dynamics in female Iberian ibex through a reduction in body weight. Veterinary Parasitology, 243: 151–156. Espinosa, J., López–Olvera, J. R., Cano–Manuel, F. J., Fandos, P., Pérez, J. M., López–Graells, C., Ráez–Bravo, A., Mentaberre, G., Romero, D., Soriguer, R. C., Granados, J. E., 2017b. Guidelines for managing captive Iberian ibex herds for conservation purposes. Journal for Nature Conservation, 40: 24–32. Espinosa, J., Pérez, J. M., López–Olvera, J. R., Ráez–Bravo, A., Cano–Manuel, F.J., Fandos, P., Soriguer, R. C., Granados, J. E., Romero, D., 2017c. Evaluation of oxidant/antioxidant balance in Iberian ibex (Capra pyrenaica) experimentally infested with Sarcoptes scabiei. Veterinary Parasitology, 242: 63–70. Espinosa, J., Ráez–Bravo, A., López–Olvera, J. R., Pérez, J. M., Lavín, S., Tvarijonaviciute, A., Cano–Manuel, F. J., Fandos, P., Soriguer, R. C., Granados, J. E., Romero, D., Velarde, R., 2017d. Histopathology, microbiology and the inflammatory process associated with Sarcoptes scabiei infection in the Iberian ibex (Capra pyrenaica). Parasites & Vectors, 10: 596. Fandos, P., 1991. La cabra montés (Capra pyrenaica) en el Parque Natural de las Sierras de Cazorla, Segura y las Villas. ICONA–CSIC, Madrid. Fernández–Morán, J., Gómez, S., Ballesteros, F., Quirós, P., Benito, J., Feliu, C., Nieto, J., 1997. Epizootiology of sarcoptic mange in a population of cantabrian chamois (Rupicapra pyrenaica parva) in Northwestern Spain. Veterinary Parasitology, 73: 163–171. Gakuya, F., Ombui, J., Maingi, N., Muchemi, G., Ogara, W., Soriguer, R. C., Alasaad, S., 2012. Sarcoptic mange and cheetah conservation in Masai Mara (Kenya): epidemiological study in a wildlife/ livestock system. Parasitology, 139: 1587–1595. Goltsman, M., Kruchenkova, E. P., Macdonald, D. W., 1996. The Mednyi arctic foxes: treating a population imperilled by disease. Oryx, 30: 251–258. Gómez–Puerta, L. A., Olazabal, J., Taylor, C. E., Cribillero, N. G., Lopez–Urbina, M. T., Gonzalez, A. E., 2013. Sarcoptic mange in vicuna (Vicugna vicugna) population in Perú. Veterinary Record, 173: 269–269.

147

González–Candela, M., León–Vizcaíno, L., Cubero– Pablo, M. J., 2004. Population effects of sarcoptic mange in Barbary sheep (Ammotragus lervia) from Sierra Espuña Regional Park, Spain. Journal of Wildlife Disease, 40: 456–465. Gortázar, C., Acevedo, P., Ruiz–Fons, F., Vicente, J., 2006. Disease risks and overabundance of game species. European Journal of Wildlife Research, 52: 81–87. Gortázar, C., Villafuerte, R., Blanco, J. C., Fernández– De Luco, D., 1998. Enzootic sarcoptic mange in red foxes in Spain. Zeitschrift für Jagdwissenschaft, 44: 251–256. Graczyk, T. K., Mudakikwa, A. B., Cranfield, M. R., Eilenberger, U., 2001. Hyperkeratotic mange caused by Sarcoptes scabiei (Acariformes: Sarcoptidae) in juvenile human–habituated mountain gorillas (Gorilla gorilla beringei). Parasitology Research, 87: 1024–1028. Granados, J. E., 2001. Distribución y estatus de la cabra montés (Capra pyrenaica, Schinz 1838) en Andalucía. PhD thesis, University of Jaén, España. Jessup, D. A., De Forge, J. R., Sandberg, S., 1991. Biobullet vaccination of captive and free–ranging bighorn sheep. In: Wildlife production: conservation and sustainable development: 429–434 (L. A. Renecker, R. J. Hudson, Eds.). University of Alaska, Fairbanks, USA. Kaden, V., Lange, E., Fischer, U., Strebelow, G., 2000. Oral immunisation of wild boar against classical swine fever: evaluation of the first field study in Germany. Veterinary Microbiology, 73: 239–252. Kalema–Zikusoka, G., Kock, R. A., Macfie, E. J., 2002. Scabies in free–ranging mountain gorillas (Gorilla beringei beringei) in Bwindi Impenetrable National Park, Uganda. Veterinary Record, 150: 12–15. Keesing, F., Holt R. D., Ostfeld, R. S., 2006. Effects of species diversity on disease risk. Ecology Letters, 9: 485–498. Kido, N., Omiya, T., Kamegaya, C., Wada, Y., Takahashi, M., Yamamoto, Y., 2014. Effective treatment for improving the survival rate of raccoon dogs infected with Sarcoptes scabiei. Journal of Veterinary Medical Science, 76: 1169–1172. Kock, R. A., Orynbayev, M., Robinson, S., Zuther, S., Singh, N. J., Beauvais, W., E. Morgan, E. R., Kerimbayev, A., Khomenko, S., Martineau, H. M., Rystaeva, R., Omarova, Z., Wolfs, S., Hawotte, F., Radoux, J., Milner–Gulland, E. J., 2018. Saigas on the brink: Multidisciplinary analysis of the factors influencing mass mortality events. Science advances, 4(1): eaao2314. León–Vizcaíno, L., de Ybánez, M. M. R., Cubero, M. J., Ortiz., J. M., Espinosa, J., Pérez, L., Simón, M. A., Alonso, F., 1999. Sarcoptic mange in Spanish ibex from Spain. Journal of Wildlife Disease, 35: 647–659. León–Vizcaíno, L., Cubero, M. J., González–Capitel, E., Simón, M. J., Pérez, L., de Ybáñez, M. M. R., Ortiz, J. M., González–Candela, M., Alonso, F., 2001. Experimental ivermectin tratment of sarcoptic mange and establishment of a mange–free population of Spanish ibex. Journal of Wildlife Disease,

148

37: 775–785. Little, S. E., Davidson, W. R., Howerth, E. W., Rakich, P. M., Nettles, V. F., 1998. Diseases diagnosed in red foxes from the southeastern United States. Journal of Wildlife Diseases, 34: 620–624. López–Olvera, J. R., Marco, I., Montané, J., Casas– Díaz, E., Mentaberre, G., Lavín, S., 2009. Comparative evaluation of effort, capture and handling effects of drive nets to capture roe deer (Capreolus capreolus), Southern chamois (Rupicapra pyrenaica) and Spanish ibex (Capra pyrenaica). European Journal of Wildlife Research, 55: 193–202. López–Olvera, J. R., Serrano, E., Armenteros, A., Pérez, J. M., Fandos, P., Carvalho, J., Velarde, R., Cano–Manuel, F. J., Ráez–Bravo, A., Espinosa, J., Soriguer, R. C., Granados, J. E., 2015. Sex–biased severity of sarcoptic mange at the same biological cost in a sexually dimorphic ungulate. Parasites & Vectors, 8(1): 583. Lyles, A. M., Dobson, A. P., 1993. Infectious disease and intensive management: population dynamics, threatened hosts, and their parasites. Journal of Zoo and Wildlife Medicine, 24: 314–326. Maxwell, B. J., Mihaljevic, J. R., Arellano, A. L., Kueneman, J. G., Preston, D. L., Cross, P. C., Johnson, P. T., 2013. Taming wildlife disease: bridging the gap between science and management. Journal of Applied Ecology, 50: 702–712. McLelland, D. J., Youl, J. M., 2005. Sarcoptic mange in agile wallabies (Macropus agilis) in the Northern Territory. Australian Veterinary Journal, 83: 744–745. Meneguz, P. G., Rossi, L., Sommavilla, G., De Martin, P., Rodolfi, M., 1996. Sulla più temibile parassitosi della fauna alpina: la rogna sarcoptica del camoscio. Large Animals Review, 2: 75–83 Menzano, A., Rambozzi, L., Milinar Min, A. R., Rossi, L., 2002. Experimental infection of chamois with Sarcoptes scabiei. In: Mange and myasis of livestock. Proc Workshop EU–COST Action, Toulouse, France, 3–6 October 2001, 833: 60–64. Mörner, T., 1992. Sarcoptic mange in Swedish wildlife. Revue Scientifique et Technique (International Office of Epizootics), 11: 1115–1121. Mwanzia, J. M., Kock, R. A., Wambua, J. M., Kock, N. D., Jarrett, O., 1995. An outbreak of sarcoptic mange in free–living cheetahs (Acinonyx jubatus) in the Mara region of Kenya. In: Proceedings of the joint conference of the American Association of Zoo Veterinarians, the Wildlife Disease Association, and the American Association of Wildlife Veterinarians: 105–114 (R. E. Junge, Ed.) American Association of Zoo Veterinarians, Cornell University, Ithaca, New York, U.S.A. Newman, T. J., Baker, P. J., Harris, S., 2002. Nutritional condition and survival of red foxes with sarcoptic mange. Canadian Journal of Zoology, 80: 154–161. Packer, C., Hol, R. D., Hudson, P.J., Lafferty, K. D., Dobson, A. P., 2003. Keeping the herds healthy and alert: implications of predator control for infectious disease. Ecology Letters, 6: 97–802. Pastoret, P. P., Brochier, B., 1999. Epidemiology and control of fox rabies in Europe. Vaccine, 17:

Espinosa et al.

1750–1754. Pedersen, A. B., Fenton, A., 2015. The role of antiparasite treatment experiments in assessing the impact of parasites on wildlife. Trends in Parasitology, 31: 200–211. Pederson, J., 1984. Mission: desert bighorn. New Mexico Wildlife Magazine, 29: 2. Pence, D. B., Ueckermann, E., 2002. Sarcoptic manage in wildlife. Revue Scientifique et Technique (International Office of Epizootics), 21: 385–398. Pence D. B., Windberg, L. A., 1994. Impact of a sarcoptic mange epizootic on a coyote population. Journal of Wildlife Management, 58: 624–633. Pérez, J. M., Castro, I., Granados, J. E., Cano–Manuel, F. J., Fandos., P., Espinosa, J., Soriguer, R. C., 2017. Does Sarcoptes scabiei synchronize its breeding cycle with that of the Iberian Ibex, Capra pyrenaica? International Journal of Acarology, 43: 199–203. Pérez, J. M., Granados, J. E., Pérez, M. C., Márquez, F. J., Ferroglio, E., Rossi, L., 2003. A survey of the gastrointestinal nematodes of Spanish ibex (Capra pyrenaica) in a high mountain habitat. Journal of Parasitology, 89: 315–318. Pérez, J. M., Granados, J. E., Sarasa, M., Serrano, E., 2011. Usefulness of estimated surface area of damaged skin as a proxy of mite load in the monitoring of sarcoptic mange in free–ranging populations of Iberian wild goat, Capra pyrenaica. Veterinary Parasitology, 176: 258–264. Pérez, J. M., Ruíz–Martínez, I., Chirosa, M., 1996. Management of sarcoptic mange and host populations. The case of the Spanish ibex from Sierra Nevada. BIPAS, 15: 69–76. Pérez, J. M., Ruiz–Martínez, I., Granados, J. E., Soriguer, R. C., Fandos, P., 1997. The dynamics of sarcoptic mange in the ibex population of Sierra Nevada in Spain–influence of climatic factors. Journal Wildlife Research, 2: 86–89. Pérez, J. M., Serrano, E., Soriguer, R. C., González, F. J., Sarasa, M., Granados, J. E., Cano–Manuel, F. J., Cuenca, R., Fandos, P., 2015. Distinguishing disease effects from environmental effects in a mountain ungulate: seasonal variation in body weight, hematology and serum chemistry among Iberian ibex (Capra pyrenaica) affected by sarcoptic mange. Journal of Wildlife Disease, 51: 148–156. Ráez–Bravo, A., Granados, J. E., Cerón, J. J., Cano– Manuel, F. J., Fandos, P., Pérez, J. M., Espinosa, J., Soriguer, R. C., López–Olvera, J. R., 2015. Acute phase proteins in Iberian ibex (Capra pyrenaica) affected by sarcoptic mange. Parasitology Research, 114: 4005–4010. Rossi, L., Fraquelli, C., Vesco, U., Permunian, R., Sommavilla, G. M., Carmignola, G., Da Pozzo, R., Meneguz, P. G., 2007. Descriptive epidemiology of a scabies epidemic in chamois in the Dolomite Alps, Italy. Eurpean Journal of Wildlife Research, 53: 131–141. Rossi, S., Fromont, E., Pontier, D., Cruciere, C., Hars, J., Barrat, J., Pacholek, X., Artois, M., 2005. Incidence and persistence of classical swine fever in free–ranging wild boar (Sus scrofa). Epidemiology

Animal Biodiversity and Conservation 43.1 (2020)

and Infection, 133: 559–568. Rowe, M. L., Whiteley, P. L., Carver, S., 2019. The treatment of sarcoptic mange in wildlife: a systematic review. Parasites & Vectors, 12: 99. Ruykys, L., Breed, B., Schultz, D., Taggart, D., 2013. Effects and treatment of sarcoptic mange in southern hairy–nosed wombats (Lasiorhinus latifrons). Journal of Wildlife Diseases, 49: 312–320. Sánchez–Isarria, M. A., Hermoso, J., Theureau de la Peña, J., Casanova, G., Burgui, J. M., Sanchís, G., Arévalo, P. R. S., 2007a. Metodología empleada en la estrategia del control de la sarna sarcóptica en la cabra montes de la Reserva Valenciana de Caza de la Muela de Cortes entre los años 2002–2007 (Valencia). In: Tendencias actuales en el Estudio y Conservación de los caprinos europeos: 255–268 (J. E. Granados Torres, J. Cano–Manuel León., P. Fandos, R. Cadenas de Llano Aguilar, Eds.). Junta de Andalucía, Granada, Spain. – 2007b. La ivermectina como tratamiento de la sarna sarcóptica en el medio natural: vías de aplicación. In: Tendencias actuales en el Estudio y Conservación de los caprinos europeos: 269–276 (J. E. Granados Torres, J. Cano–Manuel León., P. Fandos, R. Cadenas de Llano Aguilar, Eds.) Junta de Andalucía, Granada, Spain. Sarasa, M., Rambozzi, L., Rossi, L., Meneguz, P. G., Serrano, E., Granados, J., González., F. J., Fandos., P., Soriguer, R. C., González, G., Joachim, J., Pérez, J. M., 2010. Sarcoptes scabiei: specific immune response to sarcoptic mange in the Iberian ibex Capra pyrenaica depends on previous exposure and sex. Experimental Parasitology, 124: 265–271. Sarasa, M., Serrano, E., Soriguer, R. C., Granados, J. E., Fandos, P., González, G., Joachim, J., Pérez, J. M., 2011. Negative effect of the arthropod parasite, Sarcoptes scabiei, on testes mass in Iberian ibex, Capra pyrenaica. Veterinary Parasitology, 175: 306–312. Skerratt, L. F., 2003. Clinical response of captive common wombats (Vombatus ursinus) infected with Sarcoptes scabiei var. wombati. Journal Wildlife Disease, 39: 179–192. – 2005. Sarcoptes scabiei: an important exotic pathogen of wombats. Microbiology Australia, 26: 79–81. Skerratt, L. F., Skerratt, J. H., Martin, R., Handasyde, K., 2004. The effects of sarcoptic mange on the behaviour of wild common wombats (Vombatus ursinus). Australian Journal of Zoology, 52: 331–339.

149

Thomas, F., Bonsall, M. B., Dobson, A. P., 2005. Parasitism, biodiversity, and conservation. In: Parasitism and Ecosystems: 124–139 (F. Thomas, F. Renaud, J. F. Guégan, Eds.). Oxford University Press, Oxford, UK. Valldeperes, M., Granados, J. E., Pérez, J. M., Castro, I., Ráez–Bravo, A., Fandos, P., López–Olvera, J. R., Serrano, E., Mentaberre, G., 2019. How sensitive and specific is the visual diagnosis of sarcoptic mange in free–ranging Iberian ibexes? Parasites & Vectors, 12(1): 1–7. Van Wick, M., Hashem, B., 2019. Treatment of sarcoptic mange in an American Black Bear (Ursus americanus) with a single oral dose of Fluralaner. Journal of Wildlife Diseases, 55: 250–253. Verdú, J. R., Cortez, V., Martínez–Pinna, J., Ortiz, A. J., Lumaret, J. P., Lobo, J. M., Sánchez–Piñero, F., Numa, C., 2018. First assessment of the comparative toxicity of ivermectin and moxidectin in adult dung beetles: Sub–lethal symptoms and pre–lethal consequences. Scientific Reports, 8(1): 14885. Vicente, J., Höfle., U., Garrido, J. M., Fernández De–Mera, I. G., Acevedo, P., Juste, R., Barral, M., Gortazar, C., 2007. Risk factors associated with prevalence of tuberculosis–like lesions in wild boar and red deer in South Central Spain. Veterinary Research, 38: 451–464. Wobeser, G. A., 1994. Investigation and management of disease in wild animals. In: Plenum: 265 (Editors). Springer Science and Business Media, New York. – 2002. Disease management strategies for wildlife. Revue Scientifique Et Technique (International Office of Epizootics), 21: 159–178. Woodroffe, R., 1999. Managing disease threats to wild mammals. Animal Conservation, 2: 185–193. Woodroffe, R., Donnelly, C. A., Cox, D. R., Bourne, F. J., Cheeseman., C. L., Delahay, R. J., Gettinby, G., Mcinerney, J. P., Morrison, W. I., 2005. Effects of culling on badger Meles meles spatial organization: implications for the control of bovine tuberculosis. Journal of Applied Ecology, 43: 1–10. Woodroffe, R., Donnelly, C. A., Jenkins, H. E., Johnston, W. T., Cox, D. R., Bourne, F. J., Cheeseman, C. L., Delahay, R. J., Clifton–Hadley, R. S., Gettinby, G., Gilks, P., Hewinson, R. G., McInerney, J. P., Morrison, W. I., 2006. Culling and cattle controls influence tuberculosis risk for badgers. Proceedings of the National Academy of Sciences, U.S.A., 103: 14713–14717.

Muñoz and Farfán

Animal Biodiversity and Conservation 43.1 (2020)

151

'Some like it alien': predation on invasive ring–necked parakeets by the long–eared owl in an urban area E. Mori, L. Malfatti, M. Le Louarn, D. Hernández–Brito, B. ten Cate, M. Ricci, M. Menchetti Mori, E., Malfatti, L., Le Louarn, M., Hernández–Brito, D., ten Cate, B., Ricci, M., Menchetti, M. 2020. 'Some like it alien': predation on invasive ring–necked parakeets by the long–eared owl in an urban area. Animal Biodiversity and Conservation, 43: 151–158, https://doi.org/10.32800/abc.2020.43.0151 Abstract 'Some like it alien': predation on invasive ring–necked parakeets by the long–eared owl in an urban area. Predation pressure by native species may limit the spread of alien invasive species, thus playing a pivotal role in the impact and implementation of management strategies. The ring–necked parakeet Psittacula krameri is one of the most widespread alien bird species in Europe, with nearly 70 established populations. Predators of this species include diurnal raptors, synanthropic corvids, and rodents. Here we report for the first time that long–eared owls Asio otus might have preyed upon parakeets in their night roosts. Analysis of 167 owl pellets showed that ring–necked parakeets made up over 10 % of the total volume of the diet of these owls in winter (32.93 % of absolute frequency), representing the most important prey species after murid rodents and passerine birds. Further studies are needed to investigate whether parakeet consumption by long–eared owls is only a local occurrence or whether it is widespread in European cities. If so, predation by long–eared owl may eventually lead to a form of parakeet control and may limit the impact of this introduced parakeet on native biodiversity. Key words: Urban environments, Asio otus, Psittacula krameri, Invasive species, Predation pressure Resumen El gusto por lo exótico: la depredación de la cotorra de Kramer invasora por el búho chico en una zona urbana. La presión predatoria que ejercen las especies nativas puede limitar la propagación de especies invasoras exóticas y, en consecuencia, tener un papel decisivo en los efectos y la aplicación de estrategias de gestión. La cotorra de Kramer, Psittacula krameri, es una de las especies de aves exóticas más extendida de Europa, donde tiene cerca de 70 poblaciones establecidas. Entre los depredadores de esta especie se encuentran rapaces diurnas, córvidos sinantrópicos y roedores. En este estudio observamos por primera vez que el búho chico, Asio otus, puede cazar cotorras en sus dormideros. El análisis de 167 excrementos de búho chico mostró que las cotorras de Kramer constituyen el 10 % de volumen total de la dieta de estos búhos en invierno (32,93 % de frecuencia absoluta) y son la presa más importante después de los roedores múridos y las aves paseriformes. Es necesario seguir estudiando esta cuestión para analizar si el consumo de cotorras de Kramer por el búho chico es solo un fenómeno local o si se ha generalizado en las ciudades europeas. En ese caso, es posible que, la depredación por el búho chico termine suponiendo una forma de control de la cotorra y limite el impacto de esta especie introducida en la biodiversidad autóctona. Palabras clave: Entornos urbanos, Asio otus, Psittacula krameri, Especie invasora, Depredación Received: 23 IX 19; Conditional acceptance: 14 I 20; Final acceptance: 22 I 20 Emiliano Mori, Dipartimento di Scienze della Vita, Università degli Studi di Siena, Via P. A. Mattioli 4, 53100, Siena, Italy.– Emiliano Mori, Luigi Malfatti, Libero Professionista, Empoli (Florence), Italy.– Marine Le Louarn, Laboratoire Population Environnement Développement, AMU–IRD, UMR 151, Aix–Marseille Université, France.– Dailos Hernàndez–Brito, Department of Conservation Biology, Estación Biológica de Doñana (CSIC), Avda. Américo Vespucio, 41092 Sevilla, Spain.– Mattia Menchetti, Dipartimento di Biologia, Università degli studi di Firenze, Via Madonna del Piano 6, 50019 Sesto Fiorentino (Florence), Italy and Institut de Biologia Evolutiva (CSIC–UPF), Passeig Marítim de la Barceloneta 37, 08003 Barcelona, Spain. Corresponding author: E. Mori. E–mail: moriemiliano@tiscali.it ORCID ID: Emiliano Mori: 0000-0001-8108-7950; D. Hernàndez–Brito: 0000-0002-5203-3512; M. Menchetti: 0000-0002-0707-7495 ISSN: 1578–665 X eISSN: 2014–928 X

Mori et al.

152

Introduction Biological invasions are one of the main causes of the global biodiversity crisis (Nentwig et al., 2018). Predation by native species may help limit the spread of alien species and thus limit their negative effects on native environments. However, relatively few studies are available on this topic (Santos et al., 2009; Sheehy and Lawton, 2014; Pintor and Byers, 2015). Alien species introduced through the pet market are particularly appreciated (cf. Bertolino, 2009) and often fed by humans, thus facilitating the establishment of naturalized populations, mostly within human settlements, such as in urban parks (Clergeau and Vergnes, 2011; Gyimesi and Lensink, 2012; Mori et al., 2019). Once established, alien species may become part of the diet of native predators (e.g. Fajardo et al., 2018; Nardone et al., 2018; Mori et al., 2018; Macià et al., 2019). Among avian predators, nocturnal raptors have been reported to be effective control agents for alien pest management (Labuschagne et al., 2016), and their presence in urban and suburban areas is increasing (Mori and Bertolino, 2015). The ring–necked parakeet Psittacula krameri (hereafter, RNP) is the most widespread alien bird species in Europe and the Mediterranean basin, with 69 populations in at least 37 countries, mostly in urban and peri–urban areas (Menchetti et al., 2016; Pârâu et al., 2016; Grandi et al., 2018; Le Louarn et al., 2018). In addition, 116 alien populations have been identified elsewhere in the world, in the Americas, Africa, Asia, and Oceania (Menchetti et al., 2016). The RNP is a hole–nesting species that prefers tree cavities. It is gregarious with shared nocturnal roosts consisting of up to thousands of individuals (Luna et al., 2016; Pârâu et al., 2016; Le Louarn et al., 2017). Being very widespread as a pet animal and due to its bright colour (Menchetti and Mori, 2014), RNPs are widely appreciated by the general public (Clergeau and Vergnes, 2011; Le Louarn et al., 2016; Berthier et al., 2017), which may lead to precaution when planning management action (Crowley et al., 2019). Despite this global appreciation, this species has been reported to have a negative impact on native biodiversity, human activities, and the health of human/native species (for reviews, Menchetti and Mori, 2014; Menchetti et al., 2016; White et al., 2019). The most evident and severe impact of RNP is related to competition for roost sites with a threatened European species of conservation concern, the greater noctule bat Nyctalus lasiopterus (Hernández–Brito et al., 2014a, 2018). Very few anecdotal data on predators of RNPs in the invasive range occur in the scientific literature (Menchetti and Mori, 2014). In the UK and in Italy, the Eurasian sparrowhawk Accipiter nisus, the goshawk Accipiter gentilis, the peregrine falcon Falco peregrinus and the hobby Falco subbuteo may prey upon RNPs (Pithon and Dytham, 1999; Menchetti and Mori, 2014; Harris, 2015). Grey squirrels Sciurus carolinensis and Eurasian red squirrels Sciurus vulgaris may represent occasional predators on RNP chicks (Shwartz et al., 2008; Mori et al., 2013). Black rats Rattus rattus are potential nest predators and several aggressive interactions towards RNPs defending their nests have been recorded (Hernández–Brito et al., 2014b). Domestic

cats Felis catus have been reported to be effective killers of RNPs in central Italy (Menchetti and Mori, 2014). Corvids (i.e. the jackdaw Corvus monedula and the carrion crow Corvus corone) are predators of parakeet chicks in Belgium and Italy (Menchetti and Mori, 2014), but a quantification of predation rate is lacking. Predation on parrots by nocturnal raptors (e.g. the tawny owl Strix aluco, the barn owl Tyto alba and the long–eared owl Asio otus) has been suggested to occur at roosts, but no published scientific evidence is available (cf. Harms and Eberhard, 2003; Grandi et al., 2018). Despite being mainly specialized on Microtus voles (Selçuk et al., 2019), long–eared owls may adapt their feeding habits in the urban environments of European cities (where they commonly overwinter (Lövy and Riegert, 2013; Mori and Bertolino, 2015) to the most profitable prey, i.e. black rats and birds (Mori and Bertolino, 2015). In a recent study in Follonica in central Italy (Mori et al., 2017), it was reported that the population of RNPs declined notably after a winter roost in 2017 following the establishment of the long–eared owl in the immediate surroundings of the parakeet roost. The aim of present study, therefore, was to analyze the winter diet of the long–eared owl in this urban area to quantify the level of predation on RNPs. We also compared our results with the winter diet of long–eared owls in two other urban areas in central Italy to assess whether diet overlap occurred. Material and methods Study area Our study was conducted within the urban area of Follonica, in central Italy (Province of Grosseto: 42.92 ºN, 10.7 6 ºE: 0–7 m a.s.l.; fig. 1). The mean annual temperature was 16.8 ºC, with annual precipitation of 650–700 mm (Mori et al., 2017). This urban area is surrounded by farmland, mainly cereal and sunflower fields) and wide coastal pinewood forests with Pinus pinea (Mori et al., 2017). The first control site was located in the southern peripheral area of the city of Grosseto (42.77 ºN, 11.13 ºE: 14–18 m a.s.l.; fig. 1), in a small pinewood around the the city hospital (Martelli and Fastelli, 2013). A further control area was located in the plain area surrounding the International Airport of Pisa (43.68 ºN, 10.41 ºE: 1–2 m a.s.l.; fig. 1), characterized by fallows and cultivated areas surrounded by irrigation channels. Climatic conditions were similar in the study area and the two control sites (Martelli and Fastelli, 2013; Giunchi et al., 2014). The ring–necked parakeet in the study area A pair of RNP was first observed in Follonica in 1999 and the population peaked at 30–35 individuals in 2016 (Pârâu et al., 2016; Mori et al., 2017). In winter 2017, a group of eight long–eared owls established a roost at a distance of about 300 m from the parakeet roost. The following parakeet count (refer Luna et al., 2016, for methods) showed a dramatic decline in the local parakeet population (fig. 2).

Animal Biodiversity and Conservation 43.1 (2020)

153

Pisa Follonica Grosseto

500 km

Fig. 1. Location of the study area (Follonica) and the control areas (Grosseto and Pisa). Fig. 1. Ubicación de la zona de estudio (Follonica) y las zonas de control (Grosseto y Pisa).

Owl pellet analysis A total of 167 long–eared owl pellets, egested between November 2017 and February 2018, were collected once a week under the winter roost. Each collection lasted for at least one hour, for a total of 18 hours of pellet collection throughout the study. In the laboratory, pellets were dried and softened in hot water and 95% ethanol, to better separate parts corresponding to prey species, e.g. skulls, mandibles, insect fragments, beaks and feathers (Andrews, 1990; Cecere and Vicini, 2000). Food items were determined through a binocular microscope with 400× magnification (Olympus BX 51 microscope). After determination, they were stored in tubes at –20 ºC for successive analyses. Feathers and beaks of birds, mandibles of small mammals (Muridae and Soricidae) and insects (Coleopterans) were compared using national atlases (Nappi, 2001; Gaggi and Paci, 2014; Bird Skull Collections web page) and local reference collections were stored at the University of Siena (Ciampalini and Lovari, 1985). We calculated absolute (AF, number of occurrences of each prey category, when present/total number of pellets × 100) and relative (RF, number of occurrences of each prey category, when present/total number of occurrences of all prey items × 100) frequencies of occurrence for each prey category (Khan et al., 2018), using the R software (version 3.6.1., R Foundation for Statistical Computing, Vienna, Austria). In addition, we estimated the volume occupied in pellets by each prey category when present (VWP, volume of each prey category, when present, estimated by eye/total

estimated volume × 100) (Ciampalini and Lovari, 1985; Marassi and Biancardi, 2002) using the same software (Khan et al., 2018). As both AF and VWP were calculated only when the prey category was present, their summation may exceed 100 % (Khan et al., 2018). We evaluated the total volume in diet of each prey category by plotting AF and VWP in a graph (Kruuk, 1989), with isopleths connecting points of the same volume in diet (Kruuk and Parish, 1981; Marassi and Biancardi, 2002). The trophic niche breadth was measured using Levins standardised index (Bsta), which ranges from 0 (minimum breadth) to 1 (maximum breadth): Bsta = (B – 1) / Bmax – 1) where: B is the Levins index (B = 1/Σpi2 where pi is the proportion of each i–prey category identified in every pellet), and Bmax is the total number of prey categories (Krebs, 1999). Our results were then compared with those on winter feeding habits of the long–eared owl in an urban roost from nearby cities (i.e. Grosseto: Martelli and Fastelli, 2013; Pisa: N = 137 pellets collected near the airport and analyzed as described above). Diet overlap between the study area and the control area was estimated through the Pianka index (Pianka, 1974), which ranges between 0 (no overlap) and 1 (total overlap). In the control area, RNPs are not present. This index is computed by taking into account the proportion of records of each prey category at both study sites: Ojk = [3pij * pik/ 3p2ij * 3p2ik]1/2

Mori et al.

No of ring–necked parakeet

154

35 30 25 20 15 10 5 0 1999

2001

2003

2005

2007

2009 Year

2011

2013

2015

2017

Fig. 2. Ring–necked parakeet population size in Follonica between 1999 and 2018 (Mori et al., 2017, modified). Fig. 2. Tamaño de la población de la cotorra de Kramer en Follonica entre 1999 y 2018 (Mori et al., 2017, modificado).

where pij and pik are the proportions of each i–prey category identified in every pellet, respectively in the study site (j) and in the control area (k). The comparison was made by using these combined prey categories which are present at all the study sites: (i) murid rodents, (ii) voles, (iii) shrews, (iv) birds, and (v) insects. Results We obtained a total of 422 prey fragments. Eight prey categories were identified: (i) house mouse Mus domesticus; (ii) long–tailed field mouse Apodemus sylvaticus; (iii) black rat Rattus rattus; (iv) undetermined murid rodents; (v) shrews (Soricidae); (vi) RNP; (vii) other birds; and (viii) insects. The dominant dietary item in terms of relative frequency was the house mouse, followed by birds and long–tailed field mice (table 1). Insects were poorly represented (1.20 %), despite covering half of the pellets, when present. Accordingly, the house mouse represented nearly 50 % of the total volume in the owl diet, long–tailed field mice and urban avifauna represented 20 %, and RNP made up over 10 % (fig. 3). The standardised Levins index was 0.47 (i.e. 47 % of trophic niche breadth). Table 2 shows the proportion of prey species (relative frequency %) in control areas The diet of the long–eared owl in our study area overlapped with that of the control areas in Grosseto and Pisa by 96 % and 63 %, respectively (Pianka index = 0.96; 0.63). Diet overlap between Grosseto and Pisa was 65 %.

Discussion Throughout its range, the long–eared owl is reported mostly as a vole predator (see Birrer, 2009, for a review of 312 studies). Despite this, in urban environments, this raptor may shift its diet towards synanthropic, more profitable prey species (Mori and Bertolino, 2015). The long–eared owl hunts mainly in areas with low and sparse vegetation (Bertolino et al., 2001; Aschwanden et al., 2005), thus explaining presence of woodland prey species in its diet (Birrer, 2009). Urbanisation may reduce the availability of many rodent species, including forest–dwellers and semifossorial voles (Pirovano et al., 2000a; Baker et al., 2003; Angold et al., 2006). However, if winter roosts are located near woodlands or in rural areas, forest rodents (e.g. the bank vole Myodes glareolus) may be highly represented in the diet of this owl (Mori et al., 2014). Usually, in human–modified environments, this nocturnal raptor mainly feeds on large (e.g. rats) or gregarious prey species (e.g. birds: Mori and Bertolino, 2015), which provide it with highly profitable, easily captured prey items (Wijnandts, 1984; Pirovano et al., 2000b). The results from our study area are in line with previous studies showing that urban owls mainly focus on small sized murid rodents, including synanthropic house mice together with long–tailed field mice, possibly caught at ecotones with woodlands and peripheral areas (Wijnandts, 1984; Birrer 2009), followed by black rats (Mori and Bertolino, 2015). Interestingly, birds were also highly represented. Birds are usually a rare occurrence in the diet of the long–eared owl (~6 %, in the whole of its range: Birrer, 2009), However, some studies have found them to be significantly represent-

Animal Biodiversity and Conservation 43.1 (2020)

155

ed in urban areas in winter (e.g. Bezzel, 1972; Laiu and Murariu, 1998; Martelli and Fastelli, 2013). The most commonly found are colonial roosting species (e.g. the Europen serin Serinus serinus, the hawfinch Coccothraustes coccothraustes and sparrows Passer spp.) as they are easily caught in shared roosts, e.g. on bare tree branches (Laiu and Murariu, 1998; Martelli and Fastelli, 2013). Many predator species are known to adapt their diet to local prey availability and prey selection often reflects ease of capture (cf. Pavey et al., 2008; Paspali et al., 2013; Nardone et al., 2018). In our study, the RNP was an important prey for the long–eared owl, contributing over 10 % of the total volume in the diet. This parrot species shares colonial roosts (Clergeau and Vergnes, 2011; Luna et al., 2016) and is relatively large (125–135 gr; Tabethe et al., 2013), thus possibly providing owls with a large amount of food (Birrer, 2009). Only one roost of RNPs occurred in our study area (Pârâu et al., 2016), and it was located a few hundred meters from that of the long–eared owl. After winter 2017, the population size of RNPs declined sharply and six new roosting sites, used by 2–9 individuals, were detected along the coastline, from the study site to ~19 km northwards (with the northernmost currently established in Piombino, province of Livorno). However, we cannot rule out the possibility that predation by owls occurred while parakeets were still active (i.e. before sunset).

Table 1. Absolute frequency (AF in %), relative frequency (RF in %) and volume when present (VWP in %) of each prey category identified in the diet of the long–eared owl. Tabla 1. Frecuencia absoluta (AF en %), frecuencia relativa (RF en %) y volumen (VWP en %) de todas las categorías de presas identificadas presentes en la dieta del búho chico.

WWP

Prey categories

(%)

House mouse

85.03 36.44 53.79

Long–tailed field mouse 52.69 18.43 44.70 Black rat

20.96 7.42 57.03

Undetermined murids 8.38

3.18 42.00

Shrews (Soricidae)

1.06 23.00

2.99

Ring–necked parakeet 32.93 11.44 32.69 Other birds

32.34 21.61 57.47

Insects

1.20 0.42 50.00

100

Volume when present (%)

Total volume in diet (%)

60 Black rat Insects

Other birds

Undetermined murid

House mouse

Wild mouse

Ring–necked parakeet

20 Shrews

10 5 0

40 60 Absolute frequency (%)

1 100

Fig. 3. Diet of the long–eared owl: absolute frequency of occurrence (%) plotted versus the volume (%) of each food category, when present. Isopleths connect points of the same total volume in diet (%). Fig. 3. Dieta del búho chico: frecuencia absoluta de presencia (%) representada en relación con el volumen (%) de cada categoría de alimento presente. Las isolíneas conectan puntos del mismo volumen total en la dieta.

Mori et al.

156

Table 2. Relative frequency (RF %) of each prey category identified in the diet of the long–eared owl at each control area. Tabla 2. Frecuencia relativa (RF %) de todas las categorías de presas identificadas en la dieta del búho chico en cada zona de control.

Control area

Prey species

Grosseto

Pisa

Apodemus flavicollis

0.00 5.84

Apodemus sylvaticus

25.16 13.14

Apodemus sp.

0.00

Mus domesticus

49.69 8.76

Rattus norvegicus

0.00 8.03

Rattus rattus

0.63 7.30

Total murids

75.48

14.60

57.66

Arvicola italicus

0.00 22.63

Microtus savii

0.00 5.11

Total voles Crocidura leucodon Total shrews

0.00

27.74

0.00 0.73 0.00

0.73

Erithacus rubecula

0.00 0.73

Passer italiae

0.00 2.19

Turdus merula

0.00 1.46

Phoenicuros ochruros 0.00 0.73 Prunella modularis

0.00 0.73

Unidentified birds

23.27

Total birds Coleoptera Tenebrionidae Total insects

0.00

23.27

5.84

1.26

13.87

1.26

13.87

When using clustered food categories (i.e. murid rodents, voles, shrews, birds, and insects), diet habits of the long–eared owl in Follonica almost overlapped with those of the nearest other urban roosts of this species. Despite RNPs not being present, a high proportion of urban birds was also detected in the control area of Grosseto. Overlap was lower between our study area and the control area of Pisa where the main prey was the endemic Italian water vole Arvicola italicus (i.e. near 15 % of total volume in diet), as a local adaptation. Large voles (i.e. those belonging to the genus Arvicola) are mentioned as prey in 420 out of 1,215 long–eared owl prey lists (Birrer, 2009), but relative frequencies are between 1 and 9.3 %, with only seven studies showing higher frequencies, of between 10 % and 27 %. In central Italy, RNPs outcompeted native, declining scops owls from nesting sites, forcing them to occupy suboptimal breeding habitats to minimize competition (Mori et al., 2017). RNPs are

early breeders (Luna et al., 2017) and use tree cavities before they return from their African wintering grounds. Although RNPs may also be highly aggressive against larger predators (Hernández–Brito et al., 2014a), their antipredatory behaviour is mainly focused on breeding areas, showing a greater vulnerability at nocturnal roosts, which could be exploited by nocturnal predators. Hence, we confirm that long–eared owl may adapt their diet to the most profitable prey species, thus showing a wide diet plasticity ranging from small voles, when available (Birrer, 2009), to large synanthropic rodents and urban colonial birds (Laiu and Murariu, 1998; Martelli and Fastelli, 2013; Mori and Bertolino, 2015). Our findings show that this nocturnal raptor may also feed on RNPs, and that this prey species may be frequent in winter. To the best of our knowledge, this is the first report of a native owl preying on an alien parrot. Future work should better establish the magnitude of this impact on the local population of RNP and determine the potential effect on the scops owl population after the removal of the alien competitor. Acknowledgements We thank COST (European Cooperation in Science and Technology Actions) ES1304 “ParrotNet” for their support in the development of this manuscript. The contents herein are the responsibility of the authors and neither COST nor any person acting on its behalf is responsible for the use that may be made of the information contained herein. An anonymous reviewer, Chris Pavey (CSIRO, Darwin, Australia), and the Associate Editor, Montserrat Ferrer, kindly improved our first draft with their comments. To conclude, we thank Prof. Sandro Lovari and Prof. Francesco Ferretti (University of Siena), who allowed us to use laboratory facilities. References Andrews, P., 1990. Owls, cave and fossils. University of Chicago Press, Chicago, Illinois, USA. Angold, P. G., Sadler, J. P., Hill, M. O., Pullin, A., Rushton, S., Austin, K., Small, E., Wood, B., Wadsworth, R., Sanderson, R., Thompson, K., 2006. Biodiversity in urban habitat patches. Science of Total Environment, 360: 196–204. Aschwanden, J., Birrer, S., Jenni, L., 2005. Are ecological compensation areas attractive hunting sites for common kestrels (Falco tinnunculus) and long–eared owls (Asio otus)? Journal of Ornithology, 146: 279–286. Baker, P. J., Ansell, R. J., Dodds, P. A., Webber, C. E., Harris, S., 2003. Factors affecting the distribution of small mammals in an urban area. Mammal Review, 33: 95–100. Berthier, A., Clergeau, P., Raymond, R., 2017. From beautiful exotic to beautiful invasive: perceptions and appreciations of the rose–ringed parakeet (Psittacula krameri) in the metropolis of Paris. Annales de Géographie, 716: 408–434. Bertolino, S., 2009. Animal trade and non–indigenous

Animal Biodiversity and Conservation 43.1 (2020)

species introduction: the world–wide spread of squirrels. Diversity and Distributions, 15: 701–708. Bertolino, S., Ghiberti, E., Perrone, A., 2001. Feeding ecology of the long–eared owl (Asio otus) in northern Italy: is it a dietary specialist? Canadian Journal of Zoology, 79: 2192–2198. Bezzel, E., 1972. Einige Daten zur Ernährung oberbayerischer Waldohreulen (Asio otus). Anzeiger der Ornithologische Gesellschaft in Bayern, 11: 181–184. Bird Skull Collections, available online at: www. skullsite.com [Accessed on 17 March 2019] Birrer, S., 2009. Synthesis of 312 studies on the diet of the long–eared owl Asio otus. Ardea, 97: 615–624. Cecere, F., Vicini, G., 2000. Micromammals in the diet of the long eared owl (Asio otus) at the W.W.F.’s Oasi San Giuliano (Matera, Southern Italy). Hystrix, the Italian Journal of Mammalogy, 11: 47–53. Ciampalini, B., Lovari, S., 1985. Food habits and trophic niche overlap of the badger (Meles meles L.) and the Red fox (Vulpes vulpes L.) in a Mediterranean coastal area. Zeitschrift für Säugetierkunde, 50: 226–234. Clergeau, P., Vergnes, A., 2011. Bird feeders may sustain feral Rose–ringed parakeets Psittacula krameri in temperate Europe. Wildlife Biology, 17: 248–253. Crowley, S. L., Hinchliffe, S., McDonald, R. A., 2019. The parakeet protectors: understanding opposition to introduced species management. Journal of Environmental Management, 229: 120–132. Fajardo, M., Morales, M., Fontenla, E., Giordano, C., Mori, E., Mazza, G., 2018. Sighting of Southern grey shrikes preying on red palm weevil in two countries. Redia, 101: 193–196. Gaggi, A., Paci, A. M., 2014. Atlante degli Erinaceomorfi dei Soricomorfi e dei piccoli Roditori dell'Umbria. Dimensione Grafica Snc. (Editions), Spello (Perugia), Italy. Grandi, G., Menchetti, M., Mori, E., 2018. Vertical segregation by ring–necked parakeets Psittacula krameri in northern Italy. Urban Ecosystems, 21: 1011–1017. Giunchi, D., Albores–Barajas, Y. V., Baldaccini, N. E., Vanni, L., Soldatini, C., 2014. Feral pigeons: problems, dynamics and control methods. In: Integrated pest management and pest control–current and future tactics: 215–240 (S. Soloneski, M. Larramedy, Eds.). Intech Open Editions, Rijeka, Croatia. Gyimesi, A., Lensink, R., 2012. Egyptian Goose Alopochen aegyptiaca: an introduced species spreading in and from the Netherlands. Wildfowl, 62: 128–145. Harms, K. E., Eberhard, J. R., 2003. Roosting behavior of the brown–throated Parakeet (Aratinga pertinax) and roost locations on four southern Caribbean islands. Ornitologia Neotropical, 14: 79–89. Harris, A., 2015. Predation of rose–ringed parakeets by raptors and owls in Inner London. British Birds, 108: 349–353. Hernández–Brito, D., Carrete, M., Popa–Lisseanu, A. G., Ibáñez, C., Tella, J. L., 2014a. Crowding in the city: losing and winning competitors of an invasive bird. Plos One, 9: e100593. Hernández–Brito, D., Carrete, M., Ibáñez, C., Juste,

157

J., Tella, J. L., 2018. Nest–site competition and killing by invasive parakeets cause the decline of a threatened bat population. Royal Society Open Science, 5: 172477. Hernández–Brito, D., Luna, Á., Carrete, M., Tella, J. L., 2014b. Alien rose–ringed parakeets (Psittacula krameri) attack black rats (Rattus rattus) sometimes resulting in death. Hystrix, the Italian Journal of Mammalogy, 25.2: 121–123. Khan, U., Lovari, S., Ali Shah, S., Ferretti, F., 2018. Predator, prey and humans in a mountainous area: loss of biological diversity leads to trouble. Biodiversity and Conservation, 27: 2795–2813. Krebs, C. J., 1999. Ecological Methodology, Second Edition. Addison Wesley Longman, Menlo Park, California, USA. Kruuk, H., 1989. The social badger: ecology and behaviour of a group living carnivore (Meles meles). Oxford University Press, Oxford, UK. Kruuk, H., Parish, T., 1981. Feeding specialization of the European badger Meles meles in Scotland. Journal of Animal Ecology, 50: 773–788. Labuschagne, L., Swanepoel, L. H., Taylor, P.J., Belmain, S. R., Keith, M., 2016. Are avian predators effective biological control agents for rodent pest management in agricultural systems? Biological Control, 101: 94–102. Laiu, L., Murariu, D., 1998. The food of the long– eared owl (Asio otus otus L.) (Aves: Strigiformes) in wintering conditions of the urban environment in Romania. Travaux du Museum National d’Histoire Naturelle 'Grigore Antipa', 40: 413–430. Le Louarn M, Clergeau P, Briche E, Deschamps– Cottin, M., 2017. "Kill two birds with one stone": urban tree species classification using bi–temporal Pléiades images to study nesting preferences of an invasive bird. Remote Sensing, 9: 916. Le Louarn, M., Clergeau, P., Strubbe, D., Deschamps– Cottin, M., 2018. Dynamic species distribution models reveal spatiotemporal habitat shifts in native range–expanding versus non–native invasive birds in an urban area. Journal of Avian Biology, 49: jav–01527. Le Louarn, M., Couillens, B., Deschamps–Cottin, M., Clergeau, P., 2016. Interference competition between an invasive parakeet and native bird species at feeding sites. Journal of Ethology, 34: 291–298. Lövy, M., Riegert, J., 2013. Home range and land use of urban long–eared owls. The Condor, 115: 551–557. Luna, Á., Franz, D., Strubbe, D., Shwartz, A., Braun, M. P., Hernàndez–Brito, D., Malihi, Y., Kaplan, A., Mori, E., Menchetti, M., van Turnhout, C. A. M., Parrott, D., Chmielewski, F. M., Edelaar, P., 2017. Reproductive timing as a constraint on invasion success in the ring–necked parakeet (Psittacula krameri). Biological Invasions, 19: 2247–2259. Luna, Á., Monteiro, M., Asensio–Cenzano, E., Reino, L., 2016. Status of the rose–ringed parakeet Psittacula krameri in Lisbon, Portugal. Biologia, 71: 717–720. Macià, F. X., Menchetti, M., Corbella, C., Grajera, J., Vila, R., 2019. Exploitation of the invasive Asian Hornet Vespa velutina by the European Honey Buzzard Pernis apivorus. Bird Study, 66(3): 425–429,

158

doi: 10.1080/00063657.2019.1660304. Marassi, M., Biancardi, C. M., 2002. Diet of the Eurasian badger (Meles meles) in an area of the Italian Prealps. Hystrix, the Italian Journal of Mammalogy, 13: 19–28. Martelli, C., Fastelli, P., 2013. Svernamento e dieta del gufo comune Asio otus nella città di Grosseto. Gli Uccelli d’Italia, 38: 85–91. Menchetti, M., Mori, E., 2014. Worldwide impact of alien parrots (Aves Psittaciformes) on native biodiversity and environment: a review. Ethology, Ecology and Evolution, 26: 172–194. Menchetti, M., Mori, E., Angelici, F. M., 2016. Effects of the recent world invasion by ring–necked parakeets Psittacula krameri. In: Problematic wildlife. A cross–disciplinary approach: 253–266 (F. M. Angelici, Eds.). Springer, New York. Mori, E., Ancillotto, L., Menchetti, M., Romeo, C., Ferrari, N., 2013. Italian red squirrels and introduced parakeets: victims or perpetrators? Hystrix, the Italian Journal of Mammalogy, 24: 195–196. Mori, E., Ancillotto, L., Menchetti, M., Strubbe, D., 2017. ‘The early bird catches the nest’: possible competition between scops owls and ring–necked parakeets. Animal Conservation, 20: 463–470. Mori, E., Bertolino, S., 2015. Feeding ecology of long–eared owls in winter: an urban perspective. Bird Study, 62: 257–261. Mori, E., Menchetti, M., Dartora, F., 2014. Evidence of carrion consumption behaviour in the long-eared owl Asio otus (Linnaeus, 1758) (Aves: Strigiformes: Strigidae). Italian Journal of Zoology, 81: 471–475. Mori, E., Menchetti, M., Zozzoli, R., Milanesi, P., 2019. The importance of taxonomy in species distribution models at a global scale: the case of an overlooked alien squirrel facing taxonomic revision. Journal of Zoology (London), 307: 43–52. Mori, E., Zozzoli, R., Menchetti, M., 2018. Global distribution and status of introduced Siberian chipmunks Eutamias sibiricus. Mammal Review, 48: 139–152. Nappi, A., 2001. I Micromammiferi d’Italia. Simoni Editors, Naples, Italy. Nardone, V., Bosso, L., Della Corte, M., Sasso, M., Galimberti, A., Bruno, A., Casiraghi, M., Russo, D., 2018. Native red foxes depredate nests of alien pond sliders: Evidence from molecular detection of prey in scats. Mammalian Biology, 88: 72–74. Nentwig, W., Bacher, S., Kumschick, S., Pyšek, P., Vilà, M., 2018. More than '100 worst' alien species in Europe. Biological Invasions, 20: 1611–1621. Pârâu, L. G., Strubbe, D., Mori, E., Menchetti, M., Ancillotto, L., van Kleunen, A., White, R. L., Hernàndez–Brito, D., Le Louarn, M., Clergeau, P., Albayrak, T., Franz, D., Braun, M. P., Schroeder, J., Wink, M., 2016. Rose–ringed parakeet Psittacula krameri populations and numbers in Europe: a complete overview. Open Ornithology Journal, 9: 1–13. Paspali, G., Oruci, S., Koni, M., Wilson, I. F., Krystufek, B., Bego, F., 2013. Seasonal variation of small mammals in the diet of the barn owl (Tyto alba) in

Mori et al.

the Drinos River valley, southern Albania. Turkish Journal of Zoology, 37: 97–105. Pavey, C. R., Gorman, J., Heywood, M., 2008. Dietary overlap between the nocturnal letter–winged kite Elanus scriptus and barn owl Tyto alba during a rodent outbreak in arid Australia. Journal of Arid Environments, 72: 2282–2286. Pianka, E. R., 1974. Niche overlap and diffuse competition. Proceedings of the National Academy of Science, 71: 2141–2145. Pintor, L. M., Byers, J. E., 2015. Do native predators benefit from non–native prey? Ecology Letters, 18: 1174–1180. Pirovano, A., Rubolini, D., Brambilla, S., Ferrari, N., 2000a. Winter diet of urban roosting long–eared owls Asio otus in northern Italy: the importance of the brown rat Rattus norvegicus. Bird Study, 47: 242–244. Pirovano, A., Rubolini, D., de Michelis, S., 2000b. Winter roost occupancy and behaviour at evening departure of urban long–eared owls. Italian Journal of Zoology, 67: 63–66. Pithon, J. A., Dytham, C., 1999. Breeding performance of Ring–necked parakeets Psittacula krameri in small introduced populations in southeast England. Bird Study, 46: 342–347. Santos, A. F. G. N., Santos, L. N., García–Berthou, E., Hayashi, C., 2009. Could native predators help to control invasive fishes? Microcosm experiments with the Neotropical characid, Brycon orbignyanus. Ecology of Freshwater Fish, 18: 491–499. Selçuk, A. Y., Özkoç, Ö. Ü., Bilir, M. A., Kefelioğlu, H., 2019. Diet composition of the long–eared owl (Asio otus) in the Eastern Anatolia (Turkey). Turkish Journal of Forestry, 20: 72–75. Sheehy, E., Lawton, C., 2014. Population crash in an invasive species following the recovery of a native predator: the case of the American grey squirrel and the European pine marten in Ireland. Biodiversity and Conservation, 23: 753–774. Shwartz, A., Strubbe, D., Butler, C. J., Matthysen, E., Kark, S., 2008. The effect of enemy–release and climate conditions on invasive birds: a regional test using the ring–necked parakeet (Psittacula krameri) as a case study. Diversity and Distributions, 15: 310–318. Thabethe, V., Thompson, L. J., Hart, L. A., Brown, M., Downs, C. T., 2013. Seasonal effects on the thermoregulation of invasive rose–ringed parakeets (Psittacula krameri). Journal of Thermal Biology, 38: 553–559. White, R., Strubbe, D., Dallimer, M., Davies, Z. G., Davis, A. J. S., Edelaar, P., Groombridge, J., Jackson, H. A., Menchetti, M., Mori, E., Nikolov, B. P., Pârâu , L. G., Pečnikar, Ž. F., Pett, T. J., Reino, L., Tollington, S., Turbè, A., Shwartz, A., 2019. Assessing the ecological and societal impacts of alien parrots in Europe using a transparent and inclusive evidence–mapping scheme. NeoBiota, 48: 45–69, doi: 10.3897/neobiota.48.34222. Wijnandts, H., 1984. Ecological energetics of the long–eared owl (Asio otus). Ardea, 72: 1–92.

Animal Biodiversity and Conservation 43.1 (2020)

159

Yellow or transparent? Comparison of sticky traps for monitoring functional arthropod diversity in an olive agroecosystem A. Dimitrova, M. Milošević, T. Spanos, I. Livieratos, V. D. Gkisakis

Dimitrova, A., Milošević, M., Spanos, T., Livieratos, I., Gkisakis, V. D., 2020. Yellow or transparent? Comparison of sticky traps for monitoring functional arthropod diversity in an olive agroecosystem. Animal Biodiversity and Consevation, 43.1: 159–167, Doi: https://doi.org/10.32800/abc.2020.43.0159 Abstract Yellow or transparent? Comparison of sticky traps for monitoring functional arthropod diversity in an olive agroecosystem. A diverse and balanced arthropod community is known to play an important role in the olive canopy but monitoring methods are not always well defined. We monitored canopy arthropods in an olive orchard over two years, comparing the performance of yellow sticky traps and transparent sticky traps. Data used to compare the two types of traps were arthropod abundance, richness, diversity indices, species abundance distribution and aggregation of taxa in functional groups based on prioritized agroecosystem services. The total abundance of arthropods caught in the yellow traps was higher than that in the transparent traps but diversity in both traps was similar. Transparent traps may therefore be a valid option to assess biodiversity in an olive agrosystem as besides being less labor demanding than yellow traps, they are low cost and replicable, and do not damage the overall arthropod. Key words: Olive, Canopy, Agrobiodiversity, Trapping, Agroecology Resumen ¿Amarillas o transparentes? Comparación de trampas adhesivas para estudiar la diversidad funcional de los artrópodos en el ecosistema de los olivares. El valor de una comunidad de artrópodos diversa y equilibrada en el dosel de los olivares se ha descrito con frecuencia, no obstante, no siempre se ha definido debidamente una metodología adecuada para determinar dicho valor. Durante dos años se estudiaron los artrópodos del dosel de un olivar mediante trampas adhesivas amarillas y transparentes. Para comparar ambos tipos de trampa, se utilizaron los índices de abundancia, riqueza y diversidad, la distribución de la abundancia de las especies y la agregación de taxones en grupos funcionales según los servicios agroecosistémicos considerados prioritarios. Aparecieron diferencias en la abundancia total y varios taxones, y las trampas amarillas presentaron la mayor abundancia esperada, sin embargo, las trampas transparentes mostraron una comunidad de artrópodos del dosel con un grado parecido de diversidad y uniformidad. Utilizar trampas adhesivas transparentes como método de bajo costo, susceptible de ser reproducido y que exige menos trabajo puede resultar adecuado para determinar la biodiversidad sin perjudicar a la comunidad de artrópodos en su conjunto. Palabras clave: Olivar, Dosel, Agrobiodiversidad, Trampeo, Agroecología Received: 19 XII 19; Conditional acceptance: 05 II 20; Final acceptance: 19 II 20 Anastazija Dimitrova, Mina Milošević, Theodore Spanos, Ioannis Livieratos, Sustainable Agriculture Department, Mediterranean Agronomic Institute of Chania (CIHEAM–MAICh), Chania, Crete, Greece.– Vasileios D. Gkisakis, Hellenic Mediterranean University (HMU), Estavromenos, Heraklion, Crete, Greece. Corresponding author: Vasileios D. Gkisakis. E–mail: gkisakis@hmu.gr ORCID ID: A. Dimitrova: 0000-0003-2581-5487; M. Milošević: 0000-0002-8732-3237; T. Spanos: 0000-00015769-5331; I. Livieratos: 0000-0002-0494-6691; V. D. Gkisakis: 0000-0002-0602-5423 ISSN: 1578–665 X eISSN: 2014–928 X

Dimitrova et al.

160

Introduction Olive groves in the Mediterranean region often serve as the main land cover in a variety of agroecological zones, such as hilly areas and plain areas (Gkisakis et al., 2016). Their ecological impact as functioning agroecosystems is of high value. Olive groves can provide ecosystem services such as reduction of soil erosion rate and enhancement of biodiversity, and they may also have socio–economic relevance (Loumou and Giourga, 2003). However, these well–established agroecosystems are often at risk as intensive cultivation practices replace the traditional low–input olive cultivation, leading to a homogeneous landscape and significant biodiversity degradation (Sokos et al., 2013). Biodiversity in an agroecosystem provides stability and resilience (Altieri, 1999; Jackson et al., 2007), especially when considering the role of functional biodiversity (Ricotta, 2005). Indeed, the functional part of biodiversity delivers important agroecosystem services and has a greater influence on both individual and overall ecosystem processes than classical species diversity. As such, its assessment becomes a highly valuable process (Tilman et al., 1997). Additionally, abundance and diversity of arthropod communities often becomes a useful aspect to assess the short– term impact of agricultural practices and reflect on the deeper changes in the ecosystem over a longer period of time (Missa et al., 2009). The olive agroecosystem hosts a unique assembly of organisms, especially arthropods, that have the potential to enhance orchard productivity (Gkisakis et al., 2015). The canopy stratumsupports a particularly high arthropod diversity (Ozanne, 2001) that provides valuable agroecosystem services (Gkisakis et al., 2018). Through their parasitic and predatory behavior towards key pests to the olive crop, beneficial arthropods in the canopy act as a biological pest control (Ruano et al., 2004; Rei et al., 2010). Additionally, the olive canopy can provide shelter to some species that are sensitive to environmental stress and as such they can be used as an indicator of the overall agroecosystem health (Scalercio et al., 2009). Studying arthropod diversity is important for two main reasons (Clergue et al., 2005): (i) it represents a holistic agroecological approach to the biodiversity concept with accent on its beneficial aspects; and (ii) it serves to identify appropriate assessment methods to estimate, observe and manage agrobiodiversity. The ultimate goal of measuring and studying functional biodiversity is to understand the components that can improve crop productivity and achieve a satisfactory level of agricultural sustainability (Bárberi, 2013). A main challenge when assessing arthropods in habitats, such as the olive tree canopy, is the selection of suitable and standardized trapping methods (Basset et al., 1997). Three criteria must be considered for trapping canopy arthropods (Yi et al., 2012): (i) the feasibility of sampling costs; (ii) the ease of replication; and (iii) the suitability of the design both for the characteristics of the selected agroecosystem and the high mobility of the targets. Passive trapping methods are often chosen as they rely on the movement of the

arthropods towards the traps (Gullan and Cranston, 2005) and are a non–selective method that provides sufficient sampling intensity without repercussions on the arthropod populations (Missa et al., 2009). Sticky traps may be a reasonable option for continuous sampling, preserving individuals in a suitable condition for identification (Yi et al., 2012). These devices consist of a surface coated with highly adhesive glue that traps arthropods when they land or crawl on it (Basset et al., 1997). Sticky traps may be: i) non–attractive, in which case they are transparent and odorless; or ii) attractive, in which case they have appealing colors, shapes, or smells, and they may be set in specific positions to target certain arthropod group(s) (Young, 2005). When they are further characterized by low cost and ease of collection, sticky traps can be used in larger number and replications (Basset et al., 1997; Young, 2005), a significant advantage for rapid biodiversity assessments. The aim of our study was, first, to compare two types of sticky traps, transparent and yellow, as potentially passive trapping methods in order to assess the canopy arthropod community, especially its functional part, in the olive agroecosystem, and second, to evaluate the results for use in monitoring studies. Material and methods Study sites and sampling The study was conducted in a 0.7 ha sub–plot within a 28 ha olive orchard located in the region of Chania, in the north–west part of the island of Crete, Greece (35º 20' N, 24º 17' E) (fig. 1). The location is situated at an altitude of 120 m, in a zone that primarily has a Mediterranean climate with a mean annual temperature of 14.3 ºC and mean annual precipitation of 840 mm. The olive orchard under study has been cultivated for commercial purposes for the last 30 years, following the organic farming standards of EU legislation (EC) 834/2007. The orchard is planted with 'Koroneiki' olive tree variety (sp. Olea europaea var. microcarpa alba), one of the most prevailing Greek olive cultivars in Crete. Twenty weekly measurements were conducted over the course of the two–year study (in autumn 2017 and spring 2019). Sampling took place in autumn and spring and lasted five weeks, coinciding with optimal arthropod activity in the olive agroecosystem. The two–year duration provided a full observation over the biennial cycle of olive trees (Lavee, 2007).Ten study sites were used throughout the sampling period, following a randomized experimental approach, for full area coverage and higher uniformity (fig. 2). Therefore, a total of four hundred sampling units were used, as summed up by multiplying twenty sampling weeks by ten study sites and two trap types. Trapping methodology At each site, two types of traps, yellow sticky traps (YST) and transparent sticky traps (TST) were set in the central part of the canopy, on a metal wire,

Animal Biodiversity and Conservation 43.1 (2020)

20º 0' 0'' E

161

30º 0' 0'' E

40º 0' 0'' E

Crete Island

Greece

Chania region

50 km

Mediterranean Sea 0

20º 0' 0'' E

30º 0' 0'' E

285

570 km

40º 0' 0'' E

Fig. 1. Location of the olive orchard monitored. Light–shaded area on the map of the island indicates the county of Chania, where sampling took place. Fig. 1. Localización del olivar estudiado. El área sombreada en gris claro del mapa de la isla indica el condado de Chania, donde se realizó el muestreo.

at an average height of 1.70–2 m. For convenient replacement, traps were attached on the wire with metal binder clips, spaced out at 1.5 m intervals, and on an unobstructed position from branches and leaves to avoid overlapping and to assure similar trapping conditions. YST (25 x 10 cm) was a commercially available, ready–to–use product Horiver® (Kopper B. V. The Netherlands) without any volatile attractants; it is used for monitoring several different types of insects. TST was prepared before each sampling, by homogenously applying a thin layer of TemoPlastic® glue (Kollant SpA, Vigonovo, Italy) on both sides of a PVC binding cover, of A4 format (21 x 29.7 cm). The traps were collected weekly, immediately transported

in colorless plastic membranes, and preserved under laboratory conditions, thus avoiding any trap damage before taxonomization. Arthropods were identified on the traps using a stereomicroscope (Novex AP Euromax®, Holland). A central sub–part of the TST, corresponding to the size of YST, was marked and taken into consideration. Taxa were identified to taxonomic level of order as a practical and relevant approach to assess biodiversity (Cotes et al., 2011; Gkisakis et al., 2018). Following a functional diversity approach based on the prioritized agroecosystem services of biological pest control and the 'dis–services' delivered by pests, we established two separate functional groups (Barberi, 2013): the pest

6 m 6 m

Fig. 2. Map of the randomized experimental design of the sub–plot. Indicated trees refer to the monitoring sites. Fig. 2. Mapa del planteamiento del diseño experimental aleatorizado de la subparcela utilizada. Los árboles indicados se refieren a los sitios de control.

Dimitrova et al.

162

Table 1. Abundance per hectare of canopy arthropods, pests and functional taxa, richness and diversity indices of trapping methods in seasonal sampling and sum abundance for autumn 2017 and spring, 2019: YST, yellow sticky traps; TST, transparent sticky traps; BPC, biological pest control group of arthropods; Pests, main olive pests group; S: Richness; J, Pielou's index; H', Shannon index; 1–D, reverse Simpson's index; Σ, sum abundance for the whole sampling period (other taxa counted but not presented due to scarcity (< 0.05 %): Syrphidae, Margaronia Unionalis, Prays Oleae, Chrysopidae, Hemerobiidae). Tabla 1. Abundancia por hectárea de artrópodos de dosel, plagas y taxones funcionales, valores de los índices de riqueza y diversidad de diferentes metodologías de captura en muestreo estacional y suma de abundancia, para en el período comprendido entre otoño de 2017 y primavera de, 2019: YST, trampas adhesivas amarillas; TST, trampas adhesivas transparentes; BPC, grupo de control de artrópodos para plagas biológicas; Pests, principal grupo de plagas del olivar; S, riqueza; J, índice de Pielou; H', índice de Shannon; 1–D, índice recíproco de Simpson; Σ, suma de la abundancia de todo el período de muestreo (hay otros taxones contados que no se han representado debido a su escasez (< 0,05 %): Syrphidae, Margaronia Unionalis, Prays Oleae, Chrysopidae, Hemerobiidae).

Monitoring season

Autumn

Spring

Trap

YST

TST

Araneae

212

76 250 196

Diptera Asilidae Bactrocera oleae Hemiptera/Heteroptera

YST

TST

YST

TST

462 272

19,299 4,142 20,289 9,037

39,588 13,179

2,670 910 3,633 1,118

6,303 2,028

289 33 14

303 36

128

100

343

154

471

254

Other Hemiptera

1,641

138

690

217

2,331

355

Hymenoptera

3,308

1,229

4,492

2,621

7,800

3,850

Psytallia concolor

44 89

Ichneumonidae

497

307

643

186 68 1,140

405

Lepidoptera

286 248 498 383

784 631

Neuroptera

221

769 459

82 548 377

Psocoptera

8,532 995 961 302

9,493 1,297

Thysanoptera

5,413

13,814

Coleoptera Total abundance Pests BPC S

116 39,156

541

8,401

1,214

48 308 296 7,599

36,780

289 33 14

14,797 3

3,498 1,137 4,634 1,663 10

1,755

424 344 75,936

22,396

303 36 8,132 2,800 10

0.619 0.629 0.578 0.576

0.590 0.602

1.425

1.390

1–D

0.684 0.653 0.627 0.586

1.448

group (PG), and the biological pest control (BPC) group. The PG group was considered to have a negative function and included the main pests of the olive crop sp. Bactrocera oleae (Rossi) (Diptera, Tephritidae), sp. Prays oleae (Bernard) (Lepidoptera, Yponomeutidae) and sp. Margaronia unionalis (Hübner) (Lepidoptera, Crambidae) (Delrio, 1992). The BPC group was con-

1.331

1.325

1.386

0.655 0.620

sidered to have a positive function and referred to arthropods that showed parasitic and predatory behavior towards the main olive pests (Ruano et al., 2004; Rei et al., 2010; Gkisakis et al., 2018). These main pests included the order Araneae, families Syrphidae and Asilidae (order Diptera), family Ichneumonidae and species Psyttalia concolor (Szépligeti) (Hymenoptera:

Animal Biodiversity and Conservation 43.1 (2020)

163

Table 2. Results of the Mann–Whitney test, including U, Z and p values, as applied to the comparison between two trapping methods (YST and TST) in seasonal samplings (autumn and spring) and sum abundance (Σ) for autumn 2017–spring, 2019. Predetermined levels of significance used: * p < 0.05, ** p < 0.01; obtained p–values that are less than 0.001 are recorded as < 0.001. Tabla 2. Resultados de la prueba de Mann–Whitney, incluidos los valores U, Z y p, como aplicados a la comparación entre dos diferentes metodologías de captura diferentes (YST y TST) en muestreos estacionales (otoño y primavera) y la suma de abundancia (Σ) para en el período comprendido entre otoño de 2017 y primavera de 2019. Niveles predeterminados de significancia utilizados: * p < 0,05, ** p < 0,01; los valores de p obtenidos inferiores a 0,001 se registran como < 0,001.

Monitoring season Autumn

Spring

Σ p

Araneae

5.50** –3.37 < 0.001

42.50 –0.57 0.570 109.00* –2.46 0.014

Diptera

7.00** –3.25 < 0.001

18.00* –2.42 0.016

Asilidae

40.50 –0.72 0.472

32.00 –1.36 0.174 149.50 –1.37 0.172

Bactrocera oleae

3.00** –3.57 < 0.001

28.00 –1.88 0.060

96.00** –2.90 < 0.001

54.00

0.762

18.50* –2.39 0.017

152.00 –1.30 0.193

Other Hemiptera

0.00** –3.78 < 0.001

12.00** –2.88 < 0.001

14.00** –5.04 < 0.001

Hymenoptera

10.00** –3.03 < 0.001

21.00* –2.19 0.028

68.50* –3.56 < 0.001

0.730

17.00* –2.52 0.012

128.50 –1.95 0.051

13.00** –2.80 < 0.001

14.50** –2.69 < 0.001

70.00** –3.52 < 0.001

Hemiptera/Heteroptera

Psytallia concolor Ichneumonidae

0.30

45.50 –0.35

44.00** –4.22 < 0.001

Lepidoptera

49.50 –0.04 0.970

41.00 –0.68 0.496 179.50 –0.56 0.579

Neuroptera

20.00* –2.27

0.023

33.00 –1.29 0.198

Psocoptera

18.00* –2.42

0.015

30.00 –1.51 0.131 102.00** –2.65 < 0.001

Thysanoptera

19.00* –2.34

0.019

13.00** –2.80 < 0.001

Coleoptera

17.00* –2.50

0.012

44.00** –0.45 < 0.001

130,00 –1.90 0.058

Total abundance

7.00** –3.25 < 0.001

0.00* –3.78 < 0.001

17.00** –4.95 < 0.001

Pests

3.00** –3.57 < 0.001

33.50 –1.47 0.142 104.00** –2.69 0.007

BPC

28.00 –1.66

0.096

23.00* –2.04 0.041

56.00 0.45

0.650

46.50 –0.27 0.791 207.00 0.19 0.850

57.50

0.570

46.00 –0.30 0.762

1–D

49.00 –0.08 0.940

0.57

Braconidae: Opiinae), and families Chrysopidae and Hemerobiidae (order Neuroptera). Taxa of the suborder formerly known as Homoptera, were classified as 'other Hemiptera'. The hymenopteran family Formicidae was not considered due to the inappropriateness of the applied trapping method. Data analysis The two trapping methods were compared through measures of: i) total and specific taxa abundance, and abundance of functional groups (BPC and PG); ii) taxa richness (S); and iii) a set of diversity indices including Pielou's index (J), representing community

113.50* –2.34 0.019 73.50** –3.42 < 0.001

110.00* –2.44 0.015 208.50 0.23

0.818

42.50 –0.57 0.570 182.50 –0.47 0.636

evenness, Shannon index (H') and the reverse Simpson's index of diversity (1–D). These measures, except richness, were compared following a univariate statistical analysis approach, in SPSS 20.0® for Windows. Data normality was assessed by the Shapiro–Wilk test (p < 0.05) and was found to be not normally distributed, even after applying several transformation types. Therefore, the non–parametric Mann–Whitney test was run to assess the differences between the two trapping methods, with a significance reported at the predefined levels of p < 0.05 and p < 0.01. Whittaker plots (rank abundance curves) were also generated in order to visually represent the species abundance distribution (SAD) for the two trapping

Dimitrova et al.

164

100 % 90 % 80 % 70 % 60 % 50 %

YST TST

40 % 30 % 20 % 10 % 0 %

Taxa

Fig. 3. Distribution (%) of the most dominant arthropod taxa in the YST and TST: A, Araneae; B, Diptera; C, Heteroptera; D, Homoptera; E, Hymenoptera; F, Lepidoptera; G, Neuroptera; H, Psocoptera; I, Thysanoptera; J, Coleoptera. Fig. 3. Distribución (%) de los taxones de artrópodos más dominantes en las trampas amarillas (YST) y las trampas transparentes (TST).(Para las abreviaturas de los taxones, véase arriba).

methods. This approach is considered well known and informative (Magurran, 2004), being of intermediate complexity between univariate descriptors, such as species richness and diversity indices and labeled lists of species abundances, typically analyzed by multivariate statistics (McGill et al., 2007). Results Arthropod abundance and diversity We captured a total of 98,332 arthropods during the sampling over the two years. Samples were classified in 10 orders and all were found in both types of traps. A total of 75,936 individuals were captured in the YST (77.22 % of total catches) and 22,396 (22.78 %) in TST (table 1). Univariate analysis delivered a significant difference (p < 0.01) between YST and TST in total abundance. This difference remained constant during sampling seasons in both years (table 2). Diptera was the most dominant taxon throughout the sampling period and for both trapping methods. It accounted for 52,767 catches, of which 39,588 (75.02 % of total catches) were in YST and 13,179 (24.98 %) were in TST. Thysanoptera, Psocoptera and Hymenoptera followed as the most abundant species in YST, while the other orders were present in abundance < 4 %. In TST, the most abundant taxa after Diptera were also Hymenoptera, Thysanoptera and Psocoptera, while the other orders were present in abundance < 3 % (fig. 3).

Abundance was higher in YST than in TST for the orders Araneae, Diptera, other Hemiptera, Hymenoptera, Neuroptera and Thysanoptera, families Asilidae (Diptera) and Ichneumonidae (Hymenoptera) and Bactrocera oleae (table 2). Results were similar when autumn and spring captures were considered alone, with the exceptions of Coleoptera (significantly higher for YST only in autumn), Heteroptera (significantly higher for YST only in spring), and Neuroptera and Psocoptera (no significant differences between traps, in spring captures) (table 2). Bactrocera Oleae was the most dominant pest, with 339 individuals captured in both traps, accounting for only 0.34 % of the total catches (table 1). The total number of pests captured by YST was significantly higher in autumn and in terms of total catches (table 2). The BPC functional group consisted of 10,932 individuals (11.12 % of total arthropod catches). Most BPC arthropods were captured by YST (74.39 % of the BPC group total catches), being statistically higher than those captured by TST (table 2). However, the percentages of BPC catches were 12.5 % for TST and 10.71 for YST. Asilidae was the most abundant arthropod in the BPC group accounting for 77.51 % of BPC arthropods in YST and 72.43 % in TST, followed by Ichneumonidae (14.02 % in YST and 14.4 6% in TST) and Araneae (5.68 % in YST and 9.71 % in TST). Abundance in the remaining groups was > 10 % (table 1). Diversity indices did not present any statistical differences between traps in any cases, either in spring or autumn sampling, over the two–year period (table 2).

Relative abundance (log)

Animal Biodiversity and Conservation 43.1 (2020)

165

YST TST

1.000 0.100 0.010 0.001

6 Rank

YST TST

1.000 0.100 0.010 0.001

6 Rank

Fig. 4. Whittaker plots (rank abundance curves) of YST and TST for autumn (A) and spring (B) measurements. Fig. 4. Gráficos de Whittaker (curvas de rango-abundancia) para las trampas amarillas (YST) y las trampas transparentes (TST) en las mediciones de de otoño (A) y primavera (B).

Nevertheless, both Pielou's index of evenness and the Shannon index presented relatively higher absolute numbers for TST, especially in autumn (table 1). In Whittaker plots, visualizing taxa abundance distribution (fig. 4), YST and TST appeared to have similarly shallow slopes, indicating relatively high evenness and confirming the higher evenness found using Pielou’s index (table 1). Discussion The differences in catches between YST and TST were evident in our study for almost all taxa captured over the two–year study period. Significant differences were observed in both seasons for several orders, indicating a natural preference towards yellow for taxa such as Diptera (Bekker et al., 2017), Homoptera (Castro et al., 2017), Hymenoptera (Thomson et al., 2004; Gullan and Cranston, 2005) and Thysanoptera (Thomson et al., 2004). The outcome was similar when the accumulative numbers of the catches of functional groups were considered, delivering statistically significant differen-

ces for both Pests and BPC group. The significantly higher numbers of catches of Bactrocera oleae by YST was expected, as the pest is often mentioned to be attracted to the colour yellow (Petacchi and Minnocci, 1994; Bekker et al., 2017). Additionally, the pest population in both traps was higher in autumn due to its natural cycle that corresponds to the maturation of the olives (Therios, 2009). With regards to the functional BPC group, the higher abundance of arthropods caught in spring both by YST and TST is also consistent with previous studies that reported a higher arthropod abundance in this season (Morris et al., 1999; Ruano et al., 2004; Gkisakis et al., 2018). Although the overall arthropod abundance on the YST was much higher in all cases than in TST, the level of arthropod diversity and richness in both traps was similar. Indeed, no statistically significant differences occurred between the several diversity indices used, the overall richness of species, or the species abundance distribution. Also, TST interestingly captured a relatively more representative portion of functional arthropods delivering BPC, when compared to YST. In a previous study, Thomson et al. (2004) compared YST and TST in vineyards and found similar results

Dimitrova et al.

166

in terms of arthropod richness, while YST appeared more effective in sampling Hymenoptera, Thysanoptera, Hemiptera, Diptera, Araneae and Coleoptera, and TST in sampling Lepidoptera and Neuroptera. This is consistent with our results and also explains the statistically significant difference for Araneae and Ichneumonidae (Hymenoptera) in favor of the YST. The above outcomes support the acceptance of both trapping methods as suitable for providing a clear representation of the constituents of the BPC groups. TST appears to be especially appropriate for rapid diversity assessments of canopy arthropods, as its non–attractive and interceptive nature avoids very high number of catches, but diversity and evenness is well represented. It has the added advantage that it can avoid damage to a high percentage of individuals in the traps and also ease the taxonomization effort (Yi et al., 2012). Furthermore, a high number of the catches achieved with YST were arthropods that are beneficial for olive production. Indeed, parasitic arthropods in the olive canopy have been reported to have a general preference towards YST, so a destructive effect on beneficial arthropod community is generated when they are used (Neuenschwander, 1982; Mazomenos et al., 2002). Other desirable properties of a trapping method, such as low cost, adaptability and potentiality of continuous passive sampling on study sites (Özden and Hodgson, 2016) also support the use of TST. As such, TST may offer greater potential in terms of biodiversity assessment, combining representative but non–damaging sampling of functional groups of arthropods of the olive canopy with economical and practical features. Conclusions The variety of arthropods in the canopy in the olive crop implies that more than a single sampling method may be adequately used in terms of biodiversity assessment. Specifically, TST proved to be a representative, easily replicable and low cost method that may enable acquisition of a broad data set for biological pest control strategies, within an agroecological framework in olive production. Additionally, arthropod identification at higher taxonomization levels, along with the arthropod organization in functional groups, has been further proven to be convenient and with a potential for use by non– entomological experts, in rapid on–field observations. Such an approach would be potentially useful for both agronomy–oriented and biodiversity conservation studies. Additional research, where TST would be observed alongside other trapping methods, is required to yield more comparative data, both in perennial and annual crops, and to contribute towards better understanding of the agroecosystem in question. Acknowledgements We would like to thank the owners of the olive orchard and the workers for their assistance during the measurements.

References Altieri, M. A., 1999. The ecological role of biodiversity in agroecosystems. Agriculture, Ecosystems and Environment, 74: 19–31. Bárberi, P., 2013. Functional Agrobiodiversity: the key to sustainability? In: Agricultural Sustainability: Progres and prospects in crop research: 3–20 (G. S. Bhullar, N. K. Bhullar, Eds.). Academic Press, Elsevier, London. Basset, Y., Springate, N. D.., Aberlenc, H. P., Delvare, G., 1997. A review of methods for sampling arthropods in tree canopies. In: Capony Arthropods: 27–52 (N. E. Stork, J. Adis, R. K. Didham, Eds.). Chapman & Hall, New York. Bekker, G. F. H. V. G., Addison, M. F., Addison, P., 2017. Comparison of two trap types for monitoring Bactrocera oleae (Rossi) (Diptera: Tephritidae) in commercial Olive groves of the Western Cape province, South Africa. African Entomology, 25: 98–107. Castro, J., Tortosa, F. S., Jimenez, J., Carpio, A. J., 2017. Spring evaluation of three sampling methods to estimate family richness and abundance of arthropods in olive groves. Animal Biodiversity and Conservation, 40.2: 193–210, Doi: https://doi. org/10.32800/abc.2017.40.0193. Delrio, G., 1992. Integrated Pest Control in Olive Groves. In: Proceedings of an international conference organized by the IOBC/WPRS Veldhoven, Netherlands, 8–13 September 1991: 67–76 (J. C. van Lenteren, A. K. Minks, O. M. B. de Ponti, Eds.). Pudoc Scientific Publishers, Wageningen. Clergue, B., Amiaud, B., Pervanchon, F., Lasserre–Joulin, F., Plantureux, S., 2005. Biodiversity: function and assessment in agricultural areas. A review. Agronomy for Sustainable Development Springer Verlag INRA, 25: 1–15. Cotes, B., Campos, M., García, P. A., Pascual, F., Ruano, F., 2011. Testing the suitability of insect orders as indicators for olive farming systems. Agricultural and Forest Entomology, 13: 357–364. Gkisakis, V. D., Kollaros, D., Bàrberi, P., Livieratos, I. C., Kabourakis, E. M., 2015. Soil arthropod diversity in organic, integrated, and conventional olive orchards and different agroecological zones in Crete, Greece. Agroecology and Sustainable Food Systems, 39: 276–294. Gkisakis, V. D., Volakakis, N., Kollaros, D., Bàrberi, P., Kabourakis, E. M., 2016. Soil arthropod community in the olive agroecosystem: Determined by environment and farming practices in different management systems and agroecological zones. Agriculture, Ecosystems and Environment, 218: 178–189. Gkisakis, V. D., Bàrberi, P., Kabourakis, E. M., 2018. Olive canopy arthropods under organic, integrated, and conventional management. The effect of farming practices, climate and landscape. Agroecology and Sustainable Food Systems, 42: 843–858. Gullan, P. J., Cranston, P. S., 2005. The insect an outline of entomology. Blackwell Publishing Ltd., West Sussex.

Animal Biodiversity and Conservation 43.1 (2020)

Jackson, L. E., Pascual, U., Hodgkin, T., 2007. Utilizing and conserving agrobiodiversity in agricultural landscapes. Agriculture, Ecosystems and Environment, 121: 196–210. Lavee, S., 2007. Biennial bearing in olive (Olea europaea L.). Annales Series Historia Naturalis, 17: 101–112. Loumou, A., Giourga, C., 2003. Olive groves: The life and identity of the Mediterranean. Agriculture and Human Values, 20: 87–95. Magurran, A. E., 2004. Measuring Biological Diversity. Blackwell, Oxford. Mazomenos, B. E., Pantazi-Mazomenou, A., Stefanou, D., 2002. Attract and kill of the olive fruit fly Bactrocera oleae in Greece as a part of an integrated control system. IOBC/WPRS Bulletin, 25: 137–146. McGill, B. J., Etienne, R. S., Gray, J. S., Alonso, D., Anderson, M. J., Benecha, H. K., Dornelas, M., Enquist, B. J., Green, J., He, F., Hurlbert, A., Magurran, A. E., Marquet, P. A., Maurer, B. A., Ostling, A., Soykan, C. U., Ugland, K., White, E., 2007. Species abundance distributions: moving beyond single prediction theories to integration within an ecological framework. Ecology Letters, 10: 995–1015. Missa, O., Basset, Y., Alonso, A., Miller, S. E., Curletti, G., De Meyer, M., Eardley, C., Mansell, M. W., Wagner, T., 2009. Monitoring arthropods in a tropical landscape: Relative effects of sampling methods and habitat types on trap catches. Journal of Insect Conservation, 13: 103–118. Morris, T. I., Campos, M., Kidd, N. A. C., Jervis, M. A., Symondson, W. O. C., 1999. Dynamics of the predatory arthropod community in Spanish olive grove. Agriculture and Forest Entomology, 1: 219–228. Neuenschwander, P., 1982. Beneficial insects caught by yellow traps used in the mass–trapping of the olive fruit fly, Dacus oleae. Entomologia Experimentalis et Applicata, 32: 286–296. Ozanne, C. M. P., 2001. Techniques and methods for sampling canopy Insects. In: Insect Sampling in Forest Ecosystems: 146–167 (S. R. Leather, Ed.). Blackwell Publishing, Blackwell Science Ltd., Oxford. Özden, Ö., Hodgson, D. J., 2016. The impact of tillage and chemical management on beneficial

167

arthropods in Mediterranean olive groves. Israel Journal of Ecology and Evolution, 9801: 1–5. Petacchi, R., Minnocci, A., 1994. Impact of different Bactrocera oleae (Gmel) control strategies on olive–grove entomofauna. Acta Horticulturae, 356: 399–402. Rei, F. T., Campos, M., Torres, L. M., 2010. Beneficial arthropods associated with the olive tree canopy in Alentejo (South region of Portugal). IOBC/WPRS Bulletin, 53: 117–127. Ricotta, C., 2005. A note on functional diversity measures. Basic and Applied Ecology, 6: 479–486. Ruano, F., Lozano, C., Garcia, P., Pena, A., Tinaut, A., Pascual, F., Campos, M., 2004. Use of arthropods for the evaluation of the olive–orchard management regimes. Agricultural and Forest Entomology, 6: 111–120. Scalercio, S., Belfiore, T., Noce, M. E., Vizzzarri, V., Iannota, N., 2009. The impact of compounds allowed in organic farming on the above–ground arthropods of the olive ecosystem. Bulletin of Insectology, 62: 137–141. Sokos, C. K., Mamolos, A. P., Kalburtji, K. L., Birtsas, P. K., 2013. Farming and wildlife in Mediterranean agroecosystems. Journal for Nature Conservation, 21: 81–92. Therios, I., 2009. Crop Production Science in Horticulture–Olives. CABI International Publishing, Wallingford. Thomson, L. J., Neville, P. J., Hoffmann, A. A., 2004. Effective trapping methods for assessing invertebrates in vineyards. Australian Journal of Experimental Agriculture, 44: 947–953. Tilman, D., Knops, J., Wedin, D., Reich, P., Ritchie, M., Siemann, E., 1997. The influence of functional diversity and composition on ecosystem processes. Science, 277: 1300–1302. Yi, Z., Jinchao, F., Dayuan, X., Weiguo, S., Axmacher, J. C., 2012. A comparison of terrestrial arthropod sampling methods. Journal of Resources and Ecology, 3: 174–182. Young, M., 2005. Insects in Flight. In: Insect Sampling in Forest Ecosystems: 116–145 (S. R. Leather, Ed.). Blackwell Publishing, Blackwell Science Ltd., Oxford.

168

Dimitrova et al.

Animal Biodiversity and Conservation 43.1 (2020)

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.

Les proves d'impremta enviades a l'autor per a la correcció, seran retornades al Consell Editor en el termini de 10 dies. Publicar articles a Animal Biodiversity and Conservation es gratuït per als autors. Les despeses degudes a modificacions substancials introduïdes per els autors en el text original acceptat aniran al seu càrrec. El primer autor rebrà una còpia electrònica del treball en format PDF. Manuscrits Els treballs seran presentats en format DIN A–4 (30 línies de 70 espais cada una) a doble espai i amb totes les pàgines numerades. Els manuscrits han de ser complets, amb taules i figures. No s'han d'enviar les figures originals fins que l'article no hagi estat acceptat. El text es podrà redactar en anglès, castellà o català. Se suggereix als autors que enviïn els seus treballs en anglès. La revista els ofereix, sense cap càrrec, un servei de correcció per part d'una persona especialitzada en revistes científiques. En tots els casos, els textos hauran de ser redactats correctament i amb un llenguatge clar i concís. Els caràcters cursius s’empraran per als noms científics de gèneres i d’espècies i per als neologismes intraduïbles; les cites textuals, independentment de la llengua, seran consignades en lletra rodona i entre cometes i els noms d’autor que segueixin un tàxon aniran en rodona. S'evitarà l'ús de termes extrangers (p. ex.: llatí, alemany,...). Quan se citi una espècie per primera vegada en el text, es ressenyarà, sempre que sigui possible, el seu nom comú. Els topònims s’escriuran o bé en la forma original o bé en la llengua en què estigui escrit el treball, seguint sempre el mateix criteri. Els nombres de l’u al nou, sempre que estiguin en el text, s’escriuran amb lletres, excepte quan precedeixin una unitat de mesura. Els nombres més grans s'escriuran amb xifres excepte quan comencin una frase. Les dates s’indicaran de la forma següent: 28 VI 99 (un únic dia); 28, 30 VI 99 (dies 28 i 30); 28–30 VI 99 (dies 28 a 30). S’evitaran sempre les notes a peu de pàgina.

Normes de publicació

Format dels articles

Els treballs s'enviaran preferentment de forma electrònica (abc@bcn.cat). El format preferit és un document Rich Text Format (RTF) o DOC que inclogui les figures i les taules. Les figures s'hauran d'enviar també en arxius apart en format TIFF, EPS o JPEG. Cal incloure, juntament amb l'article, una carta on es faci constar que el treball està basat en investigacions originals no publicades anterior ment i que està sotmès a Animal Biodiversity and Conservation en exclusiva. A la carta també ha de constar, per a aquells treballs en que calgui manipular animals, que els autors disposen dels permisos necessaris i que compleixen la normativa de protecció animal vigent. També es poden suggerir possibles assessors.

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

ISSN: 1578–665X eISSN: 2014–928X

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

La relació de referències bibliogràfiques d’un treball serà establerta i s’ordenarà alfabèticament per autors i cronològicament per a un mateix autor, afegint les lletres a, b, c,... als treballs del mateix any. En el text, s’indicaran en la forma usual: "... segons Wemmer (1998)...", "...ha estat definit per Robinson i Redford (1991)...", "...les prospeccions realitzades (Begon et al., 1999)...". Taules. Es numeraran 1, 2, 3, etc. i han de ser sempre ressenyades en el text. Les taules grans seran més estretes i llargues que amples i curtes ja que s'han d'encaixar en l'amplada de la caixa de la revista. Figures. Tota classe d’il·lustracions (gràfics, figures o fotografies) entraran amb el nom de figura i es numeraran 1, 2, 3, etc. i han de ser sempre ressenyades en el text. Es podran incloure fotografies si són imprescindibles. Si les fotografies són en color, el cost de la seva publicació anirà a càrrec dels autors. La mida màxima de les figures és de 15,5 cm d'amplada per 24 cm d'alçada. S'evitaran les figures tridimensionals. Tant els mapes com els dibuixos han d'incloure l'escala. Els ombreigs preferibles són blanc, negre o trama. S'evitaran els punteigs ja que no es reprodueixen bé. Peus de figura i capçaleres de taula. Seran clars, concisos i bilingües en la llengua de l’article i en anglès. Els títols dels apartats generals de l’article (Introducción, Material y métodos, Resultados, Discusión, Conclusiones, Agradecimientos y Referencias) no aniran numerats. No es poden utilitzar més de tres nivells de títols. Els autors procuraran que els seus treballs originals no passin de 20 pàgines (incloent–hi figures i taules). Si a l'article es descriuen nous tàxons, caldrà que els tipus estiguin dipositats en una institució pública. Es recomana als autors la consulta de fascicles recents de la revista per tenir en compte les seves normes. Comunicacions breus Les comunicacions breus seguiran el mateix procediment que els articles y tindran el mateix procés de revisió. No excediran de 2.300 paraules incloent–hi títol, resum, capçaleres de taula, peus de figura, agraïments i referències. El resum no ha de passar de 100 paraules i el nombre de referències ha de ser de 15 com a màxim. Que el text tingui apartats és opcional i el nombre de taules i/o figures admeses serà de dos de cada com a màxim. En qualsevol cas, el treball maquetat no podrà excedir de quatre

Animal Biodiversity and Conservation 43.1 (2020)

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.

Normas de publicación Los trabajos se enviarán preferentemente de forma electrónica (abc@bcn.cat). El formato preferido es un documento Rich Text Format (RTF) o DOC, que incluya las figuras y las tablas. Las figuras deberán enviarse también en archivos separados en formato TIFF, EPS o JPEG. Debe incluirse, con el artículo, una carta donde conste que el trabajo versa sobre 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 ISSN: 1578–665X eISSN: 2014–928X

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. Publicar artículos en Animal Biodiversity and Conservation es gratuito para los autores. Los gastos debidos a modificaciones sustanciales en las pruebas de imprenta, introducidas por los autores, irán a cargo de los mismos. El primer autor recibirá una copia electrónica del trabajo en formato PDF. 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. Formato de los artículos 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 enca© 2020 Museu de Ciències Naturals de Barcelona Papers are published under a Creative Commons Attribution 4.0 International License

bezado por el título del trabajo en cursiva. Key words en inglés (un máximo de seis) que especifiquen el contenido del trabajo por orden de importancia. Resumen en castellano, traducción del abstract. Su traducción puede ser solicitada a la revista en el caso de autores que no sean castellano hablantes. Palabras clave en castellano. 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 43.1 (2020)

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. Publishing in Animal Biodiversity and Conservation is free of charge. Expenses due to any substantial ISSN: 1578–665X eISSN: 2014–928X

alterations of the proofs will be charged to the authors. 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. © 2020 Museu de Ciències Naturals de Barcelona Papers are published under a Creative Commons Attribution 4.0 International License

Author’s address will be published as they appear 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.

Muñoz and Farfán

115–121 Monroy–Vilchis, O., Luna–Gil, A. A., Endara– Agramont, A. R., Zarco–González, M. M., González– Desales, G. A. Nevado de Toluca: habitat for Romerolagus diazi?

151–158 Mori, E., Malfatti, L., Le Louarn, M., Hernández–Brito, D., ten Cate, B., Ricci, M., Menchetti, M. ‘Some like it alien’: predation on invasive ring–necked parakeets by the long–eared owl in an urban area

123–136 Cervo, I. B., Guadagnin, D. L. Wild boar diet and its implications on agriculture and biodiversity in Brazilian forest–grassland ecoregions

159–167 Dimitrova, A., Milošević, M., Spanos, T., Livieratos, I., Gkisakis, V. D. Yellow or transparent? Comparison of sticky traps for monitoring functional arthropod diversity in an olive agroecosystem

137–149 Espinosa, J., Pérez, J. M., Ráez–Bravo, A., Fandos, P., Cano–Manuel, F. J., Soriguer, R. C., López–Olvera, J., Granados, J. E. Recommendations for the management of sarcoptic mange in free–ranging Iberian ibex populations

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 43.1 (2020) ISSN 1578–665 X eISSN 2014–928 X 1–7 Gómez–Hoyos, D. A., Seisdedos–de–Vergara, R., Schipper, J., Allard, R., González–Maya, J. F. Potential effect of habitat disturbance on reproduction of the critically endangered harlequin frog Atelopus varius in Las Tablas, Costa Rica 9–26 Mendes Pontes, A. R., Guedes Layme, V. M., Rodrigues de Lucena, L. R., Chivers, D. J. Tree reproductive phenology determines the abundance of medium–sized and large mammalian assemblages in the Guyana shield of the Brazilian Amazonia 27–35 Alviz., A., Pérez–Torres, J. A difference between sexes: temporal variation in the diet of Carollia perspicillata (Chiroptera, Phyllostomidae) at the Macaregua cave, Santander (Colombia) 37–41 Muñoz, A. R., Farfán, M. Á. European free–tailed bat fatalities at wind farms in southern Spain 43–54 García–Quintas, A., Fundora Caballero, D., Parada Isada, A. Taxonomic nestedness based on guilds? Bird assemblages of the Jardines de la Reina National Park, Cuba, as study case

55–66 Ávila–Nájera, D. M., Chávez, C., Pérez–Elizalde, S., Palacios–Pérez, J., Tigar, B. Coexistence of jaguars (Panthera onca) and pumas (Puma concolor) in a tropical forest in south–eastern Mexico 67–77 Martincová, I., Aghová, T. Comparison of 12 DNA extraction kits for vertebrate samples 79–87 Procheş, Ş. Does biogeography need species? 89–96 Moreno–Rueda, G. The evolution of crypsis when pigmentation is physiologically costly 97–107 Allen, M. L., Sibarani, M. C., Utoyo, L., Krofel, M. Terrestrial mammal community richness and temporal overlap between tigers and other carnivores in Bukit Barisan Selatan National Park, Sumatra 109–114 Ruiz–García, N. Effectiveness of the aposematic Eumaeus childrenae caterpillars against invertebrate predators under field conditions

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