ABC 44-2 (2021) 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 Jordi Domènech Tursiops truncatus, Dofí mular a prop de les Illes Medes, Delfín mular cerca de las islas Medas, Common bottlenose dolphin near the Medes Islands

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 44.2, 2021 © 2021 Museu de Ciències Naturals de Barcelona, Consorci format per l'Ajuntament de Barcelona i la Generalitat de Catalunya Autoedició: Montserrat Ferrer Fotomecànica i impressió: CEVAGRAF SCCL ISSN: 1578–665 X eISSN: 2014–928 X Dipòsit legal: B. 5357–2013

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Animal Biodiversity and Conservation 44.2 (2021)

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 Beatriz Arroyo Instituto de Investigación en Recursos Cinegéticos IREC–UCLM–CSIC–JCCM, Ciudad Real, Spain Xavier Bellés Institut de Biología Evolutiva UPF–CSIC, Barcelona, Spain Agustín Camacho Instituto de Biociências–USP, São Paulo, Brasil David Canal MTA Centre for Ecological Research, Vácrátót, Hungary Gonçalo C. Cardoso CIBIO–InBIO, Universidade do Porto, Portugal Salvador Carranza Institut de Biologia Evolutiva UPF–CSIC, Barcelona, Spain Luís Mª Carrascal Museo Nacional de Ciencias Naturales MNCN–CSIC, Madrid, Spain Martina Carrete Universidad Pablo de Olavide, Sevilla, Spain Pablo Castillo Institute for Sustainable Agriculture–CSIC, Córdoba, Spain Adolfo Cordero Universidad de Vigo, Vigo, Spain Mario Díaz Museo Nacional de Ciencias Naturales MNCN–CSIC, Madrid, Spain Darío Díaz Cosín Universidad Complutense de Madrid, Madrid, Spain José A. Donazar Estación Biológica de Doñana EBD–CSIC, Sevilla, Spain Arnaud Faille Museum National histoire naturelle, Paris, France Jordi Figuerola Estación Biológica de Doñana EBD–CSIC, Sevilla, Spain Gonzalo Giribet Museum of Comparative Zoology, Harvard University, 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 MNCN–CSIC, Madrid, Spain Jacob Höglund Uppsala University, Uppsala, Sweden Damià Jaume Institut Mediterrani d'Estudis Avançats IMEDEA–CSIC–UIB, Esporles, Spain José Jiménez Instituto de Investigación en Recursos Cinegéticos IREC–UCLM–CSIC–JCCM, Ciudad Real, Spain Miguel A. Jiménez–Clavero Centro de Investigación en Sanidad Animal–INIA, Madrid, Spain Jennifer A. Leonard Estación Biológica de Doñana EBD–CSIC, Sevilla, Spain Andras Liker University of Pannonia, Veszprém, Hungary Jordi Lleonart Institut de Ciències del Mar CMIMA–CSIC, Barcelona, Spain Josep Lloret Universitat de Girona, Girona, Spain Jorge Mª Lobo Museo Nacional de Ciencias Naturales MNCN–CSIC, Madrid, Spain Pablo J. López–González Universidad de Sevilla, Sevilla, Spain Ian MacGregor–Fors University of Helsinki, Lahti, Finland Jose Martin Museo Nacional de Ciencias Naturales MNCN–CSIC, Madrid, Spain Juan F. Masello Justus Liebig University Giessen, Giessen, Germany Santiago Merino Museo Nacional de Ciencias Naturales MNCN–CSIC, Madrid, Spain Manuel B. Morales CIBC–Universidad Autónoma de Madrid, Madrid Spain Joan Navarro Institut de Ciències del Mar, CMIMA–CSIC, Barcelona, Spain Juan J. Negro Estación Biológica de Doñana EBD–CSIC, Sevilla, Spain Daniel Oro Centre d'Estudis Avançats de Blanes CEAB–CSIC, Girona, Spain Vicente M. Ortuño Universidad de Alcalá de Henares, Alcalá de Henares, Spain Miquel Palmer Institut Mediterrani d'Estudis Avançats IMEDEA–CSIC–UIB, 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 EBD–CSIC, Sevilla, Spain Silvia Pérez–Espona The University of Edinburgh, UK Juan M. Pleguezuelos Universidad de Granada, Granada, Spain Montserrat Ramón Institut de Ciències del Mar CMIMA–CSIC, Barcelona, Spain Alex Richter–Boix CREAF, Universitat Autònoma de Barcelona, Bellaterra, Spain Diego San Mauro Universidad Complutense de Madrid, Madrid, Spain Ana Sanz–Aguilar Institut Mediterrani d'Estudis Avançats IMEDEA–CSIC–UIB, Esporles, Spain Rafael Sardà Centre d'Estudis Avançats de Blanes CEAB–CSIC, Girona, Spain Ramón C. Soriguer Estación Biológica de Doñana EBD–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 EBD–CSIC, Sevilla, Spain José L. Tellería Universidad Complutense de Madrid, Madrid, Spain Simone Tenan Institute of Marine Sciences (CNR–ISMAR), National Research Council, Venezia, 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 EBD–CSIC, Sevilla, Spain Rafael Villafuerte Instituto de Estudios Sociales Avanzados IESA–CSIC, Cordoba, Spain Rafael Zardoya Museo Nacional de Ciencias Naturales MNCN–CSIC, Madrid, Spain

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Human–wildlife interactions and people's attitudes towards conservation: a case study from Central Kerala, India S. K. Govind, E. A. Jayson

Govind, S. K., Jayson, E. A., 2021. Human–wildlife interactions and people's attitudes towards conservation: a case study from Central Kerala, India. Animal Biodiversity and Conservation, 44.2: 139–151, Doi: https:// doi.org/10.32800/abc.2021.44.0139 Abstract Human–wildlife interactions and people's attitudes towards conservation: a case study from Central Kerala, India. This paper studies the human–wildlife interaction in Central Kerala, India, and attempts to understand local people's attitude toward wildlife and conservation. Data were collected from April 2009 to March 2014. A structured questionnaire survey was carried out among people living in the fringe areas of the forest (n = 210). Self–reported household crop loss was modelled as a function of agricultural, demographic and environmental factors. Wild pig (Sus scrofa) (57.1 %) was the main crop foraging species, followed by Asian elephant (Elephas maximus) (12.9 %). It was reported that 36 % of farmers' annual income was lost due to crop foraging by wild animals. Leopard (Panthera pardus) (69.76 %), Indian rock python (Python molurus) (13.95 %), dhole (Cuon alpinus) (9.3 %) and stray dogs (6.97 %) were responsible for the attacks on livestock. The factors that influenced crop loss according to the farmers were the extent of agriculture land that they owned (coefficient = 0.968), the distance to reserve forest from crop fields (–0.009), and age of respondents (0.78). Due to people's awareness concerning the importance of wildlife, reports on human–wildlife interaction in the newspapers and strict enforcement of wildlife laws, people's attitude towards conservation of wildlife was good, and they were not taking any negative precautions against wild animals. Key words: Human–wildlife interaction, Wildlife conservation, Wildlife management Resumen La interacción entre los humanos y la fauna silvestre y la actitud de las personas en relación con la conservación de la fauna silvestre: un estudio de casos en Kerala central, en la India. El presente artículo estudia la interacción entre los humanos y la fauna silvestre en Kerala central, en la India, y trata de entender la actitud de las personas en relación con la conservación de la fauna silvestre. Los datos se recopilaron entre abril de 2009 y marzo de 2014. Se llevó a cabo una encuesta estructurada entre la población que habita en los márgenes de las zonas forestales (n = 210). Se elaboró un modelo de la pérdida de cultivos comunicada por los hogares como una función de factores agrícolas, demográficos y ambientales. El cerdo (Sus scrofa) era la principal especie que se alimentaba de los cultivos (57,1 %), seguida del elefante asiático (Elephas maximus) (12,9 %). Se comunicó que el 36 % de los ingresos anuales de los agricultores se perdió a causa de los animales silvestres. El leopardo (Panthera pardus) (69,76 %), el pitón de la India (Python molurus) (13,95 %), el perro salvaje asiático (Cuon alpinus) (9,3 %) y los perros callejeros (6,97 %) atacaron al ganado. La superficie de tierra agrícola propiedad de agricultores que habitan en los márgenes (el coeficiente es de 0,968), la distancia de la parcela de cultivo a la reserva forestal (–0,009) y la edad de los encuestados (0,78) fueron los factores significativos que influyeron en la pérdida de cultivos comunicada por los agricultores que habitan en los márgenes del bosque. Debido a la concienciación de la población acerca de la importancia de la fauna silvestre, los artículos de prensa sobre la interacción entre los humanos y la fauna silvestre y el cumplimiento estricto de la legislación en materia de vida silvestre, la actitud respecto de la conservación de la fauna silvestre era buena y no se estaban tomando precauciones negativas contra los animales silvestres. Palabras clave: Interacción entre humanos y fauna silvestre, Conservación de la fauna silvestre, Gestión de la fauna silvestre Received: 27 I 20; Conditional acceptance: 16 IV 20; Final acceptance: 24 III 21 Suresh K. Govind, Department of Psychology, Christ College, Irinjalakuda, Thrissur, Kerala, 680 125 India.– E. A. Jayson, Department of Wildlife Biology, Kerala Forest Research Institute, Peechi, Kerala, 680 653 India. Corresponding author: Suresh K. Govind. E–mail: sureshavinissery@gmail.com ISSN: 1578–665 X eISSN: 2014–928 X

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Introduction Negative interaction between humans and wildlife (human–wildlife conflict) occurs "when animals pose a direct and recurring threat to the livelihood or safety of people, leading to the persecution of that species" (http://www.hwctf.org). Concern for such interaction in India is growing. This human–wildlife conflict and people´s attitudes towards conservation is a multi–disciplinary area of research that deals with the dimensions of both humans and wildlife (Conover, 2002). Resolving this negative interaction depends not only on the biology of wild animals, but on the perceptions of local people and their attitude towards wildlife (Treves et al., 2006; Sillero–Zubiri et al., 2007; Jacobs et al., 2012). Data are needed on human dimensions such as political, social, cultural, historical, economic and legal problems (Madden, 2004). Awareness programmes on the importance of wildlife will increase tolerance (the ability to suffer the economic loss incurred by farmers due to wild animals) among people, and reportedly reduce the frequency of conflict (Sutherland, 2000; Mishra et al., 2003). Naughton–Treves et al. (2003) stated that human–wildlife interaction is also influenced by the lifestyle of people in forest fringes. As wildlife conservation is a major problem worldwide, fostering the co–existence (the state of being together in marginal areas) of humans and wildlife is mandatory to ease the situation (Madden, 2004). In Kerala, negative interaction between humans and wildlife is a contentious issue, with crop foraging by wild animals representing a major problem (Veeramani and Jayson, 1995; Veeramani et al., 2004; Jayson and Christopher, 2008). Due to activities such as the large–scale conversion of forest into monoculture plantations, shifting cultivations, hydroelectric projects and encroachments, the accessible habitat of wild animals is reduced in the State (Report of the Western Ghats Ecology Expert Panel, 2011). This has increased the risk of conflict where humans and wildlife co–exist. Recently, Govind and Jayson (2018a) identified the species of wild animals involved in crop foraging in Central Kerala, and estimated the actual economic loss incurred by the farmers due to these animals. To manage the conflict, data on people's attitudes towards wildlife conservation is also required. In this paper, we attempted to study the human–wildlife interaction in Central Kerala, India, and to understand the people's attitude to the conservation of wildlife. Material and methods Study area Thrissur district (10º 46' to 10º 7' N and 75º 57' to 76º 55' E) spans an area of 3,032 km2 in the central part of Kerala, India (fig. 1). The district has a tropical humid climate and a plentiful seasonal rainfall from the south–west monsoon (June to August) and the north–east monsoon (September to November).

Different types of soil, namely laterite, sandy loam, alluvial, clayey and black soil, are found. The district is comprised of 11 forest ranges within three forest divisions, namely Thrissur (210.64 km2), Chalakudy (279.71 km2) and Vazhachal (413.94 km2), and three wildlife sanctuaries (213.44 km2). Nearby vegetation types are moist deciduous (52.86 %), riverine (10 %) and plantations (37.14 %), including teak (Tectona grandis), rubber (Hevea brasiliensis) and cashew (Anacardium occidentale). Agriculture is the main occupation of people living in the fringe areas of the forest. Coconut (Cocos nucifera), arecanut (Areca catechu), rubber, cocoa (Theobroma cacao) and plantain (Musa paradisiaca) are the major cultivated crops. Multiple crops are cultivated in the private farms adjacent to the forest. Asian elephant (Elephas maximus), wild pig (Sus scrofa), leopard (Panthera pardus), chital (Axis axis), sambar (Rusa unicolor) and Indian giant squirrel (Ratufa indica) are the wild animals most commonly found in the forest. Methods A structured questionnaire survey (see supplementary material) was carried out among the people living on the fringe to identify the wild animals involved in the interaction with people, and to understand the people's attitude towards conservation of wildlife (Christopher, 1998). Data were collected from April 2009 to March 2014 as a part of a detailed study on human–wildlife conflict in Central Kerala. The study area was divided into grids of 2 km x 2 km (fig. 1). From each forest range, 10 % of the total grids were selected using the simple random method (table 1), and the houses within these grids were selected non–randomly. From each grid, ten houses were selected for the survey. A total of 210 houses were surveyed from six forest ranges. Non–forest areas towards the western side of the district, wildlife sanctuaries (Peechi–Vazhani and Chimmony) and some selected forest ranges (Athirapilly, Charpa, Vazhachal, Kollathirumedu and Sholayar) were omitted from grid selection as the human–wildlife interaction was negligible. These omitted protected areas, however, were visited to understand the type of control measures adopted to dissuade wild animals from approaching human habitation. Sixty questions were included in the questionnaire pro–forma (see supplementary material), mainly focusing on the details of the area, respondents characteristics, crops cultivated, crop foraging animals, methods for controlling crop foraging, livestock–lifting by carnivores, human–casualties due to wild animals, people's degree of dependence on agriculture, local beliefs regarding wildlife, local knowledge about wildlife laws, local people's opinion to mitigate human–wildlife interaction, and importance of conserving wildlife. Interviews were conducted primarily with the head of the household. If a household member over 30 years of age was absent during the survey, that house was skipped and the next house was approached. The questions of the survey sheet were prepared in English but were presented in the local language.

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10.607966

75.935525

141

76.205322

76.475119

76.744915

Malappuram District

India

Wadakkancherry Machad

10.428169

Pattikkad

Peechi Vazhani WLS

Thrissur

Arabian Sea

Palakkad District

Thrissur District

Kerala

Chimmony WLS

Palapilly

Velikullangara

Charpa Vazhachal

10.158373

Pariyaram

Wildlife Sanctuary Reserve Forest Non–forest areas Thrissur District boundary Forest range boundary Selected grid

Sholayar

Kollathirumedu Athirapilly Kalady

N W

Ernakulam District

E S

0 2.5 5

15 km

Fig. 1. Locations of the grids selected for the questionnaire survey. (Grid size 2 km x 2 km). Fig. 1. Ubicación de los cuadrantes seleccionados para el cuestionario. (Tamaño de los cuadrantes 2 km x 2 km).

Self–reported household crop loss (percentage crop loss/annum reported by a household member) was modelled as a function of agricultural, demographic and environmental factors. Several candidate models were prepared based on the variables of the questionnaire survey (Karanth et al., 2012, 2013). Various hypotheses about the characteristics of crop foraging reported in the survey were used to represent the models. Before we ran the models, we calculated Pearson's correlation coefficients to find the collinearity of variables involved in each model. The corrected Akaike's Information Criterion (AICc) was used to define the models and to assess and identify the variables. The models with substantial weight (cumulative weight > 0.95) were selected as the best models (Burnham and Anderson, 2002) to identify the factors that influenced the crop loss reported by the respondents. The reports on human–wildlife interaction in the print media (national, state and regional levels) were also collected from April 2009 to March 2012 and analysed. A case study was conducted at Ernakulam district (near Thrissur district), based on newspaper reports on human–wildlife interaction, and a discussion was held with key people of the area (n = 20) to understand the predator involved in the attack on livestock.

Results Details of the respondents Both males (76.67%) and females (23.33%) responded to the survey and their mean age was 54.4 ± 9.2 (range, 35–70) (n = 210). Most participants were native to the area (82.86 %) but others had migrated from urban areas (less than 10 years ago). The educational status revealed that 33.33 % of the respondents had completed lower primary school, 43.33 % had completed upper primary school, 12 % had passed the Secondary School Leaving Certificate (SSLC) examination, and 9.5 % had graduated. Only 1.9 % of the respondents were illiterate. Wells were the main source of drinking water (99 %) and wood from the forest was the main source of fuel (51.43 %). On average, the respondents in the immediate fringe areas of the forest owned 0.49 ± 0.48 ha of land. Crop foraging and livestock–lifting Coconut (95.71 %), plantain (85.24 %), rubber (62.38 %), arecanut (37.14 %), tubers (20.48 %),

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Table 1. Grids selected using the simple random method: NH, number of houses surveyed. Tabla 1. Cuadrantes seleccionados en los que se utiliza el método aleatorio simple: NH, número de casas encuestadas. Forest ranges

Total grids

Grids

Wadakkancherry

6, 34, 50, 55, 57, 59

Machad

2, 14, 18

Pattikkad

5, 11, 12, 22

Palapilly

7, 10

Vellikulangara

5, 12, 22

Pariyaram

14, 15, 32

Total numer of houses surveyed

vegetables (8.57 %) (Cucumis melo and Cucurbita moschata) and paddy (Oryza sativa) (11.42 %) were cultivated in the fringe areas of the forests. Other cultivated crops (7.14 %) were cocoa, pineapple (Ananas comosus), turmeric (Curcuma longa) and nutmeg (Myristica fragrans). Tubers, namely Amorphophallus paeoniifolius, Colocasia esculenta, Manihot esculenta, Dioscorea alata and Ipooea batatas, were the crops most vulnerable to foraging by wild animals (fig. 2). Arecanut palms were damaged only by elephants. Seventy–nine percent of the respondents were planning to cultivate rubber on their farms. Wild pig (57.1 %) was the main crop foraging animal among all forest ranges, whereas elephants (12.9 %) in the crop fields were seasonal in the forest ranges, namely Pattikkad, Palapilly, Vellikulangara

210

and Pariyaram (fig. 3). Other crop foraging animals were Indian crested porcupine (10.5 %), Indian giant squirrel (4.8 %), Indian giant flying squirrel (Petaurista philippensis) (4 %), bonnet macaque (Macaca radiata) (3.1 %) and Indian peafowl (Pavo cristatus) (7.6 %). Feeding on coconuts by Indian giant squirrel was reported only from the forest ranges (Pattikkad, Machad and Palapilly) adjacent to Peechi–Vazhani and Chimmony wildlife sanctuaries. In the Pariyaram forest range, common palm civet (Paradoxurus hermaphroditus) was reported consuming cocoa (n = 2). Twenty percent of the respondents reported attacks on livestock by carnivores. The predators were leopard (69.76 %), Indian rock python (Python molurus) (13.9 %), dhole (Cuon alpinus) (9.3 %) and stray dogs (6.97 %).

Households (%)

100 80 60 40 20 0

Ru Crops

Fig. 2. Crops vulnerable to damage by wild animals in Thrissur District, Kerala: Tu, tubers; Co, coconut; Pl, plantain; Ru, rubber; Pa, paddy; Ve, vegetables; Ar, arecanut. Fig. 2. Cultivos vulnerables a los daños ocasionados por animales silvestres en el distrito de Thrissur, en Kerala: Tu, tubérculos; Co, coco; Pl, plátano; Ru, caucho; Pa, arroz; Ve, vegetales; Ar, nuez de areca.

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100

Households (%)

90 80 70

Occasionally

60 50

Weekly

Daily

30 20 10 0

Asian Wild elephant pig

Indian crested porcupine

Fig. 3. Incidence of crop foraging by wild animals (n = 210). Fig. 3. Incidencia de ataques a los cultivos por diferentes animales silvestres (n = 210).

Human dimensions The respondents main occupation was agriculture (n = 182). The land tenure system showed that 94.29 % of the respondents legally owned their lands. A positive correlation was observed between the extent of agriculture land the farmers owned and the percentage of loss they reported (fig. 4). It was reported that 35.6 ± 16.99 percent of the people's annual income was lost due to crop foraging by wild animals. A negative relationship was observed between the extent of agriculture land possessed by

the farmers and distance from the crop fields to the reserve forest (fig. 5). The respondents' awareness of wildlife laws was excellent (n = 210). A significant positive correlation was observed between respondents' age and self–reported household crop loss (fig. 6). They believed that conserving wildlife is an inevitable factor for a sustainable environment (n = 167) but considered that the government should protect crops from wild animals (n = 188). Hunting was a control method suggested to prevent wild pig from entering crop fields (n = 151). Delay on sanctioning ex–gratia by the wildlife authorities angered them (n = 99),

Extent of agriculture land (ha)

3 2.5 2 1.5 1 0.5

0 0

20 40 60 80 Self–reported household crop loss (%)

100

Fig. 4. Relationship between the extent of agricultural land (ha) and self–reported household crop loss (%) (n = 210) (rs = 0.361, P < 0.05). Fig. 4. Relación entre la superficie de tierra agrícola (ha) y la pérdida de cultivos comunicada por los hogares (%) (n = 210) (rs = 0,361, P < 0,05).

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Extent of agriculture land (ha)

2.5 2 1.5 1 0.5

0 0

500

1,000 1,500 2,000 2,500 3,000 Distance to the reserve forest (m)

3,500

Fig. 5. Relationship between the extent of agricultural land (ha) and the distance to the reserve forest (m) (n = 210) (rs = –0.346, P < 0.01). Fig. 5. Relación entre la superficie de tierra agrícola (ha) y la distancia a la reserva forestal (m) (n = 210) (rs = –0,346, P < 0,01).

and ex–gratia payments did not entirely satisfy their requirements (n = 84). In the Kerala scenario, the farmers received only 75 % of the total damages incurred due to wildlife till the year 2013. Moreover, 10,000 INR (Indian Rupees) (1 US Dollar = 60 INR) was the maximum amount given to a victim at a time, even if the perennial crops were damaged. As per the stipulations of the Kerala Forest and Wildlife Department, many certificates have to be submitted in order to receive the sanctioned amount. Another

suggestion the participants made was to raise the amount of ex–gratia (n = 127). Control measures Farmers employed 17 types of control measure in the study area, namely, watch and ward, sound from metallic objects, cracker, dog, trench, cable wire, bright coloured clothes, spot–light, loud–speaker and different types of fencing (table 2). Watch and ward,

Age of the respondents

70 60 50 40 30 20 10 0

20 40 60 80 100 Self–reported household crop loss (%)

Fig. 6. Relationship between respondents' age and self–reported household crop loss (%) (n = 210) (rs = 0.38, P < 0.01). Fig. 6. Relación entre la edad de los encuestados y la pérdida de cultivos comunicada por los hogares (%) (n = 210) (rs = 0,38, P < 0,01).

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Table 2. Various control measures adopted to deter wild animals from the crop fields in the marginal areas of the study area. Tabla 2. Varias medidas de control adoptadas para impedir el acceso de los animales silvestres a las parcelas de cultivo situadas en los márgenes de la zona de estudio.

Mitigative measures

Forest range

Targeted species

Watch and ward

All forest ranges

All crop foraging species

Crackers

All forest ranges

All crop raiding species

Sound from metallic objects

Vellikulangara, Pariyaram

Wild pig and sambar

Dogs

Peechi, Wadakkancherry

Asian elephant and

Indian peafowl

Trench

Kollathirumedu, Sholayar

Asian elephant

Cable wire

Wadakkancherry, Pattikkad,

Wild pig

Peechi, Machad, Vellikulangara,

Palapilly, Pariyaram

Bright coloured clothes

Wadakkancherry, Vellikulangara

Wild pig and

Indian peafowl

Spot–light

Pattikkad, Pariyaram, Peechi,

Asian elephant

Palapilly, Charpa

Loud–speaker

Pariyaram

Wild pig, sambar and

Asian elephant

Fences Stone fence (small)

Palapilly, Pariyaram, Charpa

Wild pig and sambar

Barbed fence with

Wadakkancherry, Pattikkad

Wild pig,

concrete bar

Indian crested porcupine

and sambar

Yellow plastic

Wild pig and

Pariyaram

sheet fencing

Indian crested porcupine

Bamboo fence

Wild pig and

Wadakkancherry

Indian crested porcupine

Fish–net

All forest ranges

Wild pig and sambar

Arecanut sheath fence

Peechi

Wild pig, sambar and

Indian crested porcupine

Electric fence

Peechi, Palapilly, Pariyaram,

Asian elephant

Athirapilly, Kollathirumedu, Sholayar

cracker and fish–net fence were recorded from all forest ranges. Factors influencing crop foraging Twelve priori models were prepared to identify the significant variables, including a global model with 12 explanatory variables (table 3). The variables, namely, the extent of agriculture land possessed by farmers, the distance to reserve forest from crop field, and age of the respondent, influenced crop loss reported by the farmers (table 4).

Newspaper media reports on human–wildlife interaction Three hundred and ten newspaper reports on this topic were published from April 2009 to March 2012. They included crop foraging by wild animals (14.19 %), livestock–lifting by carnivores (10.32 %), human–casualties due to wild animals (22.26 %), sightings of wild animals in human habitation (27 %), household damage by wild animals (4.84 %), poaching (5.16 %), and public opinions on human–wildlife interaction (16.13 %). The mean number of reports published per month in the first year (April 2009 to March 2010)

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Table 3. Models included in the model sets for predicting the factors influencing crop loss reported by the respondents: land, extent of the agriculture land; agripract, agricultural practice; elev, elevation; ncrop, number of crops cultivated; distrf, distance to reserve forest; nwildsps, number of wildlife species; settle, nature of settlement; age, age of the respondent; sex, sex of the respondent; edu, educational qualification; occup, occupation of the respondent; timraidbeha, time of raiding behaviour. Tabla 3. Modelos incluidos para predecir los factores que influyen en la pérdida de cultivos comunicada por los encuestados: land, superficie de tierra agrícola; agripact, práctica agrícola; elev, elevación; ncrop, número de cultivos producidos; distrf, distancia a la reserva forestal; nwildsps, número de especies de fauna silvestre; settle, tipo de asentamiento; age, edad del encuestado; sex, sexo del encuestado; edu, calificación académica; occup, ocupación del encuestado; timraidbeha, hora de los ataques; Models

AICc

distrf + elev + ncrop + land + nwildsps + agripract + settle + age + sex + edu + occup + + timraidbeha

1,361.348

distrf + elev

1,333.729

distrf + ncrop

1,329.014

distrf + land

1,336.362

distrf + nwildsps

1,338.215

land + ncrop + nwildsps

1,326.262

land + agripract

1,348.366

elev + ncrop

1,342.758

settle + age + sex + edu + occup

1,326.331

distrf + nwildsps + timraidbeha + agripract

1,342.111

distrf + ncrop + land + agripract

1,325.749

elev + nwildsps

1,339.050

distrf + land + age

1,319.259

Table 4. Best model and coefficients of the factors influencing crop loss reported by respondents (Standard errors in brackets; wi is the AICc model weight). Tabla 4. Mejor modelo y coeficientes de los factores que influyen en la pérdida de cultivos comunicada por los encuestados (errores estándar entre paréntesis; wi es la ponderación del modelo basado en el criterio de información de Akaike, AICc).

Model Intercept

Distrf + land + age wi = 0.98 –4.399 (2.587)

Extent of agriculture land

0.96 (1.38)

Distance to reserve forest

–0.009 (0.002)

Age of the respondent

0.78 (0.18)

Model AICc

1,319.259

was 5.42 ± 2.84, in the second year (April 2010 to March 2011) it was 10.58 ± 6.04 and in the third year (April 2011 to March 2012) it was 9.58 ± 3.6 (ANOVA, F = 4.73, P < 0.05). The overall mean (reports published per month) was 8.53 ± 2.6 (n = 12). The highest numbers of reports were in May and September (fig. 7). Sixty–three percent of the published reports were on the human–wildlife interaction of the district, and of these, the highest number of reports was from Pattikkad forest range (27.09 %) (fig. 8). Asian elephant had the highest media coverage (32.58 %), followed by leopard (20.32 %) and wild pig (8.06 %). Reports included stray dog killing livestock (6.78 %), spotting of common barn–owls (Tyto alba) in houses (1.61 %), sambar (1.93 %) and chital falling into wells (2.26 %), and sightings of snakes (spectacled cobra Naja naja, Russell's vipers Daboia russelii and common krait Bungarus caeruleus) in human habitations (7.42 %). Other species included in the reports were sloth bear (Melursus ursinus) (2.26 %), tiger (Panthera tigris) (1.61 %), bonnet macaque (1.29 %), mugger crocodile (Crocodylus palustris) (1.29 %), Indian peafowl (0.97 %), smooth–coated otter (Lutrogale perspicillata) (0.97 %), dhole (0.32 %), jungle cat (Felis chaus) (0.32 %), common Indian monitor (Varanus

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12 Mean number of reports published

10 8 6 4 2 0

6 7 Months

Fig. 7. Mean number of newspaper reports published from April 2009 to March 2012: 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. 7. Promedio de artículos de prensa publicados en diferentes meses (entre abril de 2009 y marzo de 2012). (Para las abreviaturas de los meses, véase arriba).

Newspaper reports (%)

30 25 20 15 10 5 0

Ma At Ve Sh Forest ranges

Fig. 8. Newspaper reports on human–wildlife interaction in forest ranges of Thrissur District (April 2009 to March 2012) (%) (n = 196): Va, Vazhachal; Ch, Charpa; Ko, Kollathirumedu; Pe, Peechi; Ma, Machad; At, Athirapilly; Ve, Vellikulangara; Sh, Sholayar; Wa, Wadakkancherry; Pl, Palapilly; Pr, Pariyaram; Pt, Pattikkad. Fig. 8. Artículos de prensa sobre la interacción entre los humanos y la fauna silvestre en diferentes regiones forestales del distrito de Thrissur (entre abril de 2009 y marzo de 2012) (%) (n = 196). (Para las abreviaturas de las regiones forestales, véase arriba).

bengalensis) (0.32 %), Indian crested porcupine (0.32 %), purple swamphen (Porphyrio porphyrio) (0.32 %), gaur (Bos gaurus) (0.32 %) and brown fish owl (Bubo zeylonensis) (0.32 %). Reporting of unidentified wild animals killing livestock (3.26 %) created much anxiety among the people. Stray dogs killed livestock in the night hours,

and the loss was blamed on 'unknown animal'. We visited Kalady (Malayattoor forest division, Ernakulam district) along with the forest officials on 14th August 2009 to identify the 'unknown animal' reported in the newspapers. A discussion was carried out among the local people (n = 20), and a plaster cast of pugmarks was collected from the area of livestock–lifting. Local

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people had only a vague knowledge of the species involved in the attack. Two persons reported that it was a leopard that attacked, and the presence of a 'leopard–like animal' was identified in the dark hours by six persons. The Kerala Forest and Wildlife Department started operating a baited trap immediately after the attack. Plaster casts revealed that pugmarks of the stray dog were similar to leopard, but without any claws. Absence of claws was also observed in the captured stray dog from the site. The proliferation of stray dogs in the locality was triggered by the availability of carrion. Discussion Crop foraging and livestock–lifting Negative interaction between humans and wildlife is an emerging issue in Kerala. In the study area, eight species of wild animals consumed 17 species of crops. The mode of feeding on these crops was previously reported by Jayson (1999). Due to the high price of rubber during the study period, cultivation of traditional crops (e.g. coconut, arecanut and plantain) in the fringe areas was replaced by the cultivation of rubber. The price of rubber in Kerala during 2008 was 100.54 ± 30.05/– INR per kg, whereas, in 2011, it increased to 207.96 ± 17.77/– INR (Source–Rubber Board, Kerala, India). Marginal farmers were cultivating plantain in between the young rubber plants (Govind, 2015). While eating plantains, wild animals also destroyed young unpalatable rubber plants. Asian elephant mainly consumed plantain and perennial crops, namely coconut palm, arecanut palm and rubber tree, from August to November. A similar trend has been reported from the other States of India (Gubbi, 2012; Lingaraju and Venkataramana, 2016). Wild pig consumed tubers in all seasons. The same was reported in Kerala (Jayson, 1999), North India (Chauhan et al., 2009) and Central India (Karanth et al., 2012). Feeding of fallen coconuts by wild pig and Indian crested porcupine was recorded. As the price of the coconut was low, farmers did not protect the fallen coconut from these animals. However, it was noticed that they protected the nut when the price increased (Govind and Jayson, 2018b). Protecting tender coconuts from Indian giant squirrel and Indian giant flying squirrel was also observed when their price was high (Govind and Jayson, 2018a). People deterred these squirrels from their farms by pelting stones and making sounds. Indian peafowls consumed paddy near Chulannur Peafowl Sanctuary, and the mode of consumption was by stripping off the grain from the panicle with their beaks (Govind and Jayson, 2018c). As the peafowl is considered sacred in Hindu mythology, poaching of this species is not reported. In the buffer areas of Kitam Bird Sanctuary (Pradhan et al., 2012) and Kanha National Park (Karanth et al., 2012), crop foraging by peafowl was reported. Carnivores attacking livestock were recorded in the area, with leopard being the main predator, followed by Indian rock python. Goat and poultry were the main prey of Indian rock python.

Previous studies recorded several human casualties due to elephant, leopard, sloth bear, tiger and gaur. In contrast, the carnivores that killed livestock were tiger, leopard and dhole (Jayson, 1999; Christopher, 1998). Tiger and leopard were the main predators of livestock in Central India (Karanth et al., 2012) and South India (Karanth et al., 2013). Livestock–lifting by dhole was reported from some areas of India (Karanth et al., 2013; Roshnath et al., 2017). Human dimensions Due to the stringent provisions of the Wildlife Protection Act of India, we found that people's response towards wildlife was generally good, and it was confirmed that the lack of public awareness was not a cause for increasing conflict. When we approached farmers with no legal documents for their lands, they reacted negatively towards us. These land–owners are not eligible for ex–gratia claims for crop foraging by wild animals. Furthermore, many farmers did not claim ex–gratia as they did not know the actual economic loss incurred due to wild animals. During the survey, they also suggested immediate sanctioning of ex–gratia. Studies indicate that speedy disbursement of ex–gratia is the significant factor in increasing co–existence between humans and wildlife (Madden, 2004; DeFries et al., 2010). Hunting of wild pig was encouraged in many areas to manage populations (Beskardes et al., 2010). In Kerala, hunting is not encouraged, even though an increase in their population is recorded (Wildlife Census Reports, Kerala Forest and Wildlife Department). As per the popular demand, the shooting of problematic wild pigs in crop fields was allowed by the Kerala Forest and Wildlife Department in 2012. Due to the stringent procedures before shooting, the farmers could not employ this method to reduce the population of wild pig so as to prevent crop foraging. It adversely affected the relationship between local people and wildlife officials. Nemtzov (2003) stated that people will turn against wildlife if the conflict is high or intolerable to humans. Due to fear and anger, many leopards were killed by local people in some regions of India (Karanth and Madhusudan, 2002). Poaching of Asian elephant and wild pig was reported when the study was being carried out in the area (Govind, 2015). Due to the awareness programs conducted, local people tolerated the intensive crop foraging species, namely nilgai (Boselaphus tragocamelus) and blackbuck (Antilope cervicapra) in north India (Sekhar, 1998; Karanth and Madhusudan, 2002). In our study area also, feeding on coconuts by Indian giant squirrel is restricted only to crop fields adjacent to the wildlife sanctuaries. Due to the stringency of wildlife laws around these protected areas, people are not taking any negative precautions towards the squirrels. Control measures Yellow plastic sheet fencing was an innovative control method used against crop foraging animals such as wild pig and Indian crested porcupine. It has been

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used to protect the newly planted rubber in the fringe areas of the forest. It is a less expensive control measure than the solar–electric fence. White plastic sheet fencing has also been reported to dissuade wild pigs from entering crop fields (Gopakumar et al., 2012). To control the damage to rubber plants, solar–electric fencing has been installed by large–scale farmers. This control method is very effective if it is properly erected and maintained (Conover, 2002; Veeramani et al., 2004). Other control measures recorded in the study have previously been reported by Veeramani et al. (2004). Lethal control measures, namely shooting, poisoning and trapping, are widely adopted to control the wildlife population, and to mitigate human–wildlife interaction (Treves and Naughton–Treves, 2005). However, these control measures may also adversely affect untargeted species (Nemtzov, 2003). The ‘hunting for fear' method is another mitigative tool practised in many countries to induce a behavioural change to crop foraging animals (Cromsigt et al., 2013). Factors influencing crop foraging Out of three variables that were found to influence the self–reported household crop loss, two variables, namely the extent of agriculture land possessed by the farmers and the distance to reserve forest from crop field were also reported to influence the buffer areas of Kanha National Park (Karanth et al., 2012). In the study area, farmers possessed large areas of land in the fringe areas of the forest, and a huge crop loss due to wild animals was reported. Sillero–Zubiri et al. (2007) reported that the conservation attitude of older generations towards wildlife was more positive than that of younger generations. Our study contradicts this hypothesis, as the perception of local people about the crop loss was slightly higher among older people and their attitude towards wildlife conservation was neutral. Kellert (1980) stated that the attitude of rural residents was more moralistic towards wildlife than urban workers. Another study revealed that the conservation attitude of people who migrated from urban areas was irreversibly negative towards wildlife (Loker et al., 1999). It was observed that 17 % of the respondents in the study area had migrated from urban areas. Newspaper media reports on human–wildlife interaction Leopards often entered human habitations and local people immediately informed such events to the newspapers. In May, several news articles on leopard sightings were episodically framed and reported. Most of these reports were from Palapilly and Sholayar forest ranges, and the sightings were recorded at the end of the summer (March to May) in Kerala. Damage to plantain by elephants was high during September in the Pattikkad forest range. This is because farmers were planning to harvest plantain in the immediate fringe areas of the forest during this month for the Onam festival in Kerala (Govind, 2015). Studies indicate that the impact generated when an elephant or a leopard or leopard enter human habitation is

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very high compared to the impact caused by other wild animals (Jhala and Sharma, 1997; Treves and Naughton–Treves, 2005). As wildlife conservation received much media coverage in the newspapers with good photographs, awareness of wildlife laws was excellent among the people, and their attitude towards the wildlife species was positive. Local people directly informed the highest forest officials about the intrusion of any wildlife species into human habitations. Local people became anxious when the media reported the predation of an 'unknown animal' on livestock. When the media amplifies uncommon events or attacks it creates a strong response from public (Crossley et al., 2014) and increases public anxiety (Sabatier and Huveneers, 2018). Stafford et al. (2018) examined the newspaper reports on human–wildlife interaction and reported that wild animals were more commonly blamed for the conflict than humans. Corbett (1992) stated that reports from the media mainly depend on bureaucratic sources, and in certain circumstances, human–human conflicts also exacerbate human–wildlife interaction (Dickman, 2010). Conclusion Eight species of wild animals are foraging crops in the study area, with the wild pig being the species most responsible for this activity. Thirty–six per cent of the annual income of farmers is lost due to crop foraging by wild animals. To deter such actions, 17 ypes of control measures are being used. Fencing made from yellow plastic sheeting are an innovative method for dissuading wild pig and Indian crested porcupine from crop fields. The predators involved in the attacks on livestock are leopard, Indian rock python, dhole and stray dog. The variables, namely, extent of agriculture land possessed by the marginal farmers, distance to reserve forest from crop field, and age of respondents influence the crop loss reported by the marginal farmers. Due to people´s growing awareness of the importance of wildlife, newspaper media reports on human–wildlife interaction, and strict enforcement of wildlife laws by the authorities, the conservation attitude of people towards wildlife is good, and they are not taking any negative actions against wild animals. Avoiding the cultivation of tubers and plantains in the immediate fringe areas of the forest is recommended. Acknowledgements We are thankful to Dr. K. V. Sankaran (Former Director), Kerala Forest Research Institute, Peechi, Kerala, India, and the forest officials, for supporting the study. References Beskardes, V., Yilmaz, E., Oymen, T., 2010. Evaluation on management of wild boar (Sus scrofa L.) population in Bolu–Sazakici hunting ground. Journal of Environmental Biology, 31: 207–212.

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Revisión del conflicto entre los seres humanos y las serpientes en México: origen, mitigación y perspectivas L. Fernández–Badillo, I. Zuria, J. Sigala–Rodríguez, G. Sánchez–Rojas, G. Castañeda–Gaytán

Fernández–Badillo, L., Zuria, I., Sigala–Rodríguez, J., Sánchez–Rojas, G., Castañeda–Gaytán, G., 2021. Revisión del conflicto entre los humanos y las serpientes en México: origen, mitigación y perspectivas. Animal Biodiversity and Conservation, 44.2: 153–174, Doi: https://doi.org/10.32800/abc.2021.44.0153 Abstract Review of the human–snake conflict in Mexico: origin, mitigation and perspectives. The conflict between humans and snakes has existed since unmemorable times. Fear of and aversion towards these animals may have an evolutionary explanation and may be justified because venomous and deadly snakes cause thousands of deaths around the world each year. Furthermore, social perception, the media, myths, and even religion, increase and feed this fear, resulting in the intentional slaughter of snakes being a common practice in many places. As Mexico is a mega–diverse country with more species of snakes than any other country, it faces a particularly difficult situation with regard to snake bites. Here we revise this human–snake conflict from different perspectives in order to better understand it, to propose possible solutions to reduce it, and to contribute towards snake conservation. Key words: Fear, Mexico, Megadiverse, Vipers, Strategies, Mitigation, Conservation Resumen Revisión del conflicto entre los seres humanos y las serpientes en México: origen, mitigación y perspectivas. El conflicto entre los humanos y las serpientes ha existido desde el principio de la humanidad y, aunque el temor y la aversión hacia estos animales podrían tener una explicación evolutiva y pueden estar justificados por la presencia de especies venenosas y mortales para el hombre, que causan miles de muertes al año, la percepción social, los medios de comunicación, los mitos e incluso la religión acrecientan y alimentan el temor hacia estos animales, por lo que en muchos lugares el sacrificio intencionado de serpientes es una práctica común. Esta situación también ocurre en México, un país megadiverso que acoge a la mayor diversidad de serpientes de todo el mundo y en el que también existe una problemática seria a causa de las mordeduras de serpientes. Aquí presentamos una revisión de este conflicto desde distintas perspectivas, para tratar de entenderlo, proponer posibles soluciones para mitigarlo y contribuir a la conservación de las serpientes. Palabras clave: Temor, México, Megadiverso, Vipéridos, Estrategias, Mitigación, Conservación Received: 28 IX 20; Conditional acceptance: 10 III 21; Final acceptance: 24 III 21 Leonardo Fernández–Badillo, Iriana Zuria, Gerardo Sánchez–Rojas, Centro de Investigaciones Biológicas, Universidad Autónoma del Estado de Hidalgo, km. 4,5 carretera Pachuca–Tulancingo, Mineral de la Reforma, Hidalgo, México.– Leonardo Fernández Badillo, Predio Intensivo de Manejo de Vida Silvestre X–Plora Reptilia. Carretera México–Tampico s/n., Pilas y Granadas, 43350 Metztitlán, Hidalgo, México.– Jesús Sigala–Rodríguez, Colección Zoológica de la Universidad Autónoma de Aguascalientes, C. P. 20131, Aguascalientes, Aguascalientes, México.– Gamaliel Castañeda Gaytán, Facultad de Ciencias Biológicas, Universidad Juárez del Estado de Durango, Av. Universidad s/n., Fracc. Filadelfia, C. P. 35020, Gómez Palacio, Durango, México. Corresponding author: Iriana Zuria. E–mail: izuria@uaeh.edu.mx

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Introducción El conflicto entre los seres humanos y los animales silvestres se define como cualquier acción de una de las partes que tiene un efecto adverso en la otra (Conover, 2001). Los efectos de este conflicto pueden ser lesiones o incluso la muerte de las personas a causa de ataques (Karanth et al. 2018), accidentes con vehículos (Saint–Andrieux et al., 2020) y transmisión de enfermedades zoonóticas (Nyhus, 2016). Sin embargo, los mayores efectos negativos que han podido cuantificarse económicamente son los daños materiales ocasionados a los cultivos, la ganadería y las infraestructuras (Conover et al., 2018). A consecuencia de este conflicto, se extermina a una gran cantidad de animales en todo el mundo, lo cual, en ocasiones, conlleva la extinción local de las especies que forman parte del conflicto, incluso de animales bien reconocidos y simbólicos, como los grandes depredadores (Hammerschlag y Gallagher, 2017). A las especies que se encuentran en zonas donde los seres humanos las consideran potencialmente peligrosas o destructivas y, por tanto, no las quieren, se las denomina especies nocivas (Nowak et al., 2002). Las personas suelen tener conflictos con las serpientes debido a que algunas de ellas representan un riesgo para la salud humana, lo que conlleva que frecuentemente se mate a estos reptiles (Pandey et al., 2016), incluso de especies inofensivas que se confunden con especies venenosas (Ávila–Villegas, 2017). En algunos estudios empíricos se sugiere que la evolución dotó a nuestros antepasados con una disposición para asociar fácilmente el miedo con amenazas recurrentes, de manera que nuestro sistema visual está predispuesto para detectar rápidamente animales peligrosos (Kawai y Koda, 2016). Esto puede explicar por qué en general los humanos temen a las serpientes; sin embargo, existen otros factores que pueden ser causantes del miedo a estos animales y que se analizan más adelante. Por otro lado, en culturas antiguas, las serpientes eran veneradas (Fernández–Rubio, 2017) e incluso siguen siendo deidades importantes para muchas culturas que las han asociado más con símbolos positivos (agua, conocimiento, fertilidad, eternidad, salud, etc.) que con símbolos negativos (Ballouard et al., 2013). Incluso actualmente, en ciertas partes del mundo como la India, el culto al Dios serpiente aún se practica, y estos animales son respetados en los templos, hasta el punto de que, en algunas regiones, se considera un pecado matar o herir a una serpiente (Yuan et al., 2020). A pesar de lo anterior, considerar a las serpientes como animales nocivos y tratar de erradicarlas parece ser un fenómeno mundial que contribuye a la disminución de sus poblaciones (Pandey et al., 2016). Por ello, resulta importante analizar las causas, las consecuencias y las posibles soluciones de este conflicto a escala local o regional, para promover acciones que permitan la conservación de estos vertebrados. En el presente trabajo se analiza el conflicto entre los seres humanos y las serpientes en México desde distintas perspectivas, partiendo desde el origen del temor hacia las serpientes, así como de los factores

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que lo acrecientan. Considerando la información y las estrategias disponibles para otros países, se sugieren soluciones aplicables a la realidad mexicana que permiten mitigar el conflicto y promover la conservación de las serpientes. Panorama de la riqueza de serpientes en México En México viven aproximadamente 438 especies de serpientes, lo que sitúa este país el primero en cuanto a diversidad de serpientes en el mundo (Midtgaard, 2021). Además, cerca de la mitad de sus especies son endémicas del país (Heimes, 2016). Estas especies se encuentran agrupadas en once familias, según la taxonomía propuesta por Zaher et al. (2019): Boidae, Charinidae, Colubridae, Dipsadidae, Elapidae, Leptotyphlopidae, Loxocemidae, Natricidae, Sybinophidae, Typhlopidae y Viperidae (Heimes, 2016); solo dos de ellas (Elapidae y Viperidae) poseen venenos potencialmente peligrosos para el hombre (Fernández–Badillo et al., 2011). Estas serpientes venenosas están representadas en México por 90 especies (17 elápidos y 73 vipéridos; Campbell y Lamar, 2004; Meik et al., 2015; Davis et al., 2016; Heimes, 2016; Blair et al., 2018; Carbajal– Márquez et al., 2020; Reyes–Velasco et al., 2020), lo que posiciona a México como el país con mayor diversidad de ofidios venenosos de América (Campbell y Lamar, 2004) y el más diverso del mundo en lo que a especies de vipéridos se refiere. Considerando esta gran riqueza de serpientes, México debe asumir el reto de lograr la conservación de estas especies a largo plazo, buscando estrategias que permitan disminuir las amenazas que afectan a este grupo en particular (Paredes–García et al., 2011). Además, es importante conservar estos reptiles debido a su importancia ecológica y a los descubrimientos que se han hecho relacionados con la salud humana (Olson et al., 2015). Existen algunas iniciativas para su conservación como el Programa para la Conservación de Especies del género Crotalus, en el que se incluyen 43 especies (SEMARNAT, 2018); sin embargo, dada la enorme riqueza de serpientes de México y su problemática de conservación, hacen falta más esfuerzos que garanticen la conservación de este diverso grupo en el país. Origen del conflicto entre los seres humanos y las serpientes El conflicto entre los seres humanos y las serpientes siempre ha existido y el miedo a estos animales se explica como un rasgo evolutivo heredado de los ancestros primates, los cuales eran depredados por grandes serpientes constrictoras (Burghardt et al., 2009; Isbell, 2009; Soares et al., 2014; Hoehl et al., 2017). Incluso se ha formulado la hipótesis de que las serpientes fueron en última instancia las responsables de la evolución del sistema visual y de la expansión del cerebro de los primeros primates, y se plantea

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que el sistema visual y el cerebro de los antropoides se modificaron aún más con la aparición de las serpientes venenosas (Isbell, 2005); sin embargo, Tierney y Connolly (2013) sugieren que no existe información suficiente para concluir lo anterior. Otros autores indican que el miedo puede ser adquirido por aspectos culturales (como los mitos y leyendas) y religiosos, por la transmisión intergeneracional de conocimientos erróneos acerca de las serpientes (Davey, 1995; Aguilar–López, 2016) o por el peligro real que representan las serpientes venenosas (Ballouard et al., 2013). Este miedo a las serpientes podría ser la fobia animal más generalizada entre los seres humanos (Anderson et al., 2013). La consecuencia más evidente de este miedo es que las serpientes son percibidas por el ser humano como animales peligrosos a los que hay que evitar o erradicar (Pandey et al., 2016; Pitts et al., 2017), por lo que es común que se las sacrifique indiscriminadamente (Whitaker y Shine, 2000; Ávila–Villegas, 2017). Por esta razón, es importante analizar con detalle los factores que influyen en el miedo hacia las serpientes y, a partir de ello, formular estrategias que permitan cambiar la percepción de las personas, a fin de minimizar el perjuicio a las poblaciones de estos animales y, consecuentemente, evitar un mayor impacto ecológico en el ecosistema. Factores que influyen en el temor hacia las serpientes Dejando de lado el posible miedo innato o instintivo hacia las serpientes, hay tres factores que propician o incrementan el miedo hacia estos animales: 1, los mitos y la religión; 2, la percepción moderna y la mala publicidad en los medios de comunicación, y 3, las mordeduras de serpientes venenosas. Todos ellos se analizan a continuación: Los mitos y la religión Las serpientes son probablemente los animales de los que más mitos y leyendas existen y resulta muy difícil desarraigar prejuicios que tienen antecedentes muy antiguos (Casas–Andreu, 2000; Ermacora, 2017; Paulino, 2018). Muchos de estos mitos y leyendas surgen de la exageración, otros a partir de observaciones o interpretaciones equivocadas (Klauber, 1982) y otros son producto de ciertos prejuicios y falta de información, ya que por lo general no se conocen aspectos básicos de la biología, la conducta y el papel que desempeñan las serpientes en los ecosistemas, principalmente debido a sus hábitos nocturnos y por ser poco detectables en algunos hábitats (Aguilar–López, 2016). Otro aspecto cultural que afecta negativamente a la percepción que se tiene de las serpientes es la interpretación literal del Génesis, donde se identifica a las serpientes como la representación del mal y se las presenta como la personificación del Diablo (Aguilar–López, 2016). En México, las serpientes tuvieron un papel muy importante en la mitología y la cosmovisión de los

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pueblos prehispánicos, y son de los animales más recurrentes en los códices (Martín del Campo, 1936). Casi todos los monumentos mayas, toltecas y mexicas tienen representaciones de serpientes (Cuesta Terron, 1931) y algunas de las deidades más importantes de estas culturas, como Quetzalcoatl o Kukulkan, fueron representadas por serpientes (Cuesta Terron, 1931; Martín del Campo, 1936). Además, desde la época prehispánica ya existían historias y mitos respecto a las serpientes, algunos de los cuales se encuentran descritos en códices (Martín del Campo, 1936, 1938) y permanecen hasta nuestros días. Por ejemplo, el mito de que las serpientes andan en pareja y que, si se mata a una de ellas, el ejemplar que queda persigue a la persona hasta vengarse, es un relato que ya existía entre los pueblos prehispánicos (Martín del Campo, 1936) y persiste aún en la actualidad en algunas localidades del estado de Hidalgo, México (obs. pers). Independientemente de si los mitos presentes en México tienen un origen muy antiguo o reciente, generalmente están arraigados en el pensamiento y la cultura de los pobladores. La percepción moderna y la mala publicidad en los medios de comunicación En la actualidad, los medios de comunicación de la sociedad occidental están más centrados en el entretenimiento que en la educación y han generado y reforzado diversas creencias negativas e irracionales sobre las serpientes (Ballouard et al., 2013). Por ejemplo, existe un subgénero cinematográfico de terror llamado “ecoterror”, en el que los personajes humanos son atacados por las fuerzas de la naturaleza, principalmente animales o plantas (Rust y Soles, 2014). Se han filmado diversas películas de este género sobre serpientes asesinas, en las que se da una visión pésima de las serpientes (Aguilar–López, 2016). Algunos títulos famosos son "Anaconda I y II", "Serpientes asesinas", "Serpientes abordo", "Serpientes en el tren", "Veneno", "Sssssss", "Pirañaconda", "Python" y "Boa", entre muchas otras. No hay duda de que estos trabajos cinematográficos de amplia distribución pueden fomentar el odio y la aversión hacia las serpientes en algunos sectores de la sociedad. En este sentido, Morris (2017) menciona que el mayor obstáculo para enseñar los comportamientos precisos de las serpientes, por ejemplo, los de las serpientes de cascabel, es luchar contra el flujo continuo de conceptos equivocados que transmiten los medios de comunicación. Incluso en algunos documentales sobre fauna silvestre se presenta a las serpientes como animales peligrosos y agresivos (Ballouard et al., 2013). En consecuencia, el miedo hacia los ofidios no siempre está relacionado con una percepción de peligro (Ballouard et al., 2013), ya que puede surgir independientemente del riesgo, y ese temor fuerte y persistente puede ser irracional o estar influenciado por los medios de comunicación y creencias religiosas arraigadas (Öhman y Mineka, 2003). Otro aspecto que ha infundido mucho temor hacia las serpientes es la existencia de especies de tallas muy grandes como las pitones y las anacondas, lo

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cual se ha reforzado con las películas en las que estas serpientes devoran personas y las noticias sensacionalistas de la prensa. Si bien hace 65 millones de años las serpientes constrictoras fueron de los primeros depredadores de los antiguos primates (Isbell, 2005), en la actualidad, la mayoría de las especies de serpientes de tallas muy grandes no percibe a los humanos como presas potenciales, y existen muy pocos casos documentados de serpientes que hayan devorado personas (Branch y Hacke, 1980). Se sabe que las anacondas en medios antropomorfizados se alimentan principalmente de animales domésticos (perros, vacas, gallinas y gatos; Miranda et al., 2016) y que, aunque los ataques a personas son sumamente infrecuentes, la idea de que estas serpientes representan un riesgo para los humanos se encuentra muy arraigada en los habitantes de las comunidades rurales y con poco acceso a información. En la mayoría de los estados de México, los pobladores aseguran que existen serpientes de 10 metros o más de longitud e indican que estas serpientes de gran tamaño no mueren al ser atropelladas y que incluso pueden volcar vehículos (Fernández–Badillo, datos no publicados). Estas afirmaciones son irreales y exageradas, debido a que en México las serpientes de mayor longitud son las boas (género Boa), que no superan los 3,20 metros (Heimes, 2016). Asimismo, en México, es común que en la prensa y en la televisión se presente una imagen negativa de las serpientes y se las identifique como criaturas peligrosas. Mordeduras de serpientes venenosas En el mundo existen 3.956 especies de serpientes y, aunque solo el 19,3 % (767 especies; Midtgaard, 2021) tienen un veneno potencialmente peligroso para el hombre, el temor a las serpientes resulta justificable si se consideran los efectos de la mordedura de una serpiente venenosa. Las mordeduras son una amenaza real para la salud humana y están catalogadas como una emergencia médica (Gil–Alarcón et al., 2011); además se las considera una enfermedad tropical desatendida (ETD: Chipaux, 2017; Gutiérrez et al., 2017; OMS, 2019), que resulta de la inyección de un veneno altamente especializado, usualmente en circunstancias accidentales (Gutiérrez et al., 2017). En este sentido, se estima que ocurren entre 1,8 y 2,7 millones de casos de envenenamiento y entre 81.000 y 130.000 muertes al año en todo el mundo (Gutiérrez et al., 2017; OMS, 2019). La mordedura de serpiente es un problema de salud pública importante en muchos países y una causa importante de morbilidad y mortalidad en las zonas empobrecidas tropicales y subtropicales del África subsahariana, el este y sureste de Asia, Papúa Nueva Guinea y Latinoamérica (Gutiérrez et al., 2017). A escala global, hay una fuerte asociación entre un estatus socioeconómico bajo y una alta tasa de mordeduras (OMS, 2019), además, es más común encontrar serpientes en las zonas rurales y las áreas de cultivo que en las grandes ciudades. Sin embargo, algunas serpientes suelen habitar cerca de zonas residenciales urbanas o suburbanas, como

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ocurre con especies del género Crotalus (Pitts et al., 2017), varias especies de elápidos australianos (Shine y Koening, 2001) y algunas especies del género Bothrops (Cândido de França et al., 2017). En México, la información disponible acerca de las mordeduras de serpiente en la década de los noventa incluía únicamente datos del Instituto Mexicano del Seguro Social, así como del Instituto Nacional de Estadística y Geografía. Entre 1994 y 1998 se registraron 2.620 mordeduras (Tay–Zavala et al., 2002; Luna– Bauza et al., 2004; González–Rivera et al., 2009), lo que seguramente es un número inferior al real, ya que muchas mordeduras quedan sin registrar o atender. A partir de 2003, se comenzaron a notificar y reportar a escala nacional todos los datos referentes a las mordeduras de serpientes en México (González–Rivera et al., 2009; Siria–Hernández y Arellano–Bravo, 2009; Zúñiga–Carrasco y Caro–Lozano, 2013; Neri–Castro et al., 2020a) y, aunque estos trabajos muestran que el promedio anual de mordeduras cambia en función del periodo de estudio, este se encuentra en un intervalo de entre 3.500 y 4.111 casos anuales. Los estados en los que se ha producido un mayor número de mordeduras son Oaxaca, Veracruz, San Luis Potosí, Puebla e Hidalgo (González–Rivera et al., 2009; Siria–Hernández y Arellano–Bravo, 2009; Zúñiga– Carrasco y Caro–Lozano, 2013; Chippaux, 2017; Neri– Castro et al., 2020a) y, en algunos periodos, Chiapas y Tabasco (Zúñiga–Carrasco y Caro–Lozano, 2013). La mayoría de los accidentes (66,3 %), así como de los fallecimientos (66,7 %), ocurre en personas del sexo masculino (Tay–Zavala et al., 2002; Zúñiga–Carrasco y Caro–Lozano, 2013; Chippaux, 2017; Neri–Castro et al., 2020a); en el periodo de 2003 a 2010, la mayor mortalidad ocurrió en personas de entre 24 y 44 años de edad y, en segundo lugar, en los mayores de 65 años, siendo más marcada en este último por la presencia de enfermedades de base y complicaciones con el ofidismo (Zúñiga–Carrasco y Caro–Lozano, 2013). Asimismo, la mayoría de las mordeduras se produce en los pies (Tay–Zavala et al., 2002; Zúñiga–Carrasco y Castro–Bravo, 2013), al menos en los accidentes que ocurren en el campo; mientras que en las ciudades, las mordeduras generalmente se producen en las manos (Gil–Alarcón, obs. pers.). Por otro lado, la mortalidad y la incidencia son menores en el norte de México y va aumentando hacia el sur del territorio, con un promedio de incidencia de dos personas por cada 100.000 en el norte, de siete por cada 100.000 en el centro y de nueve por cada 100.000 en el sur del país (Chippaux, 2017). La mortalidad por mordedura de serpiente en México ha ido disminuyendo considerablemente con el paso de los años. Frayre–Torres et al. (2006) registraron 2.728 muertes entre 1979 y 2003 y una disminución del 76 % en la tasa de mortalidad por cada 100.000 habitantes, que fue de 0,25 en la década de 1970 a 0,05 durante los años 2000; a partir del 2010, la tasa de mortalidad en México fue inferior al 0,04 (Chippaux, 2017). Por otro lado, Neri–Castro et al. (2020a) indican que en la década de 1990, el promedio de defunciones por año fue de 110,8 ± 20, mientras que en el periodo de 2010–2017, fue de 34 ± 6,6. Esto implica

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una disminución en el promedio de defunciones del 69 %, lo cual, de acuerdo con et al. (2020a), puede deberse a mejoras en la calidad y disponibilidad de los tratamientos, así como a la generalización del uso de los antídotos faboterápicos a partir de 1997. A escala estatal se han publicado pocos estudios para conocer la problemática que representan las mordeduras de serpiente y solo se cuenta con información relativa a los estados de Chiapas (Suárez–Velázquez y Luna–Reyes, 2009) y Veracruz (Yañez–Arenas, 2014; Yañez–Arenas et al., 2014; 2016). Estos estudios permiten conocer la situación general de los accidentes ofídicos en esos estados y ayudan a identificar las áreas de mayor incidencia y riesgo de mordeduras provocadas por las especies de serpientes venenosas que albergan. Este es el conocimiento que ha fundamentado la implementación de estrategias que permitan mejorar la distribución de los antídotos y reforzar la capacitación del personal médico, así como para conocer en qué zonas es más urgente realizar campañas de prevención y primeros auxilios en caso de accidente ofídico destinadas a los pobladores. No hay datos a escala nacional sobre las especies que protagonizan accidentes ofídicos y, generalmente, se desconoce la especie de serpiente que ocasiona el accidente (Neri–Castro et al., 2020a). Sin embargo, según la experiencia de estos autores, las especies que posiblemente causen más mordeduras en México son: Botrhops asper, Crotalus atrox, C. basiliscus, C. culminatus, C. tzabcan, C. mictlantecutli, C. molossus y sus subespecies, C. simus y Agkistrodon bilineatus. No obstante, es posible que otras especies de serpientes de cascabel estén implicadas en los accidentes, por ejemplo C. scutulatus, que presenta una amplia distribución en México (Heimes, 2016), o especies con distribuciones más restringidas en el centro del país, pero que son muy abundantes, incluso en zonas urbanas y zonas de cultivo, como C. aquilus, C. polystictus, C. ravus o C. triseriatus (obs. pers.). Acciones o conductas de los humanos que afectan a las serpientes Las serpientes se ven amenazadas principalmente por cinco factores antrópicos, cuyo orden de importancia es el siguiente: 1) la mortalidad asociada a la pérdida de hábitat, 2) la mortalidad causada por trabajadores del campo, 3) la mortalidad asociada al tráfico en las carreteras, 4) el tráfico ilegal de especies silvestres y 5) la captura con fines científicos (Lynch, 2012; Lynch y Angarita–Sierra, 2016). A partir de la propuesta de Lynch (2012), en la que se considera una tasa de pérdida de serpientes de 375 individuos/ha debido a la tala de coberturas boscosas, Lynch y Angarita–Sierra (2016) calculan que, en Colombia, mueren entre 500.000 y 50.000.000 ejemplares de serpiente a causa de la pérdida del hábitat. Lamentablemente, no existen datos sobre la mortalidad de serpientes asociada a la pérdida de hábitat en México. En cuanto a la mortalidad causada por trabajadores del campo, Lynch (2012) estimó que, en Colombia, los pobladores llegan a sacrificar hasta 8.000.000 ser-

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pientes cada año por aversión y miedo. Este tipo de conducta se denomina miedo activo y conlleva una acción directa en contra de estos animales con el fin de exterminarlos (Aguilar–López, 2016). El miedo activo representa entonces una amenaza sumamente importante y no tiene que ver con el peligro real que representan las serpientes (como en el caso de las especies venenosas), sino con la percepción que las personas tienen de las serpientes. En este sentido, en México existe muy poca información al respecto y se desconoce el número de serpientes que se pudieran estar sacrificando anualmente a causa del temor o la aversión, pero el sacrificio intencional de serpientes es una práctica común en cualquier comunidad rural o indígena de México. En las encuestas realizadas en los estados del norte del país (Baja California, Baja California sur, Chihuahua, Sonora y Sinaloa) se observó que el 76 % de los pobladores encuestados elimina a las serpientes por el temor que les tienen (Ávila–Villegas, 2017). Las serpientes son uno de los grupos más vulnerables a los efectos directos e indirectos de las carreteras (Andrews et al., 2008) y existen muchos estudios que evalúan la mortalidad de serpientes por atropello vehicular. Las serpientes son vulnerables a los atropellos porque en muchas ocasiones utilizan las carreteras para regular su temperatura corporal (Rincón–Aranguri et al., 2019) y también suelen utilizar microhábitats ubicados en los márgenes de las carreteras (Caletrio et al., 1996). La alta densidad y frecuencia del tráfico puede coincidir con movimientos estacionales específicos (hibernación, forrajeo, migración, búsqueda de sitios de reproducción, etc.), lo cual incrementa la susceptibilidad de mortalidad directa (Andrews y Jochimsen, 2006). Además, son animales de movimientos relativamente lentos, (Rosen y Lowe, 1994), y se ha documentado que son objeto de atropellos intencionales (Andrews y Gibbons, 2005). Existen tan solo seis estudios publicados sobre la mortalidad de serpientes en carreteras de México, los cuales se realizaron en Nuevo León (Lazcano–Villarreal et al., 2009, 2017), el istmo de Tehuantepec en Oaxaca (Grosselet et al., 2009), parte de la costa de Michoacán (Martínez–Hernández, 2011), el sur de Quintana Roo (Köhler et al., 2016a, 2026b) y un tramo de carretera en el centro de Veracruz (Cervantes– Huerta et al., 2017). La tasa de mortalidad más alta (0,57 individuos por kilómetro por día) se presentó en el estudio realizado en el istmo de Tehuatepec (Grosselet et al., 2009). Es evidente la necesidad de afrontar el rezago existente en México en referencia al conocimiento de la magnitud y las características de los efectos que la construcción de carreteras tienen en la biodiversidad en general y en la fauna silvestre en particular (Puc–Sánchez et al., 2013). El tráfico ilegal de especies es una industria potente a escala mundial, comparable únicamente con el tráfico de drogas y armas y cuyas transacciones se estima se sitúan entre 7.000 millones y 23.000 millones de dólares estadounidenses anuales (Sosa–Escalante, 2011; Masés–García et al., 2021). Asimismo, existen estudios para analizar las consecuencias de estas prácticas en las poblaciones de anfibios y reptiles

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(Schlaepfer et al., 2005). Algunos de ellos mencionan específicamente el tráfico de serpientes y pieles de reptiles (Dodd, 1986; Jenkins y Broad, 1994; Fitzgerald y Painter, 2000; Zhou y Jiang, 2004). Aunque se ha avanzado en la aplicación de la Ley General de Vida Silvestre de México, no es posible afirmar que se haya terminado el tráfico ilegal (Sosa– Escalante, 2011), y existen pocos estudios que evalúen esta práctica. Se sabe que el tráfico ilegal se debe a diversos factores y que está relacionado con el uso tradicional de las serpientes como alimento, producto medicinal o incluso con fines mágicos y religiosos, razón por la cual anteriormente era común observar la venta de grandes cantidades de serpientes en mercados populares de todo el país, así como de su carne seca o de píldoras a base de serpiente de cascabel y pomadas o cinturones hechos con piel de serpientes (Ávila–Villegas, 2017). Otro factor importante que propicia el tráfico ilegal de especies en México es su captura para la venta como mascotas (Ávila–Villegas, 2017). En este sentido, Oaxaca es uno de los principales estados de México donde se captura fauna silvestre para abastecer el mercado ilegal internacional (Masés–García et al., 2021). Estos autores indican que, en total, se capturan ilegalmente 226 especies en este estado, de las cuales, 32 son serpientes. Por otra parte, algunos autores consideran que la captura con fines científicos es una actividad que puede poner en riesgo a ciertas especies (Lynch y Angarita–Sierra, 2016; Ávila–Villegas, 2017) y aumentar el riesgo de extinción de las poblaciones pequeñas y aisladas (Mintneer et al., 2014). Aunque existen muchos descubrimientos significativos y usos establecidos para las colecciones científicas, desafortunadamente hay casos en que los investigadores no capturan los animales de forma responsable (Henen, 2016). En México, no se cuenta con un análisis del efecto de la captura con fines científicos en la biodiversidad, por lo que se desconocen sus repercusiones. Beneficios de las serpientes para los seres humanos Todas las serpientes son carnívoras (Greene, 1997) y juegan un papel fundamental como depredadores en el ecosistema al influir en el flujo de nutrientes (Pradhan et al., 2014). Además, dado que muchas de ellas se alimentan de roedores, cumplen una función importante como controladoras de plagas (Fitch, 1949; Gayen et al., 2017), sobre todo en zonas de cultivos (Ávila–Villegas, 2017), y algunas de ellas actúan también como dispersoras secundarias de semillas (Reiserer et al., 2018). Además, las serpientes son presas de otros animales (Greene, 1997), por lo que contribuyen a la estabilidad de los ecosistemas (Ávila–Villegas, 2017) y pueden ser utilizadas como indicadores de la calidad del hábitat (Beaupre y Douglas, 2009). Otro beneficio importante de las serpientes es el uso de su veneno para la salud humana. Las propiedades toxicológicas del veneno se han venido estudiando desde hace más de 400 años, y se han obtenido al menos 20 productos terapéuticos o de

diagnóstico a partir de componentes del veneno (Waheed et al., 2017). Algunos de estos productos son el Captopril® (Enalapril, Lisinopril, Ramipril o Fosinopril, dependiendo de las modificaciones estructurales del Captopril), obtenido del veneno de Botrhops jararaca y utilizado para problemas de hipertensión, y el Agrastat® (tirofibán), obtenido de Echis carinatus, así como el Integilin® (eptifibatida), obtenido de Sistrurus miliarius barbouri, utilizados para el tratamiento de la enfermedad coronaria (Chan et al., 2016; Waheed et al., 2017; El–Aziz et al., 2019). Otros componentes, como la contorstratina, obtenida del veneno de Agkistrodon contortrix, pueden inhibir el crecimiento y la metástasis del cáncer de mama y el melanoma (Waheed et al., 2017). Asimismo, se ha observado que el veneno de varias especies de serpientes (p. ej., Agkistrodon contortrix, A. rhodostoma o Naja naja, entre otras) puede actuar sobre las células tumorales, por lo que es necesario realizar estudios para garantizar la seguridad y eficacia del uso de veneno de serpiente en la elaboración de medicamentos contra el cáncer (Calderon et al., 2014) Asimismo, a partir del veneno de las serpientes se ha aislado una gran cantidad de péptidos bioactivos que representan una fuente desconocida de nuevas aplicaciones médicas, aunque pocos de ellos han logrado llegar al mercado (Chan et al., 2016). Algunas de estas toxinas son usadas como antihemorrágicos, antibióticos, inhibidores de la agregación plaquetaria, tratamientos para la esclerosis múltiple y la distrofia muscular, anticoagulantes, antitrombóticos o trombolíticos (Waheed et al., 2017). Por lo tanto, es probable que con las mejoras tecnológicas en el campo del descubrimiento de fármacos, se logre identificar nuevos compuestos terapéuticos a partir del veneno de las serpientes (El–Aziz et al., 2019). Mitigación del conflicto entre los seres humanos y las serpientes Debido a la problemática que presenta la relación entre las personas y las serpientes en distintas partes del mundo, se han realizado propuestas para minimizar los conflictos mediante su control o mitigación; algunas de ellas son: 1, educación ambiental; 2, prevención y gestión de accidentes ofídicos; 3, estrategias de manejo, conservación y aprovechamiento sostenible en México; 4, exclusión; 5, ahuyentamiento; 6, eliminación de fuentes de alimento y refugios potenciales; 7, translocación; 8, captura con trampas; y 9, control letal. Estas propuestas se detallan a continuación. Educación ambiental La difusión del conocimiento acerca de las serpientes es quizás la forma más efectiva de evitar su sacrificio, por lo tanto, la implementación de programas de educación ambiental es una alternativa a la resolución del conflicto entre los seres humanos y las serpientes (Sullivan et al., 2014). Mediante estas actividades, se

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puede capacitar a los pobladores con la finalidad de que aprendan a coexistir con las serpientes y que no sea necesario sacrificarlas ni reubicarlas. Sin embargo, cuando se trata de serpientes venenosas, esto resulta complicado. En este sentido, los programas de educación ambiental relacionados con serpientes deben tener los siguientes objetivos: 1) que las personas aprendan a diferenciar las serpientes venenosas de las inofensivas; 2) mejorar el conocimiento que las personas tienen de estos organismos y difundir información acerca su biología, hábitos, función e importancia en el ecosistema, protocolos de accidentes ofídicos, incidencia de mordeduras de serpientes y manejo seguro de serpientes; 3) inculcar amor, respeto y admiración por estos animales. La educación ambiental debería formar parte de los programas educativos de las escuelas, a fin de que todos los niños y jóvenes mexicanos puedan tener acceso a este tipo de información. Los esfuerzos adicionales deben centrarse en las comunidades donde existe mayor riesgo de contacto con serpientes o una mayor incidencia de mordeduras de serpientes venenosas. Algunas actividades de educación ambiental pueden incluir la publicación y distribución de información, lo cual es una de las herramientas más poderosas para la conservación de la diversidad biológica, además de que, en el caso de las serpientes venenosas, contribuye a la prevención de accidentes (Lynch y Angarita–Sierra, 2016; Ávila–Villegas, 2017). Asimismo, se pueden impartir cursos, charlas y talleres para tratar de cambiar la percepción que tienen las personas de las serpientes. Este tipo de actividades interactivas en las que las personas pueden tener un contacto directo con animales son una herramienta importante en el aprendizaje y la actitud de formación (Kellert, 1985) y, gracias a ellas, se puede disipar el miedo y cultivar actitudes positivas (Ballouard et al., 2012, 2013) y convertir el miedo y la ignorancia en conocimiento y admiración (Greene y Campbell, 1982; Morris, 2017). Los herpetarios contribuyen de forma importante a estas actividades de educación ambiental, pues, además de que las personas pueden observar a los organismos vivos, obtienen información sobre distintos aspectos de su biología y ecología y acerca de su importancia, lo que ayuda a erradicar mitos y fomentar su protección. Existen asociaciones independientes en México dedicadas al estudio, la protección y la conservación de los anfibios y reptiles en las que se llevan a cabo diversas actividades de educación ambiental, difusión, investigación e incluso rescate y reubicación de serpientes, así como las relacionadas con los accidentes ofídicos; algunas de ellas se exponen en el Anexo 2. Asimismo, México cuenta con dos asociaciones nacionales dedicadas al estudio y la conservación de los anfibios y reptiles: la Sociedad Herpetológica Mexicana y la Asociación para la Investigación y Conservación de los Anfibios y Reptiles de México. Además, desde hace cuatro años, se lleva a cabo en México un congreso específico sobre vipéridos mexicanos y ofidismo. Sin embargo, estas iniciativas

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están dirigidas principalmente al público especializado (investigadores y académicos), por lo que es necesario que se organice un mayor número de actividades de educación ambiental para el público en general y que estén concebidas específicamente para las comunidades donde los accidentes ofídicos son recurrentes. Otras iniciativas que se han llevado a cabo en México consisten en la formación de brigadas comunitarias de rescate y conservación de serpientes. Estas brigadas se han establecido dentro de tres áreas naturales protegidas del estado de Hidalgo y existe otra en la comunidad indígena Comcaac de Punta Chueca, en Hermosillo, Sonora. Las brigadas están conformadas por pobladores locales, capacitados tanto en aspectos biológicos de las especies como en el manejo seguro de los organismos y la atención prehospitalaria del accidente ofídico. De este modo, estos grupos organizados trabajan activamente en la conservación de las serpientes y minimizan los efectos del conflicto con estos organismos (CONANP, 2019; Savethesnakes, 2020a, obs. pers). Asimismo, en otras iniciativas han combinado el arte y la ciencia para conservar a las serpientes y reducir el conflicto entre estas y los seres humanos en la localidad de Catemaco, Veracruz (Savethesnakes, 2020b). Medidas de prevención y gestión de accidentes ofídicos Otra estrategia para mitigar el conflicto entre los seres humanos y las serpientes es la prevención de las mordeduras de serpiente y la mejora en la gestión de los accidentes ofídicos. En México, existen diversos trabajos en los que se formulan recomendaciones y medidas para evitar sufrir una mordedura de serpiente venenosa (Vázquez–Díaz y Quintero–Díaz, 2005; Fernández–Badillo et al., 2011; Ávila–Villegas, 2017). Una de las principales recomendaciones es conocer las especies de serpientes venenosas del lugar en el que se habita o el que se pretende visitar (Ávila–Villegas, 2017). Sin embargo, resulta lamentable que siendo México el país con mayor cantidad de especies de serpientes venenosas de América (Campbell y Lamar, 2004), existen pocas publicaciones a escala nacional o regional que se centren en divulgar información veraz y completa al respecto y, en ocasiones, son de difícil acceso para los pobladores rurales. En este sentido, existe un único trabajo sobre todas las serpientes de México (Heimes, 2016), hay varios trabajos de divulgación acerca de las serpientes de cascabel (Ávila–Villegas, 2017; García–Padilla et al., 2018) y un programa de acción para la conservación de las especies de cascabel del género Crotalus (SEMARNAT, 2018). Además de ello, unos cuantos estados de la República Mexicana cuentan con guías específicas de las especies de serpientes venenosas, la mayoría de los cuales contienen información referente a las mordeduras de serpiente (tabla 1). La Organización Mundial de la Salud (OMS) ha publicado un plan de acción para reducir en un 50 % las muertes por mordedura de serpiente para el 2030. Para lograrlo, la OMS plantea cuatro objetivos

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o líneas de acción: 1) dar acceso a tratamientos seguros y eficaces, 2) empoderar e involucrar a las comunidades, 3) fortalecer los sistemas de salud y 4) promover la colaboración y coordinación mundial y la obtención de recursos (OMS, 2018; Williams et al., 2019). México cuenta con la producción de antídotos que se obtienen a partir del veneno de unas cuantas especies de dos de los 10 géneros de vipéridos mexicanos. El antídoto antiviperino de BIRMEX se fabrica a partir del veneno de Bothrops asper y Crotalus basiliscus (Segura et al., 2015; Sánchez et al., 2020; Neri–Castro et al., 2020a), mientras que Antivipmyn®, producido por el Instituto Bioclon de los Laboratorios Silanes, utiliza veneno de B. asper y C. simus (Neri– Castro et al., 2020a). En el caso de los coralillos, el único antídoto disponible en México (Coralmyn®), fabricado también por Bioclon, utiliza veneno de Micrurus nigrocinctus (Neri–Castro et al., 2020a). Estos antídotos son capaces de neutralizar adecuadamente los componentes tóxicos del veneno de otras especies (Sánchez et al., 2019). Por otro lado, se ha demostrado que la composición del veneno de las serpientes varía entre especies y que incluso puede variar entre los individuos de una misma especie, dependiendo de la edad, la distribución geográfica, la dieta, las condiciones ambientales, el estrés o la variabilidad genética entre de los individuos (Chippaux et al., 1991; Borja et al., 2018a, 2018b; Healy et al., 2018; Neri–Castro et al., 2020a). Lo anterior hace necesario continuar con las investigaciones de los venenos de otras especies, no solo para entender mejor sus componentes y mecanismos de acción, sino también para identificar posibles puntos débiles de los antídotos actuales y subsanarlos (Chippaux et al., 1991). Hasta la fecha, en México se han publicado diversos estudios con veneno de ejemplares de 35 de las 90 especies de serpientes venenosas distribuidas en el país (Neri–Castro et al., 2020a). De todas las especies estudiadas (35), 29 corresponden a vipéridos y siete a elápidos (anexo 1). Por otro lado, en México se cuenta con dos redes importantes de prevención y gestión de los accidentes ofídicos: la Red de Ayuda para el Accidente Ofídico de la Universidad Nacional Autónoma de México (Red AO–UNAM) y la Red Internacional de Centros de Referencia para el Control y Tratamiento de las Intoxicaciones por Animales Ponzoñosos (Redtox). Esta última cuenta con una aplicación para celular que permite obtener información biológica de las especies y los datos de contacto de especialistas, encontrar centros de atención y acceder a información general acerca del accidente ofídico. Sin embargo, la distribución y la disponibilidad de los antídotos son un problema importante, ya que en zonas rurales o comunidades muy alejadas no se tiene acceso a ellos. En muchos lugares de México, los pobladores siguen recurriendo a remedios tradicionales, culebreros (personas que se dedican a atender las mordeduras de serpiente a partir del uso de plantas medicinales y remedios tradicionales) y prácticas poco eficaces para contrarrestar los efectos de un envenenamiento grave.

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Los antídotos cuentan con una fecha de caducidad y, de acuerdo con las leyes mexicanas (Ley General de Salud, DOF, 2020), una vez vencida dicha fecha, queda prohibida su venta y suministro y deben destruirse (Hernández–Barrios et al., 1995; DOF, 2012, 2018a). Sin embargo, la efectividad de los antídotos no disminuye aún después de 20 años de la fecha de caducidad (Sánchez et al., 2019), por lo que es necesario realizar más estudios para evaluar la efectividad neutralizante de los antídotos mexicanos caducos, con la finalidad de poder modificar los criterios actuales y seguir utilizándolos. Ante la alarmante escasez mundial de antídotos, los productos caducos pueden brindar una alternativa para el tratamiento de mordeduras de serpiente (Sánchez et al., 2019). Estrategias de manejo, conservación y aprovechamiento sostenible en México En México existen normas y leyes que permiten y regulan el manejo de la fauna silvestre y que consideran además medidas para asegurar el trato digno y respetuoso de los animales, para poder llevar a cabo acciones de conservación, reproducción, aprovechamiento sostenible o control (Hernández–Silva et al., 2018). Estas actividades se encuentran reguladas principalmente por la Ley General de Vida Silvestre y su Reglamento (DOF, 2014, 2018b) y se apoyan también en la Norma Oficial Mexicana Ecol–059–2010 (DOF, 2019). Por otro lado, en México existen las Unidades de Manejo para la Conservación de la Vida Silvestre (UMA), que permiten lograr la conservación de especies y su aprovechamiento sustentable, primordialmente de especies nativas. De esta forma y sobre la base de un plan de manejo autorizado por la SEMARNAT, las UMA tienen como objetivo general la conservación del hábitat natural, las poblaciones y los ejemplares de especies de vida silvestre nativos, y pueden tener objetivos específicos como la restauración, la protección, el mantenimiento, la reproducción, la reintroducción, la investigación, la educación ambiental y el aprovechamiento sustentable, entre otros (DOF, 2014, 2018b). Existen también los Predios Intensivos de Manejo de Vida Silvestre (PIMVS), cuyo objetivo principal es la reproducción controlada para el aprovechamiento comercial de las especies (DOF, 2014, 2018b), ya sean nativas o exóticas. En el país existen distintas UMA y PIMVS, en los que se mantienen serpientes en cautiverio (herpetarios o serpentarios). Para que un herpetario contribuya a la conservación de las especies, debe llevar a cabo actividades de educación ambiental, de formación de profesionistas en la materia, de investigación y de aprovechamiento sustentable (Ávila–Villegas, 2017). En México existen algunos herpetarios que cumplen con todos o con la mayoría de estos criterios y están ubicados en 18 estados del país (anexo 3), aunque seguramente existen algunos más que cumplen con estas características, pero que lamentablemente no conocemos y que no logramos incluir en este apartado. En cuanto a otras estrategias de conservación específicas para serpientes, la única existente en México

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Tabla 1. Relación de los estados o regiones de la República Mexicana que cuentan con guías específicas sobre serpientes venenosas. Table 1. States or regions in the Mexican Republic that have specific guidebooks on poisonous snakes.

Estado o región

Autor

Aguascalientes

Sigala–Rodríguez y Vázquez–Díaz (1996)

Chiapas

Luna–Reyes y Suárez–Valázquez (2008); Suárez–Valásquez y Luna–Reyes (2009)

Hidalgo Fernández–Badillo et al. (2011); Fernández–Badillo (2017) Querétaro

Cruz–Pérez et al. (2018)

Michoacán

Alvarado–Díaz y Suazo–Ortuño (2006)

Nuevo León

Lazcano–Villareal et al. (2010)

Tabasco

Ojeda–Morales (2004)

Yucatán Escalante–Chan et al. (2017) Península de Yucatán

Llamosa–Neumann (2005)

Noroeste de México

Gatica–Colima (2013)

es el Programa de Acción para la Conservación de las especies del género Crotalus (SEMARNAT, 2019) que, sin duda, es una herramienta de suma importancia para lograr la conservación de las serpientes de cascabel, ya que México es el país con la mayor diversidad de cascabeles (Campbell y Lamar, 2004). Medidas de exclusión Otra propuesta de control ha sido el uso de cercos para impedir que las serpientes entren a algunos sitios. Para ello se propone el uso de una malla de por lo menos 1,8 metros de altura, en la que los primeros 60 cm sean de malla de ¼ de pulgada, el exterior debe ser lo más liso posible y la cerca debe tener una inclinación de 30° para evitar que las serpientes puedan trepar. Alrededor de la cerca debe haber un área de por lo menos 60 cm sin arbustos ni árboles, y tampoco deben haber arbustos ni árboles sobresaliendo de la cerca. La cerca debe estar enterrada entre 30 y 60 cm bajo el suelo, para evitar que los mamíferos puedan cavar túneles que permitan el acceso de las serpientes y, además, se debe hacer un mantenimiento continuo a la cerca para detectar posibles huecos o túneles y sellarlos (Klauber, 1982; Parkhurst, 2019). Este método es costoso y solo es útil si el área a proteger es pequeña, como un patio de recreo infantil (Klauber, 1982), por lo que puede no resultar una medida viable, sobre todo en zonas rurales de México. Medidas de ahuyentamiento La creencia de que existen sustancias que repelen a las serpientes o evitan su mordedura es muy antigua (Klauber, 1982) y, con ese fin, se han utilizado diversos productos, plantas y remedios: ácido cianhídrico, arsénico, creosota, DDT, gas de cloro, gas de carbón,

dióxido de carbono, bolas de naftalina, sulfato de nicotina y espray de pimienta, entre muchos otros (Klauber, 1982; Julian y Woodward, 1985). También se han usado algunos productos comerciales, como Snake Away®, Snake B Gon®, Snake Defense®, Snake Out®, Snake Shield®, Snake Stopper®, Sureguard Snake Repellens®, entre otros. Otros productos naturales que se han empleado como repelentes son el ajo y cebolla (Klauber, 1982) o aceites de diversas especies de plantas (Kraus et al., 2015), incluso se ha sugerido el uso de olor artificial de zorrillo o el almizcle de serpientes como Lampropeltis getula (San Julian y Woodward, 1985). Sin embargo, no existe realmente ninguna sustancia química, gas, aceite o producto que permita ahuyentar a las serpientes (San Julian y Woodward, 1985; Mengak, 2004; Parkhurst, 2019). Otra alternativa ha sido el uso de competidores por el alimento o enemigos naturales de las serpientes venenosas. Klauber (1982) menciona, específicamente para el control de serpientes de cascabel, el uso de ciervos, tejones, halcones cola roja o serpientes inofensivas de los géneros Lampropeltis y Pituophis, así como algunos animales domésticos como los cerdos. Estas prácticas pueden resultar desastrosas y, actualmente, se sabe que las especies exóticas invasoras están causando cambios drásticos en los sistemas ecológicos, alteran profundamente las comunidades y los ecosistemas, contribuyen al descenso de la diversidad biológica (Falaschi et al., 2020) y dañan los servicios ecosistémicos (Gallardo et al., 2019). Además, junto con otras amenazas, estas prácticas están catalogadas como impulsoras de la extinción de especies (Dueñas et al., 2021). Es más recomendable evitar el sacrificio de las serpientes no venenosas de una zona determinada, ya que son las competidoras naturales de serpientes

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venenosas como las víboras de cascabel, de tal forma que, cuantos más roedores consuman las serpientes inofensivas, menos alimento habrá para las serpientes de cascabel (Klauber, 1982). En el caso de México, existen varios géneros de serpientes no venenosas que, debido a su tamaño (superior al metro de longitud), pueden resultar excelentes aliadas en el control de especies venenosas, por lo que se recomienda no sacrificarlas. Tal es el caso de las especies de los géneros Clelia y Drymarchon y de algunas especies del género Lampropeltis (L. californiae, L. polyzona, L. splendida) y del género Masticophis (ejem: M. flagellum). Medidas para eliminar alimento y refugios potenciales Uno de los métodos más eficaces que se han propuesto para minimizar la presencia de serpientes en viviendas o zonas habitadas por personas consiste en eliminar las fuentes de alimento y los posibles refugios para las serpientes y sus presas (Brock y Howard, 1962; Klauber, 1982; Parkhurst, 2019). Para ello es necesario almacenar en recipientes herméticos todo el alimento que pudiera atraer a ratas, ratones, ardillas y otros roedores que son el alimento principal de las serpientes (Klauber, 1982). También es necesario limpiar los alrededores de las viviendas; evitar la acumulación de madera, escombros, cercos de roca, basura, arbustos o cualquier cosa que pudiera funcionar como un refugio para serpientes o ratones (Klauber, 1982; Mengak, 2004), y tapar las madrigueras de roedores que se encuentren (Klauber, 1982). Aunque estas prácticas pueden resultar útiles y se han empleado en otros países, en las zonas rurales y periurbanas de México resulta imposible, porque se utilizan cercos de piedra, cercos vivos u otro tipo de elementos naturales para la delimitación de las propiedades. Además, es común que los animales de compañía, las aves de corral e incluso el ganado estén sueltos o en corrales fabricados con troncos o roca, y que se alimenten directamente en el suelo o en comederos rústicos que permiten el acceso a los roedores, de tal forma que las serpientes son visitantes o residentes comunes en estos sitios. Además, esto no disminuye inmediatamente el número de serpientes, debido a su capacidad de pasar periodos largos sin alimento (Brock y Howard, 1962). Translocación La translocación, que se refiere al traslado de animales silvestres para liberarlos en otra localidad (Nielsen y Brown, 1988), se ha utilizado extensamente con diversos fines (Craven et al., 1998), incluso como medida de mitigación del conflicto entre los seres humanos y la fauna silvestre, para lo cual se transloca la fauna nociva a zonas donde no entre en conflicto con los humanos (Sullivan et al., 2014). En este sentido, se reconocen dos tipos de translocaciones (Hardy et al., 2001): las de larga distancia (TLD) o fuera del ámbito hogareño de la especie y las de corta distancia (TCD) o dentro del ámbito hogareño.

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La efectividad de la translocación en la mitigación del conflicto entre los seres humanos y las serpientes se ha evaluado principalmente en zonas urbanas (Sealy, 1997; Reinert y Rupert 1999; Nowak et al., 2002; Kinsbury y Attum, 2009; Massei et al., 2010; Devan–Song et al., 2016; Holding et al., 2014; Sullivan et al., 2014; MacGowan et al., 2017; Pitts et al., 2017), y se ha observado que en las TLD las serpientes translocadas presentan mayor movilidad y poca fidelidad a los sitios de liberación (debido a que no reconocen el nuevo territorio), así como un incremento de la tasa de mortalidad (Reinert y Ruppert, 1999; Nowak et al., 2002; Devan–Song et al., 2016; Sullivan et al., 2014). Por ejemplo, en el caso de las serpientes de cascabel de Norteamérica se ha detectado que la mortalidad se encuentra principalmente asociada al periodo invernal (King et al., 2004; Harvey et al., 2014), debido a que las serpientes translocadas no logran encontrar hibernáculos adecuados y son menos capaces de sobrevivir al invierno (Nowak y Ripper, 1999). En otros casos, la mortalidad de los ejemplares translocados ha estado asociada a los depredadores naturales (Devan–Song et al., 2016). Sin embargo, también se ha observado que en zonas con climas más cálidos no existen diferencias en la supervivencia de ejemplares translocados a larga distancia y ejemplares residentes, por lo que se recomienda este tipo de translocación como una opción viable (con ciertas especies de serpientes) para mitigar el conflicto entre los seres humanos y las serpientes (Corbit, 2015). Además, para los ejemplares residentes, la liberación de ejemplares nuevos representa un riesgo por la posible introducción de enfermedades como el paramixovirus (Nowak et al., 2002) y por las posibles alteraciones en la genética poblacional (Massei et al., 2010). Por otro lado, con las TCD, no se ha observado que las serpientes translocadas modifiquen su conducta ni sus patrones normales de actividad, ni que se haya modificado la dimensión de su zona de actividad ni incrementado la tasa de mortalidad (Brown et al., 2009). Incluso se ha observado que las serpientes tampoco se ven afectadas por el estrés a causa de la captura y el manejo repetido durante las translocaciones (Holding et al., 2014). Sin embargo, la TCD a medio y largo plazo no resuelve el conflicto entre los seres humanos y las serpientes, ya que los ejemplares translocados tienden a regresar a las zonas de captura (Brown et al., 2009, Corbit, 2015) y, aunque al parecer tratan de evitar los sitios específicos de la captura (Sealy, 1997), permanecen en zonas donde generan conflicto. De acuerdo con Sullivan et al. (2014), las translocaciones como medida de mitigación deben tener en cuenta aspectos como la ecología, el comportamiento, la sociabilidad y los requerimientos de hábitat para asegurar la supervivencia y persistencia de los ejemplares translocados. En este sentido, contar con información acerca del uso y selección del hábitat de las serpientes permite a los investigadores conocer cuáles son los componentes del hábitat más utilizados (Reinert, 1993), así como aquellos que contribuyen al éxito reproductivo y la supervivencia (Block y Brennan, 1993). Además,

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esto permite conocer aspectos sobre la hibernación (Olson et al., 2015) y los patrones de movimiento y actividad diarios y estacionales (Nowak et al., 2002). No existe ningún estudio publicado en el que se evalúe el efecto de la translocación de serpientes en México ni su uso como medida para la mitigación del conflicto entre los seres humanos y las serpientes. Sin embargo, es común que los biólogos, las asociaciones civiles e incluso las instituciones gubernamentales (p. ej. la CONANP, la Procuraduría Federal de Protección del Ambiente [PROFEPA]), lleven a cabo translocaciones y liberaciones de ejemplares, y difundan estas acciones a través de la prensa o las redes sociales. Sin embargo, en la mayoría de los casos, estas actividades se realizan sin conocer el ámbito hogareño de la especie ni su uso del hábitat, sin un previo análisis de los impactos posibles, en ocasiones sin saber cómo, cuándo y dónde es más adecuado realizar la liberación en función de la especie de que se trate y sin un seguimiento y monitoreo de los ejemplares translocados. Por todo ello, es necesario que estas actividades las lleven a cabo expertos o que estén basadas en estudios científicos bien detallados. Lamentablemente, como existen muy pocos estudios sobre el uso del hábitat y la ecología de las especies venenosas de México, es imprescindible generar este tipo de información para poder tomar decisiones más adecuadas al respecto. Captura con trampas Se pueden utilizar distintos tipos de trampas para capturar vivas a las serpientes, como las trampas de caída, solas o acompañadas de cercos de desvío, y las trampas de embudo (Foster, 2012). Asimismo, aunque se ha recomendado el uso de trampas de pegamento (Mengak, 2004), es muy común que las serpientes mueran adheridas a las trampas, además, estas trampas no son selectivas y muchos otros animales podrían adherirse y morir, por lo que no se recomienda su uso. Debido a que los intervalos de alimentación de las serpientes son muy espaciados e irregulares, el uso de trampas con cebos no funciona tan bien como con los mamíferos y las aves (Klauber, 1982). Además, las trampas de caída deben ser lo suficientemente grandes para evitar que las serpientes que hayan caído en la trampa escapen y deben colocarse en grandes cantidades (Foster, 2012), por lo tanto, su uso puede resultar caro y laborioso y no asegura el control total de las serpientes alrededor de una casa; por consiguiente, es una práctica poco viable para la mayoría de las comunidades rurales o periurbanas de México. Cabe considerar, además, que las trampas las debe utilizar personal experimentado y que se requieren los permisos y autorizaciones correspondientes para la captura y manipulación de los ejemplares; asimismo, los ejemplares capturados deben ser translocados y se debe evitar su sacrificio. Medidas de control letal Algunos autores sugieren que la solución más práctica y menos costosa al conflicto entre los seres humanos

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y las serpientes es el sacrificio de ejemplares (Massei et al., 2010). En este sentido, se han mencionado métodos como el uso de fuego, el sacrificio de los ejemplares en su refugio, (Klauber, 1982), la detonación con dinamita de los hibernáculos de invierno (Brock y Howard, 1962), la búsqueda constante y el sacrificio de ejemplares (Brock y Howard, 1962). Por otro lado, se ha sugerido el uso de animales domésticos como gallinas, patos o gansos, que matan serpientes venenosas e inofensivas de tamaño generalmente reducido, y hasta el entrenamiento de perros y gatos para matar serpientes (Brock y Howard, 1962). Uno de los métodos más comunes es simplemente sacrificar a las serpientes en cuanto se les encuentra (Ávila–Villegas, 2017). En México existen 193 especies de serpientes que se encuentran protegidas por la Norma Oficial Mexicana NOM–ECOL–059–2010. Para llevar a cabo medidas de control letal de forma legal se debe contar con una autorización para el manejo, el control y la resolución de problemas asociados a ejemplares o poblaciones que se tornen perjudiciales (DOF, 2018b). Además, para que se puedan autorizar medidas de control, debe existir un estudio previo, se deben mostrar pruebas del daño causado y se deben proponer las medidas de manejo recomendadas y el destino o uso que se dará a los ejemplares. Sin embargo, en las comunidades rurales, los pobladores no realizan ningún trámite y simplemente sacrifican a los ejemplares. Las serpientes son depredadores clave en los ecosistemas, principalmente en los agrícolas y de pastizales, ya que realizan un control eficaz de roedores, por lo que el sacrificio intencional de serpientes puede afectar las redes tróficas, alterar la dinámica poblacional depredador–presa y contribuir a la reducción de las poblaciones de serpientes. En último término, todas estas alteraciones en el ecosistema pueden afectar negativamente la salud humana (Pandey et al., 2016). Por todo ello, el sacrificio intencional de serpientes va en contra de los esfuerzos de conservación y de la normativa que protege a las especies (Nowak y Ripper, 1999) y no debería considerarse como una opción viable para la resolución del conflicto entre los seres humanos y las serpientes, ya que las poblaciones de serpientes están decreciendo de forma alarmante debido principalmente a la pérdida de hábitat (Reading et al., 2010), aunque también como consecuencia de la reducción de las poblaciones de anfibios (Zipkin et al., 2020) o de la sobreexplotación humana (Chuanwu et al., 2019). Además de todo ello, el sacrificio intencional de serpientes incrementa enormemente el riesgo de sufrir una mordedura mortal (Whitaker y Shine, 2000), ya que muchas mordeduras ocurren cuando las personas tratan de matar, manipular o interactuar deliberadamente con una serpiente venenosa (Wasko y Bullard, 2016). Por lo tanto, resulta necesario buscar alternativas para evitar el sacrifico de serpientes, no solo dentro de Áreas Naturales Protegidas y con especies bajo alguna categoría de protección, sino con todas las especies y en todo el territorio donde se distribuyen naturalmente.

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Consideraciones finales y recomendaciones Aunque el miedo a las serpientes que son venenosas puede estar justificado, la percepción negativa de estos animales está influenciada en parte por mitos, falsas creencias y propaganda negativa, que confieren a las serpientes una imagen exageradamente dañina. Sin embargo, pese al temor popular de morir a causa de una mordedura de serpiente venenosa, la mortalidad por este motivo en México se ha reducido considerablemente y la adecuada recuperación de un accidente de este tipo depende más de la calidad y eficacia del tratamiento que del veneno de la serpiente. Para poder mitigar el conflicto entre los seres humanos y las serpientes, es necesario llevar a cabo medidas para proteger y conservar a las serpientes, mediante la investigación científica; la participación, capacitación y vinculación ciudadana; la difusión y la educación ambiental. Además, se debe trabajar activamente en el tema del accidente ofídico para promover una cultura de prevención y a su vez, fortalecer todos los aspectos relacionados con su tratamiento, tanto a nivel prehospitalario y hospitalario, así como entre los pobladores de zonas urbanas y rurales, para contar con una población más informada y con actitudes encaminadas a la protección y conservación de las serpientes. Agradecimientos Al Consejo Nacional de Ciencia y Tecnología por la beca de doctorado otorgada al primer autor (número de becario 371195), a Guillermo Gil Alarcón, Edgar Neri Castro, Melisa Benard Valle, Miguel Borja Jiménez y Sergio Bárcenas Arriaga por la información proporcionada sobre el accidente ofídico y los estudios de venenos de serpientes mexicanas. A todos los que me proporcionaron contactos o información acerca de las actividades que se realizan en los distintos herpetarios, PIMVS, UMA, redes y asociaciones que trabajan en pro de la conservación de especies de serpientes mexicanas: Raúl Hernández Arciga, Laura Briseño, Eric Centenero, Ricardo Czaplewski, Felipe Agustín Lara Hernández, Víctor Velazquez, Miguel de la Torre Loranca, Gabriel del Valle, Fernando Frego, Guillermo Gil, Jacobo García Grajales, Alejandra Ham, Manuel Kim, Víctor Jiménez Arcos, Marco Antonio Lasserre Laborde, Ezequiel Lovera Rojas, Víctor Moreno Avendaño, Luis Pedro Ocampo Hernández, Ezequiel Julio Moreno Pérez, Javier Olivos Rivera, Javier Ortiz, Edgar Reina, Mario Reyna, Mauricio Rüed, Ariel Fernando Saldivar Flores, David Santollo, Fernando Toledo, Jonathan Torres, Alejandra Versategui Mendoza y Joaquín Villegas. Agradecemos también a los dos revisores anónimos por sus comentarios y sugerencias al manuscrito. Referencias Aguilar–López, J. L., 2016. Las serpientes no son como las pintan. Ciencia, 67: 6–13.

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Anexo 1. Especies de serpientes venenosas mexicanas cuyo veneno ha sido objeto de estudio. Fuente: 1, De Roodt et al. (2004); 2, Segura et al. (2012); 3, Martínez–Romero et al. (2013); 4, Neri–Castro et al. (2013); 5, Benard–Valle et al. (2014); 6, Carbajal–Saucedo et al. (2013); 7, Macías–Rodríguez et al. (2014a, 2014b); 8, Lomonte et al. (2016); 9, Durban et al. (2017); 10, Rivas–Mercado et al. (2017); 11, Saviola et al. (2017); 12, Segura et al. (2017); 13, Arnaud–Franco et al. (2018); 14, Borja et al. (2018a); 15, Borja et al. (2018b); 16, Guerrero–Garzón et al. (2018); 17, Mackessy et al. (2018); 18, Grabowsky y Mackessy (2019); 19, Román–Domínguez et al. (2019); 20, Neri–Castro et al. (2019a); 21, Neri–Castro et al. (2019b); 22, Benard–Valle et al. (2020); 23, Neri–Castro et al. (2020b); 24, Rivas–Mercado et al. (2020); 25, Zaragoza–Bastida et al. (2020); 26, Dashevsky et al. (2020); 27, Braga et al. (2020); 28, García–Osorio et al. (2020); 29, Ponce–López et al. (2021); 30, Archundia et al. (2021). Annex 1. Species of poisonous snakes in Mexico for which studies have been published regarding their poison. (For abbreviations of sources see above).

Familia Especie

Fuente

Familia Elapidae

Familia Especie Crotalus lepidus

Fuente 3, 10, 11

Micrurus browni

16, 22, 26 27, 30

Micrurus elegans

Micrurus diastema

8, 16, 26, 30

Crotalus mitchelli 13

8, 26, 30

Crotalus polystictus 17

Micrurus distans

Crotalus mictlantecuhtli

20, 29

Crotalus molossus

7, 14

Micrurus laticollaris 1, 6, 8, 16, 26, 27, 30

Crotalus pricei

Micrurus nigrocinctus 1

Crotalus ravus 25

Micrurus tener

5, 16, 26

Familia Viperidae Agkistrodon bilineatus

Crotalus ruber 13 Crotalus scutulatus 15

1, 19

Crotalus simus

4, 9, 20

Agkistrodon russeolus 19

Crotalus totonacus 24

Agkistrodon taylori 19

Crotalus triseriatus 25

Bothrops asper

1, 2

Crotalus tzabcan

9, 20

Crotalus aquilus 10

Crotalus willardi 11

Crotalus atrox 1

Metlapilcoatlus nummifer 1, 11, 28

Crotalus basiliscus

1, 12

Mixcoatlus melanurus 23

Crotalus catalinensis 13

Ophryacus smaragdinus 21

4, 9

Ophryacus sphenophrys 21

Crotalus ehecatl 20

Ophryacus undulatus 21

Crotalus culminatus

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Anexo 2. Asociaciones independientes en México dedicadas al estudio, la protección y la conservación de los anfibios y reptiles. Annex 2. Independent associations in Mexico dedicated to the study, protection and conservation of amphibians and reptiles.

Nombre

Ubicación

Red para la Conservación y Divulgación de los Reptiles Venenosos de Chiapas

Chiapas

Red de Ayuda para el Accedente Ofídico–UNAM

Ciudad de México

Vida Silvestre Coatl AC

Ciudad de México

Herpetológica LAB

Ciudad de México

Hook it

Ciudad de México

Fundación Haghenbeck

Ciudad de México

Usea

Ciudad de México

Red de Divulgación de Anfibios y Reptiles MX

Estado de México

Coatlan ICVS

Estado de México

Herpetario de la Sierra Gorda

Guanajuato

X–Plora Reptilia

Hidalgo

HerpMex

Jalisco

Red de anfibios y reptiles de Michoacán

Michoacán

Sociedad Herpetológica del Noroeste A.C.

Nuevo León

Anfibios y reptiles de Querétaro

Querétaro

Anfibios y reptiles de San Luís de Potosí

San Luís de Potosí

Red de divulgación de herpetofauna y expediciones Sinaloa

Sinaloa

Club de herpetología UNISON

Sonora

Grupo de anfibios y reptiles del sureste mexicano

Tabasco

Instituto Lorancai

Veracruz

Vida Verde y Conservación

Veracruz

Ekuneil

Yucatán

Red para la conservación de los anfibios y reptiles de Yucatán

Yucatán

Grupo para el conocimiento y la protección de los anfibios y reptiles Yumil Kaan

Yucatán

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Anexo 3. Herpetarios que contribuyen a la conservación de las especies de serpientes mexicanas: E, exhibición; Dea, difusión y educación ambiental; I, investigación; As, aprovechamiento sostenible; Afe, apoyo a la formación académica; R, reprodución. Annex 3. Herpetariums that contribute to the conservation of snake species in Mexico: E, exhibition; Dea, diffusion and environmental education; I, research; As, sustainable use; Afe, support for academic training; R, reproduction. Nombre Colección Zoológica de la Universidad de Aguascalientes Serpentario de la Paz Herpetario de Zoomat Herpetario de la Facultad de Ciencias de la UNAM Herpetario del Zoológico de Aragón UMA Deval Animal BioMuseo Reptiles SC Herpetario La Granja

Ubicación

Dea

Afe

Aguascalientes Baja California X Chiapas X

X X X

X X X X X

X X X

X X

Ciudad Ciudad Ciudad Ciudad Ciudad

X X X X X

x X X X X

X X X X X X X X X

X X X X X

de de de de de

México México México México México

Herpetario del Museo del Desierto Herpetario Zoológico de Colima Centro Regional de Educación para la Conservación Vivario de la Facultad de Estudios Superiores Iztacala, UNAM Animal City de México Herpetario Draconis Teutle Museo de Serpientes Herpetario de la Siera Gorda X–Plora Reptilia Herpetario del Zoológico de Guadalajara Herpetario del Zoológico de Morelia Herpetario del INIRENA, UMSNH Herpetario Cantil–Instituto de Biotecnología, UNAM Barranca Honda Herpetario de la Selva Morelos Herpetario de la UANL Herpetario de la BUAP Museo Viviente Herpetario del Africam Zafari Herpetario de la UAQ Herpetario CrocoCun Herpetario del Centro Ecológico de Sonora Herpetario del Parque Museo La Venta Herpetario del Centro Ecoturístico de El Cielo Herpetario Kukulkán Herpetario Coatlique Herpetario Palancoatl Vida Verde y Conservación UMA Tsáab Kaan

Coahuila Colima

X X

Durango

UMA Parque Bicentenario Animaya

Estado de México X Estado de México Estado de México X Estado de México X Guanajuato X Hidalgo X

X X X X X X X X X X X X X X X

X X X X X X

X X X X

Jalisco Michoacán Michoacán

X X X X X X X

X X X

X X

X X X

Morelos X X Morelos X X Morelos X X Nueva León Puebla Puebla X X Puebla X X Querétaro X X Quintana Roo X X

X X X X X X X X X X X X X X X X X X X X X

X X X X X

Sonora

Tabasco

Tamaulipas X Tlaxcala X Veracruz X Veracruz Verecruz X Yucatán X

X X X X X X X X X X

X X X X X X

X X X X X

Yucatán

X X X

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Are quartzite scree slopes used by birds to promote sound transmission in the Mediterranean forest? J. Pérez–González, G. Rey Gozalo, D. Montes González, S. J. Hidalgo de Trucios, J. M. Barrigón Morillas

Pérez–González, J., Rey Gozalo, G., Montes González, D., Hidalgo de Trucios, S. J., Barrigón Morillas, J. M., 2021. Are quartzite scree slopes used by birds to promote sound transmission in the Mediterranean forest? Animal Biodiversity and Conservation, 44.2: 175–184, Doi: https://doi.org/10.32800/abc.2021.44.0175 Abstract Are quartzite scree slopes used by birds to promote sound transmission in the Mediterranean forest? Birds generate vocalisations (songs and calls) to communicate. Acoustic communication may be hindered by habitat features so birds can use several strategies to favour sound transmission. Sound transmission depends on the acoustic properties of their habitats. Scree slopes, also known as 'pedrizas', are frequent in the Mediterranean forests of south and central western Spain. As the acoustic properties of these rocky grounds might favour sound transmission, we propose that birds might actively use 'pedrizas' to increase sound transmission. We assessed the following prediction of the hypothesis: the number of vocalisations recorded should be higher near the 'pedrizas' than in forest areas far away from 'pedrizas'. Using portable recorders in the Mediterranean forest of Monfragüe National Park, we found that the number of recorded vocalisations was higher near the 'pedrizas'. As this result was not due to differences in species richness, we consider it supports the prediction of the hypothesis. This is new evidence that birds might use a natural element within their habitat to increase sound transmission. Key words: Bird communication, Sound transmission, Rocky ground, Mediterranean forest, Natural soundscape Resumen ¿Las aves utilizan las pendientes pedregosas (pedrizas) de cuarcita para fomentar la transmisión del sonido en los bosques mediterráneos? Las aves generan vocalizaciones (melodías y llamadas) para comunicarse. Como la comunicación acústica puede verse obstaculizada por ciertas características del hábitat, es posible que las aves se valgan de diferentes estrategias para favorecer la transmisión del sonido. La transmisión del sonido depende de las propiedades acústicas del hábitat. Las pedrizas son frecuentes en los bosques mediterráneos del oeste meridional y central de España. Las propiedades acústicas de este terreno pedregoso podrían favorecer la transmisión del sonido, de forma que la hipótesis que se formula es que las aves podrían utilizar las pedrizas para aumentar la transmisión del sonido. Se evaluó la siguiente predicción de la hipótesis: el número de vocalizaciones grabadas debería ser superior cerca de las pedrizas que en zonas forestales alejadas. En este estudio se utilizaron grabadoras portátiles en el bosque mediterráneo del Parque Nacional de Monfragüe, y el número de vocalizaciones grabadas fue superior cerca de las pedrizas. Este resultado no se debió a diferencias en la riqueza de especies, de manera que respalda la predicción de la hipótesis. Este es un nuevo indicio de que las aves podrían utilizar un elemento natural de su hábitat para aumentar la transmisión del sonido. Palabras clave: Comunicación entre aves, Transmisión del sonido, Terreno pedregoso, Bosque mediterráneo, Paisaje sonoro natural Received: 12 XI 20; Conditional acceptance: 14 I 21; Final acceptance: 14 IV 21

ISSN: 1578–665 X eISSN: 2014–928 X

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Javier Pérez–González, Sebastian J. Hidalgo de Trucios, Research Group on Wildlife, Game Resources, and Biodiversity (GIRFCB), Biology and Ethology Unit, Veterinary Faculty, University of Extremadura, Cáceres, Spain.– Guillermo Rey Gozalo, Juan Miguel Barrigón Morillas, INTERRA, Lambda, Departamento de Física Aplicada, Universidad de Extremadura, Cáceres, Spain.– David Montes González, INTERRA, Lambda, Departamento de Física Aplicada, Universidad de Extremadura, Cáceres, Spain; ISISE, Departamento de Engenharia Civil, Universidade de Coimbra, Luis Reis dos Santos 290, Coimbra, Portugal. Corresponding author: G. Rey Gozalo. E–mail: guille@unex.es ORCID ID: J. Pérez–González: 0000-0003-0624-835X; G. Rey Gozalo: 0000-0003-0192-0944; D. Montes González: 0000-0002-5778-2782; S. J. Hidalgo de Trucios: 0000-0001-6606-5323; J. M. Barrigón Morillas: 0000-0001-9741-8291

Animal Biodiversity and Conservation 44.2 (2021)

Introduction Vocalisations (songs and calls) are communicative signals used by birds to transmit information (Catchpole and Slater, 2008). Mate attraction, territorial defence and warning signals are functions of bird acoustic communication (Catchpole and Slater, 2008; Pärckert, 2018; Riebel et al., 2019). Despite the advantages in the evolutionary context of their functions, songs and calls might also imply costs regarding energy expenditure and detection by predators (Ward and Slater, 2005; Catchpole and Slater, 2008). It is expected that birds produce sounds when their advantages outweigh costs. Therefore, the number of discrete vocalisations they produce might depend on the presence of potential mates, competitors, or predators, as well as on physical properties such as the acoustic transmission through the habitats. Information conveyed in bird vocalisations may be hindered by habitat features. Habitat complexity or noisy environments can modify signals and reduce their communicative function (Catchpole and Slater, 2008; Barker et al., 2009; Halfwerk et al., 2018). Birds might use several different strategies to avoid the loss of information, and they are able to match their songs and calls to the acoustic properties of the habitat (Hansen, 1979). For instance, Brumm (2004) found that nightingales (Luscinia megarhynchos) sing louder in noisy areas than in quieter places. Nicholls and Goldizen (2006) found that habitat type influences variation in the advertisement calls of the satin bowerbirds (Ptilonorhynchus violaceus), with transmission qualities of habitats being the main determinant of the effect. One of the main problems for the transmission of a bird sound is that of attenuation (Wiley and Richards, 1978). The signal intensity decreases with distance from the source (Fang and Ling, 2005). However, other factors also produce attenuation and hence hinder the transmission of bird sounds (Bass, 1991). Habitat characteristics such as vegetation structure can influence the transmission of vocalisations (Proppe et al., 2010; Nasiri et al., 2015). Dense foliage, for instance, increases attenuation (Martens 1980; Blumenrath and Dabelsteen, 2004). In order to promote sound transmission, birds use different strategies, such as singing on perches high up in the vegetation (Catchpole and Slater, 2008; Barker et al., 2009). Some studies show that the benefits of long–distance transmission are more relevant to the birds than the benefits of advertising performance ability or the costs of song production (Benedict and Warning, 2017). The Mediterranean forest includes an important community of animals and plants (Myers et al., 2000). Climax vegetation has been restricted to certain areas due to the strong pressure exerted by human activity over centuries. In central–western Spain, the best conserved Mediterranean forests are mainly found in quartzite mountainous areas where large steep stones are frequent. Within the forest there are some areas without vegetation forming scree slopes of quartzite origin, known in Spain as 'pedrizas' (see fig. 1). These fragmented rocks were produced during the Quaternary period by a gelifraction process (Pulido–Fernández et al.,

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2013). A typical habitat in many mountain ranges of south and central western Spain is therefore a Mediterranean forest in which there are areas covered by 'pedrizas'. The acoustic properties of these scree slopes differ from those in the surrounding forest. The presence of rocky surfaces predominates in 'pedrizas'. The acoustic impedance is high and can be considered mainly as acoustically reflective or hard surfaces in the terms indicated by the ISO standards (ISO 9613–2, 1996; ISO 1996–2, 2017). Therefore, when the sound waves reach such surfaces, a high percentage of the sound energy is reflected. In contrast, some elements present in the forest, such as bare earth and forest mass, contribute to sound attenuation (Bucur, 2006; Swearingen et al., 2013). The lower attenuation of bird sound on reflective surfaces has been shown in previous studies (Yip et al., 2017). In addition, 'pedrizas' are found in open spaces, thus being unaffected by other factors that affect the propagation and clarity of bird vocalisations, such as reverberation (Gogoleva, 2018). Vegetation mass is expected to attenuate sounds and hinder bird communication in the Mediterranean forest. However, areas with 'pedrizas' could favour the transmission of information. Consequently, the effect of 'pedrizas' on sound transmission should be higher when birds' vocalisations are emitted on the 'pedrizas' and this effect should decrease as the distance of birds' vocalisations from 'pedrizas' increases. We therefore hypothesized that birds might use scree slopes to promote sound propagation and hence, to increase the transmission of their vocalisations. This hypothesis is not tested in this study. However, we assessed the following prediction of the hypothesis: if birds actively use scree slopes to promote sound transmission, the frequency of vocalisations recorded near the 'pedrizas' should be higher than that in forest areas far away from 'pedrizas'. A high number of vocalisations recorded near the 'pedrizas' might be due to processes other than those related to sound transmission. 'Pedrizas' might ecologically influence bird distribution. In this case, a high number of vocalisations near 'pedrizas' would be due to the biased distribution of bird biodiversity. The relationship between the distance to 'pedrizas' and bird biodiversity is a necessary control to assess the prediction of the hypothesis. The lack of this relationship would support that a high number of vocalisations recorded near 'pedrizas' can be due to processes related to sound transmission. The prediction of the hypothesis was assessed in the Mediterranean forest of Monfragüe National Park located in central–western Spain. As expected, we found that the frequency of vocalisation was higher the closer the sampling point was to 'pedrizas'. However, there was not a greater number of bird species near the 'pedrizas'. Material and methods The study was carried out in a highly conserved Mediterranean forest in Monfragüe National Park. This

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100 m

Fig. 1. A, location of Monfragüe National Park; B, study area in Monfragüe National Park: numbers indicate the location of sampling points, black points indicate the sampling points at a distance from the 'pedrizas', and grey points indicate the sampling points on the three 'pedrizas' included in this study; C, view of a quarzite scree slope or 'pedriza'. Fig. 1. A, ubicación del Parque Nacional de Monfragüe; B, zona de estudio en el Parque Nacional de Monfragüe: los números indican la ubicación de los puntos de muestreo, los círculos negros indican los puntos de muestreo alejados de las pedrizas y los círculos grises, los puntos de muestreo situados en las tres pedrizas incluidas en este estudio; C, vista de una pedriza de cuarcita.

forest is composed of cork oaks (Quercus suber) and several scrub species such as strawberry tree (Arbutus unedo), laurustinus (Viburnum tinus), myrtle (Mirtus communis), false olive (Phillyrea angustifolia) and heaths (Erica spp.). 'Pedrizas' are frequent in this area. To record bird sounds, a portable Zoom H6 recorder was used with Roland Binaural microphones. The recorder was located at nine points in Monfragüe at different distances to three 'pedrizas' (fig. 1). Sampling locations were in a forest patch with the same vegetation species composition and the same northern orientation. The sizes of the 'pedrizas' are 3,450.63 m2 ('pedriza' of point 2), 3,342.04 m2 ('pedriza' of point 4) and 5,886.85 m2 ('pedrizas' of point 6). Recordings were made on February 28th, 2020. Bird vocalizations were recorded for 5 minutes at each sampling point. The first recording was made at at 11:30 a.m. and the remaining recordings were conducted consecutively. The last recording ended

at 13:00 p.m. During recordings, the temperature was 15 ºC, with a mix of sun and clouds, and wind speed of 5 km/h. All the vocalisations of the different bird species were identified in the recordings. Recordings were analysed and managed using the free software Sonic Visualiser (www.sonicvisualiser.org/) and Audacity (www.audacityteam.org/). Species were mainly identified aurally based on the the authors' experience, although the visual inspection of spectrograms helped in the process. The Birdnet web page was consulted in case of doubt (https://birdnet.cornell.edu/api/). A small proportion of vocalisations could not be identified as belonging to a specific species (see results) and were removed from subsequent analyses. From each recording, we counted the number of recorded vocalisations produced by each bird species. In addition to the number of vocalisations, the other important variable in our study was the distance of each vocalisation to the 'pedrizas'. The

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Table 1. Vocalisations from each bird species and across all species. The table shows the number and rate of recorded vocalisations, and the mean and standard error of the distance of the vocalisations to the 'pedrizas'. It also also shows the p values of the one–sample T test for the difference between the mean distance of vocalisations to 'pedrizas' and the mean distance of all sampling points to the 'pedrizas' (0.298 km): NV, number of vocalisations; V/m, vocalisations per minute; MD, mean distance; SE, SE distance. (Standard error, SE, of distance equals 0 when all vocalisations were recorded at the same sampling point, so there was no variance in distance to the 'pedrizas'; ND if data are insufficient to obtain the parameter or to conduct the analysis; ND in SE distance was obtained when there was only one vocalisation; ND in p values was obtained when there was no SE of the distance to 'pedrizas' of vocalisations.) Tabla 1. Vocalizaciones de cada especie de ave y de todas las especies. En la tabla se muestran el número y la tasa de vocalizaciones grabadas, así como la media y el error estándar de la distancia entre el lugar de las vocalizaciones y las pedrizas. También se muestran los valores de p de la prueba de T de una única muestra para la diferencia entre la distancia media desde el lugar de las vocalizaciones hasta las pedrizas y la distancia media desde todos los puntos de muestreo hasta las pedrizas (0,298 km): NV, número de vocalizaciones; V/m vocalizaciones por minuto; MD, distancia media; SE, error estándar de la distncia. (El error estándar de la distancia es 0 cuando todas las vocalizaciones fueron grabadas en el mismo punto de muestreo, de forma que no hubo varianza en la distancia a las pedrizas; ND si no hay datos suficientes para obtener el parámetro ni para realizar el análisis; ND en el error estándar se obtuvo cuando solo hubo una vocalización; ND en los valores de p se obtuvo cuando no hubo error estándar de la distancia entre el lugar de las vocalizaciones y las pedrizas).

Species

V/m

Fringilla coelebs

470

10.444

0.213

0.002

< 0.001

Certhia brachydactyla

194

4.311

0.382

0.009

< 0.001

0.007

< 0.001

Cyanistes caeruleus

188

4.178

0.218

Aegithalos caudatus

130

2.889

0.311 0.012 0.262

Phylloscopus collybita 70 1.556 0.243 0 ND Garrulus glandarius

43 0.956 0.204 0 ND

Sitta europaea

40 0.889 0.497 0 ND

Sylvia cantillans

22 0.489 0.161 0 ND

Parus major

0.400

0.249 0.015 0.005

Sylvia atricapilla

0.378

0.253 0.034 0.209

Lullula arborea

16 0.356 0.396 0 ND

Periparus ater

0.311

0.261

Erithacus rubecula

0.267

0.327 0.030 0.352

Turdus merula

0.178

0.283 0.044 0.760

0.005

< 0.001

Carduelis spinus

5 0.111 0.495 0 ND

Serinus serinus

3 0.067 0.495 0 ND

Carduelis cannabina

Carduelis carduelis

2 0.044 0.243 0 ND

Sylvia melanocephala

0.022 0.243 ND ND

Chloris chloris

0.022 0.204 ND ND

Across all species

1,257

0.067

27.933

distance of transmission of a vocalisation to the 'pedrizas' was quantified by averaging the distances from the sampling point where the vocalisation was recorded to the centroids of the three 'pedrizas'.

0.352 0.072 0.530

0.267

0.003

< 0.001

For each bird species, we recorded the number of vocalisations and the distance to 'pedrizas' for each vocalisation. We assessed the difference between the distance to the 'pedrizas' of the vocalisations and the

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Distance to 'pedrizas' (km)

0.5

0.4

0.3

0.2

Pm Sa Species

All

Fig. 2. Distance of vocalisations to 'pedrizas' recorded for each species and across all species. Boxplots were performed for the species with the highest number of vocalisations and with variance in the distance to 'pedrizas': Fc, Fringilla coelebs; Cb, Certhia brachydactyla; Cc, Cyanistes caeruleus; Ac, Aegithalos caudatus; Pm, Parus major; Sa, Sylvia atricapilla; Pa, Periparus ater; Er, Erithacus rubecula; Tm, Turdus merula; All, across all species). Distance to 'pedrizas' was quantified by averaging the distance of each vocalisation to the centroids of the three 'pedrizas'; horizontal dotted line, mean distance to 'pedrizas' of all sampling points. Fig. 2. Distancia entre las pedrizas y el lugar de las vocalizaciones grabadas de cada especie y de todas las especies. Se generaron diagramas de caja para las especies con el mayor número de vocalizaciones y con varianza en la distancia a las pedrizas: Fc, Fringilla coelebs; Cb, Certhia brachydactyla; Cc, Cyanistes caeruleus; Ac, Aegithalos caudatus; Pm, Parus major; Sa, Sylvia atricapilla; Pa, Periparus ater; Er, Erithacus rubecula; Tm, Turdus merula; All, todas las especies). La distancia a las pedrizas se cuantificó calculando la distancia media del lugar de cada vocalización a los centroides de las tres pedrizas; línea horizontal discontinua, distancia media entre todos los puntos de muestreo y las pedrizas.

averaged distance to 'pedrizas' of all sampling points (0.298 km) using a sample T–test for each species and across species. To assess the prediction of the hypothesis, we determined the relationship between the distance of the sampling points to the 'pedrizas' and the number of recorded vocalisations. A generalized linear mixed model (GLMM) fitted by maximum likelihood with Poisson distribution was used. For this model, the number of vocalisations was included as the explanatory variable, the distance to 'pedrizas' as the fixed factor, and sampling location as the random factor. In order to assess the effect of zero–inflation, the model was repeated after removing zero values. Species richness in the recordings was used as an estimate of bird biodiversity. Species richness was a presence–absence binomial variable and was obtained after transforming the variable number of vocalisations. A zero value was assigned to species for which no vocalisations were recorded, and a value of 1 was assigned to species for which at least one

vocalisation was recorded. Therefore, to determine the relationship between the distance to 'pedrizas' and bird biodiversity, a GLMM fitted by maximum likelihood with binomial distribution was conducted with the presence–absence variable (species richness) as the explanatory variable, the distance to 'pedrizas' as the fixed factor, and sampling location as the random factor. GLMMs were performed using the lme4 package (Bates et al., 2015) in R software (R Core Team, 2019). Model residuals were checked for heteroscedasticity using residual plots. No signs of heteroscedasticity were found. Spatial analyses were conducted with QGIS (2021, www.QGIS.org). Results The averaged distance of each sampling point to the centroids of the 'pedrizas' was 0.28 km (point 1 in fig. 1), 0.197 km (point 2), 0.205 km (point 3),

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Table 2. Relationship between the mean distance from the place of vocalisations to 'pedrizas' and the number of recorded vocalisations. Results from the GLMM with the number of vocalisations as the explanatory variable, the mean distance to 'pedrizas' as the fixed factor, and the sampling location as the random factor: SD, standard deviation. Tabla 2. Relación entre la distancia media desde el lugar de las vocalizaciones hasta las pedrizas y el número de vocalizaciones grabadas. Resultados del modelo mixto lineal generalizado con el número de vocalizaciones como variable explicativa, la distancia media a las pedrizas como factor fijo y el lugar de muestreo como factor aleatorio: SD, desviación estándar.

Term

Estimate SE

Variance SD

GLMM using all data Fixed factors Intercept

3.083

0.715

4.313

< 0.001

Distance to 'pedrizas'

–4.911

2.241

–2.191

0.028

Random factor Sampling location

0.657

0.811

GLMM after removing zero values Fixed factors Intercept

4.577 0.501

Distance to 'pedrizas'

–5.209

9.129 <0.001

1.586

–3.285

0.001

Random factor Sampling location

0.161 (point 4), 0.204 km (point 5), 0.243 km (point 6), 0.396 km (point 7), 0.497 km (point 8), and 0.495 km (point 9). The mean number of recorded vocalisations per minute was 27.93 and 7.3 % of vocalisations were not identified. Unidentified vocalisations corresponded to short calls that were not determined by the authors or from the Birdnet web page. We identified 20 bird species in the recordings (table 1). Table 1 also shows the number of vocalisations from each species and across all species, and the mean distance of the recorded vocalisation to the centroids of 'pedrizas'. The mean distance to 'pedrizas' of vocalisations tended to be lower than the averaged distance to 'pedrizas' of all sampling points for several species (table 1, fig. 2). However, the mean distance to 'pedrizas' across all species was lower than the mean distance to 'pedrizas' of all sampling points (table 1, fig. 2). The number of recorded vocalisations was negatively associated with the distance to the 'pedrizas' (table 2, fig. 3). This result was obtained when all data were used (table 2a, fig. 3) and after removing zero values (fable 2b, fig. 3). Despite the general trend we found in our data, for one species (Certhia brachydactyla) we recorded high numbers of vocalisations far away from 'pedrizas' (table 1, fig. 2, and high number of vocalisations at high distance to 'pedrizas' in fig. 3).

0.317

0.563

Although the number of vocalisations was negatively associated with the distance to 'pedrizas', no relationship was found between the distance to 'pedrizas' and species richness (table 3). The lower distance to 'pedrizas' was not related to an increase in the number of bird species obtained in the recordings. Discussion In the Mediterranean forest of Monfragüe National Park, the number of vocalisations recorded near the 'pedrizas' was higher than in forest areas far away from 'pedrizas'. These results support the hypothesis that birds might actively use scree slopes to promote sound transmission. In our study, the relationship between the distance to 'pedrizas' and bird biodiversity is a necessary control to assess the hypothesis. The distance to 'pedrizas' was not related to the estimated bird biodiversity. This result supports the notion that 'pedrizas' could not ecologically determine bird biodiversity. Factors related to resource availability or predation risk could not promote the presence of birds near 'pedrizas'. Therefore, the high number of vocalisations recorded near 'pedrizas' might be due to a better transmission of sound on scree slopes.

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Number of vocalisations

182

0 0.2

0.3 0.4 Distance to 'pedrizas' (km)

0.5

Fig. 3. Relationship between the distance to 'pedrizas' and the number of recorded vocalisations. Observed values are represented in jitter points to handling overplotting: black line, model prediction with all data (see table 2a); grey line, model prediction after removing zero values (see table 2b). Fig. 3. Relación entre la distancia desde el lugar de las vocalizaciones hasta las pedrizas y el número de vocalizaciones grabadas. Los valores observados se representan mediante puntos con ajuste de posición (jitter) para evitar el solapamiento (overplotting): línea negra, predicción del modelo con todos los datos (véase la tabla 2a); línea gris, predicción del modelo tras eliminar los valores de cero (véase la tabla 2b).

Signal transmission is particularly conditioned by the propagation of sound in the environment (Winkler, 2001). Birds use different strategies to favour sound transmission: the wren (Troglodytes troglodytes) chooses higher song posts (Barker et al., 2009), male blue–black grassquits (Volatinia jacarina) leap vertically above the dense grass (Wilczynski et al., 1989) and male bluethroats (Luscinia svecica) sing in flight (Sorjonen and Merilä, 2000; Catchpole and Slater, 2008). Studies taking into account sound frequencies have found different behavioural strategies depending on ecological conditions (Boncoraglio and Saino, 2007; Ey and Fischer, 2009). The great tit produces sounds with a lower maximum frequency and frequency range in forests than in open woodland (Hunter and Krebs, 1979). The little greenbul (Andropadus virens) varies its minimum frequencies across a rainforest gradient in Africa (Slabbekoorn and Smith, 2002). The tinamou (Eudromia elegans) increases its vocal amplitude in response to the increase in background noise (Schuster et al., 2012). Additionally, woodpeckers hammer their bills rapidly against a resonating substrate to communicate, with hard surfaces (dead trees) being preferred by individuals (Miles et al., 2018). The selection of dead trees might help to increase sound intensity. The hypothesis of this study could be proposed as a strategy in which birds might actively use a natural element within their habitat to favour sound transmission. We found a general trend based on all the species for which we recorded vocalisations in the

Mediterranean forest of Monfragüe National Park. Low sample sizes and the performing of multiple comparisons do not allow us to reach conclusions at species level. However, the hypothesis proposed in this work might be particularly important in some species. For instance, commonspecies such as the common chaffinch (Fringilla coelebs) or the blue tit (Cyanistes caeruleus) seem to follow the general pattern we found across all species. In the same way, the hypothesis might not work for other species. For instance, we found a high number of vocalisations of the short–toed treecreeper (Certhia brachydactyla) in a sampling point far away from 'pedrizas'. This sampling point was characterized to the presence of a water spring, and waterresources determines the presence and distribution of birds in Monfragüe National Park (J. Pérez–González, G. Rey Gozalo, D. Montes González, S. Hidalgo de Trucios and J. M. Barrigón Morillas, unpublished data). Different acoustic and biological features might determine for which species the effect of scree slopes are important in sound transmission. A bird species use of natural elements within its habitat in the context proposed in this work could open lines of future research. Individuals might use these elements as 'instruments' favouring sound transmission. Each habitat includes an exclusive set of natural elements and the home range of a bird species might include different habitats. Therefore, different populations of the same species might use specific elements within their habitats

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Table 3. Relationship between the mean distance of vocalisations to the 'pedrizas' and species richness. Results from the GLMM with species richness (presence–absence binomial variable) as the dependent variable, mean distance to 'pedrizas' as the fixed factor, and sampling location as the random factor. Tabla 3. Relación entre la distancia media desde el lugar de las vocalizaciones hasta las pedrizas y la riqueza de especies. Resultados del modelo mixto lineal generalizado con la riqueza de especies (presencia y ausencia) como variable explicativa, la distancia media a las pedrizas como factor fijo y el lugar de muestreo como factor aleatorio. Term

Estimate SE

Variance SD

Fixed factors Intercept

–1.189

0.499

–2.382

0.017

Distance to 'pedrizas'

1.000

1.530

0.414

0.679

Random factor Sampling location

0.071

0.267

to favour sound transmission. This approach might imply the existence of cultural behaviours in birds, as previously proposed in traits related to dialects (Luther and Baptista, 2010). The findings from this study may be a useful starting point for future studies on the modulation of bird behaviour depending on habitat features such as the presence of scree slopes. These future studies should be conducted using experimental designs or the collection of large amounts of data. Both approaches require an intense effort that might be justified by the findings from this study. Despite the possible bias related to the transmission distance of bird vocalisations, the high sound transmission on scree slopes, and the different habitat quality, the findings could be taken into account when designing experiments or collecting data. The use of scree slopes with larger surfaces, assessing of the mean territory size of the recorded bird species, counting individuals and the use of call rates, or the recording how long individuals spent singing on 'pedrizas' or in the surrounding forest could also help to control all possible biases. Funding This study was supported by the Diputación de Cáceres under grant AV–6. Support was also provided by the Consejería de Economía, Ciencia y Agenda Digital of Junta de Extremadura, by the European Union and European Social Fund (ESF) through grants for the strengthening of R&D&I through the mobility of postdoctoral researchers (PO17014), and also by the Consejería de Economía, Ciencia y Agenda Digital of Junta de Extremadura through grants for attracting and returning research talent to R&D&I centres belonging to the Extremadura Science, Technology and Innovation System (TA18019), where the University of Extremadura was the beneficiary entity in both cases.

Acknowledgements We thank the Editor and reviewers for comments concerning the manuscript. The authors would also like to express their gratitude to Mr. Pedro Atanasio Moraga for his technical support in setting up the recorder, Francisco Gómez Correa for his advice on spatial analyses, and the managers of the Monfragüe National Park for the use of their facilities and the reception of the project, especially its Director Mr. Ángel Rodríguez Martín. References Barker, N. K. S., Dabelsteen, T., Mennill, D. J., 2009. Degradation of male and female rufous–and–white wren songs in a tropical forest: effects of sex, perch height, and habitat. Behaviour, 146(8): 1093–1122. Bass, H. E., 1991. Atmospheric Acoustics. In: Encyclopedia of Applied Physics, vol. 2: 145–179 (G. L. Trigg, Ed.). Wiley–VCH Verlang GmbH & Co. KGaA, New York. Bates, D., Mächler, M., Bolker, B., Walker, S., 2015. Fitting Linear Mixed–Effects Models Using lme4. Journal of Statistical Software, 67(1): 1–48. Benedict, L., Warning, N., 2017. Rock wrens preferentially use song types that improve long distance signal transmission during natural singing bouts. Journal of Avian Biology, 48(9): 1254–1262. Blumenrath, S., Dabelsteen, T., 2004. Degradation of Great Tit (Parus major) Song Before and After Foliation: Implications for Vocal Communication in a Deciduous Forest. Behaviour, 141(8): 935–958. Boncoraglio, G., Saino, N., 2007. Habitat structure and the evolution of bird song: a meta–analysis of the evidence for the acoustic adaptation hypothesis. Functional Ecology, 21(1): 134–142. Brumm, H., 2004. The impact of environmental noise on song amplitude in a territorial bird. Journal of

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erbirds. Journal of Animal Ecololgy, 75: 549–558. Päckert, M., 2018. Song: The Learned Language of Three Major Bird Clades. In: Bird Species. Fascinating Life Sciences: 75–94 (D. T. Tietze, Ed.). Springer, Cham, Switzerland. Proppe, D. S., Bloomfield, L. L., Sturdy, C. B., 2010. Acoustic transmission of the chick–a–dee call of the Black–capped Chickadee (Poecile atricapillus): forest structure and note function. Canadian Journal of Zoology, 88(8): 788–794. Pulido–Fernández, M., Lagar–Timón, D., García–Marín, R., 2013. Geosites Inventory in the Geopark Villuercas–Ibores–Jara (Extremadura, Spain): A Proposal for a New Classification. Geoheritage, 6(1): 17–27. QGIS.org, 2021. QGIS Geographic Information System. QGIS Association. http://qgis.org R Core Team, 2019. R: a language and environment for statistical computing. R Foundation for Statistical Computing, Vienna, Austria. Riebel, K., Odom, K. J., Langmore, N. E., Hall, M. L., 2019. New insights from female bird song: Towards an integrated approach to studying male and female communication roles. Biological Letters, 15(4): 20190059. Schuster, S., Zollinger, S. A., Lesku, J. A., Brumm, H., 2012. On the evolution of noise–dependent vocal plasticity in birds. Biological Letters, 8(6): 913–916. Slabbekoorn, H., Smith, T. B., 2002. Habitat–dependent song divergence in the little greenbul: an analysis of environmental selection pressures on acoustic signals. Evolution, 56(9): 1849–1858. Sorjonen, J., Merilä, J., 2000. Response of male bluethroats Luscinia svecica to song playback: evidence of territorial function of song and song flights. Ornis Fennica, 77(1): 43–47. Swearingen, M. E., White, M. J., Guertin, P. J., Albert, D. G, Tunick, A., 2013. Influence of a forest edge on acoustical propagation: Experimental results. Journal of the Acoustical Society of America, 133(5): 2566–2575. Ward, S., Slater, P. J. B., 2005. Raised thermoregulatory costs at exposed song posts increase the energetic cost of singing for willow warblers, Phylloscopus trochilus. Journal of Avian Biology, 36(4): 280–286. Wilczynski, W., Ryan, M. J., Brenowitz, E. A., 1989. The display of the blue–black grassquit: the acoustic advantage of getting high. Ethology, 80(1–4): 218–222. Wiley, R. H., Richards, D. G., 1978. Physical constraints on acoustic communication in the atmosphere: implications for the evolution of animal vocalizations. Behavioral Ecology and Sociobiology, 3(1): 69–94. Winkler, H., 2001. The Ecology of Avian Acoustical Signals. In: Ecology of Sensing: 79–104 (F. G. Barth, A. Schmid, Eds.). Springer, Berlín. Yip, D. A., Bayne, E. M., Sólymos, P., Campbell, J., Proppe, D., 2017. Sound attenuation in forest and roadside environments: Implications for avian point–count surveys. Condor, 119(1): 73–84.

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Dietary plasticity in an invasive species and implications for management: the case of the monk parakeet in a Mediterranean city J. L. Postigo, J. Carrillo–Ortiz, J. Domènech, X.Tomàs, L. Arroyo, J. C. Senar Postigo, J. L., Carrillo–Ortiz, J., Domènech, J., Tomàs, X., Arroyo, L., Senar, J. C., 2021. Dietary plasticity in an invasive species and implications for management: the case of the monk parakeet in a Mediterranean city. Animal Biodiversity and Conservation, 44.2: 185–194, Doi: https://doi.org/10.32800/abc.2021.44.0185 Abstract Dietary plasticity in an invasive species and implications for management: the case of the monk parakeet in a Mediterranean city. Behavioural flexibility may play a relevant role during invasion of a new habitat. A typical example of behavioural flexibility favouring invasion success refers to changes in foraging behaviour. Here we provide data on changes in the foraging strategies of monk parakeets Myiopsitta monachus over a period of 17 years (2001–2017) in Barcelona city. During this time, consumption of food on the ground increased by more than 25 % and the consumption of anthropogenic food increased by 8 %. Detailed information about the food consumed is provided. Feeding on the ground and consumption of low plants allow parakeets to reach not only anthropogenic food but also crops, thereby increasing the risk of crop damage as the invasion evolves. Early detection of damage to crops is crucial in order to prevent further harm, and makes the precautionary principle highly relevant. Key words: Behavioural shift, Diet, Crop damage, Spillover, Longitudinal study Resumen La plasticidad en la dieta de una especie invasora y las implicaciones para su gestión. El caso de la cotorra argentina en una ciudad mediterránea. La flexibilidad del comportamiento puede ser un factor determinante en la invasión de un nuevo hábitat. Uno de los ejemplos más típicos de flexibilidad del comportamiento que favorece la invasión son los cambios en las estrategias de alimentación. En el presente estudio proporcionamos información sobre los cambios producidos en la estrategia alimentaria de la cotorra argentina (Myiopsitta monachus) durante 17 años (2001–2017) en la ciudad de Barcelona. A lo largo de este período, el consumo de comida en el suelo y el consumo de alimentos de origen humano aumentaron, respectivamente, más del 25 % y el 8 %. Se proporciona una descripción detallada de los alimentos consumidos. Alimentarse en el suelo y en vegetación baja pone al alcance de las cotorras comida de origen humano, pero también les da acceso a los cultivos, lo que aumenta el riesgo de que en los estados avanzados de la invasión, puedan ocasionar daños a la agricultura. La detección temprana de los primeros daños que se produzcan en los cultivos es fundamental para prevenir mayores daños en el futuro y hace que el principio de precaución sea especialmente relevante. Palabras clave: Cambio de comportamiento, Dieta, Daño a cultivos, Derrame, Estudio longitudinal Received: 3 V 21; Conditional acceptance: 14 VI 21; Final acceptance: 30 XI 21 José–Luis Postigo, Diversity and Conservation Research Team, Department of Animal Biology, Universidad de Málaga, Biogeography, Campus de Teatinos s/n., 29071 Málaga, Spain.– J. L. Postigo, José Carrillo–Ortiz, Jordi Domènech, Xavier Tomàs, Lluïsa Arroyo, Juan Carlos Senar, Museu de Ciències Naturals de Barcelona, Barcelona, Spain. Corresponding author: José–Luis Postigo. E–mail: joseluispostigosanchez@gmail.com ORCID ID: Jose–Luis Postigo: 0000-0002-1754-0025; Juan Carlos Senar: 0000-0001-9955-3892

ISSN: 1578–665 X eISSN: 2014–928 X

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Introduction

Material and methods

Invasive alien species are a major driver of recent extinctions (Bellard et al., 2016). Several species of birds have shown to be successful invaders (Kark et al., 2009), with one of the most successful being parrots (Menchetti and Mori, 2014). The monk parakeet Myiopsitta monachus Bodaert is native to parts of South America but it has become one of the most successful species of introduced parrots worldwide. The species can currently be found in many countries in Europe, in North, Central, and South America, and in Israel, among others (Briceño et al., 2019; Hobson et al., 2017; Postigo et al., 2017, 2019; Pruett–Jones et al., 2007). Knowledge of the behaviour of invasive species is critical for their management (Berger–Tal and Saltz, 2016; Weis and Sol, 2016). The process through which organisms explore and adopt to new food sources is known as behavioural plasticity. Behavioural flexibility may play a relevant role during a new habitat invasion and contribute to the likelihood of an alien species establishing a naturalized population (Martin and Fitzgerald, 2005). In the introduced habitat, some species may deploy a foraging behaviour that is distinctive from that in their native habitat, possibly explaining their higher invasion success (Pintor and Sih, 2009). The monk parakeet frequently feeds on the ground both in its native range (Aramburu, 1997; Bucher and Aramburú, 2014; Pezzoni et al., 2009) and in the invasive range (Borray–Escalante et al., 2020; Di Santo et al., 2013; South and Pruett–Jones, 2000), but interestingly, it rarely does so in the early stages of the invasion (Freeland, 1973; Santos and Sol, 1995; Shields, 1974). This change in foraging strategy (not feeding on the ground) can be considered behavioural flexibility. The new behaviour could be learnt from other species exploiting anthropogenic food, such as pigeons Columba livia Gmelein (Wright et al., 2010). It could also be the result of habituation to humans, or to a mix of the two possibilities. Feeding on anthropogenic food requires habituation to humans since in many Mediterranean urban environments people traditionally feed birds in the streets and squares, throwing them bread or seeds, in contrast with the use of bird feeders. Feeding on grass also requires habituation to humans as urban grass is more exposed, requiring the birds to be closer to humans than when perching on trees. As an adaptation to new environments, habituation may play an important role in facilitating invasion success and crop damage. However, to the best of our knowledge, the time required to perform this behavioural change has not been quantified previously. Using longitudinal data on feeding behaviour of the monk parakeet over a 17–year period in Barcelona city, we documented and quantified this change. Barcelona is home to one of the largest populations of monk parakeets in Europe, a population that has quintupled in size since the study began, increasing from 1,441 in 2001 to 7,100 individuals in 2017 (Borray–Escalante et al., 2020; Domènech et al., 2003). Understanding the behavioural flexibility of monk parakeets can help predict potential impact and allow the design of tailored management strategies (Wright et al., 2010).

The study was carried out in the north east of Spain in the city of Barcelona. Barcelona belongs to the Mediterranean biogeographical region (Council Directive 92/43/EEC) and it is characterized by warm dry summers and mild humid winters (Yaalon, 1997). Between 2001 and 2017 we collected information about the feeding events of monk parakeets in the city. The sampling unit was the feeding event. Therefore, independently of the size of the group of monk parakeets feeding, each observation of a group was recorded as one feeding event. Data were collected by walking around the transects of monk parakeets established during various studies of the species in the city (Carrillo–Ortiz, 2009; Domènech et al., 2003; Molina et al., 2016; Rodriguez–Pastor et al., 2012). The walks were conducted during the hours of maximum activity, from 8 a.m. to 2 p.m., and complemented with non–systematic observations collected by staff at the Natural History Museum of Barcelona or through contributions by local birdwatchers (table 1s in supplementary material). Data recorded were date, type of food: tree or shrub (for every plant except grass), grass, anthropogenic food (i.e. human food and seeds fed to the parakeets in the streets), or other (i.e. sand or gravel). In the case of plant food, we recorded the part of the plant eaten (leaves, fruits, flowers, seeds or sprouts) and also the genus or species when possible (fig. 1s in supplementary material). The species of food were classified as native or alien. We tested for variation in the seasonal use of different food sources using contingency tables and applying Pearson's Chi–Square test (x2). To test the seasonal variation in the use of biological structures we also applied Pearson’s Chi–Square test. To test for a potential feeding behavioural variation, we divided the data into two groups; group 1 from 2001–2009 and group 2 from 2010–2017, with 2,255 and 2,062 observations, respectively. This division provided a balanced number of observations between the two periods, compensating for the difference in the number of samples between years. We then compared the observations of ground feeding versus perched feeding groups, and the use of anthropogenic food versus natural food. In both cases, we applied a crosstabs analysis using Pearson's Chi–Square test. Statistical analyses were conducted using the R software v.4.0.2 (R Core Team, 2020) Results We recorded 4,317 feeding events, of which 3,064 (71 %) corresponded to urban vegetation. Of these, 2,552 (83%) were identified at least to the genus level. These plants were distributed in 36 genera from 22 families (table 1). Alien species made up 53 % of the genus/species consumed by parakeets. The most highly consumed family was poaceae (mainly grass), with 1,744 feeding events recorded (40 % of the total observations). Other natural sources of food were sand or gravel, with 11 observations (0.25 %).

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Table 1. Family and species or category of food types consumed by monk parakeets in Barcelona (2001–2017), with the number of samples (N) and the percentage of the total number of observations (%): NA, not aplicable. Tabla 1. Tipo, familia y especies o categoría de los alimentos consumidos por la cotorra argentina en Barcelona (2001–2017), con el número de muestras (N) y el porcentaje respecto de las observaciones totales (%): Na, no aplicable.

Origin /Family

Species/group

Urban vegetation Arecaceae Phoenix dactiliphera, Chamaerops humilis, Palm sp. Asteraceae Helianthus annuus Betulaceae Betula sp. Bignoniaceae Catalpa speciosa Casuarinaceae Casuarina sp. Cupressaceae Cupressus sempervirens, Juniperus sp. Fabaceae Acacia dealbata, Arachis hypogaea, Cercis siliquastrum, Mimosa sp., Parkinsonia aculeata, Robinia pseudoacacia, Sophora japonica, Tipuana tipu Fagaceae Quercus ilex Magnoliaceae Magnolia glandiflora Malvaceae Brachychiton sp., Tilia sp. Meliaceae Melia azedarach Moraceae Ficus sp. Myrtaceae Eucaliptus sp. Oleaceae Olea europaea, Ligustrum japonicum Phytolaccaceae Phytolacca sp Pinaceae Pinus sp., Abies sp. Platanaceae Platanus hispanica Poaceae Grass sp. Rosaceae Eriobotrya japonica, Malus domestica, Prunus persica, Prunus sp. Salicaceae Populus sp. Tamaricaceae Tamarix sp. Ulmaceae Celtis australis, Ulmus sp. No id. No id. Natural NA Sand, stones Anthropogenic NA Cookies, hot dog, legumes, unidentified plants NA Bread Poaceae Oryza sativa, Triticum aestivum, Phalaris canariensis Total

Anthropogenic food was consumed in 1,289 events (30 %). The primary anthropogenic food source was the bread thrown by the public to urban birds in squares and public gardens (92 % of anthropogenic food), followed by seeds such as corn and sunflower seeds, fruit, and hot dogs (8 %). The main food type consumed throughout the study was blades of grass (39 %), which is plentiful

150 3.47 15 0.35 1 0.02 1 0.02 3 0.07 60 1.39

74 1.71 10 0.23 2 0.05 21 0.49 6 0.14 3 0.07 1 0.02 7 0.16 5 0.12 5 0.12 41 0.95 1,696 39.29 18 372 1 53 472

0.42 8.62 0.02 1.23 10.93

0.25

54 1.25 1,187 27.50 48 1.11 4,317 100

in numerous parks and avenues in Barcelona. This was followed by anthropogenic food (30 %) and then by food provided in trees and shrubs (30 %). Sixteen plant families accounted for less than 1 % of each of the feeding events (table 1). We observed a marked seasonal shift in the diet of the monk parakeet based on the type of food consumed. Grass was the most widely used food source

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

Grass Anthropogenic Tree–Shrub Sand–Gravel

Summer

Autumn

Winter

Fig. 1. Seasonal variation in food sources used by monk parakeets in Barcelona (2001–2017). Fig. 1. Variación estacional de las fuentes de alimentación utilizadas por la cotorra argentina en Barcelona (2001–2017).

in spring and autumn (55% and 52% respectively). In summer, the three main sources of food (trees–shrubs, anthropogenic food, and grass) were almost equally represented, 34 %, 33 % and 30 %, respectively. In winter, the main source of food was that provided by trees and shrubs (40 %). All food source types were used in all seasons (fig. 1). Seasonal differences were significant (x29 = 353.00, p < 0.001) (fig. 1s in supplementary material). We analyzed the seasonal variation in the use of different plant parts. Leaves were the main biological structure consumed across seasons (56 %), followed by fruits (17 %), seeds (9 %), sprouts, and flowers (8 % each). Leaves (including grass) were also the most regularly consumed structure in every season (range: 45 % in winter–78 % in autumn). The num-

ber of biological structures consumed each season varied from three in autumn to five in spring and winter. These seasonal variations were significantly different (x212= 1373.60, p < 0.001) (fig. 2). We analyzed the difference between 'ground feeding' (anthropogenic food and grass) and 'perched feeding' (leaves –excluding grass–, seeds, fruits and sprouts) during the two periods described (2001–2009 versus 2010–2017) to determine whether the feeding behaviour changed over time. 'Ground feeding' increased by 26 % in the second period (x21 = 335.26, p < 0.001) (fig. 3). We also checked the variation in the use of food of natural origin versus the anthropogenic food during the two periods. Our findings showed that the anthropogenic food increased by 8 % (x21= 26.436, p < 0.001) (fig. 3).

100% 80%

Sprout Leaves

60%

Flowers

40%

Fruits Seeds

20% 0% Spring

Summer

Autumn

Winter

Fig. 2. Seasonal variation of the biological structures consumed by monk parakeets in Barcelona (2001–2017). Fig. 2. Variación estacional de las estructuras biológicas consumidas por la cotorra argentina en Barcelona (2001–2017).

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100%

80%

60%

40%

20%

2001–2009 Ground

2010–2017 Perched

2001–2009 Natural

2010–2017 Anthropogenic

Fig. 3. Changes in the foraging behaviour of monk parakeets in Barcelona between 2001–2009 and 2010–2017: A, ground feeding versus perched feeding (in trees and shrubs); B, use of anthropogenic food versus natural food sources. Fig. 3. Cambios en el comportamiento de alimentación de la cotorra argentina en Barcelona entre los períodos de 2001–2009 y 2010–2017: A, alimentación en el suelo con respecto a la alimentación en los árboles y arbustos; B, consumo de comida de origen humano con respecto a la comida de origen natural.

Discussion Diet variation Our results are consistent with most previous studies concerning the diet of monk parakeets. Most former studies, however, had short sampling periods (table 2). Here we observed that the most common food source of the parakeets in Barcelona throughout the year was the family poaceae (grass sp.), followed by anthropogenic food, mainly bread. As bread is made of wheat, which also belongs to the poaceae family, poaceae represented 65 % of the total observations. Various studies of monk parakeet diet in their native range found the poaceae or asteraceae families were the first or second choice of monk parakeets (table 2). Nevertheless, in our study, the asteraceae family represented less than 1 % of the diet, probably due to the low availability of this family of plants in the study area. Anthropogenic food was the second choice in our study, representing 29 % of the observations. In the invasive range, anthropogenic food is one of the most commonly consumed foods in most studies (table 2). In previous studies, anthropogenic food was a main factor affecting the distribution of monk parakeets in Barcelona (Rodriguez–Pastor et al., 2012). Consumption of sand and gravel was anecdotal in this study (0.3 %), and no capture of invertebrates was detected, although they were found to represent up to 28 % of the diet of nestlings in the native range (Aramburú and Corbalán, 2000). Invertebrates are likely underrated in this study due to the difficulty in determining whether a monk parakeet apparently eating grass on the ground is actually capturing invertebrates. The dominant fa-

milies of food in the diet of the monk parakeet across studies and the extended use of the anthropogenic food when available depict a diet pattern (table 2) that is consistent with the suggestion of Di Santo et al. (2013) that the consumption of food by monk parakeet is independent of its availability. However, this is not the case in Chicago (Hyman and Pruett–Jones, 1995). We therefore consider that the monk parakeet has an opportunistically selective feeding behaviour. The seasonal variation in the diet of the monk parakeet in Barcelona appears to be more homogeneous than that in other studies carried out (Aramburu, 1997; South and Pruett–Jones, 2000). In Barcelona all four types of food are included each season. Between three and five of five biological structures are present in the diet each season. Using stable isotopes in Barcelona, Borray–Escalante et al. (2020) found that the most commonly used food by monk parakeets in summer in Barcelona was anthropogenic food and grass (42 % and 27 %, respectively). The order of preference of food sources was identical in both studies and the proportion each one represented in the diet was reasonably close to the respective 33 % and 30 % they represent in summer in this study, considering the isotope study measures assimilation during the moult season and this study measures frequency of the feeding events. It is also possible that the anthropogenic food could be more used in the area where the monk parakeets were trapped for the isotopes study, in relation to the whole city sampled in this study, but Borray–Escalante et al. (2020) ruled out the possibility that the isotopic study overrepresented anthropogenic food respective to other sources. One particularity of the seasonal diet pattern in Barcelona is that anthropogenic food and

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grass combined represent between 57 % and 83 % of the diet each season. In contrast, in colder areas of the USA, monk parakeets were found to feed exclusively on bird seeds in winter. In its native range, on the other hand, during the austral winter, up to 70 % of their food is crops, and in the austral summer, monk parakeets living in the country feed solely on wild plants (Aramburu, 1997; South and Pruett–Jones, 2000). The homogeneous diet of these parakeets in Barcelona throughout the seasons is probably due to the typically mild winters in the Mediterranean basin and the large number of alien plant species cultivated in gardens and parks, species that provide food to monk parakeets throughout the whole year. In effect, most of the taxa (species/genus, 52.8 %) consumed by monk parakeets in Barcelona are not native. Consequently, this food provided by humans, directly or indirectly, can increase the breeding success and survival of monk parakeets (Chamberlain et al., 2009). The monk parakeet population of Barcelona experiences some of the highest reproductive indexes known for the species, with more breeding attempts per season and higher hatching success than anywhere else, including in its native range (Senar et al., 2019). This finding is supported by the fact that the population growth and spread rates in Mediterranean regions are higher than those in Atlantic regions (Postigo et al., 2019). Behavioural shift Behavioural plasticity has been related to higher invasive success (Sol et al., 2013). South and Pruett–Jones (2000) described how monk parakeets in Chicago developed a foraging innovation by starting to use bird feeders in backyard gardens in winter. Here, we document and quantify how ground feeding of these parakeets increased by 26 % and the use of anthropogenic food increased by 8 % during the 17 years of our study. Such a behavioural shift greatly increases the food sources available to monk parakeets, allowing them to access grass and anthropogenic food. Such food is available throughout the year and virtually un-

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limited, as grass is a fast growing plant and because the public, providing more food as the population of parakeets grows provides more and more food (Borray–Escalante et al., 2020). Besides, this behavioural shift to ground feeding can supply parakeets with a third important source food, that is, low crops such as tomatoes, wheat, corn, and sunflowers, which may be common in suburban areas. Such crops, their third food source, are typically abundant during a particular season, and parakeets can learn to exploit them as they do in their native range (Aramburu, 1997). In its original range, the monk parakeet is considered one of the main bird pest species causing damage to crops in several countries in South America (Bruggers et al., 1998; Bucher, 2021; Spreyer and Bucher, 1998). In its invasive range (e.g. U.S., Belgium, or U.K.), monk parakeets are often not considered a threat to agriculture because they mainly occupy urban areas where the damage to crops is less extensive (Muñoz and Real, 2006; South and Pruett–Jones, 2000; Tayleur, 2010; Weiserbs, 2010). In Spain and Italy, however, monk parakeets cause relevant damage in crops at the outskirts of the cities of Barcelona and Rome (Battisti, 2019; Senar et al., 2016). This damage occurs when the parakeets leave the city to feed on the crops on the outskirts of the city, causing damage especially to low plants, such as tomatoes. The reason why a widespread invasive species like the monk parakeet causes crop damage only in a few locations after decades of presence could be that the population of monk parakeets occurring near crops must be relatively large, as predicted by Bucher (1992) for the native population. According to our results, we hypothesize here a plausible relation between the availability of anthropogenic food in cities and the potential crop damage in their surroundings. The results of this study suggest the monk parakeet population in the city of Barcelona may be subsidized by anthropogenic food resources, which could also contribute to the population increasing five–fold in 17 years in a disturbed area, and to the species being considered an 'abundant vertebrate' according to Goodrich and

Table 2. Bibliographic review of the feeding studies of monk parakeets: A, distribution area (N, native; I, Invasive); C/S, country/state (ARG, Argentine; PE, Pennsylvania; NJ, New Jersey; CH, Chicago; ITA, Italy; FL, Florida; SPA, Spain); M, methodology (Sc, stomach content; Cs, cafeteria study; OQl, Observational/ qualitative; O/Qn, observational/quantitative; Fe, feeding events; Cc, crop contents; Is, Isotopos study; Ol, observational longitudinal study; BWR, Boreal winter resource; FG, feed on the ground; Cp/c, crops present/ consumed; N, sample size; FB, feeding behaviour shift (Y, yes; N, no; Bf, use of birdfeeders; Ua, use of anthropogenic food; Fg, feeding on the ground). (Y, yes; NA, not aplicable; N, no; * Austral summer). Tabla 2. Revisión bibliográfica de los estudios de alimentación de la cotorra argentina: A, área de distribución (N, autóctona; I, invasora); C/S, país/estado (ARG, Argentina; PE, Pensilvania; NJ, Nueva Jersey; CH, Chicago; ITA, Italia; FL, Florida; SPA, España); M, metodología (Sc, contenido estomacal; Cs, estudio en una cafetería; OQl, observacional/cualitativo; O/Qn, observacional/cuantitativo; Fe, eventos de alimentación; Cc, contenido de cultivos; Is, estudio con isótopos; Ol, estudio observacional longitudinal; BWR, recurso de invierno boreal; FG, alimentarse en el suelo; Cp/c, cultivos presentes/consumidos; N, tamaño de muestra; FB, cambio de comportamiento de alimentación (Y, si; N, no; Bf, consumo de comida en comederos para aves; Ua, consumo de comida de origen humano; Fg, alimentación en el suelo). (Y, si; N, no; NA, no aplicable;* verano austral).

Preferred Study family/group period BWR FG Cp/c

Characteristics

I SPA Ol

Poaceae > anthropogenic food

I–XII Grass and (17 years) anthropogenic

Y N/N

Test the evolution of the feeding behaviour

4,317 feeding observations

Borray–Escalante et al. (2020)

Avery and Shields (2018)

Di Santo et al (2013)

South and Pruett–Jones (2000)

Shields (1974)

Freeland (1973)

Pezzoni et al. (2009)

Aramburu and Bucher (1999) Aramburú and Corbalán (2000)

Aramburu (1997)

Reference

Y This study (↑Fg)

N ARG Sc Poaceae I–III* Cultivated Y Y/Y Post–mortem 166 indiv. N (14 months) plants analysis 4 colonies N ARG Cs Asteraceae VIII–XII* NA NA Y/NA Wild MP captured 5 indiv. N > Poaceae for the study 5 tests N ARG Cs Asteraceae XII* NA NA NA/N Poaceae 11 indiv. nestlings N > Poaceae underestimated 21 samples (wild origin) N ARG Sc Poaceae XII* NA Y NA/NA Nestlings 32 indiv. N > Asteraceae diet study 3 colonies I PE O/Ql Corn, native fruit, IX–VIII Apples Y N/N First couple in 5 indiv. Y seeds and bread and pears the area, (Bf) year around. shy of feeders, Native fruit in summer feed on patio decks. I NJ O/Qn Ulmaceae III–IV Bird feeders N N/N Arrived in winter, 2 indiv. N > Cupressaceae (14 days) left in spring > Pinaceae I CH Fe Bird seeds VII–VI Bird seeds Y N/NA Grass no available 1,426 feeding Y > Rosaceae observations (Bf) > Poaceae (in winter) I ITA Fe Asteraceae VII–IX NA Y N/N Availability: 558 feeding N > Poaceae > … Poacea > Asteraceae. observations I FL Cc Asteraceae Summer NA NA N/N Food subsidized 26 indiv. nestlings Y by human activities (Bf) sunflower > millet I SPA Is Anthropogenic food IX–VIII NA Y N/N Measures assimilation 72 indiv. Y > Poaceae (herbaceous) during moult season (Ua)

A C/S M

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Buskirk (1995). Consequently, they are susceptible to produce damage to other species outside the disturbed areas by what we could term 'spillover crop–damage' if they eventually leave the city in large numbers to feed on crops in the surroundings, in a process equivalent to 'spillover predation' (Boarman, 2003). This seems to be the process occurring in Barcelona and Rome with monk parakeets (Battisti, 2019; Senar et al., 2016), and also with rose–ringed parakeets Psittacula krameri Scopoli around the city of Lihue in Hawai (Avery and Shiels, 2018). All three of these populations of monk parakeets and rose–ringed parakeets are among the largest in their country/state (Avery and Shiels, 2018; Postigo et al., 2019). Moreover, the spread of monk parakeet assisted by white storks recently, in rural areas has been described (Hernández–Brito et al., 2020). The potential consequences of this association are unknown, but could potentially increase the spreading capacity of monk parakeets and consequently increase the crops exposed to damage. Implications for management As the monk parakeet population in Barcelona continues to increase and expand, as its behavioural shift allows it to access virtually unlimited food sources (grass and anthropogenic food), and as locally relevant crop damage have been quantified, we strongly recommend population management should be considered to avoid crop–damage by spillover increasing even further in the future. This recommendation becomes even more important considering the population keeps growing and expanding. Limiting access to food sources can reduce the growth rate of urban bird populations (Haag– Wackernagel, 1995; Senar et al., 2017), but given the growth rate of the population of monk parakeets and the fact that their main food sources in Barcelona are grass and anthropogenic food, it seems unlikely that the population size will decrease without extractive methods (Conroy and Senar, 2009; Dawson Pell et al., 2021; Senar et al., 2021). Although culling parakeets has been carried out successfully in the past (Esteban, 2016; Senar et al., 2021) and would be legal given the monk parakeet was declared an Invasive Alien Species in Spain in 2011 (Real Decreto 1628/2011), control plans in various cities have been cancelled given the opposition from animal right activists who consider the species charismatic and are against any lethal method of removal (Hernández–Brito et al., 2018). Nevertheless, we stress that social considerations should not prevent relevant governmental bodies from pursuing efforts to control the species. Recently, a management plan in the Canary Islands (Spain) successfully eradicated a small number of rose–ringed parakeet from La Palma island by combining trapping and shooting with social collaboration (Saavedra and Medina, 2020). We recommend, therefore, a multidisciplinary approach, combining various methods to remove the parakeets, and building as much social learning and trust as possible by promoting effective communication and education of the public (Crowley et al., 2019; Perry and Perry, 2008; Senar et al., 2021; Shackleton et al., 2019). If our hypothesis is correct, all the populations of monk parakeets in the

Mediterranean region, independently of their present size, are susceptible to produce crop damage in the future, when they reach the appropriate size, given that they are all growing exponentially (Postigo et al., 2019). In consequence, we strongly recommend managing all the populations of monk parakeets in the Mediterranean region, independently of their size, so as to prevent future damage to crops. In addition, further efforts to identify the limiting factors affecting the monk parakeet populations in the Mediterranean region and early detection of emerging crop damage by monk parakeets in new areas are essential to prevent potential massive damage and the need for costly control measures. In our opinion, this is a fine example of where the precautionary principle could be applied (Edelaar and Tella, 2012; Kumschick Brunel et al., 2001). Acknowledgements The fieldwork for this study was supported by research project CGL-2020 PID2020-114907GB-C21 awarded to JCS from the Ministry of Economics and Enterprise, Spanish Research Council, Spain. Thanks to Parcs i Jardins of Barcelona City Council for facilitating work in Barcelona parks. Thanks also to Alba Ortega–Segalerva, Jordi Domènech, Danielle Mazzoni, Helena Navalpotro, Monica Navarro and all the volunteers for field assistance. References Aramburu, R. M., 1997. Ecología alimentaria de la cotorra (Myiopsitta monachus monachus) en la provincia de Buenos Aires, Argentina (Aves Psittacidae). Physis, 53: 29–32. Aramburu, R. M., Bucher, E. H., 1999. Food preferences in the Monk Parakeet Myiopsitta monachus (Aves: Psittacidae) in captivity. Ecología Austral.: Preferencias alimentarias de la cotorra Myiopsitta monachus (Aves: Psittacidae) en cautividad. Ecología Austral, 9: 11–14. Aramburú, R., Corbalán, V., 2000. Dieta de pichones de Cotorra Myiopsitta monachus monachus (Aves: Psittacidae) en una población silvestre, Ornitología Neotropical, 11: 241–245. Avery, M. L., Shiels, A. B., 2018. Monk and Rose– Ringed Parakeets. In: Ecology and management of terrestrial vertebrate invasive species in the United States: 333–357 (W. C. Pitt, J. C., Beasley, G. W. Witmer, Eds.). CRC Press Taylor and Francis Group, Boca Raton. Battisti, C., 2019. Impact of monk parakeet Myiopsitta monachus on commercial orchards: evidence on persimmon Diospyros kaki fruits (Rome, central Italy). Alula, 26: 139–142. Bellard, C., Cassey, P., Blackburn, T. M., 2016. Alien species as a driver of recent extinctions. Biology letters, 12(2): 20150623. Berger–Tal, O., Saltz, D., 2016. Conservation Behavior. Cambridge University Press Cambridge. Boarman, W. I., 2003. Managing a subsidized preda-

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The selection of anthropogenic habitat by wildlife as an ecological consequence of rural exodus: empirical examples from Spain A. Martínez–Abraín, X. Ferrer, J. Jiménez, I. C. Fernández–Calvo

Martínez–Abraín, A., Ferrer, X., Jiménez, J., Fernández–Calvo, I. C., 2021. The selection of anthropogenic habitat by wildlife as an ecological consequence of rural exodus: empirical examples from Spain. Animal Biodiversity and Conservation, 44.2: 195–203, Doi: https://doi.org/10.32800/abc.2021.44.0195 Abstract The selection of anthropogenic habitat by wildlife as an ecological consequence of rural exodus: empirical examples from Spain. The increasing urbanization of the landscape is a major component of global change worldwide. However, it is puzzling that wildlife is selecting anthropogenic habitats despite the availability of apparently high– quality semi–natural (i.e. less intensively modified) habitats. Definitive explanations for this process are still lacking. We have previously suggested that colonization of the urban habitat is initially triggered by ecological processes that take place outside urban areas as a consequence of past rural exodus. Here we present a diverse array of examples of selection of several types of anthropogenic habitat by wildlife in Spain (including transportation infrastructure, human–exclusion areas, urban areas under construction, cities, reservoirs, quarries and landfills) in support of this idea. Wildlife is moving out of its historical ecological refuges and losing fear of harmless urban humans. Mesopredators are rebounding by mesopredator release, due to ceased human persecution, and shrubs and trees are claiming former agricultural habitats. Together, these factors force many species to move to urbanized areas where they find open habitats, food associated with these habitats, and protection against predation. Hence, the classical balance of costs and benefits that takes place once inside urban areas, would actually be a second step of the process of colonization of urban areas. A better understanding of the initial triggers of urban colonization could help us increase the biological value of human–made habitats for wildlife in the future. Key words: Changed human attitudes, Mesopredator release, Loss of fear, Human depopulation, Shrub and tree encroachment, Urban areas Resumen La fauna silvestre selecciona hábitats antropógenos como consecuencia ecológica del éxodo rural: ejemplos empíricos de España. Uno de los principales componentes del cambio global en todo el mundo es el aumento de la urbanización del territorio. Sin embargo, es desconcertante que la fauna silvestre seleccione hábitats antropógenos a pesar de que existan hábitats seminaturales (modificados con menor intensidad) aparentemente de buena calidad. Todavía no existe una explicación definitiva para este proceso. Se ha sugerido con anterioridad que la colonización de los hábitats urbanos se produce en una primera fase a causa de procesos ecológicos que tienen lugar fuera de las zonas urbanas, como consecuencia del éxodo rural del pasado. Para respaldar esta idea, en este estudio presentamos una serie de ejemplos en los que diversas especies de fauna silvestre de España seleccionan varios tipos de hábitats antropogénicos (infraestructuras de transporte, zonas de acceso restringido, zonas urbanas en construcción, ciudades, embalses, canteras y vertederos) por razones asociadas al despoblamiento del rural. La fauna silvestre está saliendo de sus refugios ecológicos y está perdiendo el miedo a los humanos inofensivos de las zonas urbanas. Los mesodepredadores están repuntando debido a la liberación del mesodepredador y al cese de la persecución humana, y la vegetación espontánea está volviendo a colonizar los antiguos hábitats agrícolas. Estos factores obligan a muchas especies a desplazarse a zonas urbanas donde encuentran hábitats abiertos, alimento asociado a estos hábitats y protección contra la depredación. Por consiguiente, el balance de costes y beneficios en el seno de las zonas urbanas, sería en realidad un segundo componente del proceso. Entender mejor los factores iniciales que desencadenan la colonización del medio urbano podría ayudarnos a dar más valor biológico para la fauna silvestre a los hábitats creados por los seres humanos. ISSN: 1578–665 X eISSN: 2014–928 X

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Palabras clave: Cambio de actitud de las personas, Liberación del mesodepredador, Pérdida del miedo, Despoblamiento del rural, Matorralización, Zonas urbanas Received: 22 II 21; Conditional acceptance: 8 IV 21; Final acceptance: 31 V 21 Alejandro Martínez–Abraín, Universidade da Coruña, Facultade de Ciencias, Campus da Zapateira s/n, 15008 A Coruña, Spain.– Xavier Ferrer, Department of Evolutionary Biology, Ecology, and Environmental Sciences, Faculty of Biology, Universitat de Barcelona, Spain; Biodiversity Research Institute (IRBIO), Universitat de Barcelona, Av. Diagonal 643, 08028 Barcelona, Spain.– Juan Jiménez, Servicio de Vida Silvestre, Generalitat Valenciana, Ciutat Administrativa 9 d’Octubre, Torre 1, c/ Democracia 77, 46018 Valencia, Spain.– I. C. Fernández–Calvo, Sociedad Española de Ornitología (SEO/BirdLife), Delegación Territorial de Cantabria, Avenida de Chiclana 8, 39610 Astillero, Spain. Corresponding author: A. Martínez–Abraín. E–mail: a.abrain@udc.es

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Introduction One of the components of global change is the increasing urbanization of the landscape (Gil and Brumm, 2014; Murgui and Hedblom, 2017), with likely negative consequences for many animal species. Surprisingly, wildlife is selecting anthropogenic habitats made with no wildlife conservation purposes, even where natural or seminatural habitats of apparent high–quality are available (Martínez–Abraín et al. 2020). An indirect proxy of this trend is the increasing number of papers devoted to the study of the use of anthropogenic habitats or human landscapes by wildlife during the last decade, compared to the general growth of the study of wildlife ecology (fig. 1). The causes behind this phenomenon are not well known. To date, attempts to explain the presence of wildlife in anthropogenic habitats have focused on analysing the main anatomical correlates of species living in urban areas, such as brain size (Sayol et al., 2020) or on analysing the balance between costs and benefits of urban life (Møller and Díaz, 2018a, 2018b). Loss of fear to humans has also been identified as an instrumental feature of species or populations living in urban areas (Guerting et al., 2012; Sih, 2013). We analyse here a diverse set of Iberian case studies of selection of anthropogenic habitats by wildlife with the aim of providing empirical support to the idea that colonization of anthropogenic landscapes is, in first instance, an ecological consequence of the human depopulation of the rural areas in the near past. This process has led to loss of open habitat by shrub and tree encroachment, departure from ecological refuges, loss of fear to humans and importantly to mesopredator release, due to lack of direct persecution of wildlife by humans (Martínez–Abraín et al., 2009, 2019, 2020). According to Flannery (2018), in Europe humans have been substituting the role of lost Pleistocene top–predators (including lions, hyenas, leopards, bears and saber–toothed cats) during the last 14,000 years (long before the advent of agriculture) and hence the current lack of persecution of intermediate predators by humans is expected to have profound consequences on ecosystems. Already Møller and Ibáñez–Álamo (2012) showed that predation avoidance is related with colonization of urban environments. More recently, Samia et al. (2017) and Jokimäki et al. (2020) also revisited the link between urbanization and lower predation risk. Selection of anthropogenic habitats: case– studies Transportation infrastructure Wild rabbits (Oryctolagus cuniculus) are a keystone prey in Mediterranean ecosystems. Conservation of some charismatic threatened predators in Spain, such as the Iberian lynx (Lynx pardinus) and the Spanish imperial eagle (Aquila adalberti), depend on the availability of healthy rabbit populations (Delibes–Mateos et al., 2007, 2008). However, wild rabbits have been

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jeopardized by infectious diseases. They have also experienced habitat loss due to shrub and tree encroachment in open agricultural land, after rural exodus started in the country six decades ago. Many attempts to restock rabbit populations have been implemented but have failed to reverse the decline in natural areas (Carro et al., 2019). On the contrary, wild rabbits thrive in some anthropic areas, building communal dens in apparently low–quality places such as the banks of roads and railways or the median strips of highways (Planillo and Malo, 2017), even inside large cities such as Madrid. Two explanations are possible for this puzzling situation: a) fenced highways and high–speed railways provide protection against predators, and b) road and highway verges create open areas where grass grows. Indeed, a lot of energy is put annually into maintaining communication infrastructure verges free of tall vegetation, and this routine is exploited by rabbits to obtain food. Human services for the maintenance of roads and highways (often paradoxically called 'conservation services', Martínez–Abraín, 2019) could have great conservation value within a landscape matrix in which open land is increasingly scarce due to shrub and tree encroachment. This is not only true for vertebrates (rabbits, corvids, birds of prey) (Dean and Milton, 2009), but also for plants and invertebrates (Jakobsson et al., 2018). Another animal species commonly associated with road verges in Spain is the white stork (Ciconia ciconia). It selects the metal poles of traffic signals and electronic signboards within highways to place its nests (sometimes close to traffic) in spite of the abundance of trees in the surroundings. Causes behind this seemingly poor habitat selection are most likely related to avoidance of predators. Scavengers are another animal group that is clearly favoured by communication infrastructures. For example, both black and red kites (Milvus migrans and M. milvus), and sometimes griffon vultures (Gyps fulvus), as well as many corvid species, take advantage of roads, highways and railways and their associated vehicles that act as novel predators, providing carcasses (Morelli et al., 2014). This is not free of risk as scavenging birds are sometimes accidentally run over (Husby, 2016). Airports and other human exclusion zones Bird strikes are a growing threat to the safety of aviation nowadays. Great efforts are devoted to creating bird–free areas around major airports (DeVault et al., 2013; Pfeiffer et al., 2020). Paradoxically, airports themselves may generate good conditions for the occurrence of some wildlife species, especially grassland birds (Blackwell et al., 2013). Airports, being fenced open areas, with low densities of potential predators, and highly controlled human presence, favour the establishment of ground–nesting birds. In the case of Spanish airports, the list of bird species present therein includes some rare species such as little bustards (Tetrax tetrax), stone curlew (Burhinus oedicnemus), along with 50 others (Zugasti, 2004). Predator and human exclusion also fosters the establishment of dispersing and migrant birds (Milne, 2006), as well as the presence of bird roosts at airports. Unintended

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positive effects of human exclusion by fencing also occur in other anthropogenic habitats such as military training camps. One remarkable case is that of recolonizing wolves in Germany. Their success is due to the use of army land property, in a country with a human population density high (233 inhabitants/km2) (Reinhardt et al., 2019). Urban areas under construction Ground–nesting waterbird species are associated with the initial stages of ecological succession in wetlands and dune fields for nesting. They need ample visibility and low vegetation cover to avoid predators (Gómez–Serrano and López–López, 2014, 2016). Once vegetation grows, they need to move elsewhere for breeding, favouring the evolution of a nomadic behaviour (Martínez–Abraín et al., 2003). The occurrence of herbivory in the past (by domestic livestock or extirpated large wild ungulates) likely favoured the long–term occupancy of breeding habitats by ground–nesting birds. This is the case for little terns (Sterna albifrons) or Audouin’s gulls (Ichthyaetus audouinii), originally linked to river deltas, dune fields, beaches and salt marshes. Surprisingly, in recent years these species have colonized port docks, during their construction and even fenced industrial areas (fig. 1s in supplementary material). Nowadays, 40 % of the world population of Audouin’s gull breeds in ports and harbours, whilst populations of Audouin’s gulls from small Mediterranean islets have declined rapidly (Martínez–Abraín and Jiménez, 2016; Oro, 2020). These open anthropic habitats may functionally resemble newly formed land–bridge islands. Just like their original unstable habitats, areas under construction are temporary, and hence ground–nesting birds can only use them over a limited time. Another example is that of little ringed plovers (Charadrius dubius) that have been reported to nest successfully in parking lots, landfills and plots of land during the initial phases of urban development (Fernández–Calvo and González–Sánchez, 2008) (fig. 1s in supplementary material). Paradoxically, plovers performed much worse in some sites specifically restored for the species by conservation practitioners than in highly modified areas (Fernández–Calvo and González–Sánchez, 2008). Plovers were able to occupy human–made sites during several seasons as long as that they had a base made of cement, concrete, stone or asphalt, preventing shrub and tree encroachment. If the base substrate was softer and plants colonized the area, plovers rapidly deserted the site. Little ringed plovers may also use flat roofs for nesting (Baumann, 2006). Roofs are readily used by waterbirds as substitutes of original habitat to nest when the available habitat is of lower quality (e.g. there is a high risk of nest predation). In fact, several species of terns use flat roofs for nesting worldwide (Fisk, 1978; Fernández– Canero and González–Redondo, 2010). In Europe, there are records of common terns (S. hirundo) nesting on roofs in Finland, Estonia, United Kingdom, Ireland, Netherlands and France (Source: https://www.birdlife. org/europe–and–central–asia).

Cities Cities have become excellent foraging and breeding grounds for many bird species. The long list of species now making use of cities in Europe cannot be fully approached in our analysis. We cite here only the paradigmatic case of the peregrine falcon (Falco peregrinus) that started colonizing cities as early as the 19th century (Ferrer, 2016) and is now present in a large number of large towns and cities worldwide (see table 1s in supplementary material for a summary of major cities used in Spain). Some threatened bird species are now only found in cities, as is the case of the Hispaniolan amazon (Amazona ventralis) and the Hispaniolan parakeet (Psittacara chloropterus) in the Dominican Republic (Luna et al., 2018), both of which have benefited from the non– aggressive attitude of today's city dwellers, contrary to their intense persecution in rural areas. Importantly the first Special Protection Areas created in urban centres in Spain were declared in November 2020 for the protection of lesser kestrels (Falco naumanni). Reservoirs Reservoirs flood vast areas and interrupt the natural flow of rivers. Hence they are a threat for biodiversity conservation. However, reservoirs can also provide unintended benefits for many animal species. This delicate balance leans on the side of conservation when they are built on land that has low value for terrestrial fauna or/and where natural wetlands are scarce or absent. Some 880 reservoirs (functionally equivalent to lakes) were built during the 20th century throughout Spain, where natural lakes are rare. The population expansion of some formerly scarce species in the Iberian Peninsula, such as great crested grebes (Podiceps cristatus) and Eurasian otters (Lutra lutra) is linked to some extent to reservoirs (see e.g. Llinares et al., 2019). For example, in the Mediterranean river basins of Spain, reservoirs with signs of otter presence increased from 32 % in 1994–1996 to 77 % in 2015–2016, whereas the confirmed presence in rivers was 59 % and 53 % respectively, with a similar sampling effort in both periods (see sources in Martínez–Abraín and Jiménez, 2016). Likewise, the occurrence of Great crested grebes breeding in the Comunidad Valenciana (Eastern Spain) has greatly increased in inland reservoirs over the last few decades. In 1984, no grebes were known to nest on inland reservoirs (the bulk of the population was present at natural coastal wetlands), but in 2017 ca. 30 % of the breeding pairs in the region were reservoir birds (Source: unpublished data; http:// www.agroambient.gva.es/es/web/biodiversidad). Large artificial irrigation ponds located in semi–arid regions of Spain are known to play a relevant substitutive role for many aquatic bird species from coastal wetlands (see e.g. Sánchez–Zapata et al., 2005) as do major reservoirs. Osprey (Pandion haliaetus) have also discovered the advantages of this novel habitat as a food provider (Casado and Ferrer, 2005). Counts of wintering osprey in Spain between 1984 and 1996 showed that out of 522 individuals, 49 (9 %) were detected in reservoirs

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800 Number of papers

2,500

Urban + wildlife Human landscape + wildlife Anthropogenic habitat + wildlife Wildlife + ecology

2,000

700

Number of papers

900

199

1,500

600

1,000

500

400 300 2019

2018

2017

2016

2015

2014

2013

2012

2011

Fig. 1. Number of papers detected by a literature search in the Web of Science (2010–2020) under the headings 'urban + wildlife', 'human landscape + wildlife' and 'anthropogenic habitat + wildlife', compared to a control search under 'wildlife + ecology'. Fig. 1. Número de artículos encontrados en una búsqueda bibliográfica en la Web of Science (2010-2020) con los temas "fauna silvestre + urbano", "paisaje humano + fauna silvestre" y "hábitat antropogénico + vida silvestre" en comparación con una búsqueda de control de "vida silvestre + ecología".

(Migres Foundation, unpublished information). More recent data (winter 2016–17) suggest an increase in the use of reservoirs by wintering osprey of up to 17 % (Martín et al., 2019). Quarries Quarries are intrinsically associated with the destruction of habitat as rock or sand extraction is a consumptive activity. However, quarries can bring unintended benefits for wildlife as they generate artificial cliffs that can be used by obligate and facultative cliff–nesting birds in areas where cliffs are absent, scarce, have low quality or have become saturated. This is, for example,the case of ravens (Corvus corax), a species that used to nest in coastal cliffs in western France, most likely due to human persecution during the second half of the 20th century. But starting in the 1970s, ravens began to leave their coastal ecological refuges and colonize mainland quarries in the region, so that by 2003 ca. 45% of the population was nesting in quarries (Quelennec, 2004). In northern Spain it was observed that 73 % of abandoned quarries and 39 % of active quarries (n = 73) were occupied by birds of 12 species, including corvids and diurnal (three species) and nocturnal raptors (four species) (Castillo et al., 2008). Interestingly, the authors reported that several projects addressing the environmental restoration of quarries had negative consequences for wildlife. Limestone quarries in the Mediterranean side of the Iberian Peninsula are readily colonized

by cliff–nesting songbirds such as black wheatear Oenanthe leucura and blue rock thrushes Monticola solitarius. These species currently have a declining trend in NE Spain due to the abandonment of traditional rural practices and shrub and tree encroachment (Prodon, 2020). However, the proliferation of quarries in recent decades has created suitable habitat and most territories are now located in coastal quarries (Noguera et al., 2014). Sand quarries provide good substitution habitats for the collared sand martin (Riparia riparia) and European bee–eaters (Merops apiaster). For example, out of 132 collared sand martin colonies detected at Comunidad Valenciana during 2010–2018, 51 % were found in sand quarries, 21 % in excavations for construction purposes and 20 % in artificial walls. Only 8% of the colonies were located in natural substrates. Interestingly, all collared sand martin colonies in sand quarries were located in active quarries (Servicio de Vida Silvestre, 2019) suggesting predator avoidance as a major driver of selection besides the scarcity of suitable habitat. Landfills and scavenging wildlife Landfills are increasingly used by many species to obtain food (Oro et al., 2013; Meffert, 2017). Small landfills were traditionally used in Spanish rural areas by scavenging raptors, corvids and canids (foxes and wolves), whereas large city landfills started being used in mass by gulls and other bird species from the 1970s on. The four major landfills around Madrid

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Table 1. Summary of examples analysed in the text of selection of anthropogenic habitats by wildlife and likely main causes of habitat selection. Tabla 1. Resumen de los ejemplos analizados en el texto de selección de hábitats antropógenos por la fauna silvestre y principales causas probables de la selección del hábitat.

Type of anthropogenic habitat

Species affected

Benefits for wildlife

Transportation infrastructure

Oryctolagus, Ciconia,

Anti–predatory defense

Milvus, Gyps, Corvids

Food

Airports and other

Tetrax, Burhinus

Anti–predatory defense

Urban areas under construction

Sterna spp., Larus spp.,

Anti–predatory defense

Charadrius

Cities

Falco, Psittacara, Amazona Food

human–exclusion areas

Defense from rural people

Anti–predatory defense

Reservoirs

Lutra, Pandion, aquatic birds

Food

Quarries under exploitation

Corvus, Oenanthe, Monticola,

Anti–predatory defense

Riparia, Merops

Landfills

Larus, Bubulcus, Gyps,

Aegypius, Neophron

Food

city are used on a daily basis during the winter by up to 65,000 black–headed gulls (Larus ridibundus) (Del Moral et al., 2002). Over the years they have been discovered as food sources by other species and now represent a major food source for black and red kites, white storks, cattle egrets (Bubulcus ibis) and even griffon (Gyps fulvus) and black vultures (Aegypius monachus) (fig. 2s in supplementary material). The Spanish population of white storks grew from ca. 7,600 pairs in 1995 to ca. 31,000 pairs in 2004, to a large extent, due to the use of landfills as foraging grounds (Molina and Del Moral, 2005). White storks may nest inside and around landfills on poles, constructions and antennas, what might be seen as an optimal cost/benefit situation (fig. 2s in supplementary material). Landfills and reservoirs are also connected by daily activity of birds. For example, more than 10,000 lesser black–backed gulls (Larus fuscus) overnight at the Santillana reservoir around Madrid in winter (Del Moral et al., 2002). Selection of anthropogenic habitat as a two–step process Specifically, the success of ground–nesting bird species in areas in process of urban development may result from a strategy to avoid high levels of predation in the countryside after the rebound of terrestrial mesopredators and raptors (Díaz et al., 2013) and/or scarcity

of high–quality open habitats due to shrub and tree encroachment. This is most likely also the case of birds linked to sand quarries, although those linked to quarries under exploitation are also benefiting from predator avoidance. Cases of successful use of reservoirs may reflect the movement of some species out of their ecological refuges, where they had been secluded due to human persecution and alteration of their habitat, and the discovery of reservoirs as food–rich habitats (i.e. reservoirs are full of exotic crustaceans, molluscs and fish species that become a new resource when discovered by native species). Examples of landfills as food sources for scavenging species most likely reflect the scarcity of carcasses of livestock and wild ungulates. Even rare vulture species in Europe, such as the black vulture and the Egyptian vulture (Neophron percnopterus) are known to make extensive use of landfills when food from natural sources is scarce or absent (Gangoso et al., 2012; Martínez–Abraín et al., 2012; Tauler–Ametller et al., 2017). A network of small landfills scattered in the territory (imitating more closely the unpredictable distribution of carrion in nature) could be a good transition measure along the road towards closing major open–air landfills as promoted by the environmental policy of the European Union (Cortés– Avizanda et al., 2010). The ultimate causes of all case studies analysed are consistent with our hypothesis linking human depopulation of rural areas with colonization of urban areas by wildlife (Martínez–Abraín et al., 2019, 2020).

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Most of the examples analysed here come from Spain where rural exodus started relatively late within the European context, some six decades ago. Its effects are becoming evident only now, after a non–linear period of gestation. During this multidecadal period, human depopulation of rural areas has radically altered a status quo that had not experienced substantial changes over centuries or even millennia. Consequences of these changes include the movement of wildlife out of the historical refuges to which they were forced by rural human activities, loss of fear to humans, the growth of mesopredator populations by mesopredator release (where rural humans were the top predator) and of large mammalian herbivores, as well as shrub and tree encroachment, as advanced in our introduction. Importantly, all these ecological factors would not have had any practical effect if it were not for the changed human attitudes of modern urban people who do not perceive wildlife as competitors or enemies and cease to persecute wildlife, a remarkable historical landmark (Martínez–Abraín et al., 2008, 2009, 2019). Once in anthropogenic environments urban people unintentionally further protect wildlife against predation by means of the scarecrow effect (Leighton et al., 2010). Moreover, trophic opportunities are multiplied due to the role of humans as managers of large amounts of exo–somatic energy that translate in large quantities of discarded surplus food (Oro et al., 2013). Although bold individuals (with less fear and more exploratory momentum) are known to be more prone to colonizing anthropogenic areas (Díaz et al., 2013; Riyahi et al., 2015), these habitats in turn select for fearless individuals generating a positive feed–back loop (Miranda, 2017). We think that the loss of fear to humans is most likely the most relevant condition allowing approach and close coexistence with humans in human landscapes. A condition that most likely is necessary, although not sufficient. What ornithologists have traditionally labelled as 'urban birds' for many decades are most likely those species that are more prone to losing fear to humans along a gradient. Time has shown than many other species, formerly not considered as 'urban', may also colonize urban areas given the right conditions. In summary, we suggest that colonization of urban environments can be seen as a sequential two–step process. The first step occurs outside the urban environment and is triggered as an ecological consequence of the human depopulation of rural areas (Martínez–Abraín et al., 2020). The second step takes place inside urban landscapes as the result of a balance of costs and benefits of urban life (Møller and Díaz, 2018a, 2018b). It is encouraging to know that bird assemblages in highly urbanized environments are now only 20 % less functional than those in surrounding natural habitats (Sol et al., 2020). If this happens with the current low levels of interest in finding technical solutions to promote the role of anthropogenic landscapes as biodiversity–rich areas this means that there is a lot of room for future improvement. Obviously the approach of wildlife to urban areas will also create new fronts for human–wildlife conflict (see e.g. Barrett et al., 2019).

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Acknowledgements We are grateful to Daniel Oro and Pilar Santidrián who commented on a draft of the manuscript. Mario Díaz and an anonymous referee improved the content of the final version. AMA was supported by project ED431C 2018/57 funded by Xunta de Galicia. XF thanks the financial support of Fundació Barcelona Zoo. References Baumann, N., 2006. Ground–nesting birds on green roofs in Switzerland: Preliminary observations. Urban Habitats, 4: 37–50. Barrett, L. P., Stanton, L. A., Benson–Amram, S., 2019. The cognition of nuisance’ species. Animal Behaviour, 147: 167–177, Doi: 10.1016/j.anbehav.2018.05.005 Blackwell, B. F., Seamans, T. W., Schmidt, P. M., DeVault, T. L., Belant, J. L., Whittingham, M. J., Martin, J. A., Fernandez–Juricic, E., 2013. A framework for managing airport grasslands and birds amidst conflicting priorities. Ibis, 155: 199–203. Carro, F., Ortega, M., Soriguer, R. C., 2019. Is restocking a useful tool for increasing rabbit densities? Global Ecology and Conservation, 17: e00560, Doi: 10.1016/j.gecco.2019.e00560 Casado, E., Ferrer, M., 2005. Analysis of reservoir selection by wintering Ospreys (Pandion haliaetus haliaetus) in Andalusia, Spain: a potential tool for reintroduction. Journal of Raptor Research, 39: 168–173. Castillo, I., Elorriaga, J., Zuberogoitia, I., Azkona, A., Hidalgo, S., Astorkia, L., Iraeta, A., Ruiz, F., 2008. Importancia de las canteras sobre las aves rupícolas y problemas derivados de su gestión. Ardeola, 35: 103–110. Cortes–Avizanda, A., Carrete, M., Donázar, J. A., 2010. Managing supplementary feeding for avian scavengers: guidelines for optimal design using ecological criteria. Biological Conservation, 143: 1707–1715, Doi: 10.1016/j.biocon.2010.04.016 Dean, W. R. J., Milton, S. J., 2009. The importance of roads and road verges for raptors and crows in the Succulent and Nama–Karoo, South Africa. Ostrich, 74: 181–186, Doi: 10.2989/00306520309485391 Delibes–Mateos, M., Delibes, M., Ferreras, P., Villafuerte, R., 2008. Key role of European rabbits in the conservation of the western Mediterranean basin hotspot. Conservation Biology, 22: 1106–1117, Doi: 10.1111/j.1523-1739.2008.00993.x Delibes–Mateos, M, Redpath, S. M., Angulo, E., Ferreras, P., Villafuerte, R., 2007. Rabbits as keystone species in southern Europe. Biological Conservation, 137: 149–156, Doi: 10.1016/j.biocon.2007.01.024 Del Moral, J. C., Molina, B., De la Puente, J., Pérez– Tris, J. (Eds.), 2002. Atlas de aves invernantes de Madrid 1999–2001. SEO–Monticola, Madrid, Spain. DeVault, T. L., Blackwell, B. F., Belant, J. L. (Eds.), 2013. Wildlife in airport environments: preventing animal–aircraft collisions through science–based

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Reinhardt, I., Kluth, G., Nowak, C., Szentiks, C.A., Krone, O., Ansorge, H., Mueller, T., 2019. Military training areas facilitate the recolonization of wolves in Germany. Conservation Letters, 12: e12635, Doi: 10.1111/conl.12635 Riyahi, S., Sánchez–Delgado, M., Calafell, F., Monk, D., Senar, J. C., 2015. Combined epigenetic and intraspecific variation of the DRD4 and SERT genes influence novelty seeking behavior in great tit Parus major. Epigenetics, 106: 516–525. Samia, D.S.M., Blumstein, D.T., Díaz, M., Grim, T., Ibáñez–Álamo, J.D., Jokimäki, J.,Tätte, K., Markó, G., Tryjanowski, P., Møller, A. P., 2017. Rural–urban differences in escape behavior of European birds across a latitudinal gradient. Frontiers in Ecology and Evolution, 5: 66, Doi: 10.3389/ fevo.2017.00066 Sánchez–Zapata, J. A., Anadón, J. D., Carrete, M., Giménez, A., Navarro, J., Villacorta, C., Botella, F., 2005. Breeding waterbirds in relation to artificial pond attributes: implications for the design of irrigation facilities. Biodiversity and Conservation,14: 1627–1639, Doi: 10.1007/s10531-004-0534-1 Sayol, F., Sol, D., Pigot, A. L., 2020. Brain size and life history interact to predict urban tolerance by birds. Frontiers in Ecology and Evolution, 8: 58, Doi: 10.3389/fevo.2020.00058 Servicio de Vida Silvestre, 2019. Censo de colonias y parejas reproductoras de avión zapador (Riparia riparia) en la Comunidad Valenciana. Conselleria de Agricultura, Desarrollo Rural, Emergencia Climática y Transición Ecológica. Generalitat Valenciana. November, 2019. Unpublished report. Sih, A., 2013. Understanding variation in behavioural responses to human–induced rapid environmental change: a conceptual overview. Animal Behaviour, 85: 1077–1088. Sol, D., Trisos, C., Múrria, C., Jeliazkov, A., González–Lagos, C., Pigot, A. L., Ricotta, C., Swan, C. M., Tobias, J. A., Pavione, S., 2020. The worldwide impact of urbanisation on avian functional diversity. Ecology Letters, 23(6): 962–972, Doi: 10.1111/ ele.13495 Tauler–Ametler, H., Hernández–Matías, A., Pretus, J. L. L., Real, J., 2017. Landfills determine the distribution of an expanding breeding population of the Endangered Egyptian vulture Neophron percnopterus. Ibis, 159: 757–768. Zugasti, M., 2004. Guía para la identificación de las aves más frecuentes en aeropuertos españoles. AENA, Madrid.

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Can scientific laws be discussed on philosophical grounds? A reply to naïve arguments on 'predators' proposed by Bramble (2021) A. Cordero–Rivera, R. Roucourt Cezário, R. Guillermo–Ferreira, V. Marques Lopez, I. Sanmartín–Villar Cordero–Rivera, A., Roucourt Cezário, R., Guillermo–Ferreira, R., Marques Lopez, V., Sanmartín–Villar, I., 2021. Can scientific laws be discussed on philosophical grounds? a reply to naïve arguments on 'predators' proposed by Bramble (2021). Animal Biodiversity and Conservation, 44.2: 205–211, Doi: https://doi. org/10.32800/abc.2021.44.0205 Abstract Can scientific laws be discussed on philosophical grounds? a reply to naïve arguments on 'predators' proposed by Bramble (2021). A recent paper by Bramble (2021) argues that given that predators inflict pain and fear on their prey we have the moral right to act to minimize these effects. The author proposes two alternatives. The first is to transform predators by 'genetically modifying them so that their offspring gradually evolve into herbivores'. The second is simply 'painlessly killing predators', which is the title of Bramble's essay. We address the misconceptions that Bramble uses as central in his arguments and present scientific reasoning to discuss the ethical implications of disregarding scientific knowledge when addressing animal welfare and animal rights. We conclude that both Bramble's alternatives are nonsensical, not only from a scientific point of view, but also, and more importantly, from ethical grounds. Key words: Animal behaviour, Predation, Environmental ethics, Philosophy, Scientific law Resumen ¿Se puede mantener un debate filosófico sobre las leyes de la ciencia? una respuesta a los ingenuos argumentos sobre "depredadores" propuestos por Bramble (2021). En un reciente artículo, Bramble (2021) sostiene que, dado que los depredadores infligen dolor y miedo a sus presas, tenemos el derecho moral de actuar para minimizar estos efectos, y propone dos alternativas. La primera es transformar a los depredadores "modificándolos genéticamente para que sus descendientes se conviertan gradualmente en herbívoros". La segunda es simplemente "matar a los depredadores sin dolor", que es el título del ensayo de Bramble. Aquí abordamos los conceptos erróneos utilizados por Bramble y que son centrales en sus argumentos y presentamos un razonamiento científico para analizar las implicaciones éticas de ignorar el conocimiento científico al abordar el bienestar y los derechos de los animales. Concluimos que las dos alternativas de Bramble carecen de sentido, no solo desde un punto de vista científico, sino sobre todo, desde el punto de vista ético. Palabras clave: Comportamiento animal, Depredación, Ética ambiental, Filosofía, Ley de la ciencia Received: 1 III 21; Conditional acceptance: 10 V 21; Final acceptance: 7 VI 21 Adolfo Cordero–Rivera, Iago Sanmartín–Villar, Universidade de Vigo, ECOEVO Lab., E. E. Forestal, Campus A Xunqueira, 36005 Pontevedra, Galiza, Spain.– Rodrigo Roucourt Cezário, Rhainer Guillermo–Ferreira, Vinicius Marques Lopez, Universidade de São Paulo, PPG Entomologia, FFCLRP, Ribeirão Preto, SP, Brazil and Universidade Federal de São Carlos, LESTES Lab, São Carlos, SP, Brazil. Corresponding authorr: Adolfo Cordero–Rivera. E–mail: adolfo.cordero@uvigo.gal ORCID ID: 0000-0002-5087-3550

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Introduction The naïve concept of predator Studies performed in all kinds of ecosystems show that life is organised into trophic webs, with matter and energy transfers from producers to herbivores, then to first–level carnivores, and so on. 'Predation' is a general concept which includes several types of related interactions characterised by the predator’s negative effect on the prey, and the fact that the predator attacks a prey which is alive (Begon et al., 2006). It includes not only 'Attenborough's ferocious beasts' (the only animals Bramble (2021) considered as predators, fig. 1A), but also, for example, animals that eat seeds (seed predation) or eggs (egg predation) that kill 'would–be' organisms, and animals like mosquitoes, leeches, and herbivores, that eat only parts of their prey (fig. 1B). Parasitoids, insects that develop inside other insects, kill the host to complete their development and are also predators (Begon et al., 2006). Bramble (2021) restricts his definition of 'predator' arbitrarily ('ferocious beasts'), simply to serve his argument: "A vital question here is which animals count as predators for the purposes of this argument? Do insect–eating birds count? This should depend on the mental lives of insects. If insects aren’t capable of lives worth living, there would be no reason to prevent birds eating them. For what it's worth, I'm not here thinking of insect–eating birds as predators in the relevant sense." [footnote 7; the emphasis is ours] Let us examine this argument from current scientific knowledge of the 'mental lives' of insects. Can insects feel pain or fear? They can, even if the debate about this topic is complex and therefore out of the scope of this essay (contrasting views are presented for instance in Adamo (2016) and Tiffin (2016)). For instance, dragonfly larvae are known to die of fear after being exposed to chemicals released by their predators (McCauley et al., 2011). This can be considered a type of pain. The mental lives of insects allow them to count up to four (Dacke and Srinivasan, 2008), and honeybees are even able to recognize the concept of zero (Howard et al., 2018). Insects have complex personalities (Schuett et al., 2011); they can learn complex tasks (Dukas, 2008); social insects play among nestmates (Weber, 2014, p. 135) and ants use tools (Maák et al., 2017). Therefore, we conclude that insects are prey, and birds eating insects are predators. More importantly, consider the opposite case: insects eating birds (fig. 1C). There are hundreds of observations of small hummingbirds falling prey of large insects like mantises (Nyffeler et al., 2017) or other invertebrates such as spiders (Brooks, 2012). Mantises can even make a hole in the victim’s head through which the brain is extracted, but if insects do not qualify as prey because 'aren't capable of lives

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worth living', they surely cannot be also considered predators, right? However, the concepts of predator and prey are indissoluble. If the birds eaten by mantises are considered prey, then, the mantises themselves are predators. Let us now turn our attention to herbivores (fig.1B), because in Bramble's (2021) arguments they are clearly not predators given his proposal to 'herbivorise' predators. We suppose that the absence of 'mental lives' in plants would be the basic argument to consider that eating plants is 'good' whereas eating animals (except for insects) is 'bad'. But do plants feel 'pain' when they are attacked? The scientific answer is, again, yes. Trees attacked by herbivores release chemicals, which are detected by nearby trees, and trigger an anticipated response to a predictable attack. This explains why giraffes feeding on Acacia trees in Africa (fig. 1B) have to move to trees situated over 90 m away from the first attacked trees which produce chemical defences to avoid being eaten (Wohlleben, 2016). Plants also 'call for help' when attacked. Chemical compounds are released when insect herbivores eat leaves, attracting parasitoids that attack the herbivore, and indirectly help the plant. Herbivores therefore also qualify as predators. And what about plants? Surely they are the quintessence of 'prey'. They have no nervous system, they do not move and they do not prey on other organisms. In fact, as with all naïve generalizations, this is wrong. Chemicals flow within the phloem and xylem, the vessels of vascular plants, acting as a nervous system (Muday and Brown–Harding, 2018); plants move, but in a time scale different to ours. An extreme case is the so–called 'walking palm', which is able to relocate itself away from its germination point (Bodley and Benson, 1980). Recent research has shown that mycorrhizal fungal networks linking the roots of trees in forests allow abilities such as perception, learning and memory in trees, and have a topology similar to neural networks (Simard, 2018). Some plants capture insects (fig. 1D) and some 'eat' them. Even Darwin (1888) was fascinated by these insectivorous plants. Should we eliminate them according to Bramble's arguments? The arbitrary definition of 'predator' used by Bramble (2021) is scientifically flawed. Most 'herbivorous' animals do consume other animals, even if in small quantities. For instance, chimpanzees are mainly fruit and leaf–eaters, but they occasionally hunt and eat small monkeys and other animals, using highly elaborate hunting behaviours and strategies (Newton–Fisher, 2007). Using Bramble's arguments, we should extirpate chimpanzees, a species that shares a great part of our 'humanity'. Or should we make an exception if they only hunt occasionally? Many other animals are omnivorous, including our own species. Shall we 'painless kill' only those individuals that show carnivorous behaviour? Or is the whole species deemed to be eradicated if some individuals show blood–appetite? The questions to pinpoint the practical terms of Bramble’s claims are endless.

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Fig. 1. Predators are all animals –and plants– that feed on whole or parts of other organisms. This includes the 'ferocious beasts' of nature documentaries like these lions feeding on a zebra in Nairobi National Park (A), but also all herbivores (B) like the giraffe eating Acacia leaves. Bramble (2021) arbitrarily excludes insect–eating birds from his definition of predators, because he naïvely assumes that insects have no worthy 'mental lives'. However, they can act as predators of birds (C), and if they can predate on birds, they cannot be considered as irrelevant when are prey. Lastly, some plants, like this Erica ciliaris (D), have structures that capture insects and kill them slowly. Some plants even digest the insects captured. Pictures by Adolfo Cordero–Rivera (A, B, D) and Darrell Ferriss (C). Fig. 1. Los depredadores son todos los animales (y las plantas) que se alimentan total o parcialmente de otros organismos. Esto incluye las "bestias feroces" de los documentales sobre la naturaleza como estos leones que se alimentan de una cebra en el Parque Nacional de Nairobi (A), pero también todos los herbívoros (B) como la jirafa que come hojas de Acacia. Bramble (2021) excluye arbitrariamente a las aves insectívoras de su definición de depredadores, porque supone ingenuamente que los insectos no tienen "vidas mentales" que valgan la pena. Sin embargo, los insectos pueden actuar como depredadores de aves (C) y si pueden depredar aves, no se pueden considerar tan intrascendentes cuando son presas. Por último, algunas plantas, como esta Erica ciliaris (D), tienen estructuras que capturan insectos y los matan lentamente. Algunas plantas incluso digieren los insectos capturados. Imágenes de Adolfo Cordero–Rivera (A, B, D) y Darrell Ferriss (C).

Evolutionary misconceptions One of Bramble's arguments consists of the gradual herbivorisation of predators through genetic modifications. This is clearly an unjustified claim because individuals exhibit behavioural plasticity (Cordero–Rivera, 2017) to explore different foraging habits and food

items, making it impossible to genetically control what an animal will eat. Bramble (2021) apparently ignores that the phenotype evolves in an integrated way. For example, the components of the vertebrate jaw (whose structure is different in herbivorous and carnivorous mammals) are influenced by at least 33 quantitative trait loci (Klingenberg et al., 2004). Artificial selection

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on animal breeds produces many undesirable side– effects (Negro et al., 2021), illustrating the fact that deep changes like those suggested by Bramble are likely to modify most other biological attributes of the selected animals. That is, any attempt to alter this complex network of genes can lead to unexpected effects. And a change in diet needs a change in the structure of the digestive system, for example. Thus, changing the genetics of predators until they evolve into herbivores is a reductionist view of both genetic modification and the evolutionary process. Changing genetic traits of a species in such a large intervention will alter species characteristics in a way that it will result in a completely different species, and consequent new evolutionary and ecological processes. Given that the evolutionary process will not stop after our intervention in nature, even if we were able to 'herbivorise predators', natural selection would favour the evolution of new predators. Bramble (2021) surprisingly assumes that the only new species evolving would be herbivores: 'New species of herbivores might emerge without predators there to immediately cut them down'. Evolution cannot be stopped at our will. Scientific arguments Although there is an endless discussion about what a scientific law is, one common definition is the view that laws are universal statements that are so well corroborated that everyone accepts them as the basis of scientific knowledge (Krebs, 2000). In other words, each scientific paradigm is grounded on basic principles that form a coherent explanation of the field. Some eminent scientists have argued that there are no laws in Ecology (Lawton, 1999). Krebs (2000) explicitly indicates: 'there are laws in physics, chemistry, and genetics but not in ecology.' However, as Murray (2000) indicates, the theory of evolution and the dynamics of populations offer clear ecological laws. In fact, Krebs (2016) changed his mind and more recently has written: "The generalization that populations cannot increase without limits could be called a law of ecology and is a simple recognition that the Earth is finite." This is a basic ecological law, apparently ignored in Bramble's essay. Given the finite nature of resources, no population can increase endlessly, and predators are precisely one of the elements that form part of this law. Bramble (2021) attempts to revive or to elaborate the ideas propagated by McMahan, who published a similar essay in the New York Times (McMahan, 2010). McMahan's ideas received a lot of attention at the time and a great deal of replies. Likewise, Bramble's essay has spread in the social media, causing considerable turmoil among ecologists, conservationists, and defenders of animal rights. McMahan' and Bramble's arguments are absurd because they ignore the effects

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of species extinctions. Such arguments can be made about the extinction of one species of predator, but do not hold on the extinction of all predators. Let us for a moment assume that we could 'herbivorise or painlessly kill predators', obtaining a new ecosystem with only plants and plant–eating animals (and of course decomposers and detritivores, ecological roles apparently ignored by Bramble and McMahan). Would this be stable? Clearly not. The intricate trophic dynamic equilibria of all ecosystems include several types of predators. To our knowledge, there is not a single case of an ecosystem completely devoid of predators, and even more, we argue that it is theoretically impossible. A basic ecological law predicts that natural selection would always favour predators, given the high rewards in terms of efficiency (and consequently, reproduction) of using complex molecules to feed on, instead of newly assembling them using the energy of light or other sources. In the classic reference 'Origins and early evolution of predation', Stefan Bengtson (2002) comments that: "[…] whenever predatory lifestyles evolved they became a strong evolutionary force. […] Predators and prey may enter into symbiotic relationships and emerge as new organisms. Current theories on a number of major transitions in evolution (non–cellular to cellular; prokaryote to eukaryote; non–sex to sex; small to large; unicellular to multicellular; multicellular to tissue grade; sessile to motile; soft to hard; smooth to spiny) tend to focus on the introduction of predation as a decisive factor." This is true on this planet, but would also hold true in the Universe had life evolved more than once. Bramble (2021) disregards a solid body of contrary evidence to his thesis. For instance, when human activity has eradicated key predators, the effects on ecosystems have been devastating. Viruses, bacteria, and protozoa responsible for various diseases and epidemics use rodent populations to spread and invade humans, and predators can control the population density of the majority of species of zoonotic reservoirs. Therefore, the biotic homogenization advocated by Bramble (2021) can expand the incidence and distribution of infectious diseases affecting humans and increase the risk of novel diseases (Wilkinson et al., 2018). A recent study found that wolf predation can lead to a marked reduction in the prevalence of tuberculosis in wild boar, without leading to a reduction in prey population density (Tanner et al., 2019). Therefore, eliminating wolves would harm their prey, due to increased disease prevalence, with the unavoidable suffering of prey, which is the main argument to eradicate predators! The most common effect of human activities has been labelled as the 'empty forest' syndrome (Redford, 1992) or 'defaunation', which normally targets the largest forest animals, affecting plants when the animals eradicated are mainly frugivorous, which are a clear type of predators despite Bramble's arguments (Bello et al., 2015). In conclusion, the premise of McMahan (2010) that his

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arguments are only valid 'provided that this could occur without ecological upheaval involving more harm than would be prevented by the end of predation' is clearly unjustified. Eliminating predators produces more harm than good, even for herbivores. Ethical arguments In his influential book 'A sand county almanac', Aldo Leopold (1949) develops a 'land ethic', which was the seed for ecocentric ethics (Knight and Riedel, 2002). In a particularly emotive –and science–based– passage, he describes the killing of a wolf: "We reached the old wolf in time to watch a fierce green fire dying in her eyes. […] I thought that because fewer wolves meant more deer, that no wolves would mean hunters’ paradise. But after seeing the green fire die, I sensed that neither the wolf nor the mountain agreed with such a view. Since then I have lived to see state after state extirpate its wolves. […] I now suspect that just as a deer herd lives in mortal fear of its wolves, so does a mountain live in mortal fear of its deer. […] The cowman who cleans his range of wolves does not realize that he is taking over the wolf’s job of trimming the herd to fit the range. He has not learned to think like a mountain. Hence we have dustbowls, and rivers washing the future into the sea." [the emphasis is ours] Leopold (1949) describes in the above passage what is now known in Ecology as 'trophic cascades', i.e., the effect of a trophic level (wolves) on another trophic level (plants) via an intermediate level (ungulates). In Leopold's 'land ethics' mountains have moral rights, they live 'in mortal fear of its deer', and wolves precisely help mountains to maintain a good health. Have we the moral rights to extirpate predators and ignore the consequences? Should we instead learn to 'think like a mountain'? This is not simply a matter of opinion. The consequences are real, as the cases discussed above or the famous reintroduction of wolves to Yellowstone National Park exemplify (Smith et al., 2003). For instance, 'browsing by elk prior to wolf reintroduction had suppressed growth of willows across Yellowstone's Northern Range' and 'largely eliminated cottonwoods from Yellowstone with only a few old trees remaining' (Boyce, 2018), demonstrating the accuracy of Leopold's ethical ideas: wolves are crucial for the survival of trees! Bramble (2021) assumes the right of humans to judge animals by their behaviour. He proposes that humans might change other species before proposing changes in human behaviour (paradoxical, because we humans mostly fed on herbivores). He claims for behavioural censorship in order to obtain a calm environment free of harm, but following his no–harm arguments, he should prohibit all aggressiveness between herbivorous species too (e.g., male–male contests, sexual harassment…), and ultimately, all

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ways to feed on living matter (all predators, including herbivores and humans). We also should eliminate most of our pets, because they are carnivorous and cause extensive harm to wildlife, particularly cats (Marra and Santella, 2016) and dogs. Furthermore, who is predator and who becomes prey is dependent on the context (relative size, random factors, accidents), and therefore we cannot naively propose that predators are intrinsically bad, and that we have the moral duty to control or eliminate them. In his 'Anthropology beyond the human' Kohn (2013) tells a story that people from the Amazonian forests of Peru use to exemplify the intricate relationships among living beings: a jaguar attacked a turtle and became trapped with its canines in the turtle's carapace, being forced to abandon the prey and the teeth. Toothless, the jaguar died from starvation, and the turtle fed on the carrion of its former predator. Matter and energy circulate from one living being to another and to the soil. In his concluding essay 'The land ethic', Leopold (1949) gives clear advice against 'painless killing of predators' with these words: "A thing is right when it tends to preserve the integrity, stability, and beauty of the biotic community. It is wrong when it tends otherwise. […] Conservation is paved with good intentions which prove to be futile, or even dangerous, because they are devoid of critical understanding either of the land, or of economic land–use." [the emphasis is ours] At first glance, the ideas of Bramble (2021) appear to be compassionate and, even morally superior. However, the compassion proposed by Bramble is selective, and this poses many problems. What about carnivore plants? What about the suffering they cause to insects trapped and slowly digested? Bramble (2021) elaborates a succession of premises and conclusions that lead to absurdity, creating an unreal premise to manipulate the reader's feelings (appeal to emotion fallacy). Then, the author uses another fallacy called 'appeal to pity' (or 'argumentum ad misericordiam'), for the conclusion to be accepted: "I want to end by asking you to consider how predators themselves might feel about their lives were they somehow to come to understand the true nature of the harms they inflict on prey. Many of these predators, I suspect, would feel deeply sad, or even horrified, at what they are involved in–indeed, at what they are. I could even imagine them forgiving or excusing us for painlessly killing them." Conclusions Bramble (2021) confuses human ethics, a construct built in human societies to support their rules, with the behaviours of animals (naively classified as good and bad). He proposes to kill all the animals that he considers as bad, meaning only vertebrates that

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feed on vertebrates. These animals are described as evil organisms enjoying the sufferance of others. No plants, fungi and unicellular organisms are considered. Other organisms are considered as irrelevant (for instance, insects). This fact triggers hate towards some species according to personally biased criteria. In words of Leopold (1949): "It is only in recent years that we hear the more honest argument that predators are members of the community, and that no special interest has the right to exterminate them for the sake of a benefit, real or fancied, to itself." In conclusion, discussing the possibility to extirpate predators (but arbitrarily only some…) because they are intrinsically bad is a modern version of the medieval age theological discussions about the number of angels that could dance on the head of a pin. Scientific knowledge is a fundamental basis of human culture, which has long separated from philosophy and other social ways of learning (Hirsch Hadorn et al., 2008) but cannot be simply ignored in this discussion. Scientific laws, and the trophic chains are an example of an ecological law, are not susceptible to be derogated or annihilated at our will. We could feel depressed by the fact that gravity exists and therefore we cannot fly. But we cannot eliminate gravity. We can feel sad when a herd of lions captures a zebra (fig. 1A), or when a shark devours a live turtle, but we cannot eliminate predation. We must also recognize that other people might prefer to show their empathy towards starving, sick tigers or wolves, and this compassion would not be morally inferior to that for the species of herbivores. Acknowledgements We acknowledge the useful suggestions made by two anonymous referees on a first draft of this paper. We also thank Darrell Ferriss for making public his extraordinary observation of a gomphid dragonfly (Hagenius brevistylus) preying on a hummingbird (fig. 1C). ACR was supported by a grant from the Spanish Ministry of Science, including ERD funds (PGC2018–096656–B–I00). RRC, RGF and VML thank National Council for Scientific and Technological Development–CNPq (130346/2020–9, proc. 307836/2019–3 and 142299/2020–0). References Adamo, S. A., 2016. Do insects feel pain? A question at the intersection of animal behaviour, philosophy and robotics. Animal Behaviour, 118: 75–79, Doi: 10.1016/j.anbehav.2016.05.005 Begon, M., Townsend, C. R., Harper, J. L., 2006. Ecology. From individuals to ecosystems (4th edition). Blackwell Publishing, Oxford. Bello, C., Galetti, M., Pizo, M. A., Magnago, L. F. S., Rocha, M. F., Lima, R. A. F., Peres, C. A., Ovas-

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Changes in the nocturnal activity of birds during the COVID–19 pandemic lockdown in a neotropical city F. A. Estela, C. E. Sánchez–Sarria, E. Arbeláez–Cortés, D. Ocampo, M. García–Arroyo, A. Perlaza–Gamboa, C. M. Wagner–Wagner, I. MacGregor–Fors Estela, F. A., Sánchez–Sarria, C. E., Arbeláez–Cortés, E., Ocampo, D., García–Arroyo, M., Perlaza–Gamboa, A., Wagner–Wagner, C. M., MacGregor–Fors, I., 2021. Changes in the nocturnal activity of birds during the COVID–19 pandemic lockdown in a neotropical city. Animal Biodiversity and Conservation, 44.2: 213–217, Doi: https://doi.org/10.32800/abc.2021.44.0213 Abstract Changes in the nocturnal activity of birds during the COVID–19 pandemic lockdown in a neotropical city. The COVID–19 lockdown provided the opportunity to measure species biodiversity in urban environments under conditions divergent from regular urban rhythms. For 90 days, including weeks of strict lockdown and the subsequent relaxation of restrictions, we measured the presence and abundance of birds that were active at night at two sites in the city of Cali, Colombia. Our results show that species richness of nocturnal birds decreased 40 % to 58 % during the weeks with more human activity, adding further evidence to the biodiversity responses of the 'anthropause' on urban environments. Key words: Anthropause, Artificial light at night, COVID–19 lockdown, Tropical cities, Urbanization, Urban ecology Resumen Cambios en la actividad nocturna de las aves durante el confinamiento decretado con motivo de la pandemia de la enfermedad por coronavirus (COVID–19) en una ciudad neotropical. El confinamiento decretado con motivo de la pandemia por COVID–19 ofreció la oportunidad de medir la biodiversidad de especies en ambientes urbanos en condiciones diferentes a las del ritmo urbano habitual. Durante 90 días, incluidas varias semanas de estricto confinamiento y la posterior relajación de la restricción de la actividad humana, se midieron la presencia y la abundancia de aves nocturnas en dos sitios de Cali, en Colombia. Los resultados de este estudio muestran una reducción de entre el 40 % y el 58 % de la riqueza de aves nocturnas durante las semanas con mayor actividad humana, lo que suma otro indicio de la respuesta de la biodiversidad ante la "antropopausa" en ambientes urbanos. Palabras clave: Antropopausa, Luz artificial en la noche, Confinamiento debido al COVID–19, Ciudades tropicales, Urbanización, Ecología urbana Received: 26 V 21; Conditional acceptance: 28 V 21; Final acceptance: 14 VI 21 Felipe A. Estela, Camilo E. Sánchez–Sarria, Departamento de Ciencias Naturales y Matemáticas, Pontificia Universidad Javeriana, Cali, Colombia.– Camilo E. Sánchez–Sarria, Michelle García–Arroyo, Ian MacGregor–Fors, Red de Ambiente y Sustentabilidad, Instituto de Ecología, A. C. (INECOL), Xalapa, México.– Enrique Arbeláez–Cortés, Escuela de Biología, Universidad Industrial de Santander, Bucaramanga, Colombia.– David Ocampo, Instituto de Investigación de Recursos Biológicos Alexander von Humboldt, Villa de Leyva, Colombia.– Alejandro Perlaza–Gamboa, Grupo de Investigación Ecología Animal, Departamento de Biología, Universidad del Valle, Cali, Colombia.– Carlos M. Wagner–Wagner, Colombia BirdFair, Cali, Colombia. Current address: M. García–Arroyo, I. MacGregor–Fors, Ecosystems and Environment Research Programme, Faculty of Biological and Environmental Sciences, University of Helsinki, Niemenkatu 73, FI–15140, Lahti, Finland. Corresponding author: I. MacGregor–Fors. E–mail: ian.macgregor@helsinki.fi ISSN: 1578–665 X eISSN: 2014–928 X

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The 'anthropause' (Rutz et al., 2020) produced by the COVID–19 strict lockdowns provided a unique and probably unrepeatable opportunity to measure the effect of pervasive influence of human activities on urban animals. Lockdowns reduced the effect of human–induced pressures such as anthropogenic noise (Bates et al., 2020). The global reduction in activity allowed the study of the possible effects of human activities on natural and transformed ecosystems. In urban ecology, the effects of artificial light at night (ALAN, sensu Gaston, 2018) and noise on urban wildlife populations have long been a subject of study (Fröhlich and Ciach, 2019). However, the effects of human activity have received less attention and have mainly concerned differential wildlife detectability between weekdays and weekends in natural areas, for instance (Nix et al., 2018). Neotropical nocturnal birds (Strigiformes, Caprimulgiformes, Nyctibiiformes) are among the most poorly–studied avian groups (Fröhlich and Ciach, 2019) even though they play key roles in ecosystem functions and their species richness is high (Kettel et al., 2018). Additionally, some diurnal birds, such as seagulls, herons, and swifts, are also active at night (La, 2012). This nocturnal behavior can be opportunistic and rarely permanent, showing that circadian rhythms in birds can vary (Mukhin et al., 2009). Little is yet known about key issues for the conservation of nocturnal bird species, such as their sensitivity to human disturbance and environmental requirements that could impact their behavior, population dynamics, and distribution. The lack of knowledge on this subject is mainly because such study is challenging due to their low density and detectability (Fröhlich and Ciach, 2019). Nevertheless, their presence in highly–transformed environments, including cities and agro–ecosystems, is frequent and common, and with the development of new technologies of passive monitoring, we are learning more about the ecology of nocturnal birds (Marín–Gómez et al., 2020). In the case of diurnal birds that are active at night in urban environments, however, very little is known. Besides, the information available is mainly anecdotal and refers to birds foraging in night lighting (La, 2012). Despite the scarcity of information published to date on the ecology of nocturnal birds in urban environments, several characteristics have been identified as drivers of their presence and distribution (Chace and Walsh, 2006). Low urbanization levels, well–vegetated urban spaces, and high density of prey are the principal factors associated with an increase in the appearance of nocturnal birds (Kettel et al., 2018). Nevertheless, their populations are limited due to the wide spectrum of hazards in urbanized areas, such as collisions with human infrastructure, artificial lighting, noise, disease, and poisoning (Marín–Gómez et al., 2020). Specifically, anthropogenic noise at night may reduce the availability of prey and interfere with communication behavior among individuals (Marín–Gómez et al., 2020). Notwithstanding, recent studies have suggested that ALAN can benefit several nocturnal species by increasing the abundance of potential

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prey and improving their detection (Marín–Gómez et al., 2020). However, disentangling the dynamics of ALAN and anthropogenic noise, and even the consequences of other human activities, remains a challenge as it has been almost impossible to isolate the role of each factor in urbanized ecosystems. In this study, we surveyed active birds at night in the city of Cali (2.5 million people, 137.5 km2), Colombia, at different moments during the COVID–19 lockdown. The lockdown in Colombia involved varying intensities of human activity. During the first six weeks (March 23rd to June 30th, 2020), lockdown was strict. There were mandatory restrictions for the entire population regarding mobility and most economic activity was suspended. Then, for the following six weeks, until June 30th, the restrictions were gradually lifted regarding human mobility. To evaluate the relationship between nocturnal birds and human mobility in Cali during this study period, we used the most complete and detailed dataset of human mobility available during the lockdown (Google, 2020), as a proxy. Human mobility in Cali decreased by 25 % to 60 % during the time of this study, compared to pre–lockdown activity, showing a constant gradual increase (r = 0.98, P < 0.001) throughout the window of time of the study (fig. 1A). Avian surveys were performed at least five times per week in a peri–urban site (site A, 3º 27.103' N 76º 32.738' W) and in an urban site (site B, 3º 25.987' N 76º 32.746' W) from March 23rd to June 30th, 2020. Surveys consisted of point–count repetitions of 10 minutes in a circular fixed–radius (50 m) between 21:30–22:30 h (three hours after sunset) on nights without rain or strong winds. One observer per site (site A: FAE; site B: CMWW) recorded all birds seen or heard during the point–count. To assess the relationship between the recorded nocturnal birds and the time elapsed after the lockdowns started, we performed a GLM per site (Poisson error distribution, link = log) considering the number of nocturnal records as the dependent variable and time (week) as the independent variable, considering the average weekly time at which surveys were conducted as a random factor. Both GLMs were run in R (R Core Team, 2020). After conducting 194 point–count repetitions during the nocturnal surveys, we recorded nine species on 62 occasions (table 1). Five species of the recorded species are commonly classified as diurnal; three of these species were songbirds that mainly consume insects in flight (Tyrannidae), while the other two were foliage gleaners (Troglodytidae, Parulidae). The number of records and species richness of birds during the surveys at site A decreased significantly across time (GLM x2 = 8.79, P = 0.003; fig. 1B). Comparing the first six weeks of strict lockdown with the latter six weeks, we observed that in Site A the decrease in observation was 52 %, and 52 % for species richness. Site B showed a non–significant negative trend (GLM x2 = 10.46, P = 0.142; fig. 1C), with a reduction of records of 40 % and 58 % less for species richness. Although the detectability of birds at night could have changed across the surveyed time, we did not

Animal Biodiversity and Conservation 44.2 (2021)

Data for Cali

Human mobility

–20 –30 –40 –50 –60

6 8 Week

Nocturnal bird records

Site A 8

Species richness

4 2 0

6 8 Week

Site B Nocturnal bird records

directly measure it. Besides the fact that changes in the environment could have shifted avian detectability, it also could be affected by seasonal patterns of activity of the birds according to their life cycle. However, as seasonality is not as marked in a neotropical city like Cali as in other regions, there is not a clearly evident breeding season, and this could reduce potential sources of detection bias. Among the nine species recorded during this study, two of them are some of the most commonly reported birds in Colombian cities: the tropical kingbird (Tyrannus melancholicus) and house wren (Troglodytes aedon). Although the nocturnal vocalization of diurnal birds is not uncommon behavior, it is poorly understood. However, our records suggest that it could be similar to that promoted by ALAN, and where anthropogenic noise seems to be a limiting factor. Further experimental studies are therefore needed to understand this relationship in depth as it could have potentially significant importance in the management of urban systems. The small sample size of this study is a limitation that does not allow us to generalize our results. However, it is of interest that we were able to record some patterns in such an unexpected anthropause scenario. At both studied sites we recorded a higher number of records and higher species richness during the early stages of the lockdown. The fact that more species and more individuals were more active during the initial six weeks suggests that the normal rhythm of the city could represent a constraint for them. Additionally, given the inherent differences between peri–urban and urban sites (MacGregor–Fors, 2010), it was not surprising that the recorded patterns were stronger in the peri–urban site, suggesting that the city outskirts became even more diverse during lockdown than the intra–urban area of the city. It is notable that, during the lockdown in Cali, ALAN in public spaces was constant, with the same intensity, number of lights, and schedule as the previous weeks and after the lockdown. Therefore, the nocturnal birds’ exposure to ALAN during lockdown was comparable to that during 'regular' nights. However, there was a notable reduction in human activity and mobility at the beginning of lockdown, such as less traffic, fewer pedestrians, closed establishments, and general anthropogenic noise, suggesting that the changes in nocturnal bird species’ richness and frequency was enhanced by a quieter and calmer environment. Under usual conditions, it would not be easy to interpret all of the potential associations between noise and ALAN because of many factors. As mentioned above, the anthropause offered a unique opportunity to study calmer cities. Although the effect of urban noise on diurnal bird activity has received attention in the past, a recent study showed that decreases in diurnal bird numbers in noisy urban settings could be related to temporal responses of birds or our inability to detect them in highly noisy conditions (Carral–Murrieta et al., 2020). Also a recently published study from Spain showed a significant increase in bird detectability in the early morning, suggesting a rapid behavioral response of urban birds to novel environmental conditions (Gordo et al, 2021).

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Species richness

2 0

6 8 Week

Fig. 1. Increase in human mobility (A) and relationships between nocturnal bird records at night during COVID–19 lockdown in a peri– urban (B, site A) and in an intra–urban (C, site B) site in Cali, Colombia. Fig. 1. Aumento de la movilidad humana (A) y relaciones entre los registros de aves nocturnas por la noche durante el confinamiento decretado con motivo de pandemia por COVID–19 en dos zonas de Cali, en Colombia: una periurbana (B, sitio A) y una intraurbana (C, sitio B).

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Table 1. Species recorded at night in Cali during COVID–19 lockdown. Tabla 1. Especies registradas por la noche en Cali durante el confinamiento decretado con motivo de la pandemia por COVID–19.

Weeks Site A

Species

1–6 7–12

Site B 1–6 7–12

Nyctibius griseus

3 2

Nycticorax nycticorax

2 5

6 3

Tyto alba

Megascops choliba

5 2

Pyrocephalus rubinus

1 1

Myiozetetes cayanensis

Tyrannus melancholicus

Troglodytes aedon

6 5

Myiothlypis fulvicauda

2 4

In summary, the results of this study suggest that the number of birds present and active at night in a tropical city increased notable during the strict lockdown that took place in Cali. We recognize that this study has limitations of sample size and spatial representation. Furthermore, it was developed under the unique and probably unrepeatable circumstances of the anthropause brought about by the COVID–19 lockdown. According to the aforementioned observations, the reduction in numbers of pedestrians and vehicles as a direct consequence of these changes in human activity appeared to have an effect on the nocturnal activity of several species of birds. These results therefore also add to the quantitative and standardized evidence of the responses of biodiversity to the anthropause in urban environments. We consider that specific findings such as those presented here should be taken into account in future urban planning and conservation strategies. Acknowledgements The authors thank Kendra Hasenick for reviewing the manuscript. References Bates, A. E., Primack, R. B., Moraga, P., Duarte, C. M., 2020. COVID–19 pandemic and associated lockdown as a "Global Human Confinement Experiment" to investigate biodiversity conservation. Biological Conservation, 248: 108665, Doi: 10.1016/j. biocon.2020.108665

Carral–Murrieta, C. O., García–Arroyo, M., Marín–Gómez, O. H., Sosa–López, J. R., MacGregor–Fors, I., 2020. Noisy environments: untangling the role of anthropogenic noise on bird species richness in a Neotropical city. Avian Research, 11: 32, Doi: 10.1186/s40657-020-00218-5 Chace, J. F., Walsh, J. J., 2006. Urban effects on native avifauna: A review. Landscape Urban Planning, 74: 46–69, Doi: 10.1016/J.LANDURBPLAN.2004.08.007 Fröhlich, A., Ciach, M., 2019. Nocturnal noise and habitat homogeneity limit species richness of owls in an urban environment. Environmental Science and Pollution Research, 26: 17284–17291, Doi: 10.1007/s11356-019-05063-8 Gaston, K. J., 2018. Lighting up the nightime. Science, 362: 744–746, Doi: 10.1126/science.aau8226 Google, 2020. COVID–19 Community Mobility Report. www.google.com/covid19/mobility?h=es Gordo, O., Brotons, L., Herrando, S., Gargallo, G., 2021. Rapid behavioural response of urban birds to COVID–19 lockdown. Proceedings of the Royal Society B, 288: 0251320202513, Doi: 10.1098/ rspb.2020.2513 Kettel, E. F., Gentle, L. K., Quinn, J. L., Yarnell, R. W., 2018. The breeding performance of raptors in urban landscapes: a review and meta–analysis. Journal of Ornithology, 159: 1–18, Doi: 10.1007/ s10336-017-1497-9 La, V. T., 2012. Diurnal and nocturnal birds vocalize at night: a review. The Condor, 114: 245–257, Doi: 10.1525/cond.2012.100193 MacGregor–Fors, I., 2010. How to measure the urban–wildland ecotone: Redefining "peri–urban" areas. Ecological Research, 25: 883–887, Doi:

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10.1007/s11284-010-0717-z Marín–Gómez, O. H., García–Arroyo, M., Sánchez– Sarria, C. E., Sosa–López, J. R., Santiago–Alarcón, D., MacGregor–Fors, I., 2020. Nightlife in the city: Drivers of the occurrence and vocal activity of a tropical owl. Avian Research, 11: 9, Doi: 10.1186/ s40657-020-00197-7 Mukhin, A., Grinkevich, V., Helm, B., 2009. Under cover of darkness: Nocturnal life of diurnal birds. Journal of Biological Rhythms, 24: 225–231, Doi: 10.1177/0748730409335349 Nix, J. H., Howell, R. G., Hall, L. K., McMillan, B. R., 2018. The influence of periodic increases of human activity on crepuscular and nocturnal mammals:

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testing the weekend effect. Behavioral Processes, 146: 16–21, Doi: 10.1016/j.beproc.2017.11.002 R Core Team, 2020. R: A language and environment for statistical computing. R Foundation for Statistical Computing, Vienna, Austria Available online at: http://www.r–project.org Rutz, C., Loretto, M. C., Bates, A. E., Davidson, S. C., Duarte, C. M., Jetz, W., Johnson, M., Kato, A., Kays, R., Mueller, T., Primack, R. B., Ropert– Coudert, Y., Tucker, M. A., Wikelski, M., Cagnacci, F., 2020. COVID–19 lockdown allows researchers to quantify the effects of human activity on wildlife. Nature Ecology and Evolution, 4: 1156–1159, Doi: 10.1038/s41559-020-1237-z

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Natural hybridization between immigrant narrow–barred Spanish mackerel Scomberomorus commerson (Lacepède, 1800) and endemic West African Spanish mackerel Scomberomorus tritor (Cuvier, 1832) in the Egyptian Mediterranean coast S. A. Bakhoum Bakhoum, S. A., 2021. Natural hybridization between immigrant narrow–barred Spanish mackerel Scomberomorus commerson (Lacepède, 1800) and endemic West African Spanish mackerel Scomberomorus tritor (Cuvier, 1832) in the Egyptian Mediterranean coast. Animal Biodiversity and Conservation, 44.2: 219–227, Doi: https://doi. org/10.32800/abc.2021.44.0219 Abstract Natural hybridization between immigrant narrow–barred Spanish mackerel Scomberomorus commerson (Lacepède, 1800) and endemic West African Spanish mackerel Scomberomorus tritor (Cuvier, 1832) in the Egyptian Mediterranean coast. Immigrant narrow–barred Spanish mackerel, West African Spanish mackerel and specimens with an external appearance somewhere between these putative parents were collected from Abu Qir Bay, East Alexandria, Egypt. The hybrid index results and univariate and multivariate analysis indicated a natural hybridization between these two species. The discriminant function analysis successfully classified individual fish in the data to one of the three fish groups. Squared Mahalanobis distances extracted from the groups indicated the three groups were clearly distinct from each other. Moreover, distances between the hybrid and Scomberomorus tritor were longer than those of the hybrid and S. commerson. The mean values of the condition factor for the hybrid were significantly higher than those of S. commerson. Natural mortality of the hybrid was significantly lower than that of the exotic parent (S. commerson), indicating that the environmental conditions in the examined region are more suitable for the hybrid type species than for the invasive parental species. Key words: Immigrant, Scomberomorus commerson, Native Scomberomorus tritor, Natural hybridization, Egyptian Mediterranean coast Resumen Hibridación natural entre la carita estriada del Pacífico Scomberomorus commerson (Lacepède, 1800) y la carita oeste–africana Scomberomorus tritor (Cuvier, 1832) en la costa mediterránea egipcia. Se capturaron ejemplares inmigrantes de carita estriada del Pacífico y de carita oeste–africana y otros ejemplares con un aspecto externo intermedio entre estos progenitores putativos en la bahía de Abu Qir, al este de Alejandría, en Egipto. Los resultados del análisis univariado y multivariado del índice de hibridación indicaron que existía hibridación natural entre las especies parentales previstas. La función discriminante permite determinar si un determinado ejemplar del que se tienen datos pertenece a uno de los tres grupos y calcular su tasa de éxito. Las distancias cuadradas de Mahalanobis obtenidas entre los grupos analizados indicaron que los tres grupos estaban claramente diferenciados y que las distancias entre los híbridos y Scomberomorus tritor eran superiores a las de los híbridos y S. commerson. Los valores medios del factor de condición de los ejemplares híbridos fueron significativamente mayores que los de S. commerson; además, la mortalidad natural de los híbridos fue significativamente inferior a la de los progenitores exóticos (S. commerson), lo que indica que las condiciones ambientales de la región estudiada son más adecuadas para el tipo híbrido que para las especies parentales invasoras. Palabras clave: Inmigrante, Scomberomorus commerson, Scomberomorus tritor autóctono, Hibridación natural, Costa mediterránea de Egipto Received: 19 III 21; Conditional acceptance: 31 V 21; Final acceptance: 22 VI 21 Shnoudy A. Bakhoum, National Institute of Oceanography and Fisheries, Alexandria, Egypt. Corresponding author: S. A. Bakhoum. E–mail: shnoudybakhoum@yahoo.com ORCID ID: 0000-0003-4691-9218 ISSN: 1578–665 X eISSN: 2014–928 X

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Introduction The Suez Canal was opened in 1869, connecting the Red Sea to the Mediterranean and allowing numerous Indo–Pacific species to invade the Mediterranean (Golani, 1998). This process of immigration through the Canal increased alien species of Indo–Pacific origin from 12 species in 1882 to 92 in 2010 (Keller 1882; Zenetos et al., 2010). In 1985, Ben–Tuvia recorded that the narrow–barred Spanish mackerel S. commerson (Lacepède, 1800) had reached the eastern part of the Mediterranean Sea. Golani et al. (2002) later reported that this Lessepsian migrant has expanded its distribution to the Aegean Sea. Several migrant Lessepsian fishes are now well– established in the eastern Mediterranean. Halim and Rizkalla (2011) published a checklist of 42 immigrant Erythrean fish in the Egyptian Mediterranean, 17 of which are commercially exploited. Permanent change of habitat has a great effect on the biometric characters of aquatic organisms (Abd El–Gawad et al., 1995; Bakhoum, 2017). In the Egyptian Mediterranean waters, the immigrant Brush– tooth lizardfish showed a sufficiently high degree of biometric differences to recognize the Mediterranean brush–tooth lizardfish (Saurida undosquamis) as a distinct group from that of the Red Sea Fish (Bakhoum, 2000). Hybridization is defined as the crossing of genetically distinguishable groups or individuals. It includes crosses both between lineages of the same species (intraspecific) and between individuals of different species (Pinheiro et al., 2019). The invasion of closely related fish species may disturb a habitat and lead to an increasing incidence of interspecific hybridization. It may also facilitate hybrid zones where the likelihood for interbreeding between native and exotic species is seriously increased (Costedoat et al., 2005; Almodóvar et al., 2012). The aim of the present study was to detect hybrids between the native species West African Spanish mackerel S. tritor (Cuvier, 1832) and the exotic species S. commerson (Lacepède, 1800 ), and to evaluate the degree of hybrid adaptation in Egyptian Mediterranean waters east of Alexandria. Material and methods Study area Between November 2019 and January 2020, 102 specimens (27 S. tritor, 33 S. commerson and 42 possible hybrid fishes) were collected with the help of local fishermen using daytime purse seine fishing gear from Abu Qir Bay in the Egyptian Mediterranean coast, east of Alexandria (latitude from 30º 5' to 30º 20' N and longitude from 31º 15' to 31º 25' E).

were taken to the nearest mm) and eight meristic counts. To minimize any variation resulting from allometric growth, all morphometric measurements were standardized according to Reist (1985): X'i,j = log Xi – b . (log TLj – log TLi) where X'i,j is the standardized measurement of the i morphometric character; log Xi is the mean logarithm of i morphometric character measurement; TLj is the total length of the individual j; log TL is the logarithm of the mean total length of pooled individuals and b is the slope of the log X against logTL plot. To elucidate the differentiation of the species and expected hybrid we used forward stepwise discriminant analysis (DA) on the characters, based on the generalized Mahalanobis distance to determine the similarity between groups and the ability of these variables to identify the specimens correctly (Hair et al., 1998). We used univariate analysis of variance (one way ANOVA) for meristic and size–adjusted data sets and multivariate discriminant function analysis to select the important variables (Henault and Fortin, 1989). The hybrid index was calculated according to Witkowski and Blachuta (1980): Hybrid index = H – M1 / M2 – M1 where, M1 is the numerical value of the same character of S. tritor, M2 the numerical value of a character of S. commerson, and H the numerical value of hybrid characters. Characters of the hybrid approach 50 are intermediate, while characters close to 0 or 100 indicate a character state close to that of a parent species. The commonly used length–weight relationship was applied: W = a . L . b where L is total length (cm), W is weight (g), and a and b are constants. The coefficient of condition (Fulton condition factor, K) was calculated from the equation: K = 100 . W / L3 where W is the gutted weight in grams, L the total length in cm. This factor is often used as an approximation even when the allometric factor is theoretically more appropriate (Bagenal and Braum, 1971; Ricker 1975). The natural mortality coefficient 'M' was calculated by the method described by Ursin (1967). Data of condition factor and natural mortality were statistically analyzed using ANOVA. All statistical analyses were performed using the SPSS PC ver. 16 software packages.

Methods

Results

The biometric characters examined included twenty– one morphometric measurements (all measurements

Comparing morphological features of the parental species with expected hybrid showed that the back

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Fig. 1. Photographs of parent species and putative hybrids: A, Scomberomorus commerson; B, Scomberomorus tritor; C, Hybrid form. Fig. 1. Fotografías de las especies parentales y los híbridos putativos: A, Scomberomorus commerson; B, Scomberomorus tritor; C, Hybrid form.

colour of S. tritor was bluish–green. Sides were silvery, with about three rows of vertical elongated spots and lateral line gradually curving down towards the caudal peduncle. Conversely, the back of S. commerson was iridescent blue–green, and sides were silver, with numerous wavy vertical bands. The lateral line abruptly bent downward below the end of a second dorsal fin. The hybrid back was blue–green and its sides contained vertical elongated spots and numerous wavy vertical bands inherited from both parental species. Moreover, the lateral line gradually curved downwards below the end of the second dorsal fin (fig. 1). The hybrid index For some characters, the average values of hybrids were close to either S. commerson or S. tritor but

distinct from both in other characters. In contrast, the hybrid index revealed that hybrid specimens contained 4 individual characters with a hybrid index > 100 and < 0, 2 intermediate characters (hybrid index 45–55), 3 close to S. commerson (hybrid index > 55) and 19 close to S. tritor (hybrid index < 45) (see table 1). Biometric characters Table 1 summarizes a comparison of 21 morphometric measurements and eight meristic counts between parent species and hybrid groups. Univariate analyses One–way ANOVA indicated significant differences of meristic counts and morphometric measurements (p < 0.001) between the hybrid groups and parental species. Tukey's honest–significant difference (HSD) of

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Table 1. Hybrid index of meristic and morphometric measurements (mean ± SD) of Scomberomorus tritor, Scomberomorus commerson and hybrid specimens collected from Abu Qir Bay coast, east off Alexandria, Egypt: Hi, hybrid index. Tabla 1. Índice de hibridación de las mediciones merísticas y morfométricas (media ± DE de los ejemplares de Scomberomorus commerson y Scomberomorus tritor y los híbridos capturados en la bahía de Abu Qir, al este de Alejandría: Hi, índice de hibridación.

S. tritor

S. commerson

Hybrid

Biometric characters First dorsal fin 15.78 ± 1.093 (15–18) 16.17 ± 0.983 (15–18) 15.86 ± 0.727 (14–17) spines and rays Second dorsal fin 16.56 ± 0.882 (15–18) 16.67 ± 1.633 (15–19) 16.43 ± 0.676 (15–18) spines and rays Anal fin spine 17.78 ± 1.302 (16–20) 17.67 ± 1.633 (16–20) 17.50 ± 1.044 (16–20) and rays Pectoral fin spine 19.56 ± 1.509 (18–22) 22.17 ± 1.169 (21–24) 21.76 ± 1.338 (20–24) and rays Ventral fin spine I ± 5 I ± 5 I ± 5 and rays Dorsal finlets 8.56 ± 0.726 (7–9) 9.33 ± 0.516 (9–10) 9.62 ± 0.590 (8–10) Anal finlets 8.89 ± 0.333 (8–9) 9.50 ± 0.548 (9–10) 9.52 ± 0.512 (9–10) Vertebrae (total) 45.78 ± 0.667 43.00 ± 1.265 44.19 ± 1.436 (45–47) (42–45) (41–46) Morphometric measurement (mm) Total length 275.38 ± 44.201 394.17 ± 43.407 312.95 ± 56.743 (240–375) (310–430) (262–455) Forked length 248.63 ± 44.622 359.33 ± 50.162 275.10 ± 52.371 (215–350) (262–395) (233–410) Standard length 241.63 ± 42.671 349.83 ± 48.967 275.00 ± 50.838 (212–340) (264–392) (231–398) Predorsal length 62.75 ± 8.631 85.17 ± 10.610 67.76 ± 11.036 (55–80) (64–92) (51–97) Prepectoral length 58.63 ± 7.347 (51–72) 79.00 ± 10.621 (62–90) 65.29 ± 9.660 (56–91) Preanal length 132.38 ± 16.370 194.83 ± 32.591 147.90 ± 26.760 (113–138) (145–225) (123–210) Prepelvic length 61.63 ± 6.948 (55–75) 88.67 ± 4.590 (83–95) 67.71 ± 10.340 (59–92) Body depth 47.38 ± 7.671 (38–60) 64.67 ± 7.202 (60–73) 53.90 ± 7.409 (45–74) Caudal peduncle length 17.88 ± 3.523 (15–25) 25.67 ± 3.724 (20–30) 21.57 ± 3.501 (18–30) Caudal peduncle depth 10.50 ± 2.204 (9–14) 14.17 ± 1.602 (12–16) 11.19 ± 3.371 (8––13) Dorsal rays height 24.25 ± 2.816 (21–28) 35.50 ± 7.007 (23–40) 27.81 ± 27.81 (24–40) Anal base length 23.00 ± 5.732 (19–35) 36.17 ± 7.808 (22–44) 26.62 ± 5.220 (21–39) Head length 56.38 ± 5.927 (52–69) 78.00 ± 12.000 (58–88) 63.48 ± 8.418 (55–86) Head width 23.25 ± 3.655 (18–29) 32.33 ± 3.933 (26–36) 25.76 ± 4.122 (22–36) Head depth 36.50 ± 4.811 (31–46) 49.00 ± 4.817 (42–54) 40.76 ± 4.959 (30–53) Head depth passing through eyes 18.13 ± 2.850 (15–23) 30.83 ± 3.764 (25–35) 27.43 ± 4.261 (20–36) Eye diameter 11.88 ± 0.835 (10–13) 16.00 ± 2.608 (12–19) 12.62 ± 1.564 (17–11) Interorbital width 17.38 ± 1.996 (15–21) 24.17 ± 2.483 (20–27) 19.33 ± 3.055 (27–16) Snout length 21.75 ± 2.188 (20–26) 31.67 ± 3.559 (26–35) 24.52 ± 3.710 (34–21) Upper jaw length 32.50 ± 3.295 (29–39) 43.83 ± 3.251 (39–47) 35.76 ± 4.493 (48–30) Lower jaw length 31.88 ± 3.227 (29–39) 45.00 ± 3.225 (39–48) 37.86 ± 4.607 (34–49)

20.51 –118.18 254.55 84.29 – 137.66 103.28 57.19

31.63 23.91 30.84 22.35 32.70 24.85 22.49 37.71 47.37 18.80 31.64 27.49 32.84 27.64 34.08 73.23 17.96 28.72 27.92 28.77 45.58

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Table 2. Significant differences between S. tritor (St), S. commerson (Sc) and hybrid (H) specimens based on ANOVA followed by the Tukey HSD test (honestly significant difference). (Significance level * p < 0.05, ** p < 0.01). Tabla 2. Diferencias significativas entre los ejemplares de S. tritor (St), S. commerson (Sc) y los híbridos (H) basadas en el ANOVA y el test HSD de Tukey (diferencia honestamente significativa). (Nivel de significación * p < 0,05; ** p < 0,01)".

Tukey HSD test F–value

St vs. Sc

St vs. H

Sc vs. H

56.723**

0.000**

0.891

0.000**

Second dorsal fin spines and rays

4.158*

0.091

0.758

0.018*

Anal fin spine and rays

1.634

0.203

0.933

0.313

Pectoral fin spine and rays

18.37**

0.000**

0.683

Dorsal fin lets

21.742**

0.000**

0.855

Anal fin lets

13.401**

0.001**

0.000**

0.990

Vertebrae (total)

64.642**

0.000**

Biometric characters First dorsal fin spines and rays

Morphometric measurement (mm) Total length

20.126**

0.000**

0.100

0.000**

Forked length

7.137**

Standard length

19.874**

0.003**

0.015*

0.549

0.000**

0.220

Predorsal length Prepectoral length

0.084

1.000

0.934

0.941

0.658

0.665

0.529

0.996

Preanal length

45.093**

0.000**

0.186

Prepelvic length

24.289**

0.000**

0.001**

Body depth

46.902**

0.000**

0.892

Caudal peduncle length

43.119**

0.000**

1.000

Caudal peduncle depth

5.052**

0.010**

0.700

0.041*

Dorsal rays height

11.288**

0.000**

0.001**

0.714

Anal base length

10.572**

0.007**

0.000**

0.769

Head length

4.776*

0.025*

0.031*

0.891

Head width

7.850*

0.008**

0.002**

1.000

Head depth

3.115

0.065

0.123

0.803

Head depth passing through eyes

87.776**

0.000**

0.089

Eye diameter

12.91**

0.003**

0.000**

0.711

Interorbital width

25.627**

0.000**

0.483

Snout length

28.661**

0.000**

0.385

Upper jaw length

0.922

0.947

0.400

0.666

Lower jaw length

37.766**

0.000**

0.869

meristic counts indicated that the significant differences between parental species were found in five characters. The differences between hybrid and S. tritor were observed in four meristic counts, while hybrid type varied from S. commerson only in first and second dorsal fin spines and rays and vertebrae (p < 0.001). Adjusted morphometric measurements revealed 17 of 21 significant characters between parental

species. The hybrid type differed in 15 and 3 measurements from S. tritor, and S. commerson, respectively (table 2). Multivariate analysis Canonical variate analysis (CVA) was performed on meristic counts and twenty morphometric ratios to assess the shape variation between the two species

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Table 3. Discriminant analysis classification showing the percentage of specimens classified in each species, based on meristic counts: St, Scomberomorus tritor; Sc, Scomberomorus commerson; H, hibrid.

Table 4. Discriminant analysis classification showing the percentage of specimens classified in each species, based on morphometric ratios: St, Scomberomorus tritor; Sc, Scomberomorus commerson; H, hibrid.

Tabla 3. Clasificación según el análisis discriminante en la que se muestra el porcentaje de ejemplares clasificados en cada especie, según los estudios merísticos: St, Scomberomorus tritor; Sc, Scomberomorus commerson; H, híbrido.

Tabla 4. Clasificación según el análisis discriminante en la que se muestra el porcentaje de ejemplares clasificados en cada especie, según las proporciones morfométricas: St, Scomberomorus tritor; Sc, Scomberomorus commerson; H, híbrido.

Predicted species

Species St

Predicted species

Species

St Sc H

St 94.40 0.00 5.60

100.00 0.00 0.00

Sc 0.00 100.00 0.00

0.00 100.00 0.00

0.00 0.00 100.00

0.00

0.00 100.00

and hybrid form. Consequently, standardized canonical discriminant function coefficients extracted two meristic factors and seven morphometric ratios. The discriminant function successfully identified the membership of individual fish in the data to one of the three fish groups. The percentage of correctly identified specimens based on meristic counts for the hybrid was 100 % and for S. tritor and S. commerson individuals it was 94.4 % and 100 %, respectively (table 3). Identification of fish groups based on seven landmark morphometric ratios classified 100 % parental species and hybrid (table 4). CVA extracted two CVs, accounting for 100.00 % variations and showed three groups with no overlaps between groups, with a position of hybrid specimens lying between expected parental species. Moreover, CVA plots suggested that hybrids were resultant due to crossing between S. tritor and S. commerson (fig. 2). Squared Mahalanobis distances extracted among groups of S. tritor, S. commerson and hybrid specimens based on meristic and morphometric characters were highly significant (P < 0.0001), indicating that all groups were clearly distinct from each other, and squared Mahalanobis distances between hybrid and S. tritor were longer than those hybrid and S. commerson. Length–weight relationship and condition factor The relationship between total length (cm) and total weight (W) was represented by the equations: For S. tritor W = 0.0245 . L . 2.6548 (r2 = 0.9655) S. commerson W = 0.0092 . L . 2.1036 (r2 = 0.9666) Hybrid form W = 0.0192 . L . 2.3997 (r2 = 0.9455)

The mean values of condition factors of S. tritor fish (0.803 ± 0.279) were higher than those of S. commerson (0.354 ± 0.193) and hybrid (0.660 ± 0.083) fish. Analysis of variance indicated significant differences in the mean values of condition factor between parental species (F = 7.277, p < 0.01), S. tritor and hybrid fish (F = 61.921, p < 0.01) and between S. commerson and hybrid fish (F = 26.110, p < 0.01). Natural mortality The hybrid fish had the least mean value of natural mortality coefficient (0.177 ± 0.024) followed by S. tritor (0.203 ± 0.015). S. commerson had the highest mortality value (0.382 ± 0.126). Statistical analysis revealed significant differences in the natural mortality of S. commerson compared with S. tritor (F = 34.151, p < 0.01) and the hybrid (F = 46.011, p < 0.01). Analysis of variance revealed no significant difference in the mean values of natural mortality between S. tritor and the hybrid (F = 2.020, p > .05). Discussion Nature hybridization in most cases is a temporal phenomenon occurring at different scales over an extended time frame (Avise and Walker, 2000). Hybrid species, especially those leading to genomic introgression, may be an evolutionarily constructive process, as occurred in the family Salmonidae (Arnold, 1997; Dowling and Secor, 1997). Extensive hybridization and introgression occurs more commonly in fish than in other vertebrates of comparable levels of genetic divergence (Epifanio and Nielsen, 2001). Almodóvar et al. (2012) recently described a case of natural hybridization between invasive bleak

Animal Biodiversity and Conservation 44.2 (2021)

A 1 2 3 Centroid

5.0

225

100

50 2

1 0.0

–2.5

Function 2

2.5 Function 2

1 2 3 Centroid

2 3

–50

–5.0

–100 –5.0

–2.5 0.0 2.5 5.0 –100 Function 1

–50

0 50 Function 1

100

Fig. 2. Plots of canonical variate analysis based on meristic counts (A) and morphometric ratios (B) for S. tritor (1), S. commerson (2) and putative hybrid (3) showing frequency of specimen distribution in respective group on the first two axes. Fig. 2. Gráficos del análisis de variables canónicas según los caracteres merísticos (A) y según las proporciones morfométricas (B) de S. tritor (1), S. commerson (2) y los híbridos putativos (3) en los que se muestra la frecuencia de la distribución de los ejemplares en el grupo respectivo en los dos primeros ejes.

Alburnus alburnous and endemic calandino Squalius alburnoides complex following a short period of contact. This study deals with the characterization of the hybrids and their distinction and similarities with the parental species. Hybridization in nature is generally the result of interference by humans, through the construction of reservoirs, for example, or introduction of exotic species, or modification of rivers and Seas (Crivelli and Dupont, 1987; Pouyaud and Agnèse, 1995; Agnese et al., 1998; Bakhoum, 2019). The Suez Canal is an artificial waterway connecting the tropical Red Sea and the subtropical Eastern Mediterranean Sea. This example of human intervention caused a global change in the distribution of native and non–native fishes in Mediterranean waters (Ben–Tuvia, 1985; Golani et al., 2002). Natural hybridization occurs when reproductive barriers break down. These barriers may be physiological, behavioural or geographic. Hybridization between S. commerson and S. tritor in the Suez Canal is an example. Reproductive barriers between species, both pre– and post–zygotic, appear to be incomplete for many fishes (Simon and Noble, 1968; Rosenfield et al., 2000; Hendry et al., 2000). The parental species examined have pelagic eggs and larvae (Collette, 1986) and the spawning season of S. tritor from July to August interfered with the spawning of S. commerson, which extends from October to July (Collette and Nauen, 1983; Collette and Russo, 1984), thus providing great opportunity for crossbreeding.

Many researchers use biometric characters to identify natural hybrids in fish (Reist et al., 1992; Bakhoum, 2009; Jacquemin and Pyron, 2016). In the present study, the hybrid index revealed that most characters studied in hybrids were closer to either S. commerson or S. tritor, and one intermediate character revealed the hybrid nature of these specimens. Canonical variate analysis (CVA) allowed parent species to be distinguished from hybrids. No overlapping between groups was revealed, and classification in the respective group was 100 % correct. CVA placed hybrids in a position between groups of S. commerson and S. tritor, indicating that hybrids were resultant from crossing between the two species. Squared Mahalanobis distances indicated that the hybrid form was relatively closer to S. commerson than to S. tritor. This may attribute to the back–crossing between the first generation of hybrid specimens with parental S. tritor fish. The condition factor is used to compare condition, fitness or wellbeing of fish. It is based on the hypothesis that heavier fish of a particular length are in a better physiological condition (Bagenal and Tesch, 1978). It is strongly influenced by both biotic and abiotic environmental conditions and can be used as an index to assess the status of the aquatic ecosystem in which fish live (Anene, 2005). The mean values of condition factors of endemic S. tritor fish were higher than those of immigrant S. commerson and hybrid fishes due to their adaptation to environmental conditions over many years. The

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mean values of the condition factor for the hybrid were significantly higher than those of S. commerson, possibly because exotic species are strongly influenced by both biotic and abiotic situations in the new habitat. Moreover, recent studies have provided evidence for hybrid fitness when compared to parental species (Arnold and Hodges, 1995; Reyer, 2008). Natural mortality (M) is an essential parameter in determining fish stock and an indicator of the adaptation of a population of fish. IIes (1968) found that in any fish population the mortality rate and the size at which maturity is achieved determines the proportion capable of reproduction and the relative reproductive potential. In the present study, the natural mortality of the hybrid and S. tritor were significantly lower than that of the exotic parent (S. commerson), indicating that the environmental conditions in the examined region are more suitable for the hybrid type and native fishes than for invasive parental species. A comparison of the natural mortality coefficient of S. undosquamis from the two habitats indicated that the natural mortality coefficient for the emigrant to Mediterranean fish was comparatively higher than that of endemic Red Sea fish (Bakhoum, 2000). This may be attributed to variation in environmental conditions. El–shenawy et al. (2006) recorded that the range of water temperatures in the Egyptian Mediterranean coast was wide(17.14–26.31 ºC), with narrow fluctuations in salinity (37.51–39.710 ‰). In contrast, in the Red Sea coastal waters, temperatures fluctuated between 20.8 ºC and 28.1 ºC and salinity ranged between 39.0 ‰ and 40.40 ‰. Differences in temperature between habitats leads to varied natural mortality values. The relationship between growth and natural mortality is strong, but growth rates depend on temperature, explaining the effects of temperature on natural mortality (Gislason et al., 2010). We hope this paper helps to promote scientific research into the importance of natural hybridization on living resource management issues. References Abd El–Gawad, A. M, Bakhoum, Sh. A., Ragheb. E., 1995. Comparison of meristic and morphometric characters of Solea aegyptiaca in Mediterranean Sea and Lake Qarun, Egypt. Bulletin of the National Institute of Oceanography and Fisheries, ARE, 21: 451–459. Agnese, J. F., Adepo–gourene, B., Pouyaud, L., 1998. Natural hybridization in tilapias. In: Genetics and Aquaculture in Africa Collection colloques et seminires: 95–103. ORSTOM, Paris. Almodóvar, A., Nicola, G. G., Leal, S., Torralva, M. , Elvira, B., 2012. Natural hybridization with invasive bleak Alburnus alburnus threatens the survival of Iberian endemic calandino Squalius alburnoides complex and Southern Iberian chub Squalius pyrenaicus. Biological Invasions, 14: 2237–2242, Doi: 10.1007/s10530-012-0241-x Anene, A., 2005 Condition factors of four cichlid species of a man–made lake in Imo state, Southeast, Nigeria. Turkish Journal of Fisheries and Aquatic

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Sciences, 5: 43–47. Arnold, M. L., 1997. Natural Hybridization and Evolution. Oxford University Press, New York. Arnold, M. L., Hodges, S.A., 1995. Are hybrids fit or unfit relative to their parents? Trends in Ecology and Evolution, 10: 67–71, Doi: 10.1016/S01695347(00)88979-X Avise, J. C., Walker, D., 2000. Abandon all species concepts? A response Conservation Genetics, 1: 77–80, Doi: 10.1023/A:1010189805191 Bagenal, T. B., Braum, E., 1971. Eggs and early life history. In: Methods of Assessment of Fish production in fresh waters, IBP Handbook 3: 165–201 (T. Bagenal, Ed.). Blackwell Scientific Publication, Oxford. Bagenal, T., Tesch, F. W., 1978. Age and Growth. In: Methods of Assessment of Fish Production in Fresh Waters, IBP Handbook 3: 101–136 (T. Bagenal, Ed.). Blackwell Scientific Publication, Oxford. Bakhoum, S. A., 2000. Comparative study on brush– tooth lizardfish Saurida undosquamis (Richardson), from the Red Sea and Mediterranean Sea coasts of Egypt. Oebalia, International Journal of Marine Biology and Oceanography, 26: 35–48. – 2009. Biometric characteristics and some biological features of natural hybrids between Nile tilapia Oreochromis niloticus and blue tilapia Oreochromis aureus in Lake Edku, Egypt. International Journal of Ichthyology, 15(4): 191–204. – 2017. Shape and behavior of aquatic organisms. LAP Lambert Academic publishing, Germany. – 2019. Fish Assemblages in Surf Zone of the Egyptian Mediterranean Coast off Alexandria. Turkish Journal of Fisheries and Aquatic Sciences, 19(4): 351–362, Doi: 10.4194/1303-2712-v19_4_09 Ben–Tuvia, A., 1985. The impact of the Lessepsian (Suez Canal) fish migration in the Eastern Mediterranean ecosystem. In: Mediterranean Marine Ecosystem, NATO Conference Series, I Ecology, 8: 367–375 (A. Moraitou–Apostolopoulou, V. Kiortsis, Eds.). Springer, Boston, MA, Doi: 10.1007/978-14899-2248-9_17 Collette, B. B., 1986. Scombridae (including Thunnidae, Scomberomoridae, Gasterochismatidae and Sardidae). In: Fishes of the north–eastern Atlantic and the Mediterranean, volume 2: 981–997 (P. J. P. Whitehead, M. L. Bauchot, J. C. Hureau, J. Nielsen, E. Tortonese, Eds.). Unesco, Paris. Collette, B. B., Nauen, C. E., 1983. Scombrids of the world. An annotated and illustrated catalogue of tunas, mackerels, bonitos and related species known to date. FAO Fisheries Synopsis, 125(2): 137. Food and Agriculture Organization of the United Nations, Rome. Collette, B. B., Russo, J. L., 1984. Morphology, systematics, and biology of the Spanish mackerels (Scomberomorus, Scombridae). Fishery Bulletin, 82: 545–692. Costedoat, C., Pech N., Salducci, M. D., 2005. Evolution of mosaic hybrid zone between invasive and endemic species of Cyprinidae through space and time. Biological Journal of the Linnean Society, 85: 135–155. Doi: 10.1111/j.1095-8312.2005.00478.x

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Crivelli, A. J., Dupont, F., 1987. Biometrical and biological features of Alburnus alburnus X Rutilus rubilio natural hybrids from Lake Mikri Prespa, northern Greece. Journal of Fish Biology, 31: 721–733. Doi: 10.1111/j.1095-8649.1987.tb05275.x Dowling, T. E., Secor, C. L., 1997. The role of hybridization and introgression in the diversification of animals. Annual Review of Ecology, Evolution, and Systematics, 28: 593–619, Doi: 10.1146/annurev. ecolsys.28.1.593 El–shenawy, M., Farag, A., Zaky, M., 2006. Sanitary and aesthetic quality of Egyptian coastal Waters of Aqaba Gulf, Suez Gulf and Red Sea. Egyptian Journal of Aquatic Research. 32: 120–234. Epifanio, J., Nielsen, J., 2001. The role of hybridization in the distribution, conservation and management of aquatic species. Reviews in Fish Biology and Fisheries, 10: 245–251. Gislason, H., Rice, J., Niels, D., Pope, J. G., 2010. Size, growth, temperature and the natural mortality of marine fish. Fish and Fisheries, 11: 149–158. Golani, D., 1998. Impact of Red Sea fish migrants through the Suez Canal on the aquatic environment of the eastern Mediterranean. Bulletin of Yale School of Forestry and Environmental Studies, 103: 375–387. Golani, D., Orsi–Relini, L., Massuti, E., Quignard, J. P., 2002. CIESM Atlas of exotic species in the Mediterranean, vol 1. In: Fishes: pp?? (F. Briand, Ed.). CIESM publishers, Monaco. Hair, J., Anderson, R. E., Tatham, R. L., Black, W. C., 1998. Multivariate data analysis, 4th edition. Prentice Hall, Englewood Cliffs, NJ. Halim, Y., Rizkalla, S., 2011. Aliens in Egyptian Mediterranean waters. A check–list of Erythrean fish with new records. Mediterranean Marine Science, 12(2): 479–490. Henault, M., Fortin, R., 1989. Comparison of meristic and Morphometric characters among spring–and fall–spawning ecotype of Cisco (Coregonus artedii) in southern Quebec. Canadian Journal of Fisheries and Aquatic Science, 46: 166–173. Hendry, A. P., Vamosi, S. M., Latham, S. J., Heibuth, J. C., Day, T., 2000. Questioning species realities. Conservation Genetics, 1: 67–76, Doi: 10.1023/A:1010133721121 IIes, T. D., 1968. Dwarfing or strutting in the genus Tilapia (Cichlidae). Possibly recruitment mechanism. FAO Fisheries Report, Roma. Jacquemin, S. J., Pyron, M., 2016. A century of morphological variation in Cyprinidae fishes. BMC Ecology, 16: 48, Doi: 10.1186/s12898-016-0104-x Keller, C., 1882. Die fauna in Suez Canal und die diffusion der Mediterranean und Erythraischen Tierwelt. Neue Denkschriften der allgemeinen

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Schweizerischen Gesellschaft für die gesamten naturwissenschaften, 28(3): 1–28. Pinheiro, A. P. B., Melo, R. M. C., Teixeira, D. F., Birindelli, J. L. O., Carvalho, D. C., Rizzo, E., 2019. ntegrative approach detects natural hybridization of sympatric lambaris species and emergence of infertile hybrids. Scientific Reports, 9: 1–11, Doi: 10.1038/s41598-019-408 Pouyaud, L., Agnese, G. F., 1995. Phylogenetic relationships between 21 species of three tilapiine genera Tilapia, Sarotherodon and Oreochromis using allozyme data. Journal of Fish Biology, 47: 26–38, Doi: 10.1111/j.1095-8649.1995.tb01870.x Reist, J. D., 1985. An empirical evaluation of several univariate methods that adjust for size variation in morphometric data. Canadian Journal of Zoology, 63(6): 1429–1439. Reist, J. D., Vuorinen, J., Bodaly, R. A., 1992. Genetic and Morphological identification of coregonid hybrid fishes from arctic Canada. Polish Archives of Hydrobiology, 39(3–4): 551–56. Reyer, H. U., 2008. Mating with the wrong species can be right. Trends in Ecology and Evolution, 23: 289–292. Ricker, W. E., 1975. Computation and interpretation of biology statistics for fish population. Bulletin of the Fisheries Research Board of Canada, Bulletin 191. ARLIS, Alaska Resources Library and lnformation Services, Ottawa. Rosenfield, J. A., Todd, T., Greil, R., 2000. Asymmetric hybridization and introgression between pink and chinook salmon in the Laurentian Great Lakes. Transactions of the American Fisheries Society, 129: 670–679. Simon, R. C., Noble, R. E., 1968. Hybridization in Oncorhynchus (salmonidea). I. Viability and inheritance in artificial crosses of chum and pink salmon. Transactions of the American Fisheries Society, 97: 109–118. Ursin, E., 1967. A mathematical model of some aspects of fish growth respiration and mortality. Journal of the Fisheries Research Board of Canada, 24(2): 2355–2453. Witkowski, A., Blachuta, J., 1980. Natural hybrids Alburnus alburnus (L.) x Leuciscus cephalus (L.) and Rutilus rutilus (L.) x Abramis brama (L.) from the Rivers San and Biebrza. Acta Hydrobiologica, 22: 473–487. Zenetos, A., Gofas, S., Verlaque, M., Cinar, M. E., Bianchi, C. N., 2010. Alien species in the Mediterranean Sea by 2010. A contribution to the application of European Union’s Marine Strategy Framework Directive. Part I. Spatial Distribution. Mediterranean Marine Science, 11(2): 381–493, Doi: 10.12681/mms.87

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Increasing wild boar density explains the decline of a Montagu's harrier population on a protected coastal wetland J. Crespo, J. Jiménez, A. Martínez–Abraín

Crespo, J., Jiménez, J., Martínez–Abraín, A., 2021. Increasing wild boar density explains the decline of a Montagu's harrier population on a protected coastal wetland. Animal Biodiversity and Conservation, 44.2: 229–239, Doi: https://doi.org/10.32800/abc.2021.44.0229 Abstract Increasing wild boar density explains the decline of a Montagu's harrier population on a protected coastal wetland. We studied the rapid decline in the number of breeding pairs (geometric growth rate λ = 0.86; 14 % annual decrease) of a semi–colonial ground–nesting bird of prey, the Montagu's harrier (Circus pygargus), after twelve years of rapid population growth (λ = 1.15; 15 % rate of annual increase) in a protected coastal wetland in Eastern Spain. The study was conducted from 1992–2017, and the range of values in population size was: 2–37 breeding pairs. We contrasted 20 biologically–sound hypotheses (including local and regional factors) to explain the trend over time in the annual number of pairs. The most parsimonious model included a surrogate of wild boar (Sus scrofa) density in the region during the previous year and the annual number of Montagu's harrier pairs breeding inland in the study province during the focal year. Syntopic western marsh harriers (C. aeruginosus) were not found to have any effect on the numbers of Montagu's harriers either in our modelling or when we performed a quantitative and qualitative study both for years t and t–1. Our final 'best' models did not include spring rainfall, regional forest fires or local land use changes. The impact of wild boars on breeding success, together with conspecific attraction, could have resulted in the dispersal of coastal wetland birds to larger populations in dense inland shrub lands where levels of wild boar nest predation were more likely lower. Key words: Ground–nesting birds, Regional dynamics, Protected wetlands, Nest predation, Circus pygargus, Sus scrofa Resumen El aumento de la densidad de jabalíes explica la reducción de una población de aguilucho cenizo presente en un humedal costero protegido. Hemos estudiado el rápido descenso del número de parejas reproductoras (tasa de crecimiento geométrico λ = 0,86; 14 % de disminución anual) de un ave semicolonial que nidifica en el suelo, el aguilucho cenizo (Circus pygargus), tras 12 años de rápido crecimiento demográfico (λ = 1,15; 15 % de tasa de aumento anual) en un humedal costero protegido situado en el este de España. El periodo de estudio fue 1992–2017, con un intervalo de valores del tamaño de población de 2–37 parejas reproductoras. Hemos contrastado 20 hipótesis razonables desde el punto de vista biológico (teniendo en cuenta factores locales y regionales) para explicar la tendencia del número anual de parejas. El modelo más parsimonioso incluyó un indicador de la densidad regional del jabalí (Sus scrofa) durante el año anterior y el número anual de parejas de aguilucho cenizo que se reprodujeron en el interior de la provincia del estudio, durante el año en cuestión. En nuestro modelo no se observó que el aguilucho lagunero occidental (C. aeruginosus) sintópico tuviera efecto alguno en el número de aguiluchos cenizos; tampoco lo tuvo en los análisis cuantitativos y cualitativos relativos a los años t y t–1. Las precipitaciones de primavera, los incendios forestales en la región y los cambios en el uso del suelo a escala local tampoco aparecieron en nuestros mejores modelos finales. Los efectos de jabalíes en el éxito reproductor, junto con la atracción de individuos conespecíficos, podrían haber provocado la dispersión de las aves de los humedales costeros a poblaciones más numerosas presentes en zonas arbustivas de interior, donde probablemente la depredación de nidos por jabalí sea inferior. Palabras clave: Aves que nidifican en el suelo, Dinámica regional, Humedales protegidos, Depredación de nidos, Circus pygargus, Sus scrofa ISSN: 1578–665 X eISSN: 2014–928 X

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Received: 23 II 21; Conditional acceptance: 09 VI 21; Final acceptance: 27 VI 21 Jorge Crespo, Centro de Recuperación de Fauna Salvaje 'La Granja', Conselleria de Agricultura, Desarrollo Rural, Emergencia Climática y Transición Ecológica, Generalitat Valencian, Avda. Los Pinares 106, 46012 El Saler, Valencia, Spain.– Juan Jiménez, Servicio de Vida Silvestre, Conselleria de Agricultura, Desarrollo Rural, Emergencia Climática y Transición Ecológica, Generalitat Valenciana, Ciutat Administrativa 9 d'Octubre, Torre 1, 46018 Valencia, Spain.– Alejandro Martínez–Abraín, Universidade da Coruña, Facultade de Ciencias, Departamento de Bioloxía, Campus da Zapateira s/n., 15008 A Coruña, Spain. Corresponding author: A. Martínez–Abraín. E–mail: a.abrain@udc.es ORCID ID: 0000-0001-8009-4331

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Introduction Protected areas are created to secure the long–term persistence of species therein. However, animal populations in protected areas sometimes decline abruptly or are unexpectedly extirpated after years of successful breeding (Martínez–Abraín et al., 2003; Oro et al., 2004). Explanations for this phenomenon are not always evident at the local level because drivers of population decline are often multifaceted and can operate at different temporal and spatial scales (Vickery et al., 2014). For example, changes in landscape structure, in conjunction with climate change, have been listed among the most frequent causes of breeding failure in birds (Gaston et al., 2003; Luck, 2003; Hilbert et al., 2004). More specifically, habitat degradation (including the indirect effect of human infrastructure) and habitat loss (via changes in land use) are currently the main threats to raptor populations (Thiollay, 2006; Martínez– Abraín et al., 2009). Land use changes may include not only changes in agricultural practices but also the total abandonment of such activity in association with human depopulation of rural areas (Hansen et al., 2002). The consequences of such change may bring about wide and unexpected consequences at the regional level (Martínez–Abraín et al., 2020). Disentangling the relevance of local versus regional/global processes on the long–term persistence of vertebrate populations is currently a major challenge for applied ecology. We studied the possible causes of the rapid decline in the number of pairs of Montagu's harrier (Circus pygargus) in a population nesting in a coastal protected wetland. Initially, our expectations were that two local factors could be the main drivers of the decline: a) changes in land use in and around the protected wetland that might affect both the persistence of foraging and nesting habitat for Montagu’s harriers (Arroyo et al., 2002; Arroyo and García, 2005); and b) the increasing number of breeding pairs of the larger western marsh harrier (Circus aeruginosus) possibly causing competitive exclusion (Kitowski, 2008; Krupinski et al., 2012). Western marsh harriers were absent from 1992 to 1999 but colonized the site in 2000 (one pair) and grew to five pairs in 2016. However, a broader analysis of the question led us to consider that not only local but also regional factors could be involved. Ecological consequences of the human depopulation of rural areas are now emerging and are behind many population changes that have no clear proximate explanations (Martínez–Abraín et al., 2019, 2020). One major consequence of rural exodus is shrub and tree encroachment on former agricultural land, with the consequent expansion of wildlife associated to shrublands and forests and decline in species from open landscapes (Inger et al., 2015). One of the species most favoured by this process is the wild boar (Sus scrofa), a mammal that was locally extinct in the study region in the 19th century but started recolonizing the region in the 1940s (Jiménez, 2012). Wild boars have rebounded and expanded rapidly in the Iberian Peninsula over recent decades (Sáez–Royuela and Tellería, 1986; Delibes–Mateos et al., 2009) and in the entire European continent since the beginning of

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last century (Apollonio et al., 2010; Barrios–García and Ballari, 2012; Massei et al., 2014; Morelle et al., 2016). As a consequence, they have become a new agent of ecological disturbance in many ecosystems (i.e. forests, wetlands, dune fields, urban areas) and are creating a new paradigm for human–wildlife coexistence (Acevedo et al., 2011, 2014). In addition to wild boars, other regional factors must be taken into account: a) the possible role of dispersal from coastal wetland populations to other larger populations due to conspecific attraction typical of social birds (Martínez– Abraín et al., 2001); b) the regeneration of breeding habitat offering better protection from wild boar predation (mainly inland shrublands of kermes oak Quercus coccifera) after the major forest fires in the province during the 1980s; and c) changing rainfall patterns due to global climate change, as spring rainfall could be related to the level of flooding in the wetland and hence to breeding success (Corbacho et al., 1997) and lower prey capture rate (Schipper, 1973). Material and methods Study species and study population Montagu’s harrier is a semi–colonial ground–nesting raptorial bird that nests mostly in natural steppes and broad river valleys and plains within its global distribution range (Cramp and Simmons, 1980). In humanized landscapes of Europe, Montagu’s harriers often occupy anthropogenic habitats (e.g. cereal fields) that unintentionally reproduce many of the features of the species’ original habitat ('substitution habitats' sensu Martínez–Abraín and Jiménez, 2016). In Spain, around 90 % of Montagu's harriers breed in farmland (from a sample of 2,114 pairs out of 7,389 breeding pairs estimated in 2006; Arroyo and García, 2007). Similar nesting preferences have been reported for France and Portugal (Ferrero, 1995; Salamolard et al., 1999). These harriers can also use small wetlands for nesting, although the archetypical harrier species in wetlands is the western marsh harrier. The conservation status of Montagu’s harriers in the Spanish Red List of Birds (Madroño et al., 2004) is 'vulnerable'. A population decrease of 23–27 % in relation to the population in 2006 was estimated in the most recent Spanish census in 2017, indicating a declining trend at the country level (Arroyo et al., 2019). The study population was located in a protected coastal marsh in eastern Spain (Cabanes–Torreblanca Natural Park, 40º 10' N, 0º 11' E), within the province of Castellón (fig. 1). This is probably a suboptimal habitat for a ground–nesting species originally from steppe habitats as the site presents high risk of flooding if spring precipitation is high. The species might actually be present in this and other coastal marshes in the region as ecological refugees (Kerley et al., 2012, Martínez– Abraín and Jiménez, 2016, Martínez–Abraín et al., 2021) rather than by preference. This population, similarly to other populations of this species in the region, has been thoroughly monitored by the regional Department of the Environment (i.e. Conselleria de Agricultura, Desarrollo

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Table 1. Description of the variables used for modelling the number of breeding pairs of Montagu's harriers at our coastal study site. Scale, scale at which the variable acts (R, regional; L, local); Type, nature of the variable (C, climatic; H, habitat; E, ecological). The variable 'Fires' was also modelled with time delays ranging from 1 to 6 years. Overall, 12 explanatory variables were used for the modelling of number of pairs. Tabla 1. Descripción de las variables empleadas para elaborar los modelos que permiten determinar el número de parejas reproductoras de aguilucho cenizo en nuestro sitio de estudio costero. Scale, escala a la que actúa la variable: R, regional; L, local. Type, tipo de variable: C, climática; H, del hábitat; E, ecológica. En el modelo también se ha incluido la variable "Fires" con un retardo de entre 1 y 6 años. En total se emplearon 12 variables explicativas para la modelización estadística del número de parejas..

Variable Scale Type Aprp R C Wildb1 R E Ciraer L E Ciraer1 L E Montagh R E Montagh1 R E Fires R H

Description Amount of rainfall in April in the protected area (l/m2) Number of wild boars hunted regionally during the previous year Total number of western marsh Harrier pairs in the protected area in the focal year Total number of western marsh Harrier pairs in the protected area in the previous year Total number of Montagu's harrier pairs of the inland population in the focal year Total number of Montagu's harrier pairs of the inland population in the previous year Annual surface area (ha) of forest fires in the province of Castellón during the period 1986–2017

Rural, Emergencia Climática y Transición Ecológica) on a yearly basis. Field surveys and ringing of chicks by authorized and knowledgeable staff are conducted every year during the breeding season (April–July). Estimates of breeding (territorial) pairs are performed by seeking adults with territorial behaviour (or transporting nest material) and by direct nest search (Arroyo and García, 2007). Nests were placed in unflooded areas rich in Juncus maritimus and Arthrocnemum fruticosum. Study variables and statistical analyses We analysed the change in number of nesting pairs at the study site, considering a large number of biologically–sound variables with potential influence on the ecology of Montagu's harriers (table 1). The number of wild boars hunted in the province of Castellón (hunting bags) was used as a proxy of regional wild boar abundance, as in works by other authors (Massei et al., 2014). The number of hunting licenses in the province was used as surrogate to assess the change over time in hunting pressure. Importantly, wild boar hunting bags are increasing in Europe even though the number of hunting licenses is stable or declining (Massei et al., 2014), a scenario similar to that observed in our region where licenses declined from 1991–2017 (table 1s in supplementary material). The number of wild boars hunted per year was provided by the Fishing and Hunting Service of the Regional

Department of the Environment, and the number of hunting licenses was extracted from the website of this same institution (www.agroambient.gva.es). Changes in agricultural practices in the area surrounding the protected site (used by harriers to forage mainly on passerine chicks among fruit–tree orchards; G. Ros and J. Tena pers. com., see also Guixé and Arroyo, 2011) were graphically explored by comparing aerial photographs in this zone from 1996 to 2017, and by calculations from Corine Land Cover (https://visor. gva.es/visor/), quantifying variations in the surface (ha) of the different land cover classes from 1990 to 2012. For this purpose, we defined a buffer zone including all low–altitude land (< 100 m a.s.l.) around the protected area (with the rivers Sant Miquel and Xinxilla as northern and southern limits, respectively). The surface area of the buffer zone was 6,582 ha in comparison with the 848 ha of the protected wetland. Montagu's harriers are known to have foraging range sizes of over 10,000 ha in cereal steppes (Guixé and Arroyo, 2011). Data on land use changes around the protected area were not introduced in our modelling because quantitative information was only available for four years (1990, 2000, 2006 and 2012), and was hence taken into consideration in a qualitative way. Rainfall data for the study site (Torreblanca station, 40º 12' 41'' N, 00º 11' 01'' E, 35 m a.s.l.) were provided by the Spanish Meteorological Agency (www.aemet. es). The number of western marsh harrier breeding

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0º 0' 0''

1º 0' 0'' E

1 2–5 6–10 > 10

38º 0' 0'' N

39º 0' 0'' N

40º 0' 0'' N

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N Scale 1:400.000

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Fig. 1. Location of the study site (Ù , Prat de Cabanes–Torreblanca Nature Park) within the context of the province of Castellón (shaded in black in the lower left panel), Comunidad Valenciana: location and relative size of Montagu's harrier colonies within the study province and region. Fig. 1. Ubicación del área de estudio (Ù , Parque Natural Prat de Cabanes–Torreblanca) en la provincia de Castellón (sombreada en la imagen inferior izquierda), Comunidad Valenciana: ubicación y tamaño relativo de las colonias de aguilucho cenizo en la provincia y la región del estudio.

pairs and the number of Montagu's harrier breeding inland in Castellón province were provided by the regional department of the environment. The likely influence of western marsh harriers was also tested qualitatively (presence/absence) for years t and t–1. The annual land surface burned regionally by forest fires in Castellón province from 1986 to 2017 was extracted from the website of the Ministry of Agriculture and Fisheries, Food and Environment (http://www. mapama.gob.es/es/). We simultaneously tested multiple hypotheses explaining temporal variability in our dependent variable (i.e. number of pairs) by means of generalized linear mixed models with Poisson error distribution (link = log) and year as a random term. Years without available data for any of the response variables (n = 2; table 2s in supplementary material) were not considered in our modelling. Correction for small sample size of Akaike's index (AICc) was performed, and models with a difference in AICc of less than 2 units were considered to be statistically equivalent (Burnham and Anderson, 2002). The model with the lowest AICc value was considered the most parsimonious. Analyses were conducted using the R software environment (R Core Team, 2019).

Twenty models were considered. Models with time lags of up to 6 years relating surface area burned regionally were considered to account for possible delays in the response of Montagu's harriers to generation of suitable breeding habitat inland (shrublands) after forest fires (table 3s in supplementary material). Geometric growth rates (λ) were calculated by means of the equation for discrete exponential growth (λ = (Nt/N0) (1/t)). Results The coastal harrier population increased from 12 pairs in 1992 to 37 pairs in 1999 (geometric growth rate, λ = 1.15; 15 % annual increase). However, this increase was followed by a steep decline to just 2 pairs in 2017 (λ = 0.86; 14 % annual decrease) despite legal protection of the site since 1989. For comparison, the annual population growth rate of the inland Montagu's harrier population was 1.1 (10 % annual increase) for 1991–2016. Figure 2 shows the contrasting patterns of change in the number of pairs between these two populations.

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Number of pairs inland

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60 10

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0 1992

1991

Number of pairs coastal wetland

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Fig. 2. Annual number of pairs of Montagu's harrier (Circus pygargus) on the inland (shrubland) population of the study region (Comunidad Valenciana) during the period 1991-2016 (solid line and solid black dots), compared to the growth of the coastal wetland population (broken line and solid black dots). Data source: Conselleria de Agricultura, Desarrollo Rural, Emergencia Climática y Transición Ecológica, Generalitat Valenciana. Fig. 2. Número anual de parejas de aguilucho cenizo (Circus pygargus) en la población interior (arbustiva) de la región de estudio (Comunidad Valenciana) durante el período 1991-2016 (línea continua y puntos negros), en comparación con la evolución temporal de la población del humedal costero (línea discontinua y puntos negros). Fuente de los datos: Conselleria de Agricultura, Desarrollo Rural, Emergencia Climática y Transición Ecológica, Generalitat Valenciana.

The surface of land devoted to fruit tree cultivation around the protected marsh remained approximately stable from 1990 to 2000 (ca. 4,750 ha). It decreased to 3,300 ha from 2000 to 2006 and to 2,800 ha from 2006 to 2012. Fruit trees were mostly transformed into a diverse crop mosaic (increase from 230 ha to 1,280 ha from 1990 to 2012) and only a small proportion of the land was abandoned (increase from 7 to 266 ha for the same period as above). The results of our modelling (table 2) showed that the two most parsimonious models included wild boar density (during the previous year) together with the number of Montagu's harrier pairs breeding inland during the focal year (relative model probability of 0.51). These were followed by the model including only wild boar density during the previous year (relative model probability of 0.29). Differences in AICc for all the other models were higher than 2. Multi–model inference indicated that models containing wild boar density as an explanatory variable had a cumulated relative probability of 0.9.

Figure 3 shows the inverse linear relationship (r = –0.73; 95 % CI –0.45, –0.87) between the annual number of coastal harrier breeding pairs and the density of wild boars during the previous year. The annual population growth rate of Montagu's harriers showed a decrease (λ = 0.66; 14 % annual decrease) from 2007–2017 when annual regional hunting bags were above 4,000 wild boars. Our modelling did not detect any effect of the local number of western Marsh Harrier pairs (considered either quantitatively or qualitatively), forest fires at the regional level (with several time delays explored) or the amount of precipitation in April. Discussion Major changes in agricultural practices did not coincide in time with major changes in harrier trends. Whereas surface devoted to fruit tree cultivation remained roughly stable from 1990 to 2000 (4,740 ha

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Wild boars Harriers

7,000

5,000

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Fig. 3. Number of Montagu's harrier breeding pairs at our coastal study site (black solid line) over time in relation to regional wild boar hunted during the previous year (t–1) (dotted dash line) for the period 1992–2017. Data source: Conselleria de Agricultura, Desarrollo Rural, Emergencia Climática y Transición Ecológica, Generalitat Valenciana. Fig. 3. Número de parejas reproductoras de aguilucho cenizo en nuestro lugar de estudio costero (línea negra) a lo largo del tiempo con respecto a los jabalíes abatidos en la región durante el año anterior (t–1) (línea punteada) en el período 1992–2017. Fuente de los datos: Conselleria de Agricultura, Desarrollo Rural, Emergencia Climática y Transición Ecológica, Generalitat Valenciana.

in 1990 vs. 4,726 ha in 2000), harriers experienced an abrupt change in trend from 2000 on. Additionally, the fruit tree surface declined 30 % from 2000 to 2006, and harriers decreased from 35 pairs in 2000 to only 9 in 2005 and 17 in 2006. Even if fruit trees are instrumental as a source of passerine chicks for harriers, changes in agricultural crops alone cannot explain the change in population trend from a continuous increase to a rapid decline. The crop patchwork and abandoned lands that substituted the orchards were most likely a good source of food for harriers too (mostly passerines, but also large insects, lizards and small mammals according to the results of an unpublished study performed by the park staff from 1997 to 2004; Ros and Tena, 2004). Since changes in the foraging habitat could not explain the observed decline in the harrier population we focused on the analysis of perturbations in the nesting habitat. By the early 1990s, wild boars were still scarce in the region. Their overall low abundance most likely allowed the settlement and rapid growth of the Montagu's harrier population, as has been suggested previously for similar areas (Väli, 2017). However, the population growth and expansion of wild boar over the last

15 years could have caused the rapid decline in the coastal colony of Montagu's harrier. The role of wild boars as nest predators on ground–nesting birds in Mediterranean coastal wetlands has been previously reported (Herrero et al., 2004), and their predation of harrier nests has been recorded at the study site, with at least seven nests predated from 2007 to 2015 (Staff of the Cabanes–Torreblanca Natural Park, own unpublished data). The mechanism by which wild boar could have impacted the wetland population might involve permanent dispersal of birds to other breeding populations where the wild boar density or their impact on nests is lower. In effect, many protected wetlands in Spain have become an ecological refuge for wild boars because water, food and shelter are guaranteed therein (Galán, 2015, Barasona et al., 2021). According to unpublished data from the regional government (based on surveys targeting hunters from 2015 to 2020), densities of wild boars killed in the hunting states located inside the coastal nature park (where wild boars were absent 30 years ago) were ca. four times higher than those in adjacent hunting areas outside the coastal wetland within the same munici-

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Table 2. Multiple hypotheses testing of the change in number of Montagu's harrier breeding pairs (N) over time as a function of several explanatory variables (see table 1). Only the three most parsimonious models are shown. The remaining set of models tested are shown in table 3s in supplementary material: K, number of estimable parameters; AICc, Akaike's Information Criterion corrected for small sample size; LL, Log likelihood; ΔAICc, difference in AICc between each model and the most parsimonious model; wi, Akaike's weight; r2, variance–function–based coefficient of determination. The best models are highlighted in bold. Models with ΔAICc < 2 were considered statistically equivalent. Tabla 2. Múltiples hipótesis para explicar el cambio en el número de parejas reproductoras de aguilucho cenizo (N) a lo largo del tiempo en función de varias variables explicativas (véase la tabla 1). Sólo se muestran los tres modelos más parsimoniosos. El resto de los modelos se muestra en la tabla 3s del material complementario: K, número de parámetros estimables; AICc, criterio de información de Akaike corregido para un tamaño muestral reducido; LL, logaritmo neperiano de la verosimilitud; ΔAICc, diferencia en el AICc entre cada modelo y el modelo más parsimonioso; wi, pesos de Akaike; r2, coeficiente de determinación. Los mejores modelos se destacan en negrita. Los modelos con una ΔAICc < 2 se han considerado estadísticamente equivalentes.

Hypotheses

AICc

ΔAICc

N~Wildb1 + Montagh

190.81

-91.80

0.51

0.53

N~Wildb1

191.96

-93.69

1.15

0.29

0.49

N~Wildb1 + Montagh1

194.01

-93.41

3.20

0.10

0.50

palities (10.6 individuals/km2 vs. 2.7 individuals/km2) (C. Gómez, Conselleria de Agricultura, Desarrollo Rural, Emergencia Climática y Transición Ecológica, pers. com.). Dispersal could have been triggered by the continuously decreasing breeding success of harriers (see fig. 1s in supplementary material), likely caused by wild boar directly predating on nests. Conflicts between predator abundance and harriers' breeding success have been described previously (Simmons, 2000; Wiącek, 2015; Ludwig et al., 2016; Väli, 2017). Cases of abandonmnet of territories and dispersal, motivated by successive failures caused by predation, have been reported previously in harriers (Soutullo et al., 2006; McMillan, 2014 ) and other social bird species (Fernández–Chacón et al., 2013; Payo–Payo et al., 2017 ). Moreover, rapid population responses of Montagu's harriers to disturbance have been described in the study region (Oro et al., 2012). Wild boar population growth and expansion represent a regime perturbation theory with a likely threshold density value for prey such as harriers to disperse and leave the coastal wetland. This process may promote decision–making in dispersal with positive density dependence via social copying and the additional impact of the Allee effect on anti–predatory defence (Arroyo et al., 2004; Kitowski, 2008; Wiącek, 2015). As a consequence, runaway dispersal of harriers from the coastal wetland would follow a non–linear pattern after a tipping point of progressive deterioration of environmental conditions caused by wild boar (Oro, 2020). Because we only marked chicks in the monitoring program, we do not have data concerning the movements of adult birds (breeding dispersal). Howe-

ver, there is evidence of dispersal of birds from our coastal study population to other harrier populations, as three birds ringed as chicks (two females and one male) were observed breeding in inland populations four and six years after ringing. Moreover, three individuals (also ringed as a chicks) performed long–distance dispersal and were observed during the breeding season hundreds of kilometers away in Loja (Granada), Totana (Murcia) and Castuera (Badajoz), two, three and five years after marking, respectively (Conselleria de Agricultura, Desarrollo Rural, Emergencia Climática y Transición Ecológica, internal reports). Although breeding dispersal is likely to be less common than natal dispersal (which is known to be high in the focal species; see Limiñana et al., 2011), long distance movements (> 100 km) of females have been reported following breeding failure (Arroyo et al., 2004). Dispersal could have targeted areas with lower levels of wild boar predation, which could be the case of the dense shrublands of kermes oak. Montagu's harriers are known to select dense vegetation habitats for nesting as protection from predators (Claro, 2000, Limiñana et al., 2006). Dense shrublands could explain the growth of inland harrier populations despite the regional population growth of wild boars. Importantly, the differential trend of coastal and inland colonies also suggests that mortality in wintering grounds cannot explain the decline of the coastal colony (fig. 2). The local trend of other ground–nesting species in the protected wetland such as the collared pratincole (Glareola pratincola) showed a marked decline as of 2006, coinciding with the rapid increase in numbers of wild boars (r = –0.64; 95 % CI –0.32,

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–0.85). However, this was not the case for other local ground–nesting species (the black–winged stilt Himantopus himantopus) This difference, however, could be explained by the different nesting habitats of each species. Pratincoles nest on cobble dunes by the sea front whereas stilts nest on flooded terrain, which may provide some protection against wild boar predation, as in the case of western marsh harriers. The parallel growth over time of collared pratincoles and Montagu's harriers is shown in figure 2s supplementary material. The impact of wild boar on ground–nesting waterbirds, including pratincoles, has recently been reported (Barasona et al., 2021). Their impact on other vertebrate groups, such as amphibians and reptiles, has also been reported in the literature recently (Jolley et al., 2010; Barrios–García and Ballari, 2012; Galán, 2012, 2015; Carpio et al., 2016; Graitson et al., 2018; Barasona et al., 2021). Moreover, predation by wild boar and naturalized domestic pigs has been recognized as a major risk factor for imperiled species worldwide (Massei and Genov, 2004; Engeman et al., 2016; McClure et al., 2018; Wehr et al., 2018). Finally, our results highlight that predation of ground–nesting bird species by wild boars may be habitat–specific, and that dispersal to lower–predation sites or habitats is likely a major buffering mechanism to avoid nest predation (see Barros et al., 2016 for an example with seabirds). In summary, our results illustrate a sequence of changing scenarios associated with changes in the landscape and in human use of the territory at a regional level. The abandonment of agriculture and livestock farming and the early protection of coastal wetlands (in the 1980s) favoured the colonization of these areas by Montagu's harriers as they found refuge therein. Subsequently, shrub and tree encroachment favoured the expansion of wild boar outside their mountain refuges, until they recolonized the coastal areas, from where intense human activity (i.e. agriculture, livestock, hunting) likely expelled them centuries ago. The protection of coastal wetlands, including the exclusion of traditional human activities such as livestock raising and vegetation burning, has led to the recovery of spontaneous vegetation and to the growth of coastal wild boar populations, which thrive in humid areas. The movement of both harriers and wild boars out of their respective historical refuges may largely explain the rapid and unexpected changes that have occurred at the local study site and at other similar protected wetlands in the region. Acknowledgements Field work was initially carried out by V. Tena and X. del Señor, and was continued for next twenty years by G. Ros (ranger) and J. Tena (Prat de Cabanes–Torreblanca Natural Park staff). Without their commitment the data analyzed in this study would not have been available. We are also grateful to Pilar Santidrián and Helen Regan who commented on an early draft of the manuscript, and to Catherine Andrés who helped build figure 1. Daniel Oro suggested some

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Natural factors but not anthropogenic factors affect native and non–native mammal distribution in a Brazilian National Park R. A. Duarte Silveira, H. H. Marques da Rosa, A. A. Pereira, M. Passamani, R. D. Zenni

Duarte Silveira, R. A., Marques da Rosa, H. H., Pereira, A. A., Passamani, M., Zenni, R. D., 2021. Natural factors but not anthropogenic factors affect native and non–native mammal distribution in a Brazilian National Park. Animal Biodiversity and Conservation, 44.2: 241–250, Doi: https://doi.org/10.32800/abc.2021.44.0241 Abstract Natural factors but not anthropogenic factors affect native and non–native mammal distribution in a Brazilian National Park. Protected areas, designed for biodiversity conservation, are currently affected by invasive species as most of them have documented biological invasions. This study aimed to test whether non–native mammal species richness influences the local distribution of native mammals and how distance from human settlement, elevation and vegetation characteristics influence native and non–native mammal richness in a national park in Brazil. We recorded 20 mammal species in the park, 17 native species and three non–native species. Native mammal richness was higher at intermediate elevations and in forests with lower tree densities and tree basal area. Non–native mammal richness was greater at higher elevations and in forests with low tree densities. Non–native mammals did not influence native mammal presence. In conclusion, the distribution of both native and non–native mammal species was affected by elevation and vegetation but not by distance from human settlements or non–native mammal presence. Key words: Biological invasions, Domestic animals, Itatiaia National Park, Protected areas, Wild boar Resumen Son factores naturales y no antropogénicos los que afectan a la distribución de mamíferos autóctonos y alóctonos en un parque nacional del Brasil. En la actualidad, las zonas protegidas, que están concebidas para la conservación de la biodiversidad, se ven afectadas por especies invasoras, ya que en la mayoría de ellas se han documentado invasiones biológicas. Con el presente estudio tratamos de comprobar si la riqueza de especies de mamíferos alóctonos incide en la distribución local de mamíferos autóctonos y determinar la influencia de la distancia a asentamientos humanos, la altitud y las características de la vegetación en la riqueza de mamíferos autóctonos y alóctonos en un parque nacional del Brasil. Registramos 20 especies de mamíferos en el parque, de las que 17 eran autóctonas y tres, alóctonas. La riqueza de mamíferos autóctonos fue mayor en altitudes intermedias y en bosques poco densos y con escasa área basimétrica. La riqueza de mamíferos alóctonos fue mayor en altitudes intermedias y en bosques poco densos. Los mamíferos alóctonos no influyeron en la presencia de los mamíferos autóctonos. En conclusión, la altitud y la vegetación son los factores que afectaron a la distribución de las especies de mamíferos autóctonas y alóctonas, y no la distancia a asentamientos humanos. Palabras clave: Invasiones biológicas, Animales domésticos, Parque Nacional de Itatiaia, Zonas protegidas, Jabalí Received: 7 XII 20; Conditional acceptance: 27 IV 21; Final acceptance: 30 VI 21 Raphaela Aparecida Duarte Silveira, Hugo Henrique Marques da Rosa, Adriele Aparecida Pereira, Marcelo Passamani, Rafael Dudeque Zenni, Programa de Pós–Graduação em Ecologia Aplicada, Departamento de Ecologia e Conservação, Universidade Federal de Lavras, Campus Universitário, CEP 37200–900, Lavras, MG, Brasil. Corresponding author: R. A. Duarte Silveira. E–mail: rapha_24@hotmail.com ORCID ID: 000-0002-6044-9294

ISSN: 1578–665 X eISSN: 2014–928 X

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Introduction The introduction of non–native species is a cause of great concern among conservation biologists (Bellard et al., 2016; Seebens et al., 2017). Invasive species can have direct and indirect economic, environmental and social impacts (Charles and Dukes, 2007). Such impacts can threaten biological diversity through competition, predation, disease transmission, hybridization, physical disturbance of the environment, and destruction of crops and pastures (Doherty et al., 2015; Gompper, 2014; Paini et al., 2016; Rosa et al., 2019; Wyatt et al., 2008). Among non–native species, mammals are one of the groups that cause the most damage to global diversity and their strong impact as an invasive species has already been demonstrated (Bellard et al., 2016; Blackburn et al., 2004; Doherty et al., 2015). Mammals such as dogs, cats, horses and cattle, wild boar, primates, opossums and hares (Long, 2003) have been introduced into new countries around the world for a number of reasons, such as for hunting game, for biological control, and for domestication and commercialization as livestock or pets (Long, 2003). Their high impact capacity, such as competition with native species, disease and pathogen transmission, hybridization, genetic changes, and damage to crops, may be the result of their high ecological plasticity and great capacity for habitat modification (i.e. ecosystem engineers) (Jones et al., 1994; Long, 2003). The creation of protected areas is an important strategy to maintain habitat integrity and conserve biodiversity (Gray et al., 2016) but protected areas have also been affected by the introduction of invasive species. Invasive mammals worldwide have a history of impact in protected areas in relation to predation, competition, pathogen transmission, soil disturbance and exposure, vegetation damage, and non–native seed dispersal (Ballari et al., 2014; da Rosa et al., 2017; Lessa et al., 2016; Parsons et al., 2016). Moreover, the presence of non–native species in protected areas is often associated with human presence since most protected areas are either located near urban centers, or humans take non–native species to these places with them (Paschoal et al., 2018). As the presence of these animals in protected areas may then reduce the effectiveness of biodiversity conservation strategies, knowing their distribution across those areas is fundamental in order to manage biological invasion. To determine the distribution of these mammals across areas we can use environmental factors, human occupation, and vegetation traits (Ahumada et al., 2011; Dias et al., 2019; Lyra–Jorge et al., 2009; Pereira, 2017; Sampaio et al., 2010). Regarding environmental factors, the structure of the landscape is an important variable for mammal communities (Lyra– Jorge et al., 2010; Sampaio et al., 2010). There is less species richness and functional diversity and higher dominance in highly fragmented sites than in partially fragmented sites and continuous forest landscapes (Ahumada et al., 2011). It has been observed that forest cover and management intensification affect

the distribution of mammals in a cacao agroforestry system (Cassano et al., 2014). However, regarding human occupation, probability of occupancy of carnivores can be influenced by the distance to forest, human infrastructure, watercourses, and the proportion of anthropized areas (e.g pasture, crops) (Cruz et al., 2019; Dias et al., 2019). It has also been seen that areas with higher human occupancy had lower species richness, with omnivorous and insectivorous species being the most common species (Bogoni et al., 2016). Vegetation traits can also influence mammal communities. Focal tree connectivity and canopy cover are most likely the most important predictors of occupancy for the arboreal community whereas forest loss and canopy height are the strongest predictors for the terrestrial mammal community (Whitworth et al., 2019). Information about non–native mammal distribution across protected areas and environmental and other factors affecting their distribution is necessary so as to develop effective control and management measures. The aim of this study was to test whether distance from human settlements, elevation, absolute tree density, absolute tree coverage, mean tree basal area, mean tree height, and mean tree canopy cover affected the richness of native and non–native mammals in a protected area in Brazil, the Itatiaia National Park. We sought to answer the following questions: (a) do non–native mammals in the Park affect native mammal presence?; (b) does non–native mammal richness decrease the greater the distance from human settlements?; (c) does native mammal richness increase the greater the distance to these settlements?; and (d) which environmental variables influence native and non–native mammal distribution? Material and methods Study area We conducted the study in the Itatiaia National Park (22º 22' 31'' S 44º 39' 44'' W, fig. 1), a strict protection conservation area, meaning only the indirect use of its natural resources is allowed and it cannot be inhabited by humans (Brasil, 2011). Inserted in the Atlantic forest hotspot domain, in the Mantiqueira Complex, the Itatiaia National Park was created in 1937. It was the first area in Brazil to be given protected status. It comprises four municipalities, Bocaina de Minas and Itamonte in Minas Gerais State, and Resende and Itatiaia in Rio de Janeiro State. The Park is non–officially divided in two main parts, the highland and the lowland. The highland has about 16,395 ha, with an altitude ranging from 1,500 m to 2,791 m a.s.l., a type of climate Cwb (temperate climate with dry winter and warm summer), mean temperature ranges from 8.2 ºC to 13.6 ºC, and annual precipitation of about 2,600 mm. There are small rural producers living within the boundaries of the Park. The lowland, where the Park administration is located has an area of 6,414 ha. The altitude ranges from 540 m to 1,500 m a.s.l., and the climate is Cwa type (temperate with dry winters

243

22º 15' 0''S

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Pacific Ocean

RJ Lowland

N 0 1 2

Highland

22º 25' 0''S

Settlement Sampling points Trails Itatiaia NP State borders

22º 20' 0''S

Atlantic Ocean

4 km

44º 50' 0'' W

44º 45' 0'' W

44º 40' 0'' W

44º 35' 0'' W

44º 30' 0'' W

Fig. 1. Location of sampling points (black crosses) in two trails of the highland and three trails of the lowland of the Park. Black triangles indicate the nearest human settlement from the first camera–trap in each trail. The highland has about 16,395 ha, with an altitude range from 1,500 m to 2,791 m a.s.l., a Cwb type of climate (temperate climate with dry winter and warm summer), mean temperature ranges from 8.2 ºC to 13.6 ºC, annual precipitation is about 2,600 mm, and there are small rural producers living within the boundaries of the Park. The lowland, where the Park administration is located, has about 6,414 ha, with an altitude range from 540 m to 1,500 m a.s.l., a type of climate Cwa (temperate climate with dry winter and hot summer), mean temperature ranges from 20 ºC to 24 ºC, the annual precipitation is about 1,800 mm and the main occupation is related to cottages available for holidays and weekends and summer houses. Fig. 1. Ubicación de los puntos de muestreo (cruces negras) en dos pistas en la Parte Alta y tres pistas en la Parte Baja del Parque. Los triángulos negros indican la construcción antrópica más cercana a la primera trampa en cada pista. La Parte Alta tiene alrededor de 16.395 ha, un rango de altitud de 1.500 m a 2.791 m s.n.m. y un tipo de clima Cwb (clima templado con invierno seco y verano cálido). La temperatura media varía de 8,2 ºC a 13,6 ºC y la precipitación anual es de unos 2.600 mm. Asimismo, hay pequeños productores rurales que viven dentro de los límites del Parque. La Parte Baja, donde se ubica la administración del Parque, tiene alrededor de 6.414 ha, con un rango de altitud de 540 m a 1.500 m s.n.m. y un tipo de clima Cwa (clima templado con invierno seco y verano caluroso), la temperatura media varía de 20 ºC a 24 ºC, la precipitación anual es de unos 1.800 mm y el principal tipo de ocupación son cabañas para vacaciones y fines de semana y casas de veraneo.

and hot summers). Mean temperature ranges from 20 ºC to 24 ºC and the annual precipitation is about 1,800 mm. The main occupation is related to weekend and holiday accommodation (Barreto et al., 2014; Köppen, 1936; Tomzhinski, 2012). The Park has various phytophysiognomies, ranging from high altitude grasslands (above 2,000 m), dense ombrophilous forest (submontane, montane and high–montane), montane mixed ombrophilous forest (with the presence of Araucaria trees) and montane semideciduous forest

(Barreto et al., 2014). An issue of concern in Itatiaia National Park is related to land tenure regularization. This is a problem not only in the this park but also in many strict protection conservation areas (Cheade, 2015; INEA, 2010). Only 51% of the Park area was regularized until 2016 (ICMBio, 2016). Some private landowners still have properties and even live within the boundaries of the Park. This is because they did not receive compensation for the land by the federal government. In the remaining 49 % of non–regularized

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areas, there are summer houses and hotels, mainly in the lowlands, and small rural producers with grazing being the main economic activity in the highlands (Barreto et al., 2014). Mammal data collection To survey the mammals of the Itatiaia National Park we installed camera–traps in the forest at 17 sampling points, 50 m perpendicular to man–made trails in the highland and lowland. Each sampling point was positioned at a minimum distance of 150 m from the closest human settlement and at a distance of 500 m from each other in a straight line (fig. 1). Human settlements in the highland consisted of inhabited houses without large crops but with backyards covered in grasses. In the lowland, these settlements are touristic points. There is a hotel with high human traffic, and an abandoned hotel with low or no human traffic. Seven camera–traps were installed along two trails (Picu and Araucárias trails) in the highland, whereas in the lowland we installed ten camera–traps in three trails (Três Picos, Rui Braga and Hotel Donati). These cameras remained in the field from September 2018 to December 2018 in the highland and from December 2018 to March 2019 in the lowland sites. Each sampling point had one camera–trap (Bushnell©, Digital Hunting Camera© or Trail Camera©) placed on trees larger than 15 cm in diameter and approximately 45 cm from the ground (Srbek–Araujo et al., 2012). The cameras were active day and night and were set up to take three pictures every 30 seconds, whenever the motion sensor was triggered. To avoid selection of species we did not use baits (Srbek–Araujo et al., 2012). Every 45 days approximately we revisited the cameras to collect data, adjust equipment and replace batteries. After the collection period, the photos were analyzed. To avoid repetition of data with the same individual when the individualization was not possible, we considered all the photographs with intervals of at least one hour as new independent records (Srbek–Araujo et al., 2012). We included all mammals able to be photographed and identified. We therefore included some unrestricted arboreal mammals, such as the Southeastern four–eyed opossum Philander frenatus and the Southeastern squirrel Guerlinguetus ingrami. It is important to mention that the records of species frequency can depend on various factors at a site. The location of the cameras can have a strong influence on this variable and even bias the probability of detection of species (Di Bitetti et al., 2014). However, we followed a pattern among our sampling sites, installing camera traps within the forests and near animal–made trails only, trying to equalize detection probability between sites. Environmental data collection To test whether environmental factors and other factors influenced native and non–native mammal distributions, we measured elevation, absolute tree density, absolute tree coverage, mean tree basal area, mean tree height and mean tree canopy cover at each

sampling point (table 1). To measure the proportion of tree canopy cover, we made a 50 m transect in the north–south direction from the camera trap and measured the percentage of canopy every 10 m using the CanopyApp © version 1.0.4. We used the average proportion of tree canopy cover as a predictor variable for each sampling point. Elevation was obtained using a Garmin 62s GPS and the distance from the sampling point was measured to the nearest human settlement using the ruler function in Google Earth Pro software version 7.3. To obtain the predictor variables absolute tree density, absolute tree coverage, mean tree basal area and mean tree height, we used the point–centered quarter method (Mitchell, 2007). For this method, we formed a circle divided into four quadrants using the sampling point as center. The four–quadrants were divided followed the main cardinal points (North, South, East and West). For each quadrant, we found the closest tree from the sampling point and measured the height, the circumference at breast height and the distance from the tree to the center. We repeated this procedure for the other three quadrants. Using these measures, we calculated absolute tree density, absolute tree coverage, mean tree basal area, mean tree height and mean tree canopy cover based on the methodology described in Mitchell (2007). Data analysis To test if non–native mammals had influenced the presence of native mammals in the Park, we ran a generalized linear model (GLM) with Quasipoisson error distribution using the response variable native mammal richness and the predictor variable non–native mammal richness. We ran a set of GLM with Quasipoisson error distribution to test how the non–native and native mammal richness related to each of the environmental variables. First, in order to see how mammals responded to anthropogenic influence, we performed two GLM considering non–native mammal richness and native mammal richness as response variables and distance to the nearest human settlement as a predictor variable (table 1). Second, we ran another two GLM with Quasipoisson error distribution using the same response variables of the first set of analyses to see if they responded to the environmental factors. Predictor variables were investigated together in the models; these were elevation, absolute tree density, absolute tree coverage, mean tree basal area, mean tree height, and mean tree canopy cover. Spearman's correlation matrix was used to identify the variables with a strong correlation, i.e. ρ > 0.7. Absolute tree density and mean tree canopy cover were strongly correlated, the latter being excluded from the analysis because tree diversity appears to be more important for mammals (Pereira, 2017). Next, we performed model selections using the Akaike information criterion modified for small samples (AICc), considering equally plausible those models with AICc < 2 to compare the relative importance of environmental variables (Burnham and Anderson, 2002). If we had more than one selected model with AICc < 2 explaining the response variables and more than one variable in each model,

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Table 1. Environmental variables collected at each sampling point in the highland and lowland of the Itatiaia National Park: L, location; T, trail (3P, 3 Picos; RB, Rui Braga; HD, Hotel Donati; PI, Picu; AR, Araucárias); D, distance from human settlement (in m); E, elevation (in m); Atd, absolute tree density (tree/ha, number of trees per area); Atc, absolute tree coverage (m2/ha, occupied area by trees per unit area); Mtba, mean tree basal area (cm2, transversal area of the tree trunk); Mth, mean tree height (in m); Mtcc, mean tree canopy cover (%, overlapping proportion of tree branches and leaves). Tabla 1. Variables ambientales determinadas en cada punto de muestra de la Parte Alta y la Parte Baja del Parque Nacional de Itatiaia: L, ubicación; T, pista (3P, 3 Picos; RB, Rui Braga; HD, Hotel Donati; PI, Picu; AR, Araucárias); D, distancia del asientamiento humano (en m); E, altitud (en m); Atd, densidad absoluta de árboles (tree/ha, número de árboles por hectárea); Atc, cobertura absoluta de árboles (m2/ha, superficie ocupada por árboles por hectárea); Mtba, área basal media de los árboles (cm2, promedio de la sección transversal del tronco); Mth, altura media de los árboles (en m); Mtcc, cubierta media del dosel (%, proporción de superposición de las ramas y las hojas de los árboles).

D L T (m)

E (m)

Atd (tree/ha)

Atc (m²/ha)

Mtba (cm²)

Mth (m)

Mtcc (%)

Lowland

HD 1,075.64 1,082 5,548.44

6.67 12.02 4.45 79.35

Lowland

HD 598.38 1,012 5,175.71 102.2 197.46 9.25 67.42

Lowland

HD 165.45 954 12,075.84 33.83 28.01 5.25 58.72

Lowland

3P 329.45 1,108 5,548.47 33.54 60.45 5.75 63.04

Lowland

3P 827.54 1,220 779.16 15.72 201.77 7.75 45.19

Lowland

3P 1,166.95 1,228 11,138.88 129.85 116.57 5.75 62.08

Lowland

RB 779.11 1,178 5,327.93 85.12 159.76 6.38 59.84

Lowland

RB 898.63 1,365 7,901.23 38.04 48.14 4.5 61.22

Lowland

RB 1,407.67 1,501 4,643.48

6.09 13.12 3.25 56.6

Highland

199.95

1,903

3,894.07

5.98

15.35

3.5

57.82

Highland

672.69

1,893

1,242.02

7.33

5.68

37.49

Highland

1,112.1

1,896

783.53

8.3

106.01

6.88

38.36

Highland

1,588.33

1,973

2,246.13

23.32

103.83

5.75

32.94

Highland

AR 162.77 1,977 5,590 28.56 51.09 4.88 59.6

Highland

AR 992.77 2,212 4,549.99 11.2 24.62 5.38 56.87

Highland

AR 1,474.39 2,236 4,772.69 39.38 82.51 6.5 42.7

298

1,159 2,148.32 230.72 1073.94 2.38 51.66

we performed the Relative Importance of Regressors in Linear Models, using the package 'relaimpo' version 3.5.3. This analysis quantifies which of these predictor variables of the selected models were more important, based on the higher percentage of explanation to the response variable. All analyses were performed in R version 3.5.0, using packages 'ggplot2', 'MuMIn', 'lme4' and 'relaimpo'. The tests were considered significant at p < 0.05. Results With a sampling effort of 1,543 trap–days, we recorded 20 mammal species, three non–native and 17 native species. In the highland we recorded seven

native mammals (Brazilian common opossum Didelphis aurita, spotted–paca Cuniculus paca, striped hog–nosed skunk Conepatus semistriatus, tayra Eira barbara, lesser grison Galictis cuja, southern tiger cat Leopardus guttulus, tapeti Sylvilagus brasiliensis) and three non–native mammals, one wild and two domestic animals (wild boar Sus scrofa, domestic dog Canis lupus familiaris and domestic cattle Bos taurus) In the lowland, we recorded 17 native (D. aurita, C. paca, C. semistriatus, E. barbara, G. cuja, L. guttulus, S. brasiliensis, white–lipped peccary Tayassu pecari, southeastern squirrel Guerlinguetus ingrami, nine–banded armadillo Dasypus novemcinctus, crab–eating fox Cerdocyon thous, ocelot Leopardus pardalis, South American coati Nasua nasua, collared peccary Pecari tajacu, southeastern

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Number of records

Bos taurus Canis lupus familiaris Sus scrofa

30 20 10 0 400 800 1,200 1,600 Distance to nearest human settlement (m)

Fig. 2. Number of records of wild boar Sus scrofa, domestic cattle Bos taurus and domestic dogs Canis lupus familiaris in relation to distance of camera–traps from human settlement in the highland in the Itatiaia National Park. The points, crosses and x represent the sampling points where we installed the camera–traps. Fig. 2. Numero de registros de jabalíes Sus scrofa, ejemplares de ganado doméstico Bos taurus, y perros domésticos Canis lupus familiaris, en relación con la distancia de las trampas fotográficas situadas en las construcciones antrópicas en el Parque Nacional de Itatiaia. Los puntos, los triángulos y los cuadrados representan los puntos de muestreo donde se instalaron las cámaras.

four–eyed opossum Philander frenatus, Paraguayan hairy dwarf porcupine Coendou spinosus, Leopardus sp.) and no non–native species (appendix 1s and 3s). Non–native mammals did not affect the presence of native mammals (p = 0.3). We only had records of the non–native S. scrofa at distances of 1 km or more from human settlements (fig. 2). The same did not happen for B. taurus and C. lupus familiaris. We recorded B. taurus near and far from human settlements (fig. 2) whereas C. lupus familiaris were found at distances of 200 and 1,500 m only (fig. 2). Despite these differences, distances to human settlements did not influence either native (p = 0.6) or non–native mammal richness (p = 0.1). Native mammal richness decreased from eight to four, six to four, eight to zero and four to zero species as elevations increased (1,178 to 1,501 m; 1,082 to 1,501 m; 1,178 to 2,236 m; 1,903 to 2,236 m, respectively). Native mammal richness also decreased, from eight to four and four to zero in forests where mean tree basal area (159.76 to 201.77 cm² and 15.35 to 1,073.94 cm²) was higher, and from eight to four and four to one in forests where absolute tree density (5,327.93 to 7,901.23 tree/ha and 2,246.13 to 12,075.84 tree/ha) and mean tree height (6.38 to 7.75 m and 3.5 to 9.25 m) were also higher (p < 0.001). In addition, native mammal richness increased from one to five in forests with greater absolute tree coverage (33.83 to 129.85 m²/ha; p < 0.001). The AICc test selected three models that best explained native mammal richness (appendix 2s).

However, according to the relativeimportance of regressors in linear models (relaimpo), elevation had more than half of the influence (61 %) in native mammal richness followed by mean tree basal area (32 %) and absolute tree density (7 %, fig. 3). Non–native mammal richness decreased from three to zero species with an increase in tree density (3,894.07 to 5,590.0 tree/ ha) and in forests with higher trees (3.5 to 6.5 m; p < 0.001) and increased from zero to three species with an increase in elevation (954 to 1,973 m; p < 0.001). The AICc test selected two models that better explained non–native richness (appendix 2s). However, according to the relaimpo package, elevation influenced 81 % whereas absolute tree density influenced only 19 % in non–native richness (fig. 3). Discussion Native mammal richness did not increase the greater distance to human settlements. Also, non–native mammal richness did not increase the closer to these settlements. The way it is written is different from the original question. It was expected to increase native mammal richness and decrease non-native mammal richness the further from human settlements. These results may be explained by the fact that these species have more extensive home range sizes than the distance we analyzed, moving long distances in search of resources (Kasper et al., 2016; Pereira, 2017). Moreover, most of the recorded species are capable

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% of R2

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40 20

Mtba

Atd

Fig. 3. Relative importance of environmental variables for native (A) and non–native (B) mammal richness at Itatiaia National Park. Bar plots of method lmg were built with 95 % bootstrap confidence interval (lines in the bars). Variables are elevation (E), mean tree basal area (Mtba) and absolute tree density (Atd). Fig. 3. Importancia relativa de las variables ambientales para la riqueza de mamíferos nativos (A) y exóticos (Ba) en el Parque Nacional de Itatiaia. Las barras del método lmg se construyeron con un intervalo de 95 % de confianza 'bootstrap'. Las variables son la altitud (E), el promedio del área basal media de los árboles (Mtba) y densidad absoluta de árboles (Atd).

of living both in native environments and in those with some kind of anthropogenic influence (Zanzini et al., 2018). We found more native mammal richness in intermediate elevations in the Park (954–1,501 m) as has also been reported in other studies (Brown, 2001; Geise et al., 2004; Moreira et al., 2009; Rosa et al., 2014). Native mammal richness was also higher in forests where mean basal area was low. Mammal occupancy patterns in line with our findings in relation to vegetation structure have been observed previously. For instance, Pecari tajacu occupancy patterns were found to be negatively related with low basal area of fruiting trees, indicating low productivity of these trees, and, consequently, few available resources (Thornton et al., 2011). Occupational patterns of other species were negatively associated with the basal area of small trees, which could indicate difficulty in moving and foraging in these areas (Thornton et al., 2011). However, other studies have found that vegetation traits such as vertical structure index, tree species diversity, percentage of forest and grassland (Andrade–Núñez and Aide, 2010), and forest cover (Ferreira et al., 2020) have influenced mammal species richness. Based on these studies and the fact that the occurrence of mammals can be influenced by tree fructification, we can assume that one of the reasons for the negative association between native mammals in Itatiaia National Park and mean tree basal area was due to the fact that we did not consider only fruiting trees in this study. Other factors may also have influenced the study. In Brazil, all but one of the 17 invasive mammal species reported in the literature for the country (Indian

sambar Cervus unicolor) are currently present in protected areas (da Rosa et al., 2017) and for the Itatiaia National Park we recorded three non–native mammal species: wild boar Sus scrofa, considered invasive in the Itatiaia National Park, domestic dogs C. lupus familiaris, and cattle Bos taurus, both considered casual in the Park (Ziller et al., 2020). Non–native mammal richness was greater at higher elevations and in forests with lower tree densities. Additionally, with the increase in elevation in the Park there is also a change in the vegetation's phytophysiognomy, from forest to native grasslands. Thus at higher elevations, forests are less dense, which may explain the increase of non–native richness in forests with lower tree density (Barreto et al., 2014). Moreover, the greater number of non–native mammal richness in high elevations can be due to being close to Itamonte neighborhoods, such as Fragária and Serra Negra. Some parts of these neighborhoods still share areas with the Park, facilitating access by domestic animals (Barreto et al., 2014). In addition, because of land tenure regularization issues, there are still residents living in this region within the Park (see methods). These residents have domestic animals such as dogs and cattle, which can range freely in the areas of the Park. Furthermore, we know that wild boar S. scrofa, prefer high elevations due to lack of sweat glands (Allwin et al., 2016), important in body thermoregulation. In this study, we recorded S. scrofa at minimum distances of 1 km from human settlements. This could be to avoid areas with human presence since hunting is still a current activity in the Park (Morais et al., 2019) and because this species is

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wild and not directly related to humans (Long, 2003). The absence of non–native mammals in the lowland may be related to the type of occupation being mainly related to cottages available for holidays and weekends and the presence of summer houses. Furthermore, the Park administration located there controls the entrance of domestic animals in this area (Barreto et al., 2014; Rosa et al., 2014). Domestic free–ranging animals are also an important issue for the conservation of biodiversity in the Park. The presence of domestic dogs C. lupus familiaris has also been reported in other protected areas (Paschoal et al., 2018). We found dogs were kept indoors in several residents' houses, which was a surprise as most dog owners allow them to roam freely (Gompper, 2014). Even so, as camera–traps have recorded some free–ranging dogs in the forest, the population needs to be aware of the consequences these animals can have on wildlife in the Park if they are released (Long, 2003) because the effects of a few individuals can be catastrophic, as in New Zealand, for example, where it was estimated that a single dog decimated approximately half of the North Island brown kiwi bird population (Taborsky, 1988). In addition to population consciousness, other measures such as vaccination and neutering could diminish the harm of free–ranging dogs to native communities (Lacerda et al., 2009). Domestic cattle B. taurus, which were found free ranging in the Park, have been identified elsewhere as a major cause of extinction of several native plants and animals (Gurevitch and Padilla, 2004). This is because the introduction of large herbivores into an environment imposes a new herbivore regime, especially due to different dietary patterns and body size (Hobbs and Huenneke, 1992), as well as degrading habitat due to grazing and trampling (Gurevitch and Padilla, 2004). The impact of domestic livestock can be negative, positive, or neutral, but with a tendency to negative impacts on vegetation, mainly related to conservation of vegetation structure, composition and dynamics (Mazzini et al., 2018). A study in a Patagonian forest showed that in places where cattle were alone, the impacts were higher than those caused by wild boar alone or than those in places where both were present (Ballari et al., 2020). Several studies have already shown that the impact of domestic animals can be significant. Therefore, as they are free ranging in the Park, they could cause disastrous consequences for native species as the impacts they have on this native community are as yet unknown. Conclusion The present study in the Itatiaia National Park showed that the diversity of mammal species was high, but it also revealed the presence of three non–native species. Although we found no evidence that non–native mammals influence the presence of native mammals at present, their potential threat to native biodiversity worldwide is well–known. The distribution of native and non–native mammals in the Park was affected by elevation and there were differences between wild

and domestic free–ranging animals in the way they use the areas of the Park. In summary, elevation and vegetation, but not distance from human settlements or presence of non–native mammals, affect the distribution of both native and non–native mammal species in the Itatiaia National Park. Acknowledgements The authors would like to thank the Itatiaia National Park and their team as well as the Alto Montana Institute for support during field activities. The authors also would like to thank Marcelo Motta for preparation of figure 1. Gustavo Hering and Paolo Ramoni assisted with statistical analysis. RADS and HHMR received support from the Coordenação de Aperfeiçoamento de Pessoal de Nível Superior –Brasil (CAPES)– Finance Code 001. RDZ acknowledges the support from CNPq–Brazil (grant 304701/2019–0). This study was funded by ICMBio and CNPq (grant 421254/2017–3). References Ahumada, J. A., Silva, C. E. F., Gajapersad, K., Hallam, C., Hurtado, J., Martins, E., McWilliam, A., Mugerwa, B., O'Brien, T., Rovero, F., Sheil, D., Spironello, W. R., Winarni, N., Andelman, S. J., 2011. Community structure and diversity of tropical forest mammals: data from a global camera trap network. Philosophical Transactions of the Royal Society B: Biological Sciences, 366: 2703–2711, Doi: 10.1098/rstb.2011.0115 Allwin, B., Swaminathan, R., Mohanraj, A., Suhas, G. N., Vedaminckam, S., Gopal, S., Kumar, M., 2016. The Wild Pig (Sus scrofa) Behavior – A Retrospective Study. Journal of Veterinary Science and Technoly, 7: Doi: 10.4172/2157-7579.1000333 Andrade–Núñez, M. J., Aide, T. M., 2010. Effects of habitat and landscape characteristics on medium and large mammal species richness and composition in northern Uruguay. Zoologia, 27(6): 909–917, Doi: 10.1590/S1984-46702010000600012 Ballari, S. A., Cuevas, M. F., Cirignoli, S., Valenzuela, A. E. J., 2014. Invasive wild boar in Argentina: using protected areas as a research platform to determine distribution, impacts and management. Biological Invasions, 17(6): 1595–1602, Doi: 10.1007/s10530-014-0818-7 Ballari, S. A., Valenzuela, A. E. J., Nuñez, M. A., 2020. Interactions between wild boar and cattle in Patagonian temperate forest: cattle impacts are worse when alone than with wild boar. Biological Invasions, 22: 1681–1689, Doi: 10.1007/s10530-020-02212-w Barreto, C. G., Campos, J. B., Roberto, D. M., Roberto, D. M., Schwarzstein, N. T., Alves, G. S. G., Coelho, W., 2014. Plano de Manejo. Accesible online at: http://www.icmbio.gov.br/portal/component/content/article?id=2181:parna-do-itatiaia [Accessed on September 16 2019). Bellard, C., Genovesi, P., Jeschke, J. M., 2016. Global patterns in threats to vertebrates by biological

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Analysis of conflicts with wild carnivores in the Humid Chaco, Argentina M. P. Kihn, N. Caruso, K. Iaconis, M. J. Palacios González, L. Soler

Kihn, M. P., Caruso, N., Iaconis, K., Palacios González, M. J., Soler, L., 2021. Analysis of conflicts with wild carnivores in the Humid Chaco, Argentina. Animal Biodiversity and Conservation, 44.2: 251–265, https://doi. org/10.32800/abc.2021.44.0251 Abstract Analysis of conflicts with wild carnivores in the Humid Chaco, Argentina. Interactions between humans and carnivores tend to be conflictual, especially due to predation on domestic animals. As certain landscape characteristics predispose the occurrence of carnivore attacks, spatial modelling of predation events can be particularly useful when developing management plans. In this study we determined the incidence of predation on the mortality of domestic animals by interviewing local inhabitants. In addition, we identified the spatial variables that explain the distribution of the conflicts and we created a two–scale model based on the Maxent algorithm. The results showed that Puma concolor (41.2 %) and the foxes Lycalopex gymnocercus and Cerdocyon thous (33.3 %) were the most conflictive species. Predation accounted for only 5.6 % of the causes of domestic animal mortality. The distribution models showed that the most probable variables for predicting conflicts were the distance from the roads, livestock density and the proportion of anthropized areas. High–risk areas represented 28 % of the study area and were distributed in broad patches around the protected areas and in the eastern sector of the area. Keys words: Predation, Livestock, Maxent, Spatial variables Resumen Análisis de los conflictos con carnívoros silvestres en el Chaco Húmedo de Argentina. Las interacciones entre humanos y carnívoros suelen tornarse conflictivas, en especial debido a la depredación de animales domésticos. Ciertas características del territorio favorecen que se produzcan ataques de carnívoros, por lo que puede ser muy útil elaborar modelos espaciales de los episodios de depredación a la hora de preparar planes de manejo. En este trabajo determinamos la incidencia de la depredación en la mortalidad de los animales domésticos a través de entrevistas a pobladores locales. Además, determinamos las variables espaciales que explican la distribución de los conflictos y construimos un modelo en dos escalas basado en el algoritmo de Maxent. Los resultados mostraron que el puma, Puma concolor (41,2 %) y los zorros Lycalopex gymnocercus y Cerdocyon thous (33,3 %) fueron las especies más conflictivas. La depredación representó solo el 5,6 % de las causas de mortalidad de los animales domésticos. Los modelos de distribución mostraron que las variables más probables para predecir los conflictos eran la distancia a carreteras, la densidad de ganado y la proporción de superficie antropizada. Las zonas de alto riesgo representaron el 28 % del área de estudio y se distribuyeron en amplios parches alrededor de las zonas protegidas y en el sector oriental del área. Palabras clave: Depredación, Ganado, Maxent, Variables espaciales Received: 1 X 20; Conditional acceptance: 7 I 21; Final acceptance: 14 VII 21

ISSN: 1578–665 X eISSN: 2014–928 X

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M. P. Kihn, N. Caruso, L. Soler, Grupo de Ecología Comportamental de Mamíferos (GECM), Depto. de Biología, Bioquímica y Farmacia, Universidad Nacional del Sur (UNS), San Juan 670, Bahía Blanca 8000, Buenos Aires, Argentina; and Asociación Huellas, Asociación para el Estudio y la Conservación de la Biodiversidad, Bahía Blanca 8000, Buenos Aires, Argentina.– N. Caruso, L. Soler, Instituto de Ciencias Biológicas y Biomédicas del Sur, Universidad Nacional del Sur (UNS) – CONICET, San Juan 671, Bahía Blanca 8000, Buenos Aires, Argentina.– K. Iaconis, Asociación Huellas, Asociación para el Estudio y la Conservación de la Biodiversidad, Bahía Blanca 8000, Buenos Aires, Argentina.– M. J. Palacios González, Dirección General de Medio Ambiente, Extremadura, Spain. Corresponding author: M. P. Kihn. E–mail: melisakihn@gmail.com ORCID ID: 0000-0002-4529-3084

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Introduction The expansion of human activities in many ecosystems has decreased the geographic range and populations of numerous carnivores and has also led to the fragmentation of their habitats (Morrison et al., 2007; Inskip and Zimmermann, 2009). As a consequence, predators are forced to live in anthropized environments where interactions with man may become conflictive (Rippel et al., 2014). Although there are other causes, such as the transmission of diseases, competition for game and direct attacks on humans, the predation of domestic animals is the greatest source of conflict between humans and carnivores, with lethal control being the most common method used to reduce the impact (Inskip and Zimmermann, 2009). Such control can have a devastating effect on the size and distribution of carnivore populations (Treves et al., 2011) and modify ecosystems, since they play an important role in their regulation (Prugh et al., 2009). Numerous studies on conflicts between humans and carnivores due to livestock predation have identified characteristics of the landscape that favour such attacks (Zarco González et al., 2012; Miller, 2015; Sarmiento Giraldo et al., 2016). As these characteristics are distributed in non–random patterns, their study can be used to create predictive models and diagrams to develop conflict–mitigation strategies (Treves et al., 2011). Many researchers have used interviews or surveys with experts to determine the location of livestock attack events (Van Bommel et al., 2007; Zarco González, et al. 2012; Broekhuis et al., 2017). Surveys can provide valuable information that is often impossible to obtain from other sources at a relatively low cost (Masenga et al., 2017). In the Humid Chaco of Argentina, much of the economy is based on agricultural production, generally carried out in natural environments where conflicts between humans and carnivores are often part of daily life. In many cases the subsistence of rural people is linked to raising livestock and poultry (Morello et al., 2012). The presence of these carnivores is therefore potentially conflictive, especially that of larger species. However, information on the relationship of the inhabitants with the carnivores in this region is scarcely documented. Soler et al. (2004) carried out the first diagnosis of the conservation situation of wild carnivores in the provinces of Chaco and Corrientes, showing that foxes Cerdocyon thous and Lycalopex gymnocercus were the species mentioned by the inhabitants as being the most conflictive, followed by small felines, Herpailurus yagouaroundi and Leopardus geoffroyi. In several regions of Argentina, Puma concolor is the carnivore that attacks domestic livestock most frequently. Farmers consider it is highly harmful and admit to hunting it regularly (e.g., Luengos Vidal et al., 2016). In this study we explored the scope of conflicts between humans and carnivores and their spatial distribution in the Humid Chaco ecoregion in northern Argentina. Our main objectives were to identify the most conflictive species of wild carnivore and analyze their incidence on domestic animal mortality, and to develop spatial distribution models at different scales

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to determine which environmental variables are most associated with these conflicts. Material and methods Study area The research was carried out in the northeast of Chaco province, an area belonging to the Humid Chaco ecoregion (fig. 1). In Argentina, this ecoregion encompasses the eastern half of the provinces of Chaco and Formosa, the north of Santa Fe and northwest of Corrientes (Morello et al., 2012). It is a plain with a slope slightly inclined towards the east in which depressed environments predominate, so it is prone to significant flooding. The predominant landscape is a mosaic of strips of well–drained high land with forests, accompanying the course of the rivers and alternating with low interfluves, estuaries and ravines, with features of grassland, savanna and scrubland (Naumann, 2006). The climate is temperate to humid and rainfall follows a longitudinal gradient, with maximum records in the east of more than 1,300 mm that decline in the west to 750 mm on the border with the Dry Chaco (Ginzburg and Adámoli, 2006). The ecoregion presents a remarkable diversity of wild fauna due to the heterogeneity of habitats, among which the community of carnivores includes fourteen species (Ginzburg and Adámoli, 2006). The main livestock activity in Chaco province is the extensive breeding of cattle, which mostly graze on natural grassland. The other livestock species are mostly reared to complement other activities such as cotton cultivation, hunting, forestry and fishing. Livestock production in this area is characterized by a lack of planning, a lack of facilities, and and sanitary deficiencies (Subsecretaria de Planificación Económica, 2016). Surveys of local inhabitants Data were obtained through surveys carried out in two stages. Surveys in 2016 carried out in rural areas were aimed at local inhabitants and were oral and semi–structured (see annex). The survey sites were chosen opportunistically based on the possibility of access with the vehicle used, and in each of them the location was recorded with a GPS device. In 2019, surveys were sent by the Google forms application to producers, extension workers of the National Institute of Agricultural Technology (INTA) and park rangers. The questions were structured or closed. The survey included a gridded and numbered image of the study area so that the respondents could identify cells with conflicts with carnivores (see annex). Each grid covered an area of 400 km2, with a total of 44 grids covering an area of 17,600 km2. The exact conflict site within the cell was located based on the information provided by the respondents about the particular environments where conflicts occurred (e.g. areas with dense vegetation, close to protected areas, far from towns, etc.) and with the help of a satellite image extracted from Google Earth.

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60º 0' O

59º 0' O

26º 0' S

27º 0' S

Formosa

Paraguay

100 km

Conflict location Protected area Road Human settlement

Corrientes

Fig. 1. Map showing the location of the study area in the province of Chaco, Argentina, and the location of the sites with conflicts, together with some spatial characteristics used in the analysis. Fig. 1. Mapa de la ubicación de la zona de estudio en la provincia del Chaco, en Argentina. Se muestran la ubicación de los lugares donde se produjeron conflictos y algunas características especiales utilizadas en el análisis.

Spatial variables Spatial data were processed using the QGIS program (version 3.10.0, QGIS Development Team, 2019) that included the geographical location of the points of conflicts indicated by the respondents, and the natural and anthropogenic variables, potentially determining factors of the spatial distribution of conflicts. To select the variables, we considered the interviewees’ responses and various related studies (Treves et al., 2011; Karanth et al., 2013; Miller, 2015; Rostro García et al., 2016; Broekhuis et al., 2017). Populated areas with more than 200 inhabitants, estuaries and permanent water bodies were excluded from the analysis, as they are environments where there are no carnivores or livestock. For each selected variable, we constructed raster maps of 100 m resolution. Eight variables were included in the analysis (table 1). To quantify the livestock density predictor variable, the total heads of cattle, sheep, goats, pigs and horses was extracted from the National Agricultural Census 2008 (INDEC, 2008). This information was used to create a vector map (heads/km2) of the study area. All the variables were reprojected to the WGS 84/ UTM zone 21S coordinate system, which corresponds to the reference system for the study area. The proportions of anthropized environment, herbaceous vegetation, arboreal vegetation, and livestock density

were calculated using the neighborhood analysis with the average function in QGIS using two sizes of radius (846 m and 2877 m). The different radius distances were selected to represent the approximate size of the home range of P. concolor females: 26 km2 (De Angelo et al., 2011), and of mesocarnivores: 2.25 km2 (Maffei and Taber, 2003; Luengos Vidal, 2009; Castillo et al., 2019), because the carnivores’ perception of the landscape is often related to the size of their home range (De Angelo et al., 2011). We evaluated the correlation between pairs of variables using a Pearson correlation analysis (Legendre and Legendre, 2012) and no pair showed a correlation greater than 60 %. It was thus decided to use the entire set of variables in the models. They were then cropped to the same geographic extension and transformed into ASCII format for manipulation in the Maxent program. Modeling and mapping of the probabilities of conflicts The modeling of the spatial distribution of conflict risk was developed in the MaxEnt 3.4.1 program (Maximum Entropy Species Distribution Modeling: Phillips et al., 2006). This software uses the algorithm of maximum entropy (the most uniform distribution possible) to model the most probable geographic distribution of a species from data of occurrence. In this study we used

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Table 1. Variables used in the spatial models of predation risk to predict the probability of the presence of conflicts between humans and carnivores. Tabla 1. Variables utilizadas en los modelos espaciales del riesgo de depredación para predecir la probabilidad de que se produzcan conflictos entre humanos y carnívoros.

Predictor variable (unit)

Prediction

Reference

Data source

Population Lower risk of There is a strong association between density conflicts (number high human density and the loss of of people/km2) carnivore populations (Woodroffe, 2000)

INDEC (2010)

Distance Greater risk at from greater distance towns (m) from towns

The greatest number of attacks on livestock occurs further away from human settlements because predators avoid contact with humans (Davie et al., 2014; Soh et al., 2014; Constant, et al., 2015; Loveridge et al., 2016)

IGN (2017)

Distance Greater risk at from greater distance roads (m) from roads

The risk of predation is positively associated with the distance from roads (Zarco González et al., 2012; Balbuena Serrano, 2017; Soh et al., 2014; Constant et al., 2015; Miller et al., 2015)

IGN (2017)

Livestock Higher risk at Livestock density is one of the strongest density higher density predictors of predation by carnivore (number of of livestock (Karanth et al. 2013; Carvalho et al., 2015) heads/km2)

INDEC (2008)

Distance from protected areas (m)

Greater risk at a shorter distance from protected areas

Human–wildlife conflicts of all kinds are concentrated on the borders of protected areas (Van Bommel et al. 2007; Karanth et al., 2013; Constant et al., 2015)

UNEP–WCMC (2019)

Proportion of anthropized environment

Lower risk to higher proportion of anthropized environment

Carnivores avoid highly modified environments preferring natural or relative conserved sites (Caruso et al., 2017)

IGN (2017)

Proportion of Lower risk Open areas, such as grassland and wetlands, herbaceous to higher do not offer cover for hunters that stalk their vegetation proportion of prey, such as pumas and other cats herbaceous (Miller et al., 2015; Zarco Gonzales et al., 2012) vegetation

IGN (2017)

Proportion of Higher risk arboreal at intermediate vegetation proportion of arboreal vegetation (land covered by approximately 50 % trees and shrubs)

IGN (2017)

Too much coverage can reduce the chance of finding prey and prevents the growth of grass consumed by livestock (Rostro García et al., 2016)

it in an alternative context to measure the spatial risk of predation. The MaxEnt model requires two types of input data: georeferenced carnivore conflict cases and raster maps with environmental and anthropogenic

data for the geographic area of interest. Based on this information, MaxEnt estimates the distribution of those areas that present the conditions for the occurrence of conflicts.

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90 % 80 % 70 % 60 %

Presence

50 %

Prey on domestic animals

40 % 30 % 20 % 10 % 0 %

Pca

Cc Ll Lg Lp Carnivore species

Pco

Fig. 2. Percentages of the responses of people interviewed concerning the presence of carnivore species and predation on domestic animals in the study area (N = 51): Pca, P. cancrivorus; Gc, G. cuja; Eb, E. barbara; Cc, C. chinga; Ll, L. longicaudis; Lg, L. geoffroyi; Lp, L. pardalis; Hy, H. yagouaroundi; Pco, P. concolor; F, foxes. Fig. 2. Porcentaje de las respuestas de las personas encuestadas sobre la presencia de especies de carnívoros y la depredación de animales domésticos en la zona de estudio (N = 51). (Para las abreviaturas de las especies de carnívoro, véase arriba.)

Performance of the model was evaluated using Area Under the Curve (AUC) of Receiver Operating Characteristic (ROC). This tool is widely used to measure the predictive capacity of a logistic regression model, with the result obtained being a direct measure of the discrimination capacity of the model. AUC takes values close to 1 when there is a good fit with the evaluation data and close to 0.5 when the fit is not better than that obtained by chance (Benito de

Pando and Peñas de Giles, 2007). The models with AUC values between 0.7–0.9 can be considered as moderate discrimination, whereas values > 0.9 indicate high discrimination (Rostro García et al., 2016). Two models were made, one using the variables calculated with the scale less than 846 m in radius and the other with those obtained for the scale greater than 2,877 m in radius. The models were run using the automatic 'features' option, which uses an algorithm to

Livestock areas Far from towns Close to protected areas Close to wooden areas Dense vegetation 0 %

20 %

40 %

60 %

Fig. 3. Environments most conducive to attacks of domestic animals by carnivores based on the perception of the respondents (N = 16). Fig. 3. Entornos más propicios para los ataques de carnívoros a animales domésticos según la percepción de los encuestados (N = 16).

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100 % 90 % 80 % 70 % 60 %

Carnivore attack Flood/drought Theft/loss Diseases

50 % 40 % 30 % 20 % 10 % 0 %

1st

2nd

3rd

4th

Fig. 4. Main causes of mortality of domestic animals ranked from 1st to 4th according to the respondents (N = 16). Fig. 4. Principales causas de mortalidad de los animales domésticos ordenadas de la primera a la cuarta por los encuestados (N = 16).

determine the most appropriate complexity based on the number of presence records (Syfert et al., 2013). Default settings were used and 10 replicates were performed. A random subset corresponding to 75 % of the presence data was used as training to create the model and the remaining 25 % was used as test data to assess the precision of the training model. The 'bootstrap' resampling technique was selected and the 'Cloglog' output was obtained, which is proportional to the probability of conflicts (Rostro García et al., 2016). In addition, the response curves of each variable were obtained with graphs to illustrate how the prediction of the model changes with each variable studied. The maps obtained for each model were analyzed in QGIS. The probability of conflicts was divided into three quantiles to obtain the categories of conflict probability: high, medium, and low. Finally, we calculated the surface area that covered the category of high probability of the presence of conflicts in the study area. Results Surveys A total of 51 inhabitants were surveyed, 35 in 2016 (62.8 % rural inhabitants; 11.4 % park rangers and security; 25.8 % farm employees), and 16 in 2019 (75 % producers and INTA extension workers and 25 % park rangers). These surveys indicated that foxes were the most frequently observed carnivores (78.4 %), but the respondents did not distinguish between the two species present in the area (C. thous and L. gymnocercus). Secondly, they mentioned P. concolor (66.7 %) and

P. cancrivorus (49 %). Regarding the predation of domestic animals, 64.7 % of those surveyed knew of attacks by carnivores, with P. concolor being the most conflictive (41.2 %), followed by foxes (33.3 %) (fig. 2). Among the total of respondents who mentioned cases of conflict (N = 43), 51 % indicated that these were due to attacks on poultry, 44.2 % on small livestock (goats, sheep and pigs) and 9.3 % on large livestock (cows and horses). The latter cases occurred occasionally and the animals attacked were calves and foals. In relation to the perception of carnivores, 41.5 % of those surveyed expressed a positive perception, 39.2 % considered them harmful, 9.8 % showed indifference and 5.9 % did not respond. In general, the negative opinions came from respondents linked to animal husbandry. In the 2019 surveys, participants were also asked if the carnivore attacks occurred in particular environments, to which 68.75 % answered yes, 12.5 % no, and 18.75% did not know of any particular associations. The environments mentioned by those surveyed as being the most conducive to carnivore attacks were areas with dense vegetation (54.5 %), areas far from towns (36.4 %), and environments close to protected areas (27.3 %) (fig. 3). However, the attack of carnivores was the least mentioned cause of loss of domestic animals; 61 % of those surveyed (N = 16) cited floods and droughts as the most frequent cause, 16.7 % named theft or loss and, in the same percentage, diseases that affect animals after shortage of resources due to floods and droughts. Only 5.6 % ranked carnivore attacks as the first cause of mortality (fig. 4). .

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Presence of conflicts Probability of conflicts High Medium Low

50 km

Fig. 5. Distribution of the probability of conflicts with carnivores for the smaller–scale model (846 m radius) in the northeast sector of Chaco province, Argentina. Fig. 5. Distribución de la probabilidad de que se produzcan conflictos con carnívoros según el modelo de menor escala (radio de 846 m) en el sector nororiental de la provincia del Chaco, en Argentina.

Conflict probability distribution models A total of 57 sites of presence of conflict with carnivores were obtained from both surveys, 20 of which corresponded to the surveys of 2016 and 37 to those of 2019. In the latter, the majority of respondents indicated more than one site of conflict. For the smaller scale model, the AUC value of the receiver operated characteristic curve (ROC) was of moderate discrimination (0.865) and the standard deviation was low (0.022). The area corresponding to the 'high probability' category covered approximately 5,005 km2, which represents 28.4 % of the total study area (fig. 5). The variables that most contributed to the distribution model of the probability of conflicts with carnivores were: distance from roads (20.5 %), proportion of anthropized area (18.2 %) and livestock density (17.4 %) (table 2). The response curves of each variable (fig. 6) showed that the probability of conflicts increased at a greater distance from the roads, but this happened up to 3 km, from when on the probability decreased but increased again after 8 km. On the other hand, the probability of conflicts was higher at proportions between 0.2 and 0.3 of anthropized areas, after which it decreased as the proportion increased. A decrease

in the probability of conflicts was also observed with the increase in the livestock density and with the increase in the distance from protected areas. The larger scale model (radius 2,877 m) also presented moderate discrimination in relation to the AUC of the ROC curve (0.888) and a low standard error (0.012). The area corresponding to the 'high probability' category covered approximately 4,958 km2, which represents 28.2 % of the study area (fig. 7). The variables that contributed most to this model were livestock density (20.4 %), distance from roads (20.2 %) and proportion of anthropized area (13.9 %) (table 3). The response curves of each variable indicated that the probability of conflicts decreased with the increase in both the livestock density and the proportion of anthropized area. The variable distance from roads generated the same response as for the previous model (fig. 8). The predictive maps made it posible to see that the areas with a high probability of conflict were distributed in wide patches and in various sectors of the study area, at the eastern end, around the protected areas, and on the edges of main roads, showing considerable similarity for the two models.

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Discussion Our study provides novel data on the conflicts between humans and carnivores in the Argentine Humid Chaco. On the one hand, it provides complementary results to those previously obtained in the study area (Soler et al., 2004) on the perceptions and attitudes of rural inhabitants about carnivores and the identification of the most problematic species. On the other hand, it provides the first data on the environmental variables associated with conflicts and their spatial distribution, which arose mainly as a result of the predation of domestic animals whose distribution was associated with anthropogenic variables, such as distance from roads, livestock density and the proportion of anthropized environment. Characterization of the conflicts Through the surveys we found that the mountain lion P. concolor and foxes C. thous and L. gymnocercus were considered the most conflictive species due to attacks on small livestock and poultry. Similar results were found in research carried out in other regions of the country, such as the central mountainous area (Pia, 2013), the central east (Caruso et al., 2017), the high Andean area of the northwest (Lucherini et al., 2016), and Patagonia Argentina (Novaro et al., 2017), and also in other areas of South America (Uruguay: Cravino et al., 1999; Bolivia: Pacheco et al., 2004). In contrast to the results obtained by Soler et al. (2004), in the present study the cases of predation by P. concolor represented a high proportion. This might be due to a possible increase in the number of pumas in recent years. Recent studies indicate that in nearby regions the species has recolonized areas where it had previously been eliminated, for example in the provinces of Entre Ríos (Bonnot et al., 2011; Muzzachiodi, 2012; Carmarán, 2013), Corrientes (Soler and Cáceres, 2008) and Buenos Aires (Chimento and De Lucca, 2014), as well as in Uruguay (Martínez et al., 2010) and Brazil (Mazzoli, 2012), where it has been possible to detect the

Probability of conflict

Distance from road 1.0 0.5 0.0

1.0

Table 2. Percentage of contribution (P) of each variable to the smaller–scale Maxent model (846 m radius). This model had a mean ± SE area under the curve (AUC) of 0.865 ± 0.022. Tabla 2. Porcentaje de la contribución (P) de cada variable al modelo Maxent de menor escala (radio de 846 m). Este modelo tenía un área media ± EE debajo de la curva de 0,865 ± 0,022.

Variable Distance from roads Proportion of anthropized area Livestock density Distance from protected areas Distance from towns Proportion of arboreal vegetation Population density Proportion of herbaceous vegetation

species in recent years in areas where it had previously been thought to be extinct. However, Quiroga et al. (2016) found a low density of pumas in the Western Chaco, assuming that this could be mainly due to retaliation by local ranchers in response to goat predation. Therefore, the density of the species should be corroborated with specific studies in the different areas. Pumas can cause considerable economic loss when an attack involves the death of several animals (Pacheco et al., 2004) as is common behavior of the female during the breeding season, according to several respondents. Its impact on the livestock, especially when they are with young, can be very harmful, since a single individual can kill several sheep and goats (Ruth and Murphy, 2009).

Proportion of anthropized area

0.5

0.0 39320.447 m

P 20.5 18.2 17.4 13.2 12.0 7.5 7.1 4.1

1.0

Livestock density

0.5

0.0

34.106

101.54

Fig. 6. Response curves of the Maxent model on the smaller scale, showing the relationship between the variables that contribute most to the model and the probability of the presence of conflicts. Fig. 6. Curvas de respuesta del modelo de Maxent en menor escala que muestran la relación entre las variables que más contribuyen al modelo y la probabilidad de que se produzcan conflictos.

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Presence of conflicts Probability of conflicts High Medium Low

50 km

Fig. 7. Distribution of the probability of conflicts with carnivores for the larger–scale model (2,877 m radius) in the northeast sector of Chaco province, Argentina. Fig. 7. Distribución de la probabilidad de que se produzcan conflictos con carnívoros según el modelo de mayor escala (radio de 2.877 m) en el sector nororiental de la provincia del Chaco, en Argentina.

In the case of the other species mentioned as conflictive in this study, and as reported by Soler et al. (2004), foxes are the most problematical predators, followed by H. yagouaroundi, L. geoffroyi, C. brachyurus and C. chinga. Both species of foxes, L. gymnocercus in particular, would be the most common carnivores in the study area, in the Dry Chaco (Paulucci, 2018) and in other ecoregions of Argentina such as the Monte and the Pampa (Luengos Vidal et al., 2019). These species are generally considered predators of lambs and poultry, despite the fact that previous studies on their trophic niche, in the same study area, did not report any signs of domestic livestock in their diet (Iaconis, 2015). On the contrary, the most commonly consumed items mentioned were insects, small mammals, and fruit (Iaconis, 2015). Likewise, studies carried out in Brazil (Pradella Dotto, 1997) and in Uruguay (Cravino et al., 1999) did not present sufficient evidence to consider L. gymnocercus as an important predator of livestock. Other studies carried out in the Dry or Western Chaco of Argentina showed that fruit predominated in the diet of L. gymnocercus, and the predation on domestic cattle was insignificant (Varela et al., 2008). The predation mentioned by respondents would thus appear to be due to a perceived threat rather than

to an actual threat. People's perceptions do not always match the real behavior of carnivores as they can be shaped by social and cultural influences, economic pressure, personal values, and historical events (Bruskotter and Wilson, 2014; Suryawanshi et al., 2013). Moreover, farmers may overestimate the presence of conflictive species (Caruso et al., 2017) and the levels of mortality caused by predation due to confusion with post–mortem mutilation (Cravino et al., 1999). In general, the respondents had not kept track of the number of lost animals, and thiscould also have led them to overestimate losses caused by predators. Besides, predation of poultry by domestic dogs might account for the harm caused by foxes being overestimated. Our sampling to measure perceptions included diverse perspectives from the people living in the study area. The results showed that the perception of carnivores varied according to the occupation of the respondent, with positive opinions coming from conservation agents and agricultural advisors, probably due to their better understanding of the ecological role of carnivore species and appreciation of nature. Several studies have shown that human perceptions of wildlife are affected not only by educational level (Conforti and de Azevedo, 2003; Røskaft et al.,

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2007), but also by economic interests, so it can be expected that people who rear animals may have negative opinions. The prevalence of causes of mortality of domestic animals that are not related to predation indicates that conflicts with carnivores in the study area are relatively less relevant than in previously mentioned areas such as northern Patagonia (Novaro et al., 2017) and the southwest of Buenos Aires province (Guerisoli et al., 2017) where predation was considered the main cause of livestock loss. In the Humid Chaco the rearing of goats and sheep is lower than in these two regions (INDEC, 2008) and large livestock –that are predominant in our study area– are less vulnerable to attack by carnivores than small livestock, which is a feasible explanation for the low rate of predation. Diseases and cycles of floods and droughts under conditions of scarce adoption of agricultural technology created the greatest losses. Cravino et al. (1999) reported that producers in Uruguay recognized that the mortality of lambs due to climatic causes far exceeded that ascribed to predation even though hunting and the placement of poison and traps for foxes was a common practice. Spacial distribution of the conflicts The results of the spatial risk modeling of the conflicts between humans and carnivores showed that regardless of the scale of analysis, the variables that contributed most to explaining the distribution of the conflicts were the distance from roads, the proportion of anthropized environment, and livestock density. All of these variables are associated with human presence, suggesting that carnivore behavior could be strongly determined by human activities and infrastructure. The distance from the roads showed a maximum probability of conflict within 3 km. This result coincides with that documented by Miller et al. (2015) in India where the risk of attacks on livestock by tigers (Panthera tigris) reached its maximum point at 1 km from the roads, a value which could represent a threshold distance, as here livestock can access quality pastures and carnivores can access prey without the inhibition of human

Probability of conflict

Distance from road 1.0 0.5 0.0

1.0

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Table 3. Percentage of contribution (P) of each variable to the larger–scale Maxent model (2,877 m radius). This model had a mean ± SE area under the curve (AUC) of 0.888 ± 0.012. Tabla 3. Porcentaje de la contribución (P) de cada variable al modelo Maxent de mayor escala (radio de 2.877 m). Este modelo tenía un área media ± EE debajo de la curva de 0,888 ± 0,012.

Variable Livestock density Distance from roads Proportion of anthropized area Proportion of herbaceous vegetation Distance from protected areas Distance from towns Proportion of arboreal vegetation Population density

presence. On the other hand, our results might reflect the use of roads by carnivores for their dispersal within their territories, which could be particularly true in areas with dense vegetation, such as in crop fields, grasslands, and scrublands. Local roads may enable permeability through habitat structures (Červinka et al., 2013), resulting in a higher proportion of conflicts. Many large species of predators move on roads with low traffic, as reported found by Forman and Alexander (1998). Wolves, for example, may select roads of low use as travel routes (e.g., Whittington et al., 2005). Research on the use of the roads by carnivores in the area could shed light on this assumption. In our study area, the livestock density showed a negative relationship with respect to the probability of conflicts, that is, the higher the livestock density, the lower the risk of

Proportion of anthropized area

1.0

0.5

39894.359 m

0.0

P 20.4 20.2 13.9 12.8 9.0 8.7 7.7 7.3

Livestock density

0.5

0.0

34.106

68.441

Fig. 8. Response curves of the Maxent model on the larger scale, showing the relationship between the variables with the greatest contribution to the model and the probability of the presence of conflicts. Fig. 8. Curvas de respuesta del modelo de Maxent en mayor escala que muestran la relación entre las variables que más contribuyen al modelo y la probabilidad de que se produzcan conflictos.

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predation. This may seem contradictory and contrary to that recorded by Karanth et al. (2013) and Carvalho et al. (2015), who concluded that high livestock densities were related to a higher risk of predation. However Zarco González et al. (2012) showed the existence of a negative relationship for the cases of puma predation, as in our research, which in their case might be due to the fact that the livestock were kept in protected yards and far from wooded areas. On the other hand, we think that these results may also be due to the fact that a higher livestock density is associated with degraded environments and a higher density of human settlements. However, our results could be explained on the basis of the sampling method implemented in 2016, where the surveys were concentrated in the eastern sector of the study area where the livestock density was lower than in the western sector. Uniform sampling would be necessary to define the influence of this variable on the distribution of conflict probabilities. Finally, it is necessary to bear in mind that we are studying a community of carnivores where, in general, mesocarnivores prey preferentially on poultry, whose distribution and density were not analyzed. Future research should consider a more homogeneous sampling design, considering the probabilities of attacks by pumas and mesocarnivores separately and accompanied by a distribution map of the density of poultry. The lower probability of conflict in highly altered environments could indicate that carnivores avoid degraded areas, which supports our prediction. Although human activities can affect all species of carnivores, this effect varies depending on the ecological and behavioral attributes of each species (Caruso et al., 2016). For example, the puma prefers less degraded sites and is seriously affected by habitat destruction, although it is able to tolerate some degree of fragmentation of natural environments (De Angelo et al., 2011). On the other hand, L. gymnocercus and L. geoffroyi can inhabit highly modified areas (Pereira et al., 2012; Caruso et al., 2016), demonstrating a degree of ecological plasticity that allows them to tolerate human disturbance, and they survive even in strictly agricultural areas (Pereira et al., 2012). However, it can not be overlooked that the absence, or lower density, of carnivores in anthropized environments may also be due to their elimination by the inhabitants and their dogs, or to a lower abundance of their natural prey (Pereira et al., 2012). Conclusions Our analysis of the distribution of conflicts is an estimate of the probability of their presence, and therefore, it is subject to the initial data that we decided to incorporate as predictor variables. In addition, the geographic location of presence data may exhibit spatial autocorrelation, and biases due to sampling.The selection of the options offered by the Maxent software can also affect the results. The location and intensity of data collection in wildlife studies are usually strongly influenced by accessibility to the terrain. Samples are often collected in relatively accessible locations near to roads, urban settlements, and rivers, as was the case

in our study. This potential bias can have an impact on the modeling process and give results that reflect sampling effort rather than the actual distribution of a species or process (Syfert et al., 2013). On the other hand, it is possible that other factors, such as the abundance of prey, could be important predictors of the distribution of conflicts, despite the fact that they were not taken into account in this study due to the lack of such information. Moreover, our study did not consider any possible differences, in particular, of the predation by each of the species that make up the carnivore guild in this region, so it would be important to take this into account in future studies. The nature of the surveys did not allow the interviewers to verify which species of carnivores were responsible for the attacks and, therefore, we grouped all the data to generate a map of the general risk of conflicts. The environmental variables had almost the same influence in both scales of analysis, so we can assume that the presence of conflicts in our study area does not depend on the scales we used, unlike those found by several studies that analyzed the dependency of the spatial scale on the predation events (Miller et al., 2015; Rostro García et al., 2016; Broekhuis et al., 2017). On the other hand, no significant differences were observed in the distribution of conflicts in the study area, corroborating the importance of the three variables that contributed most to the models. Although the conflict probability maps showed a wide distribution of the areas with the highest probability, the respondents identified other causes of mortality of domestic animals that produce more losses than predation. Dissemination campaigns that provide information related to improving livestock management practices are thus recommended. Monitoring cases of predation and community workshops should be organized to agree on strategies aimed at preventing predation and promoting coexistence with the native fauna. Acknowledgements To the Cátedra de Fisiología Animal of the Departamento de Biología Bioquímica y Farmacia, Universidad Nacional del Sur (Argentina) for facilitating the space for the analysis of the information. We also thank Dr. Boló Bolaños and Sr. Céspedes for helping with the logistics and accommodation for the development of the project in the Eastern Chaco, and Juan de Franco and Rebecca Greenberg for their volunteering support in the field. Finally, we wish to thank the inhabitants for their time and collaboration with the information they provided. The field study had the financial support of the Zoo des Sables d’Olonne (France) and the ZACC Conference, Zoos and Aquariums: Commiting to Conservation (USA). References Balbuena Serrano, A., 2017. Modelos espaciales de riesgo de depredación de animales domésticos por grandes carnívoros en Brasil. M. Sc. Thesis,

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Daily activity pattern of pumas (Puma concolor) and their potential prey in a tropical cloud forest of Colombia J. C. Cepeda–Duque, B. Gómez–Valencia, S. Alvarez, D. R. Gutiérrez–Sanabria, D. J. Lizcano Cepeda–Duque, J. C., Gómez–Valencia, B., Alvarez, S., Gutiérrez–Sanabria, D. R., Lizcano, D. J., 2021. Daily activity pattern of pumas (Puma concolor) and their potential prey in a tropical cloud forest of Colombia. Animal Biodiversity and Conservation, 44.2: 267–278, Doi: https://doi.org/10.32800/abc.2021.44.0267 Abstract Daily activity pattern of pumas (Puma concolor) and their potential prey in a tropical cloud forest of Colombia. Ecosystems in the northern Andes face unprecedented habitat loss. Pumas are the top predators in the region and exert key ecological functions, such as population control and resource facilitation. However, little is known about the temporal niche of the species or its effects on behaviour of prey in the tropics. We hypothesized that there is a link between the activity patterns of pumas and their prey in a cloud forest of the Central Andes of Colombia. We installed 61 camera traps to estimate the degree of overlap between the daily activity curves of pumas and seven potential prey species, using conditional kernel density functions. Pumas, armadillos, mountain pacas, and white–eared opossums were mainly nocturnal, with little crepuscular activity and high temporal overlap. Central American agouti, mountain coati, little red brocket deer, and Cauca guan displayed a predominantly diurnal activity and temporal partitioning with pumas. As opportunistic predators, pumas were able to maximize foraging efficiency by preying on the crepuscular and nocturnal species. Conservation of this highland predator will largely depend on the suitable management of its native prey. Key words: Activity, Behaviour, Colombia, Conservation, Northern Andes, Top predator Resumen Patrón de actividad diaria del puma (Puma concolor) y sus posibles presas en un bosque nublado tropical de Colombia. Los ecosistemas de los Andes del Norte afrontan una pérdida de hábitat sin precedentes. Los pumas son el predador superior de la región y ejercen funciones ecológicas claves como el control poblacional y la facilitación de recursos. No obstante, se conoce poco sobre el nicho temporal de la especie y sus efectos en la conducta de sus presas. Nuestra hipótesis es que existe una relación entre los patrones de actividad del puma y de sus presas en un bosque nublado de los Andes centrales de Colombia. Instalamos 61 cámaras trampa para estimar el grado de solapamiento entre las curvas de actividad diaria de los pumas y las de siete presas potenciales utilizando funciones condicionales de densidad de kernel. El puma, el armadillo, la paca de montaña y la zarigüeya de orejas blancas fueron principalmente nocturnos, con escasa actividad crepuscular y un alto solapamiento temporal. El agutí centroamericano, el coatí de montaña, el venado soche rojo y la pava caucana mostraron una actividad predominantemente diurna y una división temporal con el puma. Como predadores oportunistas, los pumas son capaces de optimizar la eficiencia de la alimentación al cazar presas nocturnas y crepusculares. La conservación de este predador superior dependerá en gran medida del manejo sostenible de sus presas autóctonas. Palabras clave: Actividad, Conducta, Colombia, Conservación, Andes del Norte, Predador tope Received: 6 II 21; Conditional acceptance: 19 IV 21; Final acceptance: 16 VII 21

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Juan Camilo Cepeda Duque, Laboratorio de Ecología de Bosques Tropicales y Primatología, Departamiento de Ciencias Biológicas, Universidad de los Andes, cra. 1 #18a–12, Bogotá, Colombia.– Bibiana Gomez Valencia, Instituto de Investigaciones Alexander von Humboldt. Sede Venado de oro, Avenida Paseo Bolívar 16–20, Bogotá, Colombia.– Silvia Alvarez, Wildlife Conservation Society, Avenida 5 Norte # 22N–11, Cali, Valle del Cauca, Colombia.– Diego Gutiérrez Sanabria, Fundación Reserva Natural La Palmita, Centro de Investigación, Grupo de investigaciones territoriales para el uso y conservación, cra 4 # 58–59, Oficina 201, Chapinero Alto, Bogotá, Colombia.– Diego J. Lizcano, Fundación Caipora, Transversal 8 No. 9–55.T6, Cajica, Cundinamarca, Colombia. Corresponding author: J. C. Cepeda–Duque. E–mail: acinonyxjubatus96@gmail.com ORCID ID: Juan C. Cepeda–Duque: 0000-0003-0572-6268; Bibian Gómez–Valencia: 0000-0002-5963-0221; Silvia Álvarez: 0000-0002-7397-4151; Diego Gutiérrez–Sanabria: 0000-0003-3642-0499; Diego J. Lizcano: 0000-0002-9648-0057

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Introduction Large felids are considered key drivers of community structure as they have the potential to suppress prey populations and release plants from herbivory pressure (Sergio et al., 2008). To meet their energy requirement, predators have evolved specialized traits to feed on either a diverse guild or a specific type of prey (MacDonald and Loveridge, 2010). Searching for prey can be energetically expensive, especially when they are distributed heterogeneously in the habitat across space and time (MacArthur and Pianka, 1966). Light changes throughout the 24h cycle drive the activity of a predator to be synchronized with that of its prey so as to increase the probability of encounter (Kronfeld–Schor and Dayan, 2003; Foster et al., 2013; Hernández–Sánchez and Santos–Moreno, 2020). To avoid potential injury by defensive behaviours or morphological armory of the prey, the predator needs to forage while the prey is inactive (Harmsen et al., 2011). Predator activity, however, may decrease due to greater human presence in a given area so as to minimize the risk imposed by encounters with dogs or poachers (Guerisoli et al., 2019). From the prey's perspective, increasing activity during periods when the predator is inactive may decrease the risk of mortality and maximize resource intake (Brown et al., 1999; Kronfeld–Schor and Dayan, 2003). To assess predator–prey interactions from a temporal niche perspective camera traps have proved useful (Harmsen et al., 2011; Foster et al., 2013). Evidence for temporal interactions is expected when camera traps detect convergence–divergence in the distribution of daily activity curves between two or more species (Oliveira–Santos et al., 2013; Frey et al., 2017). The puma (Puma concolor, Linnaeus, 1771) is the most widely distributed felid in the Americas. It can be found from the temperate forests of Canada to the dry Chaco in Argentina (Nielsen et al., 2015). It inhabits a large variety of ecosystems given its wide patterns of movement (mean dispersal distances of 2.2–76.6 km for females and 19.0–139.8 km for males). Its behaviour is elusive (Nielsen et al., 2015) and it is opportunistic in its prey selection (Moss et al., 2016). Nevertheless, habitat loss, retaliatory killing, and illegal hunting have led to continual declines in its populations and local extinctions (Nielsen et al., 2015). In the Andes, pumas have been intensively studied at the southern extreme of their range (Walker and Novaro, 2010). Most studies have focused on trophic ecology (Rau and Jiménez, 2002; Osorio et al., 2020), resource selection (Elbroch and Wittmer, 2013; Gelin et al., 2017), habitat use (Quiroga et al., 2016), human–wildlife conflicts (Kissling et al., 2009; Ohrens et al., 2016), population density (Guarda et al., 2017) and temporal activity pattern (Lucherini et al., 2009; Osorio et al., 2020). In the Northern Andes, the natural history, ecology and behaviour of pumas remain largely unknown. In the Central Andes of Colombia, pumas are the dominant predator in the cloud forests and paramos (Castillo et al., 2020) and the few available studies describe puma feeding habits (Hernández–Guzmán et al., 2011; Jaimes et al.,

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2018; Castillo et al., 2020), habitat use (Boron et al., 2019), human wildlife conflicts (Valderrama–Vásquez et al., 2016), spatiotemporal coexistence with jaguars (Figel et al., 2021), and activity patterns (Zapata–Ríos and Branch, 2016; Ramírez–Mejía and Sánchez, 2016; Cáceres–Martínez et al., 2016). Currently, studies about pumas and prey activity patterns in this ecoregion are lacking, preventing elucidation of the temporal strategies underpinning the survival of this predator (Soria–Díaz et al., 2016). Our goal was to describe the activity patterns of pumas in a cloud forest of the Central Andes of Colombia and assess their relationship with the activity patterns of seven prey species: Central American agoutis, white–eared opossums, mountain coatis, mountain pacas, armadillos, little red brocket deer, and Cauca guans. We hypothesized that to maximize encounters, pumas will present a greater temporal overlap of activity with their presumed favoured prey in the region. Material and methods The study was conducted on the western slope of the Central Andes of Colombia, within a forest remnant in the 3,986 ha Ucumari Natural Regional Park and the southern portion of the 21,131 ha Campoalegre Soil Conservation District (fig. 1). Forests within these protected areas were strongly degraded during the early twentieth century. Before the protected area was established, cattle ranching was widespread, even on steep slopes (up to 45º), relegating the forest to yet steeper terrain (Murcia, 1997). In the 1960s, local farms were acquired by the regional public authorities and reforestation efforts were conducted at various sites for soil and watershed protection (Kattan et al., 2006). As part of this program, plantations of Chinese ash (Fraxinus chinensis) were established in degraded lands along the upper portion of the middle Otún basin (Kattan et al., 2006). A large strip was planted in the valleys and later abandoned for secondary recovery (Rangel, 1994). Currently, a mixture of native secondary forest patches remains, forming a complex habitat mosaic of native cloud forest trees, Chinese ash, Colombian pine (Podocarpus oleifolius), Andean alder (Alnus acuminata), and exotic pine (Pinus patula) plantations (Lentijo and Kattan, 2005; Murcia, 1997). The region has a bimodal precipitation pattern with the rainiest seasons occurring between March–May and October–November, respectively. Elevation in the zone ranges from 1,750 to 2,600 m (Kattan et al., 2006). The average annual temperature is 14 ºC (range 12–18 ºC) and the average annual relative humidity is 87 % (Corporación Autónoma Regional de Risaralda, 2000). We placed 61 camera traps (Bushnell Trophy Cam HD), originally to monitor mountain tapir (Tapirus pinchaque, Roulin, 1829) populations at a maximum distance between stations of 500 m following the TEAM protocol (TEAM Network, 2011). Cameras were placed at a height of 40 cm above ground level along natural trails and distanced at least 50 m from the main trails

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to minimize theft (Kelly et al., 2012). The cameras were set to record a sequence of three pictures with a trigger interval of 1 second; no lures were used. The sampling period encompassed four months during the dry season, from December 2016 to March 2017, with a total sampling effort of 3,070 traps/nights. In the Central Andes, the northern pudu (Pudu mephistophiles, Winton, 1986) is the most consumed prey by pumas, accounting for 57 % of its diet. This is followed by the forest rabbit (Silvilagus brasiliensis, Linnaeus 1758), 54 %; mountain coati (Nasuella olivacea, Gray, 1865), 27 %; white–eared opossum (Didelphis pernigra, Allen, 1900), 10 %; little red brocket deer (Mazama rufina, Pucheran, 1851), 9 %; and mountain paca (Cuniculus taczanowskii, Stolzmann, 1865), 8 % (Hernández–Guzmán et al., 2011; Castillo et al., 2020). In the Eastern Andes, the most frequently consumed prey within the puma's diet are the brown–nosed coati (Nasua nasua, Linnaeus 1776), representing 20 % of the diet, followed by the little red brocket deer, 13 %; mountain paca, 11 %; armadillo (Dasypus novemcinctus, Linnaeus, 1758), 5 %; common opossum (Didelphis marsupalis, Linnaeus, 1758), 5 %; and Central American agouti (Dasyprocta punctata, Gray, 1842), 3 % (Jaimes et al., 2018). In the current analysis, we did not include data of northern pudu, brown–nosed coati, or forest rabbits due to the small sample size (n = 0–10 captures). We included the Cauca guan (Penelope perspeicax, Bangs, 1911) as potential prey for pumas based on a previous anecdotal report on the likely consumption of this endangered and endemic cracid near the protected area (Ríos et al., 2006). Cauca guans could be a potential prey item for pumas as these birds have ground–dwelling habits, and conspicuous behaviors. Furthermore, populations in the protected area reach the highest densities within its known geographical range (10–40 ind/km2), especially during the dry season (Kattan et al., 2014). To describe the activity patterns of pumas and their potential prey, we considered all consecutive records of a species obtained in 60 minutes as a single, independent event (Di Bitetti et al., 2006; Ridout and Linkie, 2009; Oliveira–Santos et al., 2013; Zanón Martínez et al., 2016). We assigned each independent event to one of the following periods: day (from 1 h after sunrise to 1 h before sunset), night (from 1 h after sunset to 1h before sunrise), dusk (from 1 h before sunset to 1 h after sunset), and dawn (from 1 h before sunrise to 1 h after sunrise) (Monterroso et al., 2014). We classified species as diurnal or nocturnal depending on whether their activity was concentrated within daylight or night–time hours, respectively. Habits were considered crepuscular if the species were mostly detected during the dawn and dusk hours. If a species showed no tendencies of activity for a given period, it was classified as cathemeral (van Schaik and Griffits, 1996). For each species, we transformed hours and minutes of single detection events into circular data and obtained the mean, standard deviation, and the 95 % confidence intervals of the activity patterns using a von Mises distribution (Avendaño, 2019). To assess whether the activity patterns were randomly

dispersed throughout the 24 cycles, we used the Rao Spacing test, which is more sensitive to non–unimodal distributions (Avendaño, 2019). We used a non–parametric model of kernel density function to estimate activity curves and the degree of overlap between pumas and their prey (Oliveira–Santos et al., 2013; Monterroso et al., 2014). The density of activity records for a species is measured as a continuous distribution of probabilities throughout the 24h cycle (Frey et al., 2017). These models assume that while a species is active, it has the same probability of being detected in the camera traps at any time in the day (Linkie and Ridout, 2011). We conducted pairwise comparisons of puma activity patterns and those of their prey by using the conditional overlap coefficient Δ1 for small samples (Ridout and Linkie, 2009). This coefficient has a range of 0 (no overlap) to 1 (complete overlap) and is obtained by taking the minimum value of the density functions of two daily cycles that are compared at each time point and represented as the overlapped area occupied by two specific density curves (Ridout and Linkie, 2009). To calculate the accuracy of the Δ1 estimator, we used the confidence intervals of the 1,000th bootstrap sample (Ridout and Linkie, 2009). The choice of a given conditional density isopleth can affect the interpretability of the overlap measures derived from two kernel estimators (Frey et al., 2017). For instance, the 95% density isopleth accounts for the whole temporal range where animal activity takes place, whereas the 50% isopleths account for its peak activity (Oliveira–Santos et al., 2013). Under a conditional density function, we estimated both the 95 % and 50 % isopleths to determine the 'activity range' and 'activity core', respectively, from the time records (Oliveira–Santos et al., 2013). Finally, we searched for dispersion of activity throughout the daily cycle between pumas and prey using the Watson Two–test, with a significance level of 5 %. The analysis was carried out using the circular and overlap packages (Meredith and Ridout, 2014) in the R software (R Core Team, 2017). Results Our sampling effort in the Ucumari Regional Natural Park and the southern portion of the Campoalegre Soil Conservation District yielded a total of 1,445 independent records corresponding to 66 vertebrate species, 33 mammals and 33 birds. Overall, pumas accounted for only 9% of the detections during our four survey months. Some prey, such as the Cauca guan (21 %), the mountain coati (18 %), little red brocket deer (18 %), white–eared opossum (12 %) and armadillo (12 %) were better represented, whereas other species, such as the Central American agouti (5 %) and the mountain paca (3 %) accounted for a lower number of detections (table 1). We obtained 39 independent puma records and the most frequently detected potential prey were the Cauca guan, mountain coati, and little red brocket deer with 84, 71, and 64 records respectively, followed by armadillos with 51 records,

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75.5

75.4

A La Marcada Campoalegre

4.8

Los Nevados

Otun-Quimbava

Ucumarí

4.7

Barbas Bremen

0 75.6

2.5

5 km 75.5

75.4

Fig. 1. Map of the camera trap survey conducted from 16 December 2016 to 20 March 2017 located at Ucumari Regional Natural Park and Campoalegre Soil Conservation District (A), Risaralda department (B), Colombia (C): white dots represent the camera trap stations. Fig. 1. Mapa del estudio realizado con cámaras de trampeo entre el 16 de diciembre de 2016 y el 20 de marzo de 2017 en el Parque Natural Regional Ucumari y el Distrito de Conservación de Suelos Campoalegre (A), en el departamento de Risaralda (B), Colombia (C): los puntos blancos representan las estaciones de fototrampeo.

white–eared opossums with 50 records, Central American agoutis with 22 records, and mountain pacas with 12 records. Puma activity was predominantly nocturnal or crepuscular. The onset of activity was late afternoon (17:00–18:00 h), peaking early in the night (20:00–22:00 h) and decreasing before sunrise (01:00–05:00 h). Subtle increases in puma activity were recorded during the day, especially at midday and early afternoon (12:00–13:00 h). Regarding the activity patterns of prey, the Cauca guan, mountain coati, and Central American agouti were primarily diurnal, and the white–eared opossum, mountain paca, and armadillo were mainly nocturnal (fig. 2). Cauca guans were the only species with activity mainly around noon (11:0–12:00 h), with a decreasing trend towards sunrise and sunset (fig. 2A). The activity of Central American agoutis started in the early morning (07:00–09:00 h), decreased around midday (11:00– 12:00 h) and peaked near dusk (16:00–18:00 h).

Little red brocket deer were mostly diurnal and crepuscular (fig. 2F), with two activity peaks, one during the morning with periodic increases in the early (07:00–09:00 h) and late morning (10:00–11:00 h). The other peak was around the afternoon and dusk (14:00–18:00 h). Mountain coatis also showed two distinct peaks, one in the late morning (10:00–11:00 h) and the other between the early afternoon and dusk (14:00–18:00 h). White–eared opossums showed a predominant crepuscular activity, with a distinctive peak at dusk (18:00–19:00 h) that dramatically decreased during early night hours (20:00–23:00 h) before increasing again at midnight (fig. 2E). Armadillos and mountain pacas showed two distinctive nocturnal peaks, one concentrated before midnight (21:00–23:00 h) and one after midnight (00:00– 02:00 h), with a gap in its activity between the two periods (fig. 2). Circular central tendency measures and directionality tests for the activity records of each species are provided in table 1.

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Table 1. Activity patterns of pumas and their prey in a cloud forest of the Central Andes, Colombia. Number of 60–minute independent events (IE), mean and circular deviation (SD) from the activity vectors of each species are shown. Confidence intervals (CI) and the Rao (U) test were set at an alpha = 0.05. Tabla 1. Patrones de actividad del puma y de sus presas en un bosque nuboso de los Andes centrales, Colombia. Se indican el número de episodios independientes de 60 minutos (IE), la media y la desviación estándar circular (SD) de los vectores de actividad de cada especie. Los intervalos de confianza (CI) y la prueba de Rao (U) se calibraron a un alfa = 0,05. Species

Mean (SD)

CI (95%)

U–test

P–value

Puma

20:35 (01:49)

01:39–14:22

131.47

> 0.001

Cauca guan

11:30 (00:37)

10:55–12:01

222.77

< 0.001

Central American agouti

13:19 (00:57)

11:28–15:03

200.64

< 0.001

Mountain coati

12:54 (00:59)

11:48–13:54

187.28

< 0.001

Armadillo

23:31 (00:56)

22:20–00:33

202.95

< 0.001

White–eared oppossum

22:32(00:56)

21:25–23:43

221.25

< 0.001

Mountain paca

23:09 (00:42)

21:19–00:40

216.15

< 0.001

Little red brocket deer

12:54 (00:52)

11:59–13:47

180.82

< 0.001

Coefficients of overlap in the temporal activity patterns varied according to the range and core activity (table 2). We noted a higher overlap of activity between pumas and nocturnal prey species than between pumas and diurnal species (fig. 2). Cauca guans, Central American agoutis, and mountain coatis with a mostly diurnal temporal activity, exhibited low overlap with puma activity patterns (table 2). Regarding to temporal partitioning, we found statistical differences in all the paired associations between the activity patterns of pumas vs. Cauca guans (U2 = 1.2125, P < 0.001), vs. Central American agoutis (U2 = 0.4061, P < 0.001), vs. mountain coatis (U2 = 0.6772, P < 0.001), vs. little red brocket deers (U2 = 0.7031, P < 0.001), vs. mountain pacas (U2 = 0.2207, P < 0.001), vs. armadillos (U2 = 0.3415, P < 0.001), and vs. white–eared opossums (U2 = 0.3125, P < 0.001). Discussion We assessed the daily activity of pumas and their prey within a well–preserved Andean Forest in which other large predators such as jaguars (Panthera onca, Linnaeus 1758) are absent. Pumas showed a nocturnal and crepuscular activity that is geographically consistent with studies made in other Neotropical areas (Scognamillo et al., 2003; Monroy–Vilchis et al., 2009; Harmsen et al., 2009; Blake et al., 2012; Foster et al., 2013; Zanón Martínez et al., 2016; Cáceres–Martínez et al., 2016; Porfirio et al., 2017; Azevedo et al., 2018; Guerisoli et al., 2019; Osorio et al., 2020). Therefore, factors other than habitat features (such as human disturbance and prey availability) may more likely drivers of the puma’s activity patterns (Harmsen et al., 2011; Suraci et al., 2019). In view of the small sample size, the puma activity

patterns we observed should be interpreted with caution. Furthermore, we were unable to discriminate between sexes in our records. Recent research emphasized it is necessary to have more than 100 records to reduce bias in Δ1 estimates (Lashley et al., 2018) and that there are intersexual differences in puma activity, with males being more nocturnal and crepuscular and females more cathemeral (Azevedo et al., 2018). Regarding diurnal prey, our results for mountain coatis differed from reported populations in the Central and Eastern Andes (Ramírez–Mejia and Sánchez, 2016, Cáceres–Martínez et al., 2016) of Colombia, and the Tabaconas Namballe reserve in Peru (Mena and Yagui, 2019) where they are primarily nocturnal. This could be attributed to a behavioral strategy to avoid potential intraguild predation by pumas (Castillo et al., 2020), and feral dogs (Mena and Yagui, 2019). Evidence from the Central Andes of Ecuador (Zapata–Ríos and Branch, 2016) suggests that the presence of feral dogs around the protected areas can deplete mountain coati abundance. Yet given the low number of dogs (n < 5 observed during our survey, we opted to exclude them from our inferences to avoid skewed comparisons (Mena and Yagui, 2019). When compared with other procyonids, coatis have an extra cone–class on their retinas, which confers them dichromatic colour vision (Jacobs and Deegan, 1992). Greater activity during diurnal hours may thus increase their visual perception and feeding intake while foraging on the forest floor (Whiteside, 2009). Moreover, diurnal activity in coatis prevents heat loss while foraging, and foster social interactions by increasing intraspecific recognition (Costa et al., 2009; Mena and Yagui, 2019). Central American agoutis showed a diurnal behaviour, concentrated in the early morning and the late afternoon. This observation matches findings from

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Density

0.15

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0.10 0.05 0.00

Density

0.15 0.10 0.05

0.00 0:00 6:00 12:00 18:00 24:00 Time G

0:00 6:00 12:00 18:00 24:00 Tme

Density

0.20 0.10 0.00

0:00 6:00 12:00 18:00 24:00 Time

Fig. 2. Activity patterns and degree of overlap between pumas (red line) and prey (blue line) at the Ucumari Regional Natural Park and the Campoalegre Soil Conservation District: A, Cauca guans; B, Central American agoutis; C, mountain coatis; D, mountain pacas; E, white–eared opossums; F, armadillos; G, little red brocket deer. (The shaded area represents overlap; from December 2016 to March 2017 we obtained 393 independent activity records for pumas and prey from 3,070 trap/nights.) Fig. 2. Patrones de actividad y grado de solapamiento entre el puma (línea roja) y sus presas (línea azul) en el Parque Natural Regional Ucumari y el Distrito de Conservación de Suelos Campoalegre: A, pava caucana; B, agutí centroamericano; C, coatí de montaña; D, paca de montaña; E, zarigüeya de orejas blancas; F, armadillo; G, venado soche rojo. (El área sombreada reprresenta el solapamiento; entre diciembre de 2016 y marzo de 2017 obtuvimos 393 registros independientes de actividad para pumas y sus presas con un esfuerzo de 3.070 trampas noche.)

previous studies in the tropical rainforests of Mexico (Mendoza et al., 2019), Panamá (Suselbeek et al., 2014), the Brazilian Amazon (Gómez et al., 2005), and the Central Andes of Colombia (García–R. et al., 2019). This diurnal behaviour in Central American agoutis could increase its success in searching for seeds while avoiding the midday heat (Suselbeek et al., 2014; Duquette et al., 2017). Cauca guans were the only species whose activity peaked at midday and decreased towards the sunset and sunrise hours. Populations of this endemic cracid migrate from adjacent forests in the region during the dry season (November–December) to forage on Chinese Ash leaves (Muñóz et al., 2007), which provide more food than native trees (Kattan et al., 2014). In these open plantations, up to 30 individuals of Cauca guans have been observed foraging at a single location and the open canopy can increase their exposure to predators

(Muñóz et al., 2007). Thus, being active at the hottest hours of the day when pumas are less active might reduce predation risk. Likewise, terrestrial habits in this guan are linked to opportunistic foraging on army ants (Labidus praedator), and increased activity during midday hours could increase resource intake when other birds are less active (Rios et al., 2008). The diel activity of little red brocket deer that we observed to peak in the morning is consistent with the activity patterns found for this species in cloud forests and paramos of Ecuador (Zapata–Ríos and Branch, 2016). However, it differs from the findings for a neighbouring region, which showed a greater increase in activity during crepuscular hours (Ramírez–Mejía and Sánchez, 2016). Our results depart from the phylogenetic signal in activity observed towards nocturnality for other red Mazama species (Oliveira et al., 2016) such as the South American red brocket deer (M. americana,

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Table 2. Overlap estimates at the range (0.95) and core (0.50) activity for the paired associations of pumas and prey in the Ucumari Regional Natural Park and the Campoalegre Soil Conservation District, Central Andes, Colombia. We obtained an additional overlap estimate by resampling the transformed data at 1,000 iterations. Confidence intervals of the overlap coefficients were set at an alpha = 0.05. Tabla 2. Estimación del solapamiento en el período total de actividad (0,95) y el período de actividad máxima (0,50) para las asociaciones pareadas de pumas y sus presas en el Parque Natural Regional de Ucumari y el Distrito de Conservación de Suelos Campoalegre, en los Andes centrales, Colombia. Obtuvimos una estimación adicional de solapamiento al volver a muestrear los datos transformados con 1.000 iteraciones. Los intervalos de confianza de los coeficientes de solapamiento se calibraron a un alfa = 0,05. Species pair

Δ1 (0.95)

Δ1 (0.50)

Δ1 resampled

CI (95%)

Puma vs. Cauca guan

0.35

0.01

0.42

0.23–0.51

Puma vs. Central American agouti

0.51

0.15

0.54

0.34–0.66

Puma vs. mountain coati

0.51

0.07

0.56

0.37–0.67

Puma vs. armadillo

0.68

0.50

0.63

0.49–0.78

Puma vs. white–eared opossum

0.69

0.63

0.60

0.46–0.73

Puma vs. mountain paca

0.58

0.44

0.54

0.35–0.72

Puma vs. little red brocket deer

0.47

0.01

0.47

0.25–0.68

Erxleben, 1777), the Brazilian dwarf brocket deer (M. nana, Hensel, 1872), and the small red brocket deer (M. bororo, Duarte, 1996). Deer are central taxa within the puma’s diet (Ackerman, 1982), and the diurnal behaviour observed for little red brocket deer may reflect behavioural avoidance of encounters with this and other predators (Zapata–Ríos and Branch, 2016). Additional constraints in the activity of little red brocket deer may be attributed to resource partitioning with other sympatric ungulates (Blake et al., 2012) and thermal constraints related to closed habitats (Oliveira et al., 2016). Nocturnal prey species like white–eared opossums showed an activity pattern that is congruent with other studied populations of the Central (Ramírez–Mejía and Sánchez, 2016) and Eastern Andes (Cáceres–Martínez et al., 2016) of Colombia, and the Central Andes of Ecuador (Zapata–Ríos and Branch, 2016). As white– eared opossums can exploit various human–related resources, its nocturnal and arboreal activity allows it to minimize encounters with humans and dogs (Zapata–Ríos and Branch, 2016). Mountain pacas showed a nocturnal activity consistent with previous studies in Perú (Jiménez et al., 2010) and the Central Andes of Colombia (Ramírez–Mejia and Sánchez, 2016). This large rodent has historically been hunted by humans for its meat, even within protected areas where poaching is practically absent; its nocturnal behaviour could be a conditioned response to the presence of humans and dogs (Zapata–Ríos and Branch, 2016). Our initial hypothesis of temporal overlap between puma and prey was supported only by armadillos and white–eared opossums, two nocturnal mammals frequently targeted by this predator in the region (Castillo et al., 2020). This pattern is consistent with other

studies where pumas adjust their activity to decrease the energy costs of searching for food (Scognamillo et al., 2003; Harmsen et al., 2011; Foster et al., 2013; Zanón Martínez et al., 2016; Azevedo et al., 2018; Osorio et al., 2020). Moreover, it has been reported that higher nocturnality of predators in areas with intense human activity can create temporal human shields for locally abundant prey by reducing its perceived risk of predation (Gaynor et al., 2018; Suraci et al., 2019). In the Pampas (Zanón Martínez et al., 2016) and Triangulo Mineiro regions (Azevedo et al., 2018) pumas were found to be strictly nocturnal in areas with higher diurnal activity of humans. A trail that longitudinally crosses the Ucumarí Natural Regional Park is frequently visited by tourists and local farmers, mostly during daytime (Rangel, 1994). Within the park, mountain coatis, little red brocket deer, and Cauca guans have the potential to reach high densities (Kattan et al., 2014; unpublished data), which may increase their predation risk (Osorio et al., 2020). Thus, despite being common in our survey, the temporal partitioning of mountain coatis, little red brocket deer, and Cauca guans can decrease the predation pressure exerted by pumas. The theory of limiting similarity (Abrams, 1983) states that coexistence within assemblages or communities drives differences among one or more niche dimensions (but see Jaksić, 1982). Further understanding of how the shared use of the light hours can drive competition is needed to determine future conservation actions. Some relevant prey in the diet of puma populations from the Northern Andes, such as the northern pudu, and forest rabbit, were rare in our survey (Hernández–Guzman et al., 2011; Castillo et al., 2020). More research focused on the puma feeding habits in these

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Andean forests is needed to reliably couple temporal predator–prey relationships with trophic interactions (Azevedo et al., 2018). Historically, the Andean region has been characterized by a prevalence of human activities that have caused dramatic changes in the natural heritage of Colombia (Sánchez–Cuervo et al., 2012). As mentioned above, human–mediated fear in pumas may reduce the range of prey they can reach when compared to the entire diversity of prey taxa available (Suraci et al., 2019; Blecha et al., 2018; Wilmers et al., 2013). More rigorous assessment on the spatial and temporal niche of both the puma and its available prey assemblage, perhaps using radio telemetry, should be carried out to confirm our results. The conservation of forested habitats is critical for the puma’s survival and the maintenance of its prey base (Paviolo et al., 2018). Puma activity patterns were mainly nocturnal in our study, and given its opportunistic foraging behaviour, we suggest that the more nocturnal potential prey will be more relevant as a feeding resource. Although armadillos and white–eared opossums provide less biomass (3–7 kg and 1–6 kg, respectively) for pumas when compared with larger prey such as the little red borcked deer (11 kg) (Hernández–Guzmán et al., 2011; Foster et al., 2013) we suggest that, even within the protected area, human activities (trekking, unmanaged tourism) may create a landscape of fear that constricts the trophic niche of the puma to less nutritive prey (Suracy et al., 2019). We found that activity overlap analysis was a very useful (albeit complementary) approach to suggest likely key prey items for pumas when dietary studies are lacking (Azevedo et al., 2018). Several important questions remain unsolved. For instance, how can tourism within these protected areas influence the puma's prey selection, hunting success, and energy intake? And are there any seasonal or intersexual differences in the temporal overlap between pumas and prey in these threatened highland forests? Finding answers to these questions will be key for the development of management and conservation measures and for the protection of the remaining puma populations and wildlife that share the tropical cloud forests. Acknowledgements We thank Arnobis Garcia and Jairo Garcia for their significant contribution to the fieldwork. We are also grateful to the Fundación Caipora, Segre Foundation, and the IUCN Tapir Specialist Group for funding this research. JCC–D is especially grateful to Jim Patton, Christian Osorio, Tadeu G. de Oliveira, Eva López– Tello, and Salvador Mandujano for their valuable comments to improve the manuscript. References Abrams, P., 1983. The theory of limiting similarity. Annual review of ecology and systematics, 14(1): 359–

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Modelling European turtle dove (Streptopelia turtur L. 1758) distribution in the south eastern Iberian Peninsula A. Bermúdez–Cavero, J. A. Gil–Delgado, G. M. López–Iborra Bermúdez–Cavero, A., Gil–Delgado, J. A., López–Iborra, G. M., 2021. Modelling European turtle dove (Streptopelia turtur L. 1758) distribution in the south eastern Iberian Peninsula. Animal Biodiversity and Conservation, 44.2: 279–287, Doi: https://doi.org/10.32800/abc.2021.44.0279 Abstract Modelling European turtle dove (Streptopelia turtur L. 1758) distribution in the south eastern Iberian Peninsula. The European turtle dove population and breeding range has declined sharply in Spain. This study reanalyses data from the Atlas of Breeding Birds in Alicante (SE Spain), aiming to identify the main variables related to its occurrence and abundance. We used hierarchical partitioning analysis to identify important environmental variables associated with natural vegetation, farming, hydrological web, anthropic presence, climate, and topography. Analysis combining the most explicative variables of each group identified the mixture of pines and scrubland in the semiarid areas and the length of unpaved roads as the most important variables with a positive effect on occurrence, while herbaceous crops and scrublands in dry ombrotype climate areas had the most important negative effect. Abundance was related only to the availability of water points. We discuss the implications of these findings for habitat management in conservation of this species. Key words: Agriculture intensification, Habitat change, Hierarchical partitioning analysis, Pinewood, Population decline, Mediterranean Resumen Elaboración de modelos de la distribución de la tórtola europea (Streptopelia turtur L. 1758) en el sureste de la península ibérica. La población y el área de reproducción de la tórtola europea han disminuido considerablemente en España. En el presente estudio realizamos un nuevo análisis de los datos obtenidos para la elaboración del Atlas de Aves Reproductoras de Alicante (SE de España) con el objetivo de identificar las principales variables relacionadas con la presencia y la abundancia de esta especie. Utilizamos el análisis de partición jerárquica para identificar estas variables ambientales (vegetación natural, agricultura, red hidrológica, presencia antrópica, clima y topografía). El análisis que combinó las variables más explicativas de cada grupo permitió determinar que la mezcla de pinos y matorrales en zonas semiáridas y a lo largo de las carreteras sin asfaltar es la variable que tuvo el mayor efecto positivo en la presencia de la tórtola, mientras que la mezcla de cultivos herbáceos y matorrales en el ombrotipo seco es la que tuvo el mayor efecto negativo. La abundancia solo se relacionó con la disponibilidad de puntos de agua. Se discuten las implicaciones de estos resultados con respecto a la gestión del hábitat para la conservación de esta especie. Palabras clave: Intensificación de la agricultura, Cambio de hábitat, Análisis de partición jerárquica, Pinar, Disminución de la población, Mediterráneo Received: 7 I 20; Conditional acceptance: 13 V 20; Final acceptance: 23 VII 21 Alan Bermúdez–Cavero, José A. Gil–Delgado, Instituto Cavanilles de Biodiversidad y Biología Evolutiva, c/Cate– drático José Beltrán 2, 46980 Paterna, Valencia, España (Spain).– Germán M. López–Iborra, Departamento de Ecología–IMEM Ramón Margalef, Universidad de Alicante, Apdo. 99, 03080 Alicante, España (Spain).– Alan Bermúdez–Cavero, Laboratorio de Biología y Genética, Universidad Nacional de San Cristóbal de Huamanga, Av. Independencia s/n., Ayacucho, Peru. Corresponding author: A. Bermúdez–Cavero. E–mail: alan.bermudez@unsch.edu.pe ORCID ID: A. Bermúdez–Cavero: 0000-0002-9567-2701; J. A. Gil–Delgado: 0000-0002-0244-0769 G. M. López–Iborra: 0000-0003-3045-5498 ISSN: 1578–665 X eISSN: 2014–928 X

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Introduction The European turtle dove (Streptopelia turtur) is a migratory species that winters in the sub–Saharan region but breeds from North Africa to the Urals (Cramp, 1985). As its populations have declined substantially in recent years, particularly in Western Europe (Jiguet, 2016; Harris et al., 2018), it has been classified as threatened (Bird Life International, 2016). Its population trend in Spain has been monitored since 1996 when assessment of populations of common birds began (SEO/BirdLife, 2010). A declining trend has been observed (Saenz de Buruaga et al., 2012), leading the species to be listed as vulnerable in the Red Data Book of Birds of Spain (Madroño et al., 2004). The main causes of the decline in turtle dove populations are loss of nesting habitat and reduced food availability (Browne et al., 2005). Another cause is unsustainable hunting during the spring migration and the late breeding season (Boutin and Lutz, 2007). The habitat of European turtle dove is mainly associated with open forests and agricultural environments (Gibbs et al., 2010). These two habitat types have substantially changed in the last decades due to forest closures, land abandonment and intensification of agricultural land (Hanane, 2017). In the Mediterranean region, the European turtle dove prefers a mixture of wooded areas and scrubland close to grassland or farmland (Balmori, 2003). In eastern Spain, orange crops show a widely cover area and breeds in such orchards (Gil–Delgado, 1981). Furthermore, the population of this dove decreases as altitude and tree cover increases (Saenz de Buruaga et al., 2012). In Spain, higher European turtle dove densities have been found in thermo–meso–mediterranean environments, especially in pine forests and on wooded farmlands (Carrascal and Palomino, 2008). A previous study about land use of European turtle dove throughout mainland Spain found that localities dominated by complex cultivation presented the most favourable trend for this species, while the extensions of several forest types were related to more negative trends (Moreno–Zarate et al., 2020). According to Moreno–Zarate et al. (2020), most of the province of Alicante is located in the area with medium–high favourability in the transition area between meso–mediterranean and thermo–mediterranean zones, and it is currently undergoing major landscape changes driven by increased urbanisation, abandonment of traditional crops, and intensification and irrigation of new crops (Serra et al., 2008; López–Iborra et al., 2011). Therefore, knowing the relationship between occurrence of this dove and habitat characteristics may help understand the causes of the decline of this species in south–east Spain and provide insights into possible management measures that can benefit its conservation. A detailed atlas of breeding birds in Alicante, with birds and habitat data available on the 1 km2 scale (López–Iborra et al., 2015), provides the opportunity to identify the main habitat variables that affect the distribution of the European turtle dove. The aim of this paper was to evaluate the relative importance of natural vegetation (forests, shrubland),

crops of several types, and human pressure on the probability of the dove´s occurrence and abundance. These analyses also check for the potential effect of the physical environment (climate, topography, hydrological web) interacting with vegetation anthropogenic structures. However, it is more difficult to analyse the impact of modifications caused by humans. If natural or cultivated vegetation is more important than physical variables in determining the presence or abundance of the European turtle dove, there is greater potential to improve its conservation through habitat management. Material and methods Study area The study area is located in the province of Alicante (SE Spain). This province covers about 5,800 km2. It includes mountainous areas in the north and west, and wetlands and plains in the south–eastern and southern sectors. The study area has a semi–arid Mediterranean climate (Rivas–Martinez, 1987). Average annual precipitation varies from 300 to 600 mm, and rainfall is highest in autumn and winter (Rivas–Martinez, 1987). Landscapes are characterised by the dominance of scrub with various levels of development, and by mixtures of scrub and forest areas where Aleppo pines (Pinus halepensis) predominate, interspersed with agricultural fields (Rigual, 1972). Presence and abundance of European turtle dove The occurrence data came from the surveys conducted for the Atlas of Breeding Birds in the province of Alicante (López–Iborra et al., 2015). The census was conducted in a stratified random sample of 132 2 × 2 km squares defined according to UTM grid and covered approximately 10 % of this province. In each cell of 1 km2, a transect of 1 km was walked twice during the breeding season between 2001 and 2004. The species was detected in 147 cells of 1 km2 (see López–Iborra et al., 2015 for details). Predictive variables Of the available 113 variables (López–Iborra et al., 2015), we selected those of higher biological significance. The number of variables used thus decreased to 25 (table 1). These variables were classified into six groups (López–Iborra et al., 2015): i) size of the natural vegetation area (forest, scrubs, and a mixture of both habitats); ii) agricultural areas; iii) hydrological web (artificial and natural water bodies); iv) topographic variables; v) climate variables; and vi) variables related to human disturbance, such as urbanized surfaces and parks (table 1). As topography and water availability defined the development of vegetation cover, the forest variables, scrub, and mixtures of the two, were subdivided into ombrotypes (semi–arid, dry, and subhumid), bioclimatic belts (thermo–mediterranean, meso–mediterranean,

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Table 1. Environmental variables used as predictors in HP analyses. A more detailed list is found in López– Iborra et al. (2015). The total forest area was the sum of three developmental stages of pine (young, medium, mature), pines associated with other tree crops and new, repopulated pines. Three variables were obtained from the total extent of woody crops: the first was the sum of almond trees (almonds, mixture of almond trees and vineyards, mixture of almond trees and other crops); the second was the total olive area (olive trees, mixture of olive trees and other crops); and the third was the total vineyard areas (vineyard, mixture of vineyards and other crops), citrus crops area, cherry, pomegranate and fig trees. The total area of herbaceous crops consisted of intensive labour on dry land, which may also be associated with tree crops and vineyards, herbaceous crops on irrigated fields, forced crops, and other crops (López–Iborra et al., 2015). Tabla 1. Variables ambientales utilizadas como predictores en los análisis de partición jerárquica. Para obtener una lista más detallada de las variables, veáse López–Iborra et al. (2015). La superficie forestal total se obtuvo de la suma de las fases de desarrollo de los pinos (joven, intermedio y maduro), los pinos asociados a otros cultivos arbóreos y los nuevos pinos de repoblación. De la extensión total de cultivos leñosos se obtuvieron tres variables: la primera fue la suma de almendros (almendros, mezcla de almendros y viñedos, y mezcla de almendros y otros cultivos), la segunda fue la superficie total de olivos (olivos y mezcla de olivos y otros cultivos) y la tercera fue la superficie total de viñedos (viñedos y mezcla de viñedos y otros cultivos) y la superficie de cítricos, cerezos, granados e higueras. La superficie total de cultivos herbáceos comprendió las tierras de secano con labor intensiva, que también pueden estar asociadas a cultivos arbóreos y viñedos, cultivos herbáceos en campos de regadío, cultivos forzados y otros cultivos (López–Iborra et al., 2015).

Variable Description Natural vegetation (area in ha, except for diversity indices) Scrub Area covered by any kind of scrub ScrubPine Area covered by a mixture of pines and scrub Forest Area covered by forest DivForest Shannon diversity index for forest types DivScrub Shannon diversity index for scrub types DivScruPines Shannon diversity index for the scrub–pine mixture Farming (area in ha, except diversity indices) DivWoodCrop Shannon diversity index for woody crops DivHerbCrop Shannon diversity index for herbaceous crops WoodCrop Area covered by woody crops HerbCrop Area covered by herbaceous crops Hydrological web River Length (m) of rivers RavGullies Summation of length of ravines and gullies NWB Summation of water bodies number (pounds and pools) TotalChan Length (m) of channels and ditches Anthropic Unproductive Area covered by unproductive vegetation IsolHouses Area (m2) covered by isolated houses HousDevel Area (m2) covered by housing developments Urban Area (m2) occupied by cities PavRoad Length (km) of paved roads UnPavRoad Length (km) of unpaved roads Climate OmbrIndex OmbrothermicIndex ThermIndex ThermicityIndex Topograpphy DistCoast Distance to the coast (km) AltMean Average altitude (m) Slope Average slope

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and supra–mediterranean) and aspect (north, south). Thus 12 types of scrub and 10 mixtures of scrub– pine were obtained and used to calculate the Shannon–Wiener diversity index (Margalef, 1973) in each surveyed cell. The Shannon–Wiener diversity index values calculated for forests, herbaceous cultures and woody crops were obtained from the forest covers and crop types taken from digital land use maps. Collinearity between the predictive variables was assessed using Pearson’s correlation coefficient and collinearity was considered high when r > 0.7 (Mason and Perreault, 1991). For instance, pond and pool water bodies correlated highly and were summed to create a new variable: number of water bodies (NWB). The sub–humid thermo–mediterranean scrub on the southern and northern slopes showed a high correlation (r = 0.89), and scrubland were taken as a group. The pine–scrub and scrub mixture showed no collinearity. Hence, both variables were used in our analysis. Regarding anthropic variables, as the numbers of isolated houses and their surfaces also correlated (r = 0.74) we used the latter (table 1). Statistical analysis A hierarchical partitioning (HP) analysis was run to assess the variables that best explained the presence and abundance of European turtle dove. HP is used in ecological studies to identify the environmental variables that are most likely related to the occurrence or abundance of species to control for collinearity with other variables (López–Iborra et al., 2011). HP computes all the possible hierarchical models in a set of independent predictors, and its explicative capacity is divided into the individual effect I of each variable and its joint effect J through other variables (MacNally, 2002). A negative J can be possible for variables that act as suppressors of other variables (Chevan and Sutherland, 1991). HP was applied separately to dove presence and abundance data. Only the data from the second visit to transects (1 May–15 June) were used because earlier presences could correspond to migrating birds. During this period, the European turtle dove was detected in 126 squares of 1 km2. Abundance was analysed only for the presence squares and the number of doves detected in the 1 km transect. Given the limitation in the number of variables that HP can handle (Olea et al., 2010), this analysis was applied in a first step using the variables in each group shown in table 1. Some of the variables of natural vegetation cover (scrub, scrub–pine and forest) and farming (TreeCrop, HerbCrop) are available as more detailed variables according to the degree of pine development, the main crop type or as a result of crossing scrub or scrub– pine cover with the slope orientation, thermotype or ombrotypes (see López–Iborra et al., 2015 for details; table 1s in supplementary material). To test if it was possible to identify the subtypes of these habitats that are relevant for European turtle dove presence or abundance, HP analyses were also performed for each group with these more detailed variables.

Finally, an HP analysis was performed for each response variable (occurrence or abundance) by combining the variables that were significant in the analysis of each group. As the HP analyses may give rise to some errors when there are more than nine variables (Olea et al., 2010), each analysis was performed with a maximum of eight independent variables, plus a spatial term (see below). For occurrence analyses with more than eight candidate variables for the final analysis, predictors were decreasingly ranked according to their percentage of explained deviance and the eight variables that explained a higher percentage of deviance were used in the final analysis. Generalized logistic linear models were used for occurrence HP (Jongman et al., 1995). For abundance HP, Poisson regressions were used and pseudo–R2 as the goodness–of–fit measure (Jongman et al., 1995). A spatial term was included in all the analyses as the probability of occurrence or the abundance predicted by a cubic function of the geographic coordinates (X+Y+X2+Y2+XY+X3+Y3+X2Y+Y2X) to control for spatial autocorrelation (Legendre, 1993). The significance of the independent contribution of the environmental variables was obtained by a bootstrap test based on 999 randomisations (MacNally, 2002). Given that HP analysis does not provide information on the sign (positive or negative) of the effect of the independent variables, univariate regressions were carried out with each significant predictor and the spatial term to obtain the regression coefficient and its sign was used to describe the direction of the effect. The 'hier.part' package (Walsh and MacNally, 2015) in the R software (R Core Team, 2018) was used to perform the HP analyses. Results Occurrence models According to the HP analysis of the occurrence data, the group of human pressure variables explained the highest percentage of deviance, followed closely by natural vegetation and farming variables (table 2). Natural vegetation was the group with most significant variables. This group was disaggregated into seven subgroups that were analysed separately (table 1s in supplementary material). At least one significant variable was identified in all the groups. Variables with a significant positive effect were (variable group between parentheses): Pin_Mid (forest), ScPineS (mixture scrub and pine orientation), ScPineTme (mixture scrub and pine thermotype), ScPineSa (mixture scrub and pine ombrotype). Five significant variables had a negative effect, all relative to types of scrub defined by orientation, thermotype or ombrotype (table 1s in supplementary material). We also disaggregated farming variables into two groups: woody and herbaceous crops. The woody crops HP revealed a positive effect of almond, citrus and other fruiting trees and a negative effect of the cherry orchards. Herbaceous crops HP detected a negative effect of the two categories of this crop type that were not mixed with woody crops (table 1s in supplementary material).

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Table 2. Results of the HP analysis of presence and abundance of European turtle dove performed with each group of variables: I and J, independent and joint effect of each variable in the model, respectively; %I, percentage of the independent contribution of each variable in the group; S, direction of the effect; Zs, Zscore, randomisation test of independent contribution of each variable; %DV, percentage of deviance accounted for by each variable in the group; %Dev, percentage of deviance accounted for a model including all the variables in each group. (* p < 0.05, ** p < 0.01, *** p < 0.001). Tabla 2. Resultado de los análisis de partición jerárquica de la presencia y la abundancia de la tórtola europea realizado con cada grupo de variables: I y J, efecto independiente y conjunto de cada variable del modelo, respectivamente; %I, porcentaje de contribución independiente de cada variable del grupo; S, dirección del efecto; Zs, Zscore, test de aleatorización de la contribución independiente de cada variable; %DV, porcentaje de desviación explicada por cada variable del grupo; %Dev, porcentaje de desviación explicada por el modelo que incluye todas las variables de cada grupo. (* p < 0,05, ** p < 0,01, *** p < 0,001).

Presence

S %DV

Abundance I

%DV

Natural vegetation Scrub

3.77 2.95 14.62 4.68*** – 1.30

0.00 0.00 0.47 –0.69

ScrubPines

4.29 3.25 16.67 5.53*** + 1.48

0.01 0.00 5.53 –0.15

Forest

3.58 –0.44 13.88 4.3*** + 1.23

0.00 0.00 2.07 –0.53

DivForest

0.03 –0.03 0.13 –0.65 0.01

0.01 0.00 6.58 –0.02

DivScrub

1.69 1.56 6.54 1.8* – 0.58 0.00 0.00 1.07 –0.62

DivScrubPines 2.37 1.90 9.20 2.56** + 0.82 0.01 0.00 9.07 0.2 SpatialTerm

10.03 1.05 38.94 14.97*** 3.46

0.08 0.01 75.20 6.29***

%Dev

8.88

Farming DivWoodCrop 0.74 –0.31 3.25

0.3 0.26

0.00 0.00 1.32 –0.64

DivHerbCrop

0.99 0.97 4.33

0.67 0.34

0.00 0.00 1.72

WoodCrop

4.74 2.07 20.73 6.24*** + 1.63

0.00 0.00 2.39 –0.51

HerbCrop

8.00 2.29 34.99 10.32*** – 2.76

0.01 0.01 12.62

SpatialTerm

8.39 2.69 36.70 10.33*** 2.89

0.08 0.01 81.94 5.91***

%Dev

7.88

–0.57 0.38

Hydrological web River

2.05 –0.13 11.91 1.94* + 0.71 0.00 0.00 0.48 –0.65 0.09

RavGullies

0.72 –0.06 4.20 0.3 0.25 0.01 0.00 7.66 0.42 1.46

NWB

0.66 0.08 3.85 0.21 0.23 0.04 0.00 28.26 3.22* + 5.40

TotalChan

2.17 –0.12 12.64 2.24* – 0.75 0.02 0.00 9.56 0.61 1.83

SpatialTerm

11.57 –0.49 67.40 15.3*** 3.99

%Dev

5.92 19.09

0.09 0.00 54.04 7.56*** 10.32

Human pressure Unproductive 0.58 0.90 2.19 0.1 0.20 0.00 0.00 3.24 –0.36 IsolHouses

0.15 –0.08 0.57

HousDevel

2.06 1.78 7.72

–0.55 0.05

Urban

0.09 0.14 0.35 –0.62 0.03 0.01 0.00 4.71 –0.21

PavRoad

2.53 0.72 9.50 2.71** – 0.87 0.00 0.00 2.05 –0.47

UnPavRoad

12.25 0.03 45.97 18***

+ 4.22

0.00 0.00 2.92 –0.41

SpatialTerm

8.98 2.10 33.70 10.61*** 3.10

0.09 0.00 76.89 8.19***

1.7*

– 0.71

0.01 0.00 9.86

0.29

0.00 0.00 0.34

–0.63

%Dev 9.19

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Table 2. (Cont.)

Presence

Abundance

%DV

Climate OmbrIndex

0.63 0.75 4.82 0.23 0.22 0.01 0.00 9.87 0.2 1.28

ThermIndex

2.56 1.16 19.44 2.95** + 0.88

SpatialTerm

9.97 1.11 75.74 14.57*** 3.44 0.07 0.02 62.27 5.09*** 8.04

%Dev

4.54 12.92

0.03 0.02 27.86 2.01* + 3.60

Topography DistCoast

0.71 0.79 5.32 0.26 0.24 0.01 0.01 5.48 –0.2 0.67

AltMean

2.41 1.05 18.10 2.56** – 0.83

Slope

0.36 –0.35 2.69 –0.21 0.12 0.00 0.00 3.12 –0.46 0.38

SpatialTerm

9.84 1.25 73.89 11.97*** 3.39 0.06 0.02 61.14 5.74*** 7.50

% Dev

4.59

In the final model, which included the best variables from all the groups, the predictor with the highest independent contribution was scrub–pine mixture in the semi–arid (ScPineSa), with a positive effect. Furthermore, the cover of forests, woody crops and length of unpaved roads also had a positive effect, while herbaceous crops and dry scrublands had a negative influence (table 3). Abundance models Three variables were found to have a significant effect on abundance. NWB and the thermicity index had a significant positive effect, while altitude had a negative effect (table 2). When these three variables were analysed jointly in the final model, only NWB had a significant positive effect (table 3). Discussion We evaluated the habitat variables that could explain the occurrence and abundance of the European turtle dove in the province of Alicante. The groups of variables that explained a higher proportion of deviance in the occurrence analysis were natural vegetation, farming, and human pressure, while the most important groups in analyses of abundance were hydrological web, climate and topography. When considering the natural vegetation variables, the models for presence found that European turtle dove was associated with pine woods in an intermediate development stage. The mixture of scrubs and pines had a positive significant effect only on southern slopes and for a thermo–mediterranean and semiarid climate. This agreed with the results found by Saenz de Buruaga et al. (2012) in the Basque Country (north-

0.03 0.02 30.25 2.08* – 3.71

12.26

ern Spain), where this dove has been associated with patches of trees and scrublands. In countries such as the UK and Greece, this species prefers forest covered by medium–sized trees and abundant shrub vegetation (Browne et al., 2005; Bakaloudis et al., 2009). These results agree with studies proposing that its presence is limited by the availability of the resources that the habitat can provide for nesting (Browne et al., 2004). Our analyses also agree with other studies that identified a preference for a landscape covered by medium–aged pines (Bakaloudis et al. 2009; Hanane and Yassin, 2017). In Alicante, the European turtle dove also selects sunny spaces to breed, as found in studies in other countries (Browne et al., 2005; Bakaloudis et al., 2009). In Morocco, this species of dove shows a preference for Thuya forest (Tetraclinis articulata), an environment that is characterised by high temperatures and less precipitation (Hanane and Yassin, 2017). The length of unpaved roads is a positively selected variable that is related to the degree of human modification of the territory. This variable showed the highest independent contribution among the human pressure variables and presents the second contribution in the final analysis. The positive effect of this variable can be explained by the abundance of ruderal plant species that may be consumed by European turtle doves (Dunn and Morris, 2012; Gutiérrez–Galán and Alonso, 2016; Cramp, 1985; Browne and Aebischer, 2003) along edges of unpaved roads. This variable was included in the analysis as a measure of human impact and was expected to have a negative effect, but our analysis revealed that it had a positive effect. This can be explained if edges with ruderal species have disappeared from other places in cultivated areas owing to modern farming techniques, but remain mainly along unpaved roads.

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Table 3. Results of the HP analysis for presence and abundance of European turtle dove performed with the selection of the most explicative variables from all groups. (For abbreviations, see tables 1 and 2). Tabla 3. Resultado del análisis de partición jerárquica de la presencia y la abundancia de la tórtola europea realizado con la selección de las variables más explicativas de cada grupo. (Para las abreviaturas, véanse tablas 1 y 2) Presence Variable ScrubDry

Abundance

S %DV Variable

8.49 5.27 14.73 11.15*** – 2.92 NWB

UnPavRoad 9.85 2.43 17.10 12.85*** + 3.39 AltMean

Zs S %DV

0.04 0.01 27.66 2.68** + 4.48 0.02 0.03 16.17 1.32 2.62

ScPineSa 11.80 –1.06 20.49 15.87*** + 4.07 ThermIndex 0.02 0.03 12.71 0.85 2.06 HerbCrop

9.75 0.53 16.93 11.67*** – 3.36 SpatialTerm 0.06 0.02 43.46 4.8*** 7.04

WoodCrop 3.74 3.07 6.49 4.39*** + 1.29 %Dev

16.19

Forest

4.86 –1.73 8.44 6.08*** + 1.68

AltMean

1.78 1.68 3.09 1.61 0.61

River

1.12 0.79 1.95 0.84 0.39

SpatialTerm 6.20 4.88 10.77 7.59*** 2.14 %Dev 19.85

Length of rivers was significantly and positively related to the presence of the European turtle dove only when this variable was analysed with the hydrological web variables. It was not significant in the final analysis, where it was the least important variable. This positive effect therefore seems to be explained by the vegetation associated with rivers rather than by the presence of water itself (Saenz de Buruaga et al., 2012). Apart from the aforementioned unpaved roads, the human pressure variables contain two other variables (HousDevel and PavRoad) but both have a significantly negative effect. However, this effect was weak and these variables were not selected for the final analysis. These results suggest that the species can tolerate the presence of scattered houses and paved roads to a certain extent provided that the surrounding vegetation is suitable. Mason and Macdonald (2000) found that its presence in the UK was associated with residential areas. European turtle doves only use crops with trees (Gil–Delgado, 1981; Carrascal and Palomino, 2008; Hanane and Baamal, 2011). Thus, herbaceous crops have a negative impact on its probability of presence. Within woody crops, orchards of almonds and citrus and other fruiting trees groves have a positive effect on its presence. Cherry tree orchards, with a reduced distribution limited to some mountainous areas in this province, appear to have a negative effect. Dry and irrigated tree crops are known to be positively selected by the species throughout its distribution (Antón–Recasens, 2004; Gil–Delgado, 1981; Hanane and Baamal, 2011) because they offer both food and potential nesting places. Thus, it was surprising to

find that the effect of olive groves was not detected. This might be explained by the fact that in our study area olive trees are mostly grown in small groves in the midst of other crops that cover a larger extension. For abundance, the analyses of groups of variables revealed a potentially positive effect of the thermicity index and NWB, along with a negative effect of altitude. However, the combined analysis of these variables indicated that only NWB presented a significant contribution. Doves need proximity to water to avoid dehydration and weight loss (Bartholomew and Macmillen 1960; Macmillen 1962; Willoughby 1966; McKechnie et al., 2016), which may explain the positive effect of NWB, and thus the presence of water bodies should favour its populations. Today, the European turtle dove shows a generalised population decline in the Western Palearctic (Dias et al., 2013; Hanane, 2017). The habitat effects revealed by this and other studies may be useful for developing measures to mitigate its decline. Our results point out that the vegetation types present are the main determinants of the probability of this turtle dove being found in a Mediterranean province whose climate is mostly semiarid. The species is most likely to be present in mosaics of Aleppo pines and scrub, mid–sized pines and orchards. Thus, the conservation of this kind of landscape would contribute to maintaining appropriate habitats for this species. The European turtle dove seems relatively tolerant to disturbances caused by scattered houses and paved roads, and even seems to benefit from the presence of unpaved roads. The most negative effect is caused by pure scrub cover areas and herbaceous crops. These results suggest that management may contri-

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bute to significantly improve habitat quality for these turtle doves. Despite forest fires, forests are growing in Spain and Alicante. Forest expansion occurs in some places in this province at the expense of tree orchards on terraces. Thus, substituting orchard crops for pines may generate a suitable habitat for the dove at early or intermediate phases, when trees are half grown, but they would be unsuitable when the trees are mature. In lowlands, abandoned orchards are substituted by scrubland that this is unsuitable for the European turtle dove. This implies that their survival in such areas would depend on the length of survival of the abandoned crop. Acknowledgements These data are based on the census for the Atlas of Breeding Birds from the Province of Alicante. The authors gratefully acknowledge S. Hanane (Centre National de la Recherche Forestière) and H. Lormée (Office National de la Chasse et de la FauneSauvage) for their comments and recommendations on this paper. We are grateful to E. Mellink for the text review. Reference Antón–Recasens, M., 2004. Tórtora, Streptopelia turtur. In: Atles dels Ocells nidificants de Catalunya 1999–2002: 268–269 (J. Estrada, V. Pedrocchi, L. Brotons, S. Herrando, Eds.). Institut Català d’Ornitologia. Lynx Edicions, Barcelona. Balmori, A., 2003. Tórtola Europea, Streptopelia turtur. In: Atlas de las aves reproductoras de España: 306–307 (R. Martí, J. C. del Moral, Eds.). Dirección General de Conservación de la Naturaleza. Sociedad Española de Ornitología, Madrid. Bakaloudis, D. E., Vlachos, C. G., Chatzinikos. E., Bontzorlos, V., Papakosta, M., 2009. Breeding habitat preferences of the turtledove (Streptopelia turtur) in the Dadia–Soufli National Park and its implications for management. European Journal of Wildlife Research, 55(6):597–602. Bartholomew, G. A., Macmillen, R. E., 1960. The water requirements of Mourning Doves and their use of sea water and NaCl solutions. Physiological Zoölogy, 33(3): 171–178. BirdLife International, 2016. Species factsheet: Streptoplelia turtur. Available online at: http:// datazone.birdlife.org/species/factsheet/european– turtle–dove–streptopelia–turtur/text [Accessed on 23 October 2016]. Boutin, J. M., Lutz, M., 2007. Management Plan for Turtle dove (Streptopelia turtur) 2007–2009. European Commission, Luxembourg. Browne, S. J., Aebischer, N. J., 2003. Habitat use, foraging ecology and diet of Turtle Doves Streptopelia turtur in Britain. Ibis, 145(4): 572–582. Browne, S. J., Aebischer, N. J., Crick, H. Q. P., 2005. Breeding ecology of Turtle Doves Streptopelia turtur in Britain during the period 1941–2000: an analysis of BTO nest record carsd. Bird Study, 52(1): 1–9.

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Browne, S. J., Aebischer, N. J., Yfantis, Y., Marchant, J. M., 2004. Habitat availability and use by Turtle Doves Streptopelia turtur between 1965 and 1995: an analysis of Common Birds Census data. Bird Study, 51(1): 1–11, Doi: 10.1080/00063650409461326 Carrascal, L. M., Palomino, D., 2008. Las aves comunes reproductoras en España. Población en 2004–2006. SEO/BirdLife, Madrid. Chevan, A., Sutherland, M., 1991. Hierarchical Partitioning. The American Statistician, 45(2): 90–96, Doi: 10.1080/00031305.1991.10475776 Cramp, S., 1985. The birds of the Western Palearctic. Oxford University Press, Oxford. Dias, S., Moreira, F., Beja, P., Carvalho, M., Gordinho, L., Reino, L., Oliveira, V., Rego, F., 2013. Landscape effects on large scale abundance patterns of turtle doves Streptopelia turtur in Portugal. European Journal of Wildlife Research, 59(4): 531–541. Dunn, J. C., Morris, A. J., 2012. Which features of UK farmland are important in retaining territories of the rapidly declining Turtle Dove Streptopelia turtur? Bird Study, 59(4): 394–402. Gibbs, D., Barnes, E., Cox, J., 2010. Pigeons and doves: A guide to the pigeons and dove of the world. A&C Black Publishers, London. Gil–Delgado, J. A., 1981. Bird community in orange groves. In: Bird Census and Mediterranean landscape, 100–106 (F. J. Purroy, Ed.). Proceedings VII Int. Con. Bird Census IBCC V Meeting EOAC, Leon, Spain. Gutiérrez–Galán, A., Alonso, C., 2016. European turtle dove Streptopelia turtur diet composition in southern Spain: the role of wild seeds in Mediterranean forest areas. Bird Study, 63(4): 490–499. Hanane, S., 2017. The European Turtle–Dove Streptopelia turtur in Northwest Africa: A review of current knowledge and priorities for future research. Ardeola, 64(2): 273–287. Hanane, S., Baamal, L., 2011. Are Moroccan fruit orchards suitable breeding habitats for Turtle Doves Streptopelia turtur? Bird Study, 58(1): 57–67. Hanane, S., Yassin, M., 2017. Nest–niche differentiation in two sympatric columbid species from a Mediterranean Tetraclinis woodland: Considerations for forest management. Acta Oecologica, 78: 47–52. Harris, S. J., Massimino, D., Gillings, S., Eaton, M. A., Noble, D. G., Balmer, D. E., Procter, D., PearceHiggins, J. W., Woodcock, P., 2018. The Breeding Bird Survey 2017. British Trust for Ornithology, Thetford. Jiguet, F., 2016. Les résultats nationaux du programme STOC de 1989 à 2015. Available online at: http://vigie nature.mnhn.fr/page/tourt erell e-desbois [Accessed on 2 February 2020]. Jongman, R. H. G., Ter Braak, C. J. F., Van Tongeren, O. F. R., 1995. Data analysis in community and landscape ecology. Cambridge University Press, Cambridge. Legendre, P., 1993. Spatial autocorrelation: Trouble or new paradigm? Ecological Modelling, 74(6): 1659–1673. López–Iborra, G. M., Bañuls–Patiño, A., Zaragozí– Llenes, A., Sala–Bernabeu, J., Izquierdo– Rosique,

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A., Martínez–Pérez, J. E., Ramos–Sánchez, J., Bañuls–Patiño, D., Arroyo–Morcillo, S., Sánchez– Zapata, J. A., Campos–Roig, B., Reig–Ferrer, A., 2015. Atlas de las aves nidificantes en la provincia de Alicante. Publicacions de la Universitat d'Alacant–SEO/Alicante, Alicante. López–Iborra, G. M., Limiñana, R., Pavón, D., Martínez–Pérez, J. E., 2011. Modelling the distribution of short–toed eagle (Circaetus gallicus) in semi–arid Mediterranean landscapes: Identifying important explanatory variables and their implications for its conservation. European Journal of Wildlife Research, 57(1): 83–93, Doi: 10.1007/ s10344-010-0402-0 Macmillen, R. E., 1962. The minimum water requirements of Mourning doves. The Condor, 64(2): 165–166. MacNally, R., 2002. Multiple regression and inference in ecology and conservation biology: further comments on identifying important predictor variables. Biodiversity and Conservation, 11: 1397–1401, Doi: 10.1023/A:1016250716679 Madroño, A., González, C., Atienza, J. C., 2004. Libro Rojo de las aves de España. Dirección General para la Biodiversidad–SEO/BirdLife, Madrid. Margalef, R., 1973. Some critical remarks on the usual approaches to ecological modelling. Investigación Pesquera, 37(3): 621–640. Mason, C. F., Macdonald, S. M., 2000. Influence of landscape and land–use on the distribution of breeding birds in farmland in eastern England. Journal of Zoology, 251(3): 339–348. Mason, C. H., Perreault, W. D., 1991. Collinearity, power, and interpretation of multiple regression analysis. Journal of Marketing Research, 28(3): 268–280. McKechnie, A. E., Whitfield, M. C., Smit, B., Gerson, A. R., Smith, E. K., Talbot, W. A., McWhorter, T. J., Wolf, B. O., 2016, Avian thermoregulation in the heat: efficient evaporative cooling allows for extreme heat tolerance in four southern hemisphere columbids. The Journal of Experimental Biology, 219: 2145–2155, Doi: 10.1242/jeb.138776

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Moreno–Zarate, L., Estrada, A., Peach, W., Arroyo, B., 2020. Spatial heterogeneity in population change of the globally threatened European turtle dove in Spain: The role of environmental favourability and land use. Diversity and Distributions, 26: 818–831, Doi: 10.1111/ddi.13067 Olea, P. P., Mateo–Tomás, P., de Frutos, Á., 2010. Estimating and Modelling Bias of the Hierarchical Partitioning Public–Domain Software: Implications in Environmental Management and Conservation. Plos One, 5(7): e11698, Doi: 10.1371/journal. pone.0011698 R Core Team, 2018. R: A language and environment for statistical computing. Vienna, Austria. Rigual, A., 1972. Flora y vegetación de la provincia de Alicante. Instituto de Estudios Alicantinos Diputación Provincial de Alicante, Alicante. Rivas–Martinez, S., 1987. Memoria del mapa de series de vegetación de España. Instituto para la Conservación de la Naturaleza, Madrid. Saenz de Buruaga, M., Onrubia, A., Fernández–García, J. M., Campos, M. A., Canales, F., Unamuno, J. M., 2012. Breeding habitat use and conservation status of the turtle dove Streptopelia turtur in northern Spain. Ardeola, 59(2): 291–300. SEO/BirdLife, 2010. Aves exóticas invasoras en España: propuesta inicial de lista para el catálogo nacional de EEI. Grupo de Aves Exóticas–SEO/ BirdLife, Madrid. Serra, P., Pons, X., Sauri, D., 2008. Land–cover and land–use change in a Mediterranean landscape: a spatial analysis of driving forces integrating biophysical and human factors. Applied Geography, 28: 189–209, Doi: 10.1016/j.apgeog.2008.02.001 Walsh, C., MacNally, R., 2015. The hier.part package. Hierarchical Partitioning. Documentation for R: A language and environment for statistical computing. R Foundation for statistical Computing, Vienna, Austria. Available online at: http://www.rproject.org [Accessed on 10 February 2016]. Willoughby, E. J., 1966. Water requirements of the Ground Dove. The Condor, 68: 243–248.

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Machine learning as a successful approach for predicting complex spatio–temporal patterns in animal species abundance B. Martín, J. González–Arias, J. A. Vicente–Vírseda Martín, B., González–Arias, J., Vicente–Vírseda, J. A., 2021. Machine learning as a successful approach for predicting complex spatio–temporal patterns in animal species abundance. Animal Biodiversity and Conservation, 44.2: 289–301, Doi: https://doi.org/10.32800/abc.2021.44.0289 Abstract Machine learning as a successful approach for predicting complex spatio–temporal patterns in animal species abundance. Our aim was to identify an optimal analytical approach for accurately predicting complex spatio– temporal patterns in animal species distribution. We compared the performance of eight modelling techniques (generalized additive models, regression trees, bagged CART, k–nearest neighbors, stochastic gradient boosting, support vector machines, neural network, and random forest –enhanced form of bootstrap. We also performed extreme gradient boosting –an enhanced form of radiant boosting– to predict spatial patterns in abundance of migrating Balearic shearwaters based on data gathered within eBird. Derived from open–source datasets, proxies of frontal systems and ocean productivity domains that have been previously used to characterize the oceanographic habitats of seabirds were quantified, and then used as predictors in the models. The random forest model showed the best performance according to the parameters assessed (RMSE value and R2). The correlation between observed and predicted abundance with this model was also considerably high. This study shows that the combination of machine learning techniques and massive data provided by open data sources is a useful approach for identifying the long–term spatial–temporal distribution of species at regional spatial scales. Key words: Balearic shearwater, Machine learning, Random forest, Chlorophyll, NAO index Resumen Aprendizaje automático: una buen método para predecir patrones espacio–temporales complejos en la abundancia de especies animales. Nuestro objetivo fue determinar un método analítico óptimo para predecir de manera precisa patrones espacio–temporales complejos en la distribución de especies animales. En concreto, utilizando los datos recopilados en el proyecto eBird sobre la pardela balear, que es un ave migratoria, se compararon ocho técnicas diferentes de elaboración de modelos (modelos aditivos generalizados, árboles de regresión, bagged CART, k–nearest neighbors, stochastic gradient boosting, máquinas de vectores de soporte, redes neuronales, así como el bosque aleatorio –una forma mejorada de bootstrap– y extreme gradient boosting –una forma mejorada de gradient boosting) con objeto de predecir los patrones espaciales observados en la abundancia de esta especie. Utilizando conjuntos de datos de código abierto, se han cuantificado los indicadores de los sistemas frontales y la productividad de los océanos que ya se habían empleado con anterioridad para caracterizar los hábitats oceánicos de las aves marinas y, posteriormente, se han utilizado como predictores en los modelos. El bosque aleatorio resultó ser el modelo que ofreció el mejor rendimiento de acuerdo con los parámetros evaluados (RMSE y R2). La correlación obtenida con este modelo entre la abundancia observada y la predicha también fue considerablemente alta. En este trabajo, mostramos la utilidad de combinar técnicas de aprendizaje automático y los datos masivos proporcionados por diferentes fuentes de código abierto para determinar la distribución espacio–temporal a largo plazo de las especies a escalas espaciales regionales. Palabras clave: Pardela balear, Aprendizaje automático, Bosque aleatorio, Clorofila, Índice NAO Received: 27 V 21; Conditional acceptance: 30 VI 21; Final acceptance: 24 VIII 21 Beatriz Martín, Randbee Consultants, c/ Carretería 67, Málaga, Spain.– Julio González–Arias, Juan A. Vicente– Vírseda, Departamento de Economía de la Empresa y Contabilidad, Paseo de la Senda del Rey 11, 28040 Madrid Corresponding author: B. Martín. E–mail: beatriz.martin@randbee.com ORCID ID: 0000-0001-6893-2187 ISSN: 1578–665 X eISSN: 2014–928 X

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Introduction The continuous technical development of electronic devices such as geolocators, GPSs, and PTT devices for tracking animal movements have promoted the understanding of the distribution of many species (Katzner and Arlettaz, 2020). However, for migratory species, even with good sample sizes, tracking data cannot give a fully representative sample from all the possible migratory route alternatives. The reason for this is that data derived from these studies are temporally constrained to a few years of monitoring of a few individuals, but migratory species usually exhibit a great capacity to alter migratory behavior in response to environmental variability and specific individual traits (Martín et al., 2016). Consequently, even though they can provide highly detailed information on an individual's movements, electronic devices have a limited ability to improve our understanding of the adaptation of migration strategies to deal with a changing environment at both population and species levels (Martín et al., 2019). In contrast, although direct observation methods such as census are unable to detect birds when they use areas out of human sight, they can provide a valuable overall picture despite the missing information if we apply predictive models to fill the spatial and temporal gaps in the original datasets (Gouraguine et al., 2019). In this sense, active public involvement in scientific research (citizen science) has become an excellent source of data for scientists and policymakers (Strasser et al., 2019). Citizen science projects provide observations of millions of species each year (Kelling et al., 2013; Chandler et al., 2017). Compared with the detailed data gathered from electronic devices, the massive datasets from citizen science projects have the advantage of offering low–cost information on long–term temporal and large spatial extents from many individuals belonging to several populations. Prominent among citizen science projects devoted to birds and used by a large number of scientists is eBird (eBird, 2017), a biodiversity–related citizen science project gathering records of birds provided by professional and amateur ornithologists around the world. Caveats on the use of datasets from citizen science projects are related to errors in species identification, biases in count estimates and uneven sampling effort, both in terms of temporal and spatial coverage, among others, which make volunteer data highly variable in terms of precision. Recent advances in Big Data analysis and, more specifically, in Data Mining techniques, thanks to the application of artificial intelligence (AI) and, more specifically, machine learning (ML) techniques, allow to obtain robust predictions from these 'noisy' and incomplete datasets (Schain, 2015). For this reason, these techniques have become an extremely useful tool to extract relevant information in many disciplines of knowledge. As model species, we focused on the Balearic shearwater (Puffinus mauretanicus), a critically endangered seabird species (BirdLife International, 2018) which spends about one quarter of the year on migration (Guilford et al., 2012). Previous research

founded on boat–based survey counts showed Balearic shearwater abundance to be extremely difficult to predict, and it failed to provide reliable predictions on the spatial distribution of shearwater numbers (Oppel et al., 2012). Our aim was to identify an optimal analytical approach for accurately predicting spatio–temporal patterns in the abundance of this species during migration from observations collected by professional and amateur ornithologists in citizen science projects, namely eBird database. To this end, we built several predictive models applying traditional statistical analysis (generalized additive models) and the latest machine learning techniques (neural networks, gradient boosting, random forest, support vector machines, among others). With these models we predicted the abundance of the Balearic shearwater along its migratory route over the Mediterranean and Atlantic coasts. These models were founded on the assumption that there is a relationship, direct or indirect, between environmental variables and the shearwater distribution. In this way, from a selected sect of environmental predictors derived from large–scale open datasets (NCEP/NCAR, AIS, NOAA, among others), we modelled the abundance of shearwaters over time and across space. The various models were then compared in order to determine the best one in terms of accuracy and predictive ability. Our final aim was to describe a successful analytical approach that can be extended to other animal species for which citizen science projects are collecting abundance distribution data. Material and methods Study species The Balearic shearwater is included as 'Critically Endangered' on the IUCN Red List (BirdLife International, 2018) and is also considered as threatened bird for the European Union (rare or vulnerable bird species as listed in annex I of the E.U. Bird Directive). Most of the population of Balearic shearwater leaves the Mediterranean each year after breeding (Guilford et al., 2012) and stays mainly in the Atlantic during the non-breeding season (Arcos, 2011). At sea, it usually occurs in productive shelf areas related to oceanographic frontal systems (Louzao et al., 2006; Oppel et al., 2012; Pérez–Roda et al., 2017). During migration, Balearic shearwaters tend to fly very close to the shoreline (Arroyo et al., 2014), with an average off-shore distance of about 1,190 m at the Strait of Gibraltar (Mateos and Arroyo, 2011). Fluctuations in Balearic shearwater migration, both seasonally and inter-annually, seem to be related to changes in food resources (Wynn et al., 2007; Jones et al., 2014). The diet of the Balearic shearwater includes small pelagic and also demersal fish, frequently obtained from trawling discards. The species can also feed on plankton and macrozooplankton, specifically krill (Arcos and Oro, 2002; Louzao et al., 2015).

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North Sea Irish Sea

1–97 98–417 418–1120 1121–2059

Atlantic Ocean

Balearic Islands

Mediterranean Sea Strait of Gibraltar

Fig. 1. Spatial distribution of the shearwater abundance data during the post–breeding migration period (May–August). Years 2005–2017; n = 1,881. Fig. 1. Distribución espacial de los datos sobre la abundancia de la pardela correspondientes al periodo de migración post–nupcial (mayo–agosto). Años 2005–2017; n = 1.881.

The extent of the study area (fig. 1) covers the whole distribution range of the Balearic shearwater throughout the year (BirdLife International, 2018). Variables in the models Response variable Our response variable was Balearic shearwater abundance (abundance considering only presence, thus, absence of abundance –i.e., zeros– was not considered in the analysis) (Pearce and Boyce, 2006), expressed as the number of birds sighted on a given date at a given latitude/longitude, during migration. Daily abundance of the species was recorded from 1964 to 2018, both as opportunistic sighting records and within systematic effort–based surveys (with a standardized duration of the sampling effort) obtained from the online database eBird (eBird, 2017). Due to the opportunistic nature of some of these data, we needed powerful modelling techniques to obtain robust results (see below). Although it was possible

to differentiate opportunistic and systematic surveys within the dataset, we opted to keep all records for our analysis in order to test the robustness of the modelling approach in case our methods will be extended to other species for which this information is not available. Specifically, we considered abundance during the 'post–breeding migration' period, defined as the northward migration of birds leaving the Mediterranean and molt, between May 1st–August 30th (Mourino et al., 2003; Yésou, 2003). Environmental predictors Shearwater abundance was modelled using 15 environmental predictors (table 1) that have been previously described to be related with the spatial distribution at sea of this and other seabird species (Yésou, 2003; Dias et al., 2012; Oppel et al., 2012; Jones et al., 2014). As migrating birds need to replenish energy reserves during stopover periods at key locations where they maximize their refueling

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opportunities (Wynn et al., 2007; Benoit–Bird et al., 2013), fluctuations in bird abundance during migration are closely related to changes in food resources (Wynn et al., 2007). Therefore, most of the variation in seabirds appears to be mediated by changes in prey abundance (Frederiksen et al., 2006; Benoit–Bird et al., 2013). We used chlorophyll concentration (Chla, measured in mg/m3) as a proxy of marine productivity (Wakefield et al., 2009). Specifically, we downloaded satellite–based monthly products at 4 km spatial resolution for the 2003–2017 period (JRC Data Catalogue; http://gmis.jrc.ec.europa.eu/satellite/4km/). Discard availability may influence the at–sea distribution of shearwaters (Cortés et al., 2018; Genovart et al., 2018). As a proxy of food availability (both discards and fishes) we used information on fisheries. This information was inferred from the distribution of fishing vessels (Natale et al., 2015) between 2014 and 2015, sourced by the JRC Data Catalogue at 1 km resolution (http://gmis.jrc.ec.europa.eu/dataset/ jrc–fad–ais1415). This product identifies the areas where fishing is most frequent. Vessel tracking data were derived from the Automatic Identification System (AIS), an open source data system that allows analysis of the relation between fishing communities and fishing areas at high spatial resolution across Europe. Specifically, data used to build the map consists of 150 million positions from European fishing vessels above 15 m in length. In spite of its limited temporal coverage, the main strength of this dataset is its fine spatial resolution. This proxy on food availability in spatial terms, however, is complemented in our analysis with the high temporal resolution in marine productivity provided by chlorophyll concentration. Together with chlorophyll concentration, sea surface temperature (SST) is also a proxy of water mass distributions, frontal systems, and ocean productivity (Ramos et al., 2012; Robinson et al., 2013; Afán et al., 2014). In shearwaters, which perform dynamic soaring flight, oceanic winds are also of major importance in modulating migratory behavior (González–Solís et al., 2009). Wind speed and direction usually affect both migratory behavior (Catry et al., 2011; Dias et al., 2012; but see Dell'Ariccia et al., 2018) and seabird detectability by the observers (Martín et al., 2019). Data on SST and wind were obtained through RNCEP package (Kemp et al., 2012) in R (R Core Team, 2018) which allowed to request data from the NCEP/NCAR Reanalysis dataset (NOAA ESRL Physical Sciences Division; https://www.esrl.noaa.gov/psd/data/gridded/ data.ncep.reanalysis2.html) for a specified range of space and time based on the observations of shearwaters. Specifically, we requested daily values at midday at the surface level, since Balearic shearwaters tend to migrate at a very low height (Martín et al., 2019). Variables requested were 'air.995' (air temperature), 'uwind.sig995' (u–wind component) and 'vwind.995' (v–wind component). RNCEP interpolates the nearest data value (in terms of location and time) in the NCEP/ NCAR dataset to provide the requested information. We also considered the standard deviation (associated standard deviation of the points used to perform the interpolation) of the meteorological values (u–wind,

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v–wind and temperature) used for this interpolation as additional predictors in our models (table 1). General flight activity of seabirds increases during moonlit nights, and moon phase has been shown to affect shearwater migration behaviour (Dias et al., 2012). As changes in the flying patterns during the night may also affect daytime flights, we considered the daily fraction of the moon illuminated at midnight as an additional predictor, sourced by the U.S. Naval Observatory and Astronomical Applications Department (http://aa.usno.navy.mil/data/). Together with the usual inter–annual variability in food resources, long–term climate change may also affect the at–sea distribution of this species (Wynn et al., 2007; Votier et al., 2008; Luczak et al., 2011), likely through effects on fish stocks (Tsikliras et al., 2019). The possible impact of climate change on the distribution of shearwaters at a regional level was taken into account through the North Atlantic Oscillation Index (NAO; Visbeck et al., 2001) obtained from the Climate Prediction Center (U.S. National Weather Service, NOAA), as monthly data from 1950 to 2018 (http:// www.cpc.ncep.noaa.gov/products/precip/CWlink/ pna/nao.shtml). As SST and wind predictors already provided a daily resolution of the weather conditions at specific locations, monthly values rather than daily NAO index were preferred, as a surrogate of regional climate conditions affecting shearwater migration at a more general and longer time frame. Water depth has been shown to influence seabird distributions (Yen et al., 2004). Specifically, Balearic shearwaters occur in relatively shallow, coastal waters along the shoreline during migration (Martín et al., 2020). Bathymetry was used as a surrogate of coastal–pelagic areas (Afán et al., 2019). Bathymetric data were sourced by European Marine Observation and Data Network (EMODnet; http://www.emodnet-bathymetry.eu/data-products) at ~200 m resolution although they were later aggregated through a moving–average window to 10 km resolution in order to obtain a more general bathymetry pattern regarding the shearwater observation location. This cell size was selected as a compromise between this general pattern in bathymetry and the sufficient resolution for conservation purposes. Fluctuations in seabird abundance during migration are closely related to changes in food resources (Frederiksen et al., 2006; Benoit–Bird et al., 2013). However, photoperiodic cues and/or endogenous rhythms may also modulate seabird breeding and migration periods (Marshall and Serventy, 1959). Therefore, apart from food availability, migration decisions in Balearic shearwaters might be partially dictated by day length and/or by an internal rhythm that make the bird move instinctively into the north–west, while the post–breeding season progresses, and then into the south–east during the pre–breeding period. Date, longitude and latitude variables allow us to include the endogenous rhythm of the bird in the models. In addition, longitude and latitude can be indirect proxies of the effects that variable wintering sites, length of the route and en–route environmental conditions may pose to migrant shearwaters. Finally, due to potential

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Table 1. Descriptive summary of the predictors in the models. Sources: EMODnet, European Marine Observation and Data Network; AIS, Automatic Identification System (JRC Data Catalogue); NCEP/ NCAR, Reanalysis dataset (NOAA ESRL Physical Sciences Division); MERIS, JRC Data Catalogue; CPC, Climate Prediction Center (US National Weather Service, NOAA); USNO & AAD, U.S. Naval Obervatory and Astronomical Applications Department. Tabla 1. Resumen descriptivo de las variables utilizadas en los modelos. (Para las abreviaturas de las fuentes, véase arriba). Name Description

Unit

Source

batim

Bathymetry

meters

EMODnet

fish

Fishing intensity

number

AIS

tmmean

Mean temperature

Period 2012 2014–2015

of vessels ºK

NCEP/NCAR

1964–2018

tmstd

Atandard deviation of mean temperature

uwmean

Mean wind speed (u–wind:

east–west direction)

uwstd

Standard deviation of mean wind speed

(u–wind: east–west direction)

vwmean

Mean wind speed (v–wind: north–south

ºK

NCEP/NCAR

1964–2018

m/s

NCEP/NCAR

1964–2018

m/s

NCEP/NCAR

1964–2018

m/s

NCEP/NCAR

1964–2018

NCEP/NCAR

1964–2018

direction) vwstd

Standard deviation of mean wind speed

(v–wind: north–south direction)

clorof

Chlorophyll concentration

mg/m3 MERIS

2004–2017

NAO

NAO index

–

CPC

1950–2018

moon

Daily fraction of the moon illuminated

at midnight

USNO & AAD

2005–2018

differences in the rates of change of the environmental predictors across space, interactions between predictors and 'latitude' and longitude, and between 'year' and 'latitude', may allow to quantify both the spatial and temporal heterogeneity in the migratory responses. Statistical analysis As usual in count data, abundance of shearwaters followed a Poisson distribution (Hilbe, 2014). Although there is a lack of assumptions in the distribution of the data in machine learning models, when we use approaches where data partitioning is applied, we can obtain better results if we model a dependent variable homogenously distributed, because the model dispersion increases as long as the variable increases. Supporting the previous statement, we observed that this log transformation increased the variance explained by all the models (up to 1.4 times in the case of the RF model). Therefore, before building our models we log–transformed (natural log) the abundance data. The estimates of abundance

m/s

obtained from the models were then scaled back to a linear scale before assessing the model performance. In addition, before modelling, predictor variables were pre–processed: centered and scaled (subtracting the mean of the predictor's data from the predictor values and then dividing by the standard deviation). To evaluate the different performance, up to eight different modelling techniques were carried out on the post–nuptial migration dataset. The set of models constitutes a representative sample of the various machine learning techniques available for predictive analyses in which the dependent variable is quantitative (regression approaches). Specifically, regression trees (CART), bootstrap aggregation (bagged CART), extreme gradient boosting, stochastic gradient boosting, K–nearest neighbours (KNN), support vector machine (SVM) and multilayer perceptron neural network (MLP). In addition to these machine learning approaches, we include Generalized Additive Models (GAM). GAMs are simple, transparent, and flexible models which do not assume a linear relationship between independent and dependent variables.

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This approach is suitable to model count data such as animal abundance (Zuur et al., 2009). Apart from flexibility, in contrast with other multivariate models, such as multivariate adaptive regression splines (MARS) and machine learning techniques, it provides an interpretable solution, thus providing a good balance between the interpretable linear model and the extremely flexible, but 'black box' iterative/learning algorithms. To enhance model performance, according to the nature of the response variable (see above), we applied a GAM with Poisson distribution and log–link function. Prior to modelling we assessed collinearity from the total set of environmental predictors through the correlation between pairs of variables (Pearson correlation coefficient). Although machine learning techniques are not as highly subjected to the effects of collinearity as traditional statistical approaches, computation timing and model results can benefit from the removal of any redundant features from the training dataset, irrespectively of the model's algorithm. Comparisons between specific modelling techniques and particular models were based on the Root Mean Square Error (RMSE; the average difference between the observed known values of the outcome and the predicted value by the model; Gareth et al., 2014), and on R2, measured in the caret package in R (R Core Team, 2018) as the squared correlation between observed and predicted values (Kuhn, 2008). The validation procedure of the models was based on a random split of the abundance data into training and test data (setting aside 20 % of the data for testing the models; Araújo et al., 2006). Performance of the built models was estimated using 5–fold cross–validation on the training data. Models were calibrated on the training data and then evaluated on the test data to determine the model’s ability to generalize to other datasets. Except for the GAM analysis, for model building and assessment we used the caret package (Kuhn, 2008) in R (R Core Team, 2018). Compared to a Gaussian identity–link GAM approach, a Poisson log–link GAM on the same training dataset increased both RMSE (from 124.13 to 139.55) and R2 (from R2 = 0.19 to R2 = 0.36). As Poisson log–link GAMs are not supported by caret, predictions (and R2) for these models were derived using the mgcv package (Wood, 2011) also in R. However, to ensure minimum RMSE values that were comparable between modelling techniques, RMSE for a Gaussian GAM model was quantified using caret. All the models were built using the total set of predictors described in table 1. Most of the models assessed have parameters that must be tuned to obtain an optimal fitting. To determine the parameter values offering the best fit, we specified a set of tuning values to be tested during the calibration of the models. Generally, we applied a grid search method, thus we evaluated the model over different combinations of parameters included in the grid (table 2). Specifically for the GBM models, the range for the parameter tuning was based on recommendations derived from previous research (Friedman, 1999a, 1999b; Friedman et al., 2001; Ridgeway, 2005). To identify the model with the optimal parameter combi-

nation (providing the best fit) we compared the RMSE values (see supplementary material) of the models (Gareth et al., 2014). To detect significant differences in the performance between modeling techniques, all pair–wise differences in RMSE and R2 over the model resamples were quantified and tested to assess whether the difference was equal to zero. The resulting confidence level was adjusted using Bonferroni correction (Kuhn, 2008). We also assessed the relative importance of the variables used for predicting shearwater abundance in the different models using the varImp function in caret package (Kuhn, 2007). Results Assessment of the available eBird data showed that records before 2005 (from 1964 to 2004) were scarce and did not properly cover the migration periods. For this reason, to ensure a minimum sample size for estimating predictions, models were built only with data from after 2005. Similarly, we did not consider data for 2018 in the analysis as they were seasonally incomplete. After data filtering, a total of 1,881 post–breeding records (ranging from 1 to 2,059 birds; median = 5) on 58,901 individual observations of shearwaters remained for the years 2005–2017. According to the migration areas which are known to be used by the study species (BirdLife International, 2018), the spatial representation of the remaining observations was good (fig. 1). The assessment of collinearity, measured as Pearson correlation coefficients between pairs of predictors, showed non–significant results at the 0.05 p–level for any of the pairs of variables, thus all predictors included in the initial set were considered for modelling. According to the cross–correlation resamples (fig. 2), the best modelling technique was Random Forest (RF), showing the lowest RMSE mean value among resamplings (mean RMSE = 1.11) and the largest R2 (mean R2 = 0.47), closely followed by the results from XGBM (RMSE = 1.13; R2 = 0.44). Differences in RMSE between models were only significant in the case of MLP against the best modelling techniques (RF, XGBM and GBM, in that order). The worst techniques both considering RMSE and R2 were MLP followed by GAM, although the GAM model with Poisson distribution and log–link function almost doubled the R2 value (mean R2 = 0.36; min = 0.33; max = 0.41). Differences in RMSE and also in R2 of KNN against RF, XGBM and SVM were almost significant at the Bonferroni p–level (p–value < 0.1). CART was also significantly better than MLP in terms of R2 (fig. 3). The resamplings of all the modelling techniques assessed, including RF, showed high variability in terms of RMSE and R2 values, indicating that both errors and the predictive ability of the models are considerably variable depending on the data subset used in each resampling. Mainly for this reason, the difference in terms of performance

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Table 2. Parameter values used for model calibration. Tabla 2. Valores de los parámetros utilizados en la calibración de los modelos. Model Parameters

Tuning sequence

Tuning by

Method

Bagging # baggings

10, 20, 30, 35, 40

Manually assessed

0–0.07

0.001

Grid search

CART cp Extreme gradient boosting # rounds

200–1000

eta

0.01, 0.015, 0.025, 0.05, 0.1, 0.3

Grid search

maximum depth

3, 5, 7, 9, 12, 15, 17, 25

Grid search

gamma

0.05, 0.1, 0.3, 0.5, 0.7, 0.9, 1.0

colsample_bytree

0.6, 0.7, 0.8, 0.9, 1.0

min_child_weight

1, 3, 5, 7

subsample

0.6, 0.7, 0.8, 0.9, 1.0

Gradient boosting interaction depth

4–10

# trees

0–2,500

Grid search

500

Grid search

shrinkage

0.001, 0.01, 0.1

Grid search

minimum number of

5, 10, 20

Grid search

0.2, 0.3, 0.5

Manually assessed through

observations in tree's terminal nodes (n.minobsinnode) sampling fraction

(bagg.fraction)

iteratively repeating the grid

search (see suppl. material)

KNN k

1–25

Grid search

1–25

Grid search

mtry

1–25

Grid search

# trees

1000, 1500, 1600, 1625,

1650, 1700, 1725, 1750, 2000

Neural network # layers Random forest 500

Manually assessed through iteratively repeating the grid search (see suppl. material)

Support vector machines sigma

60–180/400; 80:100/400;

100:120/400; 120–140/400;

refined based on RMSE

140:180/400

results (see suppl. material)

1:10/10; 10:20/10; 20:30/10;

Grid search that was iteratively

30/10

Grid search that was iteratively

refined based on RMSE results (see suppl. material)

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RMSE

Rsquared

MLP

CAM

KNN

CART

BAGG

SVM

XGBM

GBM

RF 1.1

1.2

1.3

1.4

1.5

0.1

0.2

0.3

0.4

0.5

Confidence level 0.95 Fig. 2. Results from the resampling (5–fold cross–validation) analysis of the various models assessed. Range (minimum–maximum) and mean values; 95 % confidence intervals for the medians. Models based on data for 2005–2017: * GAM, Gaussian distribution and identity–link function. Fig. 2. Resultados de los remuestreos (validación cruzada sobre cinco conjuntos) de los diferentes modelos evaluados. Inter (valores mínimo y máximo) y valores medio; intervalos de confianza del 95 % para las medianas. Modelos basados en la muestra de los años 2005–2017: * GAM, distribución normal y función de enlace de identidad.

(assessed through RMSE and R2) between models was mostly non–statistically significant. The relative importance of the variables was variable between models. However, standard deviations of eastward and westward wind speed (uwstd and vwstd, respectively) were important variables in most of the models assessed. Year (year) of the observation, chlorophyll concentration (clorof) and bathymetry (batim) were also variables of high importance in predicting shearwater abundance (fig. 4). Discussion This study showed that the combination of Random Forest techniques and massive data provided by various open data sources, including data gathered in citizen science initiatives, is a useful approach to identify the long–term spatial–temporal distribution of species at regional spatial scales. This machine learning technique was highly successful in capturing both spatial and temporal patterns in shearwater abundance at sea. The results from Random Forest

techniques clearly outweighed those obtained with other traditional statistical modelling techniques such as GAM. In observational studies in Ecology and Conservation Biology, where the environmental variability is considerably high and it is not easy to take into consideration all the factors affecting the research subject, the ability of statistical models to predict the existing variability in the dependent variable is usually poor (<10 %), both because of the inherent 'noise' in the data as well the randomness of the natural phenomena that is being modelled (Møller and Jennions, 2002) but randomness and noise may reduce this amount considerably in biological studies. In contrast to traditional statistical techniques, we showed that Random Forest can largely increase the variability explained (up to 53 %) and the predictive ability (70 %) of the models calibrated with noisy data in complex scenarios where both temporal and spatial variation is present. The variance explained by our random forest model is not only larger than usual in Ecology and Conservation Biology studies (about 10 % according to Møller and Jennions, 2002) but it is also larger than the explanatory ability found

Fig. 3. Differences in the performance (in terms of RMSE and R2) of the assessed models. Range and median; 95 % confidence intervals for the means: * GAM, Gaussian distribution and identity–link function. Fig. 3. Diferencias en los resultados obtenidos (en cuanto RMSE y R2) con los modelos evaluados. Intervalo y mediana. Intervalos de confianza del 95 % para las medias: * GAM, distribución normal y función de enlace de identidad.

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Rsquared

RMSE XGBM – RF

XGBM – RF

XGBM – KNN

XGBM – GAM

XGBM – CART

SVM – XGBM

SVM – RF

SVM – KNN

SVM – GBM

SVM – GAM

SVM – CART

RF – KNN

RF – GAM

RF – CART

MLP – XGBM

MLP – SVM

MLP – RF

MLP – KNN

MLP – GBM

MLP – GAM

MLP – CART

GBM – XGBM

GBM – RF

GBM – KNN

GBM – GAM

GBM – CART

GAM – KNN

CART – KNN

CART – GAM

BAGG – XGBM

BAGG –XGBM

BAGG – SVM

BAGG – RF

BAGG – MLP

BAGG KNN

BAGG – GBM

BAGG – GAM

BAGG – CART

BAGG – CART –0.4

–0.2

0.0

0.2

0.4

–0.4

–0.2

0.0

0.2

0.4

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YLAT year XLONG vwstd vwmean uwstd uwmean tmstd tmmean NAO moon julian fish clorof batim 0% GAM

20% KNN

40% SVM

MLP

60% GBM

80%

EXGBM

BAGG

100% CART

Fig. 4. Relative variable importance, as percentage, in the models (from varImp function in caret package; (Kuhn, 2007). Fig. 4. Importancia relativa (en porcentaje) de las variables en los diferentes modelos (a partir de la función varImp en el paquete caret (Kuhn, 2007).

in similar studies (15 %) applying machine learning approaches to traditional survey (less 'noisy') data (i.e, boat–based surveys; Oppel et al., 2012). Machine learning techniques are increasingly applied in order to obtain valid and accurate information from massive data sets that, due to their volume, noise and variety, until not too long ago it was not possible to analyze. However, according to our results, machine learning techniques in general are not a panacea, since other models also assessed in this study, particularly MLP, did not show good results in predicting shearwaters during post–breeding. Our results also highlighted that the predictive ability of the random forest model was highly conditioned on the data subset used to calibrate the model. For this reason, although this technique can deal with datasets with small sample size thanks to the bagging procedure implemented, its application requires suitable data pre–processing to ensure that the data for the analysis is fully representative of the phenomena to be modelled. Conclusions This study shows that the combination of machine learning techniques and massive data provided by

open data sources is a useful approach to identify the long–term spatial–temporal distribution of species at regional spatial scales. Our research showed that machine learning techniques, and specifically random forest approaches, can be a good choice for the analysis of highly noisy massive datasets such as those collected by volunteers in citizen science projects. According to our results, these models allow the successful capture of the spatio–temporal variation in continuous biological variables such as bird migratory abundance recorded over long–time frames, thereby enabling long–term spatial monitoring of mobile species. Acknowledgements This study is part of the research project ('Environmental factors determining the interannual variation in the migration of Balearic and Scopoli’s shearwaters in the Mediterranean') which formed part of the Annual Programme of Grants of the Instituto de Estudios Ceutíes (IEC, Autonomous City of Ceuta, Spain), 2018–2019. Data sourced by eBird can be obtained after registration and request at https:// ebird.org/.

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2004. Marine bird and cetacean associations with bathymetric habitats and shallow–water topographies: implications for trophic transfer and conservation. Journal of Marine Systems, 50: 79–99, Doi: 10.1016/j.jmarsys.2003.09.015 Yésou, P., 2003. Recent changes in the summer distribution of the Balearic shearwater Puffinus mauretanicus off western France. Scientia Marina, 67: 143–148. Zuur, A. F., Ieno, E. N., Walker, N. J., Saveliev, A. A., Smith, G. M., 2009. GLM and GAM for Count Data. In: Mixed effects models and extensions in ecology with R. statistics for Biology and Health: 209–243 (Alain F. Zuur, Elena N. Ieno, Neil J. Walker, Anatoly A. Saveliev, Graham M. Smith, Eds.). Springer, New York. Doi: 10.1007/978-0387-87458-6_9

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A higher incidence of moult–breeding overlap in great tits across time is linked to an increased frequency of second clutches: a possible effect of global warming? I. Solís, J. J. Sanz, L. Imba, E. Álvarez, E. Barba

Solís, I., Sanz, J. J. Imba, L., Álvarez, E., Barba, E., 2021. A higher incidence of moult–breeding overlap in great tits across time is linked to an increased frequency of second clutches: a possible effect of global warming? Animal Biodiversity and Conservation, 44.2: 303–315, Doi: https://doi.org/10.32800/abc.2021.44.0303 Abstract A higher incidence of moult–breeding overlap in great tits across time is linked to an increased frequency of second clutches: a possible effect of global warming? The rise of temperatures due to global warming is related to a lengthening of the breeding season in many bird species. This allows more pairs to attempt two clutches within the breeding season, thus finishing their breeding activity later in the season and therefore potentially overlapping these with post–breeding moult. We tested whether this occurred in two Spanish great tit Parus major populations. The proportion of pairs laying second clutches increased from 1 % to 32 % over the study period in one of the populations (Sagunto, 1995–2019), while it did not change in the other (Quintos, 2006–2019; mean 5 %). We did not find any temporal trend for moult start date of late–breeding birds in any population. The proportion of individuals of both sexes that overlapped moult and breeding increased in Sagunto. For this latter population, sex and age, but not clutch type, contributed to the variability in the probability of overlapping in late–breeding individuals, this being higher for first–year males and lower for older females. Key words: Post–breeding moult, Phenological changes, Climate change, Parus major, Great tit, Spain Resumen El incremento del solapamiento entre la muda y la reproducción en el carbonero común está ligado a un aumento en la frecuencia de segundas puestas: ¿Un posible efecto del calentamiento global? El ascenso de las temperaturas debido al cambio climático está relacionado con un aumento de la duración de la temporada reproductiva de muchas especies de aves. Esto permite que más parejas intenten poner dos puestas durante la temporada reproductiva y conlleva que terminen sus actividades reproductivas más tarde; por tanto, estas actividades se podrían solapar con la muda postnupcial. Hemos comprobado si esto ocurre en dos poblaciones de carbonero común (Parus major) de España. La proporción de parejas con segundas puestas se ha incrementado del 1 % al 32 % durante el periodo de estudio en una de las poblaciones (Sagunto, 1995–2019), mientras que en la otra no ha cambiado (Quintos, 2006–2019; media 5 %). No hemos encontrado ninguna tendencia temporal en cuanto a la fecha de inicio de muda de los individuos que están criando en fechas tardías en ninguna de las dos poblaciones. La proporción de individuos de ambos sexos cuya muda y actividad reproductiva se solaparon se ha incrementado en Sagunto. En esta última población, el sexo y la edad, pero no el tipo de puesta, contribuyeron a explicar la variabilidad en la probabilidad de solapamiento entre los reproductores tardíos, ya que esta es mayor en los machos de primer año y menor en hembras adultas. Palabras clave: Muda post–nupcial, Cambios fenológicos, Cambio climático, Parus major, Carbonero común, España Received: 09 III 21; Conditional acceptance: 16 IV 21; Final acceptance: 31 VIII 21 I. Solís, L. Imba, E. Álvarez, E. Barba, Cavanilles Institute of Biodiversity and Evolutionary Biology, University of Valencia, c/ Catedrático José Beltrán 2, 46980 Paterna, Spain.– J. J. Sanz, National Museum of Natural Sciences–CSIC, c/ José Gutiérrez Abascal 2, 28006 Madrid, Spain. Corresponding author: I. Solís. E–mail: irisolis.hz@gmail.com ORCID ID: I. Solís: 0000-0001-7557-1463; J. J. Sanz: 0000-0003-2576-4050; E. Álvarez: 0000-0001-8256-443X; E. Barba: 0000-0003-2882-9788 ISSN: 1578–665 X eISSN: 2014–928 X

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Introduction The rise in temperatures due to global warming has affected ecosystems in various ways in recent decades (Walther et al., 2002), and is predicted to have dramatic effects in the near future (Trisos et al., 2020). One of the most evident effects is the alteration of phenological events of plants and animals (Menzel and Fabian, 1999; Parmesan and Yohe, 2003). In birds, for example, the timing of reproduction (Both et al., 2004; Källander et al., 2017), moult (Morrison et al., 2015) and, for migratory species, migration (Charmantier and Gienapp, 2014; Tomotani et al., 2017) are all being affected by global warming. Reproduction and moult both require the investment of substantial amounts of energy (Wilkinson, 1983). For reproduction, birds need to look for a mate, defend their territory, construct a suitable nest, form the eggs, incubate them and, lastly, take care of the chicks. As an example, the field metabolic rate of a sample of birds during incubation was 2.93 times higher than the basic metabolic rate (BMR), while that during the chick–rearing period was 3.38 times higher (Nord and Williams, 2015). On the other hand, for moult, birds must replace (partially or completely) their feathers to adequately maintain the main functions of the plumage, such as thermal insulation, flight, and appearance (Payne, 1972; Jenni and Winkler, 2020). Small birds, such as passerines, have the highest cost of moult per body mass (Hoye and Buttermer, 2011), and BMR could be up to 2.11 times higher than that during the moulting period (Lindström et al., 1993). In small passerines, moult and breeding compete for time, energy, and performance, so birds usually avoid overlapping these activities (Payne, 1972; Wilkinson, 1983; Hemborg and Lundberg, 1998; Jenni and Winkler, 2020). Most European passerines have annual cycles with regular, distinct, and relatively short periods for breeding and moult (Jenni and Winkler, 2020). Small passerines undergo a sequential moult at least once a year, and this usually occurs just after breeding (usually known as post–breeding moult), when food resources are still abundant (Hemborg et al., 2001; Moreno, 2004; Morrison et al., 2015; Tomotani et al., 2017; Jenni and Winkler, 2020). Notwithstanding the, some overlap between the end of breeding (while feeding the young or even during incubation) and the onset of moult has been repeatedly documented, both in migratory (pied flycatchers Ficedula hypoleuca; Morales et al., 2007) and resident species (great tits Parus major; Svensson, 1995), so a physiological trade–off could be expected in these overlapping birds (Svensson and Nilsson, 1997; Sanz et al., 2004). The consequences of this overlap include a reduction in breeding success, survival rate, and/or the quality of the feathers (Hemborg and Lundberg, 1998; Morales et al., 2007; Tomotani et al., 2017). However, some positive effects have been also reported, for example in body condition and stress levels, that might result in better survival probabilities for these birds (Morales et al., 2007). Local environmental conditions might affect the degree of overlapping of breeding and moult in particular populations within a species. However, direct

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comparisons of moult–breeding overlap between populations are scarce (Hemborg et al., 1998, 2001). Explanations for the between-population differences rely on latitudinal variation inthe length of the breeding season. Thus, moult–breeding overlap is more common in northern populations, where the breeding season is short and many birds have to start moulting while still breeding, but it is also significant in southern Europe, probably because birds are time constrained by the hot, dry summer (Sanz, 1999; Hemborg et al., 2001). Similar situations might be expected when comparing populations living at different altitudes, as may occur in countries like Spain where altitudinal gradients are large. Halupka and Halupka (2017) have shown that an average bird population inthe northern hemisphere has increased its breeding season by 1.4 days per decade in the last 45 years, and this lengthening is independent from changes in mean laying dates. More specifically, the breeding season in multi-brooded species has increased by 4 days per decade, while it has decreased by two days per decade in single-brooded species. The lengthening of the breeding season in potentially multi–brooding species allows more individuals to lay two clutches within the same breeding season (Monroe et al., 2008; Townsend et al., 2013). This, in turn, increases the number of pairs involved in breeding activities late in the season. Overlapping is more frequent in late–breeding birds (Svensson and Nilsson, 1997; Hemborg et al., 2001), even though birds breeding late in the season start moulting later than those finishing their breeding activities earlier (Dhondt, 1973; Orell and Ojanen, 1980; Jenni and Winkler, 2020). Therefore, if more individuals are breeding late in the season and moulting dates do not change, the number of individuals overlapping moult and reproduction would increase (Moreno, 2004; Tomotani et al., 2017). At the within–population level, not all birds have the same probability of overlapping. Males, for example, usually start moulting earlier than females (Orell and Ojanen, 1980; Tiainen, 1981; Ojanen and Orell, 1982) and have consequently been found to overlap breeding and moult more frequently (Orell and Ojanen, 1980; Hemborg, 1999; Moreno et al., 2001; Jenni and Winkler, 2020). On the other hand, the two activities often overlap more frequently in first–year birds than in older birds, maybe because the poorer condition of the primaries in younger birds inducing earlier moult or because younger birds start breeding later than older birds (Siikamäki et al., 1994; Hemborg et al., 2001). Finally, in potentially double–brooding populations, replacement clutches (those laid after a failed first one) are generally laid earlier than true second clutches (i.e. those laid after a successful one), so overlapping would probably occur more frequently among those individuals involved in second clutches. In their analysis, Haupka and Halupka (2017) considered great tits as multi–brooded species. However, the variability in the proportion of pairs laying second clutches (those laid after a successful first one) is high between populations, and between years within populations. For example, means from several studies

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compiled by Cramp and Perrins (1993) ranged from 2.6 % in England to 54 % in Ukraine, and reported a range of 2–93 % of second clutches in different years in the Netherlands. Senécal et al. (2021) even showed significant variation in the proportion of second clutches between nearby plots within a general study area (from less than 5 % in densely populated plots to about 30 % in low–density plots). Therefore, depending on their propensity to lay second clutches, different populations might respond differently to lengthening of the breeding season. We selected two Spanish great tit populations located in areas with different climate conditions and a different incidence of second clutches. The first population) was in Sagunto, on the Mediterranean coast, just above sea level, and 36 % of the pairs laid two clutches (see Results). The second population was in Quintos, in Central Spain, at about 900 m a.s.l., and only 7 % of the pairs laid two clutches (see Results). Our main objectives were (1) to check whether the proportion of individuals with overlapping postbreeding moult and reproduction has increased over the last 25 years, and (2) to identify individual traits related to the probability ofthis increase. We aimed to answer the following specific questions for each population: (1) is there a temporal trend in the proportion of pairs attempting replacement or second clutches? (2) is there atemporal trend in the date of moulting of late–breeding pairs? (3) what proportion of the breeding population shows overlaps in breeding and moult? (4) is there a temporal trend in the proportion of individuals showing overlapping of breeding and moult? (5) which individual traits (sex, age) are related to the probability of overlapping? (6) does the probability of overlapping differ between birds attempting replacement clutches and those involved in second clutchew? To classify species as single- or multi-brooded, Halupka and Halupka (2017) considered multi-brooded as those species where, at least in some populations, more than 30 % of the individuals laid more than one clutch. If we apply this criterium to the population level, our population of Sagunto would fit within the multi–brooded category, while that from Quintos would fit the single–brooded category. Following the results of Halupka and Halupka (2017), the breeding season would be expected to increase in Sagunto, with more pairs being able to lay two clutches. Milder climate conditions by the coast might allow birds to delay moulting to avoid overlap. In contrast, the breeding season in Quintos would be expected to contract, causing a decrese (or at least not promoting an increase) of second clutches (as occursin central European populations; Visser et al., 2003; Husby et al., 2009). If the favourable season were shorter reduced, moult may tend to occur earlier, potentially increasing the incidence of a moult–breeding overlap. Concerning individual traits, our predictions were that (1) males would start moulting before females, and the overlap would thus be more frequent; and (2) first–year individuals would overlap more frequently. Finally, we would expect that (3) birds involved in second clutches would overlap more frequently than those raising replacement clutches.

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Methods The study was conducted in two wild great tit populations breeding in nestboxes in Spain One was located within an extensive orange Citrus aurantium monoculture near Sagunto (Valencia, eastern Spain; 39º 42' N, 0º 15' W, 30 m a.s.l.), and data relevant for the present study are available between 1995 and 2019 (Álvarez and Barba, 2014; Rodríguez et al., 2016). The study area in Sagunto has changed over the years, from about 150 ha in the first years to about 450 ha in recent years, but the density of nestboxes has remained almost constant at about 1 nestbox per ha. The other population was located in deciduous forest patches at the Reserve of Quintos de Mora (Toledo, central Spain; 39º 32' 44'' N, 4º19' 41'' W, 800–1235 m a.s.l.), where relevant data are available from 2006 to 2019 (Bueno–Enciso et al., 2017). The study area in Quintos is about 50 ha, and the density of nestboxes was about 4 nestboxes per ha. Mean surface temperatures in Spain have increased at a rate of 0.3 ºC per decade since 1960 (Vicente–Serrano and Rodríguez, 2017; see also Pérez et al., 2015 for a more detailed study of the eastern part of the country). Six of the ten hottest years from 1965 to 2019 are characterised by a Mediterranean climate, though the weather in Sagunto, on the Mediterranean coast, is milder than that in Quintos (Chazarra et al., 2018). Each nestbox was inspected at least weekly, and daily in some periods, from mid–March until the end of the breeding season in mid–July, and basic breeding parameters (date of laying of the first egg, clutch size, number of hatchlings and fledglings) wererecorded. Dates are presented as 'April date' (1 April = day 1). Adults were captured with door–traps at the nestboxes when feeding 10–12 day–old nestlings and fitted with individually numbered metal rings. The sex and age (first–year or older) of each individual was noted (Svensson, 1992). We also we recorded whether the bird was moulting its primary feathers when trapped (Tomotani et al., 2017). Adult great tits have a complete post–breeding moult and, while breeding, rarely moult feathers other than their primaries (Orell and Ojanen, 1980). In our case, the most advanced birds were moulting their fourth primary. For each year, we considered that moult started the first day we caught the first adult moulting at least one of its primary wing feathers. All birds found to be moulting in Sagunto, and most in Quintos, were caught when raising their replacement or second brood. As the occurrence of moult was assessed when capturing the birds while feeding nestlings, we were unable to determine the date of start of moulting of birds laying only one clutch (with the exception of 9 individuals that had late first clutches in Quintos). Since birds laying only one clutch could have started moulting earlier (Dhondt, 1973; Orell and Ojanen, 1980), the date of start of moult should be virtually valid for the subset of the population laying two clutches within the season. Each year, the breeding population at each study area was estimated as the number of first clutches laid. We considered as first clutches those started

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within the first 30 days after the start of the first clutch of that year (Van Noordwijk et al., 1995; Álvarez and Barba, 2014). The rest of the clutches of the season were allocated into three types: (1) replacement clutches (laid after failure of a first clutch); (2) second clutches (laid after successful fledging of a first clutch); and (3) unknown clutches. Replacement and second clutches were identified either because at least one of the parents was ringed, or by spatial proximity and phenological concordance (Borgman and Wolf, 2016). From data on ringed birds, we know that pair bond is virtually stable in great tits in both study populations throughout the season. We therefore know that most birds did not change their partners between the first and second clutches if both were alive, and that second or replacement clutches are laid in nearby nestboxes,or, on many occasions, in the same nestbox (pers. obs.). As an example of within season mate and site fidelity, in 2019, from 18 pairs where both adults were identified in two consecutive breeding attempts, 16 pairs (89 %) maintained their composition (same male and female), while in the other two pairs the female bred with a different male. The two missing males were not observed to be breeding again with other females this season, or in subsequent seasons (2020 and 2021). Concerning site fidelity, 11 of these pairs laid their second clutch in the same nestbox, 6 did so in a contiguous nestbox (about 50 m distance), and 1 about 200 m away from the first box. This within–season mate and site fidelity is well–known in great tits, and has been quantified even between seasons in one of the study populations (Andreu and Barba, 2006). We were as careful as possible with these criteria, so some breeding attempts each year could not be clearly attributed to a specific pair. Thus, 'unknown' clutches are those which were laid late in the season (after the temporal limit to be considered true first clutches) but could not be clearly classified as replacement or second clutches of any particular pair. Although we did not check for temporal trends in these unknown clutches, the individuals involved were used for analyses when appropriate (see below). We are aware that despite taking maximum care in identifying clutch types, we may have misclassified some. However, we believe such possibilities would be relatively few, and would not bias the results. To study the proportion of individuals in each population for whom breeding and moulting stages overlapped, we identified the laying date of the clutch of the first individual we caught moulting each year. We considered that all the pairs that started laying (replacement, second or unknown clutches, or even first clutches in the case of Quintos) after this date could potentially overlap breeding and moult that year. From these potentially overlapping pairs, we caught only a fraction of individuals, mostly because nests failed before the date of trapping, but also because on some occasions parents avoided the traps. There may also have been logistic reasons. Thus, to estimate the number of individuals probably overlapping each year, we extrapolated the proportion of individuals which were actually moulting from those trapped,

to the number of individuals which could potentially overlap. Thus, for each year, we computed: (1) the total number of breeding pairs (i.e. the number of first clutches); (2) the total number of individuals that could potentially overlap breeding and moult (those belonging to pairs starting a clutch later than the first individual found moulting), and (3) the proportion of individuals, from the total breeding population, in which breeding and moulting probably overlapped. We collected data from mean daily temperatures from March and April in each of the 34 study years in Sagunto and 14 in Quintos from the meteorological stations of 'Sagunto–Pontazgo', 4 km from the study area of Sagunto, and Ciudad Real (Instituto), 45 km from the study area of Quintos. From these, we calculated monthly temperatures. To explore the temporal variation in the proportion of second and replacement clutches, we used Generalized Linear Mixed Models (GLMM) fitted with a binomial distribution and logit function (Zuur et al., 2009), in which the response variable was a data frame containing two columns: the annual number of pairs with and without second or replacement clutches. These models included the year as an explanatory variable. The study year as a categorical variable was also included as a random effect. As available years differed, we performed separate analyses for each population. The small number of birds overlapping breeding and moult in Quintos precluded the analyses relating individual characteristics and overlapping probability in this population. For the population of Sagunto, we performed exploratory analyses to determine the potential relationship between individual characteristics and clutch type and the occurrence of overlapping. As above, we used Generalized Linear Mixed Models (GLMM) fitted with a binomial distribution and logit function to test for the effect of sex (fixed factor) and year (covariate) on the proportion of individuals that overlapped moult and breeding each year. GLMM analyses fitted with a binomial distribution (moulting vs. no moulting) and logit function were also carried out to evaluate the relative contribution of the relevant variables (sex, age, and clutch type) as predictors of the probability of overlapping breeding and moult. In these GLMM, study year as categorical variable was also included as a random effect. The models were run in r (R Core Team, 2014) using the lme4 packages (Bates et al., 2015) to calculate fixed term estimates and the car package (Fox and Weisberg, 2011) to calculate p– values from the analyses of deviance of the models on the basis of Wald x2 tests. Results Breeding attempts per pair per year In Sagunto, breeding data were recorded for 2701 pairs between 1995 and 2019. Of these, 51% laid only one clutch, 16 % made a replacement clutch after a failed first clutch, and 20 % laid a second clutch after a successful first clutch. Thirteen percent of the

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Table 1. Number of breeding pairs (NBP) and proportion of pairs laying only one clutch (POOC), a first and a replacement clutch (FRC), or a first and a second clutch (FSC), from 1995 to 2019 in Sagunto. Unknown clutches (UC) are those which could not be attributed to any of the previous groups (probably late first, or early replacement). Tabla 1. Número de parejas reproductoras (NBP) y proporción de parejas que solo hacen una puesta POOC), parejas que hacen una primera puesta y una reposición (FRC) y parejas que hacen una primera puesta y una segunda puesta (FSC), entre 1995 y 2019 en la población de Sagunto. Las puestas desconocidas (UC) son aquellas que no pudieron atribuirse a ninguno de los grupos anteriores (probablemente primeras puestas tardías o reposiciones tempranas).

Year

NBP

POOC (%)

1995 91

FRC (%)

FSC (%)

UC (%)

67.03 14.29 6.59 12.09

1996 113 76.99 22.12 0.88 0.00 1997 106 67.92 21.70 6.60 3.77 1998 97

55.67 22.68 8.25 13.40

1999 115 66.96 19.13 0.87 13.04 2000 122 72.95 11.48 4.10 11.48 2001 98

50.00 28.57 11.22 10.20

2002 98

11.22 62.24 6.12 20.41

2003 113

53.98 6.19 20.35 19.47

2004 40

47.50 0.00 17.50 35.00

2005 25

68.00 0.00 16.00 16.00

2006 37

62.16 8.11 10.81 18.92

2007 49

30.61 34.69 18.37 16.33

2008 81 77.78 3.70 3.70 14.81 2009 75

36.00 5.33 24.00 34.67

2010 117

36.75 11.11 31.62 20.51

2011 134 61.19 8.21 20.90 9.70 2012 133

43.61 12.78 32.33 11.28

2013 134

35.82 29.10 27.61 7.46

2014 142

48.59 11.97 22.54 16.90

2015 146

58.22 15.07 21.23 5.48

2016 138

36.96 10.87 32.61 19.57

2017 163

35.58 15.95 38.65 9.82

2018 164

48.17 10.98 27.44 13.41

2019 170 50.59 4.71 35.29 9.41 Total 2701 51.24 15.85 19.77 13.14

breeding pairs laid a 'second' clutch, but we could not determine whether it was a replacement clutch or a second clutch, and they were classified as 'unknown' (table 1). In this study area, there was no temporal trend in the percentage of replacement clutches (R2 = 8.88 %; Estimate = –0.035 ± 0.024; Z = 1.48; P = 0.14). In contrast, the proportion of pairs laying a second clutch increased across the study years (R2 = 76.9 %; Estimate = 0.1108 ± 0.015; Z = 7.55; P < 0.001) from 1 % to 32 % (fig. 1).

Data for Quintos were recorded from 505 pairs from 2006 to 2019. From these, 92 % laid only one clutch, 2 % laid a replacement clutch, and 5 % laid a second clutch, with the remaining 1 % being unknown (table 2). Neither the percentage of replacement clutches (R2 =12.16 %; Estimate = 0.108 ± 0.894; Z = 0.89; P = 0.37) nor the percentage of second clutches (R2 = 1.28 %; Estimate = 0.038 ± 0.102; Z = 0.37; P = 0.70; fig. 1) showed a significant temporal trend.

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Second clutches (%)

40 35

Sagunto Quintos

30 25 20 15 10 5 0 1994

1999

2004 2009 Study year

2014

2019

Fig. 1. Proportion of pairs that laid a second clutch after a successful first attempt between 1995 and 2019 in Sagunto (continuous line result from GLMM) and Quintos (dotted line result from GLMM). Fig. 1. Proporción de parejas que hicieron una segunda puesta después de una primera puesta exitosa entre 1995 y 2019 en la población de Sagunto (línea continua como resultado del modelo lineal generalizado mixto del GLMM) y la de Quintos (línea discontinua como resultado del GLMM).

Temporal trends in moult–breeding overlap The date at which we captured the first moulting bird varied between June 2 and June 28 in Sagunto, and between May 15 and June 24 in Quintos. We did not find a significant temporal trend over the years in either population (Sagunto: R2 = 10.7 %; F1,21 = 2,39; P = 0.14; Quintos: R2 = 8.8 %; F1,6 = 1.482; P = 0.518). Considering the whole population each year, the proportion of birds estimated to overlap breeding and moulting activities in Sagunto varied between 0 % (in 1996, 2000 and 2004) and 28 % (in 2002), and increased over the years for both sexes (table 3, fig. 2). Globally, more males (18 %) than females (7 %) overlapped moult and reproduction, and this difference remained across the years. In Quintos, however, there were no differences between sexes in the proportion of individuals overlapping, and no significant temporal trend in the proportion of individuals overlapping (table 4). The interaction between sex and year was not significant in either study (P > 0.05). Characteristics of birds overlapping breeding and moult The number of birds found overlapping in Quintos was too small for meaningful statistical analyses, so only data from Sagunto were included in this section. Of the data included, we considered only those birds caught after the capture of the first moulting bird each year, and we refer to them here as 'late–breeding' birds. In most years, the first individual observed to be moulting was a male (x21 = 7.68, P = 0.01). Consequently, many more late–breeding males (67 %)

than females (27 %) overlapped their moulting and breeding activities (table 5; fig. 3), with this occurring more frequently in first–year breeding birds than in older birds (table 5; fig. 3). However, the type of clutch (replacement vs. second clutch) did not affect the probability of overlapping moulting while still feeding young (table 5; fig. 3). Discussion Proportion of second clutches, timing of moulting and overlap The general picture gives the impression that European passerines are generally advancing their breeding season in response to global warming, which would supposedly lengthen the breeding season, which would promote more pairs to attempt a second brood. However, this chain of events does not always occur. Visser et al. (2003), for example, examined the effect of raising ambient temperatures in 13 European great tit populations and found that only five of them had advanced their laying dates. Second, an advance of the onset of laying does not necessarily entail a longer breeding season (Møller et al., 2010; Gullett et al., 2013). Moreover, the breeding season could be lengthened independently of changes in mean laying dates (Halupka and Halupka, 2017). Finally, the three studies exploring temporal trends in the proportion of great tits presenting double brooding in several European populations showed a decline over the last years (Visser et al., 2003; Husby et al., 2009; Matthysen et al., 2011).

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Table 2. Number of breeding pairs (NBP) and proportion of pairs laying only one clutch (POOC), a first and a replacement clutch (FRC), or a first and a second clutch (FSC), from 2006 to 2019 in Quintos. Unknown clutches (UC) are those which could not be attributed to any of the previous groups (probably late first, or early replacement clutches). Tabla 2. Número de parejas reproductoras (NBP) y proporción de parejas que solo hacen una puesta POOC), parejas que hacen una primera puesta y una reposición (PRC) y parejas que hacen una primera puesta y una segunda puesta (FSC), entre 2006 y 2019 en la población de Quintos. Las puestas desconocidas (UC) son aquellas que no pudieron atribuirse a ninguno de los grupos anteriores (probablemente primeras puestas tardías o reposiciones tempranas).

Year

NBP

POOC (%)

FRC (%)

FSC (%)

UC (%)

2006 7 100.00 0.00 0.00 0.00 2007 33 81.82 0.00 9.09 9.09 2008 40

82.50 0.00 17.50 0.00

2009 58 94.83 1.72 3.45 0.00 2010 32 96.88 3.13 0.00 0.00 2011 51 96.08 1.96 1.96 0.00 2012 72 95.83 2.78 1.39 0.00 2013 48 100.00 0.00 0.00 0.00 2014 27 96.30 3.70 0.00 0.00 2015 29 93.10 0.00 3.45 3.45 2016 25

84.00 4.00 12.00 0.00

2017 25 96.00 4.00 0.00 0.00 2018 27 96.30 0.00 3.70 0.00 2019 31

74.19 0.00 25.81 0.00

Total 505 92.28 1.58 5.35 0.79

In absolute contrast with the above studies on great tits, however, the proportion of pairs laying a second clutch in Sagunto increased from 1 % to 32 % over a 25–year period. Although an increase of second clutches has been reported for other species (Monroe et al., 2008; Townsend et al., 2013), this is, to our knowledge, the first time that it has been found for great tits. The Quintos population showed an intermediate behaviour, with no significant trend in the proportion of pairs laying second clutches over the years. The frequency of second clutches is generally low in great tits (Cramp and Perrins, 1993) and is likely to change with environmental conditions (Husby et al., 2009; Reed et al., 2013; Senécal et al., 2021). To explain the reduction of second clutches in several European populations, Visser et al. (2003) suggested that, as caterpillar development accelerates with increasing ambient temperatures, food for the nestlings is scarcer late in the season in warm years, so the reproductive value of second clutches decreases. At least some Mediterranean great tit populations do not depend as much on a single caterpillar peak to feed their

Table 3: Results of the GLMM fitted with a binomial distribution testing for a temporal trend in the proportion of males and females overlapping at Sagunto throughout the study years (R2 = 53.25 %). Tabla 3. Resultados del modelo lineal generalizado mixto (GLMM) con distribución binomial para comprobar si existe alguna tendencia temporal en la proporción de machos y hembras con solapamiento en la población de Sagunto durante los años de estudio (R2 = 53,25 %).

Source

Estimate ± SE

Intercept –4.57024 ± 0.495 9.23 < 0.001 Sex

-0.569 ± 0.046

12.31 < 0.001

Year

0.134 ± 0.031

4.27 < 0.001

Solís et al.

310

Overlapping (%)

45 40

Males

Females

30 25 20 15 10 5 0 1994

1999

2004 2009 Study year

2014

2019

Fig. 2. Variation in the proportion of males (continuous line result from GLZ) and females (dotted line result from GLMM) overlapping moulting and breeding activities across the study years in Sagunto. Fig. 2. Variación en la proporción de machos (línea continua como resultado del modelo lineal generalizado, GLZ)) y hembras (línea discontinua como resultado del modelo lineal generalizado mixto, GLMM) cuya muda y actividad reproductiva se solaparon durante los años de estudio en la población de Sagunto.

nestlings (Barba and Gil–Delgado, 1990; Blondel et al., 1991; Pagani–Núñez et al., 2011), so the 'food' constraint late in the season might not be such a problem for late–breeding. Our population from Sagunto is known to feed its nestlings mainly with moths, using caterpillars only for the very early clutches (Barba and Gil–Delgado, 1990; Barba et al., 1996). Feeding on moth species avoids the dependence on a single food peak and probably provides acceptable breeding conditions late in the season (Barba et al., 1994, 2004). The scenario in Quintos is probably more similar to other populations in central and northern European, with parents depending more on a relatively short caterpillar peak to feed their nestlings (80 % of the nestling diet; García–Navas et al., 2013). The role that food plays in the case of the population in Sagunto would be worth studying in detail, since available data are relatively old and caterpillar and moth phenology might well have changed. Adult great tits perform a complete moult once a year, starting between early May and late June in different European populations (Flegg and Cox, 1969; Orell and Ojanen, 1980; Dhondt, 1981). Although it generally starts after the nestlings have fledged, some authors have reported moult initiation while rising second broods, or even when still feeding first–brood nestlings (especially in males) (Flegg and Cox, 1969; Orell and Ojanen, 1980; Dhondt, 1981). Dates found in Quintos and Sagunto seem to be at the end of this time window, but it should be noted that birds laying only one clutch were

not caught thereafter, so they might have started moulting before double–brooding birds (Orell and Ojanen, 1980; Dhondt, 1973) but gone undetected. In Sagunto, no individual was captured moulting while feeding first–brood nestlings, but some individuals were found moulting when raising their first brood in Quintos. Some studies have reported an advance of the timing of moult, related to global warming, in different species (Helm et al., 2019; Kiat et al., 2019; Nadal et al., 2021). To the best of our knowledge, this advancement has not been reported for European great tit populations. In our study areas, we did not find any consistent trend in the onset of moulting over the years, though we should keep in mind that we only have data on late–breeding birds. Great tit populations in general, and the two populations studied here in particular, are resident, so migration pressure is not a potential factor governing moulting dates. A potential strategy for those pairs attempting two clutches would be to delay the moulting dates, thus avoiding or minimizing moult–breeding overlap. This should have been especially noted in our Sagunto population, where many more pairs are now laying second clutches. However, the dry, hot summers of the Mediterranean region could be an important limiting factor for developing energetically demanding activities by these dates (Dhondt, 1981; Hemborg et al., 2001). For example, heat stress might reduce the immune response and interfere with the moult–immunity trade–off by constraining seasonal delays in moulting (Moreno et al., 2001; Moreno, 2004). Thus, in Sagunto, the continuous

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Table 4. Results of the GLMM fitted with a binomial distribution testing for a temporal trend in the proportion of males and females overlapping at Quintos throughout the study years (R2 = 86.13 %). Tabla 4. Resultados del modelo lineal generalizado mixto (GLMM) con distribución binomial para comprobar si existe alguna tendencia temporal en la proporción de machos y hembras con solapamiento en la población de Quintos durante los años de estudio (R2 = 86,13 %).

Source

Estimate ± SE

Intercept

–4.051 ± 0.882

4.60 < 0.001

Sex

–0.204 ± 0.458

0.45

0.66

Year

–0.187 ± 0.145

1.29

0.20

311

Table 5. Results of the GLMM fitted with overlap status (moulting vs. non–moulting) as the dependant variable (binomial distribution) and year, sex, age (first year vs. older birds), and clutch type (replacement or second clutch) from birds at Sagunto (R2 = 25.41 %). Study year as categorical variable was also included as a random effect. Tabla 5. Resultados del modelo lineal generalizado mixto (GLMM) con el solapamiento (muda o no muda) como variable dependiente (distribución binomial) y el año, el sexo, la edad (primer año o mayor) y el tipo de puesta (reposición o segunda) como variables independientes en las aves de la población de Sagunto (R2 = 25,41 %). El año de estudio fue también incluido como un efecto aleatorio en forma de variable categórica.

Source

Estimate ± SE

Intercept

–1.760 ± 0.500

Clutch type –0.097 ± 0.101 increase of the proportion of pairs laying two clutches, and then finishing their breeding activity later in the season, and the maintenance of the moulting dates, made the increase of individuals overlapping moult and reproduction over time inevitable. We have shown that the proportion of individuals that present breeding overlap and moult has increased over the years, from 1 % to 23 % for males and from 0 % to 14 % for females (fig. 2). The continuous increase of individuals following this strategy suggests that the benefits of raising two clutches per year override the potential costs of overlapping in the population of Sagunto. Characteristics of bird which show overlapping This section concerns only the great tit population of Sagunto, where sample size was sufficiently large to perform adequate statistical analyses. In agreement with previous studies on several passerine species (Flegg and Cox, 1969; Orell and Ojanen, 1980; Tiainen, 1981; Ojanen and Orell, 1982; Hemborg and Merilä, 1998; Hemborg and Lundberg, 1998; Hemborg, 1999; Hemborg et al., 2001; Jenni and Winkler, 2020), we observed that males started moulting earlier, and overlapping was more common in males than in females. Some authors have suggested that males 'need' to start moulting before females, either because they take longer time than females to complete the moult (e.g. Hemborg, 1999), or because they need to finish moulting earlier to be able to allocate energy to territory defence in autumn (Dhondt, 1973). The relative cost of overlap is another potential reason for sexual differences. For example, Hahn et al. (1992) stated that males invest less energy in reproduction than females (Westneat and Sherman, 1993; Queller, 1997), so they can afford

3.52 < 0.001 0.96

0.34

Sex

–0.968 ± 0.083 11.62 < 0.001

Age

0.221 ± 0.086

2.57

0.010

Year

0.086 ± 0.028

3.10

0.002

to allocate energy in breeding and moulting activities simultaneously. Hemborg (1998) also suggested that females may have larger fitness costs than males from a moult–breeding overlap. Finally, females might be more constrained than males to overlap for physiological reasons.In this sense, Miller (1961) and King (1973) suggested that gonadal activity delays moult in females especially. Siikamäki et al. (1994) and Hemborg et al. (2001) found that first–year pied flycatchers Ficedula hypoleuca from several European populations overlapped more often than adults. In agreement with these age differences in flycatchers, first–year males and females in Sagunto overlapped more frequently than older birds. Siikämaki et al. (1994) suggested that plumage condition might be inferior in first–year individuals, and they could thus need to start moulting earlier than older birds, and that as first–year individuals start breeding later, overlapping is more probable. Although Siikämaki et al. (1994) themselves considered these explanations unlikely, Morales et al. (2007) found that early moult increases survival chances, so perhaps young birds, with more survival prospects, could priorize moulting over reproductive investment. As with the differences between sexes, asymmetries in the relative costs, hormonal differences, or even lack of experience, might be behind these 'age effects', and more research is

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Individuals overlapping (%)

100 % 90 %

Non–moulting Moulting

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

First year Older First year Older First year Older First year Older (n = 56) (n = 65) (n = 29) (n = 74) (n = 138) (n =211) (n = 75) (n = 209) Female Male Female Male Replacement clutch Second clutch

Fig. 3. Proportion of individuals overlapping moult and breeding caught during the breeding season depending on their sex (male vs, female), age (first year vs. older) and clutch type (replacement vs second clutch) over the study years in Sagunto. Fig. 3. Proporción de individuos cuya muda y actividad reproductiva se solaparon y que fueron capturados durante la temporada reproductiva dependiendo de su sexo (macho o hembra), edad (primer año o más de un año) y tipo de puesta (puesta de reposición o segunda puesta) durante los años de estudio en Sagunto.

clearly needed. It should also be noted that only two species, and a handful of populations, support this conclusion, and differences between age classes are not always found (Sanz, 1999; Morales et al., 2007). Replacement clutches can be laid at any moment after the failure of a first clutch, while second clutches are laid after a first brood flies off. Second clutches therefore usually start later than replacement clutches. Moreover, birds laying second clutches have made a greater previous effort than those laying replacement clutches, since the former have raised their first brood up until independence, while the effort of the latter might have been ended at any moment between egg laying and fledging. Nevertheless, we found no differences between birds in their propensity to overlap moult and breeding according to the type of their second breeding attempt. Thus, other factors, such as date, the current physiological state of the birds, and/ or their sex and age, might govern the propensity to overlap moulting and breeding activities. We are not aware of any previous study dealing with this aspect of the moult–breeding overlap issue. Summarizing, we observed that sex and age, but not clutch type, contributed to explain the variability in overlapping probability between individuals in our populations. This probability was higher for males and for first–year individuals. These results are consistent with

our initial hypotheses regarding sexual and age–related differences in overlapping breeding and moult. With the current global warming scenario, we can expect an increase in the proportion of pairs attempting second clutches, and therefore an increasing proportion of individuals (even females) overlapping moulting and breeding in the near future in the population of Sagunto, contrasting sharply with other European populations, including thosee of Quintos studied here. The challenge now is to test whether this increase in the proportion of individuals overlapping two energetically demanding activities, such as breeding and moult, has consequences on the breeding performance and survival of adults and their nestlings. Acknowledgements We wish to thank all the people who collaborated with the fieldwork over these years, and the Spanish Meteorological Agency (AEMET) for providing temperature records from our study sites. This study was partially supported by the projects CGL2013–48001– C2–1–P (Spanish Ministry of Science and Innovation) and CGL2016–79568–C3–1–P (Spanish Ministry of Economy and Competitiveness) thanks to the European Social Fund.

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Description of a new valvatoid Pikasia smenensis n. gen. n. sp. (Gastropoda, Hydrobiidae) from Morocco A. F. Taybi, P. Glöer, Y. Mabrouki

Taybi, A. F., Glöer, P., Mabrouki, Y., 2021. Description of a new valvatoid Pikasia smenensis n. gen. n. sp. (Gastropoda, Hydrobiidae) from Morocco. Animal Biodiversity and Conservation, 44.2: 317–320, Doi: https:// doi.org/10.32800/abc.2021.44.0317 ZooBank LSID: http://zoobank.org/urn:lsid:zoobank.org:pub:AB457940-1B1A-47CD-8DEA-1C417CDD5D4E Abstract Description of a new valvatoid Pikasia smenensis n. gen. n. sp. (Gastropoda, Hydrobiidae) from Morocco. Recent field surveys conducted in the northern part of Morocco have led to the discovery of a new species belonging to a new genus Pikasia n. gen. described here. Photos of the holotype and paratype are presented in addition to the penis morphology and the female sex tract, the map of the sampling area with the type localities, and the habitat description. Key words: Morocco, Hotspot, Springsnail, Pikasia smenensis n. gen. n. sp. Resumen Descripción de un nuevo valvátido Pikasia smenensis n. gen. n. sp. (Gastropoda, Hydrobiidae) en Marruecos. Los recientes estudios de campo llevados a cabo en el norte de Marruecos han permitido descubrir una nueva especie perteneciente a un nuevo género, Pikasia n. gen., que se describe en este artículo. Se presentan fotografías del holotipo y el paratipo, así como la morfología del pene y del aparato genital femenino, el mapa de la zona de muestreo con las localidades tipo y la descripción del hábitat. Palabras clave: Marruecos, Zona de elevada diversidad, Caracol de agua dulce, Pikasia smenensis gen. n. sp. n. Received: 24 V 21; Conditional acceptance: 30 VII 21; Final acceptance: 9 IX 21 Abdelkhaleq Fouzi Taybi, Équipe de Recherche en Biologie et Biotechnologie Appliquées, Faculté Pluridisciplinaire de Nador, Université Mohammed Premier, Morocco.– Peter Glöer, Schulstr. 3, D–25491 Hetlingen, Germany.– Youness Mabrouki, Laboratoire de Biotechnologie, Conservation et Valorisation des Ressources Naturelles, Faculté des Sciences de Dhar El Mehraz, Université Sidi Mohamed Ben Abdellah, Fes, Morocco. Corresponding author: Y. Mabrouki. E–mail: younes_mab@hotmail.fr ORCID ID: A. F. Taybi: 0000-0001-9652-5407; P Glöer: 0000-0001-6995-3641; Y. Mabrouki: 0000-0002-7336-8717

ISSN: 1578–665 X eISSN: 2014–928 X

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Introduction Surrounded by the Mediterranean Sea, the Atlantic Ocean and the Sahara Desert, Morocco is one of the most interesting biogeographical regions in the occidental. Mediterranean Basin, representing a contact area between Europe and Africa and between the Palaearctic and Afrotropical region. The country has multiple geographical barriers, such as the Moulouya River Basin, the Sahara, the Rif Mountains and the Atlas Mountains. The latter divide the northern part of the country into two bioclimatic regions, which in turn are associated with high levels of endemism in freshwater fauna (Mabrouki et al., 2019; Taybi et al., 2020), giving the area a privileged place for taxonomical and ecological studies. The truncatelloidean family Hydrobiidae Stimpson 1865 is a major group of freshwater molluscs and supposedly the largest gastropod family in Morocco. Knowledge of this family continually improving, and many new species have recently been discovered (Taybi et al., 2017; Boulaassafer et al., 2018, 2020; Ghamizi, 2020; Glöer et al., 2020a, 2020b; Mabrouki et al., 2020, 2021a, 2021b). The valvatiform hydrobiid is a group of minute gastropods with depressed trochiform shells resembling those of the genus Valvata O. F. Müller, 1773. Owing to their limited dispersal abilities and high

degree of habitat specialization, most springsnails are narrow–range endemics and face a high risk of extinction (Delicado et al., 2019; Radea et al., 2021). To date, five valvatoid species are known to occur in Morocco, four crenobiotic species that are micro–endemic to their type localities, namely Ifrania zerroukansis Glöer, Mabrouki and Taybi, 2020, Fessia aouintii Glöer, Mabrouki and Taybi, 2020, Islamia tiferitensis Glöer, Mabrouki and Taybi, 2020 and I. karawiyiensis Mabrouki, Glöer and Taybi, 2021, and the stygobiont Rifiya yakoubii Ghamizi, 2020 inhabits the phreatic waters of the southern border of the Rif region, upstream of Moulouya, Sebou and Loukkos basins. New research conducted recently in northwestern Morocco revealed a new valvatoid genus. The aim of this paper is to describe a new springsnail genus and species. Material and methods Sampling. Field surveys were conducted in the northern part of the country, including its great natural barriers such as the Moulouya River basin and the Middle Atlas massif. The samples of benthic fauna (including gastropods) were collected using a kick net and clamps. The samples were fixed in 75 % ethanol.

1 1 mm

4 p

bc ov

Fig. 1–6. Pikasia smenensis n. sp.: 1, holotype; 2–3, paratype; 4, penis in situ; 5, penis; 6, female sex tract; bc, bursa copulatrix; ov, oviduct; p, penis; t, tentacle. Fig. 1–6. Pikasia smenensis n. sp.: 1, holotipo; 2–3, paratipo; 4, pene in situ; 5, pene; 6, aparato genital femenino; bc, bolsa copulatriz; ov, oviducto; p, pene; t, tentáculo.

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6º 0' 0'' W

4º 30' 0'' W

la nt ic

O ce or an oc co

Mediterranean Sea

Atl a

nti

cO ce

35º 0' 0'' N

Taounate Taza Fez

80 km

Fig. 7. The sampling sites and habitat of Pikasia smenensis n. gen. n. sp. Fig. 7. Los sitios de muestreo y el hábitat de Pikasia smenensis n. gen. n. sp.

The genital organs were dissected and the shells were measured using a stereo microscope (Leica M205C); photos were taken with a Leica M205C microscope with a digital camera Leica DMC5400. The type material is stored in the Zoological Museum of Hamburg (ZMH). Results Phylum Mollusca Cuvier, 1795 Class Gastropoda Cuvier, 1795 Superorder Caenogastropoda Cox, 1960 Superfamily Truncatelloidea Gray, 1840 Family Hydrobiidae Stimpson, 1865 Pikasia n. gen. ZooBank LSID http://zoobank.org/urn:lsid:zoobank. org:act:3A261C25-8002-4045-8991-69104B043BE2 Pikasia smenensis n. sp. ZooBank LSID: http://zoobank.org/urn:lsid:zoobank. org:act:4A1C6BC3-8529-4832-B315-67473083D69B Holotype From Ain Smen spring (33º 57' 56.5'' N–5º 01' 18.5'' W): shell height 0.97 mm, shell width 1.1 mm, ZMH 140880.

Paratypes From Ain Chqef spring (34º 00' 05.8'' N 5º 01' 51.6'' W): 5 specimens ZMH 140881, 18 specimens in coll. Glöer, from site 5: 2 specimens coll. Glöer, from site 6: 30 specimens coll. Glöer. Description Valvatoid shell with 3.5 slightly convex whorls and a prominent body whorl (fig. 1, 2). The spire low and conical, diameter of the whorls fast and regularly increasing. Aperture roundish from frontal view and touches in some specimens the body whorl over a short distance or is detached from the shell wall. From lateral view the border of the aperture appears clearly inclined (fig. 3). The diameter of the body whorl near the aperture about 2.5 times broader than the umbilicus. Operculum dark yellowish. Dimensions: shell height 0.97–1.0 mm, shell width: 1.0–1.1 mm. Animal Eyes present, head and mantle blackish, pigmented, as were parts of the tentacles. Female sex tract The bursa copulatrix spherical (fig. 6), receptaculum absent. Penis In its relaxed state the simple penis broad and elon-

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gate flat with an acute penis tip. The distal part of the penis with an elongated triangular blackish spot (fig. 4, 5). Differentiated characters The shells are a little similar to Aretiana wolfi (Boeters and Glöer, 2007) from S–Spain but A. wolfi is much larger (1.6–1.8 mm in diameter), and the penis in A. wolfi is smaller and hook shaped (Delicado et al., 2021). In addition, A. wolfi has a receptaculum, Pikasia smenensis has not. Etymology The genus name Pikasia n. gen. is given in honor of the late father of the first author (El Pikas). The specific name refers to one of the type localities Ain Smen. Habitat The species occurs in rheocrenous springs only, in the upper part of Ain Chkef catchment area (a tributary of Sebou River). The grain size of the bottom consists of stones, pebbles and sand, the banks are covered with dense vegetation (fig. 7). The entire area is under anthropogenic impact. Distribution Morocco; only known from type localities. Acknowledgements We thank the editor and reviewers for their helpful comments that improved the present paper. References Boulaassafer, K., Ghamizi, M., Delicado, D., 2018. The genus Mercuria Boeters, 1971 in Morocco: first molecular phylogeny of the genus and description of two new species (Caenogastropoda, Truncatelloidea, Hydrobiidae). ZooKeys, 782: 95–128. Boulaassafer, K., Ghamizi, M., Machordom, A., Delicado, D., 2020. Phylogenetic relationships within Pseudamnicola Paulucci, 1878 (Caenogastropoda: Truncatelloidea) indicate two independent dispersal events from different continents to the Balearic Islands. Systematics and Biodiversity, 18(4): 396–416. Delicado, D., Arconada, B., Aguado, A., Ramos, M. A., 2019. Multilocus phylogeny, species delim-

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itation and biogeography of Iberian valvatiform springsnails (Caenogastropoda: Hydrobiidae), with the description of a new genus. Zoological Journal of the Linnean Society, 186: 892–914. Delicado, D., Pešić, V., Ramos, M. A., 2021. Arganiella Giusti & Pezzoli, 1980 (Caenogastropoda: Truncatelloidea: Hydrobiidae): a widespread genus or several narrow–range endemic genera? European Journal of Taxonomy, 750: 140–155, Doi: 10.5852/ ejt.2021.750.1369 Ghamizi, M., 2020. New stygobiont genus and new species (Gastropoda, Hydrobiidae) from the Rif (Morocco). Ecologica Montenegrina, 31: 50–56. Glöer, P., Mabrouki, Y., Taybi, A. F., 2020a. A new genus and two new species (Gastropoda, Hydrobiidae) from Morocco. Ecologica Montenegrina, 28: 1–6, Doi: 10.37828/em.2020.28.1 – 2020b. Two new valvatoid genera (Gastropoda, Hydrobiidae) from Morocco. Ecologica Montenegrina, 30: 124–128. Doi: 10.37828/em.2020.30.12 Mabrouki, Y., Taybi, A. F., Skalli, A., Sánchez–Vialas, A., 2019. Amphibians of the Oriental Region and the Moulouya River Basin of Morocco: distribution and conservation notes. Basic and Applied Herpetology, 33: 19–32. Mabrouki, Y., Taybi, A. F., Glöer, P., 2020. New additions to gastropod fauna (Gastropoda: Hydrobiidae, Lymnaeidae) of Morocco. Ecologica Montenegrina, 31: 40–44. – 2021. Two new species of the genera Islamia and Mercuria (Gastropoda, Hydrobiidae) from Morocco. Ecologica Montenegrina, 39: 76–80. – 2021. Further records of freshwater Gastropods (Mollusca: Hydrobiidae, Lymnaeidae, Planorbidae) from Morocco. Bonn Zoological Bulletin, 70(2): 273–279, Doi: 10.20363/BZB-2021.70.2.273 Radea, C., Lampri, P., Bakolitsas, K., Parmakelis A., 2021. A new hydrobiid species (Caenogastropoda, Truncatelloidea) from insular Greece. Zoosystematics and Evolution, 97(1): 111–119. Taybi, A. F., Mabrouki, Y., Ghamizi, M., Berrahou, A., 2017. The freshwater malacological composition of Moulouya’s watershed and Oriental Morocco. Journal of Materials and Environmental Science, 8(4): 1401–1416. Taybi, A. F., Mabrouki, Y., Berrahou, A., Dakki, A., Millán, A., 2020. Longitudinal distribution of macroinvertebrate in a very wet North African basin: Oued Melloulou (Morocco). International Journal of Limnology, 56(17): 1–11, Doi: 10.1051/limn/2020016

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A comparison of the diet of urban and forest great tits in a Mediterranean habitat J. C. Senar, A. Manzanilla, D. Mazzoni

Senar, J. C., Manzanilla, A., Mazzoni, D., 2021. A comparison of the diet of urban and forest great tits in a Mediterranean habitat. Animal Biodiversity and Conservation, 44.2: 321–327, Doi: https://doi.org/10.32800/ abc.2021.44.0321 Abstract A comparison of the diet of urban and forest great tits in a Mediterranean habitat. The low breeding performance and body condition of nestling passerine birds in urban environments has been attributed to the poor quality and low abundance of food in these settings. However, detailed data on prey provided by parents to their chicks in the urban habitat is scarce. Here we used video cameras set in nest boxes to compare the diet of urban and forest great tits Parus major when provisioning their chicks in a Mediterranean area. We additionally analysed brood size and fledgling success. Breeding success of urban great tits was lower than that of forest birds. Urban parents displayed a lower average hourly feeding rate per nestling than forest parents. Among the three prey item categories, the percentage of spiders did not vary according to habitat. However, the percentage of caterpillars delivered to the nest by great tit parents was higher in the forest than in the urban habitat while the percentage of 'other' prey showed a reverse pattern. 'Other' prey were mainly adult butterflies and wasps in the urban habitat. Our paper adds to the view that the low feeding rates and scarcity of caterpillars in urban environments may be the underlying cause constraining the growth of great tit nestlings in these areas. Key words: Urbanization, Diet, Great tits, Parental provisioning, Prey composition, Prey size Resumen Comparación de la dieta de los carboneros comunes en entornos urbanos y forestales en un hábitat mediterráneo. El escaso éxito reproductor y la condición física deficiente de los pollos de paseriformes en entornos urbanos se han atribuido a la escasez de alimentos en estos ambientes y a la mala calidad de estos. No obstante, existen pocos datos detallados sobre las presas que los progenitores llevan a sus pollos en el hábitat urbano. En este estudio, empleamos videocámaras instaladas en cajas nido con objeto de comparar la dieta que los carboneros comunes, Parus major, del medio urbano y forestal proporcionan a sus pollos en una zona del Mediterráneo. Asimismo, analizamos el tamaño de la nidada y el éxito de los volantones. El éxito reproductor de los carboneros comunes del medio urbano fue inferior al de las aves forestales. Se observó que la tasa media de alimentación por hora y por nidada de los progenitores del medio urbano fue inferior a la de los progenitores del medio forestal. Entre las tres categorías de presas, el porcentaje de arañas no varió en función del hábitat. Sin embargo, el porcentaje de orugas que los progenitores de carbonero común llevaron a los nidos fue mayor en el bosque que en el hábitat urbano, mientras que el porcentaje de "otras" presas mostró la pauta inversa. En el hábitat urbano, la categoría "otras" presas estuvo principalmente compuesta por mariposas y avispas adultas. Nuestro artículo se suma a la opinión de que las tasas de alimentación bajas y la escasez de orugas en los entornos urbanos pueden ser los factores que limitan el crecimiento de los pollos de carbonero común en estas zonas. Palabras clave: Urbanización, Dieta, Carboneros comunes, Aprovisionamiento parental, Composición de las presas, Tamaño de las presas Received: 21 VI 21; Conditional acceptance: 2 IX 21; Final acceptance: 20 IX 21 J. C. Senar, A. Manzanilla, D. Mazzoni, Museu de Ciències Naturals de Barcelona, Parc Ciutadella, Passeig Picasso s/n., 08003 Barcelona, España (Spain). Corresponding author: J. C. Senar. E–mail: jcsenar@bcn.cat ORCID ID: J. C. Senar: 0000-0001-9955-3892; D. Mazzoni: 0000-0001-7342-4857 ISSN: 1578–665 X eISSN: 2014–928 X

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Introduction Breeding performance and nestling body composition in birds in urban environments are inferior to those in birds in more natural habitats (Solonen, 2001; Chamberlain et al., 2009; Bailly et al., 2016; Demeyrier et al., 2016; Seress et al., 2018). According to the food limitation hypothesis (Newton, 1998), food shortage and low food quality in urban environments may be a main proximate reason for the lower breeding success in urban passerine populations (Eeva et al., 1997; Robb et al., 2008; Remacha and Delgado, 2009; Seress et al., 2018) Analyses of prey composition delivered by parents to their chicks comparing forest and urban insectivorous birds stress that forest parents provide a higher proportion of caterpillars to their chicks than urban parents (Riddington and Gosler, 1995; Pollock et al., 2017; Seress et al., 2018). Urban birds seem instead to rely mostly on adult Diptera, Coleoptera or Aranea (Riddington and Gosler, 1995). However, except for this paper, data on alternative prey delivered by parents to their chicks in the urban habitat are scarce. Food supplementation experiments should be a good approach to solve whether the lower breeding success of urban birds is due to a limitation in the quantity or quality of food collected by urban birds. However, food supplementation experiments in urban birds have found positive (Bańbura et al., 2011; Seress et al., 2020), negligible (Meyrier et al., 2017), and even negative (Demeyrier et al., 2017) impact on body size and/or nestling survival. It is therefore unclear the extent to which reductions in breeding success of urban birds are driven by a reduced abundance of natural 'high quality' dietary components. As stated by Demeyrier et al. (2017), further detailed knowledge on the diet of passerine birds in cities is needed. The aim of this paper was to compare the breeding success and diet of urban and forest great tits Parus major in a Mediterranean locality, and to analyse in detail the composition of prey delivered by urban great tit parents to their chicks. We used a digital micro–camera attached to the nest–box roof and focused on the entrance so as to record delivered prey. We focused on the great tit because it is a clear model species in studies of the effects of urbanization on breeding ecological parameters (see previous references). Material and methods We analysed the breeding success and diet of great tit nestlings in an urban habitat and in a forest habitat during the breeding season in 2018 and 2019. Forest data were collected at the Can Catà field station, located in the Collserola Natural Park (Cerdanyola, Barcelona, 90 NE of the Iberian Peninsula, 45º 27' N, 2º 8' E). At this location there were a total of 182 nest boxes. Urban birds were studied in three sub–urban parks in the city of Barcelona: Sentmenat, Laberint d’Horta and Desert de Sarrià (see Björklund et al., 2010 for details on the location of the parks). We

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placed 14 nest boxes in Sentmenat, 12 in Labertint d’Horta and 8 in Desert de Sarrià. The forest area is 3 km from Laberint d’Horta and about 7 km from Sentmenat and Sarrià. We visited nest boxes 2–3 times a week during the reproduction season in order to control the construction status of the nests and to determine laying date and clutch size. Hatching date was determined from daily nest checks starting 2 days before the expected hatch date. Once the hatching date was determined, the nests were visited as little as possible until the day of recording to minimize the negative effects of human presence. We ringed nestlings at 14–17 days old (around five days before fledging), and posterior checkings allowed to determine fledling success as number of chicks abandoning successfully the nest. Data on clutch size and on fledgling success was analyzed with GLM, including factor habitat (forest/ city) and also year (2018/2019) to standarize for its effect. A digital micro–camera (Mini Colour Sony IR Camera SK–C170IR) attached to the nest–box roof was located and focused on the entrance, so that delivered prey could be observed. These cameras have an infrared vision and a motion sensor. The cameras were installed on an afternoon when the chicks were between 7 and 13 days old (with a median age of 9 days old), and continuous recordings were made until at least 12 p.m. the next day. The afternoon of the first day was excluded from the analyses so as to accustom the birds to the presence of the camera. We counted the nestlings again when we installed the video camera. We did not observe any desertion because of the presence of the camera. We used the 5–hour recording of the second day, from 7:00 a.m. to 12:00 p.m., to collect the data concerning diet (Pagani–Núñez and Senar, 2013). Recordings were obtained from 29 of April to 30 of June, thus being representative of the whole breeding season. Once all boxes were recorded, the videos were analysed (n = 83 forest, n = 10 city). To avoid any bias, all the videos were analyzed by the same person (AM). We determined the parents' sex, prey type, prey size, and exact time for each feeding action using Micro D Player software. To differentiate males from females, we used the shininess of the black cap, which is glossier in males. This sexual dichromatism is accentuated under infrared light (Pagani–Núñez and Senar, 2014). We classified prey into three categories: caterpillars, spiders, and 'others'. The 'others' included Hymenoptera, Coleoptera, Orthoptera, Phasmida, Diptera, fruits and other unidentified prey. 'Artificial' food was not detected, and we should stress that it is rare for people in Barcelona to have bird tables. In our area it also ssems that great tits do not use bird feeders when they are rearing chicks. Although not all prey could be clearly identified, we were able to categorise around 90 % of the prey. Prey size was estimated in relation to the length of the bill (average 9 mm) and according to a semi–quantitative scale: small (less than 9 mm), medium (9.1 mm–12 mm) and large (longer than 12.1 mm) (García–Navas and Sanz, 2010; Pagani–Núñez et al., 2011).

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We computed the absolute number of total feeding actions brought by each parent to each nest box. In most (99 %) of the feeding actions, parents brought a single prey. We thus assume one prey per visit when estimating the number of prey per hour. Feeding rate (number of prey per hour per nestling) was used as a dependent variable in a general linear model (GLM). We used a natural logarithm transformation to fit the dependent variable to a normal distribution and homogenize its variance (Guisande González et al., 2013). Variables sex, habitat and year were used as categorical variables and date (days from April 1st) and the age of the chicks (in days) during videotaping as continuous variables. Sex, year, date and the age of the chicks were added into analyses to control for their effect when analysing habitat differences. We tried to add nest box identification as a random factor, to control for the effect that each nest box appeared twice (for the male and for the female), but this resulted in a over–parameterized model, so we deceided not to include this variable. This was also the case with the next analyses. Prey composition was also analysed through a multiple general linear model (MGLM). Prey composition, taken as the percentage of the three main prey types, was logit transformed (Guisande González et al., 2013). Variables sex, habitat and year were used as categorical variables and date (days from April 1st), age of the chicks (in days) during videotaping, and the number of chicks in the nest were used as continuous variables. The average size of each of the groups of prey brought to each nest box was similarly analyzed using an MGLM, where sex, habitat and year were the categorical variables and date, number of chicks in the nest, and their age were the continuous variables. The specific composition of 'other' prey in the urban area was computed for each nest box and values then were averaged. In this analyses we added data from a nest box recorded in 2015 in Desert de Sarria. This nest box was not included in the previous analyses because we had only one nest box and this would not allow to include factor year in analyses. Prey composition of that nestbox was similar to that of other urban nestboxes recorded in 2018 and 2019, not biasing results but allowing to increase sample size. Mean values are provided with ± S.E. Results Clutch size was larger in the forest than in the city (forest: 8.2 ± 0.11 SE eggs; city: 6.4 ± 0.46; F1,236 = 13.48, p < 0.001). The number of fledglings successfully leaving the nest did not differ between the forest and the city when taking into account abandoned nests or nests in which chicks died from starvation (i.e. breeding success = 0) (forest: 4.3 ± 0.20 fledglings; city: 3.9 ± 0.90; F1,204 = 0.24, p = 0.62). However, considering only nests where at least one chick fledged, the number of fledglings successfully leaving the nest was higher for the forest than for the city (forest: 6.0 ± 0.17 fledglings; city: 4.2 ± 0.73; F1,136= 5.59, p < 0.05).

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Table 1. GLM analysis of the variation in provisioning rate (/chick /hour) according to habitat (urban vs. forest), date (days from 1st April), age of the chicks (in days), sex of the parents and year (2018 or 2019). Year was included as a random factor. Tabla 1. Análisis mediante un modelo lineal generalizado de la variación de la tasa de aprovisionamiento (por pollo y por hora) según el hábitat (urbano o forestal), la fecha (días a partir del 1 de abril), la edad de los pollos (en días), el sexo de los progenitores y el año (2018 o 2019). El año se incluyó como factor aleatorio.

Habitat

0.18

7.62

< 0.01

Chick age

0.06

0.95

0.33

Parent sex

0.09

1.81

0.18

Date

–0.42

41.52

< 0.001

Year

–0.39

36.95

< 0.001

Urban parents displayed a lower average hourly feeding rate per nestling than forest parents (urban: 1.26 ± 0.32; forest: 2.28 ± 0.11 feedings/chick/hour; mean ± SE) (table 1). The chick provisioning rate was not affected by the age of the chicks or the sex of the parents (table 1). The feeding rate decreased significantly aover the season, and in 2019 the feeding rate was higher than that in 2018 (sites combined) (table 1). The median number of prey items provided by an individual during the 5 hours of recording was 59 (range 8–195) for the forest habitat and 27 (range 8–58) for the urban habitat. The percentage of caterpillars delivered to the nest by great tit parents was higher in the forest (0.70 ± 0.02) than in the urban habitat (0.42 ± 0.04) (table 2, fig. 1). The median number of caterpillars provided by an individual in the five hours in which they were recorded was 38 (range 2–185) in the forest and 8 (range 0–29) in the city. The percentage of caterpillars also increased with the age of the chicks, and brood size, and was higher in 2018 than in 2019 (table 2). Date and sex of parents had no significant effect (table 2). The percentage of 'other' prey delivered to the nest by great tit parents was lower in the forest (0.24 ± 0.01) than in the urban habitat (0.49 ± 0.04) (table 2, fig. 1). The median number of 'other' prey provided by an individual was 8 (range 0–110) in the forest and 12 (range 3–43) in the city. The percentage of spiders did not vary according to habitat (0.06 ± 0.01 vs. 0.08 ± 0.01) and was only affected by brood size (higher in smaller broods) and year (higher in 2018). The median number of spiders provided by an individual was 3 (range 0–28) in the forest and 2 (range 0–8) in

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Table 2. MGLM analysis comparing the percentage (ln transformed) of caterpillars, spiders and 'other' prey according to date, age of the chicks (in days), brood size, habitat (forest vs. city), year (2018 or 2019) and sex of the parents. Tabla 2. Análisis mediante un modelo lineal generalizado multivariante para comparar el porcentaje (transformado logarítmicamente) de orugas, arañas y "otras" presas en función de la fecha, la edad de los pollos (en días), el tamaño de la nidada, el hábitat (forestal o urbano), el año (2018 o 2019) y el sexo de los progenitores.

Caterpillars

Spiders

Other

ß t p ß t p ß t p Date

0.01

0.12

Chick age

0.20

Brood size

0.26

Habitat

0.91

0.01

0.99

0.05

0.68

2.78 < 0.01

–0.16 –1.93

0.06

–0.18

–2.64 < 0.01

2.69 < 0.01

–0.26 –2.28

0.02

–0.25

–2.76 < 0.01

–0.22

3.02

0.15

–1.79

0.08

0.20

–2.83 < 0.01

Year

–0.23

2.55 < 0.01

–0.37

3.49 < 0.001

0.36

–4.18 < 0.001

Parent sex

–0.11

–1.60

0.15

1.84

0.08

1.18

< 0.01 0.11

0.00

the city. Other prey in the urban habitat consisted mainly of Lepidoptera (butterflies, 50 % ± 12.1) and Hymenoptera (31 % ± 12.0) adults, and to a lesser extent to Orthoptera (11 % ± 8.0) and Diptera adults (8 % ± 4.4). Caterpillar size and 'other' prey size did not vary according to habitat. Average spider size was larger in the city than in the forest (forest: 2.4 ± 0.05; city: 2.5 ± 0.13) (table 3). Caterpillar size increased throughout the season, was larger in larger broods and larger in 2018. 'Other' prey size also increased over the season and was also larger in 2018 (table 3). Females provided larger spiders than males, and spider size was also larger in 2018 (table 3). Discussion Our paper supports previous data indicating that the productivity of urban great tits is lower than that of their forest counterparts (Solonen, 2001; Chamberlain et al., 2009; Bailly et al., 2016; Demeyrier et al., 2016; Seress et al., 2018). Our analyses also showed that the prey composition of parents provisioning nestlings in the forest habitat in Barcelona was dominated by caterpillars (70 % of all delivered prey). This contrasted with data from the urban habitat where caterpillars made up only 42 % of delivered prey. 'Other' prey followed the reverse pattern, with 24 % of delivered prey in the forest being 'other' insects, while values increased to 49 % in the urban area. We acknowledge that our sample size was low for the city area, but despite the low power the effect sizes were large. Our results are fundamentally similar to data from other studies where caterpillars are the main prey in forests and adult stages of other insect groups constitute a great percentage of prey delivered in

0.07

0.50

0.24

urban areas (Riddington and Gosler, 1995; Pollock et al., 2017). Data from Barcelona have shown that alternative prey in the city were mainly butterflies, wasps and grasshoppers, findings that differ from those in Britain, where 'other' prey mainly consisted of beetles and flies (Riddington and Gosler, 1995). This difference is probably related to the relative abundance of different prey between areas, acknowledging that different urban areas may differ in the relative proportion of 'green' areas, plant composition, and managing practices, so that availability of types of prey may differ greatly between cities. In relation to the size of the prey, we did not find differences between the forest and the urban habitat. This contrasts with data from blue tits Cyanistes caeruleus in Glasgow, where caterpillars provided to chicks in the forest area were larger than those in the urban area (Pollock et al., 2017) and data from house sparrows Passer domesticus in Hungary, where rural house sparrow parents provided larger prey items than urban parents (Seress et al., 2012). In contrast, the length of caterpillar prey provided by great tit parents in Belgium did not vary with urbanization (Satgé, 2016), similarly to our data from Barcelona. Again, the difference between studies is probably related to the relative abundance of different prey sizes in different areas. When analysing feeding rates, we found that urban parents in Barcelona displayed a lower average hourly feeding rate per nestling than forest parents. This result is similar to that observed in starlings Sturnus vulgaris where nestlings in the city center received less food (Mennechez and Clergeau, 2006). However, provisioning rates in great tits in Belgium were not found to vary with urbanization (Satgé, 2016), and data from blue tits in Glasgow and great tits in Sweden showed the reverse pattern, with urban

Animal Biodiversity and Conservation 44.2 (2021)

325

0.8 0.7

Caterpillars

% of prey

0.6 0.5 0.4 0.3 'Others'

0.2 0.1

Spiders

0.0

Forest

Urban

Fig. 1. Proportions of different prey provided by great tits to chicks according to forest or urban habitat. The most abundant food provided in the forest environment was caterpillars, followed by the 'others' group, and finally, spiders. In the urban environment, the proportion of caterpillars decreased significantly compared to forest habitat and the proportion of 'others' increased. The abundance of spiders did not vary significantly. Sample size of nests: forest N = 83, city N = 10. Fig. 1. Proporción de las distintas presas proporcionadas por los carboneros comunes a los pollos según el hábitat sea forestal o urbano. El alimento más abundante en el entorno forestal fueron las orugas, seguidas de "otras" presas y por último, de las arañas. En el entorno urbano, la proporción de orugas descendió significativamente en comparación con el hábitat forestal y la proporción de "otras" presas aumentó. La abundancia de arañas no varió de forma significativa. Tamaño de la muestra de nidos: bosque (N = 83), ciudad (N = 10).

Tabla 3. MGLM analysis of the variation in the size of caterpillars, spiders and 'other' prey provided by great tit parents to their chicks. Variation is analyzed according to date (days from 1tst April), age of the chicks, brood size, habitat (forest vs. Urban), year (2018 or 2019) and sex of the parents. Tabla 3. Análisis mediante un modelo lineal generalizado multivariante de la variación del tamaño de las orugas, las arañas y "otras" presas proporcionadas por los progenitores de carbonero común a sus pollos. La variación se analiza en función de la fecha (días a partir del 1 de abril), la edad de los pollos, el tamaño de la nidada, el hábitat (forestal o urbano), el año (2018 o 2019) y el sexo de los progenitores.

Caterpillars

Spiders

Other

ß t p ß t p ß t p Date

0.38

5.73

< 0.001

0.12

1.59

0.11

0.27

3.41

< 0.001

Chick age

0.08

1.21

0.23

0.06

0.79

0.43

0.09

1.13

0.26

Brood size

0.18

2.00

< 0.05

0.08

0.76

0.45

–0.03 –0.27

0.79

Habitat

0.05

–0,79

0.43

0.21

–2.70 < 0.01

0.09

–1.1

0.27

Year

–0.51

6.14

< 0.001

–0.44

4.55 < 0.001

–0.39

3.94

< 0.001

Parent sex

–0.04 –0.67

0.50

0.2

2.75

–0.14

–1.8

0.07

< 0.01

326

nestlings being fed more often than forests nestlings (Isaksson and Andersson, 2007; Pollock et al., 2017). Differences between studies could perhaps be due to differences in traveling distances across different areas when provisioning their nestlings (Demeyrier et al., 2017). This difference could also be the result of the trade–off between prey size and feeding rates: when meals provided by parents include large prey, the number of trips is lower, and when prey are small, the number of trips is higher (Grieco, 2001, 2002; Navalpotro et al., 2016). To conclude, our paper adds to the growing view that the diet of insectivorous nestlings in urban areas is deficient in caterpillars. This may be the reason underlying reduced breeding success in the urban habitat (Bańbura et al., 1999; Pollock et al., 2017; Seress et al., 2018). However, without more exact nutritional data regarding the various species of prey items provided to nestlings it is not clear whether a combination of other insects and related arthropods could fullfill the nutritional requirements of insectivorous birds in urban habitats. We therefore urge urban ecologists to analyze in detail the nutritional profile of the diet of urban and forest insectivorous birds in order to understand why, as suggested (see above), caterpillars, and only caterpillars, can make the difference. Acknowledgements We are most grateful to Mark Finke for reviewing the paper. We also thank Lluisa Arroyo, Helena Navalpotro and Xavier Altarriba for their help in the field and in the lab. We thank the Gil family, owners of Can Catà, for allowing us to work on their property. Birds were handled with the permission of the Departament de Medi Ambient, Generalitat de Catalunya (SF/542). The Catalan Institute of Ornithology (ICO) provided the rings. This work was supported by funds from the Ministry of Economy and Competitiveness, Spanish Research Council to JCS (CGL–2016–79568–C3–3–P). References Bailly, J., Scheifler, R., Berthe, S., Clément–Demange, V.–A., Leblond, M., Pasteur, B., Faivre, B., 2016. From eggs to fledging: negative impact of urban habitat on reproduction in two tit species. Journal of Ornithology, 157(2): 377–392, Doi: 10.1007/s10336-015-1293-3 Bańbura, J., Bańbura, M., Gladalski, M., Kalinski, A., Markowski, M., Michalski, M., Nadolski, J., Skwarska, J., Zielinski, P., 2011. Body Condition Parameters of Nestling Great Tits Parus major in Relation to Experimental Food Supplementation. Acta Ornithologica, 46(2): 207–212, Doi: 10.3161/000164511X625991 Bańbura, J., Lambrechts, M. M., Blondel, J., Perret, P., Cartan–Son, M., 1999. Food Handling Time of Blue Tit Chicks: Constraints and Adaptation to Different Prey Types. Journal of Avian Biology, 30(3): 263–270, Doi: https://doi.org/10.2307/3677352. Björklund, M., Ruiz, I., Senar J. C., 2010. Genetic differentiation in the urban habitat: the great tits

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Animal Biodiversity and Conservation 44.2 (2021)

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ball serà establerta i s’ordenarà alfabèticament per autors i cronològicament per a un mateix autor, afegint les lletres a, b, c,... als treballs del mateix any. En el text, s’indicaran en la forma usual: "... segons Wemmer (1998)...", "...ha estat definit per Robinson 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. Grans quantitats de dades o taules numèriques molt llargues es publicaran com a Material suplementari. Aquest material suplementari només acompanyarà a la versió online de l'article, en cap cas a la versió impresa. 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 i 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 les quatre pàgines.

Animal Biodiversity and Conservation 44.2 (2021)

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

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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 en Animal Biodiversity and Conservation es gratuito para los autores, sin embargo 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 © 2021 Museu de Ciències Naturals de Barcelona Papers are published under a Creative Commons Attribution 4.0 International License

especulaciones y las citas bibliográficas. Irá encabezado por el título del trabajo en cursiva. Key words en inglés (un máximo de seis) que especifiquen el contenido del trabajo por orden de importancia. Resumen en castellano, traducción del abstract. Su traducción puede ser solicitada a la revista en el caso de autores que no sean castellano hablantes. Palabras clave en castellano. 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. Grandes cantidades de datos o tablas numéricas muy largas se publicarán como material suplementario. Este material suplementario sólo acompañará a la versión online del artículo, en ningún caso a la versión impresa. 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 44.2 (2021)

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, but 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. © 2021 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. Large amounts of data or long tables will be published as supplementary material. This supplementary material will accompany the online version of the article only, not the printed version. Main headings (Introduction, Material and methods, Results, Discussion, Acknowledgements and References) should not be numbered. Do not use more than three levels of headings. Manuscripts should not exceed 20 pages including figures and tables. If the article describes new taxa, type material must be deposited in a public institution. Authors are advised to consult recent issues of the journal and follow its conventions. Brief communications Brief communications should follow the same procedure as other articles and they will undergo the same review process. They should not exceed 2,300 words including title, abstract, figure and table legends, acknowledgements and references. The abstract should not exceed 100 words, and the number of references should be limited to 15. Section headings within the text are optional. Brief communications may have up to two figures and/or two tables but the whole paper should not exceed four published pages.

Animal Biodiversity and Conservation 44.2 (2021)

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

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Animal Biodiversity and Conservation 44.2 (2021)

251–265 Kihn, M. P., Caruso, N., Iaconis, K., Palacios González, M. J., Soler, L. Analysis of conflicts with wild carnivores in the Humid Chaco, Argentina

289–301 Martín, B., González–Arias, J., Vicente–Vírseda, J. A. Machine learning as a successful approach for predicting complex spatio–temporal patterns in animal species abundance

267–278 Cepeda–Duque, J. C., Gómez–Valencia, B., Alvarez, S., Gutiérrez–Sanabria, D. R., Lizcano, D. J. Daily activity pattern of pumas (Puma concolor) and their potential prey in a tropical cloud forest of Colombia

303–315 Solís, I., Sanz, J. J., Imba, L., Álvarez, E., Barba, E. A higher incidence of moult–breeding overlap in great tits across time is linked to an increased frequency of second clutches: a possible effect of global warming?

279–287 Bermúdez–Cavero, A., Gil–Delgado, J. A., López– Iborra, G. M. Modelling European turtle dove (Streptopelia turtur L. 1758) distribution in the south eastern Iberian Peninsula

317–320 Taybi, A. F., Glöer, P., Mabrouki, Y., Description of a new valvatoid Pikasia smenensis n. gen. n. sp. (Gastropoda, Hydrobiidae) from Morocco 321–327 Senar, J. C., Manzanilla, A., Mazzoni, D. A comparison of the diet of urban and forest great tits in a Mediterranean habitat

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 Abstracts, BIOSIS Previews, CiteFactor, Current Primate References, Current Contents/Agriculture, Biology & Environmental Sciences, Essential Science Indicators, 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–ICYT, International Abstracts of Biological Sciences, International Bibliography of Periodical Literature, International Developmental Abstracts, Latindex, Marine Sciences Contents Tables, MIAR, Oceanic Abstracts, RACO, Recent Ornithological Literature, REBIUN, REDIB, Referatirnyi Zhurnal, ResearchGate, Responsible Journals, Science Abstracts, Science Citation Index Expanded, Scientific Commons, SCImago, SCOPUS, Serials Directory, SHERPA/RoMEO, Transpose, Ulrich's International Periodical Directory, WoS, Zoological Records

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

Índex / Índice / Contents Animal Biodiversity and Conservation 44.2 (2021) ISSN 1578–665 X eISSN 2014–928 X 139–151 Govind, S. K., Jayson, E. A. Human–wildlife interactions and people’s attitudes towards conservation: a case study from Central Kerala, India 153–174 Fernández–Badillo, L., Zuria, I., Sigala–Rodríguez, J., Sánchez–Rojas, G., Castañeda–Gaytán, G. Revisión del conflicto entre los humanos y las serpientes en México: origen, mitigación y perspectivas 175–184 Pérez–González, J., Rey Gozalo, G., Montes González, D., Hidalgo de Trucios, S. J., Barrigón Morillas, J. M. Are quartzite scree slopes used by birds to promote sound transmission in the Mediterranean forest? 185–194 Postigo, J. L., Carrillo–Ortiz, J., Domènech, J., Tomàs, X., Arroyo, L., Senar, J. C. Dietary plasticity in an invasive species and implications for management: the case of the monk parakeet in a Mediterranean city 195–203 Martínez–Abraín, A., Ferrer, X., Jiménez, J., Fernández– Calvo, I. C. The selection of anthropogenic habitat by wildlife as an ecological consequence of rural exodus: empirical examples from Spain

205–211 Cordero–Rivera, A., Roucourt Cezário, R., Guillermo– Ferreira, R., Marques Lopez, V., Sanmartín–Villar, I. Can scientific laws be discussed on philosophical grounds? a reply to naïve arguments on 'predators' proposed by Bramble (2021) 213–217 Brief communication Estela, F. A., Sánchez–Sarria, C. E., Arbeláez–Cortés, E., Ocampo, D., García–Arroyo, M., Perlaza–Gamboa, A., Wagner–Wagner, C. M., MacGregor–Fors, I. Changes in the nocturnal activity of birds during the COVID–19 pandemic lockdown in a neotropical city 219–227 Bakhoum, S. A. Natural hybridization between immigrant n a r r o w – b a r r e d S p a n i s h m a c ke r e l Scomberomorus commerson (Lacepède, 1800) and endemic West African Spanish mackerel Scomberomorus tritor (Cuvier, 1832) in the Egyptian Mediterranean coast 229–239 Crespo, J., Jiménez, J., Martínez–Abraín, A. Increasing wild boar density explains the decline of a Montagu’s harrier population on a protected coastal wetland 241–250 Duarte Silveira, R. A., Marques da Rosa, H. H., Pereira, A. A., Passamani, M., Zenni, R. D. Natural factors but not anthropogenic factors affect native and non–native mammal distribution in a Brazilian National Park

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

FECYT–113/2021 FECHA DE CERTIFICACIÓN: 06 de octubre 2014 (4ª convocatoria) VÁLIDO HASTA: 13 de julio de 2022