ABC 42-1 (2019) by Museu Ciències Naturals de Barcelona MCNB

Ben J. Hatchwell, Univ. of Sheffield, UK

Dibuix de la coberta / Dibujo de la portada / Drawing of the cover: Lynx pardinus, linx ibèric, lince ibérico, Iberian lynx (Jordi Domènech)

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 42.1, 2019 © 2019 Museu de Ciències Naturals de Barcelona, Consorci format per l'Ajuntament de Barcelona i la Generalitat de Catalunya Autoedició: Montserrat Ferrer Fotomecànica i impressió: CEVAGRAF SCCL ISSN: 1578–665 X eISSN: 2014–928 X Dipòsit legal: B. 5357–2013

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

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Animal Biodiversity and Conservation 42.1 (2019)

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

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Diet composition of the Karpathos marsh frog (Pelophylax cerigensis): what does the most endangered frog in Europe eat? P. Pafilis, G. Kapsalas, P. Lymberakis, D. Protopappas, K. Sotiropoulos

Pafilis, P., Kapsalas, G., Lymberakis, P., Protopappas, D., Sotiropoulos, K., 2019. Diet composition of the Karpathos marsh frog (Pelophylax cerigensis): what does the most endangered frog in Europe eat? Animal Biodiversity and Conservation, 42.1: 1–8, https://doi.org/10.32800/abc.2019.42.0001 Abstract Diet composition of the Karpathos marsh frog (Pelophylax cerigensis): what does the most endangered frog in Europe eat? The Karpathos marsh frog (Pelophylax cerigensis) is considered the most endangered frog in Europe. Here we assess its feeding ecology and examine 76 individuals from the two known populations using the stomach flushing method. We also measured body weight, snout–vent length, mouth width and prey width and length. Pelophylax cerigensis follows the feeding pattern of green frogs of the adjacent areas, with Coleoptera, Araneae, Isopoda and Hymenoptera being the main prey groups. The two populations differed in body size but had similar values of prey abundance and frequency. It seems that P. cerigensis follows a strict feeding strategy. Further research on prey availability in its habitats will provide valuable insight. Key words: Diet, Endangered species, Islands, Frogs, Mediterranean Resumen Composición de la dieta de la rana de Kárpatos (Pelophylax cerigensis): ¿qué come la rana más amenazada de Europa? La rana de Kárpatos (Pelophylax cerigensis) es considerada la rana más amenazada de Europa. Aquí evaluamos su ecología alimentaria y examinamos 76 individuos de las dos poblaciones conocidas usando el método del lavado de estómago. También medimos el peso corporal, la longitud desde el hocico hasta la cloaca y el ancho de la boca de las ranas y el ancho y largo de las presas. La dieta de Pelophylax cerigensis, compuesta principalmente por Coleoptera, Aranean, Isopoda e Hymenoptera, es similar a la de otras especies de ranas verdes de las zonas adyacentes. Las dos poblaciones difieren en el tamaño corporal, pero presentan valores similares de abundancia y frecuencia de presas. Parece que P. cerigensis sigue una estricta estrategia de alimentación. El estudio de la disponibilidad de presas en sus hábitats aportará información valiosa. Palabras clave: Dieta, Especies en peligro de extinción, Islas, Ranas, Mediterráneo Received: 09 XII 17; Conditional acceptance: 12 II 18; Final acceptance: 12 III 18 Panayiotis Pafilis, Grigoris Kapsalas, Department of Biology, National and Kapodistrian University of Athens, Panepistimioupolis, Ilissia, 15784 Greece.– Petros Lymberakis, Natural History Museum of Crete, University of Crete, Irakleio, Greece.– Dinos Protopappas, Management Body of Karpathos–Saria, Olympos, 85700 Greece.– Konstantinos Sotiropoulos, Department of Biological Applications and Technology, University of Ioannina, Bizani, 45500 Greece. Corresponding author: P. Pafilis. E–mail: ppafil@biol.uoa.gr

ISSN: 1578–665 X eISSN: 2014–928 X

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Introduction The extensive anthropogenic activity in the last century has changed natural ecosystems and put many species at stake (Wilson, 2002; Pimm et al., 2014). Successful protection and effective conservation of threatened species require good knowledge of their overall biology (Tracy et al., 2002; Wikelski and Cooke, 2006; Bertolero and Oro, 2009). Feeding ecology is one of the most important biological parameters as it shapes numerous aspects of animal life (Vervust et al., 2010; Brown et al., 2017; Olsen, 2017). Assessing the dietary regimes of

endangered animals provides important insight into the identification of critical food resources and may contribute to integrated conservation plans for many animal taxa (Palazón et al., 2008; Pagani–Núñez et al., 2011; Butler et al., 2012).

The Karpathos marsh frog (Pelophylax cerigensis) (Beerli et al., 1994) is endemic to the island of Karpathos, south Aegean Sea, Greece (Valakos et al., 2008). Categorized as Critically Endangered by the IUCN (Beerli et al., 2009), it is considered the most endangered anuran amphibian in Europe because its range is restricted to y two small rivers (rivulets or brooks in reality) in the north part of the island (Temple and Cox, 2009). Such small, insular wetlands are nowadays considered the most endangered ecosystems in the Mediterranean Sea, representing isolated oases for birds, amphibians, aquatic reptiles and invertebrates (Cuttelod et al., 2008). The general biology of P. cerigensis is largely understudied since the first description of the species (Beerli et al., 1994). Feeding ecology is a classic scientific topic in frog studies as food quality and availability reveal the important position of the group in food webs (Duellman and Trueb, 1994). Frogs represent a considerable portion of the riparian biomass and serve as energy redirectors to higher trophic levels (Burton and Likens, 1975). Thus it is important to understand where they stand in food webs and to unravel how habitat global energy fuels riparian communities (Çiçek, 2011; Bogdan et al., 2013). Pelophylax cerigensis remains an unknown animal in terms of ecology, and as such, no specific protection measures have been taken so far. Here we studied the two known populations of P. cerigensis. We aimed to (1) assess the diet of the species for the first time, (2) examine possible differences between the two populations, and (3) compare trophic niches and food composition to those of other frogs in the Balkans. Material and methods We sampled both river sites at Argoni (35.6948 º N, 27.1523 º E) and Nati (35.7018 º N, 27.1786 º E) in the northern part of Karpathos Island. Frogs were collected during late spring (last week of May 2015, 2016 and 2017; average temperature and rainfall did not differ between the three years) from small ponds along riverbeds. Fieldwork was carried out in May, as this is the time of the year when the frogs are easier to observe and capture on Karpathos Island. During

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summer and early fall, the rivulets dry out and frogs resort to the nearby vegetation, being hard to catch. On the contrary, during late fall and winter, the rivulets turn to torrents, and frogs, once more, avoid them. The landscape in general is characterized by sparse and low vegetation in a rocky background with high erosion. The vegetation around the rivers consists of tall nerium shrubs (Nerium oleander), sparse pine trees (Pinus brutia) and low shrubs such as spiny rush (Juncus articulatus) and thyme (Thymbra capitata). Seventy–six individuals (51 from Argoni and 25 from Nati) were captured by net or hand and were anesthetized with an MS–222 solution. To anaesthetize the frogs we followed the instructions of the Amphibian Research and Monitoring Initiative (U.S. Geological Survey) (ARMI SOP No. 104–Standard Operating Procedure) for safe anesthesia (Downes, 1995). We placed the animals in a plastic water bath (2 cm deep) containing a tricaine methane sulfonate solution (50 mg/L) for 15 min. After this period, we rinsed its skin with fresh water to avoid deeper levels of anesthesia. Frogs started to recover after 10 min. For each frog we took the basic morphometric measurements: body weight (W) with a digital scale (i500 Backlit Display, My Weight, accurate to 0.1 g) and snout–vent length (SVL) and mouth width (MW) with a digital caliper (Silverline 380244, accurate to 0.01 mm). To remove stomach content we used the stomach flushing method (Solé et al., 2005). Besides being simple and effective, this method provides high quality results without sacrificing animals as it can be applied in live individuals. It is the most widely used, non–invasive technique in frogs (e.g. Lamoureux et al., 2002; Çiçek, 2011; Rebouças et al., 2013; Bogdan et el., 2013; Plitsi et al., 2016) with significantly less impact than other methods (Bondi et al., 2015) and thus it can be used even in endangered species (Bower et al., 2014; Watson et al., 2017). Holding the animal with one hand, we gently opened the mouth with a spatula and then carefully introduced the infusion tube (made of supple silicon to avoid perforations of oesophagus) of a 20 ml syringe that contained water from the pond where the frogs were captured. We flushed the content of the syringe into the stomach forcing out the consumed prey items till no more stomach content appeared. The water with stomach content was stored in a plastic glass and then decanted into a sieve. Prey items were collected with forceps and preserved in 70 % alcohol in small eppendorf tubes. After measuring the frogs and collecting stomach contents, we waited for the captured individuals to recover. We kept the frogs into a plastic bin for 30 min to ensure that all of them were in good condition. None of the individuals died during this procedure and all of them were released in their habitat after fully recovering. The personnel of the Management Body of Karpathos that regularly patrolled the river did not encounter any dead individual during the days following the measurements. Stomach contents were preserved in 70 % alcohol and then transported to the lab (Dept. of Biology, National and Kapodistrian University of Athens) where

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Table 1. Values for snout–vent length (SVL) and mouth width (MW) (in cm) and body weight (BW) (in g): means ± standard deviation; range (between brackets): N, sample size. Tabla 1. Valores para la longitud hocico–cloaca (SVL), el ancho de la boca (MW) (en cm) y el peso corporal (BW) (en g): media ± desviación estándar; rango (entre paréntesis); N, tamaño de muestra.

Site

SVL

Argoni

4.29 ± 0.81 (2.5–5.6)

1.68 ± 0.36 (0.9–2.3)

10.39 ± 5.44 (1.7–22.3)

Nati

3.76 ± 0.89 (2.3–5.4)

1.42 ± 0.38 (0.8–2.2)

7.43 ± 4.89 (1.5–18.2)

they were identified to order with a stereomicroscope (Wild Heerburg M38). Prey item width (W) and length (L) were measured to the nearest 0.1 mm. Following Dunham (1983), prey volume (V) was calculated using the ellipsoid volumetric formula: V = 4/3

o (L/2) · (W/2)2

At this point, we should point out that although this is the typical method used in similar studies, the volume of prey groups such as Diplopoda or Formicidae calculated with this approach is rather unrealistic. However, for the sake of comparison with other studies on frog feeding ecology, we apply it here as well. For each individual whose stomach contained prey items (contrary to empty stomachs), we calculated the minimum, mean and maximum prey item width, length and volume, while also counting the total number of prey items. For every identified prey category we calculated its relative abundance (%A), frequency of occurrence (%F) and relative volume (%V), per population as well as overall. We assessed food niche breadth using Levin's standardized index Bi (Levins, 1968), where pi is the relative abundance of every prey category in each population: Bi = 1/

S(pi)2

Niche overlap (O) was evaluated with Pianka's (1973) index, where pi is the relative proportion of prey category i in each of two populations A and B: n

S i piA · piB n

S i piA2 · Si piB2

We used t–tests to compare frog and prey morphometric variables between the two populations, unless the assumptions of normality (Shapiro–Wilk test, p < 0.05) and homogeneity of variances (Levene's test, p < 0.05) were violated, in which case the non–parametric Mann–Whitney test was used instead. A Mann– Whitney test was used to compare the number of prey items between the two populations. The relationship between frog body size and prey size was assessed

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with Pearson's product–moment correlation. In order to examine the similarity of diet composition between the two populations, x2–tests of independence were performed on absolute and relative abundance (N, %A) and frequency of occurrence (F, %F) of the five most common overall prey categories respectively. All statistical analyses were performed according to Zar (2010) using R 3.4.2 (R Core Team, 2017). Results The Argoni population consisted of larger (SVL: t–test, t = 2.574, df = 74, p = 0.0121) and heavier (W: t–test, t = 2.301, df = 74, p = 0.0242) individuals than those of Nati. Also, frogs from the latter study had smaller mouth widths (MW: t–test, t = –2.943, df = 74, p = 0.0043) (table 1). Stomach analyses yielded 296 prey items (199 from Argoni and 97 from Nati). Fifteen frogs, nine from Argoni and six from Nati, had empty stomachs. The two populations did not differ in the number of prey items consumed (Mann–Whitney U = 640, p = 0.98). A frog from the Argoni River had the highest number of prey items in a single individual (11). The main prey category was insects, with an overall value between the two populations reaching 64.65 %. Coleoptera and Araneae topped the list of the most abundant and the most frequently consumed prey taxa (table 2). Finally, Coleoptera occupied the largest global relative volume in frog stomachs (35 %). Eleven prey categories were present in both sites, while six taxa were found only at Argoni and three only at Nati. These unique prey groups, however, were of minor importance as they were represented by low relative abundance and frequency in the stomachs examined (table 2). Argoni frogs ate larger prey items than the Nati population (mean length: Mann–Whitney U = 703.5, p < 0.0001; mean width: Mann–Whitney U = 696, p < 0.0001) (table 3). When we additionally compared the mean minimum and mean maximum prey item lengths and widths, we found that the two populations differed considerably in all prey size variables (Mann–Whitney U, p < 0.05 for all comparisons). As a consequence, prey volume was also higher in the

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Table 2. The relative abundance (A%), frequency of occurrence (F%) and relative volume (V%) of prey consumed by the two P. cerigensis populations. Tabla 2. Abundancia relativa (A%), frecuencia de presencia (F%) y volumen relativo (V%) de las presas consumidas por las dos poblaciones de P. cerigensis.

Argoni River

Overall

Prey type

Araneae

14.07

50.00

2.51

13.40

57.89

5.65

13.85

52.46

2.81

Coleoptera

29.15

69.05

38.19

13.40

52.63

5.09

23.99

63.93

35.00

Coleoptera larvae

1.01

2.38

0.08

1.03

5.26

1.19

1.01

3.28

0.19

Dermaptera

2.51

11.90

1.23

3.09

15.79

8.38

2.70

13.11

1.92

Dictyoptera

1.01

4.76

0.37

1.03

5.26

1.64

1.01

4.92

0.49

Diplopoda

2.51

11.90

0.47

3.09

15.79

2.78

2.70

13.11

0.69

Diplura

0.00

1.03

5.26

0.00

0.34

1.64

0.00

Diptera

10.05

42.86

10.32

8.25

31.58

3.49

9.46

39.34

9.66

Ephemeroptera

0.50

2.38

0.00

0.34

1.64

0.00

Gasteropoda

5.03

21.43

2.94

7.22

31.58

16.60

5.74

24.59

4.25

Hemiptera

9.05

28.57

1.46

15.46

42.11

8.41

11.15

32.79

2.13

Hymenoptera

7.04

21.43

2.84

18.56

52.63

11.90

10.81

31.15

3.72

Isopoda

12.06

47.62

11.59

12.37

42.11

33.92

12.16

45.90

13.74

Lepidoptera

1.51

7.14

17.54

0.00

1.01

4.92

15.85

Lepidoptera larvae

1.01

4.76

0.27

0.00

0.68

3.28

0.24

Odonata

1.51

7.14

1.27

0.00

1.01

4.92

1.15

Odonata larvae

1.51

7.14

1.06

0.00

1.01

4.92

0.96

Orthoptera

0.00

1.03

5.26

0.96

0.34

1.64

0.09

Trichoptera

0.00

1.03

5.26

0.00

0.34

1.64

0.00

Lizards

0.50

2.38

7.86

0.00

0.34

1.64

7.10

Argoni population (mean volume: Mann–Whitney U = 709, p < 0.0001). Overall, there was a positive correlation between frog SVL and mean prey length [r = 0.603, t(59) = 5.81, p < 0.0001], as well as frog SVL and mean prey width [r = 0.463, t(59) = 4.02, p = 0.0002]. Levin's index was lower at Argoni than Nati (B = 6.787 and B = 8.063 respectively) (table 2). According to Pianka's index, the two populations share a high diet similarity (O = 0.840) (table 3). The most abundant prey taxa and the most frequently eaten prey items were similar in both populations (table 2). The two populations differed significantly when absolute abundance of the five most common prey groups was tested (x2 = 17.004, df = 4, p = 0.0019). However, relative abundance differences were not statistically significant (x2 = 0.13, df = 4, p = 0.998). No statistical differences were found between the two populations, either in absolute (x2 = 1.91, df = 4, p = 0.753) or in relative (x2 = 0.07, df = 4, p = 0.999) frequencies of occurrence.

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Nati River

Finally, a notable deviation was observed in the relative volume between the two populations: while Coleoptera were the dominant group in frog stomachs at the Argoni River (38.19 %), they reached only a small percentage (5.09 %) at Nati, where Isopoda was the first pray taxon in volume (33.92 %) (table 2). Discussion Karpathos is one of the oldest islands in the eastern Mediterranean (around 8 my). This long isolation is reflected in the two endemic amphibians occurring on the island, the Karpathos Lycian salamander (Lyciasalamandra helverseni) and P. cerigensis. However, although L. helverseni maintains dense populations, P. cerigensis is known from only two sparse populations in the north of Karpathos (Lymberakis et al., 2018). Climate change strongly affects the east Mediterranean, with increasingly fewer rainfalls (Giorgi and Lionello, 2008), further exacerbating the general

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Table 3. The number of examined individuals, the total number of prey items, mean dimensions of the consumed prey, and the feeding diversity (Levin's, B) and niche overlap estimation (Pianka) indices. Tabla 3. Número de individuos examinados, número total de presas, dimensiones medias de las presas consumidas, e índices de diversidad de la dieta (Levin, B) y de superposición de nicho (Pianka).

No of individuals

Nati

Empty stomachs (%)

17.6

Total number of prey

199

3.90/0–11

3.88/0–10

Mean prey length/range

9.43/2.62–32.00

5.56/1.30–25.00

Mean prey width/range

Average no of prey/range

3.60/0.59–20.20

2.13/0.20–5.50

Simpson's diversity index

6.787

8.063

Pianka's index

0.840

water scarcity on Karpathos. Thus, and also due to its restricted range, P. cerigensis faces severe risks in the immediate future. Here we present the first data on its diet. These findings should be taken into account to effectively protect the species. We should point out, however, that the most important protection task undertaken on the island should be to conserve the Karpathos environment as a whole. The Karpathos marsh frog adopted a rather simple diet, comprising fewer prey taxa than other ranid frogs in the area (Cogălniceanu et al., 2000; Mollov et al., 2006; Bisa et al., 2007; Bogdan et al., 2013). The dominant prey groups were insects, spiders and isopods. Frogs from the two examined populations prey on the same invertebrate taxa (11 common groups), with limited differentiation between them (table 2). A clear difference arose from the body size measurements: the frogs from Argoni were larger and heavier, and had a larger mouth width than their Nati peers (table 1). As a consequence, Argoni frogs ate larger food items despite the fact that the number of prey items per stomach was similar (Cogălniceanu et al., 2000) (table 3). Intrapopulation analysis revealed that body size also affected diet: larger frogs consumed larger prey in both populations. Argoni and Nati share the same ecological and abiotic parameters (vegetation, substrate, slope, exposure to winds and sunlight), and hence it is difficult to identify the reason underlying discrepancies in their diets. The strong relation between body and prey size may account for the differences in the taxonomy (limited though, as mentioned above) and size of the prey consumed. Diet compositions were quite similar between the two populations, suggesting a high dietary/niche overlap (Pianka's index, O = 0.840). The reason for such a high dietary overlap should be sought in habitat similarity: the two focal populations share the same vegetation type and main abiotic characteristics. The values of Levin's index (Bi) fall within the range of other

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Argoni

Balkan frogs (Covaciu–Marcov et al., 2010; Çiçek, 2011; Cicort–Lucaciu et al., 2011; Bogdan et al., 2012), indicating a rather broad feeding niche for the species. The two populations did not differ in the frequency of consumed prey. Argoni frogs primarily consumed Coleoptera (76.19 %), whereas Araneae was the most frequent prey taxon in Nati (table 2). Our findings confirm previous research reporting that Coleoptera and Araneae are a typical food resource for Pelophylax frogs (Sas et al., 2007; Balint et al., 2010; Mollov et al., 2010; Bogdan et al., 2012; Plitsi et al., 2016). Frequency, taken together with relative abundance in stomach content, provides a reliable evaluation of feeding homogeneity (Cogălniceanu et al., 2000). In both populations, frequency and relative abundance received high values (table 2). It is worth highlighting two particularities in the diet of P. cerigensis: the small percentage of aquatic taxa and the finding of a lizard tail in a single stomach from a frog in Nati. The limited consumption of aquatic prey is not a surprising feature among ranid frogs that typically seek their food in the banks of their habitats (Çiçek and Mermer, 2007; Mollov, 2008; Covaciu–Marcov et al., 2010; Çiçek, 2011). However, the particularly low portion in the P. cerigensis stomach content should be attributed to the physical water scarcity on the island. From April to October the rivers practically survive by small, shallow pools that cannot support a typical aquatic invertebrate fauna. Thus, frogs largely consume terrestrial and flying prey, and may include unusual food, such as the observed lizard tail. The latter belongs to the European copper skink (Ablepharus kitaibelii), a small lizard of the Scincidae family. Though the consumption of vertebrate prey is not uncommon within the genus Pelophylax (Ruchin and Ryzhov, 2002; Covaciu–Marcov et al., 2005; Çiçek and Mermer, 2006), this is only the second time an incident of saurophagy is reported (Nicolaou et al., 2014). Mediterranean insular herpetofaunas are

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known for extreme feeding behaviors in response to low food resources (Castilla et al., 2009; Brock et al., 2014; Cooper et al., 2015). Our finding may echo such behaviors but it is probably of an accidental nature. We stress here that our results refer only to spring. Frog diet is known to change through the seasons, depending on food availability and microhabitat use (Das, 1996; Covaciu–Marcov et al., 2005). To assess the annual diet of the species, similar work should be carried out even during the unfavorable periods of the year that were mentioned earlier. The dramatic changes in the Miocene eastern Mediterranean and their impact on Pelophylax phylogenetic history have been well studied (Lymberakis et al., 2007; Akin et al., 2010; Plötner et al., 2010; Poulakakis et al., 2015). Though the feeding ecology of the mainland Pelophylax frogs is also well studied (Çiçek and Mermer, 2006, 2007; Sas et al., 2007; Mollov, 2008; Sas et al., 2009; Mollov et al., 2010; Bogdan et al., 2012; Plitsi et al., 2016), in striking contrast there is an intense lack of general biology studies in the case of insular species and populations. To the best of our knowledge, this is the first systematic work on the diet of an island frog in the eastern Mediterranean. The significance of the dietary profile of species that requires protection has been highlighted in a comparative frame (Pope and Matthews, 2002; Fisher and Owens, 2004; Wiens et al., 2010). Our results stress the importance of beetles and spiders as primary food sources in the focal habitats. Measurements that will attract these two prey taxa (e.g. specific plants, suitable microhabitats) will ensure the smooth energy flow to the frogs. Fresh water is a rare commodity on Mediterranean islands, directly affecting frogs' distribution and future survival (Vervust et al., 2013). The conservation of these unique populations is a demanding task of the highest priority. Acknowledgements This work was fully supported by a grant from the Mohamed bin Zayed Species Conservation Fund (Project Number: 14259483). Sampling took place according to the Hellenic National Law (Presidential Decree 67/81) and under a special permit (code: 6ΨT04653Π8–ΥΟ3) issued by the Ministry of the Environment. References Akın, Ç., Bilgin, C. C., Beerli, P., Westaway, R., Ohst, T., Litvinchuk, S. N., Uzzell, T., Bilgin, M., Hotz, H., Guex, G.–D., Plötner, J., 2010. Phylogeographic patterns of genetic diversity in eastern Mediterranean water frogs have been determined by geological processes and climate change in the Late Cenozoic. Journal of Biogeography, 37: 2111–2124. Balint, N., Indrei, C., Ianc, R., Ursuţ, A., 2010. On the diet of the Pelophylax ridibundus (Anura, Ranidae) in Ţicleni, Romania. South Western Journal of Horticulture, Biology and Environment, 1: 57–66.

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Levins, R., 1968. Evolution in changing environments. Princeton University Press. Princeton, New Jersey. Lymberakis, P., Pafilis, P., Poulakakis, N., Sotiropoulos, K., Valakos, E. D., 2018. The Amphibians and Reptiles of the Aegean Sea. In: Biogeography and Biodiversity of the Aegean. In honour of Prof. Moysis Mylonas: 169–189 (S. Sfenthourakis, P. Pafilis, A. Parmakelis, N. Poulakakis, K. A.Triantis, Eds). Broken Hill Publishers Ltd, Nicosia, Cyprus. Lymberakis, P., Poulakakis, N., Manthalou, G., Tsigenopoulos, C. S., Magoulas, A., Mylonas, M., 2007. Mitochondrial phylogeography of Rana (Pelophylax) populations in the Eastern Mediterranean region. Molecular Phylogenetics and Evolution, 44: 115–125. Mollov, I. A., 2008. Sex based differences in the trophic niche of Pelophylax ridibundus (Pallas, 1771) (Amphibia: Anura) from Bulgaria. Acta Zoologica Bulgarica, 60: 277–284. Mollov, I. A., Boyadzhiev, P. S., Donev, A. D., 2006. A synopsis on the studies of the trophic spectrum of the Amphibians in Bulgaria. Scientific Studies of the University of Plovdiv–Biology, Animalia, 42: 115–131. – 2010. Trophic role of the marsh frog Pelophylax ridibundus (Pallas, 1771) (Amphibia, Anura) in the aquatic ecosystems. Bulgarian Journal of Agricultural Science, 16: 298–306. Nicolaou, H., Zogaris, S., Pafilis, P., 2014. Frog vs. lizard: An unusual feeding behaviour in the Levantine Marsh Frog, Pelophylax bedriagae from Cyprus. North–Western Journal of Zoology, 10: 221–222. Olsen, A. M., 2017. Feeding ecology is the primary driver of beak shape diversification in waterfowl. Functional Ecology, 31: 1985–1995. Pagani–Núñez, E., Ruiz, Í., Quesada, J., Negro, J. J., Senar, J. C., 2011. The diet of Great Tit Parus major nestlings in a Mediterranean Iberian forest: the important role of spiders. Animal Biodiversity and Conservation, 34: 355–361. Palazón, S., Ruiz–Olmo, J., Gosálbez, J., 2008. Autumn–winter diet of three carnivores, European mink (Mustela lutreola), Eurasian otter (Lutra lutra) and small–spotted genet (Genetta genetta), in northern Spain. Animal Biodiversity and Conservation, 31: 37–43. Pianka, E. R., 1973. The structure of lizard communities. Annual Review of Ecology and Systematics, 4: 53–74. Pimm, S. L., Jenkins, C. N., Abell, R., Brooks, T. M., Gittleman, J. L., Joppa, L. N., Raven, P. H., Roberts, C. M., Sexton, J. O., 2014. The biodiversity of species and their rates of extinction, distribution, and protection. Science, 344: 1246752. Plitsi, P., Koumaki, M., Bei, V., Pafilis, P., Polymeni, R. M., 2016. Feeding ecology of the Balkan Water frog (Pelophylax kurtmuelleri) in Greece with emphasis on habitat effect. North–Western Journal of Zoology, 12: 292–298. Plötner, J., Uzzell, T., Beerli, P., Akın, Ç., Bilgin, C. C., Haefeli, C., Ohst, T., Köhler, F., Schreiber, R., Guex, G.–D., Litvinchuk, S. N., Westaway, R., Reyer, H.–U., Pruvost, N., Hotz, H., 2010. Genetic divergence and

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Ontogeny of feeding by Astyanax paris in streams of the Uruguay River Basin, Brazil L. W. Cavalheiro, C. B. Fialho

Cavalheiro, L. W., Fialho, C. B., 2019. Ontogeny of feeding by Astyanax paris in streams of the Uruguay River Basin, Brazil. Animal Biodiversity and Conservation, 42.1: 9–18, https://doi.org/10.32800/abc.2019.42.0009 Abstract Ontogeny of feeding by Astyanax paris in streams of the Uruguay River Basin, Brazil. We studied the ontogeny of feeding by Astyanax paris, an insectivorous characid fish from streams of the Uruguay River Basin in southern Brazil. Size–based differences in diet composition were evaluated using permutational multivariate analysis of variance (PERMANOVA). Six streams surveyed over twelve months yielded twenty specimens for analysis of stomach contents. Smaller individuals (SL ≤ 25 mm) consumed mainly aquatic insects. As body size increased, there was a gradual shift to a diet dominated by terrestrial insects. Ontogeny of feeding habitats thus changes the species’ position in stream food webs. Key words: Characiformes, Diet, Insectivorous, Freshwater fish Resumen Ontogenia de la alimentación de Astyanax paris en los arroyos de la cuenca fluvial del río Uruguay, en Brasil. Estudiamos la ontogenia de la alimentación de Astyanax paris, un pez carácido insectívoro de los arroyos de la cuenca del río Uruguay, en el sureste de Brasil. Las diferencias de la composición de la dieta en función del tamaño se analizaron mediante el análisis de varianza multivariante con permutaciones (PERMANOVA). Se muestrearon seis arroyos durante 12 meses y se obtuvieron 20 especímenes para analizar el contenido del estómago. Los individuos más pequeños (longitud estándar ≤ 25 mm) consumieron principalmente insectos acuáticos. A medida que aumentaba el tamaño corporal, se pasaba gradualmente a una dieta compuesta principalmente por insectos terrestres. En consecuencia, la ontogenia de los hábitats de alimentación cambia la posición de las especies en las redes tróficas de los arroyos. Palabras clave: Caraciformes, Dieta, Insectívoro, Peces de agua dulce Received: 18 XII 17; Conditional acceptance: 05 III 18; Final acceptance: 10 IV 18 Laísa Wociechoski Cavalheiro, Clarice Bernhardt Fialho, Departamento de Zoologia, Universidade Federal do Rio Grande do Sul, Av. Bento Gonçalves 9500, Prédio 43435, CEP 91501–970, Porto Alegre, RS, Brazil. Corresponding author: Laísa Wociechoski Cavalheiro. E–mail: isa_woci@hotmail.com

ISSN: 1578–665 X eISSN: 2014–928 X

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Introduction Fish are important components of stream food webs and have been shown to influence ecosystem dynamics (Vanni, 2010; Rodríguez–Lozano et al., 2016). Studies of fish feeding habits have shown how trophic strategies can affect intraspecific and interspecific interactions and energy flux through ecosystems (Pompanon et al., 2012; Pendleton et al., 2014). The differential use of food resources is a well– known intraspecific strategy to avoid trophic niche overlap between juveniles and adults in several freshwater fish species (Bonato and Fialho, 2014; Cavalheiro and Fialho, 2016; Dala–Corte et al., 2016). Some species may therefore concentrate on different food resources at different stages of their development (Rudolf and Lafferty, 2011). The shift from soft–bodied, small prey to large and difficult–to– swallow prey is a common pattern among fish and reduces intraspecific competition for food resources (Russo et al., 2014). Ontogenetic niche shifts are the key to the functional variation among the life history stages in a population (Rudolf and Rasmussen, 2013). In practice, a species’ taxonomic identity alone is not sufficient to a priori predict its ecological interactions. Additional information on its biology should also be collected because fish are known for their high phenotypic plasticity in life–history traits, including body shape and trophic morphology in response to different food types (Kerschbaumer et al., 2011; Rudolf et al., 2014; Karjalainen et al., 2016). In this context, studies of diet variation in relation to modifications in body size are essential not only to characterize species as generalists or specialists, but also to identify their trophic strategies. According to Rudolf and Lafferty (2011), major challenges in studying trophic nets lie in determining the different functional roles within species and integrating such information into their trophic identity. Studies addressing the ecological relationships below the species level are therefore necessary to better understand natural communities. Ontogenetic development in fish affects morphological structures associated with feeding, thus allowing different sized individuals to consume different sized prey. Larger individuals often consume larger prey to maximize energy consumption (Keppeler et al., 2014). The trophic strategy of shifting diet composition according to ontogenetic changes allows smaller, less competitive fish to explore several food resources until they reach a size where they can compete with larger individuals of the same, or other, species (Russo et al., 2014). The genus Astyanax, the most speciose of the family Characidae, currently contains 158 valid species distributed in rivers from southern USA to Argentina, including the Uruguay River basin (Lima et al., 2003; Eschmeyer et al., 2016). In the Uruguay River basin, there are 13 valid species of the genus Astyanax, including A. paris Azpelicueta, Almirón and Casciotta, 2002 (Lucena et al., 2013b). This species was originally described from Fortaleza and Yabotí– Miní streams, both tributaries of the Yabotí–Guazú

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River, Upper Uruguay River in Misiones province, Argentina (Azpelicueta et al., 2002). Astyanax paris was considered endemic to locality (Lima et al., 2003; López et al., 2003; Liotta, 2005) until it was recorded in Upper Uruguay in the Brazilian state of Santa Catarina (Bertaco et al., 2016). The taxonomy and distribution of Astyanax species have been relatively well studied in the Uruguay River basin (Azpelicueta and Garcia, 2000; Bertaco and Malabarba, 2001; Azpelicueta et al., 2002; Casciotta et al., 2003; Bertaco and Lucena, 2010; Lucena et al., 2013a, 2013b; Bertaco et al., 2016). However, no information on the diet or any ecological data of A. paris isavailable in the literature so far. This paper thus increases understanding of the species’ biology and ecology in relation to how different age classes consume different food resources. Other species of Astyanax are considered generalists, such as A. aff. fasciatus (Cuvier, 1819), A. eigenmanniorum (Cope, 1894), A. lacustris (Lütken, 1875) and A. intermedius Eigenmann, 1908 in Tibagi River basin, Brazil (Bennemann et al., 2005) and A. lacustris in Maquiné River, Brazil (Vilella et al., 2002). Omnivory has been reported for A. bifasciatus Garavello and Sampaio, 2010, A. dissimilis Garavello and Sampaio, 2010 (Neves et al., 2015), and A. lacustri in Iguaçú River basin, Brazil (Cassemiro et al., 2002). Astyanax eigenmanniorum has been considered herbivorous in Lago del Fuerte Dam, Argentina (Grosman, 1999). This study aimed to investigate the feeding habits of A. paris from streams of the Ijuí River sub–basin in the state of Rio Grande do Sul. Hypothesis tested: A. paris follows the pattern of other freshwater neotropical fishes, modifying prey according to the sequences of life cycle states (ontogeny). Material and methods Study area The Uruguay River drains an area of about 365,000 km² and extends 1,838 km from the Serra Geral in southern Brazil to La Plata River estuary in Uruguay/Argentina (Di Persia and Neiff, 1986; Cappato and Yanosky, 2009; Bertaco et al., 2016). The basin can be divided into upper, middle, and lower courses (Bertaco et al., 2016). The river’s main tributaries are the Negro (Uruguay/Brazil), Quaraí (Uruguay/Brazil), Ibicuí (Brazil), and Ijuí (Brazil) Rivers (Carvalho and Reis, 2009). The Ijuí River is a tributary of the upper portion of the Uruguay River basin in the north–northwestern state of Rio Grande do Sul. It has a drainage area of 10,649.13 km² extending over 20 municipalities. Surveys were carried out at six streams along Ijuí River, from near its headwaters to near its confluence with the Uruguay River (Três Negrinhos: 28.432277778 ºS, 53.970805556 ºW; Nock: 28.316222222 ºS, 53.904972222 ºW; Santa Bárbara: 28.201722222 ºS, 54.218583333 ºW; Ibicuá: 28.394861111 ºS, 54.451638889 ºW; Araçá: 28.228 ºS, 54.956888889 ºW; Lajeado Grande: 28.170416667 ºS, 55.065944444 ºW).

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Surveys We carried out surveys at each point bimonthly over one year, from July 2015 to May 2016. Fish were captured by the electric fishing technique along a 100 m stretch in each stream, with a sampling effort of one hour per site. In the field, individuals were anaesthetized and euthanized with 10 % eugenol (Chair et al., 2014) and fixed with 10 % formalin. In the laboratory, specimens were identified according to taxonomic literature (Azpelicueta et al., 2002; Lucena et al., 2013b). Specimens were measured (Standard length, SL in mm) and dissected for diet examination. Voucher specimens were deposited in the fish collection of the Departamento de Zoologia, Universidade Federal do Rio Grande do Sul, Porto Alegre, Brazil (vouchers: UFRGS 21927, 21928, 21929). Fieldwork and sampling were carried out under a scientific collection permit (Permit Number 48291–1) issued by the Instituto Chico Mendes de Conservação da Biodiversidade, Ministério do Meio Ambiente, Brasília–Federal District, Brazil. Stomach contents were analyzed under a dissecting microscope and identified according to the standard taxonomic references (McCafferty, 1998; Mugnai et al., 2010; Segura et al., 2011). Food items were quantified by the volumetric method (VO %) (Hynes, 1950), associated with the frequency of occurrence (FO %) (Hyslop, 1980). Data analysis Changes in the diet according to the sampling site and intraspecific ontogenetic influences on the diet composition of A. paris were tested with permutational multivariate analysis of variance (permanova; α < 0.05) (Anderson, 2001), based on a Bray–Curtis dissimilarity matrix (Borcard et al., 2011). The Bray– Curtis index was used to construct the dissimilarity matrix as it considers data both of presence/absence and of abundance (Borcard et al., 2011). To assess possible ontogenetic variations, specimens were arbitrarily divided into three body size categories: small (SL ≤ 25 mm; n = 10), medium (SL 25 to 51.5 mm; n = 4), and large (SL ≥ 51.5 mm; n = 6). These three categories were determined according to the grouping in small, medium and large fish, of standard length classes defined by the Sturges' rule using the equation: h = (X – x) / (1 + 3.222 * log (n)) where: h = class interval; X = maximum standard length; x = minimum standard length; n = number of individuals (Sturges, 1926). A canonical analysis of principal coordinates (CAP) was used to compare diet composition in relation to standard length size classes (Legendre and Anderson, 1999). This method of ordering was chosen because of the possibility to apply a distance matrix between objects (the Bray– Curtis metrics), because this enables assessment of the relationships between the principal coordinates

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(dietary data) and variables (size categories) by redundancy analysis (RDA), and because it performs a permutation test that does not depend on the usual assumptions of data normality (Legendre and Anderson, 1999). The analysis of variance (ANOVA) with permutation tests was used to test the significance of the ordering analysis and its respective axis and terms (α < 0.05) (Legendre and Anderson, 1999). The indicator value index (IndVal) with randomization (Borcard et al., 2011) was used to determine whether any food items were associated with particular body size categories of A. paris. IndVal compares abundances and relative frequencies of food items in the diet of the studied groups (Cardoso et al., 2013). The statistical significance of such associations of food items and body size categories is confirmed by a permutation test (De Caceres, 2013). The higher the IndVal (Stat), the higher the association between a given food item and a specific group (De Caceres, 2013). Components A (comp A) and B (comp B) in the test vary from 0–1, and respectively indicate the probability of a food item being restricted to a given group and the probability of all sampled stomachs of a given group containing that food item (De Caceres, 2013). Statistical tests were carried out using R Project for Statistical Computing software, version 3.4.1. PERMANOVA, CAP, and ANOVA analyses were implemented in the statistical package Vegan, version 2.4–5 (Oksanen et al., 2017), whereas IndVal test was conducted done in the package Indicspecies, version 1.7.6 (De Caceres and Legendre, 2009). Results Twenty specimens of A. paris were collected (13 at Lajeado Grande and seven at Ibicuá). They measured between 21.9 and 80.5 mm in standard length. The diet was composed of 20 food items, classified according to their characteristics, origins, and relevance (table 1). The species presented an insectivorous feeding habit with insects making up 94 % of the volume of items consumed. This pattern did not vary between the two sampled streams (PERMANOVA; F = 0.97, R² = 0.04, p = 0.51). This study reports the first occurrence of A. paris in Rio Grande do Sul, thus extending its geographical distribution to the Medium Uruguay River Basin (fig. 1). The species was captured at two of the six sampled streams, namely, Ibicuá (28.394861111 ºS, 54.451638889 ºW municipality of Vitória das Missões) and Lajeado Grande (28.170416667 ºS, 55.065944444 ºW municipality of Dezesseis de Novembro; fig. 2). Both these streams are 1–1.5 m deep and have strong currents. Ibicuá stream is narrower, with muddy dark water and small stones on bottom (fig. 2A). Lajeado Grande stream is the widest. It has clear water and a rocky bottom with flat slippery stones (fig. 2B). The diet of A. paris is affected by ontogeny (PERMANOVA; F = 3.68, R² = 0.30, p = 0.0007). There is a marked shift in the species’ diet as it grows. The stomach contents of small specimens (SL ≤ 25 mm) consisted of 69.78 % (VO) aquatic insects and 28.97 %

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Table 1. Indicator values (IndVal) of food items consumed by standard length size classes (SL) of Astyanax paris: small, SL ≤ 25 mm; medium, SL 25 to 51.5 mm; large, SL ≥ 51.5 mm. The components A (Comp A) and B (Comp B) in the test vary from 0–1, and respectively indicate the probability of a food item being restricted to a given group and the probability of all sampled stomachs of a given group containing that food item. The Stat (test statistic) is the association between a given food item and a specific group: * α < 0.05. Tabla 1. Valores del indicador (IndVal) de los alimentos consumidos por clase de longitud estándar (LE) de Astyanax paris: pequeña, LE ≤ 25 mm; mediana, LE 25 a 51,5 mm; y grande, LE ≥ 51,5 mm. Los componentes A (Comp A) y B (Comp B) de la prueba se sitúan entre 0 y 1, e indican, respectivamente, la probabilidad de que un alimento esté limitado a un grupo determinado y la probabilidad de que dicho alimento se encuentre en todos los estómagos analizados. La Stat (prueba estadística) es la asociación entre un alimento determinado y un grupo específico: * α < 0,05.

Food item Small Aquatic Diptera

Comp A

Comp B

1.00

0.10

Stat

p–value

0.32

1.00

Medium Aquatic Plecoptera

0.64

0.75

0.69

0.10

Plant fragments

0.81

0.50

0.64

0.07

Fragments of terrestrial insects

0.73

0.50

0.61

0.19

Aquatic Acarina

1.00

0.25

0.50

0.20

Large Terrestrial Hymenoptera

0.98

1.00

0.99

0.0001*

Terrestrial Coleoptera

0.93

0.33

0.56

0.14

Terrestrial Araneae

1.00

0.17

0.41

0.49

Terrestrial Lepidoptera larvae

1.00

0.17

0.41

0.49

Aquatic Odonata

1.00

0.17

0.41

0.49

Terrestrial Odonata

1.00

0.17

0.41

0.49

Terrestrial Orthoptera

1.00

0.17

0.41

0.49

Small and medium Aquatic Ephemeroptera

1.00

0.71

0.85

0.02*

Aquatic Diptera

1.00

0.36

0.60

0.33

Fragments of aquatic insects

1.00

0.36

0.60

0.46

Seeds

1.00

0.14 0.38 0.84

Small and large Terrestrial Hemiptera

1.00

0.25

0.50

0.75

Terrestrial Coleoptera

1.00

0.13

0.35

1.00

Medium and large Aquatic Trichoptera

1.00

0.30

0.55

0.29

Fish scales

1.00

0.20

0.45

0.47

(VO) of terrestrial insects. Medium–size specimens consumed 46.56 % (VO) of aquatic and 42.11 % (VO) of terrestrial insects. Large specimens mainly consumed (VO = 91.81 %) terrestrial insects, with aquatic insects representing only 2.7 % (VO) of the stomach contents. Other prey items were consumed only occasionally or in low numbers. Aquatic ticks (Acarina) were found

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in the stomach of a medium–size specimen only, corresponding to 0.4 % (VO) of prey consumed by ﬁsh in this group size. Similarly, terrestrial spiders (Araneae) were found in the stomach of only one large specimen. Plant items (fragments and seeds) were found in the stomach of one small specimen (VO = 1.25 %), in the stomach of two medium–size specimens

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56 ºW

55 ºW

54 ºW

53 ºW

27 ºS

26 ºS

50 ºW

29 ºS

55 ºW

2010

28 ºS

km 0

km 0 25 50 75

30 ºS

25 ºS

50 ºS 30 ºS 10 ºS 10 ºN

80 ºW 60 ºW 40 ºW

km 0

270

Latin America Argentine territory Brazilian territory Ijuí River sub–basin

Hydrography Type locality–holotype Type locality–partypes New record of Astyanax paris

Fig. 1. Records of Astyanax paris in Argentinean and Brazilian regions, with indication of the species’ type locality (Azpelicueta et al., 2002) and new records from Ijuí River sub–basin, Rio Grande do Sul. Fig. 1. Registros de Astyanax paris en las regiones de Argentina y Brasil, con indicación de la localidad tipo de la especie (Azpelicueta et al., 2002) y nuevos registros de la subcuenca del río Ijuí, en Rio Grande do Sul.

A B Fig. 2. Localities of collection of Astyanax paris in Ijuí River sub–basin, Rio Grande do Sul, Brazil. Ibicuá stream, municipality of Vitória das Missões (A), Lajeado Grande stream, municipality of Dezesseis de Novembro (B). Fig. 2. Localidades de recogida de Astyanax paris en la subcuenca del río Ijuí, en Rio Grande do Sul, Brasil. Arroyo Ibicuá, municipio de Vitória das Missões (A), arroyo Lajeado Grande, municipio de Dezesseis de Novembro (B).

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Cavalheiro and Fialho

CAP CAP 2: 2: 14.93 14.93 % %

2 Medium 1

Ti Aqi

–1

Large Small

Thy

Aqe 25 mm) SmallSmall (SL ≤(SL 25<mm) Medium (SLto 2551.1 to 51.1 mm) Medium (SL 25 mm) > 51.5 mm) LargeLarge (SL ≥SL 51.5 mm)

–2 –1

0 1 CAP 1: 85.07 %

Fig. 3. Canonical analysis of principal coordinates (CAP) of food composition of Astyanax paris from Ijuí River sub–basin, Rio Grande do Sul, Brazil: Ac, aquatic Acarina; Ar, terrestrial Araneae; Aqc, aquatic Coleoptera; Tc, terrestrial Coleoptera; Aqd, aquatic Diptera; Td, terrestrial Diptera; Aqe, aquatic Ephemeroptera; Th, terrestrial Hemiptera; Thy, terrestrial Hymenoptera; Tll, terrestrial Lepidoptera larvae; Aqo, aquatic Odonata; To, terrestrial Odonata; Tor, terrestrial Orthoptera; Aqp, aquatic Plecoptera; Aqt, aquatic Trichoptera; Aqi, fragments of aquatic insects; Ti, fragments of terrestrial insects; Sc, scales; Pf, plant fragments; Se, seeds. Fig. 3. Análisis canónico de las coordinadas principales (CAP) de la composición de la dieta de Astyanax paris de la subcuenca del río Ijuí, en Rio Grande do Sul, Brasil: Ac, ácaros acuáticos; Ar, individuos del orden Araneae terrestres; Aqc, coleópteros acuáticos; Tc, coleópteros terrestres; Aqd, dípteros acuáticos; Td, dípteros terrestres; Aqe, efemerópteros acuáticos; Th, hemípteros terrestres; Thy, himenópteros terrestres; Tll, larvas terrestres de lepidópteros; Aqo, odonatos acuáticos; To, odonatos terrestres; Tor, ortópteros terrestres; Aqp, plecópteros acuáticos; Aqt, tricópteros acuáticos; Aqi, fragmentos de insectos acuáticos; Ti, fragmentos de insectos terrestres; Sc, escamas; Pf, fragmentos vegetales; Se, semillas.

(VO = 9.31 %), and in that of one large specimen (VO = 3.60 %). Fish scales were found in the stomach of one medium–size specimen (VO = 2.02 %) and in one large (VO = 1.17 %) specimen. The Canonical Analysis of Principal coordinates (F = 3.58, p = 0.0009) showed a standard length segregation of populations, especially in relation to the first axis (F = 6.12, p = 0.0001). This is further evidence of the shift in species diet from aquatic to terrestrial insects, mainly Ephemeroptera and Hymenoptera (fig. 3). These prey items are indicators of the species diet, as shown by IndVal within the 20 food items identified in this study (table 1). Aquatic Ephemeroptera is an indicator of small and medium–size specimens (Stat = 0.84, p = 0.02). This prey item has both a probability to occur in most stomachs of (comp B from IndVal = 0.71), and was restricted to (comp A from IndVal = 1.00) small and medium–size specimens. Large specimens did

ABC_42-1_pp_9-18.indd 14

not consume Aquatic Ephemeroptera. In contrast, terrestrial Hymenoptera is a strong indicator of large specimens (Stat = 0.99, p = 0.0001). Hymenoptera was restricted to large size class and was also predated by all analyzed individuals (comp B from IndVal = 1.00) (table 1). Discussion The fact the of A. paris was caught in two capture events and occurred in two streams, despite a year of intense sampling, suggests it has a low population size throughout the Ijuí River sub–basin and that it is naturally rare or extremely difficult to capture. This hypothesis is supported by the lack of previous records of A. paris in Rio Grande do Sul even though the ichthyofauna from Brazil's portion of the Uruguay River Basin has been well studied (Bertaco et al., 2016).

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The original description of the species was based on a few specimens (one holotype and 15 paratypes), currently in the Museo De La Plata (MLP) and Muséum National d’Histoire Naturelle (MHNG) (vouchers MLP 9584, 9585 e 9586 e MHNG 2623.65) (Azpelicueta et al., 2002). The other records of A. paris from scientific collections also consist of few specimens. The Pontifícia Universidade Católica do Rio Grande do Sul (MCP) has six specimens collected in the state of Santa Catarina (voucher MCP 40063), two from São Domingos River in the municipality of Cunha Porã (26.8891658783 ºS, 53.180557251 ºW) and three from Uruguay River in São Joaquim (26.8891658783 ºS, 53.180557251 ºW) (Bertaco et al., 2016). The collection of the Núcleo de Pesquisa em Limnologia Ictiologia e Aquicultura (Nupélia) da Universidade Estadual de Maringá (UEM) has six other specimens from the same state (vouchers NUP 16279 e 16282), all collected in Rio das Contas, municipality of Bom Jardim da Serra (28.4933333 ºS, 49.7825 ºW). UEM also has five specimens labeled as 'A. aff. paris' (voucher NUP 16279) from Ijuí in Rio Grande do Sul (28.3016667 ºS, 53.8927778 ºW); however, their identification should be confirmed. The lack of changes in the feeding behavior of A. paris according to the sampling site, despite the marked environmental characteristics in the two streams where it was collected, indicates the species is a probable insectivorous specialist feeder. Insectivorous fish influence both aquatic and terrestrial environments and play an important ecological role in regulating populations of their prey (Knight et al., 2005; Wesner, 2012; Xiang et al., 2016). This research confirms the hypothesis that A. paris presents ontogenetic variation in the diet with specific prey items for each life cycle stage. The variation of food items according to the standard length shows that different age classes play different functional roles in the trophic dynamics of the species habitat. Trophic webs are often studied through a traditional approach wherein species are assigned to guilds or trophic groups, without considering ontogeny (Rudolf et al., 2014). This practice has often been adopted in the existing research on Astyanax species (Bennemann et al., 2005; Silva et al., 2014). However, ontogenetic niche shifts are known from 80 % of animal taxa (Werner, 1988; Hertz et al., 2016); furthermore, the main source of intraspecific diversity in ecosystems is the variation across ontogenetic stages and size of individuals (Rudolf and Rasmussen, 2013). These aspects may lead to intraspecific functional differences in the role of individuals within ecosystems and affect the structural dynamics of communities (Hertz et al., 2016). Astyanax paris moves from a diet of aquatic to terrestrial insects as it ages. Ontogenetic shifts in the diet are often correlated with ontogenetic shifts in micro–habitat use, or preference for prey of different sizes (Rudolf and Rasmussen, 2013). Small individuals of A. paris feed mainly on aquatic Ephemeroptera, which are smaller than terrestrial Hymenoptera, Hemiptera, Coleoptera, and Orthoptera. These four items, in respective order of importance, made up most of the diet in terms of volume of large fish, although only Hymenoptera was

ABC_42-1_pp_9-18.indd 15

found to be an indicator of this size category. The shifting from soft to hard and larger prey, which aremore difficult to catch, was observed in A. paris. Mobility and higher competition capacity have often been cited as the aspects of larger fish prey selectivity (Russo et al., 2014). From this perspective, the ontogenetic shifts in the diet of A. paris can be viewed as a trophic strategy to reduce intraspecific competition, as seen in other Neotropical freshwater fish (Bonato and Fialho, 2014; Cavalheiro and Fialho, 2016; Dala–Corte et al., 2016). Species that change their diet during growth and show 'specialist phases'" may well appear generalists at species level if size is not taken in account in the study of their diets. Furthermore, these species may behave as sequential specialists as they change their tropic niche during development and are hypersensitive to food resource loss and habitat degradation (Rudolf and Lafferty, 2011). Studies investigating the ontogenetic influences on diet of species inhabiting areas vulnerable to impacts are essential. Ibicuá and Lajeado Grande streams show poorly conserved riparian vegetation. This may affect A. paris, which relies on terrestrial food resources. Astyanax paris relies on varied food resources (terrestrial and aquatic) across its life–stages. Protection of its habitats should consider not only the environmental quality of the streams, but also the integrity of adjacent riparian vegetation. The importance of riparian vegetation to fish diet is well recognized and documented for the allochthonous feeder Astyanax species (Gomiero and Braga, 2003; Borba et al., 2008; Ferreira et al., 2012; Souza and Lima–Junior, 2013; Silva et al., 2014; Leite et al., 2015). Conserving the streams is also important for those autochthonous feeders (Cavalheiro and Fialho, 2016). In conclusion, A. paris has an insectivorous tendency and plays different roles in the stream trophic web during its life–history. It shows marked ontogenetic shifts in diet, changing its food source from aquatic to terrestrial insects as it grows. Acknowledgements The authors would like to thank Gilmar Nunes Cavalheiro, Maria Ivone Wociechoski Cavalheiro, Amanda A. S. Santos, Dario F. Fuster, José Vanderlei da Silva, Juliano Ferrer, Júnior A. Chuctaya, Laura M. Donin, Leomar B. Medeiros and Rafael Angrizani for help and companionship in fieldwork, and Priscilla C. Silva for help with species identification. The first author received a PhD scholarship from Coordenação de Aperfeiçoamento de Pessoal deNível Superior (CAPES). References Anderson, M. J., 2001. A new method for non– parametric multivariate analysis of variance. Austral Ecology, 26: 32–46, doi: 10.1111/j.1442– 9993.2001.01070.pp.x Azpelicueta, M. M., Almirón, A. E., Casciotta, J. R., 2002. Astyanax paris: a new species from the

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Río Uruguay basin of Argentina (Characiformes, Characidae). Copeia, 2002: 1052–1056. Azpelicueta, M. M., Garcia J. O., 2000. A new species of Astyanax (Characiformes, Characidae) from Uruguay river basin in Argentina, with remarks on hook presence in Characidae. Revue suisse de Zoologie, 107: 245–257. Bennemann, S. T., Gealh, A. M., Orsi1, M. L., Souza, L. M., 2005. Ocorrência e ecologia trófica de quatro espécies de Astyanax (Characidae) em diferentes rios da bacia do rio Tibagi, Paraná, Brasil. Iheringia. Série Zoologia, 95: 247–254, doi: dx.doi. org/10.1590/S0073–47212005000300004 Bertaco, V. A., Ferrer, J., Carvalho, F. R., Malabarba, L. R., 2016. Inventory of the freshwater fishes from a densely collected area in South America —a case study of the current knowledge of Neotropical fish diversity. Zootaxa, 4138: 401–440, doi: doi. org/10.11646/zootaxa.4138.3.1 Bertaco, V. A., Lucena, C. A. S., 2010. Redescription of Astyanax obscurus (Hensel, 1870) and A. laticeps (Cope, 1894) (Teleostei: Characidae): two valid freshwater species originally described from rivers of southern Brazil. Neotropical Ichthyology, 8: 7–20, doi: dx.doi.org/10.1590/S1679–62252010000100002 Bertaco, V. A., Malabarba, L. R., 2001. Description of two new species of Astyanax (Teleostei: Characidae) from headwater streams of Southern Brazil, with comments on the 'A. scabripinnis species complex'. Ichthyological Exploration of Freshwaters, 12: 221–234. Bonato, K. O., Fialho, C. B., 2014. Evidence of Niche Partitioning under Ontogenetic Influences among Three Morphologically Similar Siluriformes in Small Subtropical Streams. Plos One, 9:e110999, doi: dx.doi.org/10.1371/journal.pone.0110999 Borba, C. S., Fugi, R., Agostinho, A. A., Novakowski, G. C., 2008. Dieta de Astyanax asuncionensis (Characiformes, Characidae), em riachos da bacia do rio Cuiabá, Estado do Mato Grosso. Acta Scientiarum. Biological Sciences, 30: 39–45, doi: dx.doi.org/10.4025/actascibiolsci.v30i1.1442 Borcard, D., Gillet, F., Legendre, P., 2011. Numerical ecology with R. Springer, New York. Cappato, J., Yanosky, A., Eds., 2009. Uso Sostenible de Peces en la Cuenca del Plata: Evaluación Subregional del Estado de Amenaza, Argentina y Paraguay. UICN, Gland. Cardoso, P., Rigal, F., Fattorini, S., Terzopoulou, S., Borges, P. A. V., 2013. Integrating Landscape Disturbance and Indicator Species in Conservation Studies. Plos One, 8: e63294, doi: dx.doi. org/10.1371/journal.pone.0063294 Carvalho, T. P., Reis, R. E., 2009. Four new species of Hisonotus (Siluriformes: Loricariidae) from the upper rio Uruguay, southeastern South America, with a review of the genus in the rio Uruguay basin. Zootaxa, 2113: 1–40. Casciotta, J. R., Almirón, A. E., Azpelicueta, M. M., 2003. A new species of Astyanax from río Uruguay basin, Argentina (Characiformes: Characidae). Ichthyological Exploration of Freshwaters, 14: 329–334.

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Cassemiro, F. A. Z., Hahn, N. S., Fugi, R., 2002. Avaliação da dieta de Astyanax altiparanae Garutti, Britski, 2000 (Osteichthyes, Tetragonopterinae) antes e após a formação do reservatório de Salto Caxias, Estado do Paraná, Brasil. Acta Scientiarum. Biological Sciences, 24: 419–425. Cavalheiro, L. W., Fialho, C. B., 2016. Trophic strategy of Atlantirivulus riograndensis (Cyprinodontiformes: Rivulidae), a non–annual rivulid threatened by extinction, in a perennial environment, Brazil. Neotropical Ichthyology, 14:e150068, doi: dx.doi. org/10.1590/1982–0224–20150068 Chair, J. A. J., Bart Jr, H. L., Bowker, J. D., Bowser, P. R., MacMillan, J. R., Nickum, J. G., Rachlin, J. W., Rose, J. D., Sorensen, P. W., Warkentine, B. E., Whitledge, G. W., 2014. Guidelines for Use of Fishes in Research—Revised and Expanded. Fisheries, 39: 415–416, doi: dx.doi.org/10.1080/03 632415.2014.924408 Dala–Corte, R. B., Silva, E. R., Fialho, C. B., 2016. Diet–morphology relationship in the stream–dwelling characid Deuterodon stigmaturus (Gomes, 1947) (Characiformes: Characidae) is partially conditioned by ontogenetic development. Neotropical Ichthyology, 14: e150178, doi: dx.doi.org/10.1590/1982– 0224–20150178 De Caceres, M., 2013. The R Project for Statistical Computing, The Comprehensive R Archive Network, How to use the indicspecies package (ver. 1.7.1). https://cran.r–project.org/web/packages/indicspecies/vignettes/indicspeciesTutorial.pdf [Accessed on 20 August 2016]. De Caceres, M., Legendre, P., 2009. Associations between species and groups of sites: indices and statistical inference. Ecology. http://sites.google.com/ site/miqueldecaceres/ [Accessed 20 August 2016]. Di Persia, D. H., Neiff, J. J., 1986. The Uruguay River system. In: The ecology of river systems: 599–621 (B. R. Davies, K. F. Walker, Eds.). Dr. W. Junk Publisher, Dordrecht. Eschmeyer, W. N., Fricke, R., Van Der Laan, R., Eds., 2016. Catalog of fishes: genera, species, references. California Academy of Sciences, San Francisco. http://researcharchive.calacademy. org/research/ichthyology/catalog/fishcatmain.asp [Accessed 08 August 2016]. Ferreira, A., Gerhard, P., Cyrino, J. E. P., 2012. Diet of Astyanax paranae (Characidae) in streams with different riparian land covers in the Passa–Cinco River basin, southeastern Brazil. Iheringia. Série Zoologia, 102: 80–87, doi: dx.doi.org/10.1590/ S0073–47212012000100011 Gomiero, L. M., Braga, F. M. S., 2003. O lambari Astyanax altiparanae (Characidae) pode ser um dispersor de sementes? Acta Scientiarum. Biological Sciences, 25: 353–360. Grosman, F., 1999. Estrutura da comunidade de peixes da represa 'Lago del Fuerte' Tandil, Argentina. Acta Scientiarum. Biological Sciences, 21: 267–275. Hertz, E., Trudel, M., El–Sabaawi, R., Tucker, S., Dower, J. F., Beacham, T. D., Edwards, A. M., Mazumder, A., 2016. Hitting the moving target: modelling ontogenetic shifts with stable isotopes

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of southern Brazil (Characiformes: Charcidae). Zootaxa, 3700: 226–236, doi: dx.doi.org/10.11646/ zootaxa.3700.2.2 Lucena, C. A. S., Castro, J. B., Bertaco, V. A., 2013b. Three new species of Astyanax from river drainages of southern Brazil (Characiformes: Characidae). Neotropical Ichthyology, 11: 537–552, doi: dx.doi.org/10.1590/S1679–62252013000300007 McCafferty, W. P., 1998. Aquatic entomology: the fishermen’s and ecologist’s illustrated guide to insects and their relatives. Jones and Bartlett Publishers, Boston. Mugnai, R., Nessimian, J. L., Baptista, D. F., 2010. Manual de identificação de macroinvertebrados aquáticos do estado do Rio de Janeiro para atividades técnicas, de ensino e treinamento em programas de avaliação da qualidade ecológica dos ecossistemas lóticos. Technical Books, Rio de Janeiro. Neves, M. P., Delariva R. L., Wolff, L. L., 2015. Diet and ecomorphological relationships of an endemic, species–poor fish assemblage in a stream in the Iguaçu National Park. Neotropical Ichthyology, 13: 245–254. Oksanen, J., Blanchet, F. G., Friendly, M., Kindt, R., Legendre, P., McGlinn, D., Minchin, P. R., O’Hara, R. B., Simpson, G. L., Solymos, P., Stevens, M. H. H., Szoecs, E., Wagner, H., 2017. Vegan: Community Ecology Package. R package version 2.4–5, https:// CRAN.R–project.org/package=vegan [Accessed 20 March 2017]. Pendleton, R. M., Hoeinghaus, D. J., Gomes, L. C., Agostinho, A. A., 2014. Loss of rare fish species from tropical floodplain food webs affects community structure and ecosystem multifunctionality in a mesocosm experiment. Plos One, 9: e84568, doi: dx.doi.org/10.1371/journal.pone.0084568 Pompanon, F., Deagle, B. E., Symondson, W. O. C., Brown, D. S., Jarman, S. N., Taberlet, P., 2012. Who is eating what: diet assessment using next generation sequencing. Molecular Ecology, 21: 1931–1950, doi: dx.doi.org/10.1111/j.1365–294X.2011.05403.x Rodríguez–Lozano, P., Rieradevall, M., Prat, N., 2016. Top predator absence enhances leaf breakdown in an intermittent stream. Science of the Total Environment, 572: 1123–1131, doi: doi.org/10.1016/j. scitotenv.2016.08.021 Rudolf, V. H. W., Lafferty, K. D., 2011. Stage structure alters how complexity affects stability of ecological networks. Ecology Letters, 14: 75–79, doi: dx.doi. org/10.1111/j.1461–0248.2010.01558.x Rudolf, V. H. W., Rasmussen, N. L., 2013. Ontogenetic functional diversity: size structure of a keystone predator drives functioning of a complex ecosystem. Ecology, 94: 1046–1056, doi: dx.doi. org/10.1890/12–0378.1 Rudolf, V. H. W., Rasmussen, N. L., Dibble, C. J., Van Allen, B. G., 2014. Resolving the roles of body size and species identity in driving functional diversity. Proceedings of the Royal Society of London B: Biological sciences, 281: 20133203, doi: dx.doi. org/10.1098/rspb.2013.3203 Russo, T., Scardi, M., Cataudella, S., 2014. Applications of Self–Organizing Maps for Ecomorphological Investigations through Early Ontogeny of

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Fish. Plos One, 9:e86646, doi: dx.doi.org/10.1371/ journal.pone.0086646 Segura, M. O., Valente–Neto, F., Fonseca–Gessner, A. A., 2011. Chave de famílias de Coleoptera aquáticos (Insecta) do Estado de São Paulo, Brasil. Biota Neotropica, 11: 393–412. Silva, M. R., Fugi, R., Carniatto, N., Ganassin, M. J. M., 2014. Importance of allochthonous resources in the diet of Astyanax aff. fasciatus (Osteichthyes: Characidae) in streams: a longitudinal approach. Biota Neotropica, 14: 1–10, doi: dx.doi. org/10.1590/1676–06032014001613 Souza, R. G., Lima–Junior, S. E., 2013. Influence of environmental quality on the diet of Astyanax in a microbasin of central western Brazil. Acta Scientiarum. Biological Sciences, 35: 177–184, doi: dx.doi.org/10.4025/actascibiolsci.v35i2.15570 Sturges, A., 1926. The Choice of a Class Interval, Journal of the American Statistical Association, 21(153): 65–66, doi: 10.1080/01621459.1926.10502161

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Vanni, M. J., 2010. Preface: When and Where Do Fish Have Strong Effects on Stream Ecosystem Processes? American Fisheries Society Symposium, 73: 531–538. Vilella, F. S., Becker, F. G., Hartz, S. M., 2002. Diet of Astyanax species (Teleostei, Characidae) in an Atlantic Forest River in Southern Brazil. Brazilian Archives of Biology and Technology, 45: 223–232. Werner, E. E., 1988. Size, scaling and the evolution of complex life cycles. In: Size–structured populations: 60–81 (B. Ebenman, L. Persson, Eds.). Springer, New York. Wesner, J. S., 2012. Predator diversity effects cascade across an ecosystem boundary. Oikos, 121: 53–60, doi: dx.doi.org/10.1111/j.1600–0706.2011.19413.x Xiang, H., Zhang, Y., Richardson, J. S., 2016. Importance of Riparian Zone: Effects of Resource Availability at Land–water Interface. Riparian Ecology and Conservation, 3: 1–17, doi: dx.doi. org/10.1515/remc–2016–0001

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Molecular survey of Hepatozoon infection of Teira dugesii in the Azores D. Rund, V. Neves, P. Quillfeldt

Rund, D., Neves, V., Quillfeldt, P., 2019. Molecular survey of Hepatozoon infection of Teira dugesii in the Azores. Animal Biodiversity and Conservation, 42.1: 19–29, https://doi.org/10.32800/abc.2019.42.0019 Abstract Molecular survey of Hepatozoon infection of Teira dugesii in the Azores. Hemogregarine parasites are found in many vertebrates, being most prevalent in reptiles, with lizards being the second most common hosts after snakes. Hepatozoon is the most widespread of the four genera that parasitize reptiles by infecting red blood cells. The Hepatozoon lifecycle requires blood–sucking invertebrates as vectors, and vector abundance can determine the parasite prevalence. To compare parasite prevalence between a large island and an islet without standing water, we analysed blood samples of the Madeiran wall lizard, Teira dugesii, at Praia Islet and Graciosa Island in the Azores, Portugal. We found a comparatively low prevalence of Hepatozoon, belonging to a new genetic line. The prevalence of this new parasite on the larger Graciosa Island was eight times higher than that for Praia Islet, which has no standing water sources. Our results are in line with a generally higher prevalence of blood parasites in sites with higher vector abundance. Key words: Teira dugesii, Hepatozoon, Azores, Molecular analysis, Island size Resumen Estudio molecular de la infección de Teira dugesii por Hepatozoon en las Azores. Las hemogregarinas son parásitos que se encuentran en numerosos vertebrados, principalmente en reptiles, de los cuales el grupo más común son las serpientes seguidas de las lagartijas. Hepatozoon es el género más abundante de los cuatro géneros que parasitan reptiles infectando los eritrocitos. El ciclo vital de Hepatozoon necesita invertebrados hematófagos como vectores, cuya abundancia puede determinar la prevalencia de los parásitos. Para comparar la prevalencia de los parásitos entre una isla grande y un islote sin agua estancada, analizamos muestras de sangre de lagartija de Maderia, Teira dugesii, en el islote de Praia y en la isla Graciosa en las Azores, en Portugal. Encontramos una prevalencia comparativamente baja de Hepatozoon, perteneciente a una nueva línea genética. La prevalencia de este nuevo parásito en la isla Graciosa, de mayor tamaño, fue ocho veces superior a la del islote de Praia, que carece de fuentes de agua estancada. Nuestros resultados están en consonancia con la prevalencia generalmente más elevada de parásitos hemáticos en sitios con una mayor abundancia de vectores. Palabras clave: Teira dugesii, Hepatozoon, Azores, Análisis molecular, Tamaño de isla Received: 20 IV 17; Conditional acceptance: 12 IX 17; Final acceptance: 03 V 18 Dorothee Rund, Petra Quillfeldt, Department of Animal Ecology and Systematics, Justus Liebig University Giessen, Heinrich–Buff–Ring, 35390 Giessen, Germany.– Dorothee Rund, Neuer Weg 19a, 65614 Beselich, Germany.– Verónica Neves, MARE–IMAR (Inst. of Marine Research), CIBIO Research Center in Biodiversity and Genetic Resources, CIBIO–Azores, Department of Science and Technology, Azores University, 9901–862 Horta, Azores, Portugal. Corresponding author: Petra Quillfeldt. E–mail: Petra.Quillfeldt@bio.uni–giessen.de

ISSN: 1578–665 X eISSN: 2014–928 X

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Introduction Parasites are an important, yet often poorly known, component of biodiversity (Morrison, 2009) and parasite studies have been used to answer complex questions of host–specificity and coevolution. Parasites play an important role as a crucial factor in evolution. As they constantly co–evolve with their host, they have a direct impact on natural communities and play an important role in ecosystems (Dobson and Hudson, 1986). A meta–analysis of 38 experimental studies on the cost of parasites in wild populations showed a moderately negative impact (Watson, 2013). Parasitic infections constantly challenge the immune system, and the influence of parasites was shown to be at least as strong as the influence of predation (Watson, 2013). Other stressors can be poor nutritional conditions, extreme climatic conditions and reproductive efforts (Quillfeldt et al., 2004), which can lead to a higher susceptibility to infections and even to death (Shutler et al., 1999). At the same time, parasitic infections can have a negative impact on the nutritional status, and thus lead to decreased reproductive success due to diminished manifestation of sexual ornaments and weakened physical condition (Hamilton and Zuk, 1982; Read, 1991; Zuk, 1992). From an evolutionary perspective, parasites can act as a significant selective agent and contribute to the development and maintenance of rare host genotypes and speciation (Haldane, 1992). At the genetic level, parasites are a major driving factor for the development of the complex immune system of vertebrates (Hedrick, 1994) and may structure host genotypic polymorphism (Clarke, 1979) and thus determine genetic structure (Shykoff et al., 1991). Hemogregarines are a group of the phylum Apicomplexan, Adeleorina, which are intracellular parasites. The Apicomplexa are a poorly studied group, with only about 0.1 % of species described (Morrison, 2009). Hemogregarines are the most prevalent parasites in reptiles, with lizards being the second most commonly affected hosts after snakes (Smith, 1996). Currently, four genera of hemogregarines are known to parasite reptiles: Hepatozoon (Miller, 1908), Haemogregarina (Danilewsky, 1885), Karyolysus (Labbé, 1994) and Hemolivia (Smith , 1996; Petit et al., 1990; Smith and Desser, 1997; Telford, 2009). The genus Hepatozoon is the most common genus among aquatic and terrestrial reptiles and is widely distributed among all other groups of tetrapods (Telford, 2009). In 1908, Miller described the genus Hepatozoon infesting leucocytes in rats. The first record of Hepatozoon in reptiles was made by Hoare (Hoare, 1932), who described sporogony of Hepatozoon pettiti (Thiroux, 1910, 1913) in tsetse flies (Glossina palpalis) that feed on infected Nile crocodiles (Crocodylus niloticus). This group can be very common in reptiles. Hepatozoon spp., for example, was found in 100 water pythons (Liasis fuscus) and seven other species of snakes with 100 % prevalence (Ujvari et al., 2004). In general, our knowledge of hepatozoonosis in wild animals like reptiles is far less than our understanding in pets and livestock, and further studies are required. In short–lived species, such as lizards,

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hemogregarine infections are normally non–detrimental to host condition (Amo at al., 2005b). Over 300 species are described within the genus Hepatozoon, and the wide variety of morphological characteristics, life–histories and hosts is extensive (Smith, 1996). Hepatozoon has been found to be paraphyletic (Barta et al., 2012; Maia et al., 2014). Thus, it was suggested to divide the genus into two genera, as all species of Hepatozoon except those that infect carnivores form a monophyletic group (Smith and Desser, 1997). The present study was carried out in the Azores archipelago on an introduced species of lizard, the Madeiran wall lizard Teira dugesii (Milne–Edwards, 1829). The introduction of new species can introduce new pathogens, such as parasites, which could facilitate host switching to new naïve hosts (Gurevitch and Padilla, 2004). In Iberian water frogs Pelophylax perezi, which have been introduced in the Azores, one individual from São Miguel Island was found to be infected with Hepatozoon (Harris et al., 2013). Hepatozoon is commonly found in lizards and other reptiles in Portugal, Spain and the Maghreb region of North Africa (Maia et al., 2011, 2012, 2014), while Hepatozoonosis was not found in Podarcis sp. lizards In North America (Burke et al., 2007). Studies in birds indicate that the prevalence of parasites is usually higher in freshwater inland habitats than in marine coastal habitats (Mendes et al., 2005) and higher on larger islands than on smaller ones (Lindström et al., 2004) due to the presence of more vectors. Vector distribution has been proposed as a factor affecting habitat choice in shorebirds, allowing them to make lower investments in immunofunction when living in saline areas due to lower risk of contact with vector–borne transmitted pathogens (Piersma, 1997). Given that Hepatozoon sp. Is, on one hand, prevalent in reptiles worldwide and was already present in frogs in the Azores, and, on the other hand, that more studies are needed to answer questions regarding systematics, lifecycle and distribution, our aim was to determine the prevalence of Hepatozoon sp. in T. dugesii in the Azores. Specifically, we tested the following hypotheses: Hepatozoon sp. is present in Teira dugesii in the Azores considering that it is widespread and very common among reptiles in similar habitats. The prevalence of Hepatozoon infections is higher on the main island Graciosa than on the small islet of Praia because the likelihood of an infection on the former is higher due to the presence of freshwater bodies and thus more areas allowing vector reproduction and a larger Madeiran wall lizard population. Material and methods Study site The study took place in the archipelago of the Azores, Portugal, in the subtropical northern Atlantic on Graciosa (fig. 1, 39º 3' 5'' N 28º 0' 51'' W, surface area

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N 39 º 05' N

40 ºN Corvo Flores

39 ºN

Praia islet

50 m Praia islet

Graciosa 0

Graciosa

2 km

Baixo islet 28º 55' W

100 m

Sao Jorge Faial

Terceira Pico

38 ºN

Sao Miguel

Atlantic Ocean

37 ºN

Santa Maria

Azores

31 ºW

29 ºW

27 ºW

50 km

25 ºW

Fig. 1. Location of Graciosa and Praia Islet within the Azores archipelago (adapted from Bried and Neves, 2014). Fig. 1. Ubicación de la Graciosa y del islote de Praia en el archipiélago de las Azores (adaptado de Bried y Neves, 2014).

60.66 km²) and the nearby small Praia Islet (fig. 1, 'Ilhéu da Praia', surface area 0.12 km²; stands about 1 km offshore Graciosa). Praia islet is uninhabited and has no water sources. Graciosa Island, in contrast, has about 5,000 inhabitants, several farms and open water tanks for livestock, as well as natural water sources. Graciosa is classified by UNESCO as a Biosphere Reserve, and Praia Islet is a breeding site for several seabird species. Between 1995 and 1997, rabbits (the only mammals on Praia Islet, introduced about 50 years earlier; Bried and Neves, 2014) were completely eradicated to restore seabird habitat, to stop soil erosion through overgrazing, and to protect seabird nests from being destructed. After their eradication, native plants were reintroduced and successfully spread, and invasive exogenous plants were removed. Seabird breeding numbers have since increased (Bried et al., 2009). Today Praia Islet is a natural reserve and protected area, and entry is restricted. Madeiran wall lizard Teira dugesii The Madeiran wall lizard is a relatively new species in the archipelago of the Azores, Portugal (Malkmus, 1995). Here the subspecies Teira dugesii dugesii can be found, the sturdiest of all the subspecies. Other subspecies are Teira dugesii jogeri, a native inhabitant of the archipelago of Madeira (Madeira, Deserta Grande, Bugio, Porto Santo and some smaller islets, Brehm et al., 2003; Arnold et al., 2007) and Teira

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dugesii selvagensis (Bischoff et al., 1989) from the Selvagens Islands. Teira dugesii was introduced to the Azores around 1860 via vessels and, lacking competitors, it rapidly multiplied. It is the only established terrestrial reptile in the archipelago of the Azores (Malkmus, 1995). Teira dugesii can colonize different habitats from sea level to 1850 m above sea level but it becomes rarer at higher altitudes. In the Azores, they are almost exclusively found 4 km within the coastline and not at elevations higher than 220 m (Malkmus, 1995). They can inhabit open and rocky areas but also more heavily vegetated habitats. In the Azores, T. dugesii usually lives on lava fields, agricultural areas and gardens with stonewalls and villages in close proximity to the coast. In some places, this species can reach very high densities. Due to the temperate to sub–tropical climate in the region, this species does not hibernate in the Azores. They are territorial and males can show quite aggressive behavior, especially during the mating season (Glandt, 2010). Common threats for T. dugesii are cats, dogs, rats, birds of prey, and gulls. According to the IUCN, T. dugesii is listed as of Least Concern. Hepatozoon species and their life cycle Hemoparasites have a complex lifecycle with different life stages, and they require more than one host to complete it. All Hepatozoon species in general follow

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the same lifecycle, consisting of sexual gametogony and asexual sporogony inside an hematophagous invertebrate host and merogony followed by gametogony inside a vertebrate intermediate host. Most of the life stages are haploid; only the zygote is diploid. Environmental factors determine the later sex of parasite clones (Lucius and Loos–Frank, 2008). The first invasive stadium of Apicomplexans is the wormlike sporozoite, which penetrates host cells of different tissues, e.g. blood cells, to develop into a trophozoite. After maturing, asexual schizogony occurs and the daughter cells differentiate into merozoites. Normally, several schizogony cycles occur and then the merozoites develop into gametocytes, which later differentiate into gametes. The fusion of a microgamete and a macrogamete will result in a diploid zygote. During sporogony, the sporont accrues and further differentiates into sporoblasts. The sporoblasts then divide and build sporozoits (Lucius and Loos– Frank, 2008). The meront stage of Hepatozoon can be found in different tissues and organs, depending on the species. In reptiles infected by Hepatozoon species, the gamont stage can usually be found in the erythrocytes, in birds and mammals most often in leucocytes (Lucius and Loos–Frank, 2008; but see Merino et al., 2014). In contrast to most protozoic or bacterial pathogens that are passed on by vectors, Hepatozoon is not transmitted during blood sucking through the salivary glands of a hematophagous arthropod or annelid but by swallowing and ingesting of the oocyst–carrying arthropod by the intermediate (vertebrate) host. Definitive hosts are blood–sucking invertebrates (Smith, 1996). Most studies of life cycles of Hepatozoon species infecting reptiles have been conducted on mosquitoes of the genera Culex sp., Aedes sp. and Anopheles sp., but ticks, mites, heteroptera and leeches can also act as vectors (Smith, 1996; Telford, 2009). Another pathway of transmission would be predation of an infected vertebrate host by another vertebrate. Here, the cystic stadium from the prey tissue is consumed (Smith, 1996). One example would be the transmission of Hepatozoon domerguei: A lacertid lizard that was infected with oocysts throughconsumption of an infected mosquito developed cystic forms of the parasite and was predated by a snake in which H. domerguei developed meronts and gamonts (Landau et al., 1972). In the vertebrate host, Hepatozoon infections can lead to clinical inflammation and can have a negative effect on hemoglobin concentration. The immune response is dependent on the adaptation to a specific host, the localization in the tissue, and the host´s immune status (Jacobson, 2007). In general, there is still little information about the clinical effects on reptiles (Maia et al., 2011). Field work Field work took place from May to July 2016. Blood samples of the Madeiran wall lizards on Graciosa were collected on 18 June 2016. Lizards were caught

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by hand or in traps made from empty 5 liter plastic water containers with tomato juice and pieces of fresh or dried fruit used as bait. The traps were stabilized by stones and roof tiles. Individuals were measured, weighed, sexed according to coloration and the presence of femoral pores, and examined for ticks and mites. We detected only one mite on a single male lizard. The base of the tail was disinfected using ethanol. A blood sample was then drawn from the caudal (tail) vein (see Divers and Mader, 2005), with a sterile insulin syringe (0.33 x 12.7 mm, 29G), and a drop was transferred onto a Whatman FTA classic card. We took 33 blood samples on Praia Islet and 32 on Graciosa. We also made 48 blood smears from individuals that provided sufficient blood and that were also sampled on FTA cards for genetic analysis (24 at Praia and 24 at Graciosa). We used blood as the tissue for parasite detection because, as stated above, in reptiles, the gamont stage can usually be found in the erythrocytes (Lucius and Loos–Frank, 2008). After the procedure, the lizards were released at the capture site. Laboratory analyses In the laboratory, a 2 x 2 mm piece of the dried blood sample was cut out of the FTA classic card. The DNA was then isolated using an Ammonium–acetate protocol (adapted from Martínez et al., 2009). The final DNA–concentration of the sample was determined with a NanoDrop2000c UV–Vis Spectrophotometer (NanoDrop Technologies, Wilmington, USA). In total, we successfully isolated DNA from 60 blood samples, 30 from Praia Islet and 30 from Graciosa. The analyses with the NanoDrop2000c UV–Vis Spectrophotometer (NanoDrop Technologies, Wilmington, USA) confirmed the presence of nucleic acid (DNA) in all 60 of our isolated samples (from below 10 ng/µl up to 32.5 ng/µl). These 60 DNA samples were screened for parasite presence using the polymerase–chain–reaction (PCR). We used the primers HepF300 (5‘–GTTTCTGACCTATCAGCTTTCGACG–3‘) and Hep900 (5’– CAAATCTAAGAATTTCACCTCTGAC–3‘) that target a part of the 18S rDNA gene in Hepatozoon spp. (Ujvari et al., 2004). The primers were designed to amplify Hepatozoon parasites and thus adequate for this study, but they are also found to amplify other parasite species such as Eimeria and Sarcocystis (Harris et al., 2012). The PCR reactions were run in a 16 µl mixture containing 2.5 µl of template DNA, 8 µl Hot Star TaqQiagen Plus Master Mix Kit (stock conc. 2x), and 1 µl of each primer. The reactions were cycled at the following parameters using a peqSTAR 96Q thermal cycler (Peqlab): 95 °C for 5 min (polymerase activation), 35 cycles at 95 °C for 30 sec, 60 °C for 30 sec and 72 °C for 1 min, and a final extension at 72 °C for 10 min. We also run positive and negative PCR controls. The PCR–amplicons were visualized using QIAxel (Qiagen) high–resolution capillary gel electrophoresis. Five samples showing the strongest peaks during gel electrophoresis were Sanger sequenced by Seqlab– Microsynth (Göttingen, Germany).

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Table 1. Blood parasite species, GenBank accession number (G Num), host species, location, authors, and sequences obtained in this study and reference sequences for the phylogenetic relationships in figure 2 obtained from Genbank. Author's abbreviations: 1, this study; 2, Maia et al. (2011); 3, Maia et al. (2012); 4, Maia et al. (2014); 5, Tomé et al. (2014); 6, Tomé et al. (2016); 7, Tateno et al. (2013). Tabla 1. Especie de parásito hemático, número de muestra, especie hospedante, localización, autores y secuencias obtenidas en este estudio para las relaciones filogenéticas en la figura 2 obtenidas de GenBank. (Para las abreviaturas de los autores véase arriba). Blood parasite

G Num

Host species

Location

Author

Lizards Sample L_35 MH201396 Teira dugesii Azores: Graciosa 1 Sample L_56 MH201397 Teira dugesii Azores: Graciosa 1 Sample L_58 MH201399 Teira dugesii Azores: Graciosa 1 Sample L_61 MH201398 Teira dugesii Azores: Graciosa 1 Hepatozoon sp. HQ734792 Podarcis vaucheri Morocco: Lake Tislit 2 Hepatozoon sp. HQ734794 Podarcis vaucheri Morocco: Lake Tislit 2 Hepatozoon sp. HQ734799 Timon pater tangitanus Morocco: 15 km N of Azrou (Balcon d'Ito) 2 Hepatozoon sp. HQ734801 Timon pater tangitanus Morocco: Agoudal 2 Hepatozoon sp. JX531955 Podarcis bocagei Portugal: Viana do Castelo 3 Hepatozoon sp. JX531916 Podarcis hispanica Spain: Alba de Tormes 3 Hepatozoon sp. JX531920 Podarcis lilfordi Spain: Balearics Islands, Cabrera 3 Hepatozoon sp. JX531925 Podarcis bocagei Portugal: Viana do Castelo 3 Hepatozoon sp. JX531955 Podarcis bocagei Portugal: Viana do Castelo 3 Hepatozoon sp. JX531957 Podarcis hispanica Spain: Alba de Tormes 3 Hepatozoon sp. JX531958 Podarcis bocagei Portugal: Viana do Castelo 3 Hepatozoon sp. JX531973 Podarcis lilfordi Spain: Balearic Islands, Cabrera 3 Hepatozoon sp. KJ189415 Podarcis bocagei Portugal: Gêres 4 Hepatozoon sp. KJ189416 Podarcis bocagei Portugal: Gêres 4 Hepatozoon sp. KJ189417 Podarcis hispanica Portugal: Gêres 4 Hepatozoon sp. KJ189418 Podarcis hispanica Portugal: Gêres 4 Hepatozoon sp. KJ189421 Podarcis bocagei Portugal: Gêres 4 Hepatozoon sp. KJ189422 Podarcis hispanica Portugal: Gêres 4 Skinks Hepatozoon sp. HQ734796

Morocco: close to Ouazzane

Geckos Hepatozoon sp. KJ408510 Cerastes cerastes Morocco: Figuig Hepatozoon sp. KJ408527 Psammophis schokari Morocco: Kaar Es Souk Hepatozoon sp. KU680423 Tarentola angustimentalis Spain

5 5 6

Cats Hepatozoon felis AB771546 Prionailurus iriomotensis

Eumeces algeriensis

Data analyses Forward and reverse sequences were aligned in CLC Main Workbench 7.6.4. and checked for quality. Variable bases and conflicts in nucleotides were resolved when possible by visually inspecting the traces of the aligned forward and reverse sequences. One of five sequences was discarded due to low quality. The final

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Japan: Okinawa

four sequences (length 471 bp) were used in a Blast search to find matching sequences. Sequences with the highest similarity (97–98 %) were downloaded. These sequences (N = 22) and the results from our own samples (see table 1) and Hepatozoon felis (AB771546.1, Tateno et al., 2013) as outgroup were aligned in BioEdit using CrustalW multiple alignment and very long overhangs were removed. The resulting

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0.005

KJ189421.1 Podarcis bocagei, Portugal KJ189422.1 Podarcis hispanica, Portugal KJ189418.1 Podarcis hispanica, Portugal KJ189417.1 KJ189416.1 Podarcis bocagei, Portugal 65 KJ189415.1 Podarcis bocagei, Portugal JX531958.1 JX531957.1 Podarcis hispanica, Spain 44 JX531925.1 Podarcis bocagei, Portugal JX531916.1 Podarcis hispanica, Spain HQ734792.1 Podarcis vaucheri, Morocco 67 JX531920.1 JX531973.1 Podarcis lilfordi, Spain 32 KU680423.1 Tarentola angustimentalis, Spain 70 KJ408510.1 Cerastes cerastes, Morocco 71 KJ408527.1 Psamophis schokari, Morocco HQ734796.1 Eumeces algeriensis, Morocco 50 98 HQ734799.1 Timon pater tangitanus, Morocco 37 HQ734799.1 55 L35_contig L56_contig Teira dugesii, L61_contig Azores L58_contig 48 JX531955.1 Podarcis bocagei, Portugal JX531955.1 HQ734794.1 Podarcis vaucheri, Morocco AB771546.1 Prionailurus iriometensis, Japan

Fig. 2. Phylogenetic relationships for blood parasites found in Teira dugesii on Graciosa and Praia islet inferred with maximum–likelihood. Branch support is shown above the nodes. Table 1 shows details for the samples used (22 reference samples, four study samples, one outgroup). Fig. 2. Relaciones filogenéticas de los parásitos hemáticos encontrados en Teira dugesii en la Graciosa y el islote de Praia, inferidas con la máxima probabilidad. El soporte de las ramas se indica encima de los nodos. En la tabla 1 se muestra información detallada sobre las muestras utilizadas (22 muestras de referencia, cuatro muestras del estudio y un grupo externo).

number of variable positions was 23 out of 471 and the number of parsimony–informative positions was 18 out of 471. The overall mean p–distance was d = 0.012. A matrix of pairwise distances is provided as supplementary material (table 1S). A maximum– likelihood phylogenetic tree was constructed based on 1000 bootstrap replications by fitting the best model (T92, Tamura 3–parameter, with gamma distribution) in Mega 6.0 (fig. 2). To obtain images of the parasites and infected cells (erythrocytes), blood smears were stained with Giemsa stain and a monolayer of blood cells was scanned with a light microscope (125x, oil immersion, Primo Star ZEISS). To test if the prevalence of blood parasites increased with the size of the lizards, we carried out a GLM with length as dependent parameter and island, sex and PCR result as independent factors, in R 3.4.2. (R Core Team, 2017). Results One of 30 samples from Praia Islet (3.3 %) and eight of 30 samples from Graciosa (26.7 %) were PCR positive, with a significantly higher prevalence at the larger Graciosa Island (Fisher's Exact Test, P = 0.035).

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All infected lizards were female (Fisher's exact test, P > 0.001). The lizards sampled on Graciosa were about 5 mm smaller (females: 6.3 ± 0.4 cm, males: 7.2 ± 0.3 cm SD) than the lizards sampled on Praia (females: 7.1 ± 0.2 cm, males: 7.7 ± 0.4 cm SD). The length differences between the lizards on the two islands and between sexes were statistically significant (GLM with length as dependent parameter and island, sex and PCR result as independent factors; effect of island: F1, 56 = 23.6, P < 0.001, effect of sex: F1, 56 = 32.8, P < 0.001). However, the length was not different for individuals with positive or negative PCR results (GLM as above, F1, 56 = 1.0, P = 0.312). The PCR–products from five samples with the highest signals were selected for sequencing, one from Praia (L_35) and four from Graciosa (L_55, L_56, L_58, L_61). The Blast search revealed a 97–98 % similarity with already published Hepatozoon sequences from other lacertid lizards and reptiles from Portugal, Spain and Morocco (table 1). These 22 reference sequences and one of Hepatozoon felis as outgroup were used to generate a maximum likelihood phylogenetic tree (fig. 2). The Hepatozoon sequences from the present study formed one cluster and were closest to the sequence of a parasite from a lacertid lizard from Morocco, Timon pater tangitanus (Maia et al., 2011). Microscopic examination of the slides of the sequenced

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samples confirmed an infection with a blood parasite (i.e. presence of gamonts in erythrocytes) for three samples (L_35, L_56 and L_58, fig. 3), but no infection was observed for sample L_55, probably due to a very low intensity of infection.

Discussion The aim of the present study was to examine whether Hepatozoon sp. infect Teira dugesii in the Azores and to compare the prevalence of Hepatozoon infections on the main island Graciosa with the small islet of Praia. Using genetic and microscopic methods, we detected the presence of Hepatozoon sp. in Teira dugesii red blood cells at Graciosa and Praia Islet in the Azores. On the main island Graciosa we found a prevalence of Hepatozoon sp. of 26.7 % and of only 3.3 % on Praia Islet, supporting the hypothesis for a higher prevalence on the larger island. Microscopic examinations confirmed the infection of erythrocytes as found in other reptiles (Lucius and Loos–Frank, 2008). Few studies have characterized the apicomplexan parasites in reptiles at the molecular level, and the relationships of many of these protozoan species are unresolved (Morrison, 2009; Jirku et al., 2009). Phylogenetic analyses based on microscopy methods are intricate due to the scarcity of differential phenotypic traits, which qualifies molecular phylogenetics based on genetic data as the best method to shed more light on the subject (Morrison, 2009). All of the infected lizards exhibited female–colouration but a larger sample size and genetic sex determination would be required to confirm sex as it is possible that some sub–adult males had not yet completely developed the typical male color pattern. An equal rate of infection with haemogregarines was observed among males and females of the ocellated lizard (Timon lepidus) in Spain (Amo et al., 2005a), while other studies of blood parasites of lizards suggested that males had a higher prevalence (Olsson et al., 2000; Klukowski and Nelson, 2001), probably due to the immune suppressive effects of testosterone, at least during the reproductive period (Roberts et al., 2004). Pregnant females, on the other hand, need to use a great amount of energy and metabolites for the development of eggs, energy that cannot be used for defense against parasites (Amo et al., 2005a). During our fieldwork, we found burrows of lizard eggs and observed frequent territorial fights among the males, indicating that sample collection overlapped with the breeding season. The occurrence of Hepatozoon species varies significantly among lizard families, with the highest prevalence detected in lacertids, to which T. dugesii belongs (Maia et al., 2012). However, our study on Graciosa Island and Praia islet revealed a relatively low prevalence of Hepatozoon compared to the 90 to 100 % prevalence reported in Podarcis hispanica in the Iberian peninsula (Harris et al., 2012; Maia et al., 2012), and 70 % for Podarcis lilfordi in the Balearic Islands (Harris et al., 2012). The lower prevalence

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5 µm B

5 µm

Fig. 3. Erythrocytes of the lizard T. dugesii from the Azores infected with a Hepatozoon blood parasite in samples L_35 (A) and L_56 (B), indicated with an arrow. Fig. 3. Eritrocitos de la lagartija T. dugessi de las Azores infectados con un parásito hemático del género Hepatozoon en las muestras L_35 (A) y L_56 (B), indicados con una flecha.

in the Azores and the Balearic Islands when compared to the Iberian peninsula might be partly due to a lower diversity of vectors in the Atlantic islands. The diversity of Culicoides species, for example, is much lower in the Azores than in mainland Portugal (Ramilo et al., 2012). In Timon lepiduss, 72 % of all adults but no juveniles were positive for haemogregarines (Amo et al., 2005a), and prevalence or intensity of infection in adults did not differ between seasons or in relation to body condition. A positive correlation has been found between the intensity of Hepatozoon infection and body size and thus age in different short–lived lizard species (Maia et al., 2014), suggesting that the intensity of infection increases with longevity due to more encounters with parasites and reduced immunocompetence in older animals (Amo at al., 2005b; Palacios et al., 2011). In contrast to most protozoan

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or bacterial pathogens transmitted by vectors, Hepatozoon is not transmitted during blood–sucking but by the vertebrate host swallowing the oocyst–carrying arthropod (Smith, 1996). Several vector species have been identified as infested with Hepatozoon oocysts or sporocysts of reptilian origin: the tsetse fly Glossina palpalis (Chatton and Roubaud, 2013; Macfie, 2016), ticks Amblyomma dissimile (Ball et al., 1969) and Hyalomma cf. aegyptium (Paperna et al., 2002), mites Ophionyssus sp. (Shanavas and Ramachandran, 1990) and Hirstiella sp. (Lewis and Wagner, 1964), predatory bugs Triatoma arthurneiva (da Rocha, 1975) and Triatoma rubrovaria (Osimani, 1942) and the sandfly Lutzomyia vexator occidentis (Ayala, 1970). Although to our knowledge there are no studies on arthropods parasites in the Azores, several Culicoide species recently reported to the archipelago (Ramilo et al., 2012) represent plausible vectors. Lizards may acquire mites or ticks from their conspecifics when they share favorable places such as basking spots and refuges. They have been identified as vertebrate hosts of ixodid ticks (Dantas–Torres et al., 2008). T. dugesii has been found to host immature stages of Ixodes ricinus (De Sousa et al., 2012). The lizards that we examined during our field work showed virtually no ectoparasites, except for a mite on one male. Mosquitoes could also be vectors of Hepatozoon if ingested (e.g. shown for Schellackia: Lainson et al., 1976). On Praia Islet, where there are no freshwater reservoirs, virtually no mosquitoes were present, in contrast with Graciosa, which could be an explanation for the higher prevalence of hepatozoonosis. Superficial examinations of seabirds studied on Praia Islet revealed almost no ticks or mites, and only bird lice were frequently detected. On Madeira Island, T. dugesii was infested with ticks in one of the two studied areas (De Sousa et al., 2012). Recently, four distinct main lineages of Hepatozoon spp. were found in wall geckos of the genus Tarentola from European and African countries adjacent to the Mediterranean Sea (Tomé et al., 2016). Two of these lineages clustered closely together not only with those previously known from individuals of the genus Tarentola and other species of geckos, but also with those from other reptiles and from rodents. A higher abundance of rodents could thus be another explanation for the eight times higher prevalence of Hepatozoon infections on Graciosa than on Praia Islet. The only mammals known to be present on Praia were rabbits, but they were eradicated by 1997 (Bried and Neves, 2014), while rats, ferrets, lifestock, cats and dogs are present in Graciosa, and all these mammals can also be hosts for Hepatozoon (Baneth et al., 2003). Some studies have suggested that Hepatozoon sp. infections are not host–specific and the parasite has the ability to switch easily between different host species (Maia et al., 2011). Other studies, however, have found a narrow host–specificity for some Hepatozoon sp. regarding vertebrate hosts (Telford et al., 2001), and the definitive invertebrate host (Carreno et al., 1997). Therefore, further studies are needed regarding relationships between lizard hosts, the ar-

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thropod Hepatozoon sp. vectors and other potential vertebrate hosts. In conclusion, the findings from this research supported both our hypotheses. We found a previously undescribed genetic lineage of Hepatozoon sp. infecting Teira dugesii in the Azores and a higher prevalence of Hepatozoon infections on the main island Graciosa than on the small islet of Praia. Further studies in the Azores are necessary to determine the potential role of other vertebrate species in the circulation of Hepatozoon, such as rats, cats, dogs, birds of prey and seabirds. Acknowledgements We would like to thank Stefanie Klemm for help in the field and Pedro Raposo, Luís Miguel Pereira Aguiar, Joana Cunha Lourenço and the rest of the team at the Parque Natural da Graciosa for transport between Praia Islet and Graciosa, for technical support and introduction to the fieldwork. The study, including sampling and fieldwork, was approved by the Secretaria Regional da Agricultura e Ambiente under the fieldwork permit 27/2016/DRA. This study was supported by the Fundação para a Ciência e Tecnologia (FCT), through the strategic project UID/ MAR/04292/2013 granted to MARE and the grant awarded to V. C. Neves–SFRH/BPD/88914/2012. References Amo, L., Fargallo, J. A., Martinez–Padilla, J., Millán, J., López, P., Martin, J., 2005a. Prevalence and intensity of blood and intestinal parasites in a field population of a Mediterranean lizard, Lacerta lepida. Parasitology Research, 96: 413–417. Amo, L., López, P., Martín, J., 2005b. Prevalence and intensity of haemogregarine blood parasites and their mite vectors in the common wall lizard, Podarcis muralis. Parasitology Research, 96: 378–381. Arnold, E. N., Arribas, Ó., Carranza, S., 2007. Systematics of the Palaearctic and Oriental lizard tribe Lacertini (Squamata: Lacertidae: Lacertinae), with descriptions of eight new genera. Magnolia Press. Ayala, S. C., 1970. Hemogregarine from sandfly infecting both lizards and snakes. The Journal of Parasitology, 56(2): 387–388. Ball, G. H., Chao, J., Telford Jr, S. R., 1969. Hepatozoon fusifex sp. n., a hemogregarine from Boa constrictor producing marked morphological changes in infected erythrocytes. The Journal of Parasitology, 55(4): 800–813. Baneth, G., Mathew, J. S., Shkap, V., Macintire, D. K., Barta, J. R., Ewing, S. A., 2003. Canine hepatozoonosis: two disease syndromes caused by separate Hepatozoon spp. Trends in Parasitology, 19: 27–31. Barta, J. R., Ogedengbe J. D., Martin D. S., Smith, T. G., 2012. Phylogenetic position of the adeleorinid coccidia (Myzozoa, Apicomplexa, Coccidia, Eucoccidiorida, Adeleorina) inferred using 18S rDNA

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Tomé, B., Rato, C., Harris, D. J., Perera, A., 2016. High diversity of Hepatozoon spp. in geckos of the genus Tarentola. Journal of Parasitology, 102(4): 476–480. Ujvari, B., Madsen, T., Olsson, M., 2004. High prevalence of Hepatozoon spp. (Apicomplexa, Hepatozoidae) infection in water pythons (Liasis fuscus) from tropical Australia. Journal of Parasitology, 90(3): 670–672. Watson, M. J., 2013. What drives population–level effects of parasites? Meta–analysis meets life– history. International Journal for Parasitology: Parasites and Wildlife, 2: 190–196. Zuk, M., 1992. The role of parasites in sexual selection: current evidence and future directions. Advances in the Study of Behavior, 21: 39–68.

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Comparing the predatory impact of captive–bred and free–living yellow spotted mountain newt (Neurergus microspilotus) on the larval green toad (Bufotes variabilis) T. Salehi, M. Sharifi Salehi, T., Sharifi, M., 2019. Comparing the predatory impact of captive–bred and free–living yellow spotted mountain newt (Neurergus microspilotus) on the larval green toad (Bufotes variabilis). Animal Biodiversity and Conservation, 42.1: 31–37, https://doi.org/10.32800/abc.2019.42.0031 Abstract Comparing the predatory impact of captive–bred and free–living yellow spotted mountain newt (Neurergus microspilotus) on the larval green toad (Bufotes variabilis). Captive breeding of endangered species is an important conservation tool, but it is not clear how long–term captive breeding can influence fitness attributes such as predatory ability. We experimentally investigated the predatory impact of adult captive–bred newts (CBN) and adult free–living newts (FLN) on the survival and growth of larval green toad (B. variabilis) in four predator density treatments containing none, one, two, or three newts. FLNs performed a rapid density–dependent predation, yielding average survival rates of tadpoles in no, low, medium, and high densities to 81 %, 74 %, 60 % and 17 %, respectively. CBNs had an average lower predation rate on B. variabilis tadpoles with a decrease in survival rate of tadpoles to 83 %, 81 %, 82 % and 77 % for 0, 1, 2 and 3 predator treatments, respectively. However, contrary to FLNs, they exhibited a significant increase in predation rate with time from 0.37 to 0.60 tadpoles per day. In addition, the growth rate of tadpoles reared with predators for the FLN group was significantly higher than the growth rate of tadpoles reared in control containers without the predator. In conclusion, our findings suggest that exposing captive–born adult yellow spotted mountain newts to their potential prey enriches the environment, and may be a useful approach in the development of more efficient captive breeding and reintroduction programs for this highly endangered amphibian. Key words: Captive breeding, Reintroduction, Predation, Endangered species, Tadpole, Conservation Resumen Una comparación de la capacidad de depredación de los individuos nacidos en cautividad y en libertad del tritón del Kurdistán (Neurergus micropillotus) sobre larvas de sapo verde (Bufotes variabilis). La cría en cautividad de especies amenazadas es un práctica importante en la conservación, pero no está claro si los programas de cría en cautividad a largo plazo pueden afectar a determinados atributos fundamentales de la eficacia biológica de las especies como la capacidad de depredación. En el presente estudio se estudian experimentalmente los efectos de la capacidad de depredación de tritones del Kurdistán adultos nacidos en cautividad (CBN) y de individuos obtenidos directamente de la naturaleza (FLN) en la supervivencia y el crecimiento de renacuajos de sapo verde mantenidos en cuatro tratamientos de densidad de depredadores, que contenían cero, uno, dos y tres tritones. El grupo de tritones FLN mostró una tasa de depredación rápida dependiente de la densidad de depredadores que conllevó que las tasas de supervivencia de los renacuajos en las densidades de control, baja, media y alta fueran, respectivamente, del 81 %, el 74 %, el 60 % y el 17 %. El grupo CBN tuvo una menor tasa media de depredación sobre los renacuajos de B. variabilis, cuya tasa de supervivencia disminuyó hasta el 83 %, el 81 %, el 82 % y el 77 % para los tratamientos con cero, uno, dos y tres tritones, respectivamente. Sin embargo, a diferencia del grupo FLN, esta tasa de depredación aumentó significativamente con el tiempo y pasó de 0,37 a 0,60 renacuajos por día. Por otra parte, la tasa de crecimiento de los renacuajos criados con depredadores del grupo FLN fue significativamente superior a la de los renacuajos criados en contenedores de control sin depredadores. Como conclusión, nuestros resultados sugieren que la exposición de individuos nacidos en cautividad del tritón del Kurdistán a sus presas potenciales enriquece el medio ambiente y puede ser útil para elaborar programas más eficientes de cría en cautividad y reintroducción de esta especie de anfibio muy amenazada. Palabras clave: Cría en cautividad, Reintroducción, Depredación, Especies amenazadas, Renacuajo, Conservación Received: 15 XI 17; Conditional acceptance: 12 II 18; Final acceptance: 22 V 18 Tayebe Salehi and Mozafar Sharifi, Department of Biology, Razi University, Baghabrisham 6714967346, Kermanshah, Iran. ISSN: 1578–665 X eISSN: 2014–928 X

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Introduction Captive breeding programmes have been used for a number of endangered species to save them from extinction, but the long–term outcome of these programmes has not always been satisfactory for a variety of reasons (Hedrick and Fredrickson, 2008). Some captive breeding programmes in amphibians have been characterized by poor nutritional requirements for growth (Pough, 2007), reduced natural behaviours such as foraging (including finding, identifying, acquiring, and handling food) (Burghardt, 2013), inactivity in the natural environment (Keulen and Janssens, 2017), inability to recognize natural foods (Olfert et al., 1993), a variety of diseases due to nutritional deficiency (Densmore and Green, 2007), chytridiomycosis infections (Parto et al., 2013), unsuccessful reproduction (Browne and Zippel, 2007), loss of social interactions (Rabin, 2003), variations in morphology of natural skin coloration (Ogilvy et al., 2012), loss of anti–predator response to predators (Kraaijeveld‐Smit et al., 2006), a reduced immune response (Keulen and Janssens, 2017) and losing genetic diversity due to inbreeding (Zippel et al., 2011). The potential survival of captive– born amphibians after their reintroduction into the wild is one of the concerns (Michaels et al., 2014). Although the above–mentioned studies cast doubts on the value of captive breeding and subsequent reintroduction as a conservation tool for threatened amphibian species (Griffiths and Pavajeau, 2008), there are always situations in which the captive breeding is the only conservation option available (Stuart et al., 2004). Harding et al. (2016) reviewed captive breeding programmes involving 213 amphibian species and found captive breeding affected various characteristics that could be taken into account in efforts to improve captivity management. Additionally, efforts to expand current understanding of ecology and behaviour of re–introduced species are growing (Pough, 2007). Moreover, research focusing on issues related to the release of captive–bred animals to the wild, such as acclimation to the new environments, pre–release health condition, genetic management, and long–term post–release monitoring, is greatly needed (Armstrong and Seddon, 2008). Questions regarding the importance of captive breeding and subsequent reintroduction of threatened species will continue to appear as more studies are carried out on these programmes (Armstrong and Seddon, 2008), and, especially in view of the practical irreversibility of many current threats to amphibians in their natural environments, which makes captive breeding and reintroduction indispensable conservation methods (Griffiths and Pavajeau, 2008; Harding et al., 2016). The yellow spotted mountain newt (Neurergus microspilotus) is included on the IUCN Red List as a critically endangered species for many reasons: its very small area of occupancy in its breeding streams (˂ 10 km), fragmented habitats, continuing decline in the quality and extent of aquatic habitats, habitat degradation, drought, and the pet trade (Sharifi et al., 2009). A conservation management plan funded by the Mohamed bin Zayed Species Conservation Fund

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was initiated for this species in 2010. Part of this plan included the development of a captive breeding facility at Razi University, Kermanshah, Iran (Vaissi and Sharifi, 2015). The project also launched several field and laboratory studies aiming to provide critical information for the management plan. These studies include delimiting the species range (Afroosheh et al., 2016), feeding habits (Farasat and Sharifi, 2014), activity pattern and home range (Sharifi and Afroosheh, 2014), reports on emergent diseases such as chytridiomycosis (Parto et al., 2013; Sharifi et al., 2014) and red–leg syndrome (Parto et al., 2014), and, finally, a first trial re–introduction of captive–born newts (Sharifi and Vaissi, 2014). In addition, several laboratory studies have provided information on ontogenetic changes in spot configuration (Vaissi et al., 2017), cannibalism (Vaissi and Sharifi, 2016a), and growth and development of N. microspilotus in the captive breeding facility (Vaissi and Sharifi, 2016a, 2016b). N. microspilotus in its natural environment acts as a top predator and feeds on a variety of prey, including benthic macroinvertebrates, amphibian eggs and tadpoles (Farasat and Sharifi, 2014), while the adult captive–born newts consume readily available food such as blood worm (Glycera dibranchiate), earthworm (Lumbricus terrestris) and chopped mealworm (Tenebrio molitor) ad libitum. The green toad, Bufotes variabilis, is a common toad in Iran with a wide distribution in most temperate areas, including parts of the distribution of the N. microspilotus in western Iran (Dastansara et al., 2017). We hypothesized that captive born N. microspilotus may lose their predatory ability to search, seize and engulf their natural prey (Burghardt, 2013). But we also postulated that captive born N. microspilotus could restore their predatory capability if they encountered natural prey. We therefore aimed to measure the predatory impact of captive–born adults at different densities and compare their impact with the free living newts under similar laboratory conditions in order: i) to determine whether there is a difference in the predatory impact of captive–born and free–living newts by examining the survival and growth rate of tadpoles of green toad (Bufotes variabilis) as prey, and ii) to measure whether captive newts can increase their predatory ability on tadpoles over time. Material and methods To experimentally assess the predatory ability of captive– born newts (CBN), we used eighteen adult individuals (6–year–old) with SVL (mean ± SD) of 64.37 ± 4.99 mm and body mass (mean ± SD) of 8.03 ± 1.40 g. The captive newt stock originated from two gravid females captured from Kavat stream (35º 21′ N, 46º 24′ E) on 5 February 2010. Experimental individuals were raised to maturity in aquaria boxes with electrical ventilators, waste pumps, and water temperature ranging from 15 to 21 ºC in summer and from 3 to 12 ºC in winter. Embryonic development (from oviposition to hatching) usually takes 28 d. Once hatched, larval newts were placed in separate aerated plastic containers (length x width x height: 22 × 16 × 12 cm) to

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prevent cannibalism, and they were fed by Artemia egg and shredded blood worms (Glycera dibranchiate). The larval period lasted six to eight months, until metamorphosis (loss of gills) was reached. Metamorphosed individuals were transferred to an aquarium (length x width x height: 49 × 21 × 16 cm) that contained small pebbles taken from their natural habitat to create approximately one third of their aquarium (300–350 cm2) as a terrestrial habitat. Recently metamorphosed individuals left the water, mostly occupying the terrestrial portion of their aquaria. They were fed ad libitum with gradually larger food items (blood worms, earthworm, and live mealworms). Adults were kept together, numbering six to nine newts, in a single aquarium (length x width x height: 75 × 25 × 40 cm) filled with mosses and some aquatic plants, with a sex ratio of two males to one female. Eighteen wild–caught adult individuals (FLN): SVL (60.79 ± 5.04 mm) and body mass (7.59 ± 1.30 g), were collected from the same population (Kavat stream) on 14 May 2016 and brought into the laboratory. They were kept, for one week before testing, in an aerated aquarium containing dechlorinated tap water and allowed to acclimatize in the laboratory, in similar conditions to CBN individuals. We collected three clutches of B. variabilis from Sarable wetland (34º 32' N, 47º 01' E), Kermanshah Province, Iran, in April 2016. Embryos were transferred to plastic containers until the experiments started, and they were allowed to develop and hatch in the laboratory in 12:12 light–dark cycles at approximately 20 ºC. Tadpoles were fed boiled spinach daily, and the water in all containers was changed weekly. The experiment was started when tadpoles were at the Gosner's developmental stages 25–26 (Gosner, 1960) and within the consumable size range for the predators. SVL and body mass at the start of the experiments were 1.33 ± 0.58 cm and 0.035 ± 0.004 g, n = 50. The surviving tadpoles and all eighteen FLN newts were released back to their original habitats after the experiments were finished. The study was conducted with the approval of the Razi University ethics committee under permit number: 3962022. Our experimental design tests for two sources of variation: predator origin, involving two levels: CBN and FLN, and predator density, with three levels: low (L), medium (M), high (H) and a control (1, 2, 3 and 0 newts, respectively). This design determines eight different predator origin and density combinations (including controls) and was replicated three times, for a total of twenty–four containers. Each container (49 × 21 × 16 cm) was filled with 550 ml of dechlorinated tap water. Fifty larvae of green toad, B. variabilis, were added in each experimental container. Tadpoles and predators were visually matched for size so that the initial differences in size would be minimal. We started the experiment on 21 May 2016 and randomly selected tadpoles and the newts to be used in the twenty–four experimental containers of predator origin and predator density treatment combinations: CBF–H, CBF–M, CBF–L, FLN–H, FLN– M, FLN–L and controls. For each container, we added four similar size stones to provide spatial heterogeneity. All containers were covered with mosquito cloth mesh to prevent newts from crawling out of the containers.

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Predators were weighed and measured prior to the start of the experiments. During the twenty–five days of the experiment, containers were checked daily, and the number of surviving tadpoles in each container was recorded. At each observation, dead and partly eaten tadpoles were removed from the containers and absent tadpoles were considered as preyed. Every four days, tadpole body length was measured in each container. Photographs were taken with a digital camera (Sony, DSCHX9V) on a tripod at a fixed height (30 cm). The tadpoles were put in a bowl which was located over a graph paper. Immediately after photography, the tadpoles were released into their containers. All pictures were loaded into Digimizer 4 software and the tadpole total length (TL) was measured by drawing a line from the tip of the snout to the tip of the tail. The containers were cleaned and water was renewed with dechlorinated tap water every eight days. During the experiment, tadpoles were fed boiled spinach every day ad libitum. All variables were checked for normality and homogeneity using the Kolmogorov–Smirnov normality test and Levene's test, respectively. We used two–way analysis of variance (ANOVA) to examine the effect of both predator origin and predator density on predation rate (the mean number of preyed tadpoles in each container during twenty–five days) and growth rate of tadpoles total body length (mm/day). Tukey's HSD post hoc pairwise comparisons were used to determine differences in predation rate between different newt densities. We compared the average number of preyed tadpoles in five–day periods for CBNs using repeated measure analysis of variance (repeated measure ANOVA) to determine whether predatory ability can improve with time. All statistical analyses were carried out using SPSS statistical software, version 16.0 (SPSS Inc, 2007). The statistical significance level was 0.05. Results The results of two–way ANOVA indicated a significant effect of newt origin and density and their interaction on the predation rate (table 1). In all densities of newts, the predation rates on tadpoles were higher in FLNs than in CBNs (table 2). High predator density greatly decreased tadpole survival, but this decrease was more drastic in FLN newts than in CBN newts, as revealed by the significant interaction term. Tukey's post hoc test showed that the highest predation rate was found for the FLN high density treatment. A repeated measure ANOVA at 5–day intervals showed that the predatory ability in CBNs, although lower than that of FLNs, increased with time (F4, 32 = 4.38, P = 0.006). FLN predation rate was maximal (18.11 ± 13.91 eaten tadpoles) in the first interval but did not increase later (fig. 1B), whereas the predatory ability of CBNs was low at the beginning but increased over time (3.66 ± 2.50 eaten tadpoles in the first interval to 7.33 ± 4.30 eaten tadpoles in the third interval) (fig. 1A). In addition, two–way ANOVA showed that the predator origin and predator density did not have a significant effect on the growth rate of the green toad, B. variabilis (table 1). However, there was a significant difference

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Table 1. Results of univariate ANOVA for assessing effects of predatory origin (captive– born vs free–living) and density of newts (3, 2, and 1) on predation rate and prey growth rate. Tabla 1. Resultados de la ANOVA univariada realizada para evaluar los efectos del origen del depredador (nacimiento en cautividad o en la naturaleza) y la densidad de los tritones (3, 2 y 1) en la tasa de depredación y la tasa de crecimiento de las presas. Source of variation F df P–value Predation rate Predator origin 49.24 1 0.000 Density 17.26 1 0.000 Predator origin × Density 11.46 2 0.002 Growth rate Predator origin 1.590 2 0.231 Density 0.294 2 0.751 Predator origin × Density 0.257 2 0.777

between the predator origin and the control on the growth rate of tadpoles (ANOVA, P–value = 0.037). Tukey’s post hoc test indicated that free–living predators caused a significant increase in the growth rate compared with the controls (P–value = 0.03). Discussion Our results are in agreement with our expectations. First, we found that newts raised in captivity had lower predation rates than the free–living newts. In addition, we observed that the predation rate in CBNs improved when they were exposed to the green toad B. variabilis tadpoles during the short experimental period (25 days). This finding suggests the predatory abilities of the captive stock newts increased. As in accordance with out study design we did not replace the consumed prey to keep the original prey densities constant, we could expect lower predation rates simply by reducing the chance of prey encounter by newts. However, the observation that CBNs newts increased their tadpole consumption rate is probably explained by increased capacities in prey recognition and hunting abilities over time. These results indicate that husbandry conditions of captive born newts can improve predatory ability by initiating prey–predator interaction in the breeding facilities. Under natural conditions, N. microspilotus is an active predator moving through the water column and ground feeding on macroinvertebrates, eggs and tadpoles (Farasat and Sharifi, 2014). As a result, active movements in aquatic environments increase the chances of capturing prey (Hossie and Murray, 2010). In our study, CBNs were fed

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Table 2. Mean ± SD for predation rate (%), and growth rate (mm/day) on fifty larvae of green toad, Bufotes variabilis under different predator origin (CBN and FLN) and predator density (High, Medium and Low) during the twenty–five days of the experiment. Tabla 2. Media ± DE de la tasa de depredación (%) y la tasa de crecimiento (mm/día) en 50 larvas de sapo verde, Bufotes variabilis, con depredadores de origen diferente (CBN y FLN) y distinta densidad de depredadores (alta, media y baja) durante los 25 días del experimento.

Treatment

Predation rate

Growth rate

FLN–H FLN–M FLN–L CBN–H CBN–M CBN–L Control

78.54 ± 11.13 40.29 ± 11.45 26.32 ± 8.38 22.77 ± 6.65 17.01 ± 6.40 18.10 ± 7.27 17.85 ± 7.02

0.42 0.37 0.50 0.35 0.33 0.33 0.24

± ± ± ± ± ± ±

0.33 0.10 0.10 0.05 0.06 0.09 0.04

several food items, such as earthworms, bloodworms and mealworms ad libitum, possibly meeting optimal nutritional need. However, such an approach may cause inactivity and obesity (McWilliams, 2008). Therefore, the initial assumption of predation rates being higher among FLN newts was demonstrated, probably because the prey capture efficiency and foraging was lower in the newts of captive origin. By comparing both captive–born and free–living N. microspilotus, we found both groups were similar in performing stereotype behavior of approaching, seizing and sudden engulfing tadpoles. However, the newts reared in the captive environment differed in their boldness and speed of their attack and showed less success in capturing the tadpoles than their free–living conspecifics (T. Salehi and M. Sharifi, pers. observ.).This may explain their constrasting predatory rates. Secondly, this study shows that exposing captive–born adult newts to a potential prey improve their predatory ability, suggesting that an environmental enrichment such as that used in this study, can be effective in developing more efficient captive breeding and reintroduction programs in N. microspilotus. Burghardt (2013) and Michaels et al. (2014) reviewed studies on how changes in different aspects of the environmental enrichment may improve the individual welfare of captive–born amphibian species. For instance, feeding enrichment in Dendrobates tinctorius, D. azureus, D. auratus, D. leucomelas, and Oophaga pumilio increases their activity levels and foraging behaviour (Campbell–Palmer et al., 2006). A study on Golden Mantella frogs (Mantella aurantiaca) showed that a diverse diet and a variety of live invertebrates, together with vitamins and mineral supplementations,

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40 30 20 10 0

40 30 20

3 newts

10 0

2 newts 0

15 Day

1 newt

Fig. 1. Average and standard deviation of the predation rate (total number of tadpoles consumed in 5–day intervals: by captive–bred (A) and free–living (B) Neurergus microspilotus exposed to different treatments (H, 3 newts; M, 2 newts; L, 1 newt) over the twenty–five days of the experiment. Fig. 1. Media y desviación estándar de la tasa de depredación (número total de renacuajos consumidos en intervalos de cinco días por individuos de Neurergus microspilotus criados en cautividad (A) y criados en libertad (B) expuestos a distintos tratamientos (H, 3 renacuajos; M, 2 renacuajos; L, 1 renacuajo) durante los 25 días del experimento).

can affect body condition and normal behaviour (Passos et al., 2017). Similarly, enclosures supplemented with spatial heterogeneity can increase activity levels in captive Lithobates catesbeianus (Rose et al., 2014). In addition, to hide or scatter food items throughout captive rearing enclosures can encourage natural foraging behaviour and predatory ability (Poole and Grow, 2012). Summarizing determining best feeding regimes for captive–born species can provide opportunities to mimic the situation where these species may be introduced into the wild (Keulen and Janssens, 2017). Interaction between predator origin and predator density influences tadpole predation rates. Predator density decreases prey survival in FLN newts but not in the captive reared newts. This will imply a much lower individual ingestion rate at higher densities in captive origin newts, which may have implications on net growth and breeding performance. A range of spatial organizations with increasing densities of amphibians and reptiles in the wild has been reported, but captive environment changes these organizations due to spatial constraints (Hayes et al., 1998). Thus, a lower ingestion rate in high density CBNs may help to choose lower raising densities in order to enhance growth rates and ultimately increase the viability of the captive stock.

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The effects of the density of predatory newts on anuran larvae as prey are not always predictable. Morin (1986) has shown that relative abundance of larval anurans may be positively or negatively altered by densities of predatory salamanders in pond communities. Morin (1983) showed that relative abundances of prey tadpoles of some species (Scaphiopus holbrooki, Rana sphenocephala, Bufo terrestris, and Hyla chrysoscelis) at high predator salamander densities (Notophthalmus viridescens dorsalis) were reduced, whereas the relative abundances of H. crucifer and H. gratiosa tadpoles in the presence of newts were not affected. Various explanations have been proposed for the differences found in the intensities of salamander predation on amphibian tadpoles (Morin, 1986). Individual predators may influence one another’s predation rate when foraging on prey (Ramos and Van Buskirk, 2012), or changes in prey behaviour (Charnov et al., 1976) which can reduce predator success at higher predator densities. Our results suggest that predatory ability on tadpoles increased with density in the FLN group, thus suggesting that individual predators did not affect each other. One possibility of larval salamander interference is cannibalism reported for larvae of N. microspilotus (Vaissi et al., 2017). However, neither cannibalism nor

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even apparent aggressive interactions occurred during our experiments. Overall, predator origin and density did not influence the growth rate of surviving tadpoles. However, by comparing the growth rate of tadpoles between control and predator treatments, we observed that predators induced higher and significant growth rates in FLN newts. Current evidence suggests that the presence and origin of predators have a great influence on prey growth rates (e.g. McPeek et al., 2001; Relyea, 2004; Van Buskirk, 2009). Prey responses to predators are widespread among larval amphibians. Rapid growth, for example, may reduce or even avoid larval predation by becoming too large for predators (Brodie and Formanowicz, 1983). However, other evidence has indicated that the presence of the non–lethal predator may induce reductions in growth rate (Van Buskirk, 2002). The increase in tadpole size in the presence of predators may be an adaptive strategy of prey known as the 'gap limited predator', a mechanism that reduces predation intensity when a predator cannot consume larger tadpoles. Another possible explanation for the higher growth rate in prey species is 'thinning', that is, an indirect effect of predators by removing competitors, leading to an increase in the per capita food resources for surviving prey (Abrams and Rowe, 1996). Although tadpole survival decreased at high FLN newt density, we did not find an expected increase in growth rates at high density, suggesting that no thinning mechanism is occurring. A possible explanation for the absence of competition in tadpoles under high densities of newt predators could be a strong reduction in activity by surviving tadpoles that will reduce their ingestion rates and, therefore, exhibiting a decline in their growth rates (e.g. McPeek et al., 2001). In conclusion, the significant increase in growth rate of FLN tadpoles suggests that N. microspilotus may be a 'gap limited predator' able to reduce predation intensity when they cannot consume large tadpoles. In conclusion, our results emphasize that exposing captive–born adult N. microspilotus newts to their potential prey can enrich the environment and may be a promising approach in the development of efficient captive breeding and reintroduction programs for this highly endangered amphibian. Acknowledgements We thank Razi University for its support of this study as a part of the PhD Research Project of T. Salehi (code number 19711). References Abrams, P. A., Rowe, L., 1996. The effects of predation on the age and size of maturity of prey. Evolution, 50: 1052–1061. Afroosheh, M., Akmali, V., Esmaelii, S., Sharif, M., 2016. On the distribution, abundance and conservation status of the endangered Spotted Mountain Newt Neurergus microspilotus (Caudata: Salamandridae) in Western Iran, Herpetological Conservation and Biology, 11: 52–60.

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Armstrong, D. P., Seddon P. J., 2008. Directions in reintroduction biology. Trends in Ecology and Evolution, 23: 20–25. Brodie Jr, E. D., Formanowicz Jr, D. R., 1983. Prey size preference of predators: differential vulnerability of larval anurans. Herpetologica, 39: 67–75. Browne, R. K., Zippel, K., 2007. Reproduction and larval rearing of amphibians. ILAR journal, 48(3): 214–234. Burghardt, G. M., 2013. Environmental enrichment and cognitive complexity in reptiles and amphibians: concepts, review, and implications for captive populations. Applied Animal Behaviour Science, 147(3): 286–298. Campbell–Palmer, R., Macdonald, W. C., Waran, N., 2006. The effect of feeding enrichment on the behavior of captive Dendrobatid frogs. Zoo Animal Nutrition, 3: 315. Charnov, E. L., Orians, G. H., Hyatt, K., 1976. Ecological implications of resource depression. American Naturalist, 110: 247–259. Dastansara, N., Vaissi, S., Mosavi, J., Sharifi, M., 2017. Impacts of temperature on growth, development and survival of larval Bufo (Pseudepidalea) viridis (Amphibia: Anura): implications of climate change. Zoology and Ecology, 27(3–4): 228–234. Densmore, C. L., Green, D. E., 2007. Diseases of amphibians. ILAR journal, 48: 235–254. Farasat, H., Sharifi, M., 2014. Food habit of the endangered yellow–spotted newt Neurergus microspilotus (Caudata, Salamandridae) in Kavat Stream, western Iran. Zoological Studies, 53(1): 61. Gosner, K. L., 1960. A simplified table for staging anuran embryos and larvae with notes on identification. Herpetologica, 16(3): 183–190. Griffiths, R. A., Pavajeau, L., 2008. Captive breeding, reintroduction, and the conservation of amphibians. Conservation Biology, 22(4): 852–861. Harding, G., Griffiths, R. A., Pavajeau, L., 2016. Developments in amphibian captive breeding and reintroduction programs. Conservation Biology, 30(2): 340–349. Hayes, M. P., Jennings, M. R., Mellen, J. D., 1998. Beyond mammals: environmental enrichment for amphibians and reptiles. In: Second Nature: Environmental Enrichment for Captive Animals: 205–235 (D. J. Shepherdson, J. D. Mellen, M. Hutchins, Eds.). Smithsonian Institution Press, Washington, D.C. Hedrick, P. W., Fredrickson, R. J., 2008. Captive breeding and the reintroduction of Mexican and red wolves. Molecular Ecology, 17(1): 344–350. Hossie, T. J., Murray, D. L., 2010. You can’t run but you can hide: refuge use in frog tadpoles elicits density–dependent predation by dragonfly larvae. Oecologia, 163(2): 395–404. Keulen, J., Janssens, G., 2017. Nutrition of captive amphibians. Master's Dissertation, Ghent University. Kraaijeveld–Smit, F. J. L., Griffiths, R. A., Moore, R. D., Beebee, T. J. C., 2006. Captive breeding and the fitness of reintroduced species: a test of the responses to predators in a threatened amphibian. Journal of Applied Ecology, 43: 360–365. McPeek, M. A., Grace, M., Richardson, J. M., 2001. Physiological and behavioral responses to preda-

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tors shape the growth/predation risk trade‐off in damselflies. Ecology, 82(6): 1535–1545. McWilliams, D. A., 2008. Nutrition Recommendations for some Captive Amphibian Species (Anura and Caudata). Canadian Association of Zoos and Aquariums Nutrition Advisory Research Group, http://www.caza-narg.ca/ref/amphibian%20nutrition%20report%20CAZA%202008.pdf [Accessed on September 2018]. Michaels, C. J., Downie, J. R., Campbell–Palmer, R., 2014. The importance of enrichment for advancing amphibian welfare and conservation goals. Amphibian Reptile Conservation, 8: 7–23. Morin, P. J., 1983. Predation, competition, and the composition of larval anuran guilds. Ecological Monographs, 53(2): 119–138. – 1986. Interactions between intraspecific competition and predation in an amphibian predator‐prey system. Ecology, 67(3): 713–720. Ogilvy. V., Preziosi, R. F., Fidgett, A. L., 2012. A brighter future for frogs? The influence of carotenoids on the health, development and reproductive success of the red–eyed tree frog. Animal Conservation, 16: 480–488. Olfert, E. D., Cross B. M., McWilliam, A. A. Eds., 1993. Guide to the care and use of experimental animals, Vol. 1: 51–74 and 115–124. Canadian Council on Animal Care, Ottawa. Parto, P., Siavash Haghighi, Z. M., Vaissi, S., Sharifi M., 2014. Microbiological and histological examinations in endangered Neurergus kaiseri tissues displaying Red–leg syndrome. Asian Herpetological Research, 5(3): 204–208. Parto, P., Vaissi, S., Farasat, H., Sharifi, M., 2013. First Report of Chytridiomycosis (Batrachochytrium dendrobatidis) in Endangered Neurergus microspilotus (Caudata: Salamandridae) in Western Iran. Global Veterinaria, 11(5): 547–551. Passos, L. F., Garcia, G., Young, R. J., 2017. The tonic immobility test: Do wild and captive golden mantella frogs (Mantella aurantiaca) have the same response? Plos One, 12(7): e0181972, Doi: 10.1371/journal.pone.0181972 Poole, V. A., Grow, S., 2012. Amphibian husbandry resource guide. Edition 2.0. Association of Zoos and Aquariums, Silver Spring, MD. Pough, F. H., 2007. Amphibian biology and husbandry. ILAR journal, 48(3): 203–213. Rabin, L. A., 2003. Maintaining behavioural diversity in captivity for conservation: natural behavior management. Animal Welfare, 12(1): 85–94. Ramos, O., Van Buskirk, J., 2012. Non–interactive multiple predator effects on tadpole survival. Oecologia, 169(2): 535–539. Relyea, R. A., 2004. Fine–tuned phenotypes: tadpole plasticity under 16 combinations of predators and competitors. Ecology, 85(1): 172–179. Rose, P., Evans, C., Coffin, R., Miller, R., Nash, S., 2014. Using student–centred research to evidence– base exhibition of reptiles and amphibians: three

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species–specific case studies. Journal of Zoo and Aquarium Research, 2(1): 25. Sharifi, M., Afroosheh, M., 2014. Studying migratory activity and home range of adult Neurergus microspilotus (Nesterov, 1916) in the Kavat Stream, western Iran, using photographic identification. Herpetozoa, 27(1/2): 77–82. Sharifi, M., Bafti, S. S., Papenfuss, T., Anderson, S., Kuzmin, S., Rastegar–Pouyani, N., 2009. Neurergus microspilotus. In: IUCN 2016: IUCN Red List of Threatened Species, version 2016. 3, www. iucnredlist.org [Accessed on 07 December 2016]. Sharifi, M., Farasat, H., Vaissi, S., Parto, P., Siavosh Haghighi, Z. M., 2014. Prevalence of the Amphibian Pathogen Batrachochytrium dendrobatidis in Endangered Neurergus microspilotus (Caudata: Salamandridae) in Kavat Stream, Western Iran. Global Veterinaria, 12(1): 45–52. Sharifi, M., Vaissi, S., 2014. Captive breeding and trial reintroduction of the endangered yellow–spotted mountain newt Neurergus microspilotus in western Iran. Endangered Species Research, 23: 159–166. SPSS Inc., 2007. SPSS for Windows, Version 16.0. Chicago, SPSS Inc. Stuart, S. N., Chanson, J. S., Cox, N. A., Young, B. E., Rodrigues, A. S., Fischman, D. L., Waller, R. W., 2004. Status and trends of amphibian declines and extinctions worldwide. Science, 306(5702): 1783–1786. Vaissi, S., Sharifi, M., 2015. Larval and Post–Metamorphic Growth in the Endangered Yellow Spotted Mountain Newt Neurergus microspilotus (Caudata, Salamandridae). World Journal of Zoology, 10(4): 365–373. – 2016a. Variation in food availability mediate the impact of density on cannibalism, growth, and survival in larval Yellow Spotted Newts (Neurergus microspilotus: implications for captive breeding programs. Zoo Biology, 35: 513–521. – 2016b. Changes in food availability mediate the effects of temperature on growth, metamorphosis and survival in endangered Yellow Spotted Mountain Newts: implications for captive breeding programs. Biologia, 71: 444–451. Vaissi, S., Parto, P., Sharifi, M., 2017. Ontogenetic changes in spot configuration (numbers, circularity, size and asymmetry) and lateral line in Neurergus microspilotus (Caudata: Salamandridae). Acta Zoologica, 99(1): 9–19. Van Buskirk, J., 2002. A comparative test of the adaptive plasticity hypothesis: relationships between habitat and phenotype in anuran larvae. The American Naturalist, 160(1): 87–102. – 2009. Natural variation in morphology of larval amphibians: phenotypic plasticity in nature? Ecological Monograph, 79: 681–705. Zippel, K., Johnson, K., Gagliardo, R., Gibson, R., McFadden, M., Browne, R., Martinez, C., Townsend, E., 2011. The Amphibian Ark: a global community for ex situ conservation of amphibians. Herpetological Conservation and Biology, 6(3): 340–352.

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Writing promotion and tenure evaluations for life scientists: thoughts on structure and content G. D. Grossman

Grossman, G. D., 2019. Writing promotion and tenure evaluations for life scientists: thoughts on structure and content. Animal Biodiversity and Conservation, 42.1: 39–43, https://doi.org/10.32800/abc.2019.42.0039 Abstract Writing promotion and tenure evaluations for life scientists: thoughts on structure and content. Experts in a discipline are frequently called upon to provide external evaluations of promotion and tenure (P/T) candidates at other universities and research agencies. Nonetheless, there is scant published information on the techniques and strategies used to produce a thorough and insightful P/T evaluation. External P/T evaluations must be matched to the candidate’s specific promotion criteria, and to their faculty appointment, both of which may vary substantially among institutions. A P/T evaluation letter should be based on independent and objective standards. Evaluation of the research dossier is typically easiest because of the many quantitative tools available (e.g., citation frequency, impact factors and h–index). Assessment of teaching performance should be based on comparisons of quantitative evaluation data compared to departmental means and standard deviations, although these metrics are not error–free. Written comments by students are also informative, as is evidence of successful mentoring of graduate students. Documentation of the successful use of innovative science pedagogy (e.g., flipped classrooms and active learning) also provides helpful evidence for evaluation of a teaching dossier. Finally, outreach faculty may be evaluated on the basis of: 1) publications, and 2) quantitative evaluations of workshops or other presentations. Key words: Promotion process, Tenure process, Academia, Reviewing colleagues Resumen La redacción de las evaluaciones de promoción y titularización para los biólogos: reflexiones sobre la forma y el contenido. Es frecuente que se pida a los especialistas de una determinada disciplina que realicen evaluaciones externas de los candidatos a promoción y titularización (P/T) en otras universidades y organismos de investigación. Sin embargo, la información publicada sobre las técnicas y estrategias utilizadas para realizar una evaluación exhaustiva y minuciosa es escasa. Las evaluaciones externas de la promoción y titularización deberán atenerse a los criterios específicos de promoción del candidato y a su puesto de docente, que pueden variar sustancialmente entre instituciones. La carta de promoción y titularización debería basarse en normas independientes y objetivas. La evaluación del expediente de investigación suele ser lo más fácil debido a los numerosos instrumentos cuantitativos disponibles (como la frecuencia de citas, los factores de impacto o el índice h). La evaluación del desempeño docente debería basarse en la comparación de los datos cuantitativos de la evaluación con los promedios y las desviaciones estándar de cada departamento, a pesar de que estos parámetros no están exentos de errores. Los comentarios por escrito de los estudiantes también aportan información, ya que son la prueba de la calidad de la mentoría de los estudiantes universitarios. La documentación de la utilización satisfactoria de métodos innovadores para la docencia de disciplinas científicas (aulas invertidas y aprendizaje activo) también aporta información útil para evaluar un expediente de docencia. Por último, la capacidad de promoción se puede evaluar en función de: 1) las publicaciones y 2) las evaluaciones cuantitativas de talleres y otras presentaciones. Palabras clave: Proceso de promoción, Proceso de titularización, Instituciones académicas, Examinar a compañeros Received: 08 V 18; Conditional acceptance: 26 VI 18; Final acceptance: 28 VI 18 Gary D. Grossman, Warnell School of Forestry and Natural Resources, University of Georgia, Athens, GA 30606 USA. ISSN: 1578–665 X eISSN: 2014–928 X

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Introduction The roles and responsibilities of professors in the 21st century have increased tremendously in the last 50Â years. Although several extant guides discuss how to be a productive faculty member, in general, these publications are directed towards new faculty (Boice, 2000; McGuckin and Ladhani, 2013; Wilson, 2013; Whitehead, 2016) rather than senior faculty. Besides the more obvious job responsibilities such as teaching and research, professional service is an obligation of all researchers and professors. Typically, this service falls disproportionately on senior faculty, and, not surprisingly, women and minority faculty (Guarino and Borden, 2017). One aspect of professional service that has received little attention in the published literature, especially in the life sciences, involves the strategies and techniques used to serve as an external evaluator for a candidate up for promotion/ tenure (henceforth P/T) at another university (see Goldman (2017) for a discussion of this topic with respect to law). Such evaluations represent one of the most important areas of professional service performed by faculty, and have consequences far beyond the acceptance or rejection of a manuscript or grant proposal. P/T letters play a critical role in determining job stability and current and future salary for faculty in the life sciences. In this essay I will explore the techniques and considerations that I have used in 39 external faculty/researcher evaluations for P/T. These evaluations include: 1) associate/full professors with teaching and research appointments, 2) associate professors with teaching or outreach appointments, and 3) research chairs and government scientist promotions, at institutions ranging from teaching universities to R1 research universities. Geographically, these institutions encompass locations in both the United States and Europe. My essay is based on generalities derived from these experiences, but I cannot overemphasize that any P/T evaluation must be based on the standards specific to the institution making the review request, and these may vary substantially among institutions. For example, institutional evaluation criteria may range from whether or not the candidate would be promoted and tenured at the evaluator's university, to simple reviews of the candidate's strengths and weaknesses. Because this essay is based solely on my own experience as a professor in the United States, it contains an element of subjectivity; hence, its greatest value will be as a platform for discussion among colleagues rather than as a canonical guide to the P/T evaluation process. So you have received an evaluation request A variety of factors should be considered before you agree to write a P/T evaluation. First, some faculty are concerned that writing a negative P/T letter will incur future repercussions. I know from personal experience that writing negative letters is not an enjoyable aspect of professional service, but it is a necessary one. Assuming you are familiar with the productivity

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norms of your field, one way to obtain an idea of the qualifications of a candidate is to request a summary of their accomplishments or a copy of their vita prior to agreeing to the evaluation. With this information in hand, you will be able to make a more informed decision regarding whether or not you are willing to undertake the review. Of course if everyone only writes positive P/T evaluations, then ultimately the quality of our field and the professoriate in general, will decline. Second, you should establish whether or not your evaluation letter will be kept confidential. Some public universities are bound by sunshine laws which permit the candidate to view their outside evaluation letters in an unredacted form, and this should be considered in your decisionâ&#x20AC;&#x201C;making process, especially if your field is a small one. Even when confidentiality is supposed to exist, you cannot assume that everyone privy to the decision will keep all information confidential, including the identity of external evaluators. This likely is more of a consideration for associate professors writing evaluation letters than for full professors. Finally, given the time necessary to read relevant materials and write a pithy but thorough letter, it is important to assess whether your schedule is sufficiently clear to provide a timely evaluation. Once you have agreed to serve as an outside evaluator, you should receive the candidate's P/T dossier from an administrator, but check to ensure it includes the promotion/tenure standards for the academic unit as well as the proportion of the candidate's position composed of research, teaching, and outreach/service (commonly abbreviated FTE [full time equivalent] or EFT). I have occasionally written evaluations for candidates working in units that had no specific requirements or guidance for P/T other than possession of a national or international reputation. How this is determined typically is left up to the evaluator and requires substantial thought; especially given that opinions on the matter may differ. The letter and what should be considered Unless your subdiscipline is small, and the host department composed of faculty in the same specialty, the professors reading your evaluation will likely be unfamiliar with you or your work. Consequently, I begin my letters with a short introduction of my academic record and background. Having a professional web site that includes a vita and publication list greatly aids this process, because it allows you to dispense with much descriptive text via a link to the site. I then describe my relationship to the candidate's field of research: 1) are we in the same specialty (animal community and population dynamics) or am I writing from a more general (ecologist) perspective? 2) are there portions of the candidate's research/teaching that are outside of my expertise and ability to judge? 3) finally, I describe my relationship with the candidate. Do I have a relationship with the candidate, either professional or personal, and how do I distinguish between these? Of course, if you truly have a personal relationship with the candidate then you should

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not be evaluating them for P/T unless your field is so small that everyone socializes with everyone else. Regardless of my relationship with the candidate, I describe whether or not we have met and under what circumstances (e.g. professional meeting, workshop, grant review panel, seminar reception). For example, here is a comment from an evaluation letter where I previously had a professional interaction with the candidate, 'Although I have met XXX and XXX has recently helped me with references to a chapter I have written, we do not have a personal relationship, and I believe I can evaluate their work objectively.' I consider myself to have a personal relationship with a candidate if we regularly socialize outside of our professional milieu, but do not consider a relationship personal if our socializing only occurs within a professional context (e.g., a large dinner at a meeting) and is occasional. Of course, opinions will vary on what does and does not constitute a personal relationship; consequently, I think it is important to disclose any relationship between a candidate and evaluator. I also include a statement regarding my ability to complete the evaluation in an unbiased manner. I begin my second paragraph with a statement that 'we all have too much to do and too little time to do it in' and then follow with my conclusion regarding the candidate's suitability for P/T, based on the criteria of the home unit. Here is an example, although for confidentiality I have modified the specific details, pronouns and promotion criteria: 'I fully support Dr. XXX promotion to Associate Professor with tenure at Anonymous State University. They would have no trouble being promoted to that rank in my home unit or, based on previous promotions, in Units Z, Y or Z here at University of Georgia. Their research work on the dynamics of planetary bodies and satellites is top notch and I believe will stand the test of time. In addition, their teaching and grant records are strong and indicate that their work is highly innovative. They clearly have exceeded your stated criteria for promotion to Associate Professor at Anonymous University.' I try to keep this paragraph short but thorough, and comment on all relevant criteria. I make this my second paragraph to ensure that members of the evaluation committee/faculty will be able to reach my conclusion quickly, rather than wading through what may be a dense multi–page letter. Frankly, I have witnessed more than one discussion in a P/T meeting where it was clear that faculty had not thoroughly read the outside evaluations, although structuring the letter in this manner also runs the risk of faculty not reading the entire evaluation. I begin the next few paragraphs by restating the P/T criteria of the candidate's home institution and describe my evaluation of their performance based on their specific appointment (e.g., 50 % research or 20 % research; 80 % teaching or 30 % teaching). I typically begin with something like 'Your guidelines for promotion and tenure state an individual must have achieved at least satisfactory performance ratings in all applicable areas (based upon appointment) of teaching, research/creative endeavors, and service, and an excellent rating in at least one.' If the candi-

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date has taken an unusual amount of time to apply for promotion, or there are gaps in their productivity, it may be worthwhile to enquire whether there are special circumstances that may warrant consideration such as illness or family issues, although some would argue this is inappropriate and all evaluations should be based on productivity alone. Let your professional conscience be the judge. The different portions of the P/T evaluation vary in their difficulty. In life sciences, the easiest portion to write is the research evaluation, because multiple quantitative tools are available for evaluation of a candidate's research and professional reputation. Examples of possible metrics are: 1) number of citations for a candidate's papers, which also yields the number of papers that have been highly cited (although I know of no objective definition for a highly cited paper in ecology, organismal biology or resource management, based on my experience I would consider a highly cited paper to have 50 or more citations although clearly date of publication must be considered too), 2) the number of papers in regional, national or international journals, 3) journal impact factors, although some disagree about their merit, and 4) research impact indices such as the h–index. Other possible criteria are: 1) invitations to speak at universities outside of the candidate's region, 2) regional or national awards, 3) service on agency, state or regional (e.g., Department of Natural Resources, Department of Transportation) or national (NSF, NIH or USDA) research or grant review panels, and 4) service on editorial boards for regional, national or international journals. If the criterion is an international reputation then one may use the same categories elevated to an international level, although it is likely that only a few faculty will have received international awards. Sources such as Web of Science and Google Scholar are useful tools for collecting these data and I prefer Google Scholar because it casts the widest net for citations. In addition, it is worthwhile to examine citations under both the candidate's full name and first initials, given that European journals frequently list authorship under the latter rather than the former. I have found that even Google Scholar repeatedly misses some of my highly cited papers if I do not also search under my first two initials. I also am quite wary of using online h–index calculators which are easy to use but rarely accurate, and if an h–index is needed I calculate it by hand using papers obtained via Google Scholar. This is much more time consuming but remember your evaluation will partially determine whether the candidate will keep or lose their job. Regardless of the availability of metrics to describe research productivity and quality, good P/T letters are always most effective when the reviewer can demonstrate a detailed and personal knowledge of the candidate's work. Consequently, I always try to read five to ten of the candidate's papers focusing on more recent work, in order to discuss the specifics of their research. A thorough evaluation will describe specifics of the candidate's research designs and results, as well as the implications of their findings for the field as a whole.

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Although it is easy to play the 'numbers game' (i.e., numbers of papers, numbers of citations and h–indices), an objective evaluation of a candidate's research requires thoughtful consideration and insight. Clearly, neither citations nor counts of publications alone provide a sufficient basis for an objective evaluation, especially in these days of 'pay to publish'. Although predatory journals with unethical publication policies and editorial board appointments exist, it should not be too difficult to detect publications in such journals via the many possible metrics discussed above (e.g., impact factors, number of citations). However, even restricting your evaluation to legitimate journals, one must acknowledge that publication rates vary greatly among fields and, of necessity, this must be considered in your evaluation. It always is wise, although not necessarily easy, to determine normative publication rates for fields outside your specialty. What may be even more problematical is ascertaining the candidate's actual contribution to their multi–authored publications, especially if not described in the narrative. There are many criteria used to determine authorship and author order in the life sciences, which we have recently reviewed (Grossman and DeVries, unpublished). Nonetheless, in my own work, I abide by the convention of earning either first authorship or last authorship, because these connote who did most of the work (first) and who provided the intellectual impetus for the work and provided funding (last). However, consider the case where a candidate has met the minimum requirement for research productivity but has no senior authorships. If their papers mainly represent the work of a graduate student senior author and a faculty member junior author, I would not be too concerned, given this is the typical pattern of authorship in my field for graduate student publications. But what if the candidate works in a field where large collaborative projects are the norm, has no senior authorships, and no papers with fewer than 10 coauthors; a case I have always found problematic? Here is an example of how I handle this situation 'The sole shortcoming that I find with XXX's research record is that most of XXX's papers have many authors and XXX typically is not the first. Consequently, it is difficult to judge XXX's specific contribution to the work. Nonetheless, I have given XXX the benefit of the doubt and assumed that XXX's contributions were significant.' Clearly each case may warrant a different approach and there certainly are no easy answers regarding evaluation of a candidate's publication record when many or most of the publications include many coauthors and the candidate generally is not the senior author. I will confess that for promotions to full professor or research chairs, I feel that demonstration of continued substantive scholarly performance, as evidenced by senior authorships on research papers, is required. Although a P/T letter should not be overly personal, the use of a personal 'touch' always makes it easier to read and also imparts authenticity to the author's words. Here are some statements from my own previous letters 'XXX's list of funded proposals reads

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Grossman

like a veritable Who's Who of funding sources and I wish I had as much success, especially recently!' Or 'Nonetheless, if you deem this letter inadequate, please contact me immediately. The last thing I would want to do is compromise the candidacy of such an excellent faculty member.' In conclusion, however, your letter still must retain the required objectivity of an outside evaluator. Typically, the research evaluation, at least in the United States, will require assessment of the candidate's history of obtaining competitive grants and contracts. Certainly in this day and age, proficiency at obtaining grants is a requirement for promotion in the United States, at least at most research oriented universities. The number and fiscal magnitude of grants will vary by department and institution; hence, there is no point in suggesting specific numbers and amounts. It is reasonable to assume that a candidate’s grant support should be sufficient to support graduate students or full–time technical help. Nonetheless, being proficient at obtaining grants does not necessarily demonstrate adequate scholarship for promotion. It is not uncommon to review candidates that have exceptional records at obtaining grants and contracts, but weak publication records. Conversely, you may encounter cases where the publication record is excellent but the grant record is mediocre. Again, the standards of the home institution must be followed when evaluating these cases. Nonetheless, my personal opinion is that someone who publishes in top notch journals without substantial grant support is more deserving of P/T than someone with a strong grant record but a poor to mediocre publication record. As I review previous P/T evaluation letters, I find that I have not generally been asked to evaluate a candidate’s proficiency in teaching, although summaries of teaching evaluations as well as teaching awards frequently have been included in promotion dossiers. Perhaps this is because I have only worked at a research (R1) institution. Quantitative teaching evaluations have their own set of biases including gender bias (Boring et al., 2016), although it is easy to compare them to departmental means and standard deviations. Student comments on evaluations also provide information, as do unsolicited letters of commendation from students. All of these metrics need to be considered carefully in terms of their usefulness for a tenure evaluation. There are a variety of additional criteria that are useful when evaluating a candidate’s teaching performance. For example, teaching narratives may contain information regarding the successful use of both standard and innovative tools for science pedagogy such as active or authentic learning, flipped classrooms, and new applications of educational technology such as music or karaoke videos (Grossman and Watson, 2015; Grossman and Simon, 2018). If the dossier contains little information on the candidate's teaching performance it will be useful to discuss any instances where you observed the candidate’s presentation skills such as during presentations at scientific meetings, or lectures, or from presentations at grant review panels, or simply acknowledge that the information

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in the dossier is insufficient for an evaluation. A final aspect of instructional responsibilities warrants mention, and that is the supervision and mentoring of graduate students. It is reasonable to expect that a candidate for promotion to associate professor at a research university should have an active graduate program, and should have graduated multiple students prior to being awarded promotion and tenure. For promotions to full professor at an R1 university, I am reluctant to write a positive letter for a candidate that has not graduated both multiple PhDs and Master's students. Again numbers vary too much from field to field to identify specific minima, but evidence for an active graduate program should be readily identifiable. On occasion you may be asked to evaluate faculty whose main appointment is service/outreach. For these evaluations, the main criteria I use are: 1) evidence of outreach publications, and 2) whether workshops, panels or presentations are evaluated via some quantitative instrument such as a Likert scale questionnaire. Again, the evaluation must be based on the institutional requirements for promotion. Publication may or may not be a requirement for these faculty but those who are publishing in refereed journals have clearly shown excellent performance. Outreach faculty who only publish technical bulletins may still meet P/T guidelines, and many extension service publications fall in some middle ground where they are sent out for external review but generally do not have a real chance of being rejected. Assessment of workshop productivity is often difficult because of a lack of quantitative assessment. One would hope that if outreach faculty are required to go through a P/T process that quantitative assessment data would be provided to an evaluator. External evaluation of P/T candidates is an important component of professional service for faculty, and helps to maintain a high quality professoriate. In this essay I have shared my own strategies, judgements and experiences in order to provide guidance to faculty unfamiliar with the P/T process and stimulate collegial discussion of these issues. I hope these suggestions are of use to professors young and old.

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Acknowledgements Funding for this work was provided by the Warnell School of Forestry and Natural Resources and USDA McIntire–Stennis project GEOZ–0176–MS. The manuscript was improved by the helpful comments of: D. Devries, K. Gido, B. Grossman, S. Heard, D. Orth, T. Simon and K. Woosnam. Any errors or omissions are mine alone. References Boice, R., 2000. Advice for new faculty members: nihil nimus. Allyn and Bacon, Boston MA. Boring, A., Ottoboni, K., Stark, P. B., 2016. Student evaluations of teaching (mostly) do not measure teaching effectiveness. ScienceOpen Research, 2016, doi: 10.14293/S2199–1006.1.SOR–EDU.AETBZC.v1 Goldman, E., 2017. Writing tenure–review letters. https://www.insidehighered.com/advice/2017/01/19/ advice–how–write–effective–tenure–review–letters [Accessed 23 April 2018]. Grossman, G. D., Simon, T., 2018. Student perceptions of an inquiry–based karaoke exercise for ecologically oriented classes: a multiclass evaluation. Journal of College Science Teaching, 47: 92–99. Grossman, G., Watson, E., 2015. The use of original music videos to teach natural history. The Journal of Natural History Education and Experience, 9: 1–7. Guarino, C. M., Borden, V. M. H., 2017. Faculty service loads and gender: are women taking care of the academic family? Research in Higher Education, 58: 672–694. McGuckin, D., Ladhani, M., 2013. Advice for new faculty: six lessons from the front lines. Https:// www.facultyfocus.com/articles/teaching–careers/ advice–for–new–faculty–six–lessons–from–the– front–lines/ [Accessed 23 April 2018]. Whitehead, K., 2016. Advice for young faculty. Http:// labwhitehead.blogspot.com/2016/03/my–advice– for–young–faculty.html [Accessed 21 April 2018]. Wilson, E. O., 2013. Letters to a Young Scientist. Liveright publishing, New York, NY.

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Animal Biodiversity and Conservation 42.1 (2019)

Distribution models of the Spanish argus and its food plant, the storksbill, suggest resilience to climate change A. Zarzo–Arias, H. Romo, J. C. Moreno, M. L. Munguira

Zarzo–Arias, A., Romo, H., Moreno, J. C., Munguira, M. L., 2019. Distribution models of the Spanish argus and its food plant, the storksbill, suggest resilience to climate change. Animal Biodiversity and Conservation, 42.1: 45–57, https://doi.org/10.32800/abc.2019.42.0045 Abstract Distribution models of the Spanish argus and its food plant, the storksbill, suggest resilience to climate change. Climate change is an important risk factor for the survival of butterflies and other species. In this study, we developed predictive models that show the potentially favourable areas for a lepidopteran endemic to the Iberian Peninsula, the Spanish argus (Aricia morronensis), and its larval food plants, the storksbill (genus Erodium). We used species distribution modelling software (MaxEnt) to perform the models in the present and in the future in two climatic scenarios based on climatic and topographic variables. The results show that climate change will not significantly affect A. morronensis distribution, and may even slightly favour its expansion. Some plants may undergo a small reduction in habitat favourability. However, it seems that the interaction between this butterfly and its food plants is unlikely to be significantly affected by climate change. Key words: Distribution models, Climate change, Interaction, Butterfly, Larval food plants, MaxEnt Resumen Los modelos de distribución de la morena española y las plantas nutricias de sus larvas sugieren resistencia frente al cambio climático. El cambio climático representa un importante factor de riesgo para la supervivencia de las mariposas y de otras especies. En este estudio se han elaborado modelos predictivos que muestran las zonas potencialmente favorables para un lepidóptero endémico de la península ibérica, la morena española (Aricia morronensis), y las plantas nutricias de sus larvas, los alfilerillos o agujas de pastor (género Erodium). Se ha utilizado el programa informático MaxEnt para elaborar modelos de distribución de las especies en el presente y en el futuro, bajo dos escenarios de condiciones climáticas, basados en variables climáticas y topográficas. Los resultados muestran que el cambio climático no afectará significativamente a la distribución de A. morronensis, sino que incluso podría favorecer levemente su expansión. Algunas de las plantas podrían sufrir una pequeña reducción de la favorabilidad del hábitat. Sin embargo, la interacción entre la mariposa y sus plantas nutricias probablemente no se vea afectada significativamente por el cambio climático. Palabras clave: Modelos de distribución, Cambio climático, Interacción, Mariposa, Plantas nutricias de las larvas, MaxEnt Received: 03 I 18; Conditional acceptance: 07 V 18; Final acceptance: 03 VII 18 Alejandra Zarzo–Arias, Unidad Mixta de Investigación en Biodiversidad (UO–CSIC–PA), Universidad de Oviedo, c/Gonzalo Gutiérrez Quirós, Edificio de Investigación, 5ª planta, Campus Mieres, ES–33600 Mieres, Spain.– Helena Romo, Juan Carlos Moreno, Miguel L. Munguira, Departamento de Biología, Universidad Autónoma de Madrid, c/Darwin 2, Cantoblanco, ES–28049 Madrid, Spain. Corresponding author: Alejandra Zarzo–Arias. E–mail: alejandra.zarzo@gmail.com

ISSN: 1578–665 X eISSN: 2014–928 X

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Introduction The effects of climate change are subject to much attention in conservation studies due to their influence on biodiversity, the changes produced on populations, communities and ecosystem dynamics, and the biotic interactions (Walther, 2010; Dawson et al., 2011; Giannini et al., 2013). One relevant effect is the temporal and spatial mismatch between life cycles and resource availability, particularly threatening for herbivore insects (Cornelissen, 2011; Bellard et al., 2012). These insects also appear to be negatively affected by climate change because of their sensitivity to changes in the environment, particularly temperature (Wilson and Maclean, 2011). Butterflies have proven to be good climate change indicators, and have commonly been used to assess its effects (Roy et al., 2001; Walther et al., 2002; Diamond et al., 2011). Settele et al. (2008) performed climate change models for all European butterflies of the superfamily Papilionoidea to assess the risk it could represent to the European butterfly species. Their results suggest a severe loss of climatically suitable habitat for most species. In this study we focused on the Spanish argus butterfly, Aricia morronensis (Ribbe, 1910) (Lycaenidae, Lepidoptera), an Iberian endemic species that can be found above 1,000 meters in the Peninsula, occupying most of the main mountain systems. Munguira and Martín (1988) stated that this species could have occupied lower altitude habitats, and that the postglacial rising of temperatures may have forced it to move to higher elevations as is currently occurring in several butterfly species in response to climate change (Hill et al., 2002; Konvica et al., 2003; Wilson et al., 2007). Nowadays, as A. morronensis only occupies high–altitude habitats, climate change can potentially represent a threat to this species (Dirnböck et al., 2011; Stefanescu et al., 2011; Lambers, 2015). It was listed as endangered by De Viedma and Gómez–Bustillo (1976), but was later considered out of danger by Munguira (1989), although its endemic character and the fact that it is not included in Settele's et al. (2008) atlas still raises interest from the conservation point of view. Furthermore, although most butterfly species are influenced by climate, other factors related to habitat quality and composition can determine their survival (Stefanescu et al., 2004; Brückmann et al., 2010; Krämer et al., 2012). Therefore, it seems of great importance to study how A. morronensis is able to cope with potential future shifts under global change in order to take the necessary conservation measures. Besides, recently performed genetic studies show that this species could be split into two different entities according to its distribution (Dincă et al., 2015). With this subdivision, some of the butterfly’s populations could be highly restricted and thus endangered. To increase effectiveness in predicting the evolution of these butterfly populations, in this study we considered the interaction between A. morronensis and its food plants, as suggested by Gilman et al. (2010), Romo et al. (2014) or Valiente–Banuet et al. (2015). The butterfly is a stenophagous species, and

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its larvae only feed on some perennial species of the plant genus Erodium L'Her. (storksbills, Geraniaceae). The species can survive feeding on any of the five species mentioned by Munguira and Martín (1988), but only one species is used in each location on which the species was found. Besides, in captivity larvae can survive using common annual Erodium species, but these species were never recorded in the field (Munguira, unpublished data). Therefore, its survival also depends on the future distribution of these plants, which may also be influenced by climate change, and on their interaction (Thuiller et al., 2005; Romo et al., 2014). According to systematic reviews conducted by Fiz–Palacios et al. (2010) and Alarcón et al. (2012), A. morronensis feeds on the following five Erodium species: E. carvifolium Boiss. and Reut, E. cazorlanum Heywood, E. daucoides Boiss, E. foetidum (L.) Rothm and E. glandulosum Dumort. E. cazorlanum is also interesting because it is an endemic plant catalogued as Vulnerable. The butterfly lays its eggs on the plant leaves during the summer flight period. Larvae feed on the leaves of the plant and overwinter at the third or fourth larval instar. Pupation takes place in the late spring and the pupal stage lasts 10 days on average (García–Barros et al., 2013). Distribution and habitat suitability models are useful tools in fields like ecology and biogeography (Guisan and Zimmermann, 2000; Elith et al., 2011; Titeux et al., 2016). They can provide information about the consequences of climate change on the species distribution (Elith et al., 2010). The effect of climate change has been studied for many butterfly species, showing a significant reduction on their distribution range that would eventually lead to future extinctions in most cases (Settele et al., 2008; Romo et al., 2015). These models can also serve as a basis for spatial planning and they can provide tools for an optimum conservation strategy (Kearney et al., 2010). When building potential distribution models, the main problem is the lack of information, attached to poor coverage of the territories due to insufficient sampling efforts (Ramos et al., 2001). As Romo and García–Barros (2005) concluded, the sampling effort on butterflies in the Iberian Peninsula showed a geographic bias, which supports the usefulness of potential distribution models in order to improve the knowledge about this group. However, the endemic character of the study species has attracted the interest of butterfly experts and our knowledge of the species distribution has increased substantially (around 600 %) in the last 40 years. From among the software available to perform potential distribution models, we selected MaxEnt (Philips et al., 2006) given that it is based on species presence–only data and has been widely used to successfully predict species distribution (Pliscoff and Fuentes–Castillo, 2011; Syfert et al., 2013). This program is based on maximum entropy to model the potential distribution of a species from their presence distribution points and geographical information (variables) available. These variables impose restrictions on the distribution of the species, so the obtained models will show the suitability of the predicted area for the presence of this species.

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The updated known distribution of the butterfly A. morronensis and the storksbills that serve as food plants for its larvae were considered in the main objectives of this study, which were: (1) to create potential distribution models for the butterfly and its food plants; (2) to project these into the future using different climate change scenarios; (3) to combine the two models (interaction butterfly–plant) to see how climate change will affect the biotic interaction between these species; and (4) to discuss whether conservation measures are necessary for the butterfly and/or its food plants. Methods Study area and species occurrence data The taxa of the study include an Iberian endemic butterfly (A. morronensis) and the five Erodium species on which the larvae of the butterfly feed. Although some of these plants species have a wider distribution, the study area of this paper focuses only on their range within the Iberian Peninsula. MGRS (Military Grid Reference System) network with squares of 10 x 10 km was selected as operative geographic units. The occurrence data of the butterfly was taken from García–Barros et al. (2004), and updated to 2016 from different references (Gil–T, 2009; Vicente Arranz and Parra Arjona, 2010; Manceñido González and González Estébanez, 2013; Monasterio León et al., 2014; and unpublished data) resulting in a total of 124 10 x 10 km MGRS squares, which were used to run the models. Data concerning presence of the food plants was first taken from bibliographic and public electronic sources. Several distribution atlases were revised (Aseginolaza et al., 1984; Villar et al., 1997; Uribe– Etxebarria et al., 2006; Serra Laliga, 2007; Alejandre et al., 2009) and data were incorporated from different websites as SIVIM (http://sivim.info/sivi/), ANTHOS (http://www.anthos.es/), GBIF (http://www.gbif.org/species/), BDBC (http://biodiver.bio.ub.es/biocat/), ORCA (http://biodiver.bio.ub.es/orca/) and the Atlas of Flora of Aragón (http://proyectos.ipe.csic.es/floragon/index. php). All these electronic datasets were checked in November 2014 and all data were georeferred when coordinates were not available. Doubtful occurrences were filtered and deleted following the criteria of Alarcón et al. (2012) and expert opinion, which resulted in a total of 81 presence squares for E. carvifolium, 11 for E. cazorlanum, 51 for E. daucoides, 166 for E. foetidum and 132 for E. glandulosum to build the models. Variables We selected climatic and non–climatic variables (topographic, spatial, human activity and geological related variables) to build the potential distribution models (table 1s of the supplementary material). The bioclimatic variables came from WorldClim database (http://www. worldclim.org/) and were described by Hijmans et al. (2005). The remaining variables came from the GLCF (Global Land Cover Facility) database. All the variables

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were obtained in a 5 arc–minute resolution, which is the most similar to our species presence data resolution (MGRS 10 x 10 km), and correspond to the mean value of the variable in each cell, except for landcover, which represents the main landcover type in the cell. To run the models we excluded the variables that were correlated to other variables (Pearson correlation coefficient > 0.7 with SPSS version 15.0 [SPSS, 2006] [Braunisch et al., 2013]), keeping in that case only the most biologically relevant variable for the species. With the final set of uncorrelated variables, we obtained a potential distribution model for the present situation that was then projected to future scenarios, retaining the climatic and non–climatic variables that will not change in the future. All these variables used to build the models are shown in table 1s of the supplementary material. The models were projected to two future periods: 2041–2060 (2050) and 2061–2080 (2070), using the general circulation model (GCM) CCSM4 (Community Climate System Model from the University Corporation for Atmospheric Research, UCAR). They were performed under two Representative Concentration Pathways (RPCs) scenarios (2.6 and 8.5) that differ in climate change severity. RPC 2.6 assumes that global annual greenhouse gas emissions will peak between 2010 and 2020 and then decline with a global mean temperature rise of 1 ºC. RPC 8.5 infers a continuous increase throughout the 21st century, estimating a rise between 2 ºC and 3.7 ºC until 2100 (Meinshausen et al., 2011; IPCC, 2013). Modelling potential distribution maps To build the potential distribution maps, we used the Maxent program, version 3.3.3 k (Phillips et al., 2006). Maxent works better than other techniques with low sample sizes (Hernández et al., 2006; Pearson et al., 2007; Kumar and Stohlgren, 2009) and it can also be used alone to produce accurate models (Fernández et al., 2015; Fourcade et al., 2017; Jacinto–Padilla et al., 2017). We considered the default parameters (10–6 convergence limit, 10,000 background points) recommended by Phillips et al. (2006). Fifteen replicates were performed for each model with 5,000 maximum iterations and subsample replicated run type (Young et al., 2011) using the logistic output format that is easier to interpret (Phillips and Dudik, 2008). We used 75 % of the data to build the models, and the remaining 25 % were randomly used to validate their quality, following other authors (Pawar et al., 2007; Davis and Cipollini, 2016). With E. cazorlanum, and because its number of presence occurrence data was < 25, we used the Jacknife (leave–one–out) procedure recommended by Pearson et al. (2007). In this case, we built as many models as number of known presences we had (11), and removed in each model, one of the occurrence points each time to perform it. For each species, we built a present potential distribution model that was projected to the future in two emission scenarios (2.6 and 8.5) for 2041–2060 and 2061–2080 considering all the variables specified in table 1s of the supplementary material. Then, we built

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Table 1. Percentage contribution to the distribution models of the main environmental and climatic variables included in the models of the butterfly A. morronensis and its larval food plants of the genus Erodium. The values of the variables that most contributed to the performance of the model in each species are highlighted in bold: Amr, A. morronensis; Ecr, E. carvifolium, Ecz, E. cazorlanum; Edc, E. daucoides; Eft, E. foetidum; Egl, E. glandulosum. Tabla 1. Porcentaje de contribución a los modelos de distribución de las variables ambientales y climáticas más importantes incluidas en los modelos de la mariposa A. morronensis y las plantas nutricias de sus larvas del género Erodium. Los valores de las variables que más contribuyeron a la realización del modelo de cada especie se resaltan en negrita. (Para las abreviaturas de las especies, véase arriba).

Variables

Species Amr

Ecr

Ecz

Edc

Eft

Egl

Elevation

46.6 40.3 7.1 28.3 – –

Slope

27.4

1.8 21.4 14.2 27.1 10.4

Mean temperature of wettest quarter

6.5

2.6

Latitude

1.9

21 15.5 – 37.9 –

–

12.1

–

– 15.8 25.1 20.4 2.7

Precipitation of driest month Mean temperature of driest quarter

–

– –

3.8 8.4 – 17.3 0.8 –

–

53.8

3.2

7.5

–

6.1

0.6

10.7

a model representing the interaction butterfly–food plants (in present and future scenarios) overlapping the model of the butterfly and the sum of the plant species in each period, by calculating the minimum number of squares that they had in common. As representation threshold we used the 'equal training sensitivity and specificity logistic threshold', since it is one of the five best–suited thresholds recommended by Liu et al. (2005). However, it presented very low values and did not fit the known distribution of the species. Therefore, to adjust the predictions to the most favourable areas, we chose the mean suitability value predicted by the models for the upper 75 % of all MGRS presence points, revealed to be 0.6. For this purpose, we extracted the suitability values (0–1) of each pixel given by the models with QGIS 2.6.0 (Quantum GIS Development Team, 2015) and made the average of the 75 % of the grids with higher values. This means that values given by the models higher than 0.6 were considered as very favourable areas for the presence of the species, due to the good adequacy of the considered variables. We measured the change in favourable areas for the species according to the different future scenarios. For this purpose, we compared the percentage of pixels above the threshold mentioned before (0.6) within the different models and scenarios. Finally, favourable areas for the present models of the butterfly and the plants were extracted to study their intersection with the Spanish network of protected areas (http://www. mapama.gob.es/).

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28.4 2.2

–

6.5

Annual temperature range Calcareous/siliceous

–

0.7

Precipitation of wettest quarter Longitude

–

4.5

Model validation To evaluate the models, first we used the AUC (Area Under a Receiver Operating Characteristic –ROC– Curve) value, which shows the accuracy of the model (Newbold et al., 2009). AUC values between 0.7 and 1 mean that the model is well fitted and is better than one randomly classified (Pearce and Ferrier, 2000; Philips and Dudík, 2008). We next used the equal training sensitivity and specificity logistic threshold (Liu et al., 2005) to calculate a classification percentage (obtained as the number of test locations with predicted probabilities above this threshold divided by the total number of test locations), which shows the number of squares that have been well classified (Baldwin and Bender, 2008). Finally, the statistical significance of the models was calculated using the 11 omission binomial default tests given by MaxEnt (Phillips et al., 2006). Results Significant variables Due to the minimum contribution of some of the non–climatic variables (such as landcover or sun radiation) to the present potential distribution models of all species (table 2s of the supplementary material), we did not include them in the comparison between future and current potential distribution models. This is

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because the same variables are needed to compare the different models. In our models, the variable that most contributed (47 %) to the performance of the models for A. morronensis was elevation. This variable was highly correlated to the annual mean temperature (Pearson coefficient –0.87) that was not included in the models. For two of the larval food plant species (E. carvifolium and E. daucoides), elevation was also the variable with major contribution, but latitude, annual temperature range and precipitation of the driest month were also important to build the models for E. foetidum, E. cazorlanum and E. glandulosum respectively (table 1). Slope, latitude and longitude were also important for the interaction model that considered the butterfly and its food plants. Table 3s of the supplementary material presents the contribution percentage of all the variables to the projected and not projected to the future models for the butterfly and its larval food plants. Present potential distribution models Most of the known occurrence points of the species appeared in suitable areas, in accordance with the prediction of the present distribution models obtained with MaxEnt for the butterfly (fig. 1) and for the plants (fig. 2). For A. morronensis, E. cazorlanum and E. daucoides more than 50 % of the areas predicted as suitable appear within the limits of the Spanish network of protected areas. The result of overlapping the present potential distribution model obtained with MaxEnt for the butterfly and the sum of the five plant species showed a lower amount of favourable squares (fig. 3). For this map, only the minimum number of squares that they had in common was represented (see methods). Future potential distribution models Future potential distribution models show a slight increase in the number of favourable squares, both for the butterfly and for the joint model for the plant species, especially in the most radical scenario (8.5) in 2070 (fig. 4). We overlapped these maps (butterfly and plants), using the same procedure used for the present potential distribution model (fig. 5), to obtain the future representation of the interaction between the butterfly and its food plants in 2070, which better shows the real probable scenario for A. morronensis. We calculated the percentage of favourable habitat loss for each species above the 75 % upper threshold selected to represent a major probability of occurrence (0.6) as the difference between potential present and future distribution models for the butterfly, for each plant and for their interaction (table 2). Negative values imply loss of favourable areas for the species. The species predicted to lose favourable habitat in most scenarios and periods are E. carvifolium, E. foetidum and E. glandulosum, with E. foetidum having the greatest losses (between 19 and 20 % of its favourable habitat), while for the other species and for their interaction with the butterfly, suitable areas slightly improve.

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A. morronensis

0–0.15 0.15–0.30 0.30–0.45 0.45–0.60 0.60–1

Fig 1. Occurrence data and present potential distribution model of A. morronensis. Known records and present potential distribution model obtained with MaxEnt for A. morronensis are shown. White dots represent its currently known distribution and darker colours show most favourable areas for the species. Fig. 1. Datos de distribución y modelo de distribución potencial en el presente de A. morronensis. Se muestran las presencias conocidas y el modelo de distribución potencial en el presente obtenido con MaxEnt para A. morronensis. Los puntos blancos representan la distribución actual conocida y los colores más oscuros muestran las zonas más favorables para la especie.

Model validation All AUC values were above 0.9 (table 3), showing that the developed models are suitable and have a high discriminatory power. More than 78 % of the squares presented logistical probability values with greater probability than that required for each model (classification percentage in table 3), supporting their reliability and showing that at least 78 % of the grids were correctly classified as favourable areas for the species. The 11 binomial default tests worked out with MaxEnt had values of statistical significance smaller than 0.01 for all species except E. cazorlanum (table 3), showing that the prediction can be considered reliable for most species. Discussion As ectothermic animals, butterflies appear to be highly influenced by climate, mainly temperature (Steigenga and Fischer, 2009). Supporting this idea, our study shows that elevation, which is highly correlated to temperature (Pearson coefficient of –0.87), is the variable that mainly contributes to the potential distribution models of A. morronensis (46.6% contribution). Thus, climate change that will produce global increases in temperature could potentially be a threat to this spe-

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E. carvifolium

0–0.15 0.15–0.30 0.30–0.45 0.45–0.60 0.60–1

E. daucoides

0–0.15 0.15–0.30 0.30–0.45 0.45–0.60 0.60–1

E. glandulosum

0–0.15 0.15–0.30 0.30–0.45 0.45–0.60 0.60–1

E. cazorlanum

0–0.15 0.15–0.30 0.30–0.45 0.45–0.60 0.60–1

E. foetidum

0–0.15 0.15–0.30 0.30–0.45 0.45–0.60 0.60–1

E. carvifolium E. cazorlanum E. daucoides

E. foetidum E. glandulosum

0–0.15 0.15–0.30 0.30–0.45 0.45–0.60 0.60–1

Fig. 2. Occurrence data and present potential distribution models of all Erodium species. Known records and present distribution models obtained with MaxEnt for each Erodium species that are used by A. morronensis larval stage are shown: A, E. carvifolium; B, E. cazorlanum; C, E. daucoides; D, E. foetidum; E, E. glandulosum. White dots represent the currently known distribution of each species. The last figure (F) represents the sum of all Erodium present potential distribution models, with symbols for the known distribution of each species. Darker colours show more favourable areas for the species. Fig. 2. Datos de distribución y modelos de distribución potencial en el presente de todas las especies de Erodium. Se muestran las presencias conocidas y los modelos de distribución potencial en el presente obtenidos con MaxEnt para cada especie de Erodium utilizada por A. morronensis en su estado larvario: A, E. carvifolium; B, E. cazorlanum; C, E. daucoides; D, E. foetidum; E, E. glandulosum. Los puntos blancos representan la distribución actual de cada especie. La última figura (F) representa la suma de todos los modelos de distribución potencial en el presente de las diferentes especies de Erodium, con símbolos diferentes para la distribución conocida de cada especie. Los colores más oscuros muestran zonas más favorables para la especie.

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cies. Another factor that could aggravate this situation is the fact that this species lives in high–altitude habitats, one of the most vulnerable areas to climate change (Wilson et al., 2007; Engler et al., 2011). In our models, we also considered the interaction of the butterfly with its Erodium larval food plants, in order to improve the quality of the prediction (Kissling et al., 2012; Wisz et al., 2013). We did not consider the mutualistic relationship with ants (Romo et al., 2015) because in this case, ant attendance is facultative and, as far as we know, the presence of ants does not represent a limiting factor for the survival of the butterfly (García–Barros et al., 2013). The habitat of Erodium plants is also generally restricted to mountains, but they have a wider range than the butterfly. The environmental variables that most influenced the potential distribution models of plant species were elevation (highly correlated to temperature) and precipitation. A reasonable hypothesis would therefore be that their distribution could also be affected by climate change (Grabherr et al., 1994; Dullinger et al., 2012; Gottfried et al., 2012). On the other hand, non–climatic variables such as landcover and radiation contribute little to the models. The reason for this could be that the species are mainly present in open areas (rocks, screes, grasslands) that would have similar values for the mentioned variables. Regarding the results of the present potential distribution models, and given the actual distribution of the butterfly, some known occurrence points appear in unfavourable areas. Most of these points are close to favourable patches and are likely to be the result of an expansion of nearby populations (Munguira and Martin, 1988). They are also located in areas where

0–0.15 0.15–0.30 0.30–0.45 0.45–0.60 0.60–1

Fig. 3. Present potential distribution model of the interaction between A. morronensis and the Erodium species. Overlap of the present potential distribution models for A. morronensis and the five Erodium species considered in the study worked out as the minimum number of squares that all the species have in common. Darker colours show more favourable areas for the butterfly. Fig. 3. Modelo de distribución potencial en el presente de la interacción entre A. morronesis y las especies de Erodium. Superposición de los modelos de distribución potencial en el presente de A. morronensis y las cinco especies de Erodium analizadas en el estudio, calculada como el número mínimo de cuadrículas que todas las especies tienen en común. Los colores más oscuros muestran zonas más favorables para la mariposa.

A B

0–0.15 0.15–0.30 0.30–0.45 0.45–0.60 0.60–1

Fig. 4. Future potential distribution models for A. morronensis and the set of all Erodium species. Future potential distribution models for A. morronensis (A) and Erodium plants (B) projected to 2070 using the most radical climate change scenario (RPC 8.5) in order to show the most pessimistic prediction for the species in the future. Darker colours show more favourable areas for the species in both maps. Fig. 4. Modelos de distribución potencial en el futuro de A. morronensis y el conjunto de las especies de Erodium. Modelos de distribución potencial de A. morronensis (A) plantas del género Erodium (B) extrapolados a 2070 utilizando las condiciones climáticas más drásticas (RCP 8.5) a fin de mostrar la predicción más pesimista para estas especies en el futuro. Los colores más oscuros muestran zonas más favorables para las especies en ambos mapas.

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0–0.15 0.15–0.30 0.30–0.45 0.45–0.60 0.60–1

Fig. 5. Future potential distribution model of the interaction between A. morronensis and the Erodium species. Overlap of the 2070 RPC 8.5 models for A. morronensis and the five Erodium species considered in the study worked out as the minimum number of squares the two groups have in common. Darker colours show the more favourable areas for the butterfly. This map shows the most pessimistic prediction for the butterfly in the future. Fig. 5. Modelo de distribución potencial en el futuro de la interacción entre A. morronensis y las especies del género Erodium. Superposición de los modelos previstos para 2070 con las condiciones de RCP 8.5 de A. morronensis y las cinco especies de Erodium analizadas en el estudio, calculada como el número mínimo de cuadrículas que los dos grupos tienen en común. Los colores más oscuros muestran zonas más favorables para la mariposa. Este mapa muestra la predicción más pesimista para la mariposa en el futuro.

their larval food plants are present. The Pyrenean mountain range is not particularly favourable in our models (only four squares exceed the threshold of 0.6), although there are citations for six distribution points in the area. This could be a result of historical events related to the species distribution, such as local extinctions or previous incomplete colonizations. Also, on the eastern part of the Central Mountain System, a favourable area for the butterfly is predicted; however, it may be unoccupied due to the absence of adequate larval food plants in this region. Our models only consider the presence of the species in relatively large areas (10 x 10 km2). They did not consider probable shifts of distribution ranges to higher altitudes (Walther et al., 2002; Parmesan, 2006; Wilson et al., 2007). Therefore, this effect of climate change may have been overlooked in our study. For more comprehensive knowledge on the impact of climate on this species, it would be imperative to improve the sampling effort to detect probable altitudinal shifts on this and other butterfly species. To improve the models performance it would also be necessary to update the available distribution data,

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since most of the occurrence data were collected for the 2004 atlas (García–Barros et al., 2004). However, most locations recently visited by the authors are still occupied and show large numbers of the butterfly. As for the potential distribution model of the plants, the occurrence data on areas predicted as unfavourable in our models could also be explained as an expansion from favourable patches, as they are close to the boundaries of these suitable areas. Besides, it should be taken into account that the map (fig. 2) is a result of the sum of the present potential distribution models of all plant species; the most favourable areas are thus those that include more than one species. The eastern coast areas of the Iberian Peninsula do not seem to have suitable conditions for the studied Erodium species, probably as a consequence of the high temperatures during the summer and the limited extension of mountains in these areas. Studying ecological interactions between species in terms of conservation is becoming an important field of study. Global change is shifting the cycles of the species, decoupling their interactions, that usually disappear before species extinctions themselves (Peñuelas et al., 2002; Valiente–Banuet et al., 2015). Almost all the favourable areas for the butterfly in our predicted models match favourable areas for the plants. Thus, when the overlap of models of the butterfly and the plants was considered, the favourable areas for the butterfly did not change significantly, with the only exception being a slight loss of favourability on the west side of the Cantabrian Mountains and Central Mountain System. Comparing the present and the future overlap of the potential distribution models, the prediction shows no loss of habitat favourability for the interaction between the butterfly and its larval food plants. This observation makes our study particularly interesting, because in our present and future potential distribution models the interaction does not seem to be significantly affected by shifts under climate change. Moreover, it is relevant to take into account Dincă et al's subdivision of A. morronensis populations into two genetic entities, 2015. This subdivision matches the northern and central populations in one of the entities and the southern populations in the other. We did not model these two entities separately because they probably represent different species; a more detailed molecular study based on other non–mitochondrial markers would be needed. But comparing the model predictions with the known distribution of the species, we consider our results show that neither genetic entity would experience loss of habitat adequacy: the northern populations present an increase in the future, while the southern populations remain stable. Figure 1 shows the differences between the known distribution and the present models, where some of the isolated points of the known distribution of the butterfly appear in unfavourable areas as mentioned before. While the models show that the potential distribution of the butterfly may be optimistic for the future shifts predicted under climate change, the case of some of the plants included in this study may require

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Table 2. Evolution of habitat favourability for all the species in the projected future models. Percentage of the loss (negative values) or increase (positive values) in habitat favourability on the obtained future projections for each scenario (RPC 2.6 and 8.5, see methods) regarding the present potential distribution models of A. morronensis and its larval food plants of the genus Erodium. The values come from the subtraction of the percentage of favourable areas given by the future potential distribution models minus the percentage of these areas in the present distribution models. Tabla 2. Evolución de la favorabilidad del hábitat para todas las especies en los modelos extrapolados al futuro. Porcentaje de pérdida (valores negativos) o ganancia (valores positivos) de favorabilidad del hábitat con respecto a la previsiones futuras para cada escenario (RCP 2.6 y 8.5, véanse los métodos) en relación con los modelos de la distribución en el presente de A. morronensis y las plantas nutricias de sus larvas del género Erodium. Los valores se obtienen restando el porcentaje de zonas favorables obtenido en los modelos de la distribución en el presente al porcentaje de estas zonas obtenido por los modelos de la distribución potencial en el futuro. Models Species

2050

–5.53 –1.98 –1.19 –3.95

E. cazorlanum

0.00 0.00 0.00 0.00

E. daucoides –2.59 7.33 6.03 3.45 –20.58 –20.99 –19.34 –19.75

E. glandulosum –1.78 –5.33 –2.67 –4.00 6.28 8.7 5.8 1.93

special attention. E. foetidum only occurs in the Iberian Peninsula and southern France and, although it is listed as Near Threatened (NT) by the IUCN, according to our predictions it will have its favourable area reduced to around 20 % (table 2). Furthermore, only around 40 % of the favourable areas displayed for this plant in the present potential distribution model occur inside the network of protected areas. This could have been a problem for the butterfly, but as it feeds on any of the studied Erodium species and since E. foetidum shares its habitat with two more Erodium species, our models show no negative effect on habitat suitability for the butterfly. The remaining plant species considered are predicted by our models

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Table 3. Validación de los modelos. Valores AUC (área debajo de la curva ROC), umbral logístico de la prueba de igualdad de la sensibilidad y especificidad (EtSI), porcentaje de clasificación (Cp) y significación (test p) de las 11 pruebas binomiales dadas por defecto por MaxEnt para los modelos de A. morronensis y las plantas nutricias de sus larvas (*: para E. cazorlanum el 60 % de las pruebas mostró una significación de p < 0,05 y para el resto, de p < 0,01). Species

AUC EtSl

Cp p–test

A. morronensis

0.922 0.234 80.3 < 0.01

E. carvifolium

0.936 0.258 85.7 < 0.01

E. cazorlanum

0.991 0.307 81.8 < 0.05*

E. daucoides

0.95 0.315 83.6 < 0.01

E. foetidum

0.943 0.264 78.4 < 0.01

E. glandulosum 0.931 0.284 78.8 < 0.01

2070

E. carvifolium

Interaction

Table 3. Validation of models: AUC values (area under the ROC curve), equal training sensitivity and specificity logistic threshold (EtSl), classification percentage (Cp) and significance (p–test) for the 11 binomial tests obtained by default with MaxEnt for A. morronensis and its larval food plant models (*: for E. cazorlanum 60 % of the tests had p < 0.05 while for the remaining tests p was < 0.01).

2.6 8.5 2.6 8.5

A. morronensis 5.74 9.09 5.26 2.87

E. foetidum

to stay more or less stable in their actual range in future projections, with small losses of favourable areas (E. carvifolium and E. glandulosum), or small increases (E. daucoides). E. cazorlanum, an Iberian endemic species listed as Vulnerable by the IUCN, is the species of our study with the most restricted distribution range. Nevertheless, the models do not show relevant changes for its populations and over 80 % of the potential present favourable areas appear in protected areas. We have to consider that these species with geographically restricted areas do not usually reflect all the environmental or topographic information where the species really occurs and the results can be biased somehow (Titeux et al., 2017). But even in the hypothetical case that there was a local extinction of the E. cazorlanum species, the larvae of the butterfly would be able to feed on other Erodium species in that area, such as E. daucoides or E. foetidum. In fact, recent observations (Munguira, unpublished data) show that the butterfly can use E. foetidum in the Sierra de Cazorla (SE Spain). However, it is true that the model obtained for E. cazorlanum should be interpreted not as potential areas where it could expand its actual range limit but as regions that have similar environmental conditions to those where the species was recorded (Pearson et al., 2007).

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Our results confirm the favourable status of the butterfly from a conservation point of view. Munguira (1989) stated that the species was not a priority for conservation since it was present in a large number of locations (which increased from 14 to 56 in 1975–1988 as a result of better sampling), showed strong populations and had some of its best populations inside protected areas. The present situation, with 50 % of the populations within protected areas, is even better than in 1989, and moreover, the models show stable predictions for the future. Persistence of populations will be easier for those populations living on scree slopes or rocks (most of the populations from Pyrenees, Cantabrian Mountains, Iberian Mountain System and southern sierras) where no specific management is required. Grassland populations (Galicia, Ávila, Soria and some in Burgos) need livestock grazing to keep habitat quality, so traditional land uses should be favoured in these areas. In Abejar (Soria) management already taking place to preserve the Dusky Large Blue Phengaris nausithous (Bergsträsser, 1779) populations (Vicente et al., 2013) would favour the survival of A. morronensis populations in the same grasslands. To sum up, the threat of climate change to A. morronensis could be minimal while current predictions through climatic models show considerable reductions of distribution ranges for most butterfly species (Beaumont and Hughes, 2002; Settele et al., 2008). Focusing on the Iberian Peninsula, it seems mountain butterflies will also lose habitat favourability in the future, especially when the interaction between the butterfly and the larval food plants is considered (e.g. Phengaris nausithous, Romo et al., 2014, 2015). Besides, more than 50 % of the areas considered favourable in the present potential It seems conservation of the butterfly and most of its food plants is probably not jeopardised for the time being. However, it would be interesting to focus new research on how the butterfly would counteract future challenges in climate change, and whether interaction with its larval food plants will remain stable –as predicted by our models– or increase or decline. Acknowledgements Butterfly occurrence data gathered in Barcelona, Guadalajara, León, Soria, and Teruel were completed with information kindly provided by Roger Vila, Luis Óscar Aguado, Rafael Pérez and Ángel Marco. Juan José Aldasoro and Marisa Alarcón helped with our queries regarding the distribution of Erodium species. References Alarcón, M., Vargas, P., Sáez, L., Molero, J., Aldasoro, J. J., 2012. Genetic diversity of mountain plants: Two migration episodes of Mediterranean Erodium (Geraniaceae). Molecular Phylogenetics and Evolution, 63: 866–876. Alejandre, J. A., López, J. M. G., Sanz, G. M., 2009.

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Rosalia alpina adults (Linnaeus, 1758) (Insecta, Coleoptera) avoid direct sunlight A. Castro, L. Drag, L. Cizek, J. Fernández

Castro, A., Drag, L., Cizek, L., Fernández, J., 2019. Rosalia alpina adults (Linnaeus, 1758) (Insecta, Coleoptera) avoid direct sunlight. Animal Biodiversity and Conservation, 42.1: 59–63, https://doi.org/10.32800/abc.2019.42.0059 Abstract Rosalia alpina adults (Linnaeus, 1758) (Insecta, Coleoptera) avoid direct sunlight. Adults of the threatened beetle species Rosalia alpina are usually associated with sun–exposed dead wood. In previous fieldwork, however, we frequently found adult beetles on shaded surfaces of trees. We thus studied whether adults preferred different lightning conditions depending on their behavior on 447 beech trees located in four forests in two distant locations in Europe. From a total of 542 individuals, we observed that 54 % of them occurred in shaded conditions, and 35 % in predominantly shaded conditions. This avoidance of direct sunlight could be widespread in the species because it was independent of the location and behavior. Key words: Beech forest, Behavior, Cerambycidae, Longhorn beetle Resumen Los adultos de Rosalia alpina (Linnaeus, 1758) (Insecta, Coleoptera) evitan la exposición directa a la luz solar. Por lo general, se considera que los adultos del escarabajo amenazado Rosalia alpina están asociados a la madera muerta expuesta al sol. Sin embargo, en algunos trabajos previos de los autores se observaron adultos en superficies sombreadas de árboles. Así, se comprobaron sus preferencias por distintas condiciones de iluminación dependiendo de su comportamiento en 447 hayas localizadas en cuatro bosques diferentes de dos localidades europeas lejanas. El 54 % y el 35 % de los 542 individuos observados se encontraron en condiciones de sombra total y parcial, respectivamente. Esta evitación de la luz solar directa podría estar generalizada en la especie, ya que se mostró independiente de la localización y el comportamiento. Palabras clave: Hayedo, Comportamiento, Cerambycidae, Longicornio Received: 07 II 18; Conditional acceptance: 28 III 18; Final acceptance: 03 VII 18 Alberto Castro, Jon Fernández, Department of Entomology, Aranzadi Society of Sciences, Zorroagagaina s/n., 20014–Donostia/San Sebastián, Gipuzkoa, Spain.– Lukas Drag, Lukas Cizek, Institute of Entomology, Biology Centre CAS, Branisovska 31, Ceske Budejovice 37005, Czech Republic. Corresponding author: Alberto Castro. E–mail: adecastro@aranzadi.eus

ISSN: 1578–665 X eISSN: 2014–928 X

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Introduction Sun–exposed, dead or partially dead trees are the preferred habitat of Rosalia alpina (Russo et al., 2011, 2015; Castro and Fernández, 2016), a threatened and legally protected species facing decline in Europe (Luce, 1996; Adamski et al., 2013; Bosso et al., 2013, 2018). R. alpina adults appear in summer and are diurnal. Their activity peaks in the afternoon, and they are mainly found on sun–exposed trees (Drag et al., 2011; Russo et al., 2011; Castro and Fernández, 2016). Thus, like many other saproxylic beetles (Lindhe et al., 2005, Vodka et al., 2009), R. alpina is considered to be a sun–loving species. The fact that an organism inhabits trees in sunny places, however, does not necessarily mean that its activity is concentrated on surfaces receiving direct sunlight (Kreuger and Potter, 2001; Bancroft and Smith, 2005). We analyzed the frequencies of observations of R. alpina adults on trees in relation to exposure to sun or shade. To our knowledge, this issue has not been previously studied in R. alpina. Material and methods The research took place at three sites in the Czech Republic (Maly Bezdez, Velky Bezdez and Slatinne Hills) and one site in Spain (Artaso; table 1). Maly Bezdez and Velky Bezdez are hills that are mainly covered by semi–open beech forests with rather small and crooked trees, while the beech forest in Slatinne Hills consists mainly of tall trees. The Spanish location, Artaso, is a closed forest with abandoned pollard beech and sporadic clearings mainly on the northern slope. A detailed description of the study sites and observations methods can be found in Drag et al. (2011) and Castro and Fernández (2016). Observations were conducted between July 12th to August 10th of 2008 (10–19 visits per tree) at all three sites in the Czech Republic, and also from July 5th and August 16th of 2009 (39) in Slatinne Hills, and from July 1st to August 31st of 2010 (8) in Artaso.

Over this three–year period, 157, 155 and 135 trees each year, respectively, were visually inspected for living adults. Observations were always made in suitable weather (sunny days), between 10:00 and 18:00 h in the Czech Republic, and between 11:00 and 18:00 h in Spain. Two variables were recorded for all observed beetles, exposure to sunlight and behavior. Exposure to sunlight consisted of three categories: sun, dim light, and shade. Sun and shade categories meant the individuals were in totally in sunlight or totally in the shade. Dimly light meant the individuals were in partial shade. Behavior categories were defined as resting, reproduction and movement. Rest referred to not–moving individuals; reproduction included mating, males fighting for females, and females ovipositing or looking for oviposition sites; and movement referred to individuals walking, exploring, landing on trees, and territorial fights between males. For each study site, we tested the relative frequencies of the different categories against the null hypothesis assuming all categories to be equal by performing x2–tests, goodness of fit (Zar, 2010). Whenever these tests yielded significant results at the P < 0.05 threshold, pairwise x2 comparisons were carried out. In the first step, we analyzed whether the frequencies of individuals were affected by the degree of exposure to sunlight. In the second step, we tested whether individuals selected different exposures to sunlight according to their behaviors. Due to the low sample size in Artaso, pairwise comparisons were performed applying the Yates correction for continuity (Zar, 2010), and no statistical analysis was carried out to explore interactions between sunlight exposure and behavior. Statistical analyses were performed using the PAST program (Hammer et al., 2001) version 3.06 (http:// folk.uio.no/ohammer/past/). Results We observed a total of 542 R. alpina adults. Regardless of the location, frequencies of individuals were

Table 1. Location of the study sites. Tabla 1. Localización de las zonas de estudio. Altitude Area Site Year Place Country (m a.s.l.) (ha)

East

50.540 50.539 50.553

14.713 14.720 14.707

Maly Bezdez, 2008 Ralska Upland, Czech Velky Bezdez, Northern Bohemia Republic Slatinne Hills

400–577 400–604 350–430

Slatinne Hills 2009

Ralska Upland, Northern Bohemia

350–430 12.1 50.553 14.707

Artaso

Oñati, Basque Country

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2010

Czech Republic Spain

690–940

17.9 20.3 12.1

North

33.4

42.975

–2.406

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Animal Biodiversity and Conservation 42.1 (2019)

higher in locations with less exposure to sunlight (table 2). This higher abundance in more shaded sites was also independent of the behavior shown by individuals (fig. 1). For any combination of location and behavior, the percentage of individuals exposed to direct sunlight never exceeded 19 % (fig. 1). Hence, 89 % of individuals avoided any activity on sun–exposed surfaces.

Table 2. Results of testing the equality of frequencies from the different categories of sunlight exposure. Tabla 2. Resultados de las pruebas de igualdad de frecuencias de las diferentes categorías de exposición a la luz solar.

Discussion

Categories tested d.f. Bezdez (n = 356) All 2 Sun vs. dim light 1 Sun vs. shade 1 Dim light vs. shade 1 Slatinne Hills ( n = 152) All 2 Sun vs. dim light 1 Sun vs. shade 1 Dim light vs. shade 1

The avoidance of tree surfaces exposed to sunlight by R. alpina adults was consistent in sites as far apart as southwest Europe and central Europe, and in three different habitat types, suggesting that this behavior is characteristic of the species. The causes of this behavioral pattern could be understood by evaluating several hypotheses. For example, avoidance of surfaces exposed to sunlight could be related to body thermoregulation, camouflage against predators, or a trade–off between the two. The grayish blue coloration of the body and the dark dorsal spots in the elytra and antenna seem to camouflage with the surface of trunks and branches of trees where the species lives (Luce, 1996). There is also evidence that the dark spots on the elytra and antenna can perform a thermoregulatory function to quickly absorb and retain heat (Kostić et al., 2016). R. alpina adults are relatively active and rarely feed (Drag et al., 2011). However, activity involves energy costs and greater

x42 = 6.180

Observations (%)

100

80 60

2 1 1 1

88.342 58.449 90.741 4.541

P < 0.001 < 0.001 < 0.001 0.033

56.235 < 0.001 7.053 0.008 48.485 < 0.001 20.961 < 0.001 24.060 4.9 19.36 7.84

< 0.001 0.027 < 0.001 0.005

x42 = 7.503 P = 0.112

3 4

1 1

8 8

40 20

P = 0.186

Artaso (n = 34) All Sun vs. dim light Sun vs. shade Dim light vs. shade

18 59

9 Sun

Dimly lit Shade

Rp Mv Bezdez

Rs Rp Mv Slatinne Hills

Rp Mv Artaso

Fig. 1. Numbers of individuals (inside bars) and their frequencies expressed as percentages observed per behavior and exposure to sunlight categories: Rs, resting; Rp, reproduction; Mv, movement. (The x2–test for Artaso was not performed due to statistical constraints, see text). Fig. 1. Número de individuos (dentro de las barras) y sus frecuencias expresadas como porcentajes observados por comportamiento y categoría de exposición a la luz solar: Rs, descanso; Rp, reproducción; Mv, en movimiento. (No se realizaron las pruebas x2 para Artaso debido a restricciones en los requerimientos estadísticos, véase el texto).

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risk of predation, primarily by good visual hunters such as birds, which are known to feed on R. alpina (Adamski et al., 2013). Activity in the shade may lessen the chances of predators detecting this prey (Carrascal et al., 2001; Carr and Lima, 2014) and meet R. alpina energetic demands by being active on the shaded surfaces (after short exposures to the sun) of tree trunks and branches during the hottest times of the day and year. Additionally, it is possible that females oviposit in the shaded areas to avoid exposing eggs to lethal temperatures (Keena, 2006). As the trees in our study were randomly chosen, the sun–exposed and shaded parts were probably not balanced, but we do not consider that such selection would fully explain the pattern we observed. Accordingly, in the Artaso pollard forest we observed only one out of 21 adults in sunny places in trees that provided larger surfaces exposed to sunlight ('big clearings', see Castro and Fernández, 2016). Although the individuals of R. alpina are more likely to be found more active on the shaded portions of the tree, it is highly probable that the sun–exposed habitats can still be preferred (Drag et al., 2011; Russo et al., 2011). Open habitats always offer some shaded parts of the wood, but in addition to that they may provide other benefits for the beetle. Acknowledgements The work in Artaso was part of the LIFE project NAT/E/000075 'Management and conservation of the habitats of Osmoderma eremita, Rosalia alpina and other saproxylic species of community interest in Gipuzkoa' led by the Provincial Council of Gipuzkoa. Other LIFE project partners were IKT (now Hazi), Itsasmendikoi, Basoa Fundazioa, the Basque Government, and Aranzadi Society of Sciences. Gathering of autoecological data in Artaso was granted by the Head Office of Biodiversity of the Basque Government. The data collected in Bezdez and Slatinne Hills were part of the monitoring project of R. alpina supported by Ministry of Environment (VaV/SP/2d3/153/08) and the Czech Science Foundation (17–21082S). References Adamski, P., Holly, M., Michalcewicz, J., Witkowski, Z., 2013. Decline of Rosalia longicorn Rosalia alpina (L.) (Coleoptera: Cerambycidae) in Poland – selected mechanisms of the process. In: The role and contribution of insects in the functioning of forest ecosystems: 358–372 (W. Ząbecki, Ed.). Wydawnictwo Uniwersytetu Rolniczego w Krakowie, Kraków. [In Polish.]. Bancroft, J. S., Smith, M. T., 2005., Dispersal and influences on movement for Anoplophora glabripennis calculated from individual mark–recapture. Entomologia Experimentalis et Applicata, 116: 83–92. Bosso, L., Rebelo, H., Garonna, A. P., Russo, D., 2013. Modelling geographic distribution and detect-

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ing conservation gaps in Italy for the threatened beetle Rosalia alpina. Journal for Nature Conservation, 21: 72–80. Bosso, L., Smeraldo, S., Rapuzzi, P., Sama, G., Garonna, A. P., Russo, D., 2018. Nature protection areas of Europe are insufficient to preserve the threatened beetle Rosalia alpina (Coleoptera: Cerambycidae): evidence from species distribution models and conservation gap analysis. Ecological Entomology, 43: 192–203. Carr, J. M., Lima, S. L., 2014., Wintering birds avoid warm sunshine: predation and the costs of foraging in sunlight. Oecologia, 174: 713–721. Carrascal, L. M., Díaz, J. A., Huertas, D. L., Mozetich, I., 2001. Behavioral thermoregulation by treecreepers: trade–off between saving energy and reducing crypsis. Ecology, 82(6): 1642–1654. Castro, A., Fernández, J., 2016. Tree selection by the endangered beetle Rosalia alpina in a lapsed pollard beech forest. Journal of Insect Conservation, 20(2): 201–214. Drag, L., Hauck, D., Pokluda, P., Zimmermann, K., Cizek, L., 2011. Demography and Dispersal Ability of a Threatened Saproxylic Beetle: A Mark–Recapture Study of the Rosalia Longicorn (Rosalia alpina). PLOS One, 6(6): e21345, doi:10.1371/ Journal.pone.0021345. Hammer, Ø., Harper, D. A. T., Ryan, P. D., 2001., PAST: Paleontological Statistics Software Package for Education and Data Analysis. Palaeontologia Electronica, 4(1): 9. Keena, M.A., 2006. Effects of temperature on Anoplophora glabripennis (Coleoptera: Cerambycidae) adult survival, reproduction, and egg hatch. Environmental Entomology, 35(4): 912–921. Kostić, I., Pavlović, D., Lazović, V., Vasiljević, D., Stojanović, D., Knežević, D., Tomić, L., Dikić, G., Pantelić, D., 2016. Thermal and camouflage properties of Rosalia alpina longhorn beetle with structural coloration. In: 7th International Scientific Conference on Defensive Technologies: 525–529. OTEH, Belgrade, Serbia, 6–7 October 2016. Kreuger, B., Potter, D. A., 2001. Diel feeding activity and thermoregulation by Japanese beetles (Coleoptera: Scarabeidae) with host plant canopies. Environmental Entomology, 30(2): 172–180. Lindhe, A., Lindelöw, Å., Åsenblad, N., 2005. Saproxylic beetles in standing dead wood density in relation to substrate sun–exposure and diameter. Biodiversity and Conservation, 14: 3033–3053. Luce, J. M., 1996. Rosalia alpina (Linnaeus, 1758). In: Background information on invertebrates of the Habitats directive and the Bern Convention. Part I – Crustacea, Coleoptera and Lepidoptera: 70–73 (P. J. Van Helsdingen, L. Willemse, M. C. D. Speight, Eds.). Council Europe (Nature and Environment, 79) Russo, D., Cistrone, L., Garonna, A. P., 2011. Habitat selection by the highly endangered longhorned beetle Rosalia alpina in Southern Europe: a multiple spatial scale assessment. Journal of Insect Conservation, 15(5): 685–693. Russo, D., Di Febbraro, M., Cistrone, L., Jones, G.,

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Smeraldo, S., Garonna, A. P., Bosso, L., 2015. Protecting one, protecting both? Scale–dependent ecological differences in two species using dead trees, the rosalia longicorn beetle and the barbastelle bat. Journal of Zoology, 297(3): 165–175. Vodka, S., Konvicka, M., Cìzek, L., 2009. Habitat

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preferences of oak feeding xylophagous beetles in a temperate woodland: implications for forest history and management. Journal of Insect Conservation, 13: 553–562. Zar, J., 2010. Bioestatistical Analysis. 5th edition. Pearson Prentice Hall, New Jersey, USA.

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Brief Communication 65

SNP identification and validation in two invasive species: zebra mussel (Dreissena polymorpha) and Asian clam (Corbicula fluminea) L. Peñarrubia, J. Viñas, N. Sanz, B. L. Smith, J. R. Alvarado Bremer, C. Pla, O. Vidal

Peñarrubia, L., Viñas, J., Sanz, N., Smith, B. L., Alvarado Bremer, J. R., Pla, C., Vidal, O., 2019. SNP identification and validation in two invasive species: zebra mussel (Dreissena polymorpha) and Asian clam (Corbicula fluminea). Animal Biodiversity and Conservation, 42.1: 65–68, https://doi.org/10.32800/abc.2019.42.0065 Abstract SNP identification and validation in two invasive species: zebra mussel (Dreissena polymorpha) and Asian clam (Corbicula fluminea). The development of affordable massive parallel sequencing (MPS) has reduced both time and costs of SNP identification for use in conservation and population genetic studies. After MPS, a second validation is usually required. High resolution melting analysis (HRMA) is a fast and simple method for mutation scanning, and thus a suitable validation protocol, particularly in non–model species. We present a set of nine novel polymorphic SNPs identified by MPS and validated with HRMA in two invasive species (the zebra mussel Dreissena polymorpha and the Asian clam Corbicula fluminea). These SNPs can be used in genetic studies to accurately assess and understand past and future invasion events. Key words: Corbicula fluminea, Dreissena polymorpha, Invasive species, Massive parallel sequencing, Single nucleotide polymorphisms Resumen Identificación y validación de PNU en dos especies invasoras: el mejillón cebra (Dreissena polymorpha) y la almeja asiática (Corbicula fluminea). El desarrollo de plataformas asequibles de secuenciación masiva en paralelo (SMP) ha reducido el coste y el tiempo de la identificación de marcadores de polimorfismos de nucleótido único (PNU) para su uso en estudios de genética de poblaciones y de conservación. Tras la SMP, suele ser necesaria una segunda validación. El análisis de las curvas de fusión a alta resolución (HRMA en su sigla en inglés) es un método rápido y sencillo para escanear mutaciones y, por tanto, es un protocolo adecuado de validación de dichos marcadores, especialmente en especies no modelo. En este trabajo se presenta un juego de nueve marcadores polimórficos de PNU nuevos identificados mediante SMP y validados con el HRMA en dos especies invasoras (el mejillón cebra Dreissena polymorpha y la almeja asiática Corbicula fluminea), que pueden utilizarse en estudios de genética de poblaciones para evaluar y entender correctamente los episodios de invasión pasados y los que podrían ocurrir en el futuro. Palabras clave: Corbicula fluminea, Dreissena polymorpha, Especies invasoras, Secuenciación masiva en paralelo, Polimorfimos de nucleótido único Received: 02 III 18; Conditional acceptance: 26 IV 18; Final acceptance: 04 VII 18 Luis Peñarrubia, Jordi Viñas, Nuria Sanz, Carles Pla, Oriol Vidal, Lab. d’Ictiologia Genètica, Dept. of Biology, Univ. de Girona, LEAR Building, Girona, 17073, Catalonia, Spain.– Brad L. Smith, Dept. of Natural Sciences, Brigham Young Univ. Hawaii, Laie, HI 96762, USA.– Jaime R. Alvarado Bremer, Dept. of Marine Biology, Texas A&M Univ. at Galveston, OCSB 3029 room #216, P. O. Box 1675, Galveston, TX 77553, USA and Dept. of Wildlife and Fisheries Sciences, Texas A&M Univ., TAMU 2258, College Station, TX 77840, USA. Corresponding author: Oriol Vidal. E–mail: oriol.vidal@udg.cat

ISSN: 1578–665 X eISSN: 2014–928 X

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Rapid developments in massive parallel sequencing (MPS) technologies have facilitated the use of single nucleotide polymorphisms (SNPs) in population genetic studies (Morin et al., 2009). SNPs have many advantages, including low–scoring error rates, high abundance, functional relevance, easy high–throughput genotyping (Liu et al., 2005), and more accurate estimates of population differentiation (Morin et al., 2009). SNP identification in non–model species can be performed using MPS technologies without a reference genome (Everett et al., 2011). However, after the SNP calling step, a subsequent validation is usually required. High resolution melting analysis (HRMA) is a relatively new and inexpensive technology. It is based on highly precise and accurate measures of melting temperatures (Tm) of PCR–amplified DNA achieved by recording the fluorescence of a saturating DNA dye (Wittwer et al., 2003; Reed et al., 2007). As differences in melting curve profiles are diagnostic of SNPs, homozygote and heterozygote genotypes can be distinguished (Montgomery et al., 2007). This technique has been successfully used for SNP validation in several species, such as swordfish (Smith et al. 2010) and chum salmon (Seeb et al., 2011). In this study, we used non–coding genomic MPS reads from previous studies (Peñarrubia et al., 2015a, 2015b) to identify and subsequently validate with HRMA new SNPs in two invasive species, the zebra mussel (Dreissena polymorpha, Pallas, 1771) and the Asiatic clam (Corbicula fluminea, Müller, 1774). Roche 454 GS FLX reads of D. polymorpha (GenBank SRA accession number SRP051009) and C. fluminea (GenBank SRA accession number SRP073154) were processed using CLC Genomics Workbench version 4.0 (http://www.clcbio.com/) for SNPs as described in Seeb et al. (2011) and Shahin et al. (2012). As a result, 783 sequence variants were found in D. polymorpha, of which 721 were single nucleotide polymorphisms (SNPs) (356 transversions and 365 transitions), 20 were multiple nucleotide variants (MNVs) and 42 were insertions–deletions (InDels). Conversely, in C. fluminea the analysis produced 446 sequence variants, 417 of which were SNPs (188 transversions and 229 transitions), 11 were MNVs and 18 were InDels. We selected 46 of these putative SNPs in D. polymorpha and 40 SNPs in C. fluminea for validation with short amplicon (SA) HRMA assays (Smith et al., 2013) in up to 96 individuals collected for previous studies (Peñarrubia et al., 2015a, 2015b). All the selected positions harbored a single SNP and the length of the amplicon after primer design was limited to less than 65 bp. Melting temperatures of the putative PCR products were pre–checked using uMELTSM v2.0 (https://www.dna.utah.edu/umelt/umelt. html). BLASTN analysis (https://blast.ncbi.nlm.nih. gov/Blast.cgi) was run to identify possible homologies producing non–specific amplifications in the PCR. HRMA amplifications were conducted in 10 μL reactions containing 25–100 ng of genomic DNA, 1× EconoTaq Plus Master Mix (Lucigen), 1× LCGreen+ (Idaho Technology), and 0.2 μM of each prim-

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er. Thermal cycling was performed on a LightCycler 480 Real–Time PCR system (Roche Diagnostics) with an initial denaturation of 10 min at 95 ºC followed by 35 cycles denaturing for 10 s at 95 ºC, annealing at 60 ºC for 30 s, and extension for 10 s at 72 ºC. Reactions were overlaid with 15 μL of mineral oil to ensure that evaporative losses did not affect ionic strength which may affect melting uniformity across samples (Smith et al., 2010). Twenty–five HRMA data acquisitions per ºC were collected with a ramp rate of 0.02 ºC/s between 60 and 95 ºC. All melting curve patterns were analyzed using the LightCycler 480 Gene Scanning Software v. 1.5.0 SP1 (Roche Diagnostics). SA–HRMA characterization of the 46 putative positions in D. polymorpha indicated that five were monomorphic, seven produced patterns that were not consistent with homo– and hetero–duplex curves (e.g., three or more melting peaks, potential primer dimers, etc.), 23 produced heteroduplex curves (i.e., double peaks) indicative of the heterozygous condition in every individual in the sample (n = 15) tested, six failed to amplify even after repeated attempts to optimize PCR (not shown), and five loci (10,87 %) displayed polymorphic SNPs melting curves (table 1). In C. fluminea, four were identified as monomorphic, 10 produced non–scorable melting patterns, 18 generated double–peaks in every individual tested, four failed to amplify, and four SNPs (10 %) produced melting curves of polymorphic SNPs. While 10 % may be considered a low success rate in polymorphic yield (validated/polymorphic loci), our results suggest that a complete validation of the initial 783 and 446 sequence variants in both species would generate an increase in the number of SNPs to 40 and 80 respectively for C. fluminea and D. polymorpha. Those positions with a heterozygous condition in all the first analyses (23 for D. polymorpha and 18 for C. fluminea) were genotyped in 96 individuals of each species, and all of them displayed the same result. This pattern may be explained by paralogous sequence variants (PSV) (Smith et al., 2005). PSV are a common trait of the genomes of mollusks because of the highly abundant cryptic repetitive genomic DNA (McInerney et al., 2011). All the validated SNPs (five in D. polymorpha and four in C. fluminea) and their flanking regions are available in GenBank (accession numbers KT220181–85 for D. polymorpha, and KT220186 and KT220188–90 for C. fluminea, table 1). They are the first validated SNPs in the two species and they add to the limited number of currently available loci (Peñarrubia et al., 2015a, 2015b). They can thus be used in further studies with alternative genotyping techniques. Interestingly, even such a limited number of SNPs can be useful to describe population structure. In the invasive mosquito fish (Gambusia holbrooki), 5 SNPs were used to further characterize the genetic structure in European populations (Vidal et al., 2012), and in swordfish (Xiphias gladius) the same number of SNPs was able to detect genetic differentiation in Atlantic and Mediterranean samples (Smith et al., 2013).

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Table 1. Name, nucleotide variant (NV), amplicon size (As), forward (F) and reverse (R) primers, and annealing temperature (Tª) for polymorphic SNPs in D. polymorpha and C. fluminea. Tabla 1. Nombre, variante nucleotídica (NV), tamaño del amplicón (As), cebador directo (F) e inverso (R), temperatura de unión (Tª) para los PNU polimórficos en D. polymorpha y C. fluminea. Name

Primer sequences (5' à 3')

A/T

F: TGCAACCGAGTTTACCAACGGCT

Ta (ºC)

GenBank Accesion

KT220181

KT220182

KT220183

KT220184

KT220185

KT220186

KT220188

KT220189

KT220190

D. polymorpha Dp292

R: TGCTGTTCAAATGAACCGGAGCAG

Dp367

F: TCGCCTTGCAAGTCTCGTGCT

T/G

R: GCAATTGTTCTTGCAGTAATGTCCCGC

Dp452

F: GCTGCCTGAAACGTTCAGTGGT

T/G

R: CCTCCGGGATCGGCCCACTT

Dp467

F: TGCGTGGAGCCTTTCCACCG

A/G

R: TGGCAAGAACAAAGCAGACCGC

Dp501

F: GTGTGAAATCTTGAAAGCGCCTTGT

C/T

C. fluminea Cf46

R: GGCTGCTGGTAAATAAATGGGCTCCG

C/G

F: CGAAAGCTGCGCATTTCTGCGA

R: ACCTGCGGATGGATCATTACCGA

Cf132

F: TGTAGGCGGCCACCCCATGT

T/G

R: GGTCTTCACTGACGGGCGGC

Cf190

F: AGCTTACAGTTTGCCCACTTACCTCT

T/A

R: AGATGCGAATTGGCCCCGGT

Cf270

F: GTAATGTCCGTCTGCGTATCAGATTCA

T/A

R: TGCCGGGGTGTCTTGTTTGTCG

Acknowledgements We are indebted to the following colleagues for animal samples: R. Miranda (Universidad de Navarra, Spain), J. Checa (Equipos y Suministros S. L., Zaragoza, Spain), I. Planas and R. Puig (Nyctibios Association, Barcelona, Spain), and C. Robertson (Texas Parks and Wildlife Department, TX, USA). This research was carried out within the scope of the research project CGL200909407 of the Ministry of Sciences and Innovation (MICINN) of the Spanish Government. LP received economic support with a PhD fellowship from the Spanish MICINN with reference BES–2010037446. LP also received economic support from the Spanish MICINN for a fellowship with reference EEBB–I–12–05670. References Everett, M. V., Grau, E. D., Seeb, J. E., 2011. Short reads and nonmodel species: exploring the comple-

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xities of next–generation sequence assembly and SNP discovery in the absence of a reference genome. Molecular Ecology Resources, 11: 93–108. Liu, N., Chen, L., Wang, S., Oh, C., Zhao, H., 2005. Comparison of single–nucleotide polymorphisms and microsatellites in inference of population structure. BMC Genetics, 6: S26. McInerney, C. E., Allcock, A. L., Johnson, M. P., Bailie, D. A., Prodöhl, P. A., 2011. Comparative genomic analysis reveals species–dependent complexities that explain difficulties with microsatellite marker development in molluscs. Heredity, 106: 78–87. Montgomery, J., Wittwer, C. T., Palais, R., Zhou, L., 2007. Simultaneous mutation scanning and genotyping by high–resolution DNA melting analysis. Nature Protocols, 2: 59–66. Morin, P. A., Martien, K. K., Taylor, B. L., 2009. Assessing statistical power of SNPs for population structure and conservation studies. Molecular Ecology Resources, 9: 66–73. Peñarrubia, L., Araguas, R. M., Pla, C., Sanz, N., Viñas, J., Vidal, O., 2015a. Identification of 246

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microsatellites in the Asiatic clam (Corbicula fluminea). Conservation Genetics Resources, 7: 393–395. Peñarrubia, L., Sanz, N., Pla, C., Vidal, O., Viñas, J., 2015b. Using Massive Parallel Sequencing for population genetics markers development, validation and application in the invasive bivalve zebra mussel (Dreissena polymorpha). Plos One, 10: e0120732. Reed, H. G., Kent, J. O., Wittwer, C. T., 2007. High– Resolution DNA melting analysis for simple and efficient molecular diagnostics. Pharmacogenomics, 8: 597–608. Seeb, J. E., Pascal, C. E., Grau, D. E., Seeb, W., Templin, W. D., Harkins, T., Roberts, S. B., 2011. Transcriptome sequencing and high–resolution melt analysis advance single nucleotide polymorphism discovery in duplicated salmonids. Molecular Ecology Resources, 11: 335–348. Shahin, A., van Gurp, T., Peters, S. A., Visser, R. G., van Tuyl, J. M., Arens, P., 2012. SNP markers retrieval for a non–model species: a practical approach. BMC Research Notes, 29: 5–79.

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Smith, B. L., Lu, C. P., Alvarado Bremer, J. R., 2010. High–resolution melting analysis (HRMA): a highly sensitive inexpensive genotyping alternative for population studies. Molecular Ecology Resources, 10: 193–196. Smith, B. L., Lu, C. P., Alvarado Bremer, J. R., 2013. Methodological streamlining of SNP discovery and genotyping via high–resolution melting analysis (HRMA) in non–model species. Marine Genomics, 9: 39–49. Smith, C. T., Elfstrom, C. M., Seeb, L. W., Seeb, J. E., 2005. Use of sequence data from rainbow trout and Atlantic salmon for SNP detection in Pacific salmon. Molecular Ecology, 14: 4193–4203. Vidal, O., Sanz, N., Araguas, R. M., Fernandez–Cebrián, R., Díez del Molino, D., García–Marín, J. L., 2012. SNP diversity in introduced populations of the invasive Gambusia holbrooki. Ecology of Freshwater Fish, 21: 100–108. Wittwer, C. T., Reed, G. H., Gundry, C. N., Vandersteen, J. G., Pryor, R. J., 2003. High– resolution genotyping by amplicon melting analysis using LCGreen. Clinical Chemistry, 49: 853–860.

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Forum 69

Animal Biodiversity and Conservation 42.1 (2019)

Reconsidering mammal extinctions in the Pernambuco Endemism Center of the Brazilian Atlantic Forest: a critique A. R. Mendes Pontes, A. C. Mariz Beltrão, A. M. Melo Santos

Mendes Pontes, A. R., Mariz Beltrão, A. C., Melo Santos, A. M., 2019. Reconsidering mammal extinctions in the Pernambuco Endemism Center of the Brazilian Atlantic Forest: a critique. Animal Biodiversity and Conservation, 42.1: 69–77, Doi: https://doi.org/10.32800/abc.2019.42.0069 Abstract Reconsidering mammal extinctions in the Pernambuco Endemism Center of the Brazilian Atlantic Forest: a critique. This is a reply to the critique made by Garbino et al. (2018) to our article (Mendes Pontes et al., 2016) in which we revealed an unprecedented mass extinction event in the Pernambuco Endemism Center (CEPE) and with which they disagreed. Here we critically review their arguments, and present incontrovertible evidence that the processes presented in our 2016 paper are real events. Additionally, we discuss the importance of providing up–to–date scientific data to prove the existence of a species, and the critical importance of historical records in formulating a better understanding of the mammalian diversity of the CEPE. We point out that a more rigorous approach towards historical and recent records is needed when producing checklists of CEPE mammals, given that ignoring evidence and allowing personal opinion to prevail may lead to loss of credibility and jeopardize conservation efforts. Key words: Mass extinction, Pernambuco Endemism Center, Atlantic forest of northeastern Brazil, Medium– sized and large mammals, Conservation Resumen Reconsiderando las extinciones de mamíferos en el Centro de Endemismo de Pernambuco del bosque atlántico del Brasil: una crítica. Presentamos la respuesta a la crítica formulada por Garbino et al. (2018) a nuestro artículo (Mendes Pontes et al., 2016), en el que revelamos una extinción en masa en el Centro de Endemismo de Pernambuco (CEPE), de la que discreparon. En este artículo examinamos sus argumentos de forma crítica y exponemos pruebas irrefutables de que los procesos presentados en nuestro artículos de 2016 son acontecimientos reales. Asimismo, analizamos la importancia de aportar datos científicos actualizados para demostrar la existencia de una especie y la trascendencia de mantener registros históricos para comprender mejor la diversidad de mamíferos del CEPE. Señalamos la necesidad de abordar de una manera más rigurosa los registros históricos y recientes a la hora de confeccionar listas de comprobación de los mamíferos del CEPE, ya que pasar por alto las pruebas y dejar que prevalga la opinión personal puede conllevar una pérdida de credibilidad y poner en peligro las iniciativas de conservación. Palabras clave: Extinción en masa, Centro de Endemismo de Pernambuco, Bosque atlántico del nordeste del Brasil, Mamíferos de talla mediana y grande, Conservación Received: 30 I 18; Conditional acceptance: 16 III 18; Final acceptance: 23 VII 18 Antonio Rossano Mendes Pontes, Instituto Nacional de Pesquisas da Amazônia–INPA, Núcleo de Pesquisas de Roraima, Rua Coronel Pinto, 315, Centro, Boa Vista, Roraima, RR, Brazil.– Antonio Carlos Mariz Beltrão, Universidade Federal de Pernambuco, Centro de Ciências Biológicas, Departamento de Zoologia, R. Prof. Moraes Rego 1235, Cidade Universitaria, Recife, Pernambuco, PE, Brazil.– André Maurício Melo Santos, Universidade Federal de Pernambuco, Núcleo de Biologia–Centro Acadêmico de Vitória, UFPE, Rua do Alto do Reservatório s/n., Bela Vista, Vitória de Santo Antão, PE, Brazil. Corresponding author: A. R. Mendes Pontes. E–mail: mendespontes@gmail.com ISSN: 1578–665 X eISSN: 2014–928 X

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General considerations In reply to the critique made by Garbino et al. (2018) to our article (Mendes Pontes et al., 2016) we here critically review their arguments regarding three issues. First, Garbino et al. (2018) repeatedly use the word 'locally' to refer to regional species extinctions in the Pernambuco Endemism Center. They repeatedly use the word 'locally' while simultaneously acknowledging that the species are, in fact, extinct in the entire region, which per se would be characteristic of a 'regional extinction' (linear distance 200–2,000 km; Peterson et al., 2011). In other words, if, for example, jaguars (Panthera onca) are confirmedly extinct in the CEPE, this means a total decrease in their former geographic range of ~56,000 km2, which was the original size of the CEPE. What's more, if we consider them extinct in both the CEPE (this study) and in the Bahia Endemism Center (CEBA) (Canale et al., 2012), this means a total decrease of almost 90,000 km2 in the species' former geographic range. Such figures highlight the critical importance of understanding the true dimension of these regional extinctions for effective species conservation. On correctly estimating the extent of the anthropogenic destruction of the CEPE: the detrimental use of old figures/references by Garbino et al. (2018) rather than the more recent data used by Mendes Pontes et al. (2016): Citing Ribeiro et al. (2009), Garbino et al. (2018) says that approximately 12 % of the original vegetation currently remains in the CEPE. Although we recognize the importance of the study by Ribeiro et al. (2009) and, indeed, referred to this in our 2016 study, the rate of deforestation in the CEPE is both extremely rapid and unceasing (Melo, 2009), and in the intervening near–decade the scenario will without doubt have changed for the worse. In consequence, the data used in our study (Mendes Pontes et al., 2016) were more recent and based on up–to–date satellite images. This more recent evaluation reveals that only 5.6 % of the original forest remained at the time of the publication (a total loss of 94.4 % of the original ~56,000 km2 of pristine forest), a figure that was not considered by Garbino et al. (2018), but that is a bad enough scenario to have an extremely negative effect on conservation initiatives. On the unquestionable validity of the old literature and paintings of the 16th and 17th century's first colonists of the Atlantic coast of northeastern Brazil, when analysing the composition of the former mammalian fauna in the CEPE (Gandavo, 1980; Salvador, 1975; Marcgraf and Piso, 1942; Falcão, 1964; Brandão, 1980; Vasconcelos, 1981; Marcgraf, 1995): Garbino et al. (2018) sometimes accept (e.g. the case of the blond capuchin, Sapajus flavius), sometimes refute (e.g. the case of the deer, Mazama spp.), and sometimes seem not to have had access to these above mentioned invaluable references (e.g. the case of the bush dog, Speothos venaticus). Such historical materials are invaluable testimonies of the original

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condition that these highly skilled and observant 16th and 17th century colonists/artists/naturalists found in the Atlantic coast of northeastern Brazil, in what are now the states of modern–day Pernambuco, Paraíba, Alagoas, and Rio Grande do Norte. Such valuable historical documents have been used in various regions to validate the occurrence and/or extinction of species, such as in West Indies, where several species of macaw were reported and depicted, and from which scientists became aware of their extinction (Wiley and Kirwan, 2013). But we do not have to go too far in order to find appropriate examples of the contribution of these colonists to current taxonomy and conservation. Marcgraf and Piso (1648) described what is evidently a curassow of the genus Crax with a yellow beak, which has never been recorded in the Atlantic forest of northeastern Brazil. These plates and paintings were considered unambiguous evidence of the disappearance of an undescribed species. The importance of their works is incommensurable and greatly important for our better understanding of the mammalian fauna originally in the CEPE when they first arrived. In fact, they are the only reference that we have of the occurrence of most large mammals (Jaguars, pumas, Puma concolor, tapirs, Tapirus terrestris, white–lipped peccaries, Tayassu pecari, giant ant–eater, Myrmecophaga tridactyla, among others) that were made from direct observation of the species in the wild, since they went extinct before contemporary scientists could record them. All the other references to the CEPE is via extrapolation (see Mendes Pontes et al., 2016). On the allegedly faux absences Leopardus wiedii Garbino et al. (2018) state: "Feijó and Langguth (2013) mention the margay, Leopardus wiedii, in two localities in PEC: Roteiro (Alagoas), and Alhandra (Paraiba)." Feijó and Langguth (2013) is, therefore, the reference supposedly to prove the occurrence of the species in the CEPE, but this is a reference entirely based on museum specimens (dead individuals) that have been deposited in the scientific collections and that per se does not prove that the species is still extant in the CEPE. Cabassous unicinctus Garbino et al. (2018) state: "The naked–tailed armadillo, whose species occurring in CEPE is Cabassous tatouay, not C. unicinctus as reported by Mendes Pontes et al. (2016), was recorded, based on voucher specimens, in two localities in CEPE (Feijó and Langguth, 2013)." Again, Feijó and Langguth (2013) is, therefore, the reference supposedly to prove the occurrence of the species in the CEPE, but this is a reference entirely based on museum specimens (dead individuals) that have been deposited in the scientific collections and that 'per se' does not prove that the species is still extant.

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The species has not been detected recently through proper surveys / census, or any other method that prove that the species is still extant (e.g. camera trap). Our assertion that C. unicinctus is extinct in the CEPE is therefore well supported. Cyclopes didactylus Garbino et al. (2018) assert that "There have been several recent records of the pygmy anteater, Cyclopes didactylus, for the region (Gardner, 2008; Miranda and Superina, 2010; Feijó and Langguth, 2013)." However, the use of records in Feijó and Langguth (2013) is problematic for Cyclopes didactylus since, as with previous examples, the records in the CEPE, are based solely on museum specimens, that is to say, dead individuals. Regarding the "Several recent records of the pygmy anteater" mentioned by Garbino et al. (2018), the most recent of these records is Miranda and Superina (2010), who collected their information through interviews and questionnaires in the five years prior to the publication of their study, that is to say, between 2005 and 2009 (11 years before our study (Mendes Pontes et al., 2016), in which we surveyed old and recent literature, and also carried out recent field surveys). Thus, there is more than a ten–year delay between the last opportunistic records of Cyclopes didactylus in the CEPE and the systematic line transect surveys carried out by our group. In a highly devastated scenario such as the CEPE, where deforestation and hunting is rampant and uninterrupted –while simultaneously being almost totally ignored by authorities– this is more than enough to have caused the species to have gone extinct. Our results are merely the confirmation of what should be expected in such a neglected hotspot. Alouatta belzebul Garbino et al. (2018) says: "The red–handed howler monkey, Alouatta belzebul, which is considered Vulnerable by the IUCN Red List, is still widespread at CEPE (Fialho et al., 2014)." Fialho et al. (2014) carried out literature surveys, questionnaires, and also field inventories, and could confirm the presence of Alouatta belzebul in five localities in the CEPE. Their inventories, nevertheless, took place between 2006 and 2009, around 12 years ago, which per se do not prove that the species is still extant. Recent field surveys, nevertheless, revealed a remnant population in the State of Pernambuco, which has been systematically studied (Campelo et al., 2015; Silva, 2015; Silva et al. 2015a, 2015b), which is a valid evidence of the existence of the species in the wild. This species, therefore, should be removed from the list of extinct species of the CEPE. Lontra longicaudis Garbino et al. (2018) affirm that: "Nevertheless, the Neotropical otter, L. longicaudis, was recently recorded in nine localities in CEPE by Astúa et al. (2010), Feijó and Langguth (2013) and Toledo et al. (2014)."

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Although we recognize the importance of the sightings of Lontra longicaudis that have been made over the last three decades in the CEPE (Percequillo et al., 2007; Astúa et al., 2010), it is almost a decade since that the last record was made (Astúa et al., 2010), and, as previously stated, they do not prove that the species is still present. Thus, until further extensive field surveys are carried out and direct evidence of the existence of the species is made (e.g. direct sightings; camera–trap), it is reasonable to think that the species is extinct. On the allegedly faux presences Tolypeutes tricinctus Garbino et al. (2018) state: "The open–area dweller three–banded armadillo, Tolypeutes tricinctus, do not occur naturally in the CEPE. The distribution of the genus Tolypeutes was recently revised based on interviews, direct observations, fossil, historical and recent records up to 2013 (Feijó et al., 2015), and all 168 records of Tolypeutes tricinctus were restricted to the Caatinga scrubland (Brazilian ecosystem adjacent to PEC) and Cerrado savanna of northeastern Brazil." Feijó et al. (2015), on which their arguments are based, nevertheless, recognize that: "Until now there has been no systematic mapping of the known localities, nor any reliable analysis of possible zoogeographic barriers."; "The distributional areas of T. tricinctus have been poorly surveyed, and this problem is exacerbated by the ongoing extinctions of local populations."; "Excluding the indirect records and the fossils, only 27 (16%) localities have been recorded reliably over the past 104 years."; "The geographic distribution of T. tricinctus is even less well defined than that of T. matacus... ...with extensive lacunas in peripheral areas, impeding a more conclusive interpretation of possible barriers to dispersal."; "Known localities are predominantly within the domain of the Caatinga scrub, adjacent Cerrado savanna, which suggests a preference for open and / or semi–arid habitats."; Thus, the fact that the distributional range of T. tricinctus, as given by Feijó et al. (2015), is currently limited to the east by the Atlantic forest biome could well be a sampling issue, one which is exacerbated by the ongoing and historical mass extinction processes within the CEPE (Mendes Pontes et al., 2016). We assumed that T. tricinctus once occurred in the Atlantic forest of the CEPE in Mendes Pontes et al. (2016) because it is referred and also depicted by Marcgraf in 1648, in pg. 33 of Libri Principis (Marcgraf and Piso, 1942) (fig. 1). Marcgraf says: "These are armored animals and they are able to pull in the head and paws, doubling themselves into a ball", and the only armadillos capable of rolling into a ball are those of the genus Tolypeutes (Eisenberg and Redford, 1999). Marcgraf worked specifically on the Atlantic coast of Pernambuco, describing and illustrating the

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local fauna, and producing an undeniably unique contribution to the knowledge of the former mammalian fauna of the CEPE. We believe that T. tricinctus may have been extirpated from the Atlantic forest of the CEPE before contemporary scientists could detect them, but point to the evidence in the illustrations of earlier authors. To ignore the reference and plate that Marcgraf provided of T. tricinctus in the CEPE and simply state that there is a 'lack of evidence' (Garbino et al., 2018) regarding their occurrence in the region is to ignore this key evidence and let personal opinion and hypothesis prevail over evidence. Conepatus semistriatus Although included by Mendes Pontes et al. (2016), Garbino et al. (2018) removed the striped hog–nosed skunk, Conepatus semistriatus (listed as Conepatus amazonicus) from their CEPE check list, citing lack of evidence. We contest this removal. Both Emmons and Feer (1997) and Kipp (1965) believe the species (whether treated as C. semistriatus or C. amazonicus) occurs in the CEPE region, and that it is not restricted to dry, arid environments. Emmons and Feer (1997) reported that they occur in tropical rainforests, secondary and disturbed forests, clearings, and even gardens. Thus, given their absence in our surveys, we assume that C. semistriatus (C. amazonicus) once occurred in the PEC, but has now been wiped out from the region.

occurred in the CEPE. They point out that occurrence records of the lesser long–nosed armadillo are very scarce for northeastern Brazil, and that of the states that comprise the PEC, the species has only been recorded for the Caatinga of Pernambuco (Feijó and Langguth, 2013). However, the mammal fauna of the CEPE remained virtually unknown for more than 500 years, and therefore, rarity and lack of records cannot be used as a reliable surrogate for absence. Accordingly, we follow Eisenberg and Redford (1999) and consider the species as one that was (at least originally) present in the CEPE. That all records from Feijó and Langguth (2013) are concentrated in the dry scrub Caatinga of Pernambuco is, we consider, likely to be a mere reflection of collection efforts. Mazama spp. EX BK (extinct before known) (Mendes Pontes et al., 2016).

In addition to the Conepatus, Garbino et al. (2018) also consider that the bush bog, Speothos venaticus, has no records within the CEPE. This assertion is made despite the evidence presented by Mendes Pontes et al. (2016) that the species was mentioned in historical documents from the 16th and 17th centuries. While it is true that there are no known recent historical or extant records of S. venaticus for CEPE (see Feijó and Langguth, 2013; Fernandes–Ferreira, 2014), earlier records certainly exist. In the third paragraph of page 223, book IV, of History of Quadrupeds and Serpents, in Historiae Naturalis Brasiliae (Marcgraf and Piso, 1942), a detailed description is made of a cachorro–do–mato (bush–dog). In this case, Garbino et al. (2018) seems to have overlooked this historical record. But based on this evidence, the bush dog once occurred in the CEPE and has subsequently become extinct. As Garbino et al. (2018) themselves point out, the potential occurrence of the bush dog in CEPE has been suggested by a study that inferred the habitat suitability for the species through ecological niche modelling (DeMatteo and Loiselle, 2008).

Garbino et al. (2018) state: "Mendes Pontes et al. (2016) reported two supposedly undescribed species of Mazama for CEPE. However, there is neither historical nor current evidence that there existed another species of Mazama in the CEPE besides Mazama guazoubira (Feijó and Langguth, 2013), although it is currently extinct there." As stated by Lees and Pimm (2015), the 1648 Historia Naturalis Brasiliae by Georg Marcgrave and Willem Piso (Marcgraf and Piso, 1942), represented a pioneering attempt to catalogue the vast biodiversity of the Atlantic coast of northeastern Brazil. Its pages contain compelling evidence for historical extinctions. The same can be said of the works of Zacharias Wagener, who also worked in the CEPE between 1634 and 1641, when the Dutch ruled Pernambuco (Falcão, 1964). Both books described and depicted two different species of deer from the CEPE that have never been seen by contemporary scientists. As no material was collected, these texts and plates become the only evidence of the existence of these two species. It is important to note that these two deer species had features distinguishing them very clearly from any extant species. The plate in Marcgraf and Piso (1942: see fig. 2) depicts an adult deer with a gray head and neck, and red body. The other, although the plate is in black and white, shows an adult deer with conspicuous white spots along the belly (Falcão, 1964: see fig. 3). Thus, we conclude that these two deer species went extinct without being recorded by contemporary scientists. We proposed the category EX BK 'Extinct Before Known' in Mendes Pontes et al. (2016) to distinguish it from other IUCN categories and criteria, which do not include species that went extinct before becoming known to recent scientists and being scientifically described by them.

Dasypus septemcinctus

Ateles sp. EX BK

Garbino et al. (2018) also challenge the proposal by Mendes Pontes et al. (2016), that the lesser long–nosed armadillo, Dasypus septemcinctus, once

Garbino et al. (2018) state: "The authors based the presence of a spider monkey (Ateles sp.) in northeastern Brazil in a Portuguese translation of the work of

Speothos venaticus

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Fig. 1. Tolypeutes tricinctus from Marcgraf and Piso (1942), plate 680. Fig. 1. Tolypeutes tricinctus de Marcgraf y Piso (1942), ilustración 680.

Caspar van Baarle, (latinized as Caspar Barlaeus in publications), about the Dutch possessions in Brazil, made by Claudio Brandão (Barlaeus, 1940). Brandão stated in a translation note that the name Cajatayae, used by Barlaeus to describe a long–tailed reddish monkey, was similar to the word Coatá, the name commonly used for the spider monkeys, genus Ateles." "Marcgrave (1648), however, described a monkey called Caitaia, a name that resembles Barleus's Cajatayae. The animal described by Marcgrave as Caitaia has been considered as Sapajus flavius, a capuchin monkey still extant in PEC, especially due to the reference to the yellowish color of its pelage (Oliveira and Langguth, 2006). Moreover, cay or cai, as in caitaia, is the name of capuchin monkeys in the indigenous language Tupi–Guarani." "Recent re–discovery of animal drawings made by the artist Frans Post in the Dutch Brazil area revealed an Ateles–like monkey among the depicted fauna (De Bruin, 2016). It is very improbable that naturalists such as Georg Marcgrave or Wilhelm Piso would have failed to detect a population of large–bodied spider monkeys in northeastern Brazil (Marcgrave, 1648)." "A more probable explanation is that the animal illustrated was obtained elsewhere, as transportation of primates was common in the colonial Americas (Browne, 1789; Teixeira and Papavero, 2010)." Scientific evidence from Marcgrave (1648), Barlaeus (1940), and De Bruin (2016), nevertheless, reveals that Barlaeus and Frans Post are undeniably referring to a spider monkey, genus Ateles, in the CEPE, whereas Marcgraf, in turn, is referring to a capuchin monkey, genus Sapajus. Personal opinions and interpretations should only add confusion and jeopardize the efforts to unveil the true biological diversity of this unique region. The discovery of Frans Post’s spider monkey (fig. 4) is the confirmation that spider monkeys may once have lived in the CEPE, but were extinct, as was the case also with so many other species known and unknown. Furthermore, we do not believe that "the animal illustrated was obtained elsewhere, as transportation of primates was common in the colonial Americas

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(Browne, 1789; Teixeira and Papavero, 2010) (Garbino et al., 2018)", because in the referred works it was very common that whenever a described/depicted animal was from elsewhere, that they mentioned its origin, such as is the case in many plates by Marcgraf and Wagener, in which they refer to them as from Angola, Mozambique, and Guinea (Marcgraf and Piso, 1648; Falcão, 1964). Saimiri sp. EX BK Garbino et al. (2018) say that: "According to Mendes Pontes et al. (2016), Pero de Magalhães Gandavo (Gandavo, 1924) and the Franciscan friar Vicente do Salvador (Do Salvador, 1889) mentioned the presence of squirrel monkeys, Saimiri sp., in the CEPE. In both works, there is no reference to animals morphologically similar to squirrel monkeys, and, more importantly, according to Capistrano de Abreu (in Gandavo, 1924), Gandavo never visited Pernambuco." "We believe that Mendes Pontes et al. (2016) assigned some monkeys described in Gandavo (1924) and in do Salvador (1889) to Saimiri due to the characteristic odor, mentioned by the two Portuguese authors. Probably, Mendes Pontes et al. (2016) associated the presence of odoriferous glands with the common name of Saimiri in Portuguese, 'mico–de–cheiro', which means 'monkey with odor'. However, the presence of scent glands is widely distributed in New World monkeys (Perkins, 1975; Heymann, 2006). Marcgrave (1648, p. 227), for example mentions a 'musky odor' for Sapajus flavius. The genus Saimiri is endemic from the Amazon basin and Central America (Groves, 2001), and therefore squirrel monkeys were not historically present in the PEC." We referred to Saimiri as occurring in the CEPE because, according to Barlaeus (Brandão, 1980), the not–so–large squirrel monkey that Gandavo (1980) and Salvador (1975) refer to is, in fact, a Saimiri sciureus (called by them jurupixuma, or Saimiris sciurea), which has yellow–olive fur and a very long tail. Thus, we believe that this is an unequivocal reference to squirrel monkeys in the CEPE.

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Fig. 2. A species of cf Mazama sp. depicted by Marcgraf (Marcgraf and Piso, 1942) in the CEPE. Fig. 2. Una especie de Mazama sp. dibujada por Marcgraf (Marcgraf y Piso, 1942) en el CEPE.

Sapajus (Cebus) apella Garbino et al. (2018) say: "The equivocal inclusion of 'Cebus apella' (a species that is now classified in the genus Sapajus) in CEPE fauna by Mendes Pontes et al. (2016) apparently has a simpler solution. Using a now–outdated taxonomy of capuchin monkeys, Hershkovitz (1987, p. 23) mentions 'Cebus apella libidinosus' among the mammals described by Marcgrave in CEPE."

In fact, as stated in Mendes Pontes et al. (2016), we included Sapajus apella among the species referred to the CEPE because they occur in the contiguous dry–scrub caatinga forests, and there are no physical barriers to their occurrence in the CEPE. Thus, we hypothesized that the species might once have been sympatric with Sapajus flavius. We, nevertheless, acknowledge that it is a reference based solely on a theory, and as such is prone to criticism.

Fig. 3. A species of cf Mazama sp. depicted by Wagner (Falcão, 1964) in the CEPE. Fig. 3. Una especie de Mazama sp. dibujada por Wagner (Falcão, 1964) en el CEPE.

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Fig. 4. Franz Post's spider monkey, Ateles sp. EX BK from the Pernambuco Endemism Center, Atlantic forest of northeastern Brazil. Fig. 4. Mono araña de Frans Post, Ateles sp. (extinto antes de conocerse) del Centro de Endemismo de Pernambuco, en el bosque atlántico del nordeste del Brasil.

Final considerations This paper shows that the species list in Mendes Pontes et al. (2016) is not an overestimation, as argued by Garbino et al. (2018), but rather accurately reflects the status of the mammalian fauna of the CEPE in the 21st century. Our study was the first systematic field survey of the CEPE (besides a thorough systematic literature survey) to provide robust scientific data based solely on species presence, recorded through direct sighting of the individuals. Although museum specimens are an important source of biological information (e.g. Feijó and Langguth, 2013), they cannot be used as proxy for the current and continued occurrence of a species in an area, since the species may have been long extinct at that collection site. We entirely agree with Garbino et al. (2018) on the pervasive consequences of taxonomic errors, false assumptions, and unreliable records for conservation and species management. That is why we based our identification of an Ateles sp. in the CEPE on the available evidence. And that is also why we never made assumptions that could not be validated from the literature, such as "It is very improbable that naturalists such as Georg Marcgrave or Wilhelm Piso would have failed to detect a population of large–bodied spider monkeys in northeastern Brazil" (Garbino et al., 2018), or use museum specimens to show that a species is still locally extant (Feijó and

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Langguth, 2013; Garbino et al., 2018). A crucial point that seems to have escaped Garbino et al. (2018) is that a species may be present in a museum drawer, yet no longer exist in the wild. Finally, the mass extinction process that has swept the CEPE does not benefit anyone, and it would be of incommensurable value to conservation that these extinct species were, in fact, re–discovered. In order to achieve that, however, scientists have to rely on efficient methods, such as line transects and camera traps, and not on past publications, or museum specimens. Acknowledgements We thank Dr. Alexander de Bruin, at Noord–Hollands Archief, Haarlem, inventory number: 53004662, for kindly granting the high resolution image of the spider monkey, Ateles sp. EX BK from the Pernambuco Endemism Center. We also thank Dr. Adrian Barnett for English correction and also for very useful comments on the manuscript. References Astúa, D., Asfora, P. H., Aléssio, F. M., Langguth, A., 2010. On the occurrence of the Neotropical Otter (Lontra longicaudis) (Mammalia, Mustelidae)

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People and protected areas: some issues from India J. S. Maan, P. Chaudhry

Maan, J. S., Chaudhry, P., 2019. People and protected areas: some issues from India. Animal Biodiversity and Conservation, 42.1: 79–90, Doi: https://doi.org/10.32800/abc.2019.42.0079 Abstract People and protected areas: some issues from India. India is one of the 17 mega biodiverse countries, occupying only 2.5 % of the world's geographical area and 1.8 % of the its forest area but supporting 16 % of the world’s human population and 17 % of its livestock population. Biotic pressure on the country's protected areas is tremendous and managers of these areas face an uphill task in balancing divergent needs of different stakeholders of national parks and wildlife sanctuaries. The job of managing such areas is highly challenging because of the many difficult issues such as human–wildlife conflicts, encroachments, overgrazing, tourists' pressure (including pilgrimages into the forests), poaching, and an ever–increasing demand for diversion of protected areasfor development purposes. In the present article we discuss some of these issues with reference to India and emphasise the danger of losing ecosystem services (mostly of an intangible or regulating kind of nature) emanating out of these protected areas. Key words: Protected area, Biodiversity conservation, Ecosystem services, Human–wildlife conflict, Ecotourism, Tiger Reserve Resumen La población y las zonas protegidas: algunos problemas en la India. La India es uno de los 17 países con más biodiversidad, ocupa solo el 2,5 % de la superficie del mundo y el 1,8 % de la superficie forestal mundial, y alberga el 16 % de la población humana y el 17 % del número de cabezas de ganado del mundo. La presión biótica en las zonas protegidas del país es tremenda y los gestores de estas zonas se enfrentan a la tarea cada vez más ardua de encontrar un equilibrio entre las necesidades divergentes de las diferentes partes interesadas de los parques nacionales y las reservas naturales. Existen numerosos problemas, como los conflictos entre humanos y la fauna silvestre, las invasiones, el pastoreo excesivo, la presión turística (con inclusión de los peregrinajes a los bosques), el furtivismo o la creciente demanda de zonas protegidas con fines de desarrollo, que dificultan la labor de gestión de estas zonas. En el presente artículo, hemos analizado algunas de estas cuestiones con referencia a la India, a la vez que se hace más hincapié en el peligro que supone perder los servicios ecosistémicos (en su mayoría, de carácter intangible o regulador) que se derivan de estas zonas protegidas. Palabras clave: Zona protegida, Conservación de la biodiversidad, Servicios ecosistémicos, Conflicto entre humanos y la fauna silvestre, Ecoturismo, Refugio del tigre Received: 21 VIII 17; Conditional acceptance: 15 V 18; Final acceptance: 25 VII 18 Jaideep Singh Maan, Pradeep Chaudhry,Indian Inst. of Forest Management, Nehru Nagar, 462003 Bhopal, India. Corresponding author: Jaideep Singh Maan. E–mail: jdmaan95@gmail.com

ISSN: 1578–665 X eISSN: 2014–928 X

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Introduction For existence and sustainability of life on mother earth, the foremost requirement is to conserve and maintain the equilibrium necessary for the environment, without which life on this planet will simply not be possible. Be it air, water, soil, forests, wetlands or the mountains, their collective values are undoubtedly of great significance to us. So too are the protected areas that play a vital role in the maintenance of life support systems. They are the cornerstones of biodiversity conservation and their importance can be understood in terms of their natural, ecological and cultural values. India is one of the 17 mega biodiverse countries (Mittermeier and Mittermeier, 2005). The country occupies just 2.5 % of the world's geographical area butit supports 16 % of the world's human population and 17 % of the livestock population (Mukerji, 2003; Singhal et al., 2003). India has more than 45,000 floral and 91,000 faunal species in a geographical area of 329 million ha (Reddy et al., 2016). Per capita availability of forest and productivity are among the lowest in the world, and the immense biotic pressure on the country's forests is therefore making biodiversity conservation a very challenging task. Protected areas in the form of national parks, wildlife sanctuaries, conservation reserves and community reserves in this thickly populated country are like an oasis in a desert. Managing protected areas in a democratic, large and densely populated country like India is nothing less than walking on a tight rope. The management of such areas faces constant challenges and difficulties due to issues such as e human–wildlife conflicts, encroachments, overgrazing, tourists' pressure (including pilgrimages into the forests), poaching, running of vehicular and rail traffic through these areas, and the ever–rising demand for diversion of more land in protected areas for development purposes. We discuss some of the above issues in the present article with reference to India while giving greater emphasis to the dangers of losing ecosystem services, such as the impact on water quality and aquatic fauna emanating from these protected areas. Present status There are 769 protected areas in India spread over an area of 162072.49 km2 and covering 4.93 % of the country (table 1). These protected areas are national parks, wildlife sanctuaries, conservation reserves and community reserves. In order to sustain the social, economic and cultural values of these areas, proper management is necessary. Though values are derived from these areas in many forms, it is also important to provide input in the form of proper funds. As for the budget (for 2018–2019), 2,350 million rupees (INR) have been allotted under the section 'Environment Protection, Management and Sustainable Developmen' (Ministry of Environment, Forests and Climate Change, 2018). Still, proper maintenance and evaluation of management will further help in this aspect. In India, a proper

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framework is in place to evaluate the effectiveness of management of protected areas. On the basis of this framework, protected areas in the country are being awarded ratings. The people residing near or inside these areas also play an important role in conservation of biodiversity in protected areas. The issue of 'relocation' Conservation and displacement are closely related to each other. Conservation may also lead to displacement of local people. Tension lies between human presence in or near the protected areas and the success of conservation measures. Sometimes, management objectives translate into practices that ultimately result into displacement. Lack of evidence, which tells about the extent to which the rehabilitation practices have succeeded, exemplifies this tension. Displacement may involve use of force, and may also result in impoverishment, political and social cut off, and disempowerment. The protected area may become inaccessible to local people, and so basic amenities may not be available. Livestock, too, may become negatively affected as grazing is banned in a protected area. Since the main aim of displacement is conservation, it is very important to know to what extent the conservation targets were met after displacement. It is quite difficult to know the exact effect of displacement on conservation. The balance between human costs and conservation benefits is important to maintain. Displacement activities can solve a dual purpose: proper conservation of natural resources and better living conditions for people living in the forest (Agarwal and Redford, 2009). In a study carried out in the Sariska Tiger Reserve of Rajasthan state, it was found in 2004 that the population of tigers was less than 10. It was suspected that poachers were responsible for this situation with some assistance from villagers, so the reserve was soon closed for further investigation. After the investigation and research were completed, the decision was made to relocate 11 villages situated in the reserve. The main aim was to create a 'people–free zone' in Sariska. By 2005 it was clear that a substantial proportion of Sariska had degraded to such a level that it was unable to support any mammalian prey or predator species. The connection between forest use and biodiversity decline was found to be highly complex. Most of the extractive pressure was due to the adjacent urban centres. However, no attempt was made to compare the effects caused by biomass extraction from the towns with the resident villages. Ecosystem dynamism and the human use issues are complex. In one rare case, human use has favoured biodiversity. For example, in the Kanha Tiger Reserve of Madhya Pradesh, it was found that the existence of villagers inside forests resulted in the formation of open grasslands. Thus, herbivores achieved an advantage due to the availability of forage. After the village was relocated, such formations were managed using fire and cutting techniques to sustain the biodiversity values. Thus, it shows that some form of human

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Table 1. Protected Areas in India (as on January, 2018) (www.wiienvis.nic.in) Tabla 1. Zonas protegidas en la India (en enero de 2018) (www.wiienvis.nic.in).

Number

Total area (km2)

Coverage % of country

National parks

104

40,501.03

1.23

Wildlife sanctuaries

544

118,931.80

3.62

Conservation reserves

2,594.03

0.08

Community reserves

72.61

0.002

Protected areas (total)

771 162,099.47

interference may actually favour animal biodiversity. However, on the negative side, large scale biomass extraction will impact the biodiversity in a negative way. So, under controlled extraction conditions, socio– economic and institutional conditions can be created under which the non–destructive extraction of biological resources can be carried out. However, extraction must be sustainable in nature. Some years ago, the Baigas tribe of Kanha Tiger Reserve was displaced because this tribe was considered highly destructive to the regeneration of Sal trees due to slash and burn agriculture. However, they did not make any move to settled agriculture. Also, they could not become paid labourers. They were impoverished. Some time later, the village Supkhar was relocated from Kanha, in a well–coordinated relocation. In Gir, 500 families of buffalo–herding Maldharis tribe were moved out of the core area, leading to a decrease in the number of cows predated by lions. After this, displacement became an important objective for park managers. The case of Bhadra tiger reserve is quite different. Here, relocation helped in securing fertile agriculture land for the displaced people (Rangarajan and Shahabuddin, 2006). It was thus a success move both for the residents and the authorities. Failed displacement plans have something in common. The main cause for their failure is the lack of adequate provision of technical and financial input for successful agriculture livelihoods. In addition, a cooperative environment with collaborative and sincere efforts is necessary for success in these ventures. Thus, the extent to which relocated people are satisfied with relocation is a point of concern. In one resettlement process carried out in Bhadra Wildlife Sanctuary and Tiger Reserve of Karnataka state relocation had mixed results. As a part of this study, a survey of 60% households was conducted in 2002. Later, in 2006, after the relocation of 11 villages, another survey was conducted, including 55 % of households (Karanth, 2007). Prior to resettlement, people were facing challenges such as crop loss and livestock predation. Human–wildlife conflicts weren not infrequent. People also face difficulties such as shortages of drinking water, electricity supply, and medical care. The resettlement scenario arose in 1970s when people residing inside or near the forest

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4.93

requested government help to relocate and resettle. Then, in 1974, a resettlement project was proposed by the Karnataka government. Funds were protected by the Project Tiger Steering committee between 1992 and 2002. In October 2001, 419 families were moved. Overall, the Bhadra resettlement process had a positive result. Households obtained access to electricity, drinking water, phones, solar lights, etc. A health care centre and nursery were also established. Later, people started earning from multiple sources (Karanth, 2007). There were also some shortcomings . When people lived inside the forest they had an abundant supply of firewood, non–timber forest products and grazing land. However, after the resettlement there was limited access to these forest resources. Also, not all the people who were resettled were satisfied with the plots given to them. Some faced problem with the plot size, and some found the plots less fertile. People also reported problems of living with other residents in that area. From the conservation point of view, it was found that the forest area disturbed due to human activities recovered to some extent. More forage was available for wild animals after the resettlement. Poaching and fishing activities were drastically reduced. Thus, it can be seen that on one hand there were some people who were satisfied by the resettlement process, and on the other hand there were quite a few who were not completely satisfied. While people had access to new set of resources, they had to leave some natural resources like forage and fuel wood (Karanth, 2007). The 'compensation' issue People living in close proximity to protected areas face a major problem of conflict with wildlife. It becomes troublesome for them to protect themselves, their crop and their livestock from the wild animals. In order to cater for these losses, the government runs many compensation schemes for the people. The ground reality of people receiving the proper compensation may vary in different areas. Many studies have been carried out in different national parks across the country and brought different results. By and large, despite heavy losses for those living in and around protected

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areas, few households apply for compensation in the true sense. General explanations for this type of attitude include inadequate remuneration, processing delays, corruption and red tape (Ogre and Badola, 2008). Only 31 % households in five protected areas (PAs) located in Western Ghats of India received compensation (Karanth et al., 2013). In a study carried out at Bhadra Tiger Reserve of South India by Madhusudan (2003), it was estimated that annually each household around the reserve lost 12 % of their livestock to large felines and approximately 11 % of their annual grain production to elephants. Compensation awarded to them covered less than 5 % of livestock loss and 14 % of crop loss. People were unhappy with the procedural delays for processing of claims. Similarly, in a survey conducted near Nanda Devi Biosphere Reserve of Indian Himalayas, it was found that 6 of 22 annual crops and all 4 horticultural crops on private farms were damaged by wildlife, but compensation by Reserve management for livestock killing by wildlife and compensation amounted to only 4–10 % of the total assessed monetary value of killed livestock (Maikhuri et al., 2000). A study carried out in the village Bhalalogpur (a pseudonym), located at the border of Rajaji National Park Uttarakhand, examined the experience of people with economic compensation for the losses due to human–wildlife conflict (Ogra, 2008). People living in the village had the problems of predation of livestock by leopards and tigers and crop loss by wild boars and elephants. They used techniques like wooden fencing, fire torches, home–made crackers, but these techniques did not bring effective results for them. There were very few instances when these techniques worked, but overall, villagers were at the receiving end. The compensation that was awarded after that fateful event varied widely in terms of the value, like from INR 500 to INR 100,000. The seasonal crop loss was estimated to be about 20–50 % of the total crop loss. As a result of this the amount of food grain available for domestic use declined. When it came to compensation, the villagers had to face two major problems. One was the small amount of compensation and the other was the complex procedure of applying for compensation. The amount of compensation was not enough to compensate for the losses encountered by villagers. Also, there were many damages that were not covered under the compensation scheme of government. Also, there were transaction costs associated with the filing and preparation of cases, which were unavoidable. So these costs too became a burden for the villagers. Applying for compensation involved travelling and that was logistically a complicated process. Hence, many villagers did not apply for compensation. And if someone applied, the delay caused by the whole process nullified the value of compensation. Sometimes, errors were also reported in compensation. Thus, it became quite difficult for the villagers to sustain themselves in this whole process. Suggestions have been made to deal with such issues. One, the compensation should reflect market value for the losses encountered. Two, apart from compensation in the form of 'cash', it can also be paid

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in the form of 'kind'. Participation of disadvantaged groups and local level can be effective in this whole process. On the other hand, these policies should not turn into people–centred policies. Also, it may occur that the provision of fair compensation may increase the chance of farmers becoming less cautious of the crop. So this is also a point of concern. A compensation process is necessary to provide compensation. If a proper process is not in place, then people may take unwanted advantage. But on the other hand, keeping in mind the state of poor and needy people residing in the forest area, it can be relaxed to some extent (Ogra, 2008). Sometimes the approach followed for conservation (rehabilitation of people) may put the livelihoods of local communities dwelling inside the reserve in danger. Relocation may not always be fruitful for the people who depend on forest resources. In a study performed in 2007 in Sariska Tiger Reserve (Alwar, Rajasthan), it was found that the people who were resettled as part of the conservation project were affected in many ways (Torri, 2011). Their life was changed to a completely different world. In Sariska tiger reserve, a ban was imposed on the collection of forest resources, based on a conservation point of view. Wood cutting permits were limited and removal of dry wood for the construction purposes was also forbidden. These activities affected villagers. They needed forest resources for survival. Because of fear of getting caught by the forest staff, they used to cut the branches quickly rather than spending time looking for dry wood in the forest. Thus, they started paying less attention towards sustainable extraction in forest reserve. The practices of local communities and their demographic growth were said to be the main reason for impoverishment of biodiversity of the reserve. The villagers were not satisfied with the arrangements and in some cases, the villagers who were resettled went back to their former settlements. On the other hand, the displacement was considered vital for forest dwellers as they had no access to basic necessities like education, medical care, transportation, etc. According to forest staff, grazing and lopping seriously affected the appearance of forest in Sariska, especially in the buffer area. Adverse impacts on ecological and social development, land fertility and loss of biodiversity were noticed. From the villagers' perspective, displacement of people eliminated the traditional practices that were beneficial for the maintenance of forest biodiversity. Shepherds felt that their livestock could not be sustained outside the forest area, and that their loss would have an adverse impact on their livelihood. According to the villagers, they were not provided with compensation charges for constructing the houses, and so the people had to face severe consequences for this. To achieve conservation as well as proper living standards for people, attitudes on both sides need to change. An interactive and collaborative approach in this regard can suit the process in a better way. Understanding the priorities of both sides is very important. A call for synthesis of the skills and insights of diverse communities may be helpful. Furthermore, a

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more holistic and rigorous exercise that considers the ecological and sociological insights on displacement can also be beneficial in this regard. A neutral course of action will involve steps taken to avoid involuntary displacement, keeping a track of whether the displacement–related grievances have been properly addressed or not, converting the involuntary displacements into voluntary ones, and designing such compensation packages that ensure the displaced people will not be negatively affected due to displacement. For this to happen, conservationists have to identify the interests of those who will be displaced, work with governments at local or national level or agencies to prepare suitable compensation packages, and involve local communities to determine the balance between compensation and concessions in relation to the strictness with which the conservation goals will be enforced. There is one more option, according to which conservationists may de–gazette some part of protected area so that its resources become available for development. Thus, better allocation of resources will be available, or better compensation packages can be formed out of this. Before finalizing a course of action, a balance must be sought between ethical appropriateness, monetary costs, and political feasibility (Agarwal and Redford, 2009). The dynamics of geography in and around reserves The people living in close proximity to protected areas use forest resources like fuel wood, herbs, fruits, etc. But, the people living in the core and on the periphery affect it in different manner. The way they utilise the natural resources is different. In Tadoba Andhari Tiger Reserve (TATR) of Maharashtra state, it was determined how the resources of the park are impacted by people living inside or near the park area, and the different impacts caused by the park on different communities dependent on it. It was found that as the distance from villages present in the interior increases, both sapling species richness and sapling density increase. The main incursion of villages present at the periphery is on the richness of species and trees. They affect the vegetation mostly by felling of specific trees for timber, with minimum impact on sapling regeneration. Thus, the way of affecting the forest is different in the two cases. It was found that the villages in the interior of the forest affected the forest less than villages present at the periphery. Differences were also present in terms of land cover change. Here, the peripheral part of the park was worst affected. It was the most depleted area of forest cover with the lowest percentage of stable forest and the greatest percentage of non–stable forest. The villages present in the innermost part of the forest had the highest percentage of stable forest. Also, in the case of park fragmentation, the peripheral area suffered most. This region accounted for smaller sized patches and low shape index, located at far distances from each other. The areas surrounding the innermost villages were least fragmented (Nagendra et al., 2010).

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Tourism in protected areas Tourism in protected area comes with many implications and challenges. On one hand the local people get employment opportunities, and ways of income become diversified.Ultimately, this results in the betterment of people living in and around reserves. However, some shortcomings may also creep in. The traffic rises due to the increased frequency of vehicles. This affects the wildlife too. So care is required in this regard. Land is a scarce resource in developing countries, especially in a thickly populated country like India, where demand for land for various purposes remains always high. Therefore, protected areas are under increasing pressure to provide economic justification for their existence. Ecotourism from such areas provides a platform to generate substantial benefits for both governments and the local communities. The extent to which nature–based tourism or ecotourism offsets the costs of a PA has been examined in very few cases (Walpole et al., 2000). Roads play an important role inside protected areas. There have been many instances when animals are killed by vehicles. Such damage to wildlife is difficult to control. In some cases, there are religious sites that attract a large number of visitors. So traffic increases and due to frequent passing of vehicles through the forest, the incidents of animals and insects getting killed also increases This kind of problem cannot be solved completely. However, it can be mitigated to the maximum extent possible. In a study conducted across three habitats in Kalakad Mundanthurai Tiger Reserve, Tamil Nadu, from 2008–2009, the negative influence of the presence of roads on terrestrial and aquatic ecosystems was determined (Seshadri and Ganesh, 2011). It was found that millipedes, anurans, insects and reptiles dominated the list of mortalities, whereas the mammals avoidedcollisions. By knowing this impact, various strategies can be determined to tackle this issue of roads inside protected areas. The vehicular movement at night increased the mortality rate of nocturnal animals. During the festive season, large numbers of pilgrims in visit the religious enclaves located inside the protected areas, creating huge, sudden surges in traffic. Small animals are more likely to be killed by such disturbances than large animals, as these latter normally keep away from such traffic. Among the species recorded, the millipede was the most commonly killed species both before and after the festival season. Though they are active during dusk, in wet conditions and during the rainy season they were killed in the daytime too. Smaller invertebrates like ants, forest roaches and glow worms had a higher rate of mortality than larger invertebrates such as centipede sp. and scorpion spp. Mortalities increased by a whopping 299 % during festivals. Not only diurnal but also nocturnal species were affected by this. Nocturnal species accounted for nearly 50 % of the total mortalities recorded. Apart from crawling arthropods, many flying insects were also killed. The flying insects are attracted to the light beams of vehicles at night and get crushed by them. This, leads to a cascading effect as birds like owls and nightjars come to the road to feed

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on the crushed insects. They, in turn, also get hit and killed by the vehicles. If this un–controlled movement of vehicles is not taken seriously, the situation will become worse and may lead to the local extinction of some species. The ecosystem is becoming affected. This could eventually cause a decline in the population of many species, and nocturnal species will be among the worst hit of all. Serious steps should be taken to deal with this problem. Regulations in vehicular movement can solve this problem to some extent. Though complete banning of traffic inside the protected areas is not possible, some restriction on vehicular movement –especially at night time– could be an effective option. In addition to this, public transport such as buses should be encouraged instead of private cars. This would reduce the density of vehicles on the road to some extent. The speed of vehicles also plays some role in all this. Building speed breakers would have a barrier effect on large mammals, allowing them more response time and avoid collision (Seshadri & Ganesh, 2011). It is estimated that for every rupee spent by tourists, the central and state governments in India receive 15 paise (1 rupee = 100 paise) as taxes (Seth, 1997). Furthermore, the tourism employment multiplier for India is about 1.8 and the tourism output multiplier is about 2.1 (Srivastava and Shukla, 2006), further demonstrating the high economic value of recreation services from these tiger reserves. The tourism recreational value for such spots is not truly reflected by gate fee revenue and researchers have estimated these values for some of the Tiger reserves using environmental economics methods like contingent valuation and travel cost approaches: e.g., 30 million Indian rupees (INR) per year for Corbett Tiger Reserve (Badola et al., 2010), 383.70 million INR per year for Kanha Tiger Reserve (Verma and Mishra, 2010) and 21.50 million INR per year for Kaziranga Tiger Reserve (Bharali and Mazumdar, 2012). Due to tourism, land use patterns see a shift. Land prices rise in areas close to protected areas. New tourist facilities come up. Hence, proper planning and management is required so as to maintain the ecological integrity and functionality of the protected areas. The tourism is enhanced by many factors, such as good wildlife sightings, publicity by media, enhancement of the quality and quantity of resorts, improved accessibility to urban centres, bird–watching, economic growth and betterment of middle–class conditions. In the protected areas, many problems may occur, like poaching, fishing, electrocution, fuel wood collection, etc. But the readiness and promptness showed by forest department against these problems varies. The tourism revenue is in the hands of the forest department and managed by them alone. It could be shared between the forest department and the religious institutions, so as to make the pilgrimages a better host of eco–tourism activities. However, in the Periyar Tiger Reserve only 56 % of the revenue was given to Periyar foundation to support eco–development and people’s participation. Tourism can play a very important role for protected areas by earning revenues which can be used to support proper park management plans

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and conservation efforts. If it is maintained in a proper manner, then more tourists will be attracted. In some protected areas geographical clustering of facilities is possible. This leads to better management of land use pattern than in other protected areas where facilities are spread out. Employment opportunities can be enhanced for local communities in the field of tourism. In order to sustain tourism, local communities and residents should be mixed and support should be enhanced among private enterprises for conservation initiatives (Karanth and DeFries, 2011). Eco–tourism can be beneficial for the people living inside the protected areas. These people have limited resources for consumption. They are not financially sound. Thus, tourism brings in lot of earning opportunities. However, if the tourism is stopped for some reason, locals are badly affected. Their earning opportunities suffer. Tourism generates a substantial amount of revenue not only for local people but for the country too. Foreigners and nature lovers are attracted tonational parks and wildlife sanctuaries but too much human pressure in and around protected areas may prove harmful to animal populations. For example, in a research study conducted by the scientists of Wildlife Institute of India, tigers (both male and female) in the of Sariska Tiger Reserve in Rajasthan state were found to be in an extremely stressed condition due to the excessive production of hormones (glucocorticoids). The level of these hormones in the tigers in this reserve was twice that of tigers in other reserves. The reason was found to be excessive human interference in the reserve. There are 29 villages within the reserve and more than 400 villages around the reserve. Moreover, a large number of devotees visit a religious temple located in the core area. All these factors prevent the mating environment among tigers in the reserve. This conclusion was reached by a study in five tiger reserves of the country, namely Sariska, Panna, Bandhavgarh, Kanha and Ranthambore (Yadav, 2017). In a study conducted in the Nanda Devi biosphere Reserve of Himalayan region, which sheds light on the history of expeditions and impact on local economy, the potential sites and expedition routes were identified and action plans for sustainable ecotourism were designed. After it was declared a National Park, all expeditions and trekking activities were banned, especially in the core area of the park. This had consequences on the local economy and environment. Thus, conflicts emerged between local people and authorities. This area was first approached in 1934 for trekking. After the independence of India, a huge increment in the number of mountaineers was observed. This ultimately led to adverse impacts. Both flora and fauna were seriously damaged. Entry only for the purpose of research was allowed. Stopping the flow of tourists had a negative impact on the earnings of local people. Most visitors came for trekking purposes and the people engaged as guides and porters were affected. The conservation policies affected many day–to–day activities of local inhabitants. Restrictions were placed on grazing, collection of non–timber forest products and removal of dead logs from van panchayats. Due to the presence of top–down structure, people were not involved

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much in the action. Though some portion of a buffer area was opened for trekking, it did not result well as it was considered less adventurous than the core area for trekking. For promoting eco–tourism, many factors should be taken into account. Such actions should be identified as feasible from an ecological, socio–economic and cultural point of view. The primary focus should be the role of local people, the scope of expansion and the reduction of conflict. Local people must be made aware of their legal rights and their cooperation and joint efforts should be encouraged. Visitors should be made aware of the culture, tradition, climate and ecology of the place. Many opportunities for local people can be promoted, such as homestay tourism, growing of vegetables, and poultry and milk production. Small lodges and hotels too can also be established (Maikhuri et al., 2000) The 'grazing' dilemma The people living in and around the protected areas depend largely on livestock. Grazing by cattle also affects the forage present in the area. Many programmes have been implemented for conservation and better management of biodiversity in the Himalayas. The Natural Resource Management Plan is one such plan that exhibits the biodiversity conservation and management by creating protected areas in the form of sanctuaries, national parks and biosphere reserves. As a result, grazing was banned in some regions of the Nanda Devi Biosphere Reserve and Valley of Flowers National Park in the Himalayas for many years. This led to surprising results. The population of cattle started declining rapidly. Sheep and goat populations showed a drastic decline due to bans on grazing. The population of horses and mules was also affected. However, animal population loss was highest for yak. The domesticated yak population decreased almost to zero. Negative effects were also observed in terms of vegetation dynamics. Weeds and bushes/thorny bushes started growing as the ban was enforced. These weeds and bushes started expanding across many alpine pastures. Before the implementation of the National Resource Management Plan, cattle consumed these plants and thus controlled their growth. But after the implementation of the reserve, the rate of growth increased very rapidly, changing the vegetation dynamics and posing a threat to biodiversity. These species affected the richness of medicinal and aromatic plants. They replaced the habitat of valuable alpine pastures. This interruption of traditional land use led to landscape homogenization, and chances of fire hazard increased to a high level. The excessive growth of these unwanted species created an imbalance in the ecosystem (Nautiyal and Kaechele, 2007). The Bharatpur Sanctuary was declared a National Park in 1981 and in 1982 a ban was enforced by the Government of Rajasthan. As a result of this ban, weeds took over in wetlands, reducing the fish population and also bird populations and nesting (Lewis, 2003). The excessive presence of weeds clogged

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canals, filled marshes with weeds, and led to a series of wildfire incidents in the grasslands. Weedy plants like water hyacinth grew rapidly and had negative effects. It was then concluded to allow the primary consumers back for grazing so as to avoid decline in numbers and species of birds. And for the unmanageable fires in the open grasslands, it was decided to allow the villagers to collect fodder.Villagers had to pay a fee for a cutting license and were then allowed to cut grass from any dry section of the park. In this way, the situation returned to normal. Human–wildlife conflict For conservation practitioners, one of the most challenging issues is to address human–wildlife conflicts. Many ecological and social factors can be responsible for these conflicts. There is a need to develop preventive strategies so as to avoid these conflicts. In five important reserves in Karnataka's Western Ghats, a study was conducted to examine the patterns of loss due to conflicts and compensation awarded (Karanth et al., 2013). It was found that crop raiding incidents were experienced by villagers year round. From October to December, the frequency of such incidents was found to be slightly higher. There was high crop loss in the region. Crop loss was attributed to 19 species of wild animals, mainly wild pig, elephant and chital. Lower crop loss was associated with the distance from the reserve. Fifteen percent of households reported livestock loss. Predation incidents were in the range of 0 to 3 on an annual basis. People had their own mitigation measures, but these were ineffective. Individual measures such as night watch, guarding animals, and similiar, were adopted. Like for the crop raiding pattern, the greater the distance from the reserve, the lower the livestock loss. Livestock loss was positively associated with animals grazing inside the reserve. Overall, the compensation process was a long process. Compensation payments for loss took more than one year to reach the affected people. People were inclined to report losses related to large animals like tigers and elephants. The north east region of India is significant in terms of large wild animals like tigers, rhinoceros and elephants. Of these animals, elephants have become the focal point of conflict and conservation issues. More than 1,150 humans and 370 elephants were killed between 1980 and 2003 due to these conflicts (Choudhury, 2004). The burgeoning human population, the increasing needs for housing, agriculture, etc. were the main reasons for such conflicts. Due to the squeezing habitat, elephants usually come down to paddy fields and destroy crops. In 2001, angered residents in northeast India (Assam) selectively targeted their paddy fields with poison for crop–raiding elephants; a mutilated elephant carcass was subsequently discovered in the field with the words, 'Paddy Thief Bin Laden' scrawled upon its body (Sethi, 2003; Ogra, 2008). In a household survey around the Kanha National Park of Central India, 17 species were identified as crop raiders, including 10 herbivores, 4 carnivores,

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2 primates and peacocks (Karanth et al., 2012). The highest numbers of raiding incidents were reported from September to December, peaking in the month of October. Sixty–four percent of households reported experiencing more than five incidents per year, while 32 % of households reported 2–5 incidents per year. Similarly, livestock losses to 10 carnivores were reported from April to July (dry season). Besides tiger and leopard, jackal and wolf were also found responsible. Some (34%) households reported more than 5 incidents/ year, another 34 % reported 2–5 incidents /year, and 31 % reported one incident/year. Gir National Park in the Western state of India is famous for Asiatic lions. A study determined that on average, 14.8 attacks by lions and 2.2 lion–related deaths occurred annually from 1978 to 1991 (Saberwal et al., 1994). Most of these attacks (82 %) occurred outside the protected area and during the drought season. The intensity of lion attacks also increased considerably. The researchers advocated reducing the lion population in the park by relocating or culling. While some costs are visible, others are hidden and remain poorly addressed. Hidden impacts can be many, such as food insecurity, disruption of livelihoods, diminished psychosocial wellbeing, etc. The consequences people face due to conflicts with wildlife include loss of life, livestock predation, fear of wild animals, and crop damage by wildlife.The losses incurred by poor people who are dependent on forest resources affect their lives in an adverse manner. Crop damage caused by large animals, like elephants, is widespread. Depredation of livestock is yet another impact due to human–wildlife conflict. Such damage often leads to retaliatory killing of wildlife. Overall, the ecological consequences of such conflicts lead to drastic changes in wildlife populations with changes in genetic diversity. The causes of such conflict can be mobility, displacement and increment in human populations, loss and fragmentation of the existing habitat where people have been living, etc. Some costs cannot be compensated, such as decreased psychological well–being caused to fatality, or disruption of family and food insecurity caused by crop or livestock loss. Other ill effects include opportunity loss, poor health and nutritional status, and transactional costs incurred when pursuing compensation. The visible and hidden impacts are intermixed. The degree and severity of psychosocial effects cannot be overlooked. The aftermath effects include poverty, poor access to resources and social capital, and ethnic and political marginalization. The death of a provider, generally a male from the family, leads to catastrophic results for the family as a whole. The burden of responsibility falls on the shoulders of females and children. The relatives of victims of tiger and other carnivore attacks suffer from physical and mental trauma. In most cases they are unable to recover the body of the victim, which again results in mental trauma and stigma. Post–traumatic stress disorder is found in both male and female members (Barua et al., 2013). Crop raiding by wild animals, especially elephants, can lead to the displacement of the family from that

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area. Loss of livestock due to predation by carnivores can destroy the family’s income and way of life. Livestock forms a substantial proportion of the socioeconomic capital of communities in many areas. Ultimately, people have to bear the consequences. They also have to guard their cattle at night. Thus, after a long tiring day they have to work at night too. This has a negative impact on their health (Barua et al., 2013). Many recommendations can be made. Maps can be used by park and revenue authorities as well as the non–governmental organisations to target preventive actions in the most vulnerable conflict zones. Mitigation measures should be investigated, and this investigation should be scientifically based. Compensation must be in accordance with scales and should be observed and locally monitored (Karanth et al., 2013). A suitable plan of action is required in this regard. A systematic assessment of the extent and scale of such hidden impacts is needed. Different scenarios of human–wildlife conflict should be considered for systematically assessing the extent and scale of hidden impacts. It should be determined how such conflict impact on the nutrition, physical and psychological well–being of the people. The risk of replicating other conservation conflicts must be avoided. A strong link between conservation, health and social sciences is required (Barua et. al., 2013). Road and rail kills In India, a highway bisecting the protected areas is not a rare sight. In recent years it has been realized that highways have a severe impact on wildlife and their habitats (Vijayakumar et al., 2001; Das et al., 2007; Baskaran and Boominathan, 2010). Forest departments and non–governmental organizations in India are thus protesting against the construction of these so–called temples of development. The 'hunting' bane Hunting and poaching by local people has always been a serious threat to protected areas in India. In a survey conducted in the Kudramukha protected area in south India, it was found that at least 26 species of mammals were hunted, mostly with guns, at an estimated intensity of 216 hunter days/month/village (Madhusudan and Karanth, 2002). One study found that the population of tigers in Panna Tiger Reserve in Central India decreased from 2006–2010 due to poaching (Gopal et al., 2010). Tiger deaths related to poaching reached an all–time high in 2016. These figures represent only a fraction of the true mortality figures. Electrocution and poisoning of big cats have also been recorded across tiger habitats ('Save them from the trap' (July 29, 2017 ). The Pioneer Daily News paper, Chandigarh Edition). Snares are a problem not only in India but throughout Asia. Hundreds of thousands of deadly snares are removed by rangers from India’s protected areas by forest officials annually but this is just the tip of the iceberg. Habitat degradation

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and habitat loss also have a huge impact on tiger population. Not only human population around reserves but development projects, industries and roads in and around these protected areas are potential threats. Enforcement of Wildlife Protection Act, 1972 (modified time to time), which prohibits hunting of most of the wild animals is a real challenge in India, especially in the north–east region of India. Local people in this portion of the country have a strong tradition of hunting (Hilaluddin et al., 2005; Mishra et al., 2006 and Datta et al., 2008). Eating wild birds/animals constitutes a significant part of their normal diet. Local hunters are often lured by international wildlife smugglers for derivatives from species such as tigers and elephants (Datta et al., 2008). Hunting, illegal fishing and trapping of wild fauna like tigers, barking deer, leaf deer, sambhar, wild boar, bears, wildcat and a variety of birds by local inhabitants (Lisu, Chakma and Mishmi) for bush meat and hideis a severe concern for the management of Namdapha National Park of Arunachal Pradesh of this region (Arunachalam et al., 2004). During a camera trapping survey in the Namdapha National Park, no tiger was sighted even though this park is part of Project Tiger, a centrally sponsored scheme of the Government of India since more than a decade ago. The clouded leopard was the only large carnivore detected by camera trapping. Illegal hunting seemed to be the main cause behind the disappearance of tiger from the park (Datta et al., 2008). The forests of North–East India are recognized as a global biodiversity hotspot and as an endemic bird area due to their richness in floral and faunal species. The landscape has high species diversity and endemicity as it forms the transition zone between the Indian and Malayan eco–regions. North eastern states of India account for more than one fourth of overall forest and tree cover of the country. But today, the situation on this front is of concern. Unfortunately, due to increasing anthropogenic demands and technological development, the states in this region are no longer immune to large–scale land–use change (Bhuyan et al., 2003). Discussions with local people revealed that the availability of timber species, cane and bamboo in outer region of protected areas was reduced due to encroachments, forest fires, over exploitation, habitat destruction, lack of plantations and timely regeneration activities, and invasion of exotic and weed plant species like Lantana, Mikenia, Eupatorium, Parthenium in the forest areas. Therefore, awareness must be raised among the indigenous communities, stressing the need to conserve rich biodiversity, especially plants of ethno–botanical importance and local wildlife. It was also noticed during talks with locals that villagers now have to travel larger distances in forests to hunt animals for religious ceremonial purposes than what they did a decade ago (Aiyadurai et al., 2010). The government agencies must support the conservation measures of biodiversity by the indigenous groups and must undertake vigorous awareness campaigns to protect local biodiversity and wildlife. State Biodiversity Boards, State Medicinal Plants Board, State Forest Research Institutes and Department of Environment and Forests have a major role to play in this direction.

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Vitality of protected areas in terms of ecosystem services Protected areas offer a range of ecosystem services that provide economic, social, cultural and spiritual benefits. They also help buffer climate change and contribute by storing and sequestering carbon. In India, Project Tiger was initiated in 1973, and nine national parks and sanctuaries were declared as tiger reserves. Today, there are 47 tiger reserves, covering over 2 % of the country's geographical area. These tiger reserves are of tremendous value. They support human life by protecting fish nurseries and agricultural genetic material. Not only this, but they also provide cheap and clean drinking water, which can also be used for irrigation purposes. They provide immense stock and flow whose benefits are intangible, and thus often unaccounted for in market transactions. The National Tiger Conservation Authority (NTCA), Government of India, assigned a study to the Indian Institute of Forest Management (IIFM) Bhopal to estimate the quantum of significant ecosystem services, in terms of money, so that the real worth of these ecological assets may be known to the general public, policy makers, academicians and politicians. Many ecologists feel that the establishment of such reserves could be justified in terms of emanating ecosystem services alone (Badola et al., 2010). For instance, Periyar Tiger Reserve protects the watershed of Periyar Lake that irrigates more than 900 km2 of agricultural land in neighbouring states (Shukla, 2011). A team of researchers under the leadership of Professor Madhu Verma of IIFM, Bhopal, completed the study in 2015. The study included six tiger reserves, located in different forest landscapes: (1) Corbett Tiger Reserve (Uttarakhand); (2) Kanha Tiger Reserve (Madhya Pradesh); (3) Kaziranga Tiger Reserve (Assam); (4) Periyar Tiger Reserve (Kerala); (5) Ranthambore Tiger Reserve (Rajasthan); (6) and Sunderbans Tiger Reserve (West Bengal). The findings are an eye opener for all those concerned about life supporting systems and continuance of life on earth. We summarise these values in two of these reserves, the Corbett Tiger Reserve in North India, and the Periyare Tiger Reserve in South India. The Corbett Tiger Reserve is located in three districts in the State of Uttarakhand: Pauri Garhwal, Nainital and Almora, and it extends over an area of 1,288 km2. Of this, 822 km2 is core zone and 466 km2 is buffer zone. Its total value of stock benefits was found to be INR 261.8 billion Indian rupees (INR) and for flow benefits, 14.7 billion INR per year (Verma et al., 2015; table 2). Periyar Tiger Reserve is in Western Ghats, in the state of Kerala. It is located in the Idukki district of Kerala state. It covers an area of 925 km2, of which 881 km2 is core zone and 44 km2 is buffer zone. It also includes a 26 km2 water spread area of Periyar Lake. Its total value of stock benefits is 316.5 billion INR and for flow benefits, and 17.6 billion INR per year (Verma et al., 2015; table 3). Overall, the flow benefits from these selected six tiger reserves range from 50,000 INR/ha/year (US $ 769) to 190,000 INR/ha/year (US $ 2,923). The lower value corresponds to the tropical dry deciduous

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Table 2. Few important ecosystem services emanating from Corbett Tiger Reserve Tabla 2. Pocos servicios ecosistémicos importantes originados en la Reserva del tigre de Corbett. Important ecosystem services

Economic Value (in INR)

Gene–pool protection

10.65 billion year–1

Water provision to downstream districts of Uttar Pradesh

1.61 billion year–1

Water purification services to New Delhi

550 million year–1

Generation of employment for local communities

82 million year–1

Provisioning of habitat and refugia for wildlife

274 million year–1

Sequestration of carbon

214 million year–1

Table 3. Few important ecosystem services emanating from Periyar Tiger Reserve Tabla 3. Pocos servicios ecosistémicos importantes originados en la Reserva del tigre de Periyar. Important ecosystem services

Economic Value (in INR)

Gene–pool protection

7.86 billion year–1

Water provision to districts of Tamil Nadu

4.05 billion year–1

Provisioning of habitat and refugia for wildlife

3.55 billion year–1

Generation of employment for local communities

25 million year–1

Water purification services to neighbouring towns and districts

483 million year–1

Recreation value

425 million year–1

forest region, where Ranthambore Tiger Reserve (Rajasthan) is located and the higher value corresponds to the tropical moist evergreen forest region where Periyar Tiger Reserve (Kerala) is located (Verma et al., 2017). Nearly 5 % of India's geographical area consists of protected areas and they are responsible for providing ecosystem services or flow benefits worth 2,000 trillion INR per year by taking an average figure of INR 120,000/ha/year of flow benefits from the above study. This shows the huge significance of these areas in terms of ecological and social security of humans and other living systems in the country. Conclusion In the light of growing awareness of life–supporting functions of ecosystem services and advanced technology to make use of genetic diversity, the economic value of protected areas is likely to beappreciated in the near future (Verma et al., 2015). The protected areas of India support a wide range of economic

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sectors, and investment in this natural capital will lead to maintaining ecological security and food security, thereby leading to overall sustainable development. These investments can be cost effective responses to the climate change crisis, creating jobs, supporting local economies, and maintaining ecosystem benefits on a long–term basis. An amicable and tactful handling of all contentious issues of protected areas can be a win–win situation for park managers, local communities and other stakeholders. Reaching a higher GDP through infrastructure projects is important for the country but there is a need to take up such projects together with mitigation measures such as overpasses for the passage of tigers, elephants and other animals in and around reserves. Politicians, policy makers, planners, bureaucrats and common people need to understand that the future security of the national heritage of the country is at stake. A balanced view on the country’s development, the conservation of biodiversity, and the hardships faced by people living in and around protected areas is the need of the hour.

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Age–dependent capture–recapture models and unequal time intervals A. Sanz–Aguilar, R. Pradel, G. Tavecchia

Sanz–Aguilar, A., Pradel, R., Tavecchia, G., 2019. Age–dependent capture–recapture models and unequal time intervals. Animal Biodiversity and Conservation, 42.1: 91–98, Doi: https://doi.org/10.32800/abc.2019.42.0091 Abstract Age–dependent capture–recapture models and unequal time intervals. Estimates of survival probabilities in natural populations can be obtained through capture–mark–recapture (CMR) models. However, when capture sessions are unevenly spaced, age–dependent models can lead to erroneous estimates of survival when individuals change age class during the time interval between two capture occasions. We propose a solution to correct for the mismatch between time intervals and age class duration in two age class models. The solution can be implemented in different ways. The first consists of adding dummy occasions to the encounter histories and fixing the corresponding recapture probabilities at zero. The second makes use of the log–link function available in some CMR software (e.g. program MARK). We used simulated and real data to show that the proposed solution delivers unbiased estimates of age–dependent survival probabilities. Key words: Age–dependent models, Capture–mark–recapture, Missing data, Survival probability Resumen Modelos de captura y recaptura dependientes de la edad e intervalos de tiempo desiguales. A través de modelos de captura, marcaje y recaptura (CMR) se puede estimar la probabilidad de supervivencia en poblaciones naturales. Sin embargo, cuando las sesiones de captura están espaciadas de manera desigual, los modelos dependientes de la edad pueden producir estimaciones erróneas de la supervivencia si los individuos cambian de clase de edad durante el intervalo entre dos sesiones de captura. Proponemos una solución para corregir el desajuste entre los intervalos de tiempo y la duración de las clases de edad en modelos con dos clases de edad. La solución puede aplicarse de diferentes formas. Una consiste en añadir muestreos ficticios en las historias de captura y fijar las probabilidades de recaptura correspondientes a cero. Una segunda aplicación usa la función log–link disponible en algunos programas informáticos de CMR (p. ej., el programa MARK). Usamos datos simulados y reales para mostrar que la solución propuesta produce estimaciones no sesgadas de las probabilidades de supervivencia dependientes de la edad. Palabras clave: Modelos dependientes de la edad, Captura, marcaje y recaptura, Datos incompletos, Probabilidad de supervivencia Received: 08 V 18; Conditional acceptance: 17 VII 18: Final acceptance: 25 VII 18 Ana Sanz–Aguilar, Giacomo Tavecchia, Animal Demography and Ecology Group, Institut Mediterrani d’Estudis Avançats IMEDEA (CSIC–UIB), c/Miquel Marques 21, 07190 Esporles, Baleares, Spain.– Roger Pradel, Biostatistics and Population Biology Group, Centre d’Ecologie Fonctionelle et Evolutive, CEFE–CNRS, 1919 route de Mende, F–34293 Montpellier, France. Corresponding author: Ana Sanz–Aguilar. E–mail: asanz@imedea.uib–csic.es

ISSN: 1578–665 X eISSN: 2014–928 X

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Introduction Capture–mark–recapture methods (CMR) are widely used for diagnosis of natural populations because they can be applied to obtain robust estimates of demographic parameters accounting for imperfect detection of individuals (Lebreton et al., 1992; Williams et al., 2002; Sanz–Aguilar et al., 2016). Cormack–Jolly–Seber models for the estimates of survival probability in natural populations are based on the important assumption that animals share the same parameters regardless of their past or present history (Pradel et al., 2005). When animals are marked as young this assumption does not hold because newly marked individuals typically have a lower survival probability than already marked individuals (adults). This difference can be accommodated by including age–dependent parameters into the CMR model (Pollock, 1981; Lebreton et al., 1992). In a simple two–age–class model, one parameter, noted ϕ', would apply to the survival probability of young individuals and a second, noted ϕ', would apply to the survival of adults (see examples in Hiraldo et al., 1996; Prugnolle et al., 2003). Age–dependent parameterizations have also been considered when only adults are marked to correct for an excess of animals seen only at marking, i.e. transients (Pradel et al., 1997), for example, when tags are potentially harmful (Saraux et al., 2011) or to model an effect of breeding experience (Sanz–Aguilar et al., 2008, 2012). Age–dependent survival probabilities are parameters of interest in many ecological studies (e.g. Clobert et al., 1988; Loison et al., 1999; Tavecchia et al., 2001; Bonenfant et al., 2002; Perret et al., 2003; Catchpole et al., 2004; Sanz–Aguilar et al., 2015). However, while age–classes are equally spaced, intervals between capture–recapture occasions may not be equally spaced on the same scale, leading to erroneous estimates (see the problem in, for example, Covas et al., 2002; Zabala et al., 2011; Zuberogoitia et al., 2016). This is because individuals would change their age class within the interval between two sampling occasions rather than at the end as assumed by CMR models. We briefly introduce the problem and illustrate how it can be solved by taking advantage of the flexibility of CMR models. The problem Logistic, financial or weather–dependent constraints can interrupt monitoring or modify the temporal frequency of sampling occasions, leading to different time length between capture–mark–recapture sessions. Unequal time intervals, alone, do not present a major problem in CMR models (Bears et al., 2009; Cooch, 2009; Schmidt et al., 2007). Consider a study with k sampling occasions with intervals between the occasions j = 1, 2, …, k–1. The length of the intervals between the sampling occasions is lj. The l value is taken as the exponent of the survival parameters expressed in some common time unit for the interval j, j + 1, as ϕlj. For example, the survival parameter over a unit interval (l = 1) would be ϕ1, for a two–unit interval (l = 2) it would be ϕ2, and so on. Values in

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the vector l are commonly integers, e.g., years or months, but can also be decimal numbers; for example, the survival parameter over an eighteen–month interval can be written in terms of yearly survival as ϕ1.5 (l = 1.5). The freely available software for CMR analyses, such as MARK (White and Burnham, 1999), RMARK (Laake, 2013) or ESURGE (Choquet et al., 2009), allows users to specify the vector of lj values. However, unequal time intervals pose a problem in age–dependent models because, contrarily to intervals between occasions, the age classes always retain the same length. As a consequence, an individual may 'move' through age classes during an interval of length lj and the survival parameter can no longer be written as ϕlj because the instantaneous survival probability changes with the age classes spanned by the interval of length lj. The mismatch between the duration of an age class and the time interval would, for instance, lead to an overestimation of the first–year survival probability if the sampling interval were greater than one year. The fundamental problem is that a given age–dependent parameter applies to only a part of the time interval. This can be solved by specifying the length of the interval for each parameter considered. We outline the solution and provide a step–by–step illustration of how this can be implemented in freely available software for CMR analysis (e.g. MARK, RMark or E–SURGE, see details in the Supplemental Information S1, S2 and S3). Note that the two implementations below are simply two practical approaches to solve the problem (see supplementary material S1, S2 and S3). Methods Implementation 1: adding dummy encounter occasions When intervals are of unequal lengths, a possible solution is to add dummy encounter occasions in the encounter histories to 'fill' the temporal gaps between occasions, which means, in practice, adding columns of 0s (e.g. Grosbois and Tavecchia, 2003; Sanz–Aguilar et al., 2010): the recapture probabilities corresponding to these dummy occasions should be fixed at 0. For example, let us consider a 7–year study with five capture–mark–recapture occasions in years 1, 2, 5, 6 and 7. The interval between the second and third occasions lasts three years instead of one. The l–vector of interval lengths would be 1, 3, 1, 1. The encounter history of animals released at the beginning of the study and always seen would be '1 1 1 1 1'. When columns of '0' are added to fill the temporal gaps, the encounter history above becomes '1100111' and all six elements of the l–vector are now equal to 1. The encounter probabilities at dummy occasions (3 and 4) should be fixed at 0 (figs. S2.1 and S2.3 supplementary material S2). The survival parameter of the first age class, ϕ', now always refers to an initial one–year interval. This approach can be implemented in programs MARK and E–SURGE. Adding dummy occasions permits to manipulate the correct survival parameters, but the dummy occasions come with no

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Table 1. Not separately identifiable parameters in presence of missing occasions: ϕ' juvenile survival; ϕ adult survival; p, recapture; 't' time effect; 'cov' covariate effect; '.' constant parameter; Np, number of separately identifiable parameters. Note that no individuals were marked during missing occasions and consequently juvenile survival parameters do not exist in the model for cohorts without released juveniles (i.e. ϕ'3, ϕ'4 for dataset 1 and ϕ'3 for dataset 2). Similarly, adults were not marked on the first occasion and, consequently, ϕ1 do not exist in the model. Tabla 1 . Parámetros no identificables por separado en presencia de ocasiones sin muestreo: ϕ' supervivencia juvenil; ϕ supervivencia en adultos; p, recaptura; 't' efecto del tiempo; 'cov' efecto de una covariable; '.' parámetro constante; N, número de parámetros identificables por separado. Durante las ocasiones sin muestreo no se marcó ningún individuo y, en consecuencia, no existen parámetros de supervivencia juvenil para los grupos sin juveniles liberados en el modelo (ϕ'3, ϕ'4 para el conjunto de datos 1 y ϕ'3, para el conjunto de datos 2). De igual forma, en el primer muestreo no se marcó ningún adulto y, por lo tanto, ϕ1 no existe en el modelo.

Model

Dataset 1

Dataset 2

Two missing occasions [3, 4]

One missing occasion [3]

Not identifiable

ϕ't ϕt pt 10 ϕ'2; ϕ'6; ϕ2; ϕ3; ϕ4; ϕ6; p6 13 ϕ'2; ϕ'6; ϕ2; ϕ3; ϕ6; p6 ϕ't ϕt p. 8 ϕ'2; ϕ2; ϕ3; ϕ4 10 ϕ'2 ; ϕ2 ; ϕ3 ϕ't ϕ. pt 9 11 ϕ't ϕ. p. 6 7 ϕ'. ϕt pt 9 ϕ3; ϕ4 11 ϕ'. ϕt p. 6 ϕ3; ϕ4 7 ϕ'cov ϕcov pt 8 9 ϕ'cov ϕcov p. 5 5 ϕ'cov ϕt pt 10 ϕ3; ϕ4 12 ϕ'cov ϕt p. 7 ϕ3; ϕ4 8 ϕ't ϕcov pt 10 12 ϕ't ϕcov p. 7 8 ϕ'cov ϕ. pt 7 8 ϕ'cov ϕ. p. 4 4 ϕ'. ϕcov pt 7 8 ϕ'. ϕcov p. 4 4 ϕ'. ϕ. pt 6 7 ϕ'. ϕ. p. 3 3

additional information and identifiability issues in full time–dependent models (tables 1, 2). If the parameters that appears in the gap are unrelated to known parameters from other intervals they will not be separately identifiable, e.g. if all survival parameters are time– and age–dependent, the first age–class survival probabilities ϕ' at the beginning of a gap occasion and the second age–class survival probabilities ϕ that follow them are not separately identifiable (tables 1, 2). However, the approach works well as long as one of the two types of survival parameters is kept constant or modelled as a function of environmental covariates (see simulated data results below, tables 1, 2).

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Implementation 2: using a log–link function An alternative to the above implementation and especially useful when the lj are not commensurate (or when too many dummy occasions are required) relies on the use of a logarithm transformation (this implementation is not available in program E–SURGE). The survival probability over the initial interval of a young individual can be decomposed into its initial survival as a young for a duration r with a survival probability per time unit of ϕ' followed by the survival as an adult for a duration s with a survival probability per time unit of ϕ. The survival probability over the

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Table 2. Identifiable quantities (i.e. * products of parameters) in the presence of missing occasions. Only models with redundant parameters are presented; see table 1. (Notation as in table 1). Tabla 2. Cantidades identificables (* productos de parámetros) en presencia de ocasiones sin muestreo. Solo se presentan los modelos con parámetros redundantes; véase la tabla 1. (Notación como en la tabla 1). Model

Dataset 1 Two missing occasions [3, 4]

Identifiable products

Dataset 2 One missing occasion [3]

Identifiable products

ϕ't ϕt pt 10 (ϕ'2 * ϕ3 * ϕ4); (ϕ2 * ϕ3 * ϕ4); (ϕ'6 * p6); (ϕ6 * p6) 13 (ϕ'2 * ϕ3) ; (ϕ2 * ϕ3); (ϕ'6 * p6); (ϕ6 * p6) ϕ't ϕt p. 8 (ϕ'2 * ϕ3 * ϕ4) ; (ϕ2 * ϕ3 * ϕ4) 10 (ϕ'2 * ϕ3) ; (ϕ2 * ϕ3) ϕ't ϕt pt 9 (ϕ3 * ϕ4) ϕ'. ϕt p. 6 (ϕ3 * ϕ4) ϕ'cov ϕt pt 10 (ϕ3 * ϕ4) ϕ'cov ϕt p. 7 (ϕ3 * ϕ4)

whole interval of length r + s is then ϕ'r ϕs (for year based sampling r = 1). Applying a log–link function the product ϕ'r ϕs is replaced with a linear combination of survival related quantities, log(ϕ'r ϕs) = rlog(ϕ') + slog(ϕ), to be estimated. The known quantities, r and s, can be used as covariates of the survival probability pertaining to the interval (fig. S2.2 supplementary material S2). This approach does not require changing the encounter histories contrary to implementation 1. A similar solution was used by Tavecchia et al. (2001, 2002) to estimate monthly survival of game species when marking occurred at different moments during the hunting season. Simulated cases To demonstrate the problem generated by unequal time interval in combination with age–dependent models, we considered a simple scenario with a model assuming two age classes and constant survival and recapture parameters. Note that this simple scenario is only for illustrative purposes. We simulated 100 datasets with five sampling occasions during a 7–years period (k = 7). A hundred new juvenile animals were released at each occasion. The time span elapsed between occasions was as l = 1, 3, 1, 1. We assumed constant yearly survival of newly marked juvenile individuals (ϕ' = 0.4) and constant yearly survival of adult individuals (ϕ = 0.8). We first analysed these datasets using unequal time interval (the incorrect approach) to illustrate the biases. We subsequently analysed them adding dummy columns to 'fill' the years without monitoring (implementation 1) and using the log–link implementation with r (= 1) and s (= 2) values for newly marked individuals in the second cohort to constrain the corresponding survival parameters appropriately (implementation 2). For each analysis, maximum likelihood estimates of ϕ' and ϕ were obtained using RMark (Laake, 2013). The code to simulate the data and run the analysis is provided

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in supplementary material S1. Supplementary material on S2 and S3 illustrate how to implement the solution in MARK (White and Burnham, 1999) and E–SURGE, respectively (Choquet et al., 2009). Parameter identifiability in time–dependent models In the example above we have assumed constant survival and recapture probabilities to illustrate the problem and its solution. However, in many cases, parameters are time dependent. To show the applicability of the solution to more complex parameter structures and to study parameter identifiability, we simulated two different datasets with seven intervals (k = 7, supplementary material S4). Both datasets considered temporal variation on juvenile and adult survival parameters as a function of a hypothetical temporal covariate D (adult survival was modelled as 1/(1 + exp(– (1.386 + 0.55 * D))) and juvenile survival as one half of adult survival at each occasion) and a constant recapture probability of 0.7. Datasets differed in the length of the period with no CMR information: dataset 1 considered a gap of two years (data from sessions 3 and 4 are missing), while dataset 2 considered a gap of one year only (data from session 3 are missing) so that l = 1, 3, 1, 1 and l = 1, 2, 1, 1, 1, respectively. To avoid identifiability problems associated with sample size (which is not within the scope of this note) a thousand new juvenile animals were released on each occasion. To explore parameter identifiability, we implemented 18 models to each simulated dataset, considering different combinations of temporal, covariate and constant effects on juvenile and adult survival probabilities, and temporal vs. constant effect on recapture probability. Datasets were analysed using program E–SURGE (Choquet et al., 2009), which provides detailed results on parameter identifiability using the explicit method proposed by Catchpole and Morgan (1997) to detect parameter redundancy (supplementary material S4).

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Dummy occasions/data

Log–link constraint

100

0.8

0.7

0.8

0.7

0.6

0.5

0.4

0.3

0.2

0.8

0.7

0.6

0.5

0.4

0.3

0.6

0 0.2

20 0.5

0.4

0.3

0.2

Frequency

100

Frequency

Unequal time option

First–year survival probability Fig. 1. Maximum likelihood estimates of first–year survival probability from 100 hypothetical datasets (see text). Vertical lines indicated the estimated average values (solid) and the true simulated value (dashed). Fig. 1. Estimaciones por máxima verosimilitud de la probabilidad de supervivencia en el primer año de 100 conjuntos de datos hipotéticos (véase el texto para obtener información más detallada). Las líneas verticales continuas indican el promedio de los valores estimados y las discontinuas, el valor simulado verdadero.

Application to real case We considered a dataset of capture–mark–recapture data of adult Mediterranean storm petrels (Hydrobates pelagicus melitensis) from Palomas Island (Eastern Spain). Birds were captured using mist–nets from 1996–2000 and from 2004–2006. The vector lj was 1,1,1,1,4,1,1 where the '4' stands for the 3–year gap between 2000 and 2004. A full analysis of this dataset can be found in Sanz–Aguilar et al. (2010). Here we report results obtained by using the unequal time interval option and the proposed solution for comparative purpose. We only present implementation 1 as results of both implementations are equivalent (see results). The goodness of fit test of a model assuming all parameters time dependent (Pradel et al., 2005) indicated a surplus of animals seen only at marking, i.e., transients (ϕ26 = 29.96, p < 0.05). As a consequence, survival during the first year after marking, ϕ', was considered separately from the subsequent survival, noted ϕ, in 2–age class models (Sanz–Aguilar et al., 2010). Results Survival estimates in simple two–age classes constant models Models in which we specified the unequal time interval in the l–vector delivered average estimates of first–year survival probability larger than the true value (average ϕ' = 0.52; 95 % CI = 0.46–0.60) instead of 0.4; fig. 1; supplementary material S1). This approach also led to a slightly underestimated recapture probability (average p = 0.63 (95 % CI = 0.53–0.72)

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instead of 0.70; supplementary material S1), while adult survival was close to the real value of 0.8 (average ϕ = 0.79; 95 % CI = 0.71–0.85). In contrast, the solution outlined above delivered the true values for all parameters regardless of the way the model was implemented (ϕ' = 0.40, ϕ = 0.80 and p = 0.70; fig. 1; supplementary material S1). Parameter identifiability in time–dependent models Our results indicate that models with two consecutive gaps present more problems of parameter identifiability than models with a single missing occasion (table 1). When juvenile and adult survival parameters are fully time dependent, survival parameters during the occasions without monitoring are not separately identifiable (tables 1, 2). Moreover, when the recapture probability is also time dependent, the last survival and recapture are not separately identifiable (tables 1, 2). All parameters became identifiable when juvenile and/ or adult survival is constant or modelled as a function of temporal covariates with the exception of models in which more than one consecutive occasion without monitoring and adult survival was time dependent (tables 1, 2). In this case, only the adult survival parameter corresponding to the year in which the gap begins was separately identifiable (tables 1, 2). Application to real case As in the simulated example, when using vector l as exponent of survival parameters, we obtained higher estimates for first–year survival probabilities. When gap years were not properly considered, models with the unequal time interval option delivered survival estimates of ϕ' = 0.73 and ϕ = 0.80 (transient pro-

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Table 3. AICc values of CMR models used to estimate survival of Storm petrels captured at Palomas Island using unequal time interval only (the wrong approach) and the solution using dummy occasions (see text for details). In all models, the recapture probability was kept time dependent (see Sanz–Aguilar et al., 2010): Np, number of parameters; Ui, unequal intervals; Do, dummy occasions. (The lowest AICc value in each approach is in bold.) Tabla 3. Valores de AICc de los modelos de CRM utilizados para estimar la supervivencia de los paíños europeos capturados en la isla de las Palomas usando la opción de intervalos espaciados desigualmente (el método incorrecto) y la solución añadiendo ocasiones sin datos (véase el texto para obtener información más detallada). La probabilidad de recaptura se mantuvo variable en el tiempo en todos los modelos (véase Sanz–Aguilar et al., 2010): Np, número de parámetros; Ui, intérvalos desiguales; Do, ocasiones sin datos. (El valor de AICc más bajo para cada método se muestra en negrita.) Model Np

ϕ't ϕt 19 781.73 781.73 ϕ't ϕ 15 777.31 777.31 ϕ' ϕt 14 786.60 779.86 ϕ' ϕ 9 780.61 772.91

portion p = 0.087). However, when gaps were filled with dummy data and recapture probability fixed to 0 in years 2001, 2002 and 2003 the survival estimates were ϕ' = 0.40 and ϕ = 0.83 (transient proportion p = 0.518). Moreover, in this analysis, model information theory based on Akaike’s Information Criterion (AICc, Burnham and Anderson, 2002) would have led to an over complex structure of survival selecting a model assuming a temporal variation in ϕ', and consequently in transient proportion p, when gap years were not properly considered (table 3). Discussion Obtaining unbiased estimates of demographic parameters (such as age–dependent survival) is essential for population diagnosis (Williams et al., 2002; Sanz–Aguilar et al., 2016). To achieve informed management decisions concerning biodiversity conservation, therefore, demographic parameters must be accurately estimated. Here we demonstrate that when not handled properly, the combination of unequal time intervals and age dependence in

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capture–recapture models can lead to erroneous estimates of survival and model selection (i.e. biological inference). To avoid the problem of unequal time intervals, sampling protocols should be properly designed. Here we showhow to partially overcome the s problem of uneven intervals in two age–classes capture–recapture models. Adding dummy columns to the encounter history can generally be used when interval lengths are commensurable with a same unit of time during which no change of age class occurs (one month, for instance). However, there might be practical limitations because in some cases this solution would lead to add a great number of dummy occasions and would 'push' estimate of survival probability close to the upper boundary value of 1. In these cases, a log–link function can be especially useful to accommodate heterogeneity in the duration of encounter occasions when uneven periods are of small duration and too many dummy occasions should otherwise be incorporated to account for the unequal time period. However, not all problems can be solved using the approaches suggested. For example, the log–link function works well in rich datasets but might cause numerical problems when data are sparse (Tavecchia et al., 2001). Moreover, by using the log–link function, the effect of temporal covariates cannot be modelled. Also, when survival parameters are fully time–dependent there are still some parameters that cannot be estimated separately, with only their products being identifiable (tables. 1, 2). The longer the period with missing information the higher the number of redundant parameters (tables 1, 2). However, by constraining survival using external covariates, most parameters become identifiable. Here we focused on two age class, single state models, but more complex models such as multistate models or models including multiple age–classes will present additional parameter identifiability problems. Despite these limitations, the solution presented here performed well in relatively simple situations and we recommend its use when age–dependent parameters are incorporated in models with uneven intervals between sampling occasions. Finally, the presence of transient animals can be accommodated in CMR models by using age– dependent models (see the real case, Pradel et al., 1997; Sanz–Aguilar et al., 2010). However, recently developed multi–event models (Pradel, 2005) allow to model transients as a specific uncertain category of individuals with known parameter values (survival probability = 0). In this case, age–dependent models are no longer necessary and the problem does not apply (e.g. Genovart et al., 2012; Santidrián Tomillo et al., 2017). Acknowledgements Gonzalo G. Barberá and Gustavo Ballesteros kindly allowed the re–analysis of Palomas data collected by the Asociación de Naturalistas del Sureste (ANSE) and by volunteers supported by the Dirección General de Medio Natural of the Regional Government

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of Murcia. GT was partially supported by I. Hendriks and the 'Salvador de Madriaga' fellowship (ref. 16/00101). GT and RP were partly supported by a joint grant (PICS INTERACT, ref. 272847) from CNRS in France and CSIC in Spain. ASA was supported by a postdoctoral 'Vicenç Mut' contract co–funded by the Regional Government of the Balearic Islands and the European Social Fund (ref. PD/003/2016). References Bears, H., Martin, K., White, G. C., 2009. Breeding in high‐elevation habitat results in shift to slower life‐history strategy within a single species. Journal of Animal Ecology, 78: 365–375. Bonenfant, C., Gaillard, J., Klein, F., Loison, A., 2002. Sex‐and age‐dependent effects of population density on life history traits of red deer Cervus elaphus in a temperate forest. Ecography, 25: 446–458. Burnham, K. P., Anderson, D. R., 2002. Model selection and multi–model inference: a practical information–theoretic approach. Springer, New York. Catchpole, E. A., Fan, Y., Morgan, B. J., Clutton– Brock, T. H., Coulson, T., 2004. Sexual dimorphism, survival and dispersal in red deer. Journal of Agricultural, Biological, and Environmental Statistics, 9: 1–26. Catchpole, E. A., Morgan B. J., 1997. Detecting parameter redundancy. Biometrika, 84, 187–196. Choquet, R., Rouan, L., Pradel, R., 2009. Program E– SURGE: a software application for fitting multievent models. In: Modelling demographic processes in marked populations: 845–865 (D. L. Thomson, E. G. Cooch, M. J. Conroy, Eds.), Springer series: Environmental and ecological statistics Vol. 3, New York. Clobert, J., Perrins, C. M., McCleery, R. H., Gosler, A. G., 1988. Survival rate in the great tit Parus major in relation to sex, age, and immigration status. Journal of Animal Ecology, 57: 287–306. Cooch, E., 2009. Program MARK: a gentle introduction, http://www. phidot. org/software/mark/docs/book/ Covas, R., Brown, C. R., Anderson, M. D., Brown, M. B., 2002. Stabilizing selection on body mass in the sociable weaver Philetairus socius. Proceedings of the Royal Society London Series B Biological Sciences, 269: 1905–1909. Genovart, M., Pradel, R., Oro, D., 2012. Exploiting uncertain ecological fieldwork data with multi–event capture–recapture modelling: an example with bird sex assignment. Journal of Animal Ecology, 81: 970–977. Grosbois, V., Tavecchia, G., 2003. Modeling dispersal with capture–recapture data: disentangling decisions of leaving and settlement. Ecology, 84: 1225–1236 Hiraldo, F., Negro, J. J., Donázar, J. A., Gaona, P., 1996. A demographic model for a population of the endangered lesser kestrel in southern Spain. Journal of Applied Ecology, 33: 1085–1093. Laake, J. L., 2013. RMark: An R Interface for Analysis of Capture–Recapture Data with MARK. Alaska

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Ecology, 38: 1197–1207. White, G. C., Burnham, K. P., 1999. Program MARK: survival estimation from populations of marked animals. Bird Study, 46: S120–S139. Williams, B., Nichols, J. D., Conroy, M. J., 2002. Analysis and Management of Animal Populations: Modeling, Estimation and Decision Making. Academic Press, London, UK. Zabala, J., Zuberogoitia, I., Martínez–Climent, J. A., Etxezarreta, J., 2011. Do long lived seabirds reduce the negative effects of acute pollution on adult survival by skipping breeding? A study with European storm petrels (Hydrobates pelagicus) during the "Prestige" oil–spill. Marine Pollution Bulletin, 62: 109–115. Zuberogoitia, I., Zabala, J., Etxezarreta, J., Crespo, A., Burgos, G., Arizaga, J., 2016. Assessing the impact of extreme adverse weather on the biological traits of a European storm petrel colony. Population Ecology, 58: 303–313.

Supplementary material Supplementary material S1. R–code for data simulation and analyses using RMARK. Supplementary material S2. Design matrices for solution in MARK and E–SURGE. Supplementary material S3. Implementing the solution using dummy occasions and a log–link in MARK. Supplementary material S4. Datasets 1 and 2 and models (tables 1, 2) implemented in E–SURGE.

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Mammal diversity before the construction of a hydroelectric power dam in southern Mexico M. Briones–Salas, M. C. Lavariega, I. Lira–Torres†

Briones–Salas, M., Lavariega, M. C., Lira–Torres†, I., 2018. Mammal diversity before the construction of a hydroelectric power dam in southern Mexico. Animal Biodiversity and Conservation, 42.1: 99–112, Doi: https:// doi.org/10.32800/abc.2019.42.0099 Abstract Mammal diversity before the construction of a hydroelectric power dam in southern Mexico. Hydroelectric power is a widely used source of energy in tropical regions but the impact on biodiversity and the environment is significant. In the Río Verde basin, southwestern of Oaxaca, Mexico, a project to build a hydroelectric dam is a potential threat to biodiversity. The aim of this work was to determine the parameters of mammals in the main types of vegetation in the Río Verde basin. We studied richness, relative abundances, and diversity of the community in general and among groups (bats, small mammals and medium and large–sized mammals). In the temperate forests, small mammals were the most diverse while medium–sized mammals and large mammals were the most diverse in land transformed by humans. As the Río Verde basin shelters 15 % of the land mammal species of Mexico, if the hydroelectric power dam is constructed, mitigation measures should include rescue programs, protection of the nearby similar forests, and population monitoring, particularly for endangered species (20 %) and endemic species (14 %). In a future scenario, whether the dam is constructed or not, management measures will be necessary to increase forest protection, vegetation corridors and corridors within the agricultural matrix in order to conserve the current high mammal diversity in the region. Key words: Effective number of species, Río Verde basin, Oaxaca, Deciduous forest, Temperate forests, Mitigation Resumen Diversidad de mamíferos antes de la construcción de una presa hidroeléctrica en el sur de México. En las regiones tropicales la energía hidroeléctrica es una de las fuentes de energía más utilizadas; sin embargo, también ha afectado significativamente a la biodiversidad y el ambiente. En la cuenca de Río Verde, al suroeste de Oaxaca, en México, se ha proyectado la construcción de una presa hidroeléctrica que podría poner en peligro la biodiversidad. El objetivo de este trabajo fue determinar los parámetros de la comunidad de mamíferos en los principales tipos de vegetación en la cuenca del Río Verde. Estudiamos la riqueza de especies, las abundancias relativas y la diversidad en la comunidad en general y entre grupos (murciélagos, pequeños mamíferos, y mamíferos de talla mediana y grande). Los mamíferos de talla pequeña fueron los más diversos en los bosques templados, mientras que los de talla mediana y grande lo fueron en las tierras transformadas por los humanos. La cuenca del Río Verde alberga el 15 % de las especies de mamíferos terrestres presentes en México, por lo que si la presa hidroeléctrica se construyera, las medidas de mitigación deberían comprender programas de rescate, la protección de bosques similares cercanos y un control poblacional, en particular de las especies amenazadas (el 20 %) y las endémicas (el 14 %). En el futuro, tanto si se construye la presa como si no, será necesario adoptar medidas de manejo encaminadas a aumentar la protección de los bosques y establecer corredores de vegetación y corredores dentro de la matriz agrícola con vistas a conservar la alta diversidad de mamíferos presente actualmente en la región. Palabras clave: Número efectivo de especies, Cuenca del Río Verde, Oaxaca, Bosque caducifolio, Bosques templados, Mitigación Received: 09 III 18; Conditional acceptance: 07 V 18; Final acceptance: 13 VIII 18 M. Briones–Salas, M. C. Lavariega, I. Lira–Torres†, Centro Interdisciplinario de Investigación para el Desarrollo Integral Regional, Unidad Oaxaca (CIIDIR–OAX), IPN Hornos 1003, Santa Cruz Xoxocotlán, C.P. 71230 Oaxaca, México. Corresponding author: M. Briones–Salas: miguelbrionessalas@hotmail.com ISSN: 1578–665 X eISSN: 2014–928 X

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Introduction Population growth, human activities, and development have triggered the need for greater quantities of non– renewable resources and energy in tropical regions. In these ecosystems, hydroelectricity is a major source of energy, but the serious impact of dams on biodiversity and the environment must be taken into account (Lehner et al., 2011; Tundisi et al., 2014). Dams have a direct impact on hydrology by changing the flow of water to a non–natural, lotic to lentic system. They not only alter the flux sediment, biogeochemical processes and nutrient dynamics, but also affect the thermal regime, homogenizing the system, and affecting primary production. Dams also nullify the migration of aquatic species and flood the habitat of terrestrial species. The cascade effect includes the spread of cosmopolitan non–indigenous species, affecting the native aquatic species and the base of the food web (Dudgeon, 2000; Pringle, 2003; Agostinho et al., 2004; McCartney, 2009; Poff et al., 2007; Nilsson et al., 2005; Winemiller et al., 2016). Cumulative impacts are pollution and overfishing, relocation of human populations, and expanding deforestation associated with new roads and settlements (Dudgeon, 2000; Nilsson et al., 2005; Winemiller et al., 2016). Measures of mitigation, compensation, and restoration are crucial factors to be taken into consideration to alleviate negative impacts on the environment (McCartney, 2009; Winemiller et al., 2016). For these measures to be successful, it is necessary to understand the composition of biological communities, to identify the most potentially vulnerable species, and to consider potential rescue before these facilities are be built (McCartney, 2009). Only with this knowledge can compensation and restoration measurements similar to initial conditions be designed (McCartney, 2009; Winemiller et al., 2016). Southwestern Oaxaca, Mexico is in an area with high biodiversity. However, it is also threatened by the high likelihood of losing a greater quantity of plant and vertebrate species due to habitat loss (Flores–Villera and García–Vázquez, 2014; Navarro–Sigüenza et al., 2014; Sánchez–Cordero et al., 2014). Although in this region there is a Natural Protected Area (the Lagoons of Chacahua National Park, LCNP), areas around the region have been deforested (Contreras et al., 1997; Pérez, 2002). In contrast, several inaccessible areas maintain well–conserved semi–deciduous tropical forest (Lira–Torres et al., 2005). Within this region, which constitutes the Río Verde basin, the construction of a hydroelectric dam is being planned. This dam would directly affect 3,100 hectares in 15 villages and six municipalities inhabited by Mixtec and Chatino people. Besides, it is unknown how dam construction would affect the vast biodiversity of the region. In sites near Río Verde basin, mammalian presence surveys have been conducted (Lira–Torres et al., 2005; Lira–Torres, 2006; Buenrostro–Silva et al., 2012), but a site study during the dam pre–construction phase is needed for later comparison of changes in the diversity of mammal assemblages in response to construction.

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The aims of this work were to compare the parameters of the mammal community (species richness, relative abundance, and alpha diversity) between temperate forest, deciduous forest and agricultural areas in the Río Verde basin, southwestern Oaxaca, Mexico. Information will be useful to guide mitigation, restoration and compensation measures during the implementation of the hydroelectric project. Material and methods Study site The Río Verde basin is an exorheic basin, located in the southwest of the State of Oaxaca, Mexico (15° 56' 55'' N – 6° 18' 15'' N, 97° 26' 23'' W – 97° 58' 36'' W). It has an approximate extension of 1,640 km2 (fig. 1). The climate is warm semi–humid (Aw) and semi– warm semi–humid [(A)C(w)]; annual precipitation is 2,245 mm (Trejo, 2004). The types of forest predominant in the middle of the Río Verde basin are pine forest, oak forest, montane cloud forest, deciduous tropical forest, semi deciduous tropical forest, savannas and areas of agricultural and rangelands. In the lowlands of the basin, deciduous and semi–deciduous tropical forest, areas of agricultural and pastureland, savannas and mangrove prevail (Arriaga et al., 2000; Ortiz–Pérez et al., 2004). Due to the complexity of the terrain, pine forest, oak forest and montane cloud forest fragments are interspersed up to 1,000 m a.s.l. Thus, in this study, these forests were grouped and named temperate forest, covering approximately 37.16 % of the basin (609 km2). The deciduous and semi–deciduous tropical forest was named deciduous forest (below 1,000 m a.s.l.) and cover approximately 27.06 % of the basin (444 km2). Finally, areas with corn crops, plots with fruit trees and pasture lands were grouped and named areas of agricultural areas that cover 33.68 % of the basin (552 km2). Methods We conducted seven sampling visits in the Río Verde basin from January to November 2009. During each period we took at least one sample for each vegetation group at three locations, giving a total of 20 sampled locations, and covering the rainy season and the dry season: six in temperate forest, seven in deciduous forest, and seven in agricultural areas. Each locality was surveyed for three consecutive days. The sites were selected on the basis of the vegetation type and low human presence (fig. 1). Small mammals (< 100 g) were captured using 100 Sherman traps baited with a peanut butter, vanilla essence and oats mixture. Traps were set daily along two 500 m lineal transect. Throughout the study, a total of 3,800 Sherman/trap/days were set up. We also placed 100 pitfall traps, that were separated from each other by about 2 m and situated, in places with leaf litter and near fallen logs. Bats were captured at each site using four mist nets (12 x 2.4 m) that were deployed for seven hours

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Fig. 1. Geographic location of the Río Verde basin, Mexico. Localities surveyed: circles, agricultural areas; diamonds, temperate forests; squares, deciduous forests. Type of vegetation and cover: light gray, agricultural areas; medium gray, deciduous forests; dark gray, temperate forests; black, human settlements. Fig. 1. Localización geográfica de la cuenca del Río Verde, en México. Localidades estudiadas: círculos, zonas agrícolas; rombos, bosques templados; cuadrados, bosques deciduos. Tipo de vegetación y cobertura: gris claro, zonas agrícolas; gris medio, bosques deciduos; gris oscuro, bosques templados; negro, asentamientos humanos.

every night (19:00 a 02:00 h); the total sampling effort for bats was 10,944 m net/h. In the case of medium and large–sized mammals (1,000–10,000 g), two linear transects of approximately 2.5 km in length were distributed randomly at each locality and walked for the signs of tracks and/or feces (Aranda, 2000). A total of 152 km of transects were walked. To complete the inventory, we placed five Tomahawk–type traps, with double–door folding, 24 x 6 x 6. The bait was sardine. In addition, five trap cameras (Cuddeback ®) were used. The cameras were set at a height between 30 and 50 cm from the ground along natural paths, and roads or sites where we observed tracks. Geographical coordinates and elevation were recorded with a global positioning system (GPS; Datum WGS84). Survey efforts were similar in deciduous forest and agricultural areas (1,400 Sherman/trap/ days; 4,032 m net/h and 56 km in each one), while in the temperate forest the survey effort was lower (1,000 Sherman trap/day; 2,880 m net/h and 40 km). In the temperate forest it was not possible to perform a sampling period due to security problems. Mammal individuals were taxonomically determined using specialized keys (Ceballos and Miranda, 1986; Álvarez et al., 1994; Medellín et al., 1997). Nomenclature was updated following Ramírez–Pulido et al.

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(2014). Most individuals were released at the site of capture; only a small sample was prepared as museum specimens following Hall (1981) recommendations. These specimens are deposited in the Mammals Collection at the Centro Interdisciplinario de Investigación para el Desarrollo Integral Regional (CIIDIR), Unidad Oaxaca (OAX.MA.026.0497), Instituto Politécnico Nacional. Specimens were captured and collected with the license for scientific collection issued by the Mexican Secretaría de Medio Ambiente y Recursos Naturales (FAUT–0037; SEMARNAT, 2010). Data analyses Species richness for the whole community and between ensembles was counted as the total number of species at each vegetation type. The species relative abundance was calculated as the quotient of the number of individuals of every species and the survey effort applied to record it (Davis and Winstead, 1987; Medellín, 1993). In the case of small mammals, the effort applied was measured as the number of traps/ day, whereas for bats was the number of m net/hour. Finally, for the medium and large–sized mammals, abundance was estimated considering the number of signs recorded per km walked. To compare the

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patterns of species abundance and composition between the different types of vegetation, we elaborated curves of rank abundance. These graphs are a useful tool to visualize attributes of the assemblage such as species richness (number of points), evenness (slope), number of rare species (tail of the curve), and relative abundance of each species (order of the species in the graph) (Feinsinger, 2001; Avila– Cabadilla et al., 2009). We performed species accumulation curves using the iNEXT software program. The iNEXT performs sample curves based on the Hill numbers, based on rarefaction and extrapolation (Chao et al., 2016). Using iNEXT we computed the non–asymptotic approach due to large and heterogeneous study area. Alpha diversity was estimated with the calculus of the effective number of species (q D), which measures the diversity that a virtual community would have integrated by i species. The values obtained by this diversity index could be interpreted as a virtual community in which all the species have the same abundance. The equation is (Jost, 2007): q

D = (∑S piq)1/(1–q)

where pi is the abundance of the species i divided between the sum of the total of abundances of S species that compose the whole community; the q exponent is the order of the diversity. As this estimator is affected by the abundance of the species, three orders were considered: (i) q = 0, it does not consider the abundances of the species, so it is equivalent to the species richness (0D); (ii) q = 1, all the species are included with a weight exactly proportional to their abundance in the community (1D) exponential of Shannon´s entropy index; (iii) q = 2, it is the inverse of the Simpson index and considers only the commons species, excluding the rare species (2D) (Jost, 2007). Effective number of species allows to measure magnitudes of change in communities (García–Morales et al., 2011; Moreno et al., 2011). In order to balance the variability in the survey effort due to logistic and environmental factors and low detectability of the rare species, we generated models to estimate the diversity in the communities. In the case of diversity 0D, we used the nonparametric abundance–based coverage estimator (ACE). To estimate 1D and 2D, we used the maximum likelihood estimator (MLE; Chao and Shen, 2010). The estimators and standard error were calculated for each type of vegetation and mammal ensemble with SPADE software program (Chao and Shen, 2010). The beta diversity was obtained using Jaccard's qualitative similarity index; the range of values for this index is 0 when there are no shared species between the two sites, and up to 1 when the sites have the same composition, with the unweight pair–group method for arithmetic averages (UPGMA) was used. The measure of magnitudes of change between communities was analyzed through the selection of the cover with the highest value of diversity (orders 0, 1, and 2), and calculating the percentage representing the diversity value of the two remaining covers with respect to this.

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The accumulated species richness for the whole Río Verde basin was obtained by comparing merging species of previous published works (Lira–Torres et al., 2005; Buenrostro–Silva et al., 2012). Conservation status and regulation of the species was consulted in the Norma Oficial Mexicana 059 (NOM–ECOL–059–2010; SEMARNAT, 2010), the Red List of the International Union for Conservation of Nature (IUCN, 2017), and the Appendices of the Convention on International Trade in Endangered Species of Wild Fauna and Flora (CITES, 2014). Results Species richness Species richness differed in the types of forests studied. In the deciduous forest, 43 species, 34 genera, and 16 families were recorded; in the temperate forest there were 31 species, 22 genera, and 20 families; and in the agricultural areas, there were 30 species, 25 genera, and14 families. For the ensembles of bats and small mammals, the highest species richness was observed in the deciduous forests (21 and 12 species, respectively). Finally, for the ensemble of medium and large–sized mammals, the highest species richness was recorded in the agricultural areas (11 species). Relative abundance Among the bats, Artibeus jamaicensis (Jamaican fruit–eating bat) was the most frequent species in both deciduous forest and agricultural areas, while in temperate forest it was Sturnira parvidens (little yellow–shouldered bat) (fig. 2). The small mammals, Peromyscus aztecus (Aztec deermouse), P. mexicanus (Mexican deermouse), and Heteromys pictus (painted spiny pocket mouse) were the most frequent species in the temperate forests, deciduous forest and agricultural areas, respectively (fig. 2). Finally, for medium and large–sized mammals we did not observe any pattern in their relative abundance because in temperate forest three species had the same high relative abundance value. In the deciduous forest, Dasypus novemcinctus (nine–banded armadillo) and Didelphis virginiana (Virginia opossum) were the most frequent species; and D. novemcinctus, Procyon lotor (Raccoon) and D. virginiana in the agricultural areas (fig. 2). Alpha diversity The extrapolation species curves suggest that temperate forests and agricultural areas have a similar species diversity and tend to an asymptote, whereas the deciduous have a higher species diversity and a likelihood to increase. For small mammals differences in species diversity were estimated, but confidence intervals overlap; only agricultural areas tend to an asymptote. The curves for medium and large–sized mammals suggest an asymptote for temperate forests, whereas for deciduous forests and agricultural areas the need for additional surveys was evident (fig. 3).

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Fig. 2. Rank abundance curves of mammal communities in the Río Verde basin, Mexico: A, bats; B, small mammals; C, medium and large–sized mammals; TF, temperate forests; DF, deciduous forests; AA, agricultural areas. Letters on the curves indicate species (see list of abbreviations in table 3). Fig. 2. Curvas de rango–abundancia de las comunidades de mamíferos en la cuenca del Río Verde, en México: A, murciélagos; B, mamíferos pequeños; C, mamíferos de talla mediana y grande; TF, bosques templados; DF, bosques caducifolios; AA, zonas agrícolas. Las letras encima de las curvas indican las especies (véase la lista de abreviaturas en la tabla 3).

Diversity of the order 0 was highest in the deciduous forest in the whole community, and in the ensembles of bats, small mammals, and medium and large–sized mammals. Considering the abundances of the species with the same proportional weight as their abundance in the community (diversity of the order 1) and only for the most common species (diversity of the order 2), we found that in the whole community and in bats that deciduous forest was the most diverse. However, small mammals were the most diverse in

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the temperate forest, while medium and large–sized mammals were most diverse in agricultural areas (table 1; fig. 4). When comparing magnitudes in the diversity, we found that for the order 0, temperate forests and agricultural areas represented between 40 and 77 % of the diversity estimated in deciduous forests. In the diversity of the orders 1 and 2, magnitudes between covers showed a similar pattern. First, the temperate forests had a higher proportion of the diversity esti-

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Fig. 3. Interpolation and extrapolation curves of diversity of species in mammal communities in the Río Verde basin, Mexico: A, whole community; B, bats; C, small mammals; D, medium and large–sized mammals; DF, deciduous forests; TF, temperate forests; AA, agricultural areas. Fig. 3. Curvas de interpolación y extrapolación de la diversidad de especies en las comunidades de mamíferos de la cuenca del Río Verde, en México: A, toda la comunidad; B, murciélagos; C, mamíferos pequeños; D, mamíferos de talla mediana y grande; DF, bosques caducifolios; TF, bosques templados; AA, zonas agrícolas.

Table 1. Diversity values of mammals in the Río Verde basin, Mexico: N, number of individuals; S, observed diversity; 0D, species richness, order 0; 1D, exponential of Shannon entropy index; 2D, inverse of the Simpson index. (The standard error is shown in brackets). Tabla 1. Valores de diversidad de las comunidades de mamíferos en la cuenca del Río Verde, en México: N, número de individuos; S, diversidad observada; 0D, riqueza de especies, orden 0; 1D, exponencial del índice de entropía de Shannon; 2D, inverso del índice de Simpson. (El error estándar está entre paréntesis).

0 1 2 N S D D D Whole community Temperate forests 181 31 37.7 (4.2) 19.56 (1.328) 14.16 (0.157) Deciduous forests 232 43 57.6 (7.9) 23.52 (1.579) 15.96 (0.136) Agricultural areas 189 30 35 (3.2) 17.95 (1.303) 12.09 (0.15) Bats Temperate forests 106 15 17.7 (2) 10.33 (0.872) 7.57 (0.193) Deciduous forests 136 21 23.9 (2.8) 13.24 (1.003) 9.76 (0.171) Agricultural areas 105 12 18.3 (5.4) 7.49 (0.652) 5.61 (0.256) Small mammals Temperate forests 66 10 11.1 (1.7) 6.62 (0.638) 5.35 (0.202) Deciduous forests 76 12 16.7 (4.6) 5.83 (0.703) 4.09 (0.282) Agricultural areas 46 7 7.5 (0.9) 3.87 (0.591) 2.62 (0.224) Medium and large–sized mammals Temperate forests 9 6 9 (3.2) 5.67 (0.618) 5.40 (0.237) Deciduous forests 20 10 20.5 (9.8) 7.71 (1.26) 6.06 (0.247) Agricultural areas 38 11 11.5 (0.9) 9.22 (0.865) 7.93 (0.096)

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DF Whole community

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Fig. 4. Diversity index of the mammal communities in the Río Verde basin, Mexico: TF, temperate forests; DF, deciduous forests; AA, agricultural areas. Fig. 4. Índices de diversidad de las comunidades de mamíferos en la cuenca del Rio Verde, en México: TF, bosques templados; DF, bosques caducifolios; AA, zonas agrícolas.

mated in deciduous forests for the whole community (order 1 = 83 %, order 2 = 89 %) and for bats (order 1 = 78 %, order 2 = 78 %) than the agricultural areas (76 % and 76 %, and 57 % and 58 %, respectively). We also found that deciduous forests had a higher proportion of diversity of small mammals (order 1 = 88 %, order 2 = 77 %) in temperate forests than in agricultural areas (order 1 = 58 %, order 2 = 49 %). Finally, the deciduous forests had higher a proportion of diversity of medium and large–sized mammals (order 1 = 84 %, order 2 = 76 %) in agricultural areas than in temperate forests (order 1 = 62 %, order 2 = 68 %) (fig. 5). Beta diversity Nineteen species were shared in the three types of cover studied. Six species were found in the temperate

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forest only, 15 in the deciduous forest only, and six in the agricultural areas only (table 2). In the whole community, the highest similitude was observed between temperate forest and agricultural areas (0.488). For the ensemble of bats, the highest similitude was between temperate forest and agricultural areas (0.588). For the ensemble of small mammals, the highest similitude was observed between temperate forests and deciduous forests (0.467). Finally, for the ensemble of medium and large–sized mammals, the highest similitude was between temperate forest and agricultural areas (0.416) (table 2). On the other hand, dendograms had the same shape for the whole community, for bats and for the medium and large– sized mammals: a node formed by the temperate forest and agricultural areas. In the case of the small mammals, the temperate forests and the deciduous forests formed a group (fig. 6).

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Diversity order 0 Temperate forests (65 %)

Whole community

Deciduous forests diversity = 57.6 (100 %)

Agricultural areas (61 %) Temperate forests (74 %)

Bats

Deciduous forests diversity = 23.9 (100 %)

Agricultural areas (77 %) Temperate forests (66 %)

Small mammals

Deciduous forests diversity = 16.7 (100 %)

Agricultural areas (45 %)

Temperate forests (44 %) Medium and large–sized mammals Agricultural areas (56 %) 0 10 Diversity order 1

Deciduous forests diversity = 20.5 (100 %) 60

Temperate forests (683 %)

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Deciduous forests diversity = 23.56 (100 %)

Agricultural areas (76 %) Temperate forests (78 %)

Bats

Deciduous forests diversity = 13.2 (100 %)

Agricultural areas (57 %) Temperate forests (58 %)

Small mammals

100

Agricultural areas (88 %)

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Medium and Temperate forests (62 %) large–sized mammals Agricultural areas (84 %)

Agricultural areas diversity = 9.2 (100 %)

0 10 Diversity order 2 Whole community

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100 Deciduous forests diversity = 16.0 (100 %)

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Deciduous forests diversity = 9.8 (100 %)

Agricultural areas (58 %)

Deciduous forests (77 %)

Temperate forests diversity = 5.3 (100 %)

Medium and Temperate forests (68 %) large–sized mammals Deciduous forests (76 %)

Agricultural areas diversity = 7.9 (100 %)

Small mammals

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50 60 70 80 Percentage

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Fig. 5. Magnitudes of the diversity of mammal communities in the Río Verde basin, Mexico. Fig. 5. Magnitudes de la diversidad de las comunidades de mamíferos en la cuenca del Río Verde, en México.

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Accumulated species richness In this study we report the presence of 58 mammal species, 19 of which were not recorded in previous surveys (table 1s in supplementary material). If we consider the 52 species reported by Lira–Torres et al. (2005), and the 42 species by Buenrostro–Silva et al. (2012), the mammalian accumulated species richness for the Río Verde basin is 73 species, belonging to 56 genera, 24 families and 10 orders. Conservation status According to the Mexican Official Norm 059 (SEMARNAT, 2010), five species are Endangered (Tamandua mexicana, Leopardus pardalis, L. wiedii, Potos flavus, Tapirella bairdii), six species are Threatened (Leptonycteris nivalis, L. yerbabuenae, Coendou mexicanus, Herpailurus yagouaroundi, Spilogale pygmaea and Lontra longicaudis), and two are Subject to special protection (Enchisthenes hartii, Bassariscus sumichrasti). In the Red List of the IUCN, five species are Endangered, two species are Vulnerable and two species are Near threatened. The CITES Appendices included four species in Appendix I, one species in Appendix II, and five species in Appendix III (table 1s in supplementary material). Discussion Species richness and composition The Río Verde basin is located within a region of high biodiversity (Olson and Dinerstein, 1998; Mittermeir et al., 2011). With respect to mammalian species richness, in Mexico there are 496 species (Ramírez–Pulido et al., 2014), 73 of which (14.7 %) inhabiting the Río Verde basin were collected in this study or are recorded in literature. The mammalian species richness accumulated in this basin is higher than reported for any other site along the Mexican Pacific coast (59–70 species; Ceballos, 1995; Cervantes and Yépez, 1995; Lira–Torres et al., 2008; López et al., 2009; Briones–Salas et al., 2016). This high species richness could be explained by the latitudinal pattern of mammalian species richness along the Pacific coast, which increases as latitude decreases (Ceballos, 1995). The landscape heterogeneity, with several types of vegetation in the study site, also contributes to the high species richness. In the Río Verde basin, we found differences in species richness between the forests studied, with richness being higher in the deciduous forests than in the temperate forests or agricultural areas. In particular, agricultural areas had 33 % fewer species than deciduous forests; differences between this cover land were most notable for bats (43 %) and small mammals (42 %). Furthermore, the effective number of species shows a similar pattern of loss of diversity in these mammal groups. This loss of mammal diversity due to change of land use has frequently been observed in the neotropics, where the degree of change and configuration of the landscape has been seen to affect

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Table 2. Affinity matrix of the mammalian species in the different types of vegetation and land use in the coast of Oaxaca, México: TF, temperate forests; DF, deciduous forests; AA, agricultural areas. (The numbers in bold correspond to the total species in each forest, and the exclusive species are shown in brackets). Table 2. Matriz de afinidad de las especies de mamíferos en los diferentes tipos de vegetación y uso de suelo en la cuenca del Rio Verde, México: TF, bosques templados; DF, bosques caducifolios; AA, zonas agrícolas. (Los números en negritas corresponden al total de especies en cada tipo de cobertura y ente paréntesis el número de especies exclusivas).

Whole community Temperate forests

31(6) 0.480 0.488

Deciduous forests

Agricultural areas

43(15) 0.460 4

30(6)

Bats Temperate forests

15(2) 0.565 0.588

Deciduous forests

Agricultural areas

21(6) 0.571 2

12(0)

Small mammals Temperate forests

10(3) 0.467 0.416

Deciduous forests

Agricultural areas

12(4) 0.461 0

7 (1)

Medium and large–sized mammals Temperate forests

6(1) 0.333 0.416

Deciduous forests

Agricultural areas

10(5) 0.312 1

11(5)

both species richness and their abundance (Estavillo et al., 2013; Roque et al., 2018). Although the temperate forests had few small mammal species, relative abundance was distributed more evenly than for species in the deciduous forests or in the agricultural areas. In turn, the deciduous forests presented a greater species richness and a highest number of rare species. In contrast, the agricultural areas were characterized by one very dominant species and lower species richness. In this study, the small species loss reached 50% between deciduous forests and agricultural areas. Such results fit findings from one of the most notable studies in human–altered environments, where a small number of species benefit from the disturbances while other more sensitive species disappear (McKinney and Lockwood, 1999; McGill et al., 2015).

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Similarity 0.3 0.4 0.5 0.6 0.7 0.8 0.9 1

Agricultural areas

Temperate forests

0.470

Deciduous forests Whole community

0.323

0.568

Deciduous forests Bats

Agricultural areas

Temperate forests

Deciduous forests

Medium and large–sized mammals

0.439

Deciduous forests Small mammals

Fig. 6. Specific similarity in mammal communities in the Río Verde basin, Mexico: A, whole community; B, bats; C, small mammals; D, medium and large–sized mammals. Fig. 6. Similitud específica en las comunidades de mamíferos de la cuenca del Río Verde, en México: A, toda la comunidad; B, murciélagos; C, mamíferos pequeños; D, mamíferos medianos y grandes.

With respect to the ensemble of bats, we found that the deciduous forest and the agricultural areas showed a similar pattern in the relative abundance of species, A. jamaicensis was the dominant species, followed by S. parvidens. A. jamaicensis showed higher relative abundance in sites within agricultural areas, a finding coincides with other studies that established that abundance of A. jamaicensis increases with the level of perturbation (Fenton et al., 1992; Vargas et al., 2008; Murillo–García and Bedoya–Durán, 2014). Gorresen and Willing (2004) suggest that adaptability of Artibeus to perturbation is due to its ability to perform long flights, which allows them to explore large fragments of vegetation within the landscape. Another similitude with the works cited was the high frequency of Desmodus rotundus (common vampire bat) in agricultural areas, a consequence of the highest availability of food (Fenton et al., 1992). For these reasons, A. jamaicensis and D. rotundus are recognized as able to adapt to habitat fragmentation and as indicator of sites with perturbation (Wilson et al., 1996; Galindo– González, 2004). In contrast with other neotropical regions, in this study, rare species, such as Phyllostominae subfamily species, which are good indicators of non–perturbed sites, were not recorded (Wilson et al., 1996; Castro– Luna et al., 2007). This is because on the Mexican Pacific coast there are currently no representatives of this subfamily. Instead, the composition of nectarivorous species (the Glossophaginae subfamily) changed between deciduous forests and agricultural areas, with higher species richness and abundance in the former. Although overall nectarivorous species are adaptable

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to human land–use, their response to the type of perturbation is variable, showing a preference for an agroforestry crop system when compared with well– preserved forest; these bats select well–preserved forest over monocultures, silvopastoril systems or induced grasslands (García–Morales et al., 2013). The usefulness of Glossophaginae species as an indicator group in deciduous forest throughout the Pacific coast should be explored in further studies. In the ensemble of small mammal, the species with higher relative abundance are known to be common in the forests surveyed; Peromyscus aztecus in temperate forests at elevations from 1,000 to 2,700 m a.s.l. (Vázquez et al., 2001) and P. mexicanus in deciduous forest (Trujano–Alvarez and Alvarez–Castañeda, 2010). Heteromys pictus occupied place regarding highest relative abundance in both deciduous and temperate forests but dominated broadly in the ensemble of small mammals in the agricultural areas. This species is capable of taking advantage of secondary vegetation, and agriculture and pasture lands, with higher densities due to the high availability of food (Briones–Salas and González–Pérez, 2016). Unexpectedly, the agricultural areas had the highest species richness and relative abundances of medium and large–sized mammals. These findings, however, could be an artifact of the sampling technique, the search for tracks. Tracks are more visible on uncovered terrain in the agricultural areas and species common and tolerant to perturbations (e.g. Urocyon cinereoargenteus, gray fox; Odocoileus virginianus, white–tailed deer) usually visit this types of land cover in search of food (Lira–Torres, 2006).

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Despite the sampling technique, endemic (Spilogale pygmaea, pygmy skunk) and endangered species (Potos flavus, H. yagouaroundi, jaguarondi; and Tamandua mexicana, Northern anteater) were only recorded in the deciduous forests by means of tracks and visual observations. In turn, the temperate forests had few species, all of which are common in several types of ecosystems (Briones–Salas et al., 2015). Surveys for medium and large–sized mammals in the three cover types could be improved with the use of complementary techniques such as camera trapping (Silveira et al., 2003; Cortés–Marcial et al., 2014). Species diversity Effective number of species showed a generalized loss of diversity with respect to land–use change for agricultural purposes, except for the ensemble of medium and large–sized mammals, which showed the highest diversity in this type of cover. For the ensemble of bats, both species richness and diversity were highest in the deciduous forests. Differences in diversity in perturbed and unperturbed sites have been found in Yucatán, Mexico (Fenton et al., 1992), but near the study site, Barragán et al. (2010) found no difference in the abundance and diversity of small mammals and bats with respect to perturbation. Castro–Luna et al. (2007) found similar results in successional stages of vegetation, with no differences in diversity. Likewise, in a semi–deciduous forest in Nicaragua, Medina et al. (2007) did not find any differences in diversity between perturbed and unperturbed sites. However, in these studies, estimators that do not allow a direct comparison were applied. Thus, a reanalysis could give different conclusions (Moreno et al., 2011). In this study, the loss of bat diversity was noteworthy, with a difference of nine species between deciduous forests and agricultural areas, and a decrease in the diversity of order 1 in 23.7 % in agricultural areas. On the contrary, diversity of medium and large–sized mammals was highest in sites with agricultural areas, followed by deciduous forests and temperate forests. As mentioned above, sampling favored agricultural areas, affecting measures of diversity. Using complementary methods, tracks and camera trapping in the Isthmus of Tehuantepec, Oaxaca, Cortés–Marcial et al. (2014) found that the diversity of this group of mammals was high in low–degradation environments and in environments with a low density of livestock. Conservation status Thirteen species (18.05 %) have a certain level of protection. Seven of these species belong to the Carnivora order; two are Leopardus genus, whose main threat is habitat loss and the illegal hunting for their fur (Aranda, 2005a, 2005b). The presence of Tapirella bairdii (Baird’s tapir), an endangered species, is also noteworthy because the study site could host the northern–most population (Lira–Torres et al., 2006). The protection of this and other endangered species, such as Panthera onca (jaguar), was the impetus for creating the LCNP (Mexican Government,

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1937). However, habitat loss within and around the park has been significant. Management implications Due to the planned construction of a dam and the current land–use change rates in the Río Verde basin (Salas–Morales and Casariego–Madorell, 2010), it is necessary to apply conservation policies that guarantee the functionality of the ecosystem and the perpetuity of wildlife populations and to take strong actions to protect the endangered species. Given the imminent hydroelectric development in the study site, as a mitigation measureand on the basis of the results here presented, action should be focused on avoiding loss and fragmentation of the tropical deciduous forests.Furthermore, connectivity through corridors into agricultural areas should be promoted in this type of land cover (Estrada and Coates–Estrada, 2001). As a final consideration, during the dam filling several species of low mobility such as small mammals would likely drown. A wildlife rescue program should therefore be established for this phase. Particularly for terrestrial fauna, wildlife rescue is an undeniable measure. During the filling of the Chiew Larn dam in Thailand, for example, 1,364 animals were captured and translocated (Nakhasathien, 1989), and in the construction of the Petit Saut dam in French Guiana 5,500 animals were rescued (Vié, 1999). A compensatory measure should be to establish a protected area of fauna and flora through partnership between government and local communities. The protected area should have a similar extension to the dam and provide the same type of covers. Such an area will help the translocation actions and serve as a refuge to displaced animals. In addition, considering these translocations and the displacement of medium and large–sized mammals to new areas, a post–construction survey could be advisable for management and conservation. The carrying capacity of habitats may be another issue of concern. Acknowledgements Comisión Federal de Electricidad and Secretaría de Investigación y Posgrado, Instituto Politécnico Nacional (SIP: 20090672, 20100263) by the support to field surveys. To B. Riveros–Lara, Y. Martínez–Ayón, J. García, N. Chávez, and A. Sánchez by the invaluable collaboration in field. M. Briones–Salas thanks the support of the Comisión y Operación de Fomento Actividades Académicas (COFAA) and Programa de Estímulos al Desempeño a la Investigación (EDI), Instituto Politécnico Nacional and Sistema Nacional de Investigadores (SNI). References Agostinho, A. A., Thomaz, S. M., Gomes, L. C., 2004. Threats for biodiversity in the floodplain of the Upper Paraná River: effects of hydrological regulation

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Silveira, L., Jacomo, A. T., Diniz–Filho, J. A. F., 2003. Camera trap, line transect census and track surveys: a comparative evaluation. Biological Conservation, 114(3): 351–355. Trejo, I., 2004. Clima. In: Biodiversidad de Oaxaca: 67–85 (A. J. García–Mendoza, M. J. Ordóñez, M. Briones–Salas, Eds.). Instituto de Biología, UNAM–Fondo Oaxaqueño para la Conservación de la Naturaleza–World Wildlife Fund. Mexico City. Trujano–Álvarez, A. L., Álvarez–Castañeda, S. T., 2010. Peromyscus mexicanus (Rodentia: Cricetidae). Mammalian Species, 42(1): 111–118. Tundisi, J. G., Goldemberg, J., Matsumura–Tundisi, T., Saraiva, A. C., 2014. How many more dams in the Amazon? Energy Policy, 74: 703–708. Vargas, A., Aguirre, L. F., Galarza, M. I., Gareca, E., 2008. Ensamble de murciélagos en sitios con diferente grado de perturbación en un bosque montano del Parque Nacional Carrasco, Bolivia. Mastozoología Neotropical, 15: 297–308. Vázquez, L. B., Cameron, G. N., Medellín, R. A., 2001. Peromyscus aztecus. Mammalian Species, 649: 1–4. Vié, J. C., 1999. Wildlife rescues–the case of the Petit Saut hydroelectric dam in French Guiana. Oryx, 33: 115–126. Wilson, D. E., Ascorra, C. F., Solari, S., 1996. Bats as indicators of habitat disturbance. In: Manu: The Biodiversity of Southeastern Peru: 613–625 (D. E. Wilson, A. Sandoval, Eds.). Smithsonian Institution Press, Lima. Winemiller, K. O., McIntyre, P. B., Castello, L., Fluet–Chouinard, E., Giarrizzo, T., Nam, S., Stiassny, M. L. J., 2016. Balancing hydropower and biodiversity in the Amazon, Congo, and Mekong. Science, 351(6269): 128–129.

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Haplotypic characterization of the olive ridley turtle (Lepidochelys olivacea) in northwest Mexico: the northernmost limit of its distribution S. Campista León, J. A. Beltrán Espinoza, I. Sosa Cornejo, H. Castillo Ureta, J. R. Martín del Campo Flores, J. G. Sánchez Zazueta, L. I. Peinado Guevara Campista León, S., Beltrán Espinoza, J. A., Sosa Cornejo, I., Castillo Ureta, H., Martín del Campo Flores, J. R., Sánchez Zazueta, J. G., Peinado Guevara, L. I., 2019. Haplotypic characterization of the olive ridley turtle (Lepidochelys olivacea) in northwest Mexico: the northernmost limit of its distribution. Animal Biodiversity and Conservation, 42.1: 113–126, Doi: https://doi.org/10.32800/abc.2019.42.0113 Abstract Haplotypic characterization of the olive ridley turtle (Lepidochelys olivacea) in northwest Mexico: the northernmost limit of its distribution. The olive ridley sea turtle (L. olivacea) has a pantropical distribution. In the Eastern Pacific, the official limits of its reproduction area are south of the Baja California peninsula and south of Sinaloa, Mexico. Ceuta beach in Elota, Sinaloa, has served as a protection site for L. olivacea for over three decades. In this study, the L. olivacea population from Ceuta beach was genetically characterized. Specifically, a 712–bp fragment from the control region of mtDNA was amplified from 32 olive ridley turtles. Eight haplotypes (seven after cutting to ~468 bp) were identified, and these included two novel haplotypes (Lo–T7 and Lo–T8) and five haplotypes that were previously identified in other nesting beaches. The Lo–T2 haplotype was dominant (~60 %) in the samples: h = 0.6048 (± 0.0974) and π = 0.002212 (± 0.001504). Although this study was conducted in the northernmost limit of the olive ridley turtle nesting distribution in the eastern Pacific, the sampled group presents moderate genetic diversity and belongs to a population that, on an evolutionary scale, only recently underwent demographic expansion. Because the olive ridley turtle in the eastern Pacific is considered resilient to environmental variation, nesting area studies in northwest Mexico are necessary. Key words: Endangered species, mtDNA, Control region (D–loop), Haplotypic and nucleotidic diversity, Olive ridley turtle Resumen Caracterización haplotípica de la tortuga golfina (Lepidochelys olivacea) en el noroeste de México: el límite septentrional de su distribución. La tortuga golfina (L. olivacea) tiene una distribución pantropical. En el Pacífico oriental, los límites oficiales de su zona de reproducción son la península de Baja California y el sur de Sinaloa, en México. La playa de Ceuta en Elota en Sinaloa, ha servido de sitio de protección para L. olivacea durante más de tres decenios. En este estudio, se caracterizó genéticamente la población de L. olivacea de la playa de Ceuta. Concretamente, se amplificó un fragmento de 712 pb de la región de control del ADNmt de 32 tortugas golfinas. Se identificaron ocho haplotipos (siete tras reducir a ~468 pb) y se incluyeron dos haplotipos nuevos (Lo–T7 y Lo–T8) y cinco haplotipos que se habían identificado anteriormente en otras playas de anidación. El haplotipo Lo–T2 era dominante (~60 %) en las muestras: h = 0,6048 (± 0,0974) y π = 0,002212 (± 0,001504). Si bien este estudio se realizó en el límite septentrional de la zona de anidación de la tortuga golfina en el Pacífico oriental, el grupo estudiado presenta una diversidad genética moderada y pertenece a una población que, en la escala evolutiva, ha pasado recientemente por una expansión demográfica. Debido a que la tortuga golfina del Pacífico oriental se considera resiliente a la variación ambiental, es necesario estudiar las zonas de anidación en el noroeste de México. Palabras clave: Especie amenazada, ADNmt, Región de control (bucle–D), Diversidad haplotípica y nucleotídica, Tortuga golfina

ISSN: 1578–665 X eISSN: 2014–928 X

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Received: 31 V 18; Conditional acceptance: 06 VIII 18; Final acceptance: 03 IX 18 Samuel Campista León, Juan Antonio Beltrán Espinoza, Ingmar Sosa Cornejo, Hipólito Castillo Ureta, Jorge Guillermo Sánchez Zazueta, Luz Isela Peinado Guevara, Facultad de Biología, Universidad Autónoma de Sinaloa, Av. Universitarios s/n., Ciudad Universitaria, Culiacán Rosales, Sinaloa 80040, México.– Jesús Rodolfo Martín del Campo Flore, Laboratorio de Biología Molecular, Centro de Investigación en Alimentación y Desarrollo (CIAD), Av. Sábalo Cerritos s/n., Mazatlán, Sinaloa 82110, México. Corresponding author: Luz Isela Peinado Guevara. E–mail: luzipg@uas.edu.mx

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Introduction The olive ridley turtle (Lepidochelys olivacea) is the most abundant sea turtle in the world and has a pantropical distribution. Its main nesting beaches are located on the east coast of India and in the eastern Pacific (Abreu–Grobois and Plotkin, 2008). In the Mexican Pacific, olive ridley turtles nest from the peninsula of Baja California Sur to the state of Chiapas (Márquez, 1990). The nesting beach El Verde Camacho in Sinaloa is considered the northern limit of its nesting range in the mainland portion of the eastern Pacific (Abreu– Grobois and Plotkin, 2008), with nests observed since 1974 (Ríos–Olmeda, 2005). Although some sporadic nesting has been reported in regions with greater latitude, such as the Upper Gulf of California, El Verde Camacho is by far the most common nesting site in the Gulf of California (Seminoff and Nichols, 2007). The olive ridley turtle has been observed in the Gulf of California, mainly for feeding, but fishermen and residents of the high and middle regions of the Gulf of California have reported that nests have been commonly observed for 50 years ago (Rodríguez– Valencia et al., 2005). The first reports date back to 1961 (Seminoff and Nichols, 2007), and other reports indicate observations of nests in 1995 and 1996 in Puerto Peñasco, Sonora; in 2004 in San Carlos, Guaymas, Sonora; and more recently in El Desemboque, Sonora (CEDO, 1995; COMCAÁC, 2013; Navarro, 1996; Rodríguez–Valencia et al., 2005; Seminoff and Nichols, 2007) (fig. 1). However, due to a lack of documentation, it is difficult to assess how nesting numbers have changed in these regions. Moreover, it is unclear if the nesting observed here is a result of recolonization due to management and conservation strategies implemented in the last decade or whether environmental variations have induced changes in the nesting behavior of this species. The olive ridley turtle is currently listed as a 'Vulnerable' species according to the IUCN Red List of Threatened Species (Abreu–Grobois and Plotkin, 2008) and as 'In danger of extinction' according to Mexican law (SEMARNAT, 2010). Conservation biology serves as a tool in decision–making regarding management of endangered species and focuses on the maintenance of genetic diversity at different levels, a major component of biodiversity (Hunter and Gibbs, 2007). Bottlenecks are an indicator of a loss of genetic diversity and thus constitute a threat to the conservation of many species (Mills, 2006). Genetic markers differ at the molecular level, and special attributes make these markers suitable for examining the life history and evolution of sea turtles (Bowen and Karl, 1996). One of these markers is mitochondrial DNA (mtDNA) from the maternal lineage. A rapidly evolving segment of this marker is the control region, called the D–loop (Avise, 1995), which is the replication start site of the mtDNA. The high mutation rate of this region allows fine–scale identification of populations. Phylogenetic studies of the olive ridley turtle using mtDNA have revealed four lineages at a global level: east Pacific, Indo–west Pacific, Atlantic, and east coast of India (Bowen et al., 1998; Shanker et al., 2004). Bowen et al. (1998) suggested that the haplotypic

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diversity of the olive ridley turtle is classified globally between 'moderate' and 'low' in comparison with that of other species of sea turtles. Briseño–Dueñas (1998) indicated that the genetic heterozygosity of the olive ridley turtle population in the Mexican Eastern Pacific showed no erosion due to a bottleneck effect despite its overexploitation in the second half of the twentieth century, and proposed that this species forms a panmictic population, characterized by no genetic differentiation among nesting colonies. However, Briseño–Dueñas (1998) did not include samples of the nesting area in Baja California Sur. In contrast, López–Castro and Rocha–Olivares (2005) argued that the rookery of Baja California Sur shows low genetic differences that are nevertheless significant with respect to turtles from nesting beaches in mainland Mexico (ΦST = 0.048; p = 0.006). Rodríguez–Zárate et al. (2013) analyzed 10 microsatellite loci in 18 nesting sites along Mexico’s Pacific coast (including the Baja California peninsula and mainland) and identified a clear signature of recent bottlenecks associated with changes in allelic diversity, very low levels of differentiation, and no clear geographical pattern in the population structure. An analysis of molecular variance (AMOVA) did not indicate a significant structure between rookeries of the Baja California peninsula and those of the mainland (FST = 0.0004, P = 0.595) (Rodríguez–Zárate et al., 2013). The available reports of genetic diversity in the northern nesting limits of the species originated from the nesting beach El Verde Camacho (n = 15), where four haplotypes have been identified and haplotype and nucleotide diversities appeared to be higher than those in other Mexican Eastern Pacific rookeries (h = 0.6190 ± 0.1196 and π = 0.0022 ± 0.0017, respectively) (Briseño–Dueñas, 1998; Lopez–Castro and Rocha–Olivares, 2005). In contrast, at the northern limits of Baja California Sur, López–Castro and Rocha–Olivares (2005) identified five haplotypes and lower haplotype and nucleotide diversities than those reported for other nesting beaches in the Mexican Pacific (h = 0.1613 ± 0.0715 and π = 0.0005 ± 0.0007). For conservation purposes, management units (MUs) have been defined for populations showing significant divergence of allele frequencies at nuclear or mitochondrial loci, regardless of the phylogenetic distinctiveness of the alleles (Moritz, 1994). In addition, multi–scale regional management units (RMUs) have been developed to evaluate threats, identify areas of high diversity, highlight data gaps and assess the conservation status of sea turtles above the level of nesting populations but below the level of species (Wallace et al., 2010). On the basis of mtDNA and nDNA information, eight RMUs have been defined for olive ridley turtles worldwide (Wallace et al., 2010). The solitary olive ridley nesting turtles from the eastern Pacific belong to a single RMU distinct from that of 'arribada' (Wallace et al., 2010), and are highly resilient to environmental variations (according to traits such as rookery vulnerability, population trends and genetic diversity) in comparison to the olive ridley turtles from RMU belonging to the Indian Ocean regions and to the RMUs of other species, such as the Kemp's ridley turtle (Lepidochelys kempii) from the Gulf of Mexico (Fuentes et al., 2013).

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The nesting beaches of olive ridley turtles in northwestern Mexico can be classified as follows: (a) beaches with protection activities under a legal framework; (b) beaches with protection activities but without a legally protected status; and (c) potential nesting areas that require evaluation (Márquez et al., 2004) (fig. 1). Ceuta beach, Sinaloa, has legally defined protection activities and is cataloged as a sanctuary for olive ridley nesting areas (DOF, 2002). This 37–km–long beach is located between the Cospita River (24º 10' N and 107º 20' W) and the Elota River (23º 52' N and 106º 57' W) in the central region of the State of Sinaloa, Mexico (fig. 1). According to records dating back to 1976 (Sosa et al., 2012), Ceuta beach is far north of the limit of the nesting range for olive ridley turtles. In the present study, the olive ridley turtle population from Ceuta beach was genetically characterized based on the control region of mtDNA. Material and methods Sample collection During the 2014 and 2015 seasons, 32 samples were collected from olive ridley turtles in the nesting sanctuary at Ceuta beach: five blood samples were extracted from the dorsal cervical sinus of nesting females (Owens and Ruiz, 1980), and 27 skin samples were taken from dead hatchlings. The blood samples were collected in Vacutainer tubes containing 7.2 mg of K2EDTA and stored at –20 ºC until analysis, and the skin samples were preserved in 97 % ethanol. DNA extraction, amplification and sequencing Total genomic DNA was extracted using the Wizard® SV Genomic DNA Purification System (Promega, USA) according to the manufacturer's instructions. An ∼800–bp fragment was amplified from the mtDNA control region using the primers H950 (5'GTC TCG GAT TTA GGG GTT T3') and LTEi9 (5'GAA TAA TCA AAA GAG AAG G3') designed by Abreu–Grobois et al. (2006). Amplification was performed in a thermocycler (T100TM Thermal Cycler, BioRad, USA) with an initial step of 94 ºC for 2 min followed by 30 cycles of 94 ºC for 30 s, 50 ºC for 30 s and 72 ºC for 1 min and a final extension step at 72 ºC for 7 min (modified from Abreu–Grobois et al. 2006). The amplified product was analyzed via electrophoresis in a 1.9 % agarose gel stained with GelRed (BIOTIUM) and displayed on an Imaging system (Digi Doc–It, UVP, USA). The PCR products obtained were gel extracted, purified with the Wizard® SV Gel and PCR Clean–up System (Promega, USA) and sent for sequencing in both directions to Macrogen Inc®. Data analysis An exhaustive search for L. olivacea mtDNA control region haplotypes identified a total of 72 haplotypes of varying sizes. Of these, 52 were previously published or were included in the GenBank database, 13 were identified by Briseño–Dueñas in 1998 (unpublished)

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and the other seven were haplotypes investigated in this study. The sequences were trimmed to a common length of ∼468 bp (Bowen et al., 1998) and then aligned and compared using the web–based program Multalin (Corpet, 1988) for haplotype identification. The haplotypes reported by Shanker et al. (2004), which presented a shorter length (∼399 bp), were analyzed only in the mutated nucleotide positions. To analyze the grouping of the haplotypes, the DnaSP program was used (Librado and Rozas, 2009). The haplotype (h) and nucleotide (π) diversities were calculated using Arlequin ver. 3.5.1.2 (Excoffier and Lischer, 2010). Historical demographics (colonization or bottlenecks) were analyzed based on mismatch distributions using DnaSP (Librado and Rozas, 2009), and the parameters of the sudden expansion model were calculated as follows: τ = 2μt, θ0 = 2μN0 before expansion and θ1 = 2μN1 after expansion, where μ is the fragment–specific mutation rate, t is the time since expansion in generations, and N is the effective population size (Rogers and Harpending, 1992). The evolutionary relationships of the identified haplotypes were identified using the maximum likelihood and neighbor–joining method for phylogenetic reconstruction based on the model suggested by the Model test module of MEGA ver. 7 (Kumar et al., 2016); node support for both analyses was assessed through nonparametric bootstrap analysis (1,000 replicates). The phylogenetic tree was constructed with 20 previously reported haplotypes representative of each region of the world and those identified in the present study. The L. kempii mitochondrial control region sequence was included as an outgroup (GenBank accession AF051777). Additionally, a minimum spanning network (MSN) was constructed using the median joining option in the Network ver. 4.6.1.1 program (Bandelt et al., 1999) to compare the evolutionary relationships between the identified haplotypes. Results A 712 bp mtDNA control region sequence was obtained from 32 olive ridley turtles from Ceuta beach. Comparison of these sequences revealed seven variable sites, namely, five transitions, one transversion and one indel (table 1). Eight haplotypes were identified (labeled Lo–T1 to Lo–T8) and submitted to GenBank (accession numbers KX768696 to KX768698 and KX812518 to KX812522). The dominant haplotype was Lo–T2, which was present in 62.5 % of the samples, followed by Lo–T4, which was present in 9.4 % of the samples (table 1). All haplotypes were trimmed to ∼468 bp fragments. After trimming, the 65 haplotypes identified previously (published or present in GenBank) decreased to 38, and the haplotypes identified in this study were reduced to seven because Lo–T1 and Lo–T5 were identical in the resulting segment (table 2; these haplotypes were plotted on a map, shown in fig. 2). Two of the seven haplotypes were identified as novel (Lo–T7 and Lo–T8) and were identified in 6.3 and 3.1 %, respectively, of the samples, whereas the other five were reported

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115º W

110º W

105º W

Puerto Peñasco

30º N

El desemboque

f ul

C D

A B

San Carlos

Gulf of California 0

ia rn ifo al

y wa 15 gh o Hi xic Me

25º N

ac h

Study area Ceuta beach

La Cruz

320

480 km

115º W

25º N

El Verde Camacho

Ceuta

160

Meseta de Cacaxtla

110º W

105º W

Fig. 1. Study area and main nesting beaches for the olive ridley turtle in northwestern Mexico (adapted from Márquez et al. (2004) and digitized Google Earth images): A, protected without a legal framework; B, protected under a legal framework; C, potential; D, sporadic nesting. Fig. 1. Zona de estudio y principales playas de anidación de la tortuga golfina en el noroeste de México (adaptado de Márquez et al. (2004) e imágenes de Google Earth digitalizadas): A, protegida sin un marco legal; B, protegida bajo un marco legal; C, potencial; D, anidación esporádica.

Table 1. Variable sites in the mtDNA control region and frequencies of the eight haplotypes of olive ridley turtles identified in Ceuta beach, Sinaloa. The upper numbers represent the position of the base, and the first nucleotide corresponds to site 15,554 of the mitochondrial genome of the olive ridley turtle (GenBank accession number: AM258984): – indels (the novel haplotypes are marked in bold). Tabla 1. Sitios variables en la región de control del ADNmt y frecuencias de los ocho haplotipos de tortuga golfina identificados en la playa de Ceuta, en Sinaloa. Los números de la fila superior representan la posición de la base y el primer nucleótido corresponde al sitio 15.554 del genoma mitocondrial de la tortuga golfina (número de accesión de GenBank: AM258984): – mutaciones indel (los nuevos haplotipos se marcan en negrita). Haplotype

Nucleotide position 179

202

240

351 C

420

426

Lo–T2 C

A C

Lo–T4 .

. T . . –

A –

543

Freq

T 20 62.5 . 3 9.4

Lo–T6 . G . . . A . 2 6.3 Lo–T7 A

. T T G –

. 2 6.3

Lo–T1 A

. T . G –

C 2 6.3

Lo–T3 . . . . . A . 1 3.1 Lo–T5 A

. T . G –

. 1 3.1

Lo–T8 . . . T . A . 1 3.1 32 100

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Table 2. Identification of haplotypes of the mtDNA control region from olive ridley turtles identified globally obtained from unpublished data (UnP: 1, Briseño–Dueñas, 1998) and the GenBank database (2, Bowen et al., 1998; 3, Shanker et al., 2004; 4, López–Castro and Rocha–Olivares, 2005; 5, Jensen et al., 2013; 6, Bahri et al., 2015; 7, Plot et al., 2012; 8, Revuelta et al., 2015) after trimming to ~468 bp and their correspondence with the haplotypes identified in the present study. The GenBank accession numbers are indicated in brackets and * indicates that the haplotype was only reported in GenBank. Tabla 2. Haplotipos de la región de control del ADNmt de la tortuga golfina identificados a escala mundial obtenidos de datos sin publicar (UnP: 1, Briseño–Dueñas, 1998) y de la base de datos de GenBank (2, Bowen et al., 1998; 3, Shanker et al., 2004; 4, López–Castro and Rocha–Olivares, 2005; 5, Jensen et al., 2013; 6, Bahri et al., 2015; 7, Plot et al., 2012; 8, Revuelta et al., 2015) tras reducir a ~468 pb y su correspondencia con los haplotipos identificados en el presente estudio. Los números de accesión de GenBank se indican entre paréntesis y * indica que el haplotipo solo se registró en GenBank.

This study

UnP 1

Published or listed in GenBank 2

E Lo621 (FJ795433) F Lo275 Lo–Oropesa (AF051773) (FJ795418)/ (KP117262) Lo027 (FJ795409) D G Lo–02 (JN391446) H Lo–04 (JN391448) I A J J Lo–01/Lo–15 Lo–4 (AF051774) (AF314652) (JN391459) (KM357632) K K (AF051775) (AF314651) E L E (AY920522) Lo–T4 B M Lo–27 (KX812520) (KC207830) Lo–T2 K N N K (KX812518) (AF051776) (AF514311) (AY920519) Lo–T3 N O N (KX812519) (AY920521) P M U M (AY920520) Lo–T6 O V O (KX812522) (AY920523) C Lo–T1 F R (KX768696)/ Lo–T5 (KX812521)

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Table 2. (Cont.)

This study

UnP 1

Published or listed in GenBank 2

G I P K–2 (AF314654) K–3 (AF314655) K–4 (AF513545) K–5 (AF513546) Lo–03 (JN391447) Lo–05 (JN391449) Lo–06/Lo–10/Lo–13 (JN391450) Lo–07/Lo–09 (JN391451) Lo–08/Lo–14/ Lo–18 Lo–1 (JN391452) (KM357629) Lo–11/Lo–17 Lo–3 (JN391455) (KM357631) Lo–12 (JN391456) Lo–16 (JN391460) Lo–19 (JN391463) Lo–20 (JN391464) Lo–21 (JN391465) Lo–23 (KC207829) Lo–2 (KM357630) Lo–5 (KM357633) Lo–T7 (KX768697) Lo–T8 (KX768698) 7 13 15 8 5 16 5 2

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Artic Ocean North America

This study Lo–T1 to Lo–T8

Asia Bowen et al. (1998) J

Bowen et al. (1998) E and F

López Castro and Rocha Olivares (2005) E, K, N, M, and O

South America

Bowen et al. (1998) L, M, N, O and P

Bowen et al. (1998) H, I, J, and K

Shanker et al. (2004) J, K, N, K–1 to K–5

Bowen et al. (1998) F

Bahri et al. (2015) Lo–01 to Lo–05

Atlantic Ocean

Indian Ocean

N W

Pacific Ocean Bowen et al. (1998) G and J

Africa

Plot et al. (2016) Lo621, Lo275 and Lo027

Briseño Dueñas (1998) E, B, K, N, M, O C, F, G, I and P

Pacific Ocean

Europe

Revuelta et al. (2015) Lo–Oropesa

Southern Ocean

Australia

Jensen et al. (2013) Lo–01 to Lo–21 Lo–23 and Lo–27

Briseño Dueñas (1998) A and D

S Antarctica Fig. 2. World map showing the principal haplotypes of the mtDNA control region from olive ridley turtles identified globally (obtained from the GenBank database and unpublished data). Fig. 2. Mapa mundial en el que se muestran los principales haplotipos de la región de control del ADNmt de las tortugas golfinas identificadas a escala mundial (obtenido de la base de datos GenBank y de datos sin publicar).

previously (Bowen et al., 1998; López–Castro and Rocha–Olivares, 2005) (table 2) and represent 90.6 % of the samples. Three of the haplotypes were previously observed in the northern region of the Mexican Pacific. Specifically, Lo–T2 (haplotype N GenBank AF051776 by Bowen et al., 1998 and K, GenBank AY920519 by López–Castro and Rocha–Olivares, 2005) and Lo–T3 (haplotype N, GenBank AY920521 by López–Castro and Rocha–Olivares, 2005) were found in El Verde Camacho and represented 62.5 and 3.1 % of the samples from Ceuta beach, respectively. The third haplotype (Lo–T6) was previously reported for the nesting colony of BCS (Haplotype O GenBank AY920523 by López–Castro and Rocha–Olivares, 2005) and found at low frequencies (6.3 %) in this study. The two dominant haplotypes in Ceuta beach (Lo–T2 and Lo–T4) have been reported outside the region: Lo–T2 in Madras, India (N GenBank AF514311 by Shanker et al., 2004) and Lo–T4 in Flinders Beach, Australia, in the western Pacific (Lo27 GenBank KC207830 by Jensen et al., 2013) (table 2). The remaining haplotypes (Lo–T1/ Lo–T5) have been observed in other nesting regions of the eastern Pacific (Guerrero, México) (Briseño– Dueñas, 1998). The haplotype diversity was h = 0.6048 (± 0.0974), and the nucleotide diversity was π = 0.002212 (± 0.001504). A mismatch distribution analysis indi-

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cated that the historical demography of turtles nesting in Ceuta beach fit the sudden expansion model (P > 0.05) (Rogers and Harpending, 1992): τ = 0.258, θ0 = 1091 and θ1 = infinite. Of the models suggested by the Modeltest module of MEGA ver. 7 (Kumar et al., 2016), the evolutionary model that best fitted the data based on maximum likelihood phylogenetic reconstruction was T92, also known as the Tamura 3–parameter model. Under this model, the eight haplotypes identified (seven after cutting to ∼468 bp) were found to belong to the lineage of the eastern Pacific (fig. 3). The MSN indicated that all other haplotypes present in the eastern Pacific descended from the dominant haplotype Lo–T2 (fig. 4). Discussion The olive ridley turtles that nest at Ceuta beach (the northern nesting limit in the eastern Pacific) had moderate genetic diversity compared with that found in other studies (table 3) (Bowen et al., 1998; López–Castro and Rocha–Olivares, 2005; Shanker et al., 2004). This rookery was characterized by a dominance of the haplotype Lo–T2, in agreement with the dominance of this haplotype (labeled as K; table 2) in rookeries from the Mexican Pacific, particularly from El Verde

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Lo–T1/Lo–T5 (R) Lo–T7

–

---

(L)

Lo–T4 (M) 64 64

Lo–T8 – 89 88

---

East Pacific

(P)

LO–T3 (O) Lo–T6 (V) L0–T2 (N) ---

(Lo11)

---

(Lo8)

---

Lo23)

---

(H)

---

(I)

66 67

---

(G)

---

(Lo16)

---

Lo3

---

(E)

---

(F)

---

(J)

---

(Lo2)

---

(K4)

---

(K)

---

(K3)

---

(K1)

---

(K2)

---

(K5)

Indo–West Pacific

Atlantic Indo–West Pacific

East Coast of India

LK 0.005

Fig. 3. Maximum likelihood tree obtained under the Tamura 3–parameter model of nucleotide substitution describing the relationships among olive ridley turtle haplotypes from the Ceuta nesting beach and olive ridley turtle haplotypes from other ocean basins. The nomenclature of previously published haplotypes is shown in brackets. The bootstrap values (1,000 replicates) for critical nodes were derived from the maximum likelihood (above the branch) and neighbor joining analyses (below the branch), and only values above 50 % are shown. LK (L. kempii) was used as the outgroup (AF051777). Fig. 3. Árbol de probabilidad máxima obtenido mediante el modelo Tamura de tres parámetros de substitución de nucleótidos en el que se describen las relaciones entre los haplotipos de la tortuga golfina de la playa de anidación de Ceuta y los haplotipos de la tortuga golfina de otras cuencas oceánicas. La nomenclatura de los haplotipos publicados previamente se muestra entre paréntesis. Los valores de bootstrap (1.000 réplicas) de los nodos críticos se obtuvieron a partir de la probabilidad máxima (encima de la rama) y los análisis de unión de vecinos (debajo de la rama). Solo se muestran los valores superiores al 50 %. LK (L. kempii) se empleó como grupo externo (AF051777).

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Lo–T7 G Lo23

Lo3

K5 K

I Lo16

Lo–T1/Lo–T5

Lo8

Lo2

Lo–T4

Lo11 K4

Lo–T2

Lo–T3

K2 22

Lo–T8

F Lo–T6

Fig. 4. Minimum spanning network (MSN) of haplotypes of the mtDNA control region of the olive ridley turtle. Each circle represents a haplotype, the largest circles represent the dominant haplotypes in each lineage, the lines show the connected haplotypes, and the numbers represent the number of nucleotide substitutions. Lineages: east Pacific (black), Indo–west Pacific (dark gray), Atlantic (light gray), and east coast of India (white): LK (L. kempii) was used as the outgroup (AF051777). Fig. 4. Red de expansión mínima de los haplotipos de la región de control del ADNmt de la tortuga golfina. Cada círculo representa un haplotipo, los círculos más grandes representan los haplotipos dominantes en cada linaje, las líneas muestran los haplotipos conectados y los números representan la cantidad de substituciones nucleotídicas. Linajes: Pacífico oriental (negro), Pacífico indooccidental (gris oscuro), Atlántico (gris claro) y costa oriental de la India (blanco): LK (L. kempii) se empleó como grupo externo (AF051777).

Camacho nesting beach, as reported by Briseño– Dueñas (1998). The evolutionary relationships of the identified haplotypes were grouped in the east Pacific lineage (fig. 3). Starting from the dominant haplotype (Lo–T2), all other haplotypes descend from and bind to the lineage of the Indo–west Pacific by haplotype J (AF051774 by Bowen et al., 1998), with only two variable sites. In turn, haplotype J joins the lineage of the east coast of India with 16 variable sites (fig. 4). Bowen et al. (1998) proposed that the lineage of olive ridley turtles in the eastern Pacific derives from turtles from India. These turtles colonized the eastern Pacific mainland, part of which is included in this study, approximately 300,000 years ago and more recently on an evolutionary timescale, Baja California Sur (López–Castro and Rocha–Olivares, 2005). It is possible that the haplotypes identified as novel in this study (Lo–T7 and Lo–T8) have not been previously identified in the region due to the number of samples analyzed (n = 15; Briseño–Dueñas, 1998) or perhaps due to the low philopatry that characterizes the olive ridley turtle, which has unique and complex post–reproductive migrations that vary

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annually (Morreale et al., 2007; Plotkin, 2010). The exchange of nesting beaches might form part of a complex phenomenon that the olive ridley turtle uses to colonize new areas or even completely change its nesting site (Tripathy and Pandav, 2008). A recent and novel report describes a female stranded in the Mediterranean Sea in the province of Castellon, Spain (KP117262; Revuelta et al., 2015), a region where the olive ridley turtle does not usually nest, and the haplotype identified for this individual turtle matches reports of haplotype F from the Atlantic (AF051773; Bowen et al., 1998). According to the MSN topology and a mismatch distribution analysis, the rookery of Ceuta beach belongs to a population that recently (on the evolutionary scale) expanded demographically (Grant and Bowen, 1998), and this notion agrees with that proposed by López–Castro and Rocha–Olivares (2005) for Mexican rookeries of olive ridley turtle nesting. Although this study was conducted within the northernmost nesting limits of olive ridley turtles in the eastern Pacific, the findings revealed a moderate genetic diversity (h = 0.6048), very similar to that

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Table 3. Levels of genetic diversity in the mtDNA control region of the olive ridley turtle: n, sample size; H, number of haplotypes; h, haplotype diversity; π, nucleotide diversity; and bp, base pairs. Tabla 3. Grados de diversidad genética en la región de control del ADMmt de la tortuga golfina: n, tamaño de la muestra; H, número de haplotipos; h, diversidad haplotípica; π, diversidad nucleotídica; bp, pares de bases.

Site

Bibliographic source

Ceuta beach, Sinaloa, México

0.6048

0.0022

712

This study

El Verde Camacho,

0.6190

0.0023

488

Briseño–Dueñas (1998)

Sinaloa, México

López–Castro and

Rocha–Olivares (2005)

Baja California Sur, México

0.1613

0.0005

829

López–Castro and Rocha–Olivares (2005)

Ixtapilla, Michoacán, México

0.6800

0.0029

750

Rojas–Cortés et al. (2015)

Global

0.8100

0.0108

470

Bowen et al. (1998)

East coast of India

0.2700

0.0030

399

Shanker et al. (2004)

Flinders beach, Australia

0.7493

0.0033

780

Jensen et al. (2013)

reported by Briseño–Dueñas (1998) for the El Verde Camacho nesting beach (h = 0.6190) (reported for Sinaloa by López–Castro and Rocha–Olivares, 2005; table 3) and that described in similar reports from other areas of the Mexican Pacific, such as Ixtapilla, Michoacán (Rojas–Cortés et al., 2015). In contrast, the haplotype diversity of olive ridley turtles within the nesting limits in the peninsular area of Baja California Sur was low (h = 0.1613), either due to rare genetic flows with other nesting colonies or recent colonization of these beaches (López–Castro and Rocha–Olivares, 2005). Recent studies in western Pacific rookeries reported haplotypic diversities similar to those found in the present study (Jensen et al., 2013; table 3). The genetic diversity obtained in this study was greater with respect to the ancestral lineage (Indian Ocean) (h = 0.2700) that gave rise to the turtles in the eastern Pacific (Bowen et al., 1998; Shanker et al., 2004), and this increase was likely due to genetic drift events that favor certain haplotypes (Shanker et al., 2004). The haplotype diversity identified is congruent with respect to the only worldwide study of the species (table 3; Bowen et al., 1998), which included seven colonies in three ocean basins (Pacific, Atlantic and Indian) and identified a haplotype diversity of h = 0.8100 (table 3); this haplotype diversity is greater than that found in any study conducted in any specific area of the world. The nucleotide diversity was estimated to equal π = 0.0022 (table 3), which falls within the range recognized for this parameter in a typical population (0.0005 to 0.020) (Avise, 1994) and coincides with the range reported for El Verde Camacho (π = 0.0023, the official northern nesting limit in the eastern Pacific). The low nucleotide diversity could be due to the

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recent origin of some haplotypes, which might have resulted from a recent colonization by a few individuals, resulting in the elapsed time being too short to allow the accumulation of mutations that increase nucleotide diversity (Bowen et al., 1998), as in the rookery of Baja California Sur (table 3; López–Castro and Rocha–Olivares, 2005). The olive ridley turtles that nest in Ceuta beach have moderate genetic diversity with respect to other turtles studied in the eastern Pacific and worldwide. This trait is common to resilient species facing environmental changes because it can facilitate adaptations to variable conditions (Fuentes et al., 2013; Wallace et al., 2010). According to Wallace et al. (2010), although the geographical boundaries of the RMU for arribada and solitary nesters of olive ridley turtles in the eastern Pacific show some overlap, there are differences in the abundances of populations and trends between the two behaviors. Information based on arribada events (Fuentes et al., 2013; Wallace et al., 2011) indicates that the olive ridley turtle that nests in the eastern Pacific is a species with high resilience to environmental variations. Nests have been observed in regions located north of the current study area (Seminoff and Nichols, 2007). Rodríguez–Valencia et al. (2005) suggested that most of these nests must face and overcome adverse environmental conditions, such as changes in temperature, humidity or extreme storms to reach the hatching stage. Because sea turtles have adapted by redistributing their nesting sites (Hamann et al., 2007), it is quite possible that olive ridley turtles alter the range of their latitudinal distribution depending on spatiotemporal environmental variations. It is therefore necessary to continue the maintenance and protection

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of important nesting regional beaches as well as the identification and legal protection of areas that will continue to provide suitable environments for nesting areas in the future (Fuentes et al., 2012). Currently, three zones in the state of Sinaloa have been legally recognized for the protection of olive ridley turtles: El Verde Camacho (DOF, 2002), Ceuta beach (DOF, 2002), and Meseta de Cacaxtla (DOF, 2003) (fig. 1). However, some protected areas, such as Caimanero beach, Rosario, Sinaloa, where the number of nests has increased at least fourfold in the past ten years (UAS, 2015), have not been legally recognized. Additionally, other areas with sporadic nests that can potentially be used in various types of studies (e.g., with available data regarding genetic information, indicators of abundance, and records of environmental factors) have been identified (San Carlos, Guaymas, Sonora; El Desemboque, Sonora) (Seminoff and Nichols, 2007) (fig. 1). Area in which nesting occurs, even if only sporadically, should therefore be studied and continuously surveyed. Acknowledgements We thank the Promotion Program for Research Projects (PROFAPI–2013/030 and PROFAPI–2013/189) of the Autonomous University of Sinaloa (UAS), PROMEP Project Folio: UAS–PTC–109, document number of the letter of exemption: DSA/103.5/14/10808, the UAS Turtles Program, National Commission of Natural Protected Areas (CONANP) and the H. City Council of Elota, Sinaloa. This research was conducted under an agreement between CONANP and the UAS [CONANP/RNOYAGC/CONVENIOS/03/2010 (CONANP–GES–HACE–UAS Agreement)]. We also thank M. S. Rafael Bautista, B. A. Monserrat González Gómez and M. M. Elisa Martin del Campo for editing the images. References Abreu–Grobois, A., Horrocks, J., Formia, A., Dutton, P., Leroux, R., Vélez−Zuazo, X., Soares, L., Meylan, P., 2006. New mtDNA dloop primers which work for a variety of marine turtle species may increase the resolution capacity of mixed stock analysis. In: 26th Annual Symposium on Sea Turtle Biology and Conservation, Crete, Greece: 200 (M. Frick, A. Panagopoulou, A. Rees, K. Williams, Eds.). International Sea Turtle Society, Athens, Greece. Available online at: http://www.nmfs.noaa.gov/pr/pdfs/species/ turtlesymposium2006_abstracts.pdf [Accessed on 11 December 2018]. Abreu–Grobois, A., Plotkin, P., 2008. Lepidochelys olivacea. IUCN Red List of Threatened Species, http://www.iucnredlist.org/details/11534/0 [Accessed on 13 December 2016], doi:10.2305/IUCN. UK.2008.RLTS.T11534A3292503.en Avise, J. C., 1994, Molecular Markers, Natural History and Evolution. Chapman, Hall, New York. – 1995. Mitochondrial DNA polymorphism and a

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First insights into the fecal bacterial microbiota of the black–tailed prairie dog (Cynomys ludovicianus) in Janos, Mexico I. Pacheco–Torres, C. García–De la Peña, D. R. Aguillón–Gutiérrez, C. A. Meza–Herrera, F. Vaca–Paniagua, C. E. Díaz–Velásquez, L. M. Valenzuela–Núñez, V. Ávila–Rodríguez Pacheco–Torres, I., García–De la Peña, C., Aguillón–Gutiérrez, D. R., Meza–Herrera, C. A., Vaca–Paniagua, F., Díaz–Velásquez, C. E., Valenzuela–Núñez, L. M., Ávila–Rodríguez, V., 2019. First insights into the fecal bacterial microbiota of the black–tailed prairie dog (Cynomys ludovicianus) in Janos, Mexico. Animal Biodiversity and Conservation, 42.1: 127–134, Doi: https://doi.org/10.32800/abc.2019.42.0127 Abstract First insights into the fecal bacterial microbiota of the black–tailed prairie dog (Cynomys ludovicianus) in Janos, Mexico. Intestinal bacteria are an important indicator of the health of their host. Incorporating periodic assessment of the taxonomic composition of these microorganisms into management and conservation plans can be a valuable tool to detect changes that may jeopardize the survival of threatened populations. Here we describe the diversity and abundance of fecal bacteria for the black–tailed prairie dog (Cynomys ludovicianus), a threatened species, in the Janos Biosphere Reserve, Chihuahua, Mexico. We analyzed fecal samples through next generation massive sequencing and amplified the V3–V4 region of the 16S rRNA gene using Illumina technology. The results were analyzed with QIIME based on the EzBioCloud reference. We identified 12 phyla, 22 classes, 33 orders, 54 families and 263 genera. The phyla Firmicutes and Bacteroidetes were the most abundant groups and are associated with healthy intestinal communities and high efficiency in the energy diet. Most of the bacterial genera reported here for C. ludovicianus are not pathogenic and are normally found in mammalian feces. Some of the other bacteria are associated with soil, water and plants, possibly in relation to the habitat of the black– tailed prairie dog. This is the first study to report the fecal bacteria of C. ludovicianus in Mexico and it provides a baseline for determining this species' health for use in long–term conservation strategies. Key words: 16s rRNA, Bacteria, Diversity, Metagenomics, Rodent Resumen Primeros datos sobre la microbiota bacteriana fecal del perrito de las praderas de cola negra (Cynomys ludovicianus) en Janos, México. Las bacterias intestinales son un indicador importante de la salud de su hospedero y la incorporación de una evaluación periódica de la composición taxonómica de estos microorganismos en los planes de gestión y conservación puede ser una herramienta valiosa para detectar cambios que puedan poner en peligro la supervivencia de las poblaciones amenazadas. En este estudio describimos la diversidad y abundancia de las bacterias fecales de una especie amenazada, el perrito de la pradera de cola negra (Cynomys ludovicianus), en la Reserva de la Biosfera de Janos, en Chihuahua, México. Se analizaron muestras fecales mediante secuenciación masiva de siguiente generación y se amplificó la región V3–V4 del gen que codifica el ARNr 16S utilizando la tecnología Illumina. Los resultados se analizaron con QIIME a partir de la referencia EzBioCloud. Se identificaron 12 filos, 22 clases, 33 órdenes, 54 familias y 263 géneros. Los filos Firmicutes y Bacteroidetes, que fueron los grupos más abundantes, se asocian con comunidades intestinales saludables y una alta eficiencia en la dieta energética. La mayoría de los géneros bacterianos detectados en este estudio para C. ludovicianus no son patógenos y se encuentran habitualmente en las heces de mamíferos. Algunas de las otras bacterias están asociadas al suelo, el agua y las plantas, posiblemente en relación con el hábitat del perrito de las praderas de cola negra. Este es el primer estudio que reporta las bacterias fecales de C. ludovicianus en México y que proporciona un punto de referencia para determinar la salud de esta especie con vistas a utilizar esta información en estrategias de conservación a largo plazo. Palabras clave: ARNr 16s, Bacteria, Diversidad, Metagenómica, Roedor ISSN: 1578–665 X eISSN: 2014–928 X

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Received: 24 X 17; Conditional acceptance: 15 XII 17; Final acceptance: 20 IX 18 Irene Pacheco–Torres, Cristina García–De la Peña, David R. Aguillón–Gutiérrez, Luis M. Valenzuela–Núñez, Verónica Ávila–Rodríguez, Facultad de Ciencias Biológicas, Universidad Juárez del Estado de Durango, Gómez Palacio, Durango, 34000 México.– César A. Meza–Herrera, Unidad Regional Universidad de Zonas Áridas, Universidad Autónoma Chapingo, Bermejillo, Durango, 35230 México.– Felipe Vaca–Paniagua, Lab. Nacional en Salud, Diagnóstico Molecular y Efecto Ambiental en Enfermedades Crónico–Degenerativas, Facultad de Estudios Superiores Iztacala, Tlalnepantla, Estado de México, 54090 México; Instituto Nacional de Cancerología, Ciudad de México, 14080 México; Unidad de Biomedicina, Facultad de Estudios Superiores Iztacala, UNAM, Tlalnepantla, Estado de México, 54090 México.– Clara E. Díaz–Velásquez, Laboratorio Nacional en Salud: Diagnóstico Molecular y Efecto Ambiental en Enfermedades Crónico–Degenerativas, Facultad de Estudios Superiores Iztacala, Universidad Nacional Autónoma de México, Tlalnepantla, Estado de México, 54060 México. Corresponding autor: C. García–De la Peña. E–mail: cristina.g.delapena@gmail.com

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Introduction The black–tailed prairie dog (Cynomys ludovicianus) occurs in the region of Casas Grandes and Janos, Chihuahua, Mexico (CONABIO, 2011). It is a keystone species since, together with the bison (Bison bison), it controls the proliferation of shrubby plants, maintaining the structure of a grassland ecosystem (Cid et al., 1991). The 'Norma Oficial Mexicana 059' (SEMARNAT, 2010) has listed this species as threatened since 1994. In 2005, it was estimated that populations of black–tailed prairie dog had decreased by 98 %, with 97 % of its original range being lost to urban, agricultural and livestock development (CONABIO, 2011). Currently, the largest surviving colonies occur within the Janos Biosphere Reserve, considered the primary area for this species' conservation in Mexico and the United States of America (Ceballos et al., 1993). These rodents are well suited for measuring the conservation status of a habitat. Because of their keystone function, they are environmental indicators to assess short and medium term changes in their habitat. Environmental indicator species are sensitive to variations in the environment due to their dependence on certain physical and climatic characteristics that affect their physiology, survival and reproduction (Wilson and Reeder, 2005; Aragón et al., 2012; Tzab–Hernández and Macswiney–Gónzalez, 2014). Intestinal bacteria play an important role in the health of the host by participating in the synthesis of essential vitamins, renewing the intestinal epithelium and facilitating the digestion of food (Guarner and Malagelada, 2003). However, the microbiota within the host shows a generalized variation in the composition of the intestinal bacterial community depending on the modifications in the diet over time and space (Caporaso et al., 2011). Variation occurs as a response to seasonal changes in feeding patterns, which can affect the ecology of the host itself by altering its nutritional status and overall health (Lee and Mazmania, 2010; Amato et al., 2015; Aivelo et al., 2016, Amato et al., 2016). Due to the influence of intestinal bacteria on the health of their hosts (Lee and Mazmania, 2010), it is important to document the taxonomic composition of these microorganisms. This is especially important in threatened animal species as this information will inform management and conservation plans, and allow the detection of future changes in the fecal microbiota that could risk the survival of this population. Here we report on the diversity and abundance of fecal bacteria of C. ludovicianus, determined by the 16S rRNA metagenomic technique in the Janos Biosphere Reserve, Chihuahua, Mexico. This information constitutes a first approach to knowledge of the fecal microbiota of this species, serving as a baseline and contributing to future approaches to management and conservation strategies of this threatened species in Mexico. Material and methods Study area The Janos Biosphere Reserve is located in the northeastern part of the state of Chihuahua, Mexico,

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south of the border with the United States of America (fig. 1). It forms part of the Chihuahuan Desert ecoregion and limits the southwest of the area with the Sierra Madre Occidental and the east with meadows and mountains. Sample collection In October 2015, we identified a colony with the highest activity of the black–tailed prairie dog in the Janos reserve. During the first hours of the day, observations were made using binoculars at a distance of approximately 50 m. After 15 to 20 minutes of recording the activity of the prairie dogs outside their burrows, we collected ten samples of fresh fecal scat using sterile tweezers. We sprayed each sample with an antiseptic solution composed of super oxidized water (H2O2), sodium chloride (NaCl), hypochlorous acid (HClO) and sodium hypochlorite (NaClO), Microdacyn®, to remove soil bacteria that had adhered to the scat. The collected samples were deposited in BashingBead™ Zymo Research™ cell lysis tubes, and 750 μl of lysing/stabilizing solution and 25 g of fecal material were added. Each tube was processed in a cellular disruptor (TerraLyzer™) for 20 seconds. The vibration of the processor (3600 beats/minute) breaks down the bacterial cells with the help of the silica beads contained in the tube, allowing the DNA to come into contact with the stabilizing buffer. This process conserves the genetic material at room temperature, making it viable for up to two weeks according to the manufacturer’s specifications. All the sampling procedures followed the guidelines approved by the American Society of Mammalogists (Sikes et al., 2011). Laboratory work We extracted DNA from the samples using the Xpedition™ Soil/Fecal DNA MiniPrep kit in a laminar UV flow hood in sterile conditions. The extracted DNA was then combined in a pool and run on a 1.2 % agarose gel at 80V for 45 minutes in the BIO–RAD electrophoresis chamber to visualize the presence of high molecular weight DNA. The visualization was carried out in a GelMaxTM photodocumenter (UVP®). The amount of DNA obtained was measured in a Qubit® fluorometer. The amplification of the V3 and V4 regions of the 16S rRNA gene was carried out using the following primers (Klindworth et al., 2013): 5'–CCTACGGGNGGCWGCAG–3' and 5'–GACTACHVGGGTATCTAATCC–3'; which produces an amplicon of about ~460 bp. By joining these sequences to the 'overhang' adapters of the Illumina protocol (2017a), they were as follows: 5'–TCGTCGGCAGCGTCAGATGTGTATAAGAGACAGCCTACGGGNGGCWGCAG–3'and 5'–GTCTCGTGGGCTCGGAGATGTGTATAAGAGACAGGACTACHVGGGTATCTAATCC–3´ (amplicon of ~550 bp). The Illumina PCR protocol (2017a) was performed by using 12.5 μl of MyTaqTM Ready Mix 1X (Bioline®), 1 μl of each primer (10 uM), 5 μl of DNA (50 ng total) and 5.5 μl of molecular grade H2O; the following cycle was used: 95 ºC for 3 minutes; 25 cycles of 95 ºC for 30 seconds, 55 ºC for 30 seconds, 72 ºC for 30 sec-

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Fig. 1. Geographical location of the Janos Biosphere Reserve, Chihuahua, Mexico. Fig 1. Ubicación geográfica de la Reserva de la Biosfera Janos, en Chihuahua, México.

onds; 72 ºC for 5 minutes in a Labnet MultigeneTM Gradient PCR thermal cycler. One μl of the PCR products were placed on a Bioanalyzer DNA 1000 chip to verify the size of the amplicon (~550 bp). Purification of the amplicons was performed with Agentcourt® AMPure® XP 0.8 % beads. Subsequently, Nextera XT Index KitTM was used to create the library, following the Illumina protocol (2017b), using 25 μl of MyTaqTM Ready Mix 1X (Bioline®), 5 μl of each primer (N7xx and S5xx), 5 μl of DNA and 10 μl of molecular grade H2O; the following cycle was used: 95 ºC for 3 minutes; 10 cycles of 95 ºC for 30 seconds, 55 ºC for 30 seconds, 72 ºC for 30 seconds; 72 ºC for 5 minutes. Purification of the libraries was carried out with Agencourt® AMPure® XP 1.2 % beads. One μl of the final library of some randomly selected PCR products was placed on a Bioanalyzer DNA 1,000 chip to verify amplicon size of ~630 bp. Finally, quantification, normalization (equimolarity) and next generation massive sequencing (MiSeq Illumina® of 2 x 250 paired final readings) were performed following the 16S metagenomic protocol (Illumina, 2017a).

was started by assembling the forward and reverse sequences of the samples using the PEAR program (Zhang et al., 2014) with an overlap of 50 bp, a minimum reading length of 430 bp and a maximum of 470 bp, a quality criterion Q30 (one false base for every 1,000 bases) and a value of P < 0.0001. The files were then converted to FASTA format and chimeric sequences of the samples with VSEARCH were eliminated (Edgar, 2010). The operational taxonomic units (OTUs) were selected with the VCLUST method (Edgar, 2010) at 97 % similarity; a representative sequence was obtained for each OTU and the taxonomy was assigned, taking the EzBioCloud database as reference (Yoon et al., 2017). The OTUs tables were built in Biom format (Biological observation matrix; McDonald et al., 2012) and the domains were separated. We calculated Simpson and Shannon alpha diversity indices and also Faith's phylogenetic index. The relative abundance of the taxonomic levels of phylum, class, order, family, genus and species was then obtained using Krona (Ondov et al., 2011).

Bioinformatic analysis

Results

The sequence was analysed in VM Oracle VirtualBox 5.1.14 on the MGLinux platform using the Quantitative Insights into Microbial Ecology bioinformatics software (QIIME) v.1.9.0 (Caporaso et al., 2010). The process

We obtained 131,828 non–chimeric bacterial sequences. The OTUs resulted in 12 phyla, 22 classes, 33 orders, 54 families, 263 genera and 333 species. The most abundant phyla were Firmicutes (83.14 %), Bacteroide-

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Chloroflexi 0.002% Spirochaetes 0.002% Fibrobacteres 0.002%

Fig. 2. Relative abundance of bacterial taxa found in the fecal pool of Cynomys ludovicianus in Janos, Mexico. Fig. 2. Abundancia relativa de taxones bacterianos encontrados en las heces de Cynomys ludovicianus en Janos, México.

tes (9.94 %), Cyanobacteria (1.79 %), Verrucomicrobia (1.56 %), Proteobacteria (1.45 %) and Tenericutes (1.16 %) (fig. 2). The class Clostridia was the most abundant in the fecal pool of C. ludovicianus with 80.97 %, followed by the classes Bacteroidia (9.94 %), Bacilli (2.10%), Vampirovibrio_c (1.79 %), and Verrucomicrobiae (1.55 %) (fig. 2). The order Clostridiales (80.97 %) was dominant in the pool, while Bacteroidales (9.9 %), FR888536_o (1.79 %) and Bacilliales (1.6 %) followed in abundance (fig. 2). The three most abundant families were Ruminococcaceae (41.26 %), Lachnospiraceae (28.93 %) and Christensenellaceae (10.16 %) (fig. 2). At the gender level, 30 % of the OTUs in the faecal pool was taxonomically classified (fig. 2), with Ruminoccocus being the most abundant (11.08 %); 65 % were bacteria that currently only have an identification key and the remaining 5 % were unknown OTUs. Finally, 333 species were determined, of which 75 % are not known, 23.7 % have a key nomenclature and only 0.9 % have a known name (Acetatifactor muris, Bacteroides rodentium, and Streptococcus gallolyticus). The alpha diversity of the sample was 0.99 with the Simpson index, 9.65 with the Shannon index, and 179.46 with the Faith’s phylogenetic index.

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Discussion We determined that the fecal microbiota of C. ludovicianus at the phylum level is similar to that of wild rodents and other species of the order Rodentia (Lu et al., 2012; Maurice et al., 2015) with two main groups, Firmicutes and Bacteroidetes. These two groups represented about 90 % of the total readings in the sequencing of the 16S rRNA gene. This type of bacteria have been found in healthy intestinal communities, associated with a high efficiency in the energetic diet and with high probabilities that the host individual develops obesity (Duncan et al., 2008; Ley et al., 2008a; Mai and Draganov, 2009). In areas of extreme climates such as the Chihuahuan desert, precipitation and food availability vary from one year to the next (González–Romero et al., 2005). The high proportion of Firmicutes and Bacteroidetes can therefore be attributed to the need to extract and store energy from food sources occasionally limited within the area. Within the phylum Firmicutes, the order Clostridia was identified as the most abundant. This group represents a crucial factor in the modulation of physiological, metabolic and immune processes within the intestine (Lopetuso et al., 2013). Within this class,

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the families Lachnospiraceae and Ruminococcaceae were the most abundant; this was also reported in a fecal metagenomic analysis in field mice (Microtus agrestis, M. oeconomus, and Myodes glareolus) where more than 50 % of the classified sequences belonged to those two families (Koskela et al., 2017). There are precedents that indicate that the families Lachnospiraceae and Rumiococcaceae are common in the intestinal microbiota of animals that metabolize complex carbohydrates such as cellulose (Rainey, 2009). It should be noted that C. ludovicianus feeds mainly on grasses and some species of cactus (Sánchez–Cordero, 2003). These results confirm that the composition of the intestinal and fecal microbiota of mammals is similar regardless of the host species (Ley et al., 2008a; Muegge et al., 2011), showing the limited set of microorganisms that have adapted to life in the gastrointestinal tract (Ley et al., 2008b). Genera within Lachnospiracea, such as Eubacterium, Coprococcus, and Roseburia have been associated with the production of butyrate, playing a fundamental protective role, necessary for the health of intestinal epithelial tissue in mice and other hosts (Stanton and Savage, 1983). Genera that are included in the Ruminococcaceae have been found as part of the intestinal flora in cattle, sheep and goats degrading cellulose and colonizing the rumen (Rainey, 2009). Most bacterial genera reported in the present study for C. ludovicianus are not pathogenic and are common in mammalian feces. Others are associated with soil, water and plants, which could be related to the habitat of the black–tailed prairie dog (Cai and Dong, 2010; Pindi et al., 2016). However, the possible existence of pathogenic bacteria due to climatic changes in the distribution zone of this species was not ruled out. These changes can affect bacterial diversity due to the influence of ecological and environmental factors such as temperature (Muegge et al., 2011). Although only one colony of C. ludovicianus was studied in the present study, it is important to report the alpha diversity registered for this population, since comparisons with other colonies may be carried out in future research. The diversity and abundance of microorganisms in diverse vertebrate body regions can be important to understand the symbiotic associations with their hosts; nevertheless they are currently rarely used for conservation applications. In the case of C. ludovicianus, we documented the basic composition of intestinal bacteria represented in their feces, which provides a baseline of microbiological knowledge for this species. Future research should focus on determining the factors that affect this bacterial diversity and abundance, in addition to the spatiotemporal dynamics of its intestinal microbiota (Bobbie et al., 2017). This information can extend the effectiveness of conservation strategies for species at risk by incorporating aspects of health and nutrition of the population that is being protected. Fortunately, as 16S rRNA sequencing is now an economically feasible approach, microbial analyses can be integrated into conservation strategies for the benefit of vulnerable species (Stumpf et al., 2016), as is the case with the black–tailed prairie dog in Mexico.

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Acknowledgements To Antonio Esquer and José Luis Loya for the facilities to work in the Janos Biosphere Reserve, Chihuahua; to Norma Salas–Muro and Enrique Ceniceros–Esparza for their support in the collection of samples; to Miriam Zavalza–Ávila for the map design. We also thank Cameron Barrows (University of California, Riverside) for reviewing the manuscript. The present study was carried out under the Mexican federal permit SEMARNAT DGVS/08471/15. References Aivelo, T., Laakkonen, J., Jernvall, J., 2016. Population–and individual–level dynamics of the intestinal microbiota of a small primate. Applied and Environmental Microbiology, 82(12): 3537–3545. Amato, K. R., Leigh, S. R., Kent, A., Mackie, R. I., Yeoman, C. J., Stumpf, R. M., Wilson, B .A., Nelson, K. E., White, B. A., Garber, P. A., 2015. The gut microbiota appears to compensate for seasonal diet variation in the wild black howler monkey (Alouatta pigra). Microbial Ecology, 69(2): 434–443. Amato, K. R., Martinez–Mota, R., Righii, N., Raguet– Schofield, M., Corcione, F. P., Marini, E., Humphrey, G., Gogul, G., Gaffney, J., Lovelace, E., Williams, L., Luong, A., Dominguez–Bello, M. G., Stumpf R. M., White, B., Nelson, K. E., Knight, R., Leigh, S. R., 2016. Phylogenetic and ecological factors impact the gut microbiota of two Neotropical primate species. Oecologia, 180: 717–733. Aragón, P. E. E., Muñiz–Martínez, R., Garza, H. A., 2012. Roedores del Estado de Durango, México. In: Estudios sobre la biología de roedores silvestres mexicanos: 165–183 (F. A. Cervantes, C. Ballesteros–Barrera, Eds.). Universidad Autónoma Metropolitana, México. Bobbie, C. B., Mykytczuk, N. C. S., Schulte–Hostedde, A. I., 2017. Temporal variation of the microbiome is dependent on body region in a wild mammal (Tamiasciurus hudsonicus). FEMS Microbiololgy Ecology, 93(7), doi: 10.1093/femsec/fix081 Cai, S., Dong, X., 2010. Cellulosilyticum ruminicola gen. nov., sp. nov., isolated from the rumen of yak, and reclassification of Clostridium lentocellum as Cellulosilyticum lentocellum comb. nov. International Journal of Systematic and Evolutionary Microbiology, 60: 845–849. Caporaso, J. G., Kuczynski, J., Stombaugh, J., Bittinger, K., Bushman, F. D., Costello, E. K., Flerer, N., Gonzalez–Peña, A., Googrich, J. K., Gordon, I. J., Huttley, G. A., Kelley, S. T, Knights, D., Koening, J. E., Ley, R. E., Lozupone, C. A., McDonald, D., Muegge, B. D., Pirrung, M., Reeder, J., Sevinsky, J. R., Turnbaugh, P. J., Walters, W. A., Widmann, J., Yatsunenko, T., Zaneveld, J., Knight, R., 2010. QIIME allows analysis of highthroughput community sequencing data. Nature Methods, 7(5): 335–336. Caporaso, J. G., Lauber, C. L., Costello, E. K., Berg– Lyons, D., Gonzalez, A., Stombaugh, J., Knights, D., Gajer, P., Ravel, J., Fierer, N., Gordon, J. I.,

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Head size and personality in great tits Parus major A. P. Møller

Møller, A. P., 2019. Head size and personality in great tits Parus major. Animal Biodiversity and Conservation, 42.1: 135–142, Doi: https://doi.org/10.32800/abc.2019.42.0135 Abstract Head size and personality in great tits Parus major. Behavior and decision–making depend on cognitive abilities and ultimately brain size and structure. I hypothesized that personality may be related to relative brain size adjusted for body size, and therefore, that selection acts against individuals that have small brains for their body size. I investigated standard personality scores in great tits Parus major from the field in relation to head volume, sex, age, capture date, and body size. Head volume and brain mass were strongly positively correlated, allowing for non–destructive estimation of brain size based on head volume. Personality score was positively correlated with head volume andwas higher in individuals captured later in the season. In an analysis of head volume in relation to sex, age, date of capture and body size, males had larger heads than females and older individuals had larger heads than yearlings. Head volume was larger in individuals captured later during the season. These findings are consistent with the prediction that personality is related to relative brain size and that selection acts on personality and relative head size as reflected by changes over time and between age classes. Key words: Age effect, Brain size, Head size, Personality, Selection, Sex effect Resumen Tamaño cefálico y personalidad del carbonero común, Parus major. El comportamiento y la toma de decisiones dependen de las capacidades cognitivas y, en último término, del tamaño y la estructura del encéfalo. Nuestra hipótesis es que la personalidad puede estar relacionada con el tamaño relativo del encéfalo respecto del tamaño corporal y, por consiguiente, que la selección actúa en contra de los individuos que tienen encéfalos pequeños con respecto al tamaño del cuerpo. Analizamos las puntuaciones habituales de personalidad en individuos de carbonero común, Parus major, capturados en el campo, en relación con el volumen cefálico, el sexo, la edad, la fecha de captura y el tamaño corporal. Debido a la fuerte correlación positiva existente entre el volumen cefálico y la masa encefálica, se pudo emplear una técnica no destructiva para estimar el tamaño encefálico a partir del volumen cefálico. La puntuación de la personalidad estaba positivamente correlacionada con el volumen cefálico y fue superior en individuos que se capturaron más entrada la estación. En un análisis del volumen cefálico en relación con el sexo, la edad, la fecha de captura y el tamaño corporal, se observó que la cabeza era mayor en los machos y en los individuos de años anteriores que en las hembras y los jóvenes del año, respectivamente. El volumen cefálico era mayor en los individuos que se capturaron más entrada la estación. Estos resultados coinciden con la predicción de que la personalidad está relacionada con el tamaño encefálico relativo y que la selección actúa sobre la personalidad y el tamaño cefálico relativo, tal como reflejan los cambios producidos en el tiempo y entre clases de edad. Palabras clave: Efecto de la edad, Tamaño encefálico, Tamaño cefálico, Personalidad, Selección, Efecto del sexo Received: 30 VII 18; Conditional acceptance: 31 VIII 18; Final acceptance: 23 IX 18 A. P. Møller, Ecologie Systématique Evolution, Université Paris–Sud, CNRS, AgroParisTech, Université Paris– Saclay, F–91405 Orsay Cedex, France. Corresponding author: A. P. Moller. E–mail: anders.moller@u–psud.fr ISSN: 1578–665 X eISSN: 2014–928 X

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Introduction While behavioural syndromes are generally defined as correlations between multiple behavioural traits, personality is defined as individual consistency within a single trait across time and contexts (Sih et al., 2004). Correlations between multiple behavioral traits show consistent variation over time and contexts (e. g. Drent et al., 2003; Bell et al., 2009; Dunn et al., 2011). Behavioral syndromes constitute the non–random association of different behaviors that are adaptations to specific environmental conditions or constitute links caused by pleiotropy (e.g. Sih et al., 2004; Réale et al., 2007; Møller and Garamszegi, 2012; Garamszegi et al., 2012). These aggregations of associated behavioral traits reflect differences in underlying personality (Gosling, 2001; Groothuis and Carere, 2005). Aspects of behavioural personality traits, such as the axis from shyness to boldness, may correlate with risk taking and be linked to exploration of habitats and hence to exploitation of natural resources (e.g. Sih et al., 2004; Garamszegi et al., 2012). Therefore, it is not surprising that the diversity of behavioural syndromes has implications for fitness components such as survival and mating (Klein, 2000; Smith and Blumstein, 2008; Biro, 2012; Wolf et al., 2007). Although many recent studies have described behavioral syndromes in a diverse array of species, there is limited information on how such syndromes and the underlying personalities relate to cognitive abilities and the associated neural substrate (Feldker et al., 2003; Carere and Locurto, 2011; Sih and Del Giudice, 2012). Such associations between behavior and the relative size of the brain would be expected (Jerison, 1973; Dukas, 2004). Many studies have suggested that relative brain size after adjustment for the effects of body size and its component parts play an important role in mediating innovative behavior, risk of predation, foraging and hoarding, sexual selection, social behavior and even population trends of birds (e.g. DeVoogd et al., 1993; Dunbar, 1993; Madden, 2001; Garamszegi and Eens, 2016; Garamszegi et al., 2009; Nottebohm, 2005; Shultz et al., 2015; Sol et al., 2005; Gonda et al., 2012; Roth and Pravosudov, 2009). In contrast, very few intraspecific studies have attempted to link relative brain size to behavior or morphology of individuals (Riters et al., 2004; Møller, 2010; Møller et al., 2011). Gene expression studies in mice have shown that more than 80% of the genome is expressed in the brain (Lein et al., 2007; Sunkin and Hohmann, 2007). This makes it likely that many phenotypic characters will be linked to relative brain size. It has been reported that head size is a non– destructive measure of brain size that allows investigation of the link between relative brain size and behavior (Møller, 2010; Møller et al., 2011). Because flying birds are strongly selected for aerodynamic properties and weight minimization, there is a tight correlation between external head volume and internal brain mass (Møller, 2010; this study). Head size can be measured non–destructively and hence

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provide a link between behavior and cognitive ability, as reflected by relative head size (Møller, 2010). For example, barn swallows Hirundo rustica with large heads are not simply larger individuals, and that they do not generally have larger wings, larger aspect ratios, or larger wing areas (Møller, 2010). Individuals with larger heads arrived earlier from the African winter quarters to their European breeding sites; furthermore, they were more difficult to catch and likely to be recaptured after having been caught previously than were individuals with smaller heads (Møller, 2010). Intensity of brood defense by females was stronger among individuals with larger heads (Møller, 2010). Barn swallows breeding in larger colonies had larger heads, suggesting that brain size and differential recruitment of large–headed individuals to large colonies play a role in social behavior (Møller, 2010). The same authors also found evidence of directional selection for larger head size because head size increased from yearlings to older individuals due to differential mortality of small–headed individuals (Møller, 2010). In a second study, Møller et al. (2011) showed that brain size was reduced by an average of 5 % in birds in radioactive areas at Chernobyl compared to nearby control areas with little or no contamination. This effect was hypothesized to arise from inferior environmental conditions for normal brain development, including severe depletion of antioxidant levels under radiation exposure. The same study revealed a significant increase in brain size when comparing head volume of yearlings and older individuals, suggesting differential mortality of small–headed individuals. The objectives of the present study were to investigate the relationship between personality behavior and relative brain size in a model species commonly used for studies of personality and behavioral syndromes, the great tit Parus major. First, I tested whether personality score was related to head volume, while controlling for the potentially confounding effects of sex, age, body condition and body size. I used a standard procedure to quantify exploratory behavior using a novel environment and two novel object tests (Drent et al., 2003). Second, I tested whether head volume could be predicted by rearing environment, sex, age, date, body size, body mass and condition. Furthermore, I predicted that relatively large heads would be associated with specific personality behavior because a large head would allow for more diverse or more thorough processing of behaviour. Finally, I predicted that if there was selection against specific personality traits and head sizes, there should be an increase in standard personality score and head size over time and across age classes. This argument rests on the assumption that later captured individuals would be predicted to also have been present early during the season, and individuals over one–year of age would also have been present when yearlings. Again, previous studies in other species have shown that larger–headed individuals survive better as reflected by larger heads in older birds compared to yearlings (Møller, 2010; Møller et al., 2011).

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Material and methods

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I sexed great tits using the intensity of yellow breast coloration, the width of the breast band, and the presence of a large cloacal protuberance (males) or a brood patch (females) as criteria (Svensson, 2006). Birds were aged using molt of the wing coverts as a criterion (Svensson, 2006). These criteria are sufficient to reliably sex and age all adult great tits.

The size of this test chamber was smaller than that in previous studies of great tits (Drent et al., 2003), but this should not affect the relative ranking of individuals with respect to time required to perch on the individual trees. The birds were released from a holding bag. The behavioral score was the time it took for the bird to reach four of the five trees to convert this into a score ranging from 0 (slow) to fast (10) (Drent et al., 2003). Two novel object tests were conducted: these consisted of attaching an 8–cm bendable rubber toy (a Pink Panther) to the perch on one side of the cage in the first test, and a penlight battery in the second test. The behavioral score was recorded as the latency to approach the object and the closest distance to the object during the first 2 minutes. The results from these two tests were converted linearly to a scale from 0 (slow) to five (fast). I used a score that ranged from zero (the bird did not land on the perch), to one (it landed on the distant third of the perch during the two minutes), two (it landed on the distant third during the first minute), three (it landed on the central third of the perch), four (it landed on the third closest to the object), and five (the individual pecked at the novel object). Scores from the two novelty tests were analyzed separately for exploratory and novelty tests. Therefore, a higher score implies more exploratory or neophilic behavior, respectively. The use of two measurements allowed estimation of repeatability (Becker, 1984).

Head size, tarsus and beak length and body mass

Ethical note

I measured head length including beak length, head height and head width using a digital caliper to the nearest 0.01 mm, as described in detail by Møller (2010). I subsequently measured beak length to the skull (Møller, 2010). All characters were measured three times to allow estimation of repeatability. Head volume (cm3) was then estimated as ((head length – beak length) x head width x head height x 1/6 x p), assuming that head volume can be approximated by an ellipsoid. Tarsus length was recorded with a digital caliper, while body mass was recorded with a Pesola spring balance. J. Erritzøe kindly measured the three head dimensions described above while also recording brain mass for dead great tits that he received as a taxidermist. All specimens came from the same population around Christiansfeld, Denmark. Of 21 birds, 12 were males, 9 were females, 9 were adults and 12 were juveniles. There was a strong positive relationship between head volume and brain mass in all 21 great tits, implying that head volume is a reliable index of brain mass in these birds (fig. 1; F = 242.24, df = 1, 19, adjusted r2 = 0.93, P < 0.0001, estimate (SE) = 0.289 (0.019)).

All birds were captured with permission from the local ringing center. All individuals were released at the site of capture and all flew away without difficulty when released.

Study area The study was based on wild great tits caught at Orsay (48º 42' N, 2º 11' E), France, from January 23 to February 25, 2013. This study site is a mixture of urban habitats and a 300 ha chestnut Castanea sativa forest where great tits are one of the most common residents throughout the year. Capture Great tits were captured in mist nets in a suburb and an adjacent forest. They were then tested in exploratory and novel environment tests, after which they were immediately banded and released. I took care not to damage any of the individuals; none died, and all flew away on release. Sexing and aging

Personality score I used a standard procedure to quantify exploratory behavior using a novel environment and two novel object tests (Drent et al., 2003). The novel environment test consisted of placing the birds individually in a test chamber (1 m x 1 m x 1 m) with five artificial trees.

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Statistical analyses I used generalized linear models to predict personality score. I started out with a full model that included sex (a categorical variable), age (a categorical variable), tarsus length, body mass, date, and time of day as predictors. The variables were subsequently eliminated with the least important predictor eliminated first until only predictor variables with an associated P < 0.10 remained. I made generalized linear models with head volume as the response variable and sex (a categorical variable), age (a categorical variable), tarsus length, body mass, date, and time of day as predictors. In a second model, I used personality score as a response variable. These models were subsequently reduced as described above. Repeatability (Becker, 1984) was above 0.95 for all three head dimensions , implying that measurement errors were small. I evaluated the magnitude of associations between escape behavior and predictor variables based on effect sizes using Cohen's (1988) criteria for small (Pearson r = 0.10, explaining 1 % of the variance), intermediate (Pearson r = 0.30, 9 % of the variance) and large effects (Pearson r = 0.50, 25 % of the variance). All analyses were made with JMP (SAS, 2012).

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10 Personality score

Brain mass (g)

1.0

0.9

0.8

0.7 2.0

2.2 2.4 2.6 Brain volume (cm3)

2.8

Fig. 1. Brain mass (g) in relation to head volume (cm3) in 21 great tits. Fig. 1. Masa encefálica (g) en relación con el volumen cefálico (cm3) en 21 carboneros.

8 6 4 2 0 –0.3

–0.2 –0.1 0 0.1 0.2 Residual head volume

0.3

Fig. 2. Personality score in great tits in relation to residual head volume in 46 great tits (after adjustment for sex). Fig. 2. Puntuación de la personalidad en relación con el volumen cefálico residual en 46 carboneros comunes (tras el ajuste por sexo).

Results Personality score, head size and other predictors The two novel object scores were significantly repeatable (F = 3.04, df = 44, 45, P = 0.0001, R = 0.51

(SE = 0.16)). The great tits showed significant correlations between exploratory behaviour and sex, date and head volume, respectively (fig. 2, table 1). Individuals with higher exploratory behaviour scores

Table 1. Generalized linear model of the relationship between exploratory and novelty components of behavior of great tits in relation to sex, date, time and head volume. The models had the statistics x2 = 19.66, df = 4, p = 0.0006 and x2 = 23.09, df = 4, p = 0.0001. Sample size was 46 individuals. Effect size is Pearson's product–moment correlation coefficient. Tabla 1. Modelo lineal general de la relación entre el comportamiento de exploración y el comportamiento frente a la novedad del carbonero común con respecto a el sexo, la fecha, el tiempo y el volumen cefálico. Los modelos dieron como resultados x2 = 19,66, gdl = 4, p = 0,0006 y x2 = 23,09, gdl = 4, p = 0,0001. El tamaño de la muestra fue de 46 individuos. La magnitud del efecto es el coeficiente de correlación producto–momento de Pearson.

Estimate (SE)

Sex

8.64

0.0033

Date

3.89

0.049

Variable

95 % CI

Effect size

0.841 (0.273)

0.295, 1.387

0.433

0.054 (0.027)

0.000 (0.108)

0.291

Exploratory behaviour

Time

0.072

–0.669 (0.365)

–1.400, –0.061

0.265

6.69

0.0097

8.328 (2.985)

1.283, 8.810

0.381

Sex

3.45

0.063

–0.413 (0.218)

0.024, 0.851

0.274

Date

6.02

0.014

0.054 (0.021)

0.011 (0.097)

0.362

Time

9.05

0.0026

–0.924 (0.292)

–1.509, –0.339

0.444

Head volume

4.62

0.032

3.320 (1.505)

0.307, 6.333

0.317

Head volume

3.24

Novelty behaviour

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Table 2. Generalized linear model of the relationship between head volume of great tits in relation to sex, date, time, age and body mass. The model had the statistics x2 = 35.54, df = 4, p < 0.0001. Sample size was 46 individuals. Effect size is Pearson'’s product–moment correlation coefficient Tabla 2. Modelo lineal general de la relación entre el volumen cefálico del carbonero común con respecto a el sexo, la fecha, el tiempo, la edad y la masa corporal. El modelo dio como resultado x2 = 35,54, gdl = 4, p < 0,0001. El tamaño de la muestra fue de 46 individuos. La magnitud del efecto es el coeficiente de correlación producto–momento de Pearson.

Variable Sex Date Age Body mass

Estimate (SE)

95 % CI

Effect size

23.93 5.83 10.16

< 0.0001 0.016 0.0014

–0.083 (0.015) 0.004 (0.002) 0.050 (0.015)

0.629, 2.066 0.0009, 0.0081 0.020, 0.080

0.721 0.356 0.470

3.57

0.059

0.004 (0.002)

–0.00015, 0.00768

0.279

had a larger head volume with an intermediate effect size (table 1). The interaction between sex and head size was not a significant predictor of exploratory behaviour (x2 = 1.04, df = 1, P = 0.31), implying that males and females did not show different patterns. Individuals with higher novelty score had larger head volume. In addition, novelty behaviour increased during spring (a large effect size), and individuals that were captured later during the day had higher personality scores (an intermediate effect size). The interaction between sex and head size was not a significant predictor of exploratory behaviour (x2 = 3.48, df = 1, P = 0.06), implying that males and females did not show different patterns.

measure of brain size, and head volume increased with date and age. The present study was based on correlative data from great tits captured in the field. However, data from captive reared great tits from the Netherlands showed similar relationships for personality score (K. van Oers and M. Naguib, pers. comm.). Furthermore, their great tits of lab origin had a smaller head size than field captured individuals, with consequences for personality score. This shows that the findings reported here go beyond the specific great tit population under study. I will briefly discuss these findings.

Predictors of head volume

Discussion The main findings of this study were that the personality score in a model species for studying personality was related to head volume, which is an indirect

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0.3 Residual head volume

Head volume of 46 great tits showed a significant correlation with sex, with males having larger heads than females and with later captured birds having significantly larger heads than early captured individuals (table 2, fig. 3). The latter effect size was intermediate in magnitude. In addition, older individuals had larger heads than yearlings with a large effect size (table 2). Finally, there was a non–significant correlation for birds with a higher body mass having larger heads than birds with a lower mass (table 2). There were no significant correlations between head volume and tarsus length (x2 = 0.76, df = 1, P = 0.38) or time of day (x2 = 0.0004, df = 1, P = 0.98), and these relationships all had small effect sizes. Finally, the interaction between sex and date was not a significant predictor of head volume (x2 = 1.97, df = 1, P = 0.16), implying that males and females did not show different patterns of variation in head size.

0.2 0.1 0.0 –0.1 –0.2

30 40 50 Date (1 = January 1st)

Fig. 3. Residual head volume (after adjusting for the effects of sex and age) in relation to capture date (1 = January 1) in 46 great tits (after adjustment for sex). Fig. 3. Volumen cefálico residual (tras el ajuste por los efectos del sexo y la edad) en relación con la fecha de captura (1 = 1 de enero) en 46 carboneros comunes (tras el ajuste por sexo).

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140

Head volume is a reliable index of brain mass in the great tit, as shown here for a sample of 21 individuals explaining more than 93% of variance. A similar tight relationship was previously reported for the barn swallow (Møller, 2010). Personality score was positively correlated with head size with an intermediate effect size. The significant relationships between personality score and head volume in the great tit may seem surprising, although there are no previous studies of great tits or other species linking personality behavior to morphological traits. This relationship between personality score and head volume was independent of the confounding effects of sex, age, capture date and time of day. Body size was not a confounding variable, as shown by the non–significant correlation between head size and tarsus length. This was also observed in the barn swallow where there were weak phenotypic correlations between head size and other morphological characters (Møller, 2010). Independently of head size, there was a sex difference in personality score. Personality is linked to sex in humans (e.g. reviews in Biglan et al., 1990; Sansone and Sansone, 2011) and animals (Duckworth, 2006; van Oers et al., 2008; While et al., 2009). Finally, individuals captured late during the season had higher personality scores than early captured individuals. The winter 2012–2013 was one of the coldest in France for many years, with snow falling in the Orsay study site intermittently between November 2012 and April 2013. In normal winters, snow is rare or non–existent in this area. I suggest that differential mortality during this very cold winter resulted in selective mortality among individuals with low personality scores. Such differential mortality among individuals with small brains has been reported previously (Møller, 2010; Møller et al., 2011). Here I hypothesize that the different results in mortality are linked to low personality scores. Alternatively, there may have been local movements of individuals with an influx of individuals with high personality scores late during the season. I consider this second possibility to be unlikely given that the main dispersal period is in autumn (Perrins, 1979). Furthermore, none of the 46 great tits were recaptured at one of the five study sites –separated by 5 km– other than that from which they were originally captured, suggesting that movement among sites was uncommon. I found a significantly weak correlation with time of day, suggesting that birds that were captured later in the day had lower personality scores, perhaps as mediated by diel rhythm in corticosterone that mediates activity (Carere et al., 2003). Given that head volume was an important predictor of personality score, I investigated the determinants of head volume. There was a significant sex difference, and this was independent of structural body size as reflected by tarsus length. In a study of zebra finches Taeniopygia guttata reared on either a low or a high protein diet, reduced head size indirectly affected learning capacity (Bonaparte et al., 2011). In another study, Møller et al. (2011) reported a significantly reduced head size in birds at Chernobyl, linked to vitamin E, vitamin A and carotenoid deficiency during radiation exposure, extensive production of free radicals, and oxidative stress (Møller et al., 2005).

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Natural selection may have acted on head size in great tits. Two pieces of evidence are consistent with this suggestion; correlations with date and age. Great tits captured later during winter 2012–2013 had larger heads than individuals captured earlier. Because 2012–2013 was a very cold winter with repeated snowfall, individual great tits captured later were also present at the start of the capture session due to their strictly resident status. It is unlikely that this difference is due to dispersal because I did not recapture any of the great tits at a site other than that at which they were initially captured. In contrast, I captured all of the individuals that were present at the end of the study multiple times, showing that they were indeed present in the neighborhood. The correlation between age and head volume of great tits showed that older individuals had larger heads than yearlings. Again, this finding is consistent with differential mortality of individuals with small heads. Such differential mortality of individuals with small heads has previously been reported for barn swallows (Møller, 2010) and birds in general including great tits in Chernobyl (Møller et al., 2011). We can exclude the possibility that this difference between age classes is due to growth among juveniles because juveniles are fully grown by the end of the summer, well before the capture of birds for this study. The findings reported here open the possibility to study brain size in relation to personality behavior without sacrificing birds. Many different genes are expressed in the brain (Lein et al., 2017; Sunkin and Hohmann, 2007), and, therefore, it is likely that head volume as an index of brain size will be correlated with other behavioral traits. There is also evidence suggesting current selection for larger head volume, and that the quality of the rearing environment may constitute an important constraint on achieving optimal brain size. In conclusion, personality as reflected by exploratory and novelty behavior is related to head size in a passerine bird, and head size in turn is associated with sex, age and date. These findings have important implications for our understanding of the link between personality behavior, cognition and relative brain size. Acknowledgements J. Erritzøe kindly measured head size and brain mass in great tits. K. van Oers and M. Naguib kindly provided unpublished information on their captive great tits. References Becker, W. A., 1984. Manual of quantitative genetics. Academic Enterprises, Pullman, WA. Bell, A. M., Hankison, S. J., Laskowski, L. K., 2009. The repeatability of behavior: a meta–analysis. Animal Behaviour, 77: 771–783. Biglan, A., Metzler, C. W., Wirt, R., Ary, D., Noell, J., Ochs, L., French, C., Hood, D., 1990. Social and behavioural factors associated with high–risk sexual behavior among adolescents. Journal of Behavioral Medicine, 13: 245–261.

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perspective. Philosophical Transactions of the Royal Society of London B Biological Sciences, 367: 2762–2772. Smith, B. R., Blumstein, D. T., 2008. Fitness consequences of personality: A meta-analysis. Behavioral Ecology, 19: 448–455. Sol, D., Duncan, R. P., Blackburn, T. M., Cassey, P., Lefebvre, L., 2005. Big brains, enhanced cognition, and response of birds to novel environments. Proceedings of the National Academy of Science of the USA, 102: 5460–5465. Sunkin, S. M., Hohmann, J. G., 2007. Insights from spatially mapped gene expression in the mouse brain. Human Molecular Genetics, 16: R209–R219. Svensson, L., 2006. Identification guide to European passerines. British Trust for Ornithology, Thetford. van Oers, K., Drent, P. J., Dingemanse, N. J., Kempenaers, B., 2008. Personality is associated with extra–pair paternity in great tits, Parus major. Animal Behaviour, 76: 555–563. While, G. M., Sinn, D. L., Wapstra, E., 2009. Female aggression predicts mode of paternity acquisition in a social lizard. Proceedings of the Royal Society of London B Biological Sciences, 276: 2021–2029. Wolf, M., van Doorn, G. S., Leimar, O., Weissing, F. J., 2007. Life–history trade–offs favour the evolution of animal personalities. Nature, 447: 581–584. Yamaguchi, N., Kitchener, A. C., Gilissen, E., Macdonald, D. W., 2009. Brain size of the lion (Panthera leo) and the tiger (P. tigris): Implications for intrageneric phylogeny, intraspecific differences and the effects of captivity. Biological Journal of the Linnean Society, 98: 85–93.

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Incubation temperatures, sex ratio and hatching success of leatherback turtles (Dermochelys coriacea) in two protected hatcheries on the central Mexican coast of the Eastern Tropical Pacific Ocean J. García–Grajales, J. F. Meraz Hernando, J. L. Arcos García, E. Ramírez Fuentes García–Grajales, J., Meraz Hernando, J. F., Arcos García, J. L., Ramírez Fuentes, E., 2019. Incubation temperatures, sex ratio and hatching success of leatherback turtles (Dermochelys coriacea) in two protected hatcheries on the central Mexican coast of the Eastern Tropical Pacific Ocean. Animal Biodiversity and Conservation, 42.1: 143–152, Doi: https://doi.org/10.32800/abc.2019.42.0143 Abstract Incubation temperatures, sex ratio and hatching success of leatherback turtles (Dermochelys coriacea) in two protected hatcheries on the central Mexican coast of the Eastern Tropical Pacific Ocean. Incubation temperatures, sex ratio and hatching success of leatherback turtles have received little attention in conservation programs in Mexico. This study was carried out from October 2014 to May 2017 in two enclosed hatchery sites. To determine temperature parameters in the nest chamber environment and their variation during the incubation period, we placed data loggers in the centre of the egg mass in relocated nests. We then buried other data loggers in the sand near the relocated nests, inside and outside the hatchery. A total of 46 nests were examined over three nesting seasons. Mean nest temperature showed no statistical difference between nests in either the San Juan Chacahua hatchery or in the Palmarito hatchery nests. The mean sex ratio based on average temperature during the middle third of incubation duration was 96.3 % skewed to female production. Hatching success in both San Juan Chacahua and Palmarito was high. Our findings support the common pattern of a female–dominated leatherback turtle sex ratio. Furthermore, hatching success rates in the shade–cloth hatchery were higher than those in the natural nests observed in other populations. Key words: Hatcheries, Secondary beaches, Clutch size, Nests temperatures, Shading net Resumen Temperaturas de incubación, proporción de sexos y éxito de eclosión de la tortuga laúd (Dermochelys coriacea) en dos criaderos protegidos en la costa central mexicana del océano Pacífico tropical oriental. Las temperaturas de incubación, la proporción de sexos y el éxito de eclosión de la tortuga laúd han recibido poca atención en los programas de conservación en México. Este estudio se realizó entre octubre de 2014 y mayo de 2017 en dos criaderos cercados. Para determinar los parámetros de temperatura en el ambiente de la cámara de anidación y su variación durante el período de incubación, colocamos registradores de datos en el centro de cada nido trasladado. Posteriormente, se enterraron otros registradores en la arena cerca de los nidos trasladados, dentro y fuera del criadero. Se analizó un total de 46 nidos durante tres temporadas de anidación. No se observaron diferencias estadísticas en la temperatura media entre los nidos del criadero de San Juan Chacahua ni entre los de Palmarito. La proporción de sexos basada en la temperatura media durante el segundo tercio del período de incubación fue del 96,3 % en favor de la producción de hembras. El éxito de eclosión en los criaderos de San Juan Chacahua y de Palmarito fue alto. Nuestros resultados concuerdan con el patrón habitual de proporción de sexos observado para la tortuga laúd, que está dominado por la presencia de hembras. Además, los índices de éxito de eclosión en el criadero protegido con sombra artificial fueron más altos que los observados en los nidos naturales en otras poblaciones. Palabras clave: Criaderos, Playas secundarias, Tamaño de nidada, Temperaturas de nidos, Malla de sombreado Received: 24 V 18; Conditional acceptance:10 VII 18; Final acceptance: 26 IX 18 Jesús García–Grajales, Universidad del Mar, Campus Puerto Ángel km 3.5, Puerto Ángel, 70902, San Pedro Pochutla, Oaxaca, Mexico.– Juan F. Meraz Hernando, José Luis Arcos García, Eustacio Ramírez Fuentes, Instituto de Recursos/ Instituto de Industrias, Universidad del Mar, Campus Puerto Escondido km 2.5, Carretera Federal Puerto Escondido, Sola de Vega, 71980 San Pedro Mixtepec, Oaxaca, Mexico. Corresponding autor: Jesús García–Grajales. E–mail: archosaurio@yahoo.com.mx ISSN: 1578–665 X eISSN: 2014–928 X

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Introduction The leatherback turtle (Dermochelys coriacea) inhabits a wide range of coastal and pelagic waters in tropical and temperate ecosystems. They are found in both hemispheres, from the equator to sub–polar regions, although nesting activity is confined to tropical and subtropical latitudes (Benson et al., 2015). This species is globally listed as vulnerable under the International Union for the Conservation of Nature (IUCN) criteria (Wallace et al., 2013), and trends and status in the Pacific Ocean basin have declined precipitously in recent decades, with declines of more than 90 % in Mexico (Sarti et al., 2007). For more than two decades, considerable efforts and broader conservation strategies have been devoted to the protection of sea turtles in Mexico (García et al., 2003). These efforts include the protection of nesting beaches with regular patrols against human poaching, widely implemented nest translocation to protected hatcheries, and other general strategies such as a complete ban on the exploitation of turtles and their eggs (García et al., 2003). However, specific conservation efforts for leatherback turtles in Mexico have been focused on four index beaches (Mexiquillo, Tierra Colorada, Cahuitán and Barra de la Cruz) selected due to their intense nesting activity (Sarti et al., 2007; Santidrían et al., 2017). Nevertheless, secondary beaches where turtles nest regularly can also be considered important nesting sites (Santidrían et al., 2017). In all of these beaches, nests are protected by relocating freshly laid clutches to protected hatcheries —a common practice used at sea turtle rookeries around the world— to increase hatchling recruitment (Baskale and Kaska, 2005; Maulany et al., 2012; Santidrían et al., 2017). In this context and knowing that temperature plays an important role in the life–history of sea turtles (Binckley and Spotila, 2015), it is important to understand temperature regimes in enclosed, protected hatcheries and corresponding hatchling sex ratios and hatching success because few detailed studies have been conducted on these topics and on the effectiveness of hatchery management on leatherback turtle nests laid on secondary beaches in the Mexican Pacific (Vannini and Rosales, 2009; Vannini et al., 2011) The goals of this study were: 1) to compare the temperatures of leatherback turtle nests in two protected hatcheries, 2) to estimate the sex ratio of hatchlings, 3) to compare the differences in the incubation period, the number of dead embryos, and hatching success between hatcheries, and 4) to obtain information about the relationship between nest temperature and the incubation period in hatcheries on the Central Pacific Coast of Oaxaca, Mexico. Material and methods Field work The study took place in San Juan Chacahua and Palmarito beaches in the Central Pacific Coast of Oaxaca. San Juan Chacahua beach is 12 km in length, and is part of the Lagunas de Chacahua National Park, while

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Palmarito is about 16 km in length, extending from San José Manialtepec River in the northwest to Punta Colorada in the southeast. The climate is tropical, hot, and humid, and characterized by well–defined dry and rainy seasons. The mean annual temperature is 27.5 ºC and mean annual rainfall is 800 mm, concentrated between July and October; the dry season can last 8 months, from November to June. This study was carried out from October 2014 to May 2017, comprising three annual leatherback–breeding seasons (October–May). Community groups at each beach protect and relocate the nests to increase hatching success (García et al., 2003; Vannini and Rosales, 2009; Vannini et al., 2011). The enclosed hatchery sites were constructed to cover a total area of 80 m2 (10 x 8 m), which was sufficient to accommodate 100 nests and high enough not to be inundated by high tides. The distance between nests was set at 1 m in order to reduce interaction and to allow hatchery personnel to walk without stepping on the nests. Likewise, in order to protect them from the intensity of the sun, the hatcheries were covered by a sheet of shading net at a height of 1.50 m during all breeding seasons. Hatcheries were moved each year around the area to avoid accumulation of bacteria and other kinds of contamination. Community groups patrolled both beaches at night from 21:00 to 06:00 h, using an all–terrain vehicle ATV, to record any sea turtle activity. All nests recorded were collected and numbered and cloth size was recorded. Nests were transported in clean plastic bags to the enclosed hatcheries. These sites were closely monitored daily for threats from natural predators. All relocated nests were buried in the hatchery at a depth of 80 cm, the mean depth of leatherback nesting activity reported for Pacific populations. To determine temperature parameters in the nest environment and their variation during incubation period, we placed a data logger in the center of the egg mass in some relocated nests. Two types of data loggers were used between 2014 and 2017: HOBO® Pro v2 Temp/HR (Onset Comp. Corp., Bourne, MA, USA) in 6 nests on both beaches, and HOBO® UA–002–08 (Onset Comp. Corp., Bourne, MA, USA) in 40 nests on both beaches. Additionally, to study the effect of metabolic heating, one data logger was buried outside the hatchery (approximately 10 m from the hatchery) at the same depth (called 'R1', reference 1). A second group of two temperature data loggers was distributed inside the hatchery; one was buried alone in the ground at the same depth (called 'R2', reference 2), and another was placed at the environmental level under the shading net (called 'R3', reference 3). Due to logistical situations, we put the reference data loggers only in the last two seasons of the study, and we only recorded the incubation temperature data in nests of San Juan Chacahua in the first season of the study (2014–2015). All data loggers were programmed to record temperature every 30 min. Nest contents were excavated twenty–four hours after the emergence of the first hatchling, as suggested by Patiño–Martínez et al. (2010), and data loggers were retrieved. The total number of eggs (the number of eggs laid in the nest) and hatching success were calculated

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by counting unhatched eggs, dead hatchlings in eggs, and dead hatchlings in nests, and by excluding the shelled albumen globes (SAGs). The hatching success for each nest was calculated as the percentage of hatchlings in the clutch, and the incubation period per nest was determined as the numbers of days from the date of egg deposition to the date of the first hatchling emergence (Yalçın–Özdilek et al., 2007). Data analysis Daily thermal fluctuation was calculated from the difference between maximum and minimum daily temperatures for each nest. After testing for normality and homogeneity of variances, we used a parametric one–way ANOVA test to examine differences in mean daily temperature and daily thermal fluctuation in nests and between nests, years and hatcheries. Metabolic heat is measured from the difference between clutch temperature and soil temperature (without eggs), and has a daily cyclic variation depending on clutch and the differences in the soil temperature. For this work, the metabolic heat was defined as the difference between the nest temperature and the datalogger reference temperature (without eggs, outside of the hatchery) during the incubation period (Broderick et al., 2001). Levene's t–tests were used to compare variance between nest temperatures and reference–site temperatures during the middle third of incubation, which corresponds to the temperature sensitive period (TSP), and the entire incubation duration. Welch t–tests were used to compare temperatures between nest and reference sites. The mean middle–third temperature for each monitored clutch was calculated individually. For this work, we used the mean temperature during the middle third of the incubation period to estimate the sex ratio, and used and adapted the equations of sex ratio as a function of temperature calculated by Mrosovsky et al. (2002), as follows: Y = 100.06 / (1 + Exp (+ 188.78 – 6.37*X)) where Y is the sex ratio and X is the temperature. The hatching success was determined using the following formula: [(total eggs – unhatched eggs)/total eggs] x 100 Mean hatching success was calculated by hatchery and by year. As hatching success is not a continuous variable, an arcsine transformation of data was implemented. Then we tested the normality and homogeneity of variances of data, and a parametric one–way ANOVA test was used to examine differences between years and hatcheries. Results Clutch size and nest temperatures A total of 46 nests were examined for nest temperatures during three nesting seasons in the two hatcheries from the Central Pacific Coast of Oaxaca. For each season,

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the nests were placed at different times throughout the breeding season. The mean clutch size in leatherback turtle nests was 84.1 ± 10.6 eggs (range 62–104 eggs), excluding the SAGs. The incubation period was recorded between 52 and 62 days, with a mean of 57.04 days ± 2.7 SD (standard deviation). Mean nest temperature in San Juan Chacahua hatchery showed no statistical difference among nests (ANOVA F25, 584 = 147.4, p > 0.05; fig. 1), nor was there any statistical difference among Palmarito hatchery nests (ANOVA F19, 367 = 118.7, p > 0.05; fig. 2). The overall average temperature by seasons for the San Juan Chacahua hatchery was 31.01 ± 0.46 ºC in 2014–2015, 30.74 ± 0.96 ºC in 2015–2016, and 30.43 ± 0.78 ºC in 2016–2017, but this difference was not significant (ANOVA F2, 54 = 1.33, p > 0.05).The overall average temperature by seasons for the Palmarito hatchery was 30.61 ± 0.23 ºC in 2015–2016, and 30.53 ± 0.41 ºC in 2016–2017, again without significant differences (ANOVA F1, 18 = 1.83, p > 0.05). The daily thermal fluctuation varied among nests (ANOVA F25, 38= 3.72, p < 0.05) in the San Juan Chacahua hatchery, and also varied (ANOVA F19, 36 = 3.16, p < 0.05) in the Palmarito hatchery. However, or we found no differences in daily thermal fluctuations between seasons in the San Juan Chacahua hatchery (ANOVA F2, 19 = 4.41, p > 0.05) or in the Palmarito hatchery (ANOVA F1, 16 = 3.38, p > 0.05). Length of incubation period The length of the incubation period was obtained for 46 nests with known mean incubation temperature. We found a statistical relationship between the length of the incubation period and the mean nest temperature (r = –0.97; F1,47 = 22.260; p < 0.05; fig. 3A), as well as between th e length of the incubation period and the mean temperature during TSP (r = –0.86; F1, 53 = 24.75; p < 0.05; fig 3B). Metabolic heat Metabolic heat was evident in all monitored clutches and most evident during the TSP in all clutches. However, heating for all clutches was greater during the final third of incubation (F2, 45 = 7.41, P < 0.05; fig. 4), followed by a gradual decline in nest temperature toward the end of incubation, compared with the first or the middle third. Mean temperature during the entire incubation period differed between nests and reference sites (t = 0.847, p < 0.05). Intensity of metabolic heat ranged from 0.7 ºC to 3.2 ºC, with a mean of 1.88 ºC ± 0.52 ºC for the entire incubation period. Sex ratio and hatching success Table 1 shows the sex ratios of hatchlings for all nests estimated from curve equation. The mean sex ratio based on average temperature during the middle third of incubation duration (Tº) was 96.3 %, and ranged between 44 and 100 %, but the sex ratio between nests did not vary significantly (Kruskal–Wallis test, H = 17.469, P = 0.13). Only one of the 46 nests (nest

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2014–2015 34 season

29.1 ºC (pivotal temperature)

1 2 3 4 5 6 Nest number

2015–2016 season

29 28

29.1 ºC (pivotal temperature)

35.6 º

19.8º 20.8º

7 8 9 10 11 12 13 14 15 16 R1 R2 R3 Nest number

29 28

36.2 º

2016–2017 season

29.1 ºC (pivotal temperature)

21.8 º 19.4 º

17 18 19 20 21 22 23 24 25 26 R1 R2 R3 Nest number

Fig. 1. Mean nest temperature in San Juan Chacahua hatchery during three seasons. Dark lines inside each box represent median temperature. Whiskers represent the maximum and minimum values recorded. The horizontal dashed lines correspond to the proposed nest pivotal temperature for leatherback turtle: R1, reference sensor 1; R2, reference sensor 2; R3, reference sensor 3. Fig. 1. Temperatura media de los nidos en el criadero de San Juan durante tres temporadas. Las líneas oscuras dentro de los recuadros representan la temperatura mediana. Los bigotes representan los valores máximos y mínimos registrados. Las líneas horizontales discontinuas corresponden a la temperatura umbral de nido propuesta para la tortuga laúd; R1, sensor de referencia 1; R2, sensor de referencia 2; R3, sensor de referencia 3.

37.8 º

2015–2016 season

29 28

36.4 º

2016–2017 season

Temperature (ºC)

29.1 ºC (pivotal temperature)

29 20.9º 21.3º

27 28 29 30 31 32 33 34 35 36 R1 R2 R3 Nest number

29.1 ºC (pivotal temperature)

20.5º 20.7º 16.9º

37 38 39 40 41 42 43 44 45 46 R1 R2 R3 Nest number

Fig. 2. Mean nest temperature in Palmarito hatchery during two seasons. Dark lines inside each box represent median temperature. Whiskers represent the maximum and minimum values recorded. The horizontal dashed lines correspond to the proposed nest pivotal temperature for Leatherback turtle: R1, reference sensor 1; R2, reference sensor 2; R3, reference sensor 3. Fig. 2. Temperatura media de los nidos en el criadero de Palmarito durante dos temporadas. Las líneas oscuras dentro de los recuadros representan la temperatura mediana. Los bigotes representan los valores máximos y mínimos registrados. Las líneas horizontales discontinuas corresponden a la temperatura umbral de nido propuesta para la tortuga laúd: R1, sensor de referencia 1; R2, sensor de referencia 2; R3, sensor de referencia 3.

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Length of the incubation period (days)

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62 61 60 59 58 57 56 55 54 53 52 51

28.5 29 29.5 30 30.5 31 31.5 32 32.5 Temperature (ºC)

Fig. 3. Relationship between the length of the incubation period and a) mean nest temperature, and b) mean TSP; obtained for 46 nests in both hatcheries. Filled circles represent San Juan Chacahua hatchery and open triangles represent Palmarito Hatchery. Fig. 3. Relación entre la duración del período de incubación y a) la temperatura media de los nidos y b) el período sensible a la temperatura, obtenida para 46 nidos en ambos criaderos. Los círculos negros representan el criadero de San Juan y los triángulos blancos, el de Palmarito.

3.5

Temperature (ºC)

3.0 2.5 2.0 1.5 1.0 0.5 65

75 85 95 105 115 Number of eggs per nest

125

135

Fig. 4. Metabolic heat during the final third of the incubation period, and the relationship with the number of eggs per nest. Filled circles represent San Juan Chacahua hatchery and open circles represent Palmarito Hatchery. Fig. 4. Calor metabólico durante el último tercio del período de incubación y la relación con el número de huevos por nido. Los círculos negros representan al criadero de San Juan y los blancos, al de Palmarito.

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Table 1. Sex ratios of hatchlings incubated in two protected hatcheries on the central coast of Oaxaca, Mexico: Nn, nest number; Nf, number of fertile eggs; N, number of SAG's; T, total number of eggs; Id, incubation duration (days); WIP, whole IP; Ft, first third; St, second third (PTS); Lt, last third; Mh, metabolic heating (in ºC); Hs, hatching success number (%). Tabla 1. Proporción de sexos en las crías incubadas en dos criaderos protegidos en la costa central de Oaxaca, en México: Nn, número de nidos; Nf, número de huevos fértiles; N, número SAG; T, número total de huevos; Id, Duración de la incubación (en días); WIP, IP completo; Ft, primer tercio; St, segundo tercio (PTS); Lt, último tercio; Mh, calor metabólico (en ºC); Hs, número de eclosiones con éxito (%).

Incubation temperature (ºC)

Nest

Sand

Nn Lay date Nf N T Id WIP Ft St Lt Ft St Lt Mh Hs San Juan Chacahua 2014–2015 1

13/10/2014

86 24 110 62

30.3 29.7 30.4 30.9 29.1 29.7 30.2 0.7 35 (40.7)

28/10/2014

78 18 96

30.7 29.8 31.1 32.1 29.5 29.9 30.3 1.8 28 (35.9)

05/11/2014

92 16 108 57

30.9 29.8 30.3 31.7 29.6 30.1 30.4 1.3

07/11/2014

81 12 93

31.5 30.1 31.1 32.3 29.6 30.1 30.6 1.7 36 (44.4)

18/11/2014

87 15 102 55

31.3 30.1 31.3 32.2 29.5 29.9 30.1 2.1 46 (52.9)

05/12/2014

69 28 97

31.4 30.3 31.1 31.7 29.8 30.2 30.5 1.2 32 (46.4)

29.5 28.9 29.8 30.9 28.2 28.5 28.8 2.1 57 (83.8)

34 (37)

2015–2016 7

08/11/2015

68 31 99

12/11/2015

94 14 108 61

23/11/2015

79 26 105 59

09/12/2015

85 21 106 58

13/12/2015

79 25 104 58

29.9 29.5 30.3 31.6 28.3 28.6 29 2.6 57 (72.2)

24/12/2015

93 18 111 56

31.5 30.1 30.9 32.1 28.5 28.5 29.3 2.8 71 (76.3)

27/12/2015 103 15 118 56

31.6 30.1 31.1 32

03/01/2016

65 37 102 55

31.7 30.5 31.2 32.4 28.5 28.8 29.4 3

05/01/2016

89 21 110 55

31.6 30.7 31.2 32.4 28.6 28.8 29.3 3.1 64 (71.9)

13/02/2016

78 23 101 56

31.8 30.8 31.4 32.6 28.7 28.9 29.4 3.2 59 (75.6)

97 15 112 62

29.7 28.7 30.2 31.3 28.3 28.5 29 2.3 71 (73.2)

29.7

29 30.3 31.3 28.3 28.5 29 2.3 75 (79.8)

29 29.6 30.9 28.2 28.4 28.7 2.2 63 (79.7)

30.1 29.6 30.4 31.1 28.4 28.7 29.1 2

68 (80)

28.7 28.6 29.1 2.9 76 (73.8) 52 (80)

2016–2017 17

28/10/2016

18 02/11/2016 84 12 96 61 29.6 28.6 29.9 30.8 28.1 28.4 28.9 1.9 69 (82.1) 19

05/11/2016

72 18 90

16/11/2016

96 17 113 60

27/11/2016 101 18 119 59

07/12/2016

69 31 100 58

29.1 30.1 31.3 28.6 28.8 29.1 2.2 53 (76.8)

09/12/2016

73 19 92

29.2 30.3 31.2 28.5 28.9 29.2 2

05/01/2017

87 18 105 56

31.1 29.1 30.2 31.1 28.6 28.8 29.3 1.8 62 (71.3)

10/01/2017

95 13 108 56

31.5 29.4 30.3 31.2 28.7 28.6 29.2 2

74 (77.9)

08/02/2017

62 7 69

31.6 29.7 30.5 31.6 28.4 28.7 29.4 2.2

49 (79)

29.9 28.8 30.1 31.1 28.3 28.5 28.9 2.2 58 (80.6) 29.8 28.7 30 31.1 28.5 28.7 29.1 2 30.1 28.9 30.2 31

68 (70.8)

28.4 28.7 29.2 1.8 82 (81.2) 58 (79.5)

Palmarito 2015–2016

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03/11/2015

82 23 105 59

30.2 29.9 30.4 31.1 28.8 29.1 29.5 1.6 68 (82.9)

09/11/2015

65 32 97

30.5 29.7 30.2 31.1 28.9 29 29.4 1.7 51 (78.5)

24/11/2015

86 13 99

30.4 29.9 30.6 31.2 28.9 29.2 29.4 1.8 67 (77.9)

07/12/2015

76 21 97

30.5 30.1 30.4 31.2 29.1 29.4 29.7 1.5 62 (81.6)

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Tabla 1. (Cont.)

Incubation temperature (ºC)

Nest

Sand

Nn Lay date Nf N T Id WIP Ft St Lt Ft St Lt Mh Hs 31

15/12/2015

84 19 103 52

08/01/2016

79 21 100 56

30.6 30.5 31.1 31.8 29.5 29.7 30.1 1.7 58 (73.4)

13/01/2016

94 17 111 56

30.7 30.2 30.9 31.7 29.4 29.6 30.2 1.5 73 (77.7)

26/01/2016

75 22 97

30.7 30.6 31.2 31.7 29.5 29.8 30 1.7

06/02/2016

86 17 103 55

30.6 30.4 31.1 31.6 29.7 29.9 30.1 1.5 73 (84.9)

12/02/2016

92 31 123 54

30.9 30.5 31.3 31.8 29.4 29.7 30.1 1.7 76 (82.6) 29.9 29.7 30.1 30.9 28.8 29.1 29.7 1.2 72 (85.7)

30.2 30.7 31.5 29.3 29.6 30 1.5

68 (81)

63 (84)

2016–2017 37

13/10/2016

84 13 97

17/10/2016

97 22 119 57

10/11/2016

73 11 84

18/11/2016

88 15 103 60

30.5 29.8 30.4 31.1 28.7 29.1 29.5 1.6 71 (80.7)

13/12/2016

96 19 115 58

30.4 29.7 30.3 31

28/12/2016

78 23 101 59

30.3 29.7 30.2 31.1 28.9 29.3 29.7 1.4 65 (83.3)

05/01/2017 104 9 113 53

30.7 29.8 30.2 31.2 28.9 29.2 29.8 1.4 74 (71.2)

13/02/2017

83 18 101 52

28/01/2017

96 22 118 54

30.8 30.2 30.6 31.5 29.1 29.4 29.9 1.6 77 (80.2)

08/02/2017

89 16 105 52

31.2 30.2 30.9 31.6 29.1 29.3 29.8 1.8

number seven, see table 1) was predicted to produce more males; t the majority of the nests were thus predicted to produce more females. Hatching success varied between years in both hatcheries. In the San Juan Chacahua hatchery it was 42.8 % in 2014–2015, 77.1 % in 2015–2016, and 77 % in 2016–2017 (ANOVA F = 105.84, p < 0.0001), while in the Palmarito hatchery it was 80.5 % in 2015– 2016, and 81.5 % in 2016–2017 (ANOVA F = 0.568, p < 0.05). Hatching success also varied between hatcheries (ANOVA F = 12.771, p < 0.0). Discussion In this study, the mean clutch size in leatherback turtle nests was 84.1 ± 10.6 eggs in both hatcheries. This is higher than the clutch size (62 ± 17.9) reported in the index of beaches of the Mexican Pacific (Sarti et al., 2007). Most other leatherback nesting populations have a smaller clutch size (Eckert et al., 2015; Sotherland et al., 2015). Some researchers have shown that clutch size increases with body size in other marine turtles (Wallace et al., 2007). However, we did not collect female size data.Santidrían and Swiggs (2015) mention inconsistencies in the reports of clutch size and in the terms used to describe hatching of eggs and emergence of hatchlings, mainly because some consider the total number of eggs including SAGs, which

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29.6 30.1 30.8 28.9 29.1 29.5 1.3 81 (83.5)

30.4 29.8 30.2 31.1 28.7 29 29.4 1.7 59 (80.8)

30.1 30.6 31.4

28.9 29.2 29.6 1.4 83 (86.5)

29 29.2 29.7 1.7 69 (83.1) 73 (82)

are not real eggs because they lack yolk (Sotherland et al., 2015). Temperature may be the single most important variable affecting egg development and hatchling output in leatherback turtles, influencing the developmental rate, hatching success, emergence rate, proportion of female hatchlings, and fitness of hatchlings (Santidrián and Swiggs, 2015). Our results provide evidence of daily thermal fluctuation within the egg chamber of Dermochelys coriacea nests in protected hatcheries. Likewise, it should be taken into account that the mean temperature of nests remained relatively homogeneous, probably as a result of the hatchery shading. This shading strategy has been proposed as a focused technique to mitigate the effects of temperature (Van de Merwe et al., 2006; Hill et al., 2015) and can be performed to facilitate survival from the nest and to increase reproductive output, principally because hatcheries with shade cloth decreased sand temperatures to the upper limit of the optimal incubation temperature range (Hamann et al., 2010). We found that the mean incubation period of all hatchery nests (57.04 days) was shorter than that of natural nests (59.9 days) in the East Pacific (Santidrian and Swiggs, 2015). Similarly, for temperature–recorded nests, the mean nest temperature of hatchery nests (30.6 ºC) was similar to that of natural nests (30.6 ºC) in the East Pacific (Santidrian and Swiggs, 2015).

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To the best of our knowledge, this is the first study to investigate and compare metabolic heating in hatcheries in Mexico, including the comparison of nest centre and adjoining sand temperatures (reference site). In the past, quantifying metabolic heating within nests was difficult due to the expense and unreliability of temperature loggers (Limpus et al., 1983; Broderick et al., 2001). Currently, however, the miniaturization and the capacity of the loggers make accurate measurements of nest temperatures feasible (Broderick et al., 2001). Our results show that the general pattern of metabolic heating was similar to that found in previous studies for other species (Morreale et al., 1982; Broderick et al., 2001; Tapilatu and Ballamu 2015; Candan and Kolankaya, 2016; Özdilek et al., 2016), with metabolic heating recorded mainly during the second half of incubation with a peak, followed by a gradual decline in nest temperature toward the end of incubation. Godfrey et al. (1997) recorded the temperature in clutches of the leatherback turtle (Matapica beach, Suriname) in addition to recording the sand temperature to the side of the clutch. Nest temperatures were found to vary, on average, from control temperatures by 0.82 °C during the TSP, suggesting that metabolic heating may play some role in influencing hatchling sex ratios. In addition, it has been suggested that metabolic heating can only be important if it elevates the nest temperature by > 1 ºC during the middle third of incubation (Yntema and Mrosovsky, 1980). Özdilek et al. (2016) mention that metabolic heating during incubation periods should not be ignored as a cause for the increasing nest temperatures found during incubation. This study showed that mean temperature in the nest chamber increases 3.2 ºC with respect to the reference sites, and heating was observed in all monitored nests. Yntema and Mrosovsky (1980) stated that a change of 1–2 ºC can make a considerable difference in the sex ratios of hatchlings, and some studies have documented sufficient metabolic heating in nests during TSP to significantly alter the hatchling sex ratios (Broderick et al., 2001; Kaska et al., 2006; Jribi et al., 2013; Tapilatu and Ballamu, 2015; Özdilek et al., 2016). This is because sexual differentiation in sea turtles is strongly influenced by ambient incubation temperature (Standora and Spotila, 1985; Mrosovsky, 1994); more specifically, the embryo is exposed to a continuous temperature during the middle trimester of incubation, which determines the eventual gonadal differentiation and sex of the hatchling (Wibbels, 2003). Nevertheless, incubation studies in leatherback turtles indicate that the most temperatures produce either all males or females, given the narrow transitional range of temperature (Mrosovsky and Pieau, 1991; Godfrey et al., 1997; Binckley and Spotila, 2015), and there is only a narrow time range when nest temperature determines hatchling sex (Mrosovsky and Pieau, 1991). Although in this work we provide evidence of metabolic heat, this was most pronounced during the last third of development with variations, and possibly sex was already determined (Mickelson and Downie, 2010). The reported sex ratio for leatherback turtles is generally female–dominated (Binckley and Spotila,

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2015), as we report in this study. Possibly, increases in global temperature will affect leatherback hatchling sex ratios (Binckley et al., 1998; Patiño–Martínez et al., 2012), and survival of hatchlings (Saba et al., 2012; Spotila et al., 2015). Several studies have used controlled incubation temperature to measure and explain temperature effects on sex determination in leatherback turtle eggs (Rimblot et al., 1985; Chan and Liew, 1996; Binckley et al., 1998; Chevalier et al., 1999). However, the pivotal temperature may vary with species and among populations in natural nests (Binckley and Spotila, 2015). Nevertheless, the hatchery management strategy implemented in Mexico (García et al., 2003) focuses only on increasing hatching success (Sönmez et al., 2013; Sari and Kaska, 2017) and does not contemplate the effects of temperature on the nests. With regards to the hatching success of temperature– recorded nests, we found the mean hatching success (74.4 %) to be higher than that of natural nests (47 %, Playa Grande Costa Rica; Santidrian and Swiggs, 2015), which is consistent with our overall finding on hatching success showing that hatchery management increases hatching success. However, the use of a hatchery site may not always guarantee the hatching success (Pazira et al., 2016; Vannini et al., 2011). For example, in a study on leatherback turtle eggs in the Southwestern Caribbean Sea, Patino–Martínez et al. (2012) found that the hatching rate was higher in natural nests (79.9 %) than in those transferred to the beach hatchery (67.7 %). Hence, although relocation is suggested to be a common strategy for conservation of declining sea turtle populations (Baskale and Kaska, 2005; Pfaller et al., 2008), there is no consensus among researchers about whether relocation is an effective conservation tool for sea turtles (Sari and Kaska, 2017). More details about the advantages and disadvantages of hatchery management can be found in Sari and Kaska (2017). This study attempts to elucidate the impact of nest relocation and the effectiveness of community conservation in Mexico. It also provides a basis for further studies related to reproductive ecology of the leatherback sea turtle, an endangered species. Conclusions San Juan Chacahua and Palmarito beaches are considered secondary nesting beaches for leatherback turtles on the Mexican Pacific Coast (Sarti et al., 2007; Santidrian et al., 2017). The nest temperatures recorded inside the chamber nests were near the upper tolerance limits for incubation of leatherback turtle eggs and hence the results of the present study were in agreement with the common pattern of leatherback turtle hatchling sex ratios from beaches in the Eastern Pacific, which is female–dominated. Hatching success rates in the shade–cloth hatchery were higher than the natural nest rates observed in other populations. Finally, the conservation aspects related to sex ratio and hatching success, as well as research on metabolic heating need to be continuously monitored over several years.

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Acknowledgements We thank the Universidad del Mar (UMAR) and its División de Estudios de Posgrado for logistics and facilities provided. Funding was partially provided by Promotora del Desarrollo S. C. and Fondo Oaxaqueño para la Conservación de la Naturaleza (FOCN). Research permits were obtained from Secretaría de Medio Ambiente y Recursos Naturales (SEMARNAT: DGVS/03409/16, and DGVS/05873/17). We are especially grateful to San Juan Chacahua and ViveMar communitarian turtle campsites. Special thanks also to Julia Saterlee for revision and suggestions to our English manuscript. This paper is part of the PhD dissertation of Jesús García–Grajales as a student of the División de Estudios de Posgrado at the UMAR. References Baskale, E., Kaska, Y., 2005. Sea turtle nest conservation techniques on southwestern beaches in Turkey. Israel Journal of Zoology, 51, 13–26. Benson, S. R., Tapilatu, R., Pilcher, N., Santidrian T. P., Sarti, L., 2015. Leatherback turtle populations in the Pacific Ocean. In: The Leatherback turtle. Biology and Conservation: 110–122 (J. R. Spotila, P. Santidrian, Eds.). John Hopkins University Press, U.S. Binckley, C. A., Spotila, J. R., 2015. Sex determination and hatchlings sex ratios of the Leatherback turtle. In: The Leatherback turtle. Biology and Conservation: 84–93 (J. R. Spotila, P. Santidrian, Eds.). John Hopkins University Press, U.S. Binckley, C. A., Spotila, J. R., Wilson, K. S., Paladino, F. V., 1998. Sex determination and sex ratios of Pacific leatherback turtles, Dermochelys coriacea. Copeia 2: 291–300. Broderick, A. C., Godley, B. J., Hays, G. C., 2001. Metabolic heating and the prediction of sex ratios for green turtles (Chelonia mydas). Physiological and Biochemical Zoology, 74: 161–170. Candan, O., Kolankaya, D., 2016. Sex ratio of green turtle (Chelonia mydas) hatchlings at Sugözü, Turkey: Higher accuracy with pivotal incubation duration. Chelonian Conservation and Biology, 15: 102–108. Chan, E., Liew, H., 1996. Decline of the Leatherback population in Terengganu, Malaysia, 1956–1995. Chelonian Conservation and Biology, 2: 196–203. Chevalier, J., Godfrey, M. H., Girondot, M., 1999. Significant difference of temperature–dependent sex determination between French Guiana (Atlantic) and Playa Grande (Costa Rica, Pacific) leatherbacks (Dermochelys coriacea). Ann. Sci. Nat. Zool. Biol. Anim., 20: 147–152. Eckert, K. L., Wallace, B. P., Spotila, J. R., Bell, B. A., 2015. Nesting ecology and reproductive investment of the Leartherback turtle. In: The Leatherback turtle. Biology and Conservation: 63–73 (J. R. Spotila, P. Santidrian, Eds.). John Hopkins University Press, U.S. García, A., Ceballos, G., Adaya, R., 2003. Intensive beach management as an improved sea turtle conservation strategy in Mexico. Biological Conservation, 111: 253–261.

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40.2013.795061 Sotherland, P., Wallace, B., Spotila, J., 2015. Leatherback turtle eggs and nests, and their effect on embryonic development. In: The Leatherback turtle. Biology and Conservation: 135–148 (J. R. Spotila, P. Santidrian, Eds.). John Hopkins University Press, U.S. Spotila, J. R., Saba, V., Patel, S. H., Santidrian, P., 2015. Warming climate. A new threat to the Leatherback turtle. In: The Leatherback turtle. Biology and Conservation: 185–195 (J. R. Spotila, P. Santidrian, Eds.). John Hopkins University Press, U.S. Standora, E. A., Spotila, J. R., 1985. Temperature dependent sex determination in sea turtles. Copeia, 3: 711–722. Tapilatu, R., Ballamu, F., 2015. Nest temperatures of the Piai and Sayang Islands green turtle (Chelonia mydas) rookeries, Raja Ampat Papua, Indonesia: Implications for hatchlings sex ratios. Biodiversitas, 16: 102–107. Vannini, F., P. Rosales. 2009. Leatherback nesting in Tomatal, Oaxaca, Mexico in 2007/2008. Marine Turtle Newsletter, 126: 13–14. Vannini, F., Reyes Sánchez, A., Escamilla, G., Santos, C., Cruz, E., Franco, P., Pérez, E., 2011. Sea turtle protection by communities in the Coast of Oaxaca, Mexico. Cuadernos de Investigación UNED, 3(2): 187–194. van de Merwe, J., Ibrahim, K., Whitier, J., 2006. Effects of nest depth, shading, and metabolic heating on nest temperatures in sea turtles hatcheries. Chelonian Conservation and Biology, 5: 210–215. Wallace, B., Sotherland, P., Santidrian, P., Reina, R., Spotila, J., Paladino, F., 2007. Maternal investment in reproduction and its consequences in Leatherback turtles. Oecologia, 152: 37–47, doi: 10.1007/ s00442–006–0641–7 Wallace, B. P., Tiwari, M., Girondot, M., 2013. Dermochelys coriacea. In: IUCN 2013. IUCN Red List of Threatened Species. Version 2013.2. Available on: www.iucn–redlist.org [Accessed on 27 de February 2014]. Wibbels, T., 2003. Critical approaches to sex determination in sea turtle biology and conservation. In: Biology of sea turtles: 103–134 (P. Lutz, J. Musik, J. Wynekan, Eds.). CRC Press, U.S. Yalçın–Özdilek, Ş., Özdilek, H. G., Ozaner, F. S., 2007. Possible influence of beach sand characteristics on green turtle nesting activity on Samandağ Beach, Turkey. Journal of Coastal Research, 23: 1379–1390. Yntema, C. L., Mrosovsky, N., 1980. Sexual differentiation in hatchling loggerheads (Caretta caretta) incubated at different controlled temperatures. Herpetologica, 36: 33–36.

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Temporal overlap in the activity of Lynx rufus and Canis latrans and their potential prey in the Pico de Orizaba National Park, Mexico R. Serna–Lagunes, L. R. Álvarez–Oseguera, D. M. Ávila–Nájera, O. R. Leyva–Ovalle, P. Andrés–Meza, B. Tigar Serna–Lagunes, R., Álvarez–Oseguera, L. R., Ávila–Nájera, D. M., Leyva–Ovalle, O. R., Andrés–Meza, P., Tigar, B., 2019. Temporal overlap in the activity of Lynx rufus and Canis latrans and their potential prey in the Pico de Orizaba National Park, Mexico. Animal Biodiversity and Conservation, 42.1: 153–161, Doi: https://doi.org/10.32800/abc.2019.42.0153 Abstract Temporal overlap in the activity of Lynx rufus and Canis latrans and their potential prey in the Pico de Orizaba National Park, Mexico. Species of the same trophic guild are thought to coexist through their differential use of resources, including food, space and time. Time understood as the pattern of activity is highly dynamic. Fourteen camera–traps were set up in the Pico de Orizaba National Park and active for 12 months. Frequency histograms were used to analyze their activity patterns (AP) and a coefficient of overlap (Δ) was used to determine the temporal overlap between two predators, Lynx rufus and Canis latrans, and the predators and their potential prey. A sampling effort of 5,110 traps/night obtained 217 independent records of L. rufus (45), C. latrans (27) and eight potential prey species (145). The predators were cathemeral and four potential prey mainly lagomorphs and ^ = 0.80, and the highest overlap between rodents were nocturnal. The temporal overlap between the predators Δ ^ = 0.80), and L. rufus and lagomorphs (Δ ^ = 0.58), with predators and prey were for C. latrans and rodents (Δ differences between the degree of overlap in dry and rainy seasons. The cathemeral habits of the predators likely increase their likelihood of hunting success, particularly for prey with variable activity patterns. The APs support information on dietary breadth and the differential use of resources and temporal differences as strategies for coexisting predators, continually adapting to a highly dynamic and changing environment. Key words: Coefficient of overlap, Co–predators, Coyote, Bobcat, Lagomorphs, Rodents Resumen Superposición temporal de la actividad de Lynx rufus y Canis latrans y sus presas potenciales en el Parque Nacional Pico de Orizaba, en México. Se cree que el uso diferencial de los recursos, en especial del espacio, la comida y el tiempo, permite la coexistencia de especies del mismo gremio trófico. El tiempo entendido como el patrón de actividad es altamente dinámico. En el Parque Nacional Pico de Orizaba se instalaron 14 cámarastrampa que estuvieron activas durante 12 meses. Se analizaron los patrones de actividad (PA) de las especies mediante histogramas de frecuencia y se calculó el índice de solapamiento (Δ) para determinar la superposición temporal entre dos depredadores, Lynx rufus y Canis latrans y entre los depredadores y sus presas potenciales. Con un esfuerzo de muestreo de 5.110 noches/trampa se obtuvieron 217 registros independientes de L. rufus (45), C. latrans (27) y de ocho especies de presas potenciales (145). Los depredadores fueron catamerales y cuatro presas, nocturnas, principalmente lagomorfos y roedores. La superposición temporal entre ambos depredadores ^ = 0,80 y entre estos y sus presas, los valores más altos se encontraron entre C. latrans y los roedores fue ∆ ^ = 0,58), con variaciones entre la estación seca y la de lluvias. Al ^ = 0,80) y entre L. rufus y los lagomorfos (∆ (∆ ser de hábitos catamerales, los depredadores tienen más posibilidades de cazar más presas, en especial las que tienen patrones de actividad variables. Los PA validan la información sobre la variedad de la alimentación y la utilización diferencial de los recursos y las diferencias temporales como estrategias de coexistencia de los depredadores, que se adaptan constantemente a un entorno muy dinámico y cambiante. Palabras clave: Índice de solapamiento, Codepredadores, Coyote, Gato montés, Lagomorfos, Roedores

ISSN: 1578–665 X eISSN: 2014–928 X

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Received: 28 VI 18; Conditional acceptance: 07 IX 18; Final acceptance: 01 X 18 Ricardo Serna–Lagunes, Otto R. Leyva–Ovalle, Pablo Andrés–Meza, Unidad de Manejo y Conservación de Recursos Genéticos, Facultad de Ciencias Biológicas y Agropecuarias, Región Orizaba–Córdoba, Universidad Veracruzana, c/ Josefa Ortiz de Domínguez s/n., Col. Centro, Peñuela, Municipio de Amatlán de los Reyes, Veracruz, 94945 México.– Luis Raúl Álvarez–Oseguera, Administración del Parque Nacional Pico de Orizaba, Comisión Nacional de Áreas Naturales Protegidas (CONANP), c/ Poniente 24, 80, Sta. María Tlachichilco, Orizaba, Veracruz, 94350 México.– Dulce María Ávila–Nájera, Universidad Politécnica de Huatusco, Unidad Académica de Biotecnología y Agroindustrial, Av. 1, 728, Col. Centro, Huatusco, 94100 Veracruz, México.– Barbara Tigar, School of Forensic and Applied Sciences, University of Central Lancashire, Preston, UK. Corresponding author: Dulce María Ávila–Nájera. E–mail: dul.avna@gmail.com

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Introduction Several mechanisms are used to explain the coexistence of species with similar trophic niches, including their ability to segregate shared resources in space and time (Schoener, 1974; Gordon, 2000; Di Bitetti et al., 2010). However, the impact of the time of day on community dynamics, particularly the interactions between species that share food resources, has been neglected and is probably underestimated (Morgan, 2004). The activity pattern (AP) of an animal varies according to many factors, including its feeding habits (Karanth and Sunquist, 2000; Scognamillo et al., 2003; Carrillo et al., 2009), prey availability and diversity (Sunquist and Sunquist, 2002), the presence of predators and competitors (Scognamillo et al., 2003; Delibes–Mateos et al., 2014), temperature (Hernández–SaintMartin et al., 2013) and the level of natural and human disturbances (Van Dyke et al., 1986; Paviolo et al., 2009). Therefore, studying the APs of species with overlapping distributions can help explain how they partition their shared resources (Kronfeld–Schor and Dayan, 2003), including the temporal relationships between coexisting predatory species and their potential prey, which have resulted from evolutionary changes driven by competition for food (Abrams and Cortez, 2015). Camera traps are increasingly being used to monitor and assess the biodiversity of Protected Areas (Mandujano, 2017). They are also used to study the APs of wildlife (Di Bitetti et al., 2010; Hernández–SaintMartin et al., 2013; Rowcliffe et al., 2014; Ávila–Nájera et al., 2016) and to gather data on community interactions, such astemporal niche partitioning (Steenweg et al., 2017), community dynamics and species’ responses to global climate change (Frey et al., 2017). The Pico de Orizaba National Park (PONP) contains several nationally important ecosystems that are at risk of habitat fragmentation and anthropogenic disturbance despite having a protected status (SEMARNAT and CONABIO, 2015). The PONB is also thought to have a nationally important population of bobcat (Lynx rufus), and coyote (Canis latrans), but records from PONP are few (SEMARNAT and CONABIO, 2015) and information on coexistence strategies between both predators in Mexico (Hidalgo–Mihart et al., 2009; Elizalde–Arellano et al., 2014) is lacking. The aim of this study was to assess the APs of L. rufus and C. latrans, and their main prey species in the PONP, including the temporal overlap between these co–predators and potential prey. This will provide baseline data about the level of competition between them, including any evidence of temporal segregation and shared prey. These data could be used to monitor changes in this dynamic environment, which is at high risk of anthropogenic pressure.

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Veracruz State. PONP is located between 18º 56' 30'' and 19º 09' 3'' N and 97º 12' 30'' and 97º 2' 30'' W and has an altitudinal range of 2,700 to 5,760 m a.s.l. It has three main vegetation types: pine forest (up to 3,000 m a.s.l.), subalpine vegetation (3,000 to 4,000 m a.s.l.), and mountain scrub (4,000 to 4,200 m a.s.l.), with bare volcanic rock at higher altitudes. Despite its protected status, PONP is at risk of overgrazing by livestock, forest fires, and deforestation due to illegal logging (Martínez– Vázquez et al., 2010). Its fauna include 47 mammal, 48 amphibian and reptile, and 67 bird species, many of which are endangered in Mexico (Fa and Morales, 1991). Camera trapping Fourteen passive infrared sensor cameras (Cuddeback 1231 and Black Flash E3®) were sited in areas where mammals had been seen in a pilot study, with six on the western slope and eight on the eastern slope of the Pico de Orizaba Volcano. The positions of the camera stations were taken with a GPS–Garmin (fig. 1), with a distance between 1–3 km one from another. The cameras operated 24 h/day for 12 months (January–December 2017). They were placed 40 cm above ground level on tree trunks, and at an angle that avoided direct sunlight on the lens. The camera traps were examined monthly for battery replacement and SD memory cards. Videos were downloaded and reviewed, and records of predators and/or prey were sorted into different taxa for analysis (Mandujano, 2017). Activity patterns The following rules were applied to ensure that only independent records of individual animals were used when calculating AP: (i) individuals of the same species recorded in consecutive videos could identified by a distinguishing feature; (ii) when not possible to distinguish between individuals of the same species, at least 3 h had elapsed between the photos; (iii) more than one individual of the same species was visible in a single photo (Ávila–Nájera et al., 2016). The SUN TIMES V7.1 program (Kay and Du Croz, 2008) was used to determine the time of sunrise and sunset. All records of activity were classified according to the time on the video as: nocturnal (20:00–06:00 h), diurnal (08:00–18:00 h), or crepuscular (06:00–08:00 and 18:00–20:00 h). Species records were grouped into different APs: diurnal (˂ 15 % of observations at night), mainly diurnal (15 to 35 % of observations at night), nocturnal (> 85 % of observations at night), mainly nocturnal (65 to 85 % of the observations at night), cathemeral (intermittently active both at night and day) and crepuscular (active in the early hours of sunrise and sunset) (Gómez et al., 2005).

Material and methods

Overlap in activity patterns

The study took place around the PONP, Mexico. This park consists of 19,756 ha in the Tlachichuca, Chalchicomula de Sesma and Atzitzintla municipalities in Puebla State, and the Chalcahualco and La Perla municipalities in

The species were identified with their scientific name; in case of doubts with the family name only. As we were unable to identify small rodents they were all included as order Rodentia.

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97º 21' 0'' W

97º 19' 30'' W

97º 18' 0'' W

97º 16' 30'' W

97º 15' 0'' W

97º 13' 30'' W

97º 12' 0'' W

97º 13' 30'' W

97º 12' 0'' W

19º0'0''N 19º1'0''N 19º2'0''N 19º3'0''N 19º4'0''N 19º50''N 19º6'0''N 19º7'0''N 19º8'0''N 19º9'0''N

97º 22' 30'' W

— Altitude ● Camera traps 1 cm = 1 km

18º 58'30''N

2.5

18º 57'0''N

18º 58'30''N

10 km

Projection: North_America_Lambert_Conformal_Conic Angular Unit: degree (0.0174532925199433 97º 22' 30'' W

97º 21' 0'' W

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97º 18' 0'' W

97º 16' 30'' W

97º 15' 0'' W

Fig. 1. Position of the camera traps (black dots) used to monitor wildlife in the Pico de Orizaba National Park, Veracruz, Mexico, with the park boundary shown as a thick grey line, and contour lines in black (altitude, m a.s.l.). Fig. 1. Posición de las cámaras–trampa (puntos negros) utilizadas para monitorear la fauna silvestre en el Parque Nacional Pico de Orizaba, Veracruz, en México, la línea gris gruesa señala el límite del Parque y las líneas negras, la altimetría (altitud, m s.n.m.).

The overlap in the APs between predators and their prey was calculated using the total records for the whole year, and then for the dry (December–May) and the rainy seasons (June–November). As AP data have a circular distribution (Zar, 2010), we used a coefficient (Δ) to estimate the temporal overlap in the AP between predators and prey, where Δ is between 0 (no overlap) and 1 (complete overlap) (Ridout and Linkie, 2009) using the equation: Δ min {f (t), g(t)} / dt

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where are the values of two APs, and 95 % confidence intervals of Δ were estimated using 1,000 bootstrap repetitions at 2.5 and 97.5 percentiles. Statistical analyses were performed using the overlap library of R (version 3.1.0). Results A total sampling effort of 5,110 traps/night yielded 217 independent photographs of predators and potential

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Activity of all records (%)

100 90 80 70 60 50 40

Diurnal

Crepuscular

Nocturnal

10 0

Fig. 2. Frequency histogram showing the percentage of records for three activity patterns (diurnal, crepuscular and nocturnal) of two predators, bobcat (Lr, Lynx rufus) and coyote (Cl, Canis latrans), and their potential prey species (Dn, Dasypus novemcinctus; M, Mephitidae; R, Rodentia; Sc, Sciurus sp.; Sy, Sylvilagus sp.) recorded over 12 months using camera traps in Pico de Orizaba National Park, Veracruz, Mexico. Fig. 2. Histograma de frecuencias que muestra el porcentaje de registros de tres patrones de actividad (diurnos, crepusculares y nocturnos) de los dos depredadores, el lince (Lr, Lynx rufus) y el coyote (Cl, Canis latrans) y de sus presas potenciales (Dn, Dasypus novemcinctus; M, Mephitidae; R, Rodentia; Sc, Sciurus sp.; Sy, Sylvilagus sp.) registradas durante 12 meses usando cámaras–trampa en el Parque Nacional Pico de Orizaba, Veracruz, en México.

prey species. Throughout the year, the two predators exhibited mainly cathemeral activity: L. rufus (45; 20.7 %), and C. latrans (27; 11.5 %) (number of total independent photographs/percentage of total photographs). However, the potential prey were mostly nocturnal: Sylvilagus sp. (37; 17.1 %), Mephitidae (Hooded skunk, Mephitis macroura (10; 4.6 %), hog–nosed skunk, Conepatus leuconotus (3; 1.4 %), striped hog–nosed skunk, Conepatus semistriatus (1; 0.5 %), Dasypus novemcinctus, nine–banded armadillo (3; 1.4 %), and rodents (28; 13.4 %). The few diurnal species were Sciurus sp. (Peters’s squirrel, Sciurus oculatus (40; 18.4 %) and Mexican grey squirrel, Sciurus aureogaster (24; 11.1 %). Figure 2 shows the frequency of total records grouped by AP for the two predators and the main prey species (Rodentia, Mephitidae, Sciurus sp., Sylvilagus sp. and D. novemcinctus). The temporal overlap in the total activity recorded ^ = 0.80) (fig. 3) was high, for L. rufus and C. latrans (Δ particularly between 02:00–06:00 h, and at 16:00 and 20:00 h. This overlap was slightly higher in the rainy ^ = 0.79) than in the dry season (Δ ^ = 0.60) season (Δ (table 1). The APs of the two predators and potential prey species (table 1) differed, with the highest annual temporal overlap between the predators and prey ^ occurring between C. latrans and rodents (Δ = 0.80), ^ = 0.58). The and for L. rufus with Sylvilagus sp. (Δ temporal overlap between C. latrans and rodents ^ = 0.79) than in was higher in the rainy season (Δ ^ = 0.60), while the temporal overthe dry season (Δ

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lap between L. rufus and Sylvilagus sp. was slightly ^ = 0.59) than in the rainy higher in the dry season (Δ ^ season (Δ = 0.52). Records for the total and separate wet and dry season for the four species with the highest temporal overlaps (L. rufus, C. latrans, Sylvilagus sp. and rodents, see table 1) are plotted in figure 4. There was little seasonal variation in the number of L. rufus records, with peak activity occurring between 04:00– 06:00 h and 17:00–20:00 h, and fewer records in the hottest time of day (09:00–14:00 h). There were more records for C. latrans in the rainy season (n = 22) than in the dry season (n = 5), and coyotes were active throughout the day in the rainy season. Sylvilagus sp. had the highest number of records for a single species in the rainy season (n = 31), most of these being nocturnal (active between 19:00–04:00 h). Most rodent activity occurred in the rainy season (n = 25), and was mainly nocturnal (fig. 4). Discussion Niche segregation results from the distribution of resources, including temporal resource segregation between competing species. The two predators in this study can show marked variation in their circadian rhythms (Romero–Muñoz et al., 2010, Hernández–SaintMartin et al., 2013), as seen from their APs in this study, despite a high overall temporal overlap. Some studies show that the AP of predators is not totally conditioned by the

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Density

0.15

Coyote Bob Cat

d = 0.8

0.10

0.05

0.00 0:00

6:00

12:00 Hours

18:00

24:00

Fig. 3. Overlap in the activity patterns (Δ) of bobcat (Lynx rufus) and coyote (Canis latrans) in the Pico de Orizaba National Park, Veracruz, Mexico. Density is the frequency of records from January to December 2017 using camera traps. Fig. 3. Traslape de los patrones de actividad (Δ) del lince (Lynx rufus) y el coyote (Canis latrans) en el Parque Nacional Pico de Orizaba, en México. La densidad (density) es la frecuencia de registros tomados de enero a diciembre de 2017 por fototrampeo.

^ for two Table 1. Overlap coefficient values (Δ) species of predator, bobcat (Lynx rufus) and coyote (Canis latrans), and their potential prey (grouped by order or family) in the Pico de Orizaba National Park, Mexico. ^ Tabla 1. Valores del índice de solapamiento (Δ) entre dos especies depredadoras, el lince (Lynx rufus) y el coyote (Canis latrans) y sus presas potenciales (agrupadas en orden o familia) en el Parque Nacional Pico de Orizaba.

^ Overlap coefficient (∆)

Species tested

Global

Bobcat–coyote

0.80 0.80 0.60

Rain

Dry

Bobcat–lagomorphs 0.58 0.52 0.59 Bobcat–armadillos 0.29 0.26 – Bobcat–squirrels

0.39 0.44 0.37

Bobcat–skunks

0.57 0.50 0.45

Bobcat–rodents

0.57 0.50 0.45

Coyote–lagomorphs 0.50 0.48 0.38 Coyote–armadillos 0.35 0.36 – Coyote–squirrels

0.49 0.56 0.33

Coyote–skunks

0.45 0.21 0.39

Coyote–rodents

0.80 0.79 0.60

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activity of their prey (Elizalde–Arellano et al., 2012), with temporal segregation acting as an important mechanism to help carnivores avoid intraguild predation (Fedriani et al., 2000; Monterroso et al., 2014; Ávila–Nájera et al., 2016). However, other factors, such as human or natural disturbances, can significantly alter mammal behaviour (Monroy–Vilchis and Soria–Díaz, 2013; Ramesh and Downs, 2013; Ávila–Nájera et al., 2018), although high levels of activity at the hottest time of day are thought to be associated with reproduction in predatory species (Halle, 2000; Heurich et al., 2014). The AP of a species can thus be considered a complex response to its biotic or abiotic environment, especially where it interacts with other species (Halle, 2000). In this study, L. rufus and C. latrans had cathemeral habits. L. rufus was most active around 04:00–06:00 h and 17:00–19:00 h, as seen by Elizalde–Arellano et al. (2014), although it is thought to be mainly nocturnal like its main prey, lagomorphs (Aranda, 2002). However, in PONP C. latrans was active throughout the day, with peak activity between 05:00 and 06:00 h. This finding is similar to that in a study by González et al. (1992), who also found its diet was more omnivorous in spring–summer when it ate fruits and become mainly diurnal, spending less time searching for nocturnal prey. An overlap in activity between predator and prey may reflect that predators hunt when their prey are most active (Lima, 2002; Hernández, 2008; Romero– Muñoz et al., 2010), and could account for the high overlap between bobcat and lagomorph records in PONP, as reported by Hamilton and Hunter (1939), Leopold and Krausman (1986) and Aranda (2002).

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N 8

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7 6 5

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Canis latrans Global Rain season Riny season

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Sylvilagus sp.

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4 3

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00:00 01:00 02:00 03:00 04:00 05:00 06:00 07:00 08:00 09:00 10:00 11:00 12:00 13:00 14:00 15:00 16:00 17:00 18:00 19:00 20:00 21:00 22:00 23:00

Fig. 4. Annual total, dry and rainy season records of activity of two predators, Lynx rufus and Canis latrans, and their most frequently recorded prey items, Sylvilagus sp. and rodents, in the Pico de Orizaba National Park, Veracruz, Mexic: N, number of records. Fig. 4. Registros anuales totales y por temporada seca y lluviosa de dos depredadores, Lynx rufus y Canis latrans, y sus presas más frecuentemente registradas, Sylvilagus sp. y roedores, en el Parque Nacional Pico de Orizaba, Veracruz, México: N, número de registros.

Alternatively, predators may hunt opportunistically (Emmons, 1987), which could explain the partial overlap in activity between both predatory species and squirrels in PONP. Squirrels were found to be a component of the bobcat’s diet in another Mexican study (Aranda, 2002), and may have been under recorded by the camera traps because of their arboreal habit. Finally, predators may also hunt when their prey are least active (Sunquist, 1981; Emmons, 1987; Romero–Muñoz et al., 2010), and in PONP, skunks showed little overlap with the activity of the two predators (0.2) but are an important prey item elsewhere (Cruz–Espinoza et al., 2010). Neither C. latrans nor L. rufus are globally considered endangered, and both are widely distributed (Kelly et al., 2016). However, populations of L. rufus could be at risk because of hunting and illegal trade, despite being on Appendix I and II of CITES (Kelly et al., 2016). In PONP they are also at risk of forest fires (with frequent records from 1876 to 2011, particularly

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in the spring [97 %]), and decreasing rainfall (Cerano–Paredes et al., 2016). There is also considerable habitat loss due to clandestine logging and over use/ harvesting of forest resources (Ávila et al., 1994). In conclusion, the activity of bobcats and coyotes shows low temporal segregation, with their peak activities occurring at different times in PONP. However, both species can be active at any time of day and have a high temporal overlap across the year. This preliminary study shows that even a limited number of cameras can capture significant data about predatory species, especially those that are difficult to observe because they are rare or sparsely distributed or because they avoid human activity and have cathemeral or nocturnal habits. Small–scale studies of this type can add to the quality and quantity of records of poorly known predatory species, providing information concerning how their APs are impacted by their interactions with other species, including humans, and by disturbance and environmental change.

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Acknowledgements We thank the communities of the Pico de Orizaba National Park for their help with setting and maintaining the camera traps. We also thank the Unit of Management and Conservation of Genetic Resources of the Faculty of Biological and Agricultural Sciences Orizaba–Córdoba for providing resources for data analysis. We are grateful to Noé Hernández Garcia of CONANP and to two students, Roldán Vivas Lindo and Ángel Eduardo Carrasco González, for their support in organizing the database of records. References Abrams, P. A., Cortez, M. H., 2015. The many potential indirect interactions between predators that share competing prey. Ecological Monographs, 85: 625–641. Aranda, M., 2002. Análisis comparativo de la alimentación del gato montés (Lynx rufus) en dos diferentes ambientes de México. Acta Zoológica Mexicana, 87: 99–109. Ávila, C. H., Aguirre, J. R., García, E., 1994. Variación estructural del bosque de Oyamel (Abies hickelli Flous and Gaussen) en relación con factores ambientales en el Pico de Orizaba, México. Investigación Agraria en Sistemas de Recursos Forestales, 3: 5–17. Ávila–Nájera, D. M., Chávez, C., Lazcano–Barreto, M. A., Mendoza, G. D., Pérez–Elizalde, S., 2016. Traslape en patrones de actividad y traslape entre grandes felinos y sus principales presas en el norte de Quintana Roo, México. Therya, 7: 439–448. Ávila–Nájera, D. M., Chávez, C., Pérez–Elizalde, S., Guzmán–Plazola, R. A., Mendoza, G. D., Lazcano–Barrero, M. A., 2018. Ecology of Puma concolor (Carnivora: Felidae) in a Mexican tropical forest: adaptation to environmental disturbances. Journal of Tropical Ecology, 66(1): 78–90. Carrillo, E., Fuller, T. K., Saenz, J., 2009. Jaguar (Panthera onca) hunting activity: effects of prey distribution and availability. Journal of Tropical Ecology, 25: 563–567. Cerano–Paredes, J., Villanueva–Díaz, J., Vázquez– Selem, L., Cervantes–Martínez, R., Esquivel– Arriaga, G., Guerra–de la Cruz, V., Fulé, P., 2016. Régimen histórico de incendios y su relación con el clima en un bosque de Pinus hartwegii al norte del estado de Puebla, México. Bosque (Valdivia), 372: 389–399. Cruz–Espinoza, A., Pérez, G. E. G., Santos–Moreno, A., 2010. Dieta del Coyote (Canis latrans) en Ixtepeji, Sierra Madre de Oaxaca, México. Naturaleza y Desarrollo, 8: 33–45. Delibes–Mateos, M., Díaz–Ruiz, F., Caro, J., Ferreras, P., 2014. Activity patterns of the vulnerable guiña (Leopardus guigna) and its main prey in the Valdivian rainforest of southern Chile. Mammalian Biology–Zeitschrift für Säugetierkunde, 79: 393–397. Di Bitteti, M., De Angelo, C., Di blanco, Y., Paviolo, A., 2010. Niche partitioning and species coexistence

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Revista Mexicana de Mastozoología, 12: 113–130. Hernández–SaintMartin, A., Rosas–Rosas, O., Palacios–Núñez, J., Tarango–Arámbula, L., Clemente– Sánchez, F., Hoogesteijn, A., 2013. Activity patterns of jaguar, puma and their potential prey in San Luis Potosi, México. Acta Zoológica Mexicana (n. s.), 29: 520–533. Hidalgo–Mihart, M., Cantú–Salazar, G., Carrillo–Percastegui, L., Samia E., López–González, C. A., 2009. Daily activity patterns of coyotes (Canis latrans) in a tropical deciduous forest of western Mexico. Studies on Neotropical Fauna and Environment, 44: 77–82. Karanth, K. U., Sunquist, M. E., 2000. Behavioral correlates of predation by tiger (Panthera tigris), leopard (Panthera pardus) and dhole (Cuon alpinus) in Nagarahole, India. Journal of Zoology, 250: 255–265. Kay, S., Du Croz, T., 2008. Sun Times. Version 7.1. http://www.aptl72.dsl.pipex.com/suntimes.htm Kelly, M., Morin, D., López–González, C. A., 2016. Lynx rufus. In: The IUCN Red List of Threatened Species 2016: e.T12521A50655874 Kronfeld–Schor, N., Dayan, T., 2003. Partitioning of time as an ecological resource. Annual Review of Ecology, Evolution, and Systematics, 34: 153–181. Leopold, B. D., Krausman, P. R., 1986. Diet of 3 predators in Big Ben National Park, Texas. Journal of Wildlife Management, 50: 290–295. Lima, S. L., 2002. Putting predators back into behavioral predator–prey interactions. Trends in Ecology and Evolution, 17: 70–75. Mandujano, S., 2017. Monitoreo de la biodiversidad de mamíferos en áreas naturales protegidas empleando cámaras–trampa: sugerencias de herramientas para la gestión y el análisis numérico de las fotos. Paraquaria Natural, 5: 22–31. Martínez–Vázquez, J., González–Monroy, R. M., Díaz– Díaz, D., 2010. Hábitos alimentarios del Coyote en el parque nacional Pico de Orizaba. Therya, 1: 145–154. Monroy–Vilchis, O., Soria–Díaz., L., 2013. Ecología de Puma concolor en la Sierra Nanchititla, México. Universidad Autónoma del Estado de México. México. Monterroso, P., Alves, P. C., Ferreras, P., 2014. Plasticity in circadian activity patterns of mesocarnivores in Southwestern Europe: implications for species coexistence. Behavioral Ecology and Sociobiology, 68: 1403–1417. Morgan, E., 2004. Ecological significance of biological clocks. Biological Rhythm Research, 35: 3–12.

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Paviolo, A., Di Blanco, Y., De Angelo, C., Di Bitetti, M., 2009. Protection affects the abundance and activity patterns of puma in the Atlantic forest. Journal of Mammalogy, 90: 963–964. Ramesh, T., Downs, C. T., 2013. Impact of farmland use on population density and activity patterns of serval in South Africa. Journal of Mammalogy, 94: 1460–1470. Ridout, M. S., Linkie, M., 2009. Estimating overlap of daily activity patterns from camera trap data. Journal of Agricultural Biological Environmental Statistics, 14: 322–337. Romero–Muñoz, A., Maffei, L., Cuéllas, E., Noss, A., 2010. Temporal separation between jaguar and puma in the dry forests of southern Bolivia. Journal of Tropical Ecology, 26: 303–311. Rowcliffe, J. M., Kays, R., Kranstauber, B., Carbone, C., Jansen, P. A., 2014. Quantifying levels of animal activity using camera trap data. Methods in Ecology and Evolution, 5: 1170–1179. SEMARNAT and CONABIO, 2015. Programa de manejo Parque Nacional El Pico de Orizaba. Secretaría de Medio Ambiente y Recursos Naturales y Comisión Nacional de Áreas Naturales Protegidas. México. Scognamillo, D., Maxit, E. I., Sunquist, M., Polisar, J., 2003. Coexistence of Jaguar (Panthera onca) and puma (Puma concolor) in a mosaic landscape in the Venezuela Llanos. Journal of Zoology, 259: 269–279. Schoener, T. W., 1974. Resources partitioning in ecological communities. Science, 185: 27–39. Steenweg, R., Hebblewhite, M., Kays, R., Ahumada, J., Fisher, J. T., Burton, C., Townsend, S. E., Carbone, C., Rowcliffe, J. M., Whittington, J., Brodie, J., Royle, J. A., Switalski, A., Clevenger, P. A., Heim, N., Rich, N. L., 2017. Scaling–up camera traps: Monitoring the planet's biodiversity with networks of remote sensors. Frontiers in Ecology and the Environmental, 15: 26–34. Sunquist, M. E., 1981. The social organization of tigers (Panthera tigris) in Royal Chitawan National Park, Nepal. Smithsonian Contibutions to Zoology. Smithsonian Institution Press, Washington. Sunquist, M. E., Sunquist, F., 2002. Wild cats of the world. University of Chicago Press. Chicago, U.S.A. Van Dyke, F. G., Brocke, R. H., Shaw, H. G., Ackerman, B. B., Hemker, T. P., Lindzey, F. G., 1986. Reactions of mountain lions to logging and human activity. Journal of Wildlife Management, 50: 95–102. Zar, J. H., 2010. Bioestatistical analysis. Person Prentice Hall. New Jersey.

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Wind effects on habitat use by wintering waders in an inland lake of the Iberian Peninsula M. S. S. Gonçalves, J. A. Gil–Delgado, G. M. López–Iborra, P. dos Santos Pons

Gonçalves, M. S. S., Gil–Delgado, J. A., López–Iborra, G. M., dos Santos Pons, P., 2019. Wind effects on habitat use by wintering waders in an inland lake of the Iberian Peninsula. Animal Biodiversity and Conservation, 42.1: 163–169, Doi: https://doi.org/10.32800/abc.2019.42.0163 Abstract Wind effects on habitat use by wintering waders in an inland lake of the Iberian Peninsula. We aimed to identify the effects of the direction and wind speed on feeding habitat selection of wintering dunlins and little stints in an inland lake of the Iberian Peninsula. Feeding habitat (muddy surface or shallow water) and location in the lake with respect to wind direction (windward and leeward) of feeding flocks of both species were assessed on days with different wind speed (light or strong). We also performed visual counts of potential prey items (zooplankton) in mud and water habitats. In light wind conditions, wader flocks mostly selected the shallow water on the lake’s leeward shore. On the contrary, in strong wind conditions, the birds tended to forage on the windward shore, with a similar frequency in mud and shallow water habitats. The abundance of prey items in the mud and water column varied according to wind conditions, being higher in the sites preferred by waders. Our findings advance knowledge on how small–sized waders cope with environmental dynamics of wind in non–tidal lakes. Key words: Abiotic factors, Foraging habitat, Saline lakes, Shorebirds, Wintering season Resumen Los efectos del viento en el uso del hábitat que hacen las aves limícolas invernantes en una laguna interior en la península ibérica. Nuestro objetivo fue determinar los efectos de la dirección y la velocidad del viento en el uso del hábitat de alimentación que hacen el correlimos común y el correlimos menudo en una laguna interior de la península ibérica. Se evaluaron el hábitat de alimentación (superficie fangosa o aguas poco profundas) y la ubicación en la laguna con respecto a la dirección del viento (barlovento y sotavento) de los grupos de limícolas en días con diferente velocidad del viento (suave y fuerte). También se realizaron recuentos visuales de posibles presas (zooplancton) en hábitats de lodo y agua. En condiciones de viento suave, el grupo de limícolas seleccionó principalmente la zona de aguas poco profundas ubicada a sotavento de la laguna. Por el contrario, en condiciones de viento fuerte, las aves tendieron a buscar comida en la orilla de barlovento, con una frecuencia similar en los hábitats de lodo y de aguas poco profundas. La abundancia de presas en las columnas de lodo y de agua varió según las condiciones del viento, y fue más alta en los sitios preferidos por los limícolas. Nuestros resultados se suman al conocimiento de cómo los limícolas de tamaño pequeño hacen frente a la dinámica ambiental del viento en lagunas no mareales. Palabras claves: Factores abióticos, Hábitat de forrajeo, Lagunas salinas, Aves limícolas, Período invernal Received: 03 IV 18; Conditional acceptance: 29 VI 18; Final acceptance: 02 X 18 Maycon Sanyvan Sigales Gonçalves, José Antonio Gil–Delgado, Priscila dos Santos Pons, Inst. Cavanilles de Biodiversidad y Biología Evolutiva, Univ. de Valencia, Paterna, Valencia, Spain.– Germán Manuel López–Iborra, Depto. de Ecología / IMEM Ramon Margalef, Univ. de Alicante, Alicante, Spain. Corresponding author: M. Sanyvan Sigales Gonçalves. Email: mayconsanyvan@gmail.com

ISSN: 1578–665 X eISSN: 2014–928 X

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Introduction Multiple environmental components act independently or in association with one another to influence habitat selection and use by birds (Jones, 2001). The choice of foraging sites by waders is fundamentally a consequence of the balance of costs and benefits of feeding in potential foraging habitats, defined mainly by depth of the water column and the quality and quantity of prey (e.g. Piersma, 2006; Granadeiro et al., 2007; Beerens et al., 2015a, 2015b). Other important factors, such as predation risk (Mikula et al., 2018), anthropic disturbances (Holm and Laursen, 2009), eco–physiological adaptations (Gutiérrez et al., 2012; Lourenço and Piersma, 2015), and roost location and landscape attributes (Dias et al., 2014; Santiago–Quesada et al., 2014) are associated with wader habitat selection outside the breeding season, but they have been poorly documented for inland natural wetlands. Large numbers of small– and medium–sized waders rely on inland natural wetlands to replenish their energy stores during their migrations (Verkuil et al., 1993). In these non–tidal wetlands, wind significantly affects the predator–prey relationship by altering detection or locomotion, and consequently, the foraging habitat or microhabitat selection patterns (Verkuil et al., 1993; Cherry and Barton, 2017). The main food supply for waders is aquatic invertebrates, which can be detected and captured visually or using tactile sensibility (Piersma et al., 1996). Many waders, especially scolopacid species, are long–billed and may feed efficiently on small prey items suspended in water by using distal rhynchokinesis and a feeding mechanism termed surface tension transport (Estrella and Masero, 2007; Estrella et al., 2007). Wind can act in two different ways on aquatic invertebrate distribution. On one hand, it can centre on specific sites of the foraging grounds, concentrating prey and facilitating the visual strategy (Verkuil et al., 1993). On the other hand, it can expose the organisms on the mud surface, favoring both visual and tactile detection (Verkuil et al., 1993, 2003; Masero et al., 2000). Most studies on wader habitat selection focus on coastal systems, where most migratory wader populations spend the wintering season (van de Kam et al., 2004). In the Iberian Peninsula, dunlin (Calidris alpina) and little stint (C. minuta) are two of the most abundant wintering waders and most of their populations winter in coastal areas (SEO/BirdLife, 2012). Specifically, the winter populations of dunlin and little stint in Spain are approximately 100,000 and 12,813 individuals, respectively, and for both species, less than 5 % of these individuals spend this season in continental wetlands (SEO/BirdLife, 2012; BirdLife International, 2018). Although several studies have been carried out on habitat selection by waders wintering in Iberia, they were performed on intertidal habitats or coastal areas (e.g. Masero et al., 2000; Masero and Pérez–Hurtado, 2001; Dias, 2009; Lourenço et al., 2013; Martins et al., 2016), and information on wader habitat use or selection during the winter period in continental Iberian wetlands is lacking.

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In this study, we aimed to determine how wind conditions influenced microhabitat use by two small–medium wader species –dunlin and little stint– foraging in an inland wetland located in 'La Mancha Húmeda' Biosphere Reserve (Spain). We studied the relationship between wind direction and wind speed and the selection of two foraging habitats –mud and shallow water. In accordance with Verkuil et al. (1993), we expected to find that wind direction and speed had significant effects on the spatial distribution of aquatic invertebrates and, consequently, on the habitat use by foraging waders. Material and methods Study area 'La Mancha Húmeda' Biosphere Reserve in central Spain is formed by a constellation of temporary and permanent wetlands of international importance (SEO/BirdLife, 2012; BirdLife International, 2018). This study was conducted in Lake Alcahozo (N 39º 23' 26,7'' / W 2º 52' 35,7''), a temporary, natural wetland located in this reserve (fig. 1). The lake has a roughly circular shape and covers an area of approximately 88 ha, surrounded by 19 ha of natural halophile vegetation (Gonçalves et al., 2016). It has homogeneously flat shores and a maximum depth of approximately 45 cm. There are no sedimentary islands within the lake. Data collection and sampling design To evaluate the foraging habitat use, we applied a used sampling design versus a non–used sampling design (Jones, 2001). Habitat use was recorded between December and February (winters 2014–15 and 2015–16). Previous observations showed that dunlins and little stints often foraged in mixed flocks. Twice a week, we selected the largest foraging flock inside the lake and recorded its habitat use. Observations were separated by a minimum of seven hours and a maximum of 120 hours, with a maximum of two samplings per day. The size of the mixed flock varied between 145 and 530 individuals (mean = 258.9; SD = 130.7), usually with the dominance of dunlins (60–80 % of the individuals). The largest flock usually concentrated around 90% of the individuals of both species present at the lake at counting time. When waders were observed foraging for at least five minutes in a single habitat, the foraging habitat was assigned to one of the following types: mud (muddy surface at the lake’s shore) or shallow water (< 2 cm) (fig. 2). We next identified their position in relation to the wind direction as windward or leeward (fig. 2). Following this identification, the wader flocks could be found in any of four situations: mud/leeward, mud/windward, water/leeward and water/windward. We also defined two wind speed classes, light (0–12 km/h) and strong (> 25 km/h) winds. Wind speed was measured using a handheld Brunton ADC Atmospheric Data Center (Brunton, Inc., USA). The lake was visited frequently during both winters until thirty observations were accumulated for each wind/habitat situation.

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44ºN

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40ºN

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36ºN 8ºW

4ºW

1.1 km

0ºW

Fig. 1. Location (A) and view (B and C) of Lake Alcahozo in 'La Mancha Húmeda' Biosphere Reserve, Spain. Fig. 1. Localización (A) y vista (B y C) de la laguna de Alcahozo en la Reserva de la Biosfera de La Mancha Húmeda, en España.

Anostracans and copepods are abundant in Lake Alcahozo and commonly distributed in the water column (Pons et al., 2018). Small–sized waders such as dunlins and little stints feed on these prey items (Verkuil et al., 1993). In December 2014, preliminary inspections noted that by approaching cautiously and remaining motionless, these and other groups, such as ostracods, could be observed at a close distance. Although the best method to quantify the abundance of invertebrates is the collection of water samples and identification in the laboratory (e.g. Pons et al., 2018), careful visual counts can also be good indicators of the quantity of prey in aquatic systems (e.g. McIntosh and Townsend, 1996). To observe whether prey abundance varied with wind conditions and influenced the habitat use of waders, we obtained an index of invertebrate abundance from visual counts in the four habitat/wind combinations (fig. 2). Samplings were performed between January and February 2016. Invertebrate counts were performed on six days for each wind speed class. Five plots of 40 cm x 40 cm distributed every 2 m along a line parallel to the shore were placed in mud and water, both in windward and leeward positions. Therefore, 20 plots were surveyed each counting day. The observer's visual distance to the water or mud surface was 40 cm. All invertebrates observed inside a plot for 30 seconds were counted. One set of plots was located at the habitat/wind direc-

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tion combination where the flock was observed and the others were located at places representative of the other three categories not used at that moment. The location of the plots thus varied at each sampling occasion because it depended on the location of the waders and wind direction. Data analysis We used log–linear models to evaluate whether observations of wader flocks were preferentially associated with particular habitat types as a response to wind. We created a three–dimensional contingency table with the factors habitat (mud/water), position relative to wind (windward/leeward) and wind speed (light/strong). The number of wader flocks found in each combination of the three factors was the response variable. The significance of the interactions between the factors considered was tested by removing interactions in turn from the most general model and comparing resulting models using the chi–square test (Bolker et al., 2009). Interactions were removed until a significant term was identified. To assess invertebrate abundance, we tested the effects of abundance of the same factors cited above using Generalized Linear Mixed Models (GLMMs) with Poisson distribution. The response variable was the total number of invertebrates counted in each plot.

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Windward

Leeward

Fig. 2. Scheme showing the four possible distributions of the flocks (represented by a single bird) of waders during winter at Lake Alcahozo. At each observation, only one of four situations was possible: windward or leeward and mud (light gray) or shallow water (dark gray). Fig. 2. Esquema en el que se muestran las cuatro posibles distribuciones de los grupos de limícolas (representados por un único ave) durante el invierno en la laguna de Alcahozo. En cada observación, solo podía darse una de las cuatro situaciones: barlovento o sotavento y lodo (en gris claro) o aguas poco profundas (en gris oscuro).

Habitat type, position relative to wind and wind speed category were factors included as fixed effects. The sampling day was considered a random effect. As above, the he significance of interactions between fixed effects was tested by removing interactions in turn from the general model and comparing model deviances using the chi–square test (Bolker et al., 2009). All analyses were performed using the R–programming environment (R Development Core Team, 2016). The GLMMs were performed using the package 'lme4' (Bates et al., 2013). Results In light wind conditions, 74 % of the observations of foraging flocks were made on the leeward of the lake and were mainly of birds feeding in the water column (67 %) (fig. 3A). In strong wind conditions, 80 % of foraging flocks were found on the windward side, with 50 % and 30% of flocks using the mud and the water column, respectively (fig. 3A). All possible interactions were significant (table 1). We identified three taxonomic groups of invertebrates during the visual counts: Anostraca, Copepoda, and Ostracoda. Under light wind conditions, invertebrate abundance was highest in shallow water, with the highest values located on the leeward side (fig. 3B). However, under strong wind conditions, invertebrate abundance was much higher in mud than in water on the windward side, while it was similar in both habitat types on the leeward side in both habitat types (fig. 3B). Interaction between the three factors (wind speed, direction and habitat type) was not significant (x2 = 0.95; df = 1; p = 0.32), but the two–way interactions were all significant (p < 0.01, in all cases) (table 1).

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Discussion Wind speed had a strong effect on feeding habitat selection by overwintering dunlins and little stints. The circular shape and the homogeneous edge of the lake avoided local confounding variables potentially associated with the wind effects. Our results showed that under light wind conditions, birds foraged mainly in shallow water, on the leeward side of the lake, while under strong wind conditions, birds foraged on the windward shore and used both mud and shallow water microhabitats. Invertebrate abundance in mud and water microhabitats also varied with wind conditions, being higher in the sites preferred by waders. In shallow lakes, such as our study area, strong wind movements rapidly affect the water column, altering the spatial distribution of zooplankton (Cardoso and Marques, 2009) and increasing turbidity due to sediments in suspension (G–Tóth et al., 2011). In these ecosystems, therefore, wind potentially affects wader prey abundance and detectability. Our results agree with the general patterns that Verkuil et al. (1993, 2003) observed in shallow lagoons in the Ukraine region. Under light wind conditions, most feeding wader flocks used the water habitat on the leeward coast. This pattern was possibly a response to the gentle displacement of prey items towards the leeward shore. In addition, the light wind keeps water turbidity low, facilitating visual detection and predation on invertebrates. In this situation, the distribution of feeding waders flocks mirrors the variations in invertebrate abundance almost perfectly. In contrast, under strong wind conditions, wader flocks mainly used the windward side of the lake for foraging activities. This may be due to two causes. First, the area of exposed mud is greatly reduced on

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100 90 80 70 60 50 40 30 20 10 0

Windward

Observations (%)

Leeward

Mud

Mean abundance

Strong wind

Water

Windward

Leeward

10 8 6 4 2 0

Mud

Strong wind

Water

167

100 90 80 70 60 50 40 30 20 10 0

40 35 30 25 20 15 10 5 0

Windward Leeward

Mud

Light wind

Water

Windward Leeward

Mud

Light wind

Water

Fig. 3. A, percentage of observations of the wader flock based on 30 flocks observed under each wind speed category (± SE calculated using binomial distribution); B, invertebrates abundance (± SE) between mud and water habitats in relation to the direction (windward or leeward) and wind category (strong or light). Fig. 3. A, porcentaje de observaciones del grupo de limícolas basado en 30 grupos observados en cada categoría de velocidad del viento (± DE calculada utilizando la distribución binomial); B, abundancia de invertebrados (± DE) entre los hábitats de lodo y de agua en relación con la dirección (barlovento o sotavento) y la categoría del viento (fuerte o ligero).

the leeward shore and the water column moves so rapidly that prey are exposed on the muddy surface on the windward shore (Verkuil et al., 1993), facilitating predation. Second, in addition, under strong wind conditions, the prey abundance index is relatively low at the water column in both positions relative to the wind, while the frequency of waders feeding in water does not vary so much on leeward and windward sides. This suggests that the prey abundance index may be affected by turbidity and tends to underestimate prey abundance under these conditions, while waders may overcome this problem, at least partially, using tactile senses to forage (Estrella and Masero, 2007). The visual counts of invertebrates identified Anostraca, Copepoda, and Ostracoda as the main groups. At Lake Alcahozo, these groups are represented by Branchinectella media (Anostraca), Arctodiaptomus salinus (Copepoda) (Pons et al., 2018) and Heterocypris barbara (Ostracoda) (Castillo–Escrivà et al., 2015). Information about spatial distribution and habitat selection is available only for B. media. Pons et al. (2018) investigated the spatial

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distribution of B. media on days without wind and observed that adult individuals tend to occur on the central region of the lake, while juveniles are most commonly observed on the shore. Although our visual counts were only a proxy of prey abundance and availability, they showed the direction and wind speed are important factors influencing the spatial distribution of potential prey items, including B. media. For both wind speed categories (smooth and strong), our results provide evidence that feeding flocks of dunlin and little stint seek to select sites with the highest abundance of prey. They therefore change their spatial distribution according to the wind conditions. However, we did not take into account the potential influence of predators or other birds that could influence habitat choice irrespectively of wind. Although we did not observe any predation attempts, the approximation of some species such as the lesser black–backed gull (Larus fuscus), the black–headed gull (L. ridibundus) and the Eurasian marsh harrier (Circus aeruginosus) often forced the flocks to change their foraging behavior. Specifically,

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Acknowledgements Table 1. Tests of interactions of factors in habitat use models of waders (log–linear models) and invertebrate abundance (GLMM). Interaction terms were tested by removing each interaction in turn and comparing resulting models using the x2–test: Ws, wind speed; P, position; H, habitat. Tabla 1. Prueba de interacciones de factores en los modelos de uso del hábitat de los limícolas (modelo log–lineal) y abundancia de invertebrados (modelo mixto lineal generalizado). Los términos de la interacción se pusieron a prueba eliminando todas las interacciones por turnos y comparando los modelos resultantes con el uso de la x2: Ws, velocidad del viento; P, posición; H, hábitat. Interaction

df p–value

Wader flock habitat Ws x P x H

3,95

1 0,0468

Invertebrate abundance Ws x P x H

0,95

1 0,3288

Ws x P

82,79

1 < 0,001

Ws x H

588,27

1 < 0,001

P x H

69,59

1 < 0,001

the waders presented three responses: short distance displacement before their immediate return to the same site; long distance displacement without return to the foraging site; and interruption of feeding. In addition, during the winter, we observed more than 1,500 gulls (Larus spp.) gathered to roost at the lake, and depending on the location chosen to settle, their arrival could have been a physical barrier to the habitat use of the waders. However, the gulls always roosted 50–60 m from the shoreline and therefore did not directly influence the habitat use of waders. Such effects on waders using these inland wetlands should be considered in research on the habitat selection by waders. Findings in this study add to our understanding of factors influencing the foraging patterns of small– sized waders in inland (non–tidal) lakes. Wetlands in central Spain, especially in the region of the 'La Mancha Húmeda' Biosphere Reserve, are generally surrounded by plantations and most of the lakes have no physical protection limiting the advance of the agricultural frontier (Gonçalves et al., 2018b). The shore topography of some of these lakes has been modified in the past by agricultural and infrastructure works, including transformations for receiving wastewater from nearby towns (Gonçalves et al., 2018a). Thus, artificial modifications of the topography of inland lakes that limit availability of some preferred habitats under some wind conditions could reduce the suitability of these wetlands for wintering waders.

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This study forms part of the doctoral thesis of M. S. S. G. and was supported by a grant from CAPES–Coordenação de Aperfeiçoamento de Pessoal de Nível Superior, Brazil (IBEX 0850/14–4). This study was jointly supported by the Spanish Ministry of Economy and Competitiveness and the European Regional Development Fund (FEDER) 'One way to make Europe', through the projects: CLIMAWET–Climate change mitigation and adaptation in the main types of Iberian Mediterranean wetlands: carbon budged and response models of species and habitats (CGL2015–69557–R); and ECOLAKE–Ecological patterns in endorheic lakes: keys to their conservation (CGL2012–38909). References Bates, D., Maechler, M., Bolker, B., 2013. lme4: Linear–mixed Effects Models Using S4 Classes. Available online at: http://CRAN.R–project.org/ package=lme4 [Accessed on 18 January 2018]. Beerens, J. M., Frederick, P. C., Noonburg, E. G., Gawlik, D. E., 2015a. Determining habitat quality for species that demonstrate dynamic habitat selection. Ecology and Evolution, 5: 5685–5697. Beerens, J. M., Noonburg, E. G., Gawlik, D. E., 2015b. Linking dynamic habitat selection with wading bird foraging distributions across resource gradients. Plos One, 10(6): e0128182. BirdLife International, 2018. IUCN Red List for birds. Available online at: http://www.birdlife.org [Accessed on 18 January 2018]. Bolker, B. M., Brooks, M. E., Clark, C. J., Geange, S. W., Poulsen, J. R., Stevens, M. H. H., White, J. S. S., 2009. Generalized linear mixed models: a practical guide for ecology and evolution. Trends in Ecology & Evolution, 24: 127–135. Cardoso, L. S., Marques, D. M., 2009. Hydrodynamics–driven plankton community in a shallow lake. Aquatic Ecology, 43: 73–84. Castillo–Escrivà, A, Valls, L., Rochera, C., Camacho, A., Mesquita–Joanes, F., 2015. Spatial and environmental analysis of an ostracod metacommunity from endorheic lakes. Aquatic Science, 78: 707–716. Cherry, M. J., Barton, B. T., 2017. Effects of wind on predator–prey interactions. Food Webs, 13: 92–97. Dias, M. P., 2009. Use of salt ponds by wintering shorebirds throughout the tidal cycle. Waterbirds, 32: 531–537. Dias, R. A., Blanco, D. E., Goijman, A. P., Zaccagnini, M. E., 2014. Density, habitat use, and opportunities for conservation of shorebirds in rice fields in southeastern South America. The Condor, 116: 384–393. Estrella, S. M., Masero, J. A., 2007. The use of distal rhynchokinesis by birds feeding in water. Journal of Experimental Biology, 210: 3757–3762. Estrella, S. M., Masero, J. A., Pérez–Hurtado, A., 2007. Small prey profitability: field analysis of shorebirds' use of surface tension of water to transport prey. The Auk, 124: 1244–1253.

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Gonçalves, M. S. S., Gil–Delgado, J. A., Gosálvez, R. U., López–Iborra, G. M., Ponz, A., Velasco, A., 2016. Spatial synchrony of wader populations in inland lakes of the Iberian Peninsula. Ecological Research, 31: 947–956. Gonçalves, M. S. S., Gil–Delgado, J. A., Gosálvez, R. U., Florín, M. B., López–Iborra, G. M., 2018a. RE: Spanish too low emphasis on science does not help either with wetlands [e–letter]. Available online at: http://science.sciencemag.org/content/361/6398/111/tab–e–letters [Accessed on 14 August 2018]. Gonçalves, M. S., Gil–Delgado, J. A., Gosalvez, R. U., López–Iborra, G. M., Ponz, A., Velasco, A., 2018b. Seasonal differences in drivers of species richness of waders in inland wetlands of the 'La Mancha Húmeda' Biosphere Reserve. Aquatic Conservation: Marine and Freshwater Ecosystems: 1–8, doi:10.1002/aqc.2968 Granadeiro, J. P., Santos, C. D., Dias, M. P., Palmeirim, J. M., 2007. Environmental factors drive habitat partitioning in birds feeding in intertidal flats:implications for conservation. Hydrobiologia, 587: 291−302. G–Tóth, L., Parpala, L., Balogh, C., Tàtrai, I., Baranyai, E., 2011. Zooplankton community response to enhanced turbulence generated by water level decrease in Lake Balaton, the largest shallow lake in Central Europe. Limnology and Oceanography, 56: 2211–2222. Gutiérrez, J. S., Dietz, M. W., Masero, J.A., Gill, R. E. Jr., Dekinga, A., Battley, P. F., Sánchez–Guzmán, J. M., Piersma, T., 2012. Functional ecology of saltglands in shorebirds: flexible responses to variable environmental conditions. Functional Ecology, 26: 236–244. Holm, T. E., Laursen, K., 2009. Experimental disturbance by walkers affects behaviour and territory density of nesting black–tailed godwit Limosa limosa. Ibis, 151: 77–87. Hutto, R. L., 1985. Habitat selection by non–breeding, migratory land birds. In: Habitat Selection in Birds: 455–476 (M. L. Cody, Ed.). Academic Press, New York. Jones, J., 2001. Habitat Selection Studies in Avian Ecology: A Critical Review. The Auk, 118: 557–562. Lourenço, P. M., Catry, P., Lecoq, M., Ramírez, I., Granadeiro, J. P., 2013. Role of disturbance, geology and other environmental factors in determining abundance and diversity in coastal avian communities during winter. Marine Ecology Progress Series, 479: 223–234. Lourenço, P. M., Piersma, T., 2015. Migration distance and breeding latitude correlate with the scheduling of pre–alternate body moult: a comparison among migratory waders. Journal of Ornithology, 156: 657–665. Martins, R. C., Catry, T., Rebelo, R., Pardal, S., Palmeirim, J. M., Granadeiro, J. P., 2016. Contrasting estuary–scale distribution of wintering and migrat-

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ing waders: The potential role of fear. Hydrobiologia, 768: 211–222. Masero, J. A., Pérez–Hurtado, A., 2001. Importance of the supratidal habitats for maintaining overwintering shorebird populations: how Redshanks use tidal mudflats and adjacent saltworks in southern Europe. The Condor, 103: 21–30. Masero, J. A., Pérez–Hurtado, A., Castro, M., Arroyo, G. M., 2000. Complementary use of intertidal mudflats and adjacent salinas by foraging waders. Ardea, 88: 177–191. McIntosh, A. R., Townsend, C. R., 1996. Interactions between fish, grazing invertebrates and algae in a New Zealand stream: a trophic cascade mediated by fish–induced changes to grazer behaviour? Oecologia, 108: 174–181. Mikula, P., Díaz, M., Møller, A. P., Albrecht, T., Tryjanowski, P., Hromada, M., 2018. Migratory and resident waders differ in risk taking on the wintering grounds. Behavioural Processes, 157: 309–314, doi: https://doi.org/10.1016/j.beproc.2018.07.020 Piersma, T., van Gils, J. A., Wiersma, P., 1996. Family Scolopacidae (sandpipers, snipes and phalaropes). In: Handbook of the Birds of the World: 444–533 (J. del Hoyo, A. Elliott, J. Sargatal, Eds.). Lynx Edicions, Barcelona. Piersma, T., 2006. Understanding the numbers and distribution of waders and other animals in a changing World: habitat choice as the lock and the key. Stilt, 50: 3–14. Pons, P., Gonçalves, M. S. S., Gil–Delgado, J. A., Ortells, R., 2018. Spatial distribution of Branchinectella media (Crustacea, Branchiopoda) in a saline pond from 'La Mancha Húmeda': a case of habitat selection?. Limnetica, 37: 69–83. R Development Core Team, 2016. R: A language and environment for statistical computing. R Foundation for Statistical Computing, Vienna, Austria. Santiago–Quesada, F., Masero, J. A., Albano, N., Sánchez–Guzmán, J. M., 2014. Roost location and landscape attributes influencing habitat selectionof migratory waterbirds in rice fields. Agriculture, Ecosystems and Environment, 188: 97–102. SEO/BirdLife, 2012. Atlas de las aves en invierno en España 2007–2010. Ministerio de Agricultura, Alimentacion y Medio Ambiente–SEO/BirdLife. Madrid. van de Kam, J., Ens, B., Piersma, T., Zwarts, L., 2004. Shorebirds: an illustrated Behavioural Ecology. KNNV Publishers, Utrecht Verkuil, Y., Koolhaas, A., Winden, J. V. D., 1993. Wind effects on prey availability: how northward migrating waders use brackish and hypersaline lagoons in the Sivash, Ukraine. Netherlands Journal of Sea Research, 31: 359–374. Verkuil, Y., van der Have, T. M., van der Winden, J., Chernichko, I. I., 2003. Habitat use and diet selection of northward migrating waders in the Sivash (Ukraine): the use of brine shrimp Artemia salina in a variably saline lagoon complex. Ardea, 91: 71–83.

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Mitochondrial evidence for a new evolutionary significant unit within the Gila eremica lineage (Teleostei, Cyprinidae) in Sonora, Northwest Mexico C. A. Ballesteros–Córdova, A. Varela–Romero, G. Ruiz–Campos, L. T. Findley, J. M. Grijalva–Chon, L. E. Gutiérrez–Millán Ballesteros–Córdova, C. A., Varela–Romero, A., Ruiz–Campos, G., Findley, L. T., Grijalva–Chon, J. M., Gutiérrez– Millán, L. E., 2019. Mitochondrial evidence for a new evolutionary significant unit within the Gila eremica lineage (Teleostei, Cyprinidae) in Sonora, Northwest Mexico. Animal Biodiversity and Conservation, 42.1: 171–186, Doi: https://doi.org/10.32800/abc.2019.42.0171 Abstract Mitochondrial evidence for a new evolutionary significant unit within the Gila eremica lineage (Teleostei, Cyprinidae) in Sonora, Northwest Mexico. We present the phylogenetic affinities and DNA barcode of Gila cf. eremica, a geographically isolated and morphologically divergent population from G. eremica DeMarais, 1991. Mitochondrial phylogenetic analyses of cyt–b, cox1 and nd2 show a clades pattern within the G. eremica lineage, placing G. cf. eremica in a clade of specific identity to and sharing a putative common ancestor with G. eremica from the Mátape River basin. The barcoding analysis using a character–based approach of CAOS showed seven single pure characters discriminating G. eremica from its regional congener G. purpurea, and one fixed character in G. cf. eremica discriminating it from G. eremica. These results and the recent detection of diagnostic morphological differences between G. cf. eremica and G. eremica support the hypothesis of Gila cf. eremica as an significant evolutionary unit within the G. eremica lineage. Key words: Gila, Phylogenetic analyses, DNA barcode, Evolutionary significant unit, Northwest México Resumen Evidencia mitocondrial de una nueva unidad evolutivamente significativa en el linaje de Gila eremica (Teleostei, Cyprinidae) en Sonora, en el noroeste de México. Presentamos las afinidades filogenéticas y el código de barras del ADN de Gila cf. eremica, una población morfológicamente divergente y geográficamente aislada de G. eremica DeMarais, 1991. Los análisis filogenéticos mitocondriales de cyt–b, cox1 y nd2 muestran la existencia de un patrón de clados dentro del linaje de G. eremica, que sitúa a G. cf. eremica en un clado de identidad específica que comparte un supuesto ancestro común con G. eremica, originario de la cuenca del río Mátape. El análisis de código de barras, en que se utilizó un método basado en caracteres del programa informático CAOS, mostró siete caracteres puros que diferencian a G. eremica de su congénere regional G. purpurea y un carácter fijo en G. cf. eremica que lo diferencia de G. eremica. Estos resultados y la reciente detección de diferencias morfológicas diagnósticas entre G. cf. eremica y G. eremica sostienen la hipótesis de que Gila cf. eremica es una unidad evolutivamente significativa dentro del linaje de G. eremica. Palabras clave: Gila, Análisis filogenéticos, Código de barras de ADN, Unidad evolutivamente significativa, Noroeste de México Received: 28 V 18; Conditional acceptance: 12 IX 18; Final acceptance: 09 X 18 C. A. Ballesteros–Córdova, A. Varela–Romero, J. M. Grijalva–Chon, L. E. Gutiérrez–Millán, Departamento de Investigaciones Científicas y Tecnológicas de la Universidad de Sonora, Blvd. Luis Encinas y Rosales s/n., 83000 Hermosillo, Sonora, México.– G. Ruiz–Campos, Facultad de Ciencias, Universidad Autónoma de Baja California, 22800 Ensenada, Baja California, México.– L. T. Findley, Centro de Investigación en Alimentación y Desarrollo, A. C. Unidad Guaymas, 85480 Guaymas, Sonora, México. Corresponding author: A. Varela–Romero. E–mail: alejandro.varela@unison.mx

ISSN: 1578–665 X eISSN: 2014–928 X

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Introduction The genus Gila, one of the most widespread groups of the family Cyprinidae in North America, includes morphologically heterogeneous fishes inhabiting waters of arid and semiarid regions of the western United States (USA) and northwestern Mexico (Miller et al., 2005). Using both mitochondrial and nuclear markers it has been observed that recent phylogenetic analyses of cyprinids in North Americaalign Gila with ten other nominal genera within a so–called Revised Western Clade (RWC) (Schönhuth et al., 2012), and suggest that Gila comprises an evolutionary lineage involving at least 18 species, including species currently recognized taxonomically in the monotypic genera Acrocheilus and Moapa (Schönhuth et al., 2014). Nevertheless, Schönhuth and colleagues termed he composition of the Gila lineage incomplete because of phylogenetic affinities of G. coerulea and Ptychocheilus lucius that were only resolved within Gila by using mitochondrial marker cyt–b rather than the concatenated nuclear genes rag1, rhod, and s7 or concatenated mitochondrial and nuclear markers (Schönhuth et al., 2012, 2014). Molecular and morphological analyses of Gila species occurring in northern México and western USA suggest that nominal species of the G. robusta complex in the Colorado River basin show both allopatric and sympatric distributions, with probable hybrid origins at least in part (DeMarais et al., 1992; Dowling and DeMarais, 1993; Gerber et al., 2001; Schönhuth et al., 2014; Dowling et al., 2015; Page et al., 2017). Gila species in Mexican waters of the Atlantic slope Chihuahuan Desert region, and those in the Pacific slope, however, apparently show mainly allopatric distributions associated with major river drainages, suggesting peripatric speciation events (Wiley, 1981; Schönhuth et al., 2014). Phylogenetic affinities and current distributions of all known species of Gila show that those in Mexico include an Atlantic–slope lineage referred to as the Chihuahuan Desert Group, which includes G. pulchra, G. conspersa, G. nigrescens, G. brevicauda, an undescribed species, and a lineage composed of G. modesta nested within G. pandora (Schönhuth et al., 2014). Their phylogenetic affinities, morphological characteristics, and geographical distributions suggest that species in this group share a single common ancestor, unrelated to any species belonging to the G. robusta complex of the Colorado River system (Uyeno, 1960; Schönhuth et al., 2014). The remaining nominal species of Gila in México (G. ditaenia, G. minacae, G. purpurea, G. eremica) occur in Pacific–slope drainages. These four species were not resolved in analyses by Schönhuth et al. (2014) as part of their Chihuahuan Desert Group. However, they were resolved as monophyletic within the greater Gila lineage, and G. eremica and G. purpurea were corroborated as sister species (Schönhuth et al., 2012, 2014), as proposed by previous morphological analysis (DeMarais, 1991). The study of the evolution of Gila is important because it is an opportunity to understand the evolution of freshwater fish because it relates to the geological history of western North America. Evaluations of Gila

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species in USA and México suggest a current–day lack of understanding regarding the diversity of the genus (Schönhuth et al., 2014). This is substantiated by paraphyletic groupings obtained by Schönhuth et al. (2012, 2014) and recent records of undescribed populations of the genus in several drainages of central–north and northwest Mexico (DeMarais, 1991; Varela–Romero, 2001; Norris et al., 2003; Minckley and Marsh, 2009; Bogan et al., 2014; Schönhuth et al., 2014). The rapid development of molecular taxonomic and systematic methods in recent years has provided several tools to study biodiversity. Nowadays, in addition to molecular phylogenetic methods, the DNA barcoding technique (Hebert et al., 2003a, 2003b) has been applied as a molecular taxonomic tool to support species identifications and species discoveries (Hebert et al., 2003a, 2003b; Witt et al., 2006; Hubert et al., 2008; Rach et al., 2008; Lara et al., 2010; Li et al., 2011; Zou and Li, 2016; Yi et al., 2017). Current analytical methods to assign DNA barcodes to taxa can be divided into distance–based, phylogeny–based, and character–based approaches (Hebert et al., 2003a; Pons et al., 2006; Sarkar et al., 2008; Puillandre et al., 2012; Taylor and Harris, 2012). Although genetic distance–based methods for DNA barcoding have been considered useful tools in species discrimination and cryptic species discovery (Ward et al., 2005; Hubert et al., 2008; Lara et al., 2010; April et al., 2011; Lakra et al., 2015, Zou and Li, 2016), this approach has been questioned because of the relatively high rates of evolution of mitochondrial DNA between and within species (and between different groups of species) that can result in overlaps of intra– and interspecific distances, thus suggesting an uncertain existence of a barcoding gap for all species (Kipling and Rubinoff, 2004; Rubinoff, 2006; Rubinoff et al., 2006). The character–based analytical method known as CAOS (characteristic attribute organization system) has been used to define barcodes of taxa (Rach et al., 2008; Sarkar et al., 2008; Damm et al., 2010; Yassin et al., 2010; Li et al., 2011; Reid et al., 2011; Jörger and Schrödl, 2013; Yu et al., 2014; Zou and Li, 2016; Yi et al., 2017), and has been considered to better approximate a 'real' barcode, as well as providing better resolution to distinguish species than other approaches (Reid et al., 2011; Zou et al., 2011; Yu et al., 2014). Like traditional taxonomy, species identifications using CAOS DNA barcoding operate under the premise that members of a given taxonomic group share character attributes (i. e., putative diagnostic nucleotides) that are absent from other related groups (Rach et al., 2008; Sarkar et al., 2008; Bergmann et al., 2009; Jörger and Schrödl, 2013). Recent records of unstudied populations of Gila in northwestern México include two populations inhabiting a series of large spring–fed pools (pozas) of the Arroyo El Tigre sub–basin (Varela–Romero, 2001; Bogan et al., 2014). This sub–basin pertains to the Mátape River basin to the east and includes the adjacent low–elevation subtropical canyons La Balandrona and La Pirinola, both located in the southeastern sector of the Sierra

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La Pirinola Canyon

Guaymas

Ri pe ta

ra no er So Riv

Chihuahua

Má

yo ro re Ar Tig El

San Carlos

Sa n M Ri igu ve el r

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ve r

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na le da er g v a i M R

New Mexico

Bavispe River

Arroyo San Bernardino

Yaqui River

Arizona

5 km

C Gu al lf ifo o rn f ia

n Si

Fig. 1. Collection sites for Gila specimens analyzed. Open circles represent towns, and numbers in solid black circles are locations detailed in table 1. Hydrographic drainage divides are indicated by thick lines. Dashed lines indicate contemporary intermittent drainage courses. Dotted lines are state boundaries. Fig. 1. Sitios de captura de los especímenes de Gila analizados. Los círculos representan las ciudades. Los números en el interior de los círculos negros son las localidades que figuran en la tabla 1. Las divisiones de las cuencas hidrográficas se indican con líneas gruesas. Las líneas discontinuas indican los cursos de agua intermitentes contemporáneos. Las líneas punteadas son los límites estatales.

El Aguaje coastal mountain range, near the towns of San Carlos and Guaymas (fig. 1, inset). The proximate geographical location of these newly discovered populations suggest they are part of the Gila eremica lineage (Varela–Romero, 2001; Bogan et al., 2014) and thus are referred to herein as Gila cf. eremica. However, both populations are geographically isolated from populations of the lineage of G. eremica inhabiting the Mátape and Sonora River basins to the east and northeast. This isolation was probably promoted by volcanic events occurring in the area during the Miocene (Mora–Álvarez and McDowell, 2000), causing a geographical disconnection of the Arroyo El Tigre sub–basin from the Mátape River sub–basin. Recent morphological evaluations of the Gila eremica lineage revealed the G. cf. eremica populations as distinct compared to all G. eremica and other selected congeners analyzed, and showed at least 16 morpho–linear and two meristic characters that distinguish G. cf. eremica from other G. eremica populations (Ballesteros–Córdova et al., 2016). These mensural and meristic differences detected in the G. cf. eremica populations, as well as their isolated geographic occurrence, suggest a potential evolutionary isolation event within the G. eremica lineage (Ballesteros–Córdova et al., 2016), similar to that proposed for G. eremica and G. purpurea

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in Sonora (DeMarais, 1991; Schönhuth et al., 2014). The proposal of a Gila eremica lineage comprised of populations from the Sonora and Mátape River basins, plus the G. cf. eremica from the isolated Arroyo El Tigre sub–basin in La Balandrona and La Pirinola canyons, calls for development of molecular analyses to further investigate the evolutionary affinities of G. cf. eremica within the entire Gila lineage, and to potentially detect character attributes that may discriminate it from related groups. Knowledge of the evolutionary history of Gila can contribute to elucidating speciation mechanisms involved in this taxonomically problematic genus and other related fishes from arid and semiarid regions in North America. Also, the potential recognition of an evolutionary significant unit within Gila in México would enable the development of management strategies for its conservation. Material and methods Sample collection and DNA extraction A total of 178 specimens of five species of the genus Gila occurring in four river basins (seven sub–basins)

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Table 1. Taxa, sampling localities (drainage depicted in figure 1), molecular markers, accession numbers (No), voucher specimen catalog numbers (USON, Universidad de Sonora, Hermosillo, México; MNCN, Museo Nacional de Ciencias Naturales, Madrid, Spain; UAIC, University of Alabama, Tuscaloosa, Alabama, USA; BYU, Brigham Young University, Provo, Utah, USA), and source references for specimens of Gila spp. used for molecular analyses (S: 1, this study; 2, Schönhuth et al., 2014). Tabla 1. Taxones, localidades de muestreo (cuencas representadas en la figura 1), marcadores moleculares, números de accesión (No), números de catálogo de los especímenes de referencia (USON, Universidad de Sonora, Hermosillo, México; MNCN, Museo Nacional de Ciencias Naturales, Madrid, España; UAIC, Universidad de Alabama, Tuscaloosa, Alabama, EE.UU.; BYU, Universidad Brigham Young, Provo, Utah, EE.UU.) y referencia de los especímenes de Gila spp. utilizados para realizar los análisis moleculares (S: 1, este estudio; 2, Schönhuth et al., 2014).

Taxon

Locality

Gene

Voucher

G. eremica

1 - Bacanuchi River sub–basin,

cyt–b KX855966 USON–1301–1 1

Bacanuchi River at Tahuichopa ford,

cyt–b KX855967 USON–1301–3 1

Arizpe–Cananea road, Sonora

cyt–b KX855968 USON–1301–15 1

30º 21' 59.66'' N, 110º 09' 24.54'' W

cyt–b KX855969 USON–1301–17 1

Sonora River basin

cyt–b

nd2 KX858655 USON–1301–1 1

KX855970 USON–1301–25 1

nd2

cox1 KX858664 USON–1301–1 1

cox1 MH091955 USON–1301–2 1

cox1 MH091956 USON–1301–3 1

cox1 MH091957 USON–1301–4 1

cox1 MH091958 USON–1301–5 1

cox1 MH091959 USON–1301–6 1

cox1 MH091960 USON–1301–9 1

cox1 MH091961 USON–1301–10 1

cox1 MH091962 USON–1301–11 1

cox1 MH091963 USON–1301–12 1

cox1 MH091964 USON–1301–13 1

KX858656

USON–1301–15 1

G. eremica

2 - Sonora River sub–basin,

cyt–b KF514191 MNCN–279687 2

Sonora River at El Cahui

cyt–b KF514192 MNCN–279687 2

(1 km from La Labor), Sonora

cyt–b KF514189 MNCN–279686 2

29º 36' 23.65'' N, 110º 7' 29.57'' W

G. eremica

3 - San Miguel River sub–basin,

San Miguel River at Cucurpe, Sonora cyt–b

cyt–b KX855971 USON–1304–1 1 KX855972 USON–1304–6 1

30º 20' 26.59'' N, 118º 41' 37.09'' W nd2 KX858657 USON–1304–6 1

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cox1 MH091965 USON–1304–1 1

cox1 MH091966 USON–1304–2 1

cox1 MH091967 USON–1304–4 1

cox1 MH091968 USON–1304–5 1

cox1 KX858665 USON–1304–6 1

cox1 MH091969 USON–1304–7 1

cox1 MH091970 USON–1304–8 1

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Table 1. (Cont.)

Taxon

Locality

Gene

Voucher

cox1 MH091971 USON–1304–11 1

cox1 MH091972 USON–1304–13 1

cox1 MH091973 USON–1304–14 1

cox1 MH091974 USON–1304–15 1

cox1 MH091975 USON–1304–16 1

cyt–b JX443052 UAIC–15297 2

Mátape River basin G. eremica

4 - Mátape River sub–basin,

cyt–b KX855973 USON–1376–4 1

Mátape River at Mazatán, Sonora

cyt–b KX855974 USON–1376–9 1

29º 59' 56.04'' N, 100º 8' 51.25'' W

cyt–b KX855975 USON–1376–25 1

nd2 KX858658 USON–1376–4 1

nd2 KX858659 USON–1376–9 1

nd2

cox1 MH091976 USON–1376–1 1

cox1 MH091977 USON–1376–2 1

cox1 MH091978 USON–1376–3 1

cox1 MH091979 USON–1376–4 1

cox1 MH091980 USON–1376–5 1

cox1 MH091981 USON–1376–6 1

cox1 KX858666 USON–1376–9 1

cox1 MH091982 USON–1376–11 1

cox1 MH091983 USON–1376–12 1

cox1 MH091984 USON–1376–13 1

cox1 MH091985 USON–1376–14 1

cox1 MH091986 USON–1376–15 1

G. eremica

5 - Mátape River sub–basin,

Mátape River just W San José de Pimas

on Hwy 16, Sonora

28º 43' 7.82" N, 110º 20' 53.43'' W

KX858660 USON–1376–25 1

cyt–b KF514193 UAIC–15296 2

G. cf. eremica 6 - Arroyo El Tigre sub–basin,

cyt–b KX855976 USON–1300–17 1

La Balandrona Canyon,

cyt–b KX855977 USON–1300–18 1

Sierra El Aguaje mountains, Sonora

cyt–b KX855978 USON–1300–29 1

28º 2' 38.04'' N, 111º 4' 21.98'' W

nd2

KX858649 USON–1300–17 1

nd2

KX858650 USON–1300–18 1

nd2

KX858651 USON–1300–29 1

cox1 KX858667 USON–1300–17 1

cox1 MH091928 USON–1300–18 1

cox1 MH091929 USON–1300–19 1

cox1 MH091930 USON–1300–20 1

cox1 MH091931 USON–1300–21 1

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Table 1. (Cont.)

Taxon

Locality

Gene

Voucher

cox1 MH091932 USON–1300–22 1

cox1 MH091933 USON–1300–23 1

cox1 MH091934 USON–1300–24 1

cox1 MH091935 USON–1300–25 1

cox1 MH091936 USON–1300–26 1

cox1 MH091937 USON–1300–27 1

cox1 MH091938 USON–1300–28 1

cox1 MH091939 USON–1300–29 1

cox1 MH091940 USON–1300–30 1

cox1 MH091941 USON–1300–31 1

G. cf. eremica 7 Arroyo El Tigre sub–basin,

cyt–b KX855981 USON–1302–7 1

La Pirinola Canyon,

cyt–b KX855979 USON–1302–12 1

Sierra El Aguaje mountains, Sonora

cyt–b KX855980 USON–1302–13 1

28º 5' 32'' N, 111º 2' 15'' W

nd2 KX858654 USON–1302–7 1

nd2

KX858652 USON–1302–12 1

nd2

KX858653 USON–1302–13 1

cox1 MH091942 USON–1302–2 1

cox1 MH091943 USON–1302–3 1

cox1 MH091944 USON–1302–4 1

cox1 MH091945 USON–1302–5 1

cox1 MH091946 USON–1302–6 1

cox1 KX858668 USON–1302–7 1

cox1 MH091947 USON–1302–8 1

cox1 MH091948 USON–1302–9 1

cox1 MH091949 USON–1302–10 1

cox1 MH091950 USON–1302–11 1

cox1 MH091951 USON–1302–13 1

cox1 MH091952 USON–1302–14 1

cox1 MH091953 USON–1302–15 1

cox1 MH091954 USON–1302–16 1

Yaqui River basin

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G. purpurea

8 - Bavispe River sub–basin,

Arroyo San Bernardino at US/MX border nd2

31º 19' 57.37'' N, 109º 15' 35.17'' W

cyt–b JX443020 KX858661

BYU–14072 2 USON–1378–1 1

cox1 KX858669 USON–1378–1 1

cox1 MH091987 USON–1378–2 1

cox1 MH091988 USON–1378–3 1

cox1 MH091989 USON–1378–4 1

cox1 MH091990 USON–1378–5 1

cox1 MH091991 USON–1378–6 1

cox1 MH091992 USON–1378–7 1

cox1 MH091993 USON–1378–8 1

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Table 1. (Cont.)

Taxon G. minacae

Locality 9 - Bavispe River sub–basin,

Gene No Voucher S cyt–b KF514195 UAIC–14983 2

Arroyo El Largo, 2.5 km E Ejido

nd2 KX858663 USON–1224–1 1

Arroyo El Largo, Sonora

cox1 KX858671 USON–1224–1 1

29º 44' 3.9'' N,108º 36' 48.6'' W

De la Concepción River basin G. ditaenia

10 - Magdalena River sub–basin,

cyt–b JX443022 UAIC–15299 2

Magdalena River at road crossing

nd2

to San Ignacio–Terrenate, Sonora

cox1 KX858670 USON–1377–1 1

30º 41' 50.7'' N,110º 55' 39.5'' W

cox1 MH091994 USON–1377–2 1

USON–1377–1 1

cox1 MH091995 USON–1377–3 1

cox1 MH091996 USON–1377–4 1

cox1 MH091997 USON–1377–5 1

cox1 MH091998 USON–1377–6 1

cox1 MH091999 USON–1377–7 1

cox1 MH092000 USON–1377–8 1

cox1 MH092001 USON–1377–9 1

in Sonora, Mexico, were collected between April 2000 and November 2015 (fig. 1, table 1). Individuals of the G. eremica lineage included samples of G. cf. eremica collected from large spring–fed pools in the intermittent–flowing arroyos of the two subtropical canyons, La Balandrona (n = 30) and La Pirinola (n = 30), located in the southeastern sector of the Sierra El Aguaje mountains (fig. 1, inset). Specimens of G. eremica were collected from its known distribution in the Sonora (Sonora River sub–basin, n = 30; San Miguel River sub–basin, n = 30) and Mátape (Mátape River sub–basin, n = 30) river drainages. Samples of G. purpurea were collected from the Arroyo San Bernardino (n = 8) in the Bavispe River sub–basin of the extensive Yaqui River system. Specimens of G. ditaenia (n = 10) were collected from the Magdalena River sub–basin of the De la Concepción River basin, and specimens of G. minacae (n = 10) were obtained from Arroyo El Largo in the Bavispe River sub–basin, Yaqui River system (fig. 1, table 1). In the field, specimens for genetic analyses were labeled and tissue from a pelvic fin was removed and preserved in absolute ethanol and stored at 4 ºC until DNA extraction. After fin–clipping, the specimens were preserved in 10 % buffered formalin and later transferred to 50 % ethanol for deposition as vouchers in the Native Fish Collection of the Departamento de Investigaciones Científicas y Tecnológicas (DICTUS) of the Universidad de Sonora (USON) in Hermosillo.

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KX858662

cox1 MH092002 USON–1377–10 1

DNA extraction and PCR amplification Total DNA was obtained from fin tissue of all collected specimens following protocols of an extraction kit, the QIAamp DNA Mini Kit (QIAGEN). Amplification reactions were performed in a total volume of 50 µl using GoTaq Colorless Master Mix (Promega). The mitochondrial gene cyt–b was totally amplified (1140 bp) using the primers FW–L15058 5'–TGA CTT GAA AAM CCA CCG TTG–3' and RV: H16249 5'–TCA GTC TCC GGT TTA CAA GAC–3' as reported by Kocher et al. (1989). The total sequence (1047 bp) of mitochondrial gene nd2 was amplified with primers ND2F: 5'–AAC CCA TRC YCA AGA GAT CA–3' and ND2R: 5'–ACT TCT RCT TAR AGC TTT GAA GG–3', designed for other sequences of the genus as reported in GenBank. Conditions for the amplification of cyt–b and nd2 comprised an initial denaturation for 5 min at 94 ºC followed by 35 cycles of 50 s at 94 ºC, 50 s annealing temperature at 50 ºC for cyt–b and 60 ºC for nd2, and then 2 min at 72 ºC. The final extension was performed at 72 ºC for 7 min. A region of 651 bp of the mitochondrial gene cox1 was amplified with the primers FishF2_t1: 5'–TCT ACA AAY CAC AAA GAC ATT GGT AC–3' and FishR2_t1: 5'–ACC TCT GGG TGR CCA AAG AAT CAG AA–3', modified of Ivanova et al. (2007) to make them more specific to Gila. Conditions for amplification of cox1 comprised an initial denaturation for 5 min at 94 ºC followed by 34 cycles of 50 s at 94 ºC, 50 s at 50 ºC, and 1 min at

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72 ºC. The final extension was performed at 72 ºC for 7 min. The PCR products were sent to Macrogen, Inc., Seoul, South Korea for purification and bidirectional sequencing according to the company’s specifications. Phylogenetic inferences Obtained sequences were edited and assembled by overlapping using Chromas Pro 1.6 (Technelysium Pty Ltd, South Brisbane, Queensland, Australia). Each gene was identified using BLAST searches (Altschul et al., 1990) against GenBank data. Sequence divergence for the members of the G. eremica lineage, including G. cf. eremica and G. purpurea, was analysed using MEGA v5 (Tamura et al., 2011). Phylogenetic relationships of Gila cf. eremica from both La Balandrona and La Pirinola canyons were first evaluated via mitochondrial gene cyt–b, with sequences used by Schönhuth et al. (2014) for their phylogenetic inferences of the genus, including other members of their Revised Western Clade, plus our sequences of the G. eremica lineage. The evolutionary relationships of the two G. cf. eremica populations were corroborated using concatenated sequences of the mitochondrial genes cyt–b, nd2 and cox1 of all specimens obtained in this study plus sequences of congeners available from GenBank. Phylogenetic trees using Maximum Likelihood (ML) for cyt–b and the concatenated mitochondrial gene dataset (cyt–b, nd2 and cox1) were estimated using RAxML–HPC2 on XSEDE 8.0.24 (Stamatakis, 2014). The JModeltest2 software (Darriba et al., 2012) was used to find the best nucleotide substitution models for each dataset separately. We defined data blocks based on genes, and used the Akaike information Criterion (Posada and Buckley, 2004), and the best–fit model was used for the subsequent analyses. ML trees were performed on the CIPRES Science Gateway 3.3 (Miller et al., 2010), using GTRGAMMA model, and 1,000 bootstraps pseudoreplications (Felsenstein, 1985) to estimate the node reliability. Bayesian inference (BI) analyses were conducted for each gene data set using MrBayes v3.2.6 (Ronquist and Huelsenbeck, 2003). The Akaike information criterion (AIC) implemented in JModelTest2 (Posada and Buckley, 2004; Darriba et al., 2012) was used to identify the optimal molecular evolutionary model for each partition block on each sequence data set of the analysis. For BI, ten million cycles were implemented in four simultaneous Monte Carlo Markov chains; sampling the Markov chain at intervals of 1,000 generations. Log–likelihood stability was attained after 100,000 generations; the first 1,000 trees were discarded as burn–in in each analysis. The remaining trees were used to compute a 25 % majority rule consensus tree in PAUP* (Swoford, 2002). Support for BI tree nodes was determined based on values of Bayesian posterior probabilities. Final trees of ML and BI were edited using FigTree 1.4.2 (Rambaut, 2014). DNA barcode and cox1 sequence analysis DNA barcode analyses included samples of nominal Gila eremica populations from the Sonora and Mátape

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River basins, samples of G. cf. eremica from the Arroyo El Tigre sub–basin’s La Balandrona and La Pirinola canyons, G. purpurea from Arroyo San Bernardino, and G. ditaenia as outgroup. Cox1 sequences were edited and assembled by overlapping using Chromas Pro v1.6 (Technelysium Pty Ltd, South Brisbane, Queensland, Australia). The values of haplotype (H), nucleotide diversity (π) (Nei, 1987), and the polymorphic/variable sites were estimated with DnaSP v5.0 (Librado and Rozas, 2009). Genetic uncorrected p distance analysis between groups was performed using MEGA v5 (Tamura et al., 2011). A neighbor–joining (NJ) 50 % majority–rule consensus tree was constructed in PAUP* (Swoford, 2002) over 1,000,000 bootstrap replicates, using K2P (Kimura, 1980). The obtained tree was incorporated into the NEXUS file of the DNA data matrix of Gila species using Mesquite v3.10 (Maddison and Maddison, 2016), according to the specifications of Jörger and Schrödl (2014). The NEXUS/TREE file was carried out in CAOS software package (Sarkar et al., 2008) to identify diagnostic single pure characters (sPu) which are present in all members of a clade but absent from all members of another clade. These sPu characters were used for taxa discrimination of our groups of interest. Results Phylogenetic relationships within the Gila eremica lineage Amplified sequences of the mitochondrial gene cyt–b for all specimens of the G. eremica lineage analyzed had a length of 1,140 bp with no insertions or deletions. Thirty–four positions were variable sites, 15 of which were parsimony informative. Individuals of the G. eremica lineage using mitochondrial gene cyt–b were classified into 19 haplotypes. Haplotypes 8 and 10 were shared between individuals from the Sonora and San Miguel Rivers sub–basins (Sonora River drainage), and none of the remaining haplotypes were shared with any other population. Sequence divergence (uncorrected p distance) within the G. eremica lineage ranged between 0.26 % and 1.11 % with an average of 0.7 %. Genetic divergence of G. eremica populations from the Sonora, San Miguel and Mátape Rivers sub– basins compared to G. cf. eremica from La Balandrona and La Pirinola canyons (Arroyo El Tigre sub–basin) was 0.44–1.14 %. The mean distance between all individuals of G. eremica and G. cf. eremica was 0.83 %. The genetic divergence for individuals of G. eremica (but not including G. cf. eremica) against G. purpurea was 1.93–2.72 %. Genetic divergence between G. cf. eremica and G. purpurea ranged from 2.11–2.46 %. Mean nucleotide frequencies within nominal G. eremica populations were 26.30 % A, 29.57 % T, 27.48 % C, and 16.65 % G. Mean nucleotide frequencies for G. cf. eremica populations were 26.58 % A, 29.59 % T, 27.56 % C, 16.37 % G. The estimated Transition/Transversion bias (R) for both groups was 2.27. Substitution pattern and rates were estimated under the General Time Reversible model (GTR) (Nei and Kumar, 2000).

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Phylogenetic analyses of cyt–b using ML and BI with all species analyzed of the Revised Western Clade of Schönhuth et al. (2014), and including our samples of G. eremica and G. cf. eremica, showed the same topology with variations in nodal support of bootstrap probabilities (BP) for ML and posterior probabilities (PP) for BI (fig. 2). The tree topology was consistent with that obtained by Schönhuth et al. (2014) for members of the Gila lineage (excluding G. cf. eremica). Members of the G. eremica lineage, including G. cf. eremica of the present study, were always resolved as monophyletic, with G. purpurea as the sister species (BP = 98 %, PP = 100 %) (fig. 2). The G. eremica lineage was resolved with a clade for the Sonora River sub–basins (Sonora and San Miguel Rivers), and a sister clade for the Mátape River sub–basins (Mátape River and Arroyo El Tigre) (fig. 2). Individuals of G. cf. eremica from La Balandrona and La Pirinola canyons of the Arroyo El Tigre sub–basin were nested together sharing a putative common ancestor, corroborating these two isolated populations as unequivocal members of the G. eremica lineage (fig. 2). Phylogenetic analyses by ML and BI, using the concatenated results of mitochondrial genes cyt–b, nd2, and cox1, included Gila robusta, G. ditaenia, G. purpurea, members of the G. eremica lineage (including G. cf. eremica), with G. minacae as outgroup. The tree topology resulting from the analyses was the same for both criteria, with variations in the nodal support values of BP for ML and PP for BI (fig. 3). The analyses showed that members of the G. eremica lineage are monophyletic (BP = 100 %, PP = 100 %) with G. purpurea as sister species (BP = 100 %, PP = 100 %), and corresponded with our results from cyt–b in regards to monophyly within the G. eremica lineage (figs. 2, 3). The geographical clade for all members of the Mátape River basin, including those from the two isolated Arroyo El Tigre canyons, was better supported in the concatenated genes' analyses (fig. 3, BP = 76 %, PP = 95 %) compared to using cyt–b alone (fig. 2). Individuals of G. cf. eremica from both canyons in the Arroyo El Tigre sub–basin were supported in a clade of specific identity (fig. 3, BP = 94 %, PP = 100 %) and indicating relationship with G. eremica from the Mátape River sub–basin as putative closest relative (fig. 3). DNA barcoding analysis of the Gila eremica lineage The analyses of 82 cox1 sequences of several Gila species showed a total of nine haplotypes: five for G. eremica, one for G. cf. eremica, one for G. purpurea, and two for G. ditaenia. The G. eremica lineage, including G. cf. eremica, showed a haplotype and nucleotide diversity of 0.6093 (SD = 0.034) and 0.00125 (SD = 0.00112), respectively. Individuals of G. eremica from the Sonora, San Miguel and Mátape Rivers sub–basins shared at least one haplotype, and also showed unique haplotypes for the Sonora and San Miguel rivers sub–basins. The analysis did not detect shared haplotypes among G. cf. eremica and all the G. eremica populations. The sequences of taxa

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included in the DNA barcoding analyses showed 46 polymorphic sites (table 2). The genetic uncorrected p distances analysis between groups produced a value of 1.27 % for populations of G. eremica and G. purpurea. The distance value between G. cf. eremica and G. purpurea was 1.38 %, and the value between nominal G. eremica and G. cf. eremica was 0.20 %. The DNA barcoding analysis using the character– based approach with CAOS showed eight single pure characters (sPu) to discriminate G. eremica from G. purpurea, nine sPu to discriminate G. cf. eremica from G. purpurea, and one fixed sPu in the 29 analyzed sequences of G. cf. eremica, discriminating it from G. eremica (table 2). Discussion The monophyly of the G. eremica lineage (Sonora and Mátape River basins) and G. purpurea (Bavispe River in northern headwaters of the Yaqui River basin) obtained in the present study was highly supported by both ML and BI criteria, as previously suggested by morphological analyses by DeMarais (1991) and molecular data by Schönhuth et al. (2014). Phylogenetic relationships inferred here for both ML and BI using the cyt–b gene alone and the concatenated set of genes cyt–b, nd2 and cox1, support with high values of posterior probabilities, the monophyly of the G. eremica lineage for all its members, and corroborated G. cf. eremica as a member of the lineage (figs. 2, 3). The monophyly, geographical clades, and low genetic divergence detected here within the G. eremica lineage may be explained by relatively recent isolation of once–connected drainages inhabited by this lineage, as suggested for other nominal species of Gila occurring in México (Schönhuth et al., 2014). In addition, we provide evidence for the existence of two geographical clades for the Sonora and Mátape river basins, with high scores for both ML and BI criteria (figs. 2, 3). The close relationship and low genetic divergence between G. eremica from the Mátape River and G. cf. eremica from Arroyo El Tigre sub–basin (figs. 2, 3) suggest a putative common ancestor for these populations and indicate a relatively recent connection between the two sub–basins. The phylogenetic analysis also supports G. cf. eremica as a clade of specific identity, apart from other members of the lineage, but closely related to populations of G. eremica in the Mátape River sub–basin. The morphological differences detected between G. cf. eremica and G. eremica (Ballesteros–Córdova et al., 2016), along with its phylogenetic position resolved in the present study, suggest that G. cf. eremica is an evolutionary significant unit within the G. eremica lineage that requires additional species delimitation methods such as DNA barcoding for further discrimination. The effectiveness of character–based methods (e. g., CAOS) for taxa detection relies on the use of diagnostic characters, as those used in traditional taxonomy. The character–based method is based on the premise that members of a given taxon share a distinctive combination or combinations of diagnostic

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Revised Western Clade

83/100

100/100 G. pulchra G. brevicauda 92/100 81/100 G modesta - G pandora G. nigrescens 75/100 Gila sp. G. conspersa G intermedia Chihuahuan G. nigra Desert G. nigra G. robusta Group 99/100 G. robusta G. robusta G. robusta G. nigra Moapa coriacea Moapa coriacea G. elegans G. seminuda G. elegans 100/100 G. elegans G. seminuda G. seminuda G. ditaenia G. ditaenia G. eremica (SS) G. eremica (SMS) 100/100 G. eremica (SS) 100/100 G. eremica (SS) G. eremica (SMS) G. eremica (SS-SMS) G. eremica (SS) G. eremica (SS) G. eremica lineage G. eremica (SS) G. eremica (SS) G. eremica (MS) -/100 98/100 G. eremica (MS) G. eremica (MS) G. cf. eremica* (LB) G. cf. eremica* (LB) 68/75 G. cf. eremica* (LB) G. cf. eremica* (LP) G. cf. eremica* (LP) 68/99 G. cf. eremica* (LP) G. purpurea G. purpurea G. cypha G. cypha G. cypha G. cypha 100/100G. cypha G. robusta G. cypha G. cypha G. atraria 100/100 G. atraria G. atraria G. orcutti 75/100 G. orcutti Gila lineage Ptychocheilus lucius Ptychocheilus lucius Acrocheilus alutaceus Acrocheilus alutaceus G. minacae G. minacae G. minacae 100/100 G. minacae G. minacae -/90 G. coerulea G. coerulea Relictus solitarius Ptychocheilus oregonensis Ptychocheilus oregonensis 100/100 Ptychocheilus umpquae Ptychocheilus umpquae Siphateles alvordensis 100/100 Siphateles boraxobius Mylopharodon conocephalus Myolopharodon conocephalus Lavinia exilicauda Hesperoleucus symmetricus 100/100 100/100 Orthodon microlepidotus Siphateles bicolor Siphateles bicolor 100/100 Siphateles b. mohavensis Ptychocheilus grandis Ptychocheilus grnadis Eremichthys acros Lepidomeda Agosia chrysogaster [Snydericthys] copei Algansea lacustris Platygobio gracilis 0.06

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G. eremica (SS) G. eremica (SMS) 99/100

71/100 100/100 80/100

G. G. G. G.

eremica eremica eremica eremica

G. G. G. G. G. G.

cf. cf. cf. cf. cf. cf.

(SS) (MS) (MS) (MS)

eremica* eremica* eremica* eremica* eremica* eremica*

(LB) (LB) (LB) (LP) (LP) (LP)

G. purpurea G. purpurea G. robusta

G. ditaenia G. ditaenis G. minacae

0.008

Fig. 3. Recovered phylogenetic tree via maximum likelihood and posterior probabilities of Bayesian inference of the concatenated genes cyt–b, cox1, and nd2 for populations of the Gila eremica lineage plus other three species examined in this study (SS, Sonora River sub-basin; SMS, San Miguel River sub-basin; MS, Mátape River sub-basin; LB, La Balandrona Canyon; LP, La Pirinola Canyon. Numbers on branches represent values of ML bootstap probabilities (BP > 65 %)/BI posterior probabilities (PP > 70 %). (* specimens from Arroyo El Tigre sub–basin). Fig. 3. Árbol filogenético recuperado mediante el método de la máxima verosimilitud y probabilidades a posteriori de inferencia bayesiana de los genes concatenados cyt–b, cox1 y nd2 para las poblaciones del linaje de Gila eremica y otras tres especies examinadas en este estudio (SS, subcuenca del río Sonora; SMS, subcuenca del río San Miguel; MS, subcuenca del río Mátape; LB, cañón La Balandrona; LP, cañón La Pirinola). Los números sobre las ramas representan las probabilidades de remuestreo de ML (BP > 65 %)/ probabilidades a posteriori BI (PP > 70 %). (* especímenes de la subcuenca del arroyo El Tigre).

character attributes (e. g., polymorphisms) that are absent in related groups, and these attributes can be used for species discrimination (Rach et al., 2008; Sarkar et al., 2008; Bergmann et al., 2009; Zou et al., 2011; Jörger and Schrödl, 2013; Zou and Li,

2016). Despite the proposal to use more than three characters as a DNA barcoding gap to separate natural taxonomic groupings (Rach et al., 2008, Yassin et al., 2010; Zou et al., 2011; Yu et al., 2014), Jörger and Schrödl (2013) argue that CAOS does not

Fig. 2. Recovered tree of phylogenetic relationships via Maximum Likelihood (GTR + G + I model) for haplotypes of all members of the Gila lineage, including other species of the Revised Western Clade (Schönhuth et al., 2012, 2014), using mitochondrial gene cyt–b. Numbers on branches represent ML bootstap probabilities (BP > 65 %)/BI posterior probabilities (PP > 70 %): SS, Sonora River sub-basin; SMS, San Miguel River sub-basin; MS, Mátape River sub-basin; LB, La Balandrona Canyon; LP, La Pirinola Canyon (* specimens from Arroyo El Tigre sub–basin). Fig. 2. Árbol recuperado de relaciones filogenéticas mediante el método de la máxima verosimilitud (modelo GTR + G + I) de los haplotipos de todos los miembros del linaje de Gila, incluidas otras especies del Revised Western Clade (Schönhuth et al., 2012, 2014), utilizando el gen mitocondrial cyt–b. Los números sobre las ramas representan la probabilidad de remuestreo de ML (BP > 65 %)/probabilidad a posteriori BI (PP > 70 %): SS, subcuenca del río Sonora; SMS, subcuenca del río San Miguel; MS, subcuenca del río Mátape; LB, cañón La Balandrona; LP, cañón La Pirinola (* especímenes de la subcuenca del arroyo El Tigre)

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Table 2. Polymorphic sites and composite of attribute characters obtained by CAOS in the first 651 bp of the mitochondrial gene cox1 for G. cf. eremica compared to nominal G. eremica populations, G. purpurea and G. ditaenia. Numbers at the top indicate variable sites of the fragment studied in this study. The far right column shows the number of individuals sharing each haplotype. Pure diagnostic characters among G. cf. eremica, G. eremica populations and G. purpurea are in bold. Single pure characters between G. cf. eremica and G. eremica are shaded and in bold: LB, La Balandrona Canyon; LP, La Pirinola Canyon; S, Sonora River; SM, San Miguel River; M, Mátape River.

0 0 0 0 0 0 0 0 0 1 1 1 1 1 1 1 2 2 2 2 3

0 2 2 3 3 4 4 8 8 2 4 6 7 7 7 9 9 3 6 7 2

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

G. cf. eremica (LB)

C G T G G A T C A T G G C T G T A C C G G

G. cf. eremica (LP)

C G T G G A T C A T G G C T G T A C C G G

G. eremica (S, SM, M) – – – – – – – – – – – – – – – – – – – – – G. eremica (SR, SMR) – – – – – – – – – – – – – – – – – – – – – G. eremica (SR)

– C C – – – – – – – – – – – – – G – – – –

G. eremica (SR)

– – – – – – – – – – – – G – – – – – – – A

G. eremica (SMR)

– – – – – – – – – – – – – – – C – – – – –

G. purpurea

T – – – – – C – – – A – – C – – – – – – –

G. ditaenia

T – – A A G – T G C A A – – C C – T T A A

G. ditaenia

T – – – A G – T G C A A – – C C – T T A A

possess an objective criterion with which to delimit a threshold number of distinguishing nucleotides that would indicate a species boundary (i. e., to delimit a probable new species) or an independent population belonging to the same species. On the other hand, the supposition that a characteristic attribute has been fixed within a population gains more confidence if a higher number of samples is analyzed (Rach et al., 2008). According to Zhang et al. (2010), the desired sample size for a DNA barcoding analysis should range from 9.5 to 216.6, which is within the range of samples of G. cf. eremica and G. eremica analyzed here. Our analysis revealed a single fixed polymorphism in all 29 examined sequences of G. cf. eremica that was absent from all analyzed specimens of the G. eremica populations. The unique polymorphism detected here in G. cf. eremica with respect to G. eremica represents a sPu character and thus reveals an apomorphy within the G. eremica lineage. This indicates a different pattern of genetic variation for G. cf. eremica compared with G. eremica and also with the close congener G. purpurea. The study by Rach et al. (2008) identified less than three diagnostic characters in ten very closed related sister taxa of the insect order Odonata using mitochondrial gene ndh1, showing similar results to those obtained here between populations of G. cf. eremica and G. eremica. The close relationship and low genetic divergence of the two G. cf. eremica populations with respect to

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G. eremica obtained in our several molecular analyses bolsters the morphological differences previously detected (Ballesteros–Córdova et al., 2016). Such differences may reflect a general phenotypic plasticity of freshwater fishes to adapt to environmental alterations, or variation in stream size, flow and substrate (Hubbs, 1940), leading in our case to a morphological variant within the Gila eremica lineage. However, the morphological differences seen in the G. eremica lineage members (Ballesteros–Córdova et al., 2016) are consistent with the establishment of a fixed polymorphism in the geographically isolated G. cf. eremica. Moreover, none of our phylogenetic analyses nested individuals of G. cf. eremica with samples of G. eremica, thus revealing this potentially ESU as a natural group. Similarly, the character–based method using cox1, and sequences analyses of cyt–b and its concatenation with nd2 and cox1 showed a different pattern of variation in G. cf. eremica compared with G. eremica, and with G. purpurea. However, current results will need to be further tested using nuclear data. The process of species identification through DNA barcoding has often been confused with species discovery (DeSalle, 2006). Species identification has been considered a valid use for the DNA barcode, which does not rely on any particular species concept (Rach et al., 2008). This appears to be because species identification using DNA barcode is consistent with any concept of species a taxonomist may

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Tabla 2. Sitios polimórficos y composición de los atributos de carácter obtenidos con CAOS en las primeras 651 pb del gen mitocondrial cox1 de G. cf. eremica comparados con las poblaciones de G. eremica, G. purpurea y G. ditaenia. Los números de la parte superior indican los sitios variables del fragmento estudiado. La columna de la derecha muestra el número de individuos que comparten cada haplotipo. Los caracteres diagnósticos puros entre las poblaciones de G. cf. eremica, G. eremica y G. purpurea están en negrita. Los caracteres únicos puros entre G. cf. eremica y G. eremica están sombreados, en negrita: LB, cañón La Balandrona; LP, cañón La Pirinola; S, río Sonora; SM, río San Miguel; M, río Mátape. 3 3 3 3 3 3 3 3 4 4 4 5 5 5 5 5 5 5 6 6 6 6 6 6 6 5 5 6 7 7 8 8 9 5 6 6 0 1 1 4 4 6 8 0 1 1 1 2 3 4 2 5 7 6 9 2 3 7 7 0 3 5 4 7 1 4 2 9 1 3 6 9 8 4 3 C C G G A T C G T T A G G C A G T A C T A A A C C 15 C C G G A T C G T T A G G C A G T A C T A A A C C 14 – T – – – – – – – – – A – – – – – – – – – – – – – 27 – – – – – – – – – – – A – – – – – – – – – – – – – 4 – – – – – – – – – – – A – – – – – – – – – – – – – 1 – – – – – – – – – – – A – – – – – – – – – – – – – 1 – – – – – – – – – – – A – – – – – – – – – – – – – 1 – – – – G – – A – – – A – – – – – – T – – – G – – 8 T – A A – C – A C C G – A T G A C G – C C G – T T 1 T – A A – C – A C C G – A T G A C G – C C G – T T 9

use when identifying a named species (Rach et al., 2008). The CAOS approach in DNA barcoding is considered by many as an accurate and available method for testing species boundaries. However, this method requires a priori defined groups, making it difficult or unsuitable for species discovery (Jörger and Schrödl, 2013; Kekkonen et al., 2015). Species discovery involves a more complicated task because it requires a recognized species concept along with a traditional taxonomic corroboration system (DeSalle et al., 2005; DeSalle, 2006). Accordingly, a single source of data, be it molecular, morphological, ecological or ethological, is not able by and of itself to be used for species discovery (Rach et al., 2008). However, comparison of DNA sequences can be used to detect potentially new species which then need corroboration by an integrated taxonomic approach using a species concept (Rubinoff, 2006; Rach et al., 2008). The phylogenetic data and DNA barcoding results obtained here, coupled with those from the morphological analyses for G. cf. eremica (Ballesteros– Córdova et al., 2016) and its geographic isolation supports it as a natural evolutionary significant unit within G. eremica. The low genetic distance detected between G. cf. eremica and nominal G. eremica, along with their phylogenetic affinities indicates a relatively recent disruption within this lineage. Our data thus contribute to the knowledge of systematics and evolution of the greater Gila lineage. The

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recognition of an isolated taxon of Gila in the Arroyo El Tigre sub–basin of the Sierra El Aguaje reveals a potential microendemism for the genus in subtropical canyons of this region of Northwest Mexico. The evidence presented here calls for further studies aimed at clarifying the biology, origin and history of G. cf. eremica populations and contributes to increased understanding of the evolution and conservation of fish species inhabiting arid and semiarid regions in Mexico and the USA. Acknowledgements Many thanks to Ramón Villafaña and José Ines– Ramírez for facilitating access to sampling sites for Gila cf. eremica. We also thank Michael T. Bogan, Nohemí Noriega–Félix, Sylvette L. Vidal–Aguilar, Celso Haros–Méndez, Emmanuel M. Bernal–Loaiza, and Dylann Córdova–Martínez for assistance in sampling. Dr. José Said Gutiérrez–Ortega provided advice in the phylogenetical analyses. Field collections were made under Mexican government permits DGOPA.03947.250406.1606 and SGPA/ DGVS/00505/10. The ﬁrst author received a fellowship from the Consejo Nacional de Ciencia y Tecnología (CONACyT) for doctoral studies. This work was funded primarily by CONACyT. Additional thanks to the Desert Fishes Council for partial funding for gene sequencing.

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Yu, Z., Li, Q., Kong, L., Yu, H., 2014. Utility of DNA barcoding for Tellinoidea: a comparison of distance, coalescent and character–based methods on multiple genes. Marine Biotechnology, 17(1): 55–65. Zhang, A. B., He, L. J., Crozier, R. H., Muster, C., Zhu, C. D., 2010. Estimating sample sizes for DNA barcoding. Molecular Phylogenetics and Evolution, 54(3): 1035–1039. Zou, S., Li, Q., 2016. Pay attention to the overlooked cryptic diversity in existing barcoding data: the case of Mollusca with character–based DNA barcoding. Marine Biotechnology, 18: 1–9. Zou, S., Li, Q., Kong, L., Yu, H., Zheng, X., 2011. Comparing the usefulness of distance, monophyly and character–based DNA barcoding methods in species identification: a case study of Neogastropoda. Plos One, 6: e26619.

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Assessing best practice for selecting surrogates and target–setting methods in a megadiverse country T. Urquiza–Haas, W. Tobón, M. Kolb, A. Lira–Noriega, V. Contreras, J. Alarcón, P. Koleff Urquiza–Haas, T., Tobón, W., Kolb, M., Lira–Noriega, A., Contreras, V., Alarcón, J., Koleff, P., 2019. Assessing best practice for selecting surrogates and target–setting methods in a megadiverse country. Animal Biodiversity and Conservation, 42.1: 187–202, Doi: https://doi.org/10.32800/abc.2019.42.0187. Abstract Assessing best practice for selecting surrogates and target–setting methods in a megadiverse country. Systematic conservation planning provides a framework to identify representative areas of biodiversity, but its effectiveness depends on the choice of surrogates and targets. Mexico has conducted participatory and comprehensive gap analyses. We present the results of two independent surrogate assessments to test the criteria used in Mexico's spatial conservation prioritization. We tested the surrogate efficiency of range restricted, endemic, and threatened mammals and the influence of target–setting on the spatial configuration of the conservation network, as well as the performance of taxonomic–based surrogates. Results show that target–setting heavily influences the spatial configuration and irreplaceability values of the conservation area network. Representation effectiveness and coverage of species distribution was sensitive to surrogate selection but not to target–setting. Threatened and rare species were poorly represented when other surrogate species were used, while threatened mammals represented 90 % of all species. The effectiveness of networks designed for a single vertebrate taxon varied greatly; reptiles and amphibians performed better than random achieving high species representation. Key words: Systematic conservation planning, Surrogate species, Target setting, Endemic species, Threatened species, Megadiverse country Resumen Evaluación de las mejores prácticas para seleccionar sustitutos y métodos para establecer metas de conservación en un país megadiverso. La planificación sistemática de la conservación proporciona un marco para identificar áreas representativas de la biodiversidad, pero su eficacia depende de la elección de sustitutos y de las metas de conservación. México ha realizado análisis integrales y participativos de vacíos y omisiones en conservación. En este trabajo presentamos los resultados de dos evaluaciones independientes sobre sustitutos de la biodiversidad a fin de poner a prueba los criterios utilizados en México para identificar las prioridades espaciales de conservación. Se probó el efecto sombrilla de los mamíferos de distribución restringida, endémicos y amenazados, y la influencia del establecimiento de las metas en la configuración espacial de la red de áreas de conservación, así como el desempeño de sustitutos taxonómicos. Los resultados muestran que el establecimiento de metas influye mucho en la configuración espacial y en los valores de irremplazabilidad de la red. La eficacia de la representación y la cobertura de las áreas de distribución de las especies variaron con la selección de los sustitutos, pero no con el establecimiento de las metas. Las especies amenazadas y las raras no estuvieron suficientemente representadas cuando se usaron otras especies como sustitutos, mientras que los mamíferos amenazados representaron el 90 % de todas las especies. La eficacia de las redes diseñadas para un solo taxón de vertebrados varió mucho; los reptiles y anfibios obtuvieron mejores resultados que los obtenidos al azar, y lograron una alta representación de otros grupos taxonómicos en la red diseñada para dichos taxones. Palabras clave: Planificación sistemática de la conservación, Especies sustitutas, Metas de conservación, Especies endémicas, Especies amenazadas, País megadiverso

ISSN: 1578–665 X eISSN: 2014–928 X

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Received: 20 IV 17; Conditional acceptance: 04 IX 17; Final acceptance: 16 X 18 Tania Urquiza–Haas, Wolke Tobón, Victoria Contreras, Jesús Alarcón, Patricia Koleff, Comisión Nacional para el Conocimiento y Uso de la Biodiversidad (CONABIO), Liga Periférico–Insurgentes Sur 4903, Parques del Pedregal, Tlalpan, 14010 Ciudad de México, México.– Melanie Kolb, Departamento de Geografía Física, Instituto de Geografía, Universidad Nacional Autónoma de México, Circuito exterior s/n., Ciudad Universitaria, Coyoacán 04510, Ciudad de México, México.– Andrés Lira–Noriega, Catedrático CONACyT, Instituto de Ecología A.C., Red de Estudios Moleculares Avanzados, Carretera Antigua a Coatepec 351, El Haya, 91070 Xalapa, Veracruz, México. Corresponding author: Tania Urquiza–Haas. E–mail: turquiza@conabio.gob.mx

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Introduction Conserving biological diversity in the face of an irrefutable environmental crisis remains a serious challenge as governments and societies have largely failed to keep up with the anthropogenic pace of change (Stafford–Smith et al., 2012). To minimize biodiversity loss, there is a central and urgent need to concentrate the scarce conservation resources and efforts on effective conservation area systems in regions of high biological value (Chape et al., 2005; Rands et al., 2010). Assuring adequate representation and long–term maintenance of biodiversity lies at the heart of an effective protected area (PA) network (Gaston et al., 2008). Systematic conservation planning (SCP) provides a framework to select complementary conservation areas that represent the biodiversity of the planning region (Sarkar et al., 2006). Core to the process of spatial conservation prioritizaition within the SCP framework is the selection of biological and environmental data to represent biodiversity and the treatment of socioeconomic data to consider budgetary and sociopolitical constraints for maximizing implementation efficiency (Sarkar et al., 2006; Kukkala and Moilanen, 2013). Practitioners therefore face difficult decisions when determining key aspects of their conservation plan. They have to decide, first, which datasets are sufficiently reliable to serve as surrogates for biodiversity, second, how to assign conservation targets for species, vegetation types or other features used as surrogates; and third, (3) how to avoid selecting unsuitable areas for conservation action. The field of SCP has been useful in advancing concepts and designing reserve selection tools. One of the objectives of peer–reviewed studies has been to test biological data and establish limitations (Knight and Cowling, 2007). Nonetheless, conservation planning must often be conducted in the absence of comprehensive biodiversity datasets, and the adequacy of results and decisions in real–world circumstances has rarely been tested. These challenges highlight the need to test the choice of biodiversity data, and validate their robustness using various analytical approaches in order to promote an efficient network of conservation areas. More than ten years ago, the Programme of Work on Protected Areas of the Convention on Biological Diversity (CBD) encouraged the Parties to increase representativeness and coverage of biodiversity within national PA systems. Mexico was one of the first countries to conduct comprehensive conservation gap analyses for the terrestrial, marine and freshwater environments (Koleff et al., 2009; Lira–Noriega et al., 2015). The process involved over 260 experts from numerous academic and research institutions, civil society organizations, and governmental agencies. The purpose was to conduct a spatial conservation prioritization to assess the effectiveness of PAs to adequately represent Mexico’s biodiversity, and to guide the implementation of area–based conservation measures. The aim of this paper was, first, to test the criteria used in Mexico’s spatial conservation prioritization to

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select biodiversity surrogates and set conservation targets through two independent surrogate assessments, and second, to frame our findings as lessons learned in the context of a megadiverse country. In the first surrogate assessment we explored questions regarding the appropriateness of using range restricted (hereafter referred as rare species), endemic and threatened species as surrogates for other species not previously considered. We also examined the influence of target–setting on the spatial configuration of the conservation network. In the second surrogate assessment we analyzed the performance of selected taxonomic groups as surrogates for other known species groups. We first provide a brief overview of the core methodological decisions of Mexico’s spatial conservation prioritization analyses in terrestrial environments (hereafter, gap analysis) and the overall results (described in detail in Urquiza–Haas et al., 2009). We then present and update the results of the independent surrogate assessments based on the methods of gap analysis (Koleff et al., 2011). Finally, in view of these results and the relevant and recent SCP literature we discuss whether methodological decisions were appropriate to ensure a conservation network representative of Mexico’s megadiversity. The lessons learned from the surrogate assessments in the context of the Mexican gap analysis project may be useful to guide decision makers, planners and managers in other countries in the selection of conservation targets and surrogates,. This is of foremost relevance in megadiverse developing countries that need to develop a clear spatial guide towards meeting Aichi target 11 for effectively conserving 17 % of terrestrial and inland water areas of particular importance to biodiversity, especially when spatial patterns of biodiversity are complex and represent major challenges to fulfill criteria expressed theoretically in SCP. SCP for biodiversity conservation in Mexico The most common obstacle for conducting high–resolution systematic conservation assessments is having limited data or access to data on species distributions (Kremen et al., 2008). Commission and omission errors inherent to species occurrence data can affect the comprehensiveness, representativeness, efficiency and adequacy of reserve networks in different ways (Rondinini et al., 2006). The challenge of acquiring good quality information on species distributions for the spatial conservation prioritization was overcome in Mexico with the collaboration of government agencies and numerous researchers (see Koleff et al., 2009). Species distribution modeling (hereafter, SDM) with explicit considerations on the use of reliable taxonomic determination, precise georeferencing (Soberón and Peterson, 2004) and a post–processing step after modelling that minimizes commission errors was considered the best tool to deal with scarce and biased occurrence data (despite efforts of the National Biodiversity Information System, SNIB, and worldwide of the Global Biodiversity Information Facility, GBIF). The National Commission for the Knowledge and Use

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of Biodiversity (CONABIO) commissioned expert– lead–technical groups (see references and methods in Koleff et al., 2009) to generate SDMs using niche modelling techniques from species records reviewed and curated by experts belonging mainly to the SNIB and available through GBIF. This effort corresponded to about 2,412,000 records to generate distribution maps for 2,408 species, including a broad span of taxonomic groups, mainly vascular plants and terrestrial vertebrates. The SDMs database for vertebrates was by far the most complete as it represented 86% of the vertebrate species in Mexico. The use of several types of surrogates representing different levels of biological organization and targets assigned to each based on biological knowledge and socio–ecological context are considered best practice in SCP (Groves et al., 2002; Carwardine et al., 2009; Polak et al., 2015). To accomplish this, a working group consisting of members from CONABIO, the National Protected Areas Commission (CONANP), several national and international NGOs and academic institutions discussed the criteria for conducting a spatial conservation prioritization within the SCP framework in five workshops that took place during 2005 and 2006. The final biodiversity dataset comprised 1,450 plant and vertebrate SDMs, 68 vegetation type maps, nine species richness maps (overall and endemic) and 12 ad hoc richness and endemism indices to represent flowering plant diversity (in particular of four families, and two genera; Koleff et al., 2009). Target–setting for species, which ranged from 5–40 % of their distribution area, were based on weights given to different criteria, such as the degree of rarity, in terms of geographic distribution area, country endemism, extinction risk status in the Mexican red list (NOM–059–SEMARNAT–2001) and in the international red list (IUCN), and status in CITES appendices (I and II) as a proxy of species that need conservation actions because of overexploitation and illegal trade (detailed in Urquiza– Haas et al., 2009). Integration of conservation costs was considered a key aspect in order to come up with a potentially more amenable network for long term persistence of biodiversity and viable in terms of management costs (Luck et al., 2004; Chan et al., 2006). Threats to biodiversity were used to define a suitability layer (i.e. costs) to orient priorities to sites where impediments to conservation are lower and to minimize the selection of areas that have likely lost their biodiversity value. Weights were assigned to each of the 19 threat layers selected based on data availability and known impact on species and ecosystems to obtain a final integrated cost value for each planning unit (detailed in Urquiza– Haas et al., 2009). We used Marxan software (Ball and Possingham, 2000) to identify a set of planning units (out of 8,045 hexagons of 256 km2) that meet the representation targets for biodiversity surrogates while minimizing the area and the costs of the conservation network. Marxan uses a simulated annealing algorithm to find multiple alternative good solutions to the minimum set problem; it does so with an iterative improvement method that incorporates occasional backward steps. Marxan produces a best solution that

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represents the reserve network with the lowest score from all the reserve networks generated and also provides the selection frequency of sites (Ball et al., 2009). Marxan was run for Mexico’s spatial prioritization analysis with 10,000 runs and 1,000,000 iterations. PAs were not considered a priori in the selection of priority conservation sites because not all PAs in the network have proven effective in terms of reducing or halting ecosystem degradation. Furthermore, as part of the gap analysis, all participants and stakeholders agreed to evaluate the performance of PA to represent species and other biodiversity elements efficiently (results not shown here). This algorithmic process allowed the selection of terrestrial priority sites (TPS) for conservation. These sites cover 594,894 km2 (30.6 % of the country’s continental territory) and include a subset of the best solution sites (43 % of the country) with the highest selection frequency scores that met most of the biodiversity surrogate targets (90.5 %). Irreplaceable sites, i.e. essential sites to meet conservation targets or sites where unique biodiversity elements are distributed, cover 16.6 % of the territory and accomplished conservation targets for 81 % of all biodiversity surrogates. As Mexico continues to reinforce efforts for conservation, the percentage of coverage of TPS under protected areas (federal, state, municipal and private PA) has increased 1,316,927 ha (2.1 %) in the last 10 years (CONABIO, 2015; CONANP, 2017). The PA network covered a total of 244,539 km2 (or 12.54 % of the continental surface) by the end of 2016 (Sarukhán et al., 2017). Surrogate assessments Effectiveness of biodiversity surrogates: influence of target–setting methods and selection criteria Methods We determined whether a system of priority sites for conservation based on mammal species of conservation concern (i.e. endemic, rare and threatened) was appropriate in terms of the representation of other species of this taxonomic group. Mammals were chosen for this assessment as Mexico holds the second largest number of mammal diversity worldwide (564 described taxa, including 169 endemic terrestrial species and 50 marine species). Besides, species have relatively well–known distributional data, and at least half of the taxa (291) are threatened according to the Mexican legal list of endangered species (NOM–059–SEMARNAT–2001), circumstances that explain the interest to assess whether species of conservation concern of this group deliver efficient outcomes as surrogates of other mammal and vertebrate species (cf. Di Minin et al., 2016). We used SDMs (Ceballos et al., 2006) described in Koleff et al., 2009) of 354 mammal species of 10 orders. Data available were insufficient to produce SDMs for a further 113 mammals whose distributions are highly restricted, so we used maps from occurrence records. In total, our dataset on mammal distribution covered 96 % of all terrestrial mammal species in Mexico. Species only occurring on islands were excluded

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from the analysis as island inventory was incomplete. Most of the insular territory in Mexico (84.8 %) today is already protected by the PA system, which is very relevant because of the elevated number of insular endemics and threatened species (Sarukhán et al., 2017). We used this dataset to identify ten conservation area networks based on the best solution (hereafter, CAN) using four surrogate groups and different conservation targets running Marxan (1,000,000 iterations and 10,000 runs). Surrogate groups were as follows: [1] Threatened species (TS; n = 104); species listed as critically endangered (CR), endangered (E), or vulnerable (VU) in the IUCN red list, or listed as possibly extinct in the wild (E), at risk of extinction (P), or threatened (A) in the Mexican list of endangered species (NOM–059–SEMARNAT–2001). [2] All species of conservation concern (SCC; n = 241); species that fulfill any of the following criteria: (a) endemic to the country; (b) of restricted distribution (using as a threshold the last quartile of the geographic distribution range of all mammal species, hereafter referred as rare species); (c) listed as E, P or A in the Mexican list of endangered species; (d) listed as CR, E or VU in the IUCN Red List; (e) listed in CITES Appendices I or II. [3] Other species, i.e. those that did not fulfil any of the above mentioned criteria and are usually not considered of conservation concern (OS; n = 204). and [4] All mammal species (AS; n = 445). We set targets at 10 % and 20 % of the species distribution area (OS and AS), and also used variable target levels (5–40 %) for species of conservation concern (TS and SCC), by applying the target–setting methods used for the Gap Analysis, i.e. assigning values to each criteria above mentioned and summing them to obtain the final percentage. Threatened, endemic and rare species thus had the highest conservation targets (supplementary material, see Urquiza–Haas et al., 2009). We used the same costs layer as in the Gap Analysis (see Urquiza–Haas et al., 2009) to consider the degree of impediments to conservation success at the beginning of the planning process. We further tested whether sites selected randomly were as efficient as sites selected using surrogate groups with SCP tools. We generated 100 random solutions to obtain average values of species representation and proportion of species distribution area achieved by the random CANs. We considered 799 and 1,579 planning units for analysis (scenarios R–799 and R–1579, respectively), corresponding to the average number of planning units of scenarios using 10 % and 20 % variable target values, respectively. Results The solutions of the ten different scenarios based on four indicator groups and three alternative conservation target–settings differed in total area, spatial distribution, selection frequency of planning units’, and representation of mammal species (table 1). As expected, the size of the CAN was strongly influenced by the targets assigned. It doubled from circa 10 to 20 % of the country’s continental surface, as targets were doubled from 10 to 20 % of the total species’ distribution area, irrespective of the indicator group

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used. Scenarios with conservation targets of 20 % had CAN with more planning units of higher selection frequencies (i.e. irreplaceable sites) in comparison with scenarios where conservation targets were set at 10 %. CANs were similar in size for scenarios using variable target levels (TS–V; SCC–V) and target levels of 10 % (TS–10; SCC–10; AS–10; OS–10) (table 1). An important difference between scenarios of variable target levels and targets of 10% was in respect to the selection frequency and the spatial distribution of planning units, as well as in the levels of species representation. Solutions of scenarios with variable target levels (TS–V; SCC–V) were more spatially compact and had a higher number of irreplaceable units (fig. 1). For instance, the number of clusters, i.e. adjacent and connected planning units of larger size (> 2,560 km2), was higher for the scenario designed for species of conservation concern using variable targets than for the scenario with 10 % targets (11 vs. 4 clusters, respectively). The frequency distribution of clusters differed significantly between these scenarios (one–tailed Mann–Whitney test U = 169; p = 0.04). Also, scenario SCC–V had more planning units with selection frequencies higher than 80 % (n = 819) than the scenario SCC–10 which only had 136 planning units. Considering a selection frequency of more than 90 %, the numbers decreased to 136 and 79 planning units, respectively). In scenarios TS–V and SCC– V, selected planning units with high irreplaceability scores were concentrated in several geographical areas across the country: the northern part of the Mexican plateau, the Pacific coast (along the states of Michoacan and Guerrero), and southern Mexico (Chimapalas, Lacandon Forest, Maya Forest). Many of these areas coincide with areas of high conservation importance identified in other planning exercises based on patterns of mammal species richness and the concentration of endemic and threatened species (Ceballos et al., 1998). Representation values (i.e. proportion of species in the CAN) for all mammal species was high (81.8– 100 %) irrespective of the indicator group or target levels used to design the CAN. A sample of randomly chosen planning units of ca. 10 % and 20 % of the continental surface also achieved high representation values (83.4 and 87.5 %, respectively). Nonetheless, representation values for several non–target groups varied considerably. For instance, in scenarios using species of less concern (OS–10 and OS–20), representation values varied from 23.6 % for rare species to 84.5 % for all mammal species. Rare species had a poor representation (23.6–34.9 %) in scenarios that did not explicitly consider these species in the planning process (OS–10, OS–20), attaining higher representation in scenarios designed for threatened species (66–73.6 %; table 1). Scenarios designed for non–conservation concern species (OS–10, OS–20) performed slightly worse than randomly chosen planning units of similar area (R–799, R–1579), in particular for rare species (table 1). Scenarios performed differently with regards to the average proportion of species distribution area achieved by CANs, with increments in accordance

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with representation targets. CANs designed for threatened species and species of conservation concern using variable and 10 % targets captured an average of 18–24 % of the species’ distribution area, whereas scenarios for all indicator groups with 20 % targets captured an average of 24–31 % of the species' distribution area. The scenario designed for species of less conservation concern using 10 % targets (OS–10) was outperformed by almost all others (fig. 2), and was very similar to the scenario of randomly chosen planning units of equivalent total area (R–799: 11.9, 95 % CI = 10.8–13.0). OS–10 covered on average 12 % of the species distribution, while the scenario designed for threatened species with 10 % target (TS–10) covered on average 18 %; nonetheless these scenarios did not differ statistically from each other (fig. 2). The scenario designed for species of conservation concern with 10 % and variable target levels performed as well as scenarios designed for all species using a 10 % target; they had the highest species’ distribution area average, significantly different from all other scenarios using a 10 % target level. Less than 2 % of rare species had more than 60 % of their distribution area covered by the CAN in scenario OS–10, in comparison to 25 % of rare species in scenario TS–10. On the other hand, only 5 % of rare species had less than 10 % of their distribution area covered by CANs designed for threatened species and for species of conservation concern using variable targets (TS–V and SCC–V; fig. 3) in comparison to scenario OS–10, in which 8.5 % of the rare species had less than 10 % of their distribution covered. Overall threatened species performed well as surrogates based on the high representation achieved for non–target species. When endemic and rare species were also considered as surrogates (i.e. all species of conservation concern), it was possible to represent all mammal species, including those of less concern. Moreover, CANs designed for species of conservation concern using variable target levels as in the Mexican Gap Analysis were more area efficient and more compact (i.e. CAN is more connected and its perimeter is minimized) than that of the CAN designed for all mammal species using a 10 % representation target. Testing the effectiveness of taxonomic groups as biodiversity surrogates Methods In this assessment, we explored whether a CAN designed for a given vertebrate taxon can appropriately represent all vertebrates. We used the same database and criteria used in the Mexican Gap Analysis described beforehand (i.e. size of planning units, spatial prioritization algorithm, cost and target–setting methods) and compared the performance of mammals, resident birds (hereafter referred as birds), reptiles and amphibians as taxonomic surrogates. A total of 1,146 species were selected following target–setting methods to focus on species of conservation concern (i.e. endemic, rare, and threatened) (208 amphibians, 424 reptiles, 273 birds, 241 mammals).

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Surrogate performance was measured in terms of the effectiveness of CAN designed for each indicator group to represent non–target species from other taxa. We standardized our analyses to a subset of planning units from the best solutions. From these units we calculated species accumulation curves in EstimateS (Colwell, 2006), first, to predict the number of species represented in a given number of planning units and second, to statistically compare the performance of indicator groups based on 95 % confidence intervals. We further tested whether sites selected for indicator taxa using SCP tools were more or less efficient than sites selected randomly. We generated 100 random solutions (considering 477 and 1,510 planning units) to obtain averaged accumulation curves. Results CANs selected for full coverage of indicator taxa included between 753 and 875 of all non–target species, while the number of planning units ranged from 477 for amphibians to 1,510 for birds (table 2; fig. 4). CANs designed for amphibians, reptiles and mammals did not differ significantly in their effectiveness as indicators (as measured by overlapping confidence intervals; fig. 5). On average, CANs designed for amphibians and reptiles represented the highest proportion of non–target species (on average, 80.1 % and 85.7 %, respectively), while CAN designed for birds represented the lowest number of species (74 %) over an area three times that of amphibians (table 2). With the CAN size being equal (477 planning units), amphibians were the most effective surrogate, representing on average of 80.1 % of non–target species, and 83.9 % of species from all four vertebrate groups examined, followed by the CAN designed for reptiles (78.4 % and 87.3 %, respectively; table 3). CAN designed for birds was the least effective (60.9 % and 71.1 %). In contrast, birds were represented almost entirely (on average 97.3 %) by the CAN designed for other taxa, while amphibians and mammals had the poorest average representation within CAN designed for other taxa (59.8 % and 65.5 %, respectively). Differences in the degree of effectiveness of indicator groups were more evident when the number of species that accomplished their conservation targets was measured. CAN designed for reptiles accomplished or surpassed the targets for 958 species of all four taxa (on average 73.5 %), in contrast with the CAN designed for amphibians that accomplished or surpassed the targets for only 748 species (56.1 % on average). However, the CAN for reptiles was almost twice as large as that for amphibians, and managed to increase the total number of species that accomplished their conservation targets by 18.3 %. In contrast, the CAN designed for birds only increased the number of species that accomplished their target by 7.6 % with respect to the CAN designed for amphibians, but in an area 3.2 times larger. Using the same number of planning units as the best solutions for amphibians (n = 477) and birds (n = 1,510), we found that a random selection of sites represented 71.9 % and 82.9 % of all species, respectively. Randomly selected planning units outperformed

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Table 1. Comparison of conservation scenarios for the protection of Mexican terrestrial mammals resulting from varying subsets of mammal species and conservation targets: TS–10, threatened species, 10 % conservation targets; TS–V, threatened species, variable (5–40 %) conservation targets; SCC–10, species of conservation concern, 10 % conservation targets; SCC–V, species of conservation concern, variable (5–40 %) conservation targets; OS–10, other species, i.e. of non–conservation concern, 10 % conservation targets; AS–10, all species, 10 % conservation targets; R–799, 100 random solutions, each 799 planning units, mean (SD); TS–20, threatened species, 20% conservation targets; SCC–20, species of conservation concern, 20 % conservation targets; OS–20, other species, i.e. of non–conservation concern, 20 % conservation targets; AS–20, all species, 20 % conservation targets; R–1579, 100 random solutions, each 1,579 planning units, mean (SD).

Tabla 1. Comparación de los escenarios de conservación para la protección de mamíferos terrestres en México derivados de distintos subconjuntos de especies de mamíferos y metas de conservación: TS–10, especies amenazadas, metas de conservación del 10 %; TS–V, especies amenazadas, metas de conservación variables (5–40 %); SCC–10, especies de interés para la conservación, metas de conservación del 10 %; SCC–V, especies de interés para la conservación, metas de conservación variables (5–40 %); OS–10, otras especies, es decir, de poco interés para la conservación, metas de conservación del 10 %; AS–10, todas las especies, metas de conservación del 10 %; R–799, 100 soluciones aleatorias, considerando 799 unidades de planificación, media (DE); TS–20, especies amenazadas, metas de conservación del 20 %; SCC–20, especies de interés para la conservación, metas de conservación del 20 %; OS–20, otras especies de poco interés para la conservación, metas de conservación del 20 %; AS–20, todas las especies, metas de conservación del 20 %; R–1579, 100 soluciones aleatorias, considerando 1.579 unidades de planificación, media (DE). Threatened All species (%) species (%) Planning n = 104 n = 445 units (#) Scenario

Species of conservation concern (%) n = 241

Endemic species (%) n = 128

Rare species (%) n = 106

85.1 89.1 66.0

TS–10 785 91.9 100 767 91.5 100

84.23 87.50 66.0

SCC–10 780 100 100

100 100 100

TS–V SCC– V

819

100

OS–10 826 81.8 67.3

66.4 74.2 23.6

AS–10 819 100 100

100 100 100

R–799

799

83.4 (1.3)

70.2 (3.1)

TS–20 1,560 93.7 100 SCC–20 1,571 100

100

69.6 (2.4)

77.0 (2.5)

30.0 (5.3)

88.4 93.0 73.6 100

100

OS–20 1,599 84.5 69.2

71.1 77.3 34.9

AS–20 1,588 100 100

100 100 100

R–1579

1,579

87.5 (1.1)

77.1 (3.1)

those achieved by CAN designed for birds by 7.6 %. Likewise, the confidence intervals of CAN designed for mammals overlapped those of the random selections of sites. However, the CAN for amphibians and reptiles outperformed the random solutions (fig. 5). As the effect of using costs in the performance of indicator groups was not accounted for when compared with sites selected at random, we explored their influence on the efficiency of indicator groups by comparing the random selected sites with the CAN designed for indicator groups without using cost information in the algorithm selection process. Results

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77.1 (2.1)

83.2 (2.4)

47.2 (4.7)

(not presented here) showed the same tendency as described for the CANs designed for each taxon. In general, however, they represented around 1 % more species than CANs that included costs. The minimum set of planning units (1,824, representing 23.3 % of the country’s continental surface) required to achieve all targets for vertebrate species of conservation concern was almost as large as the CAN designed for birds, and four times as large as the CAN designed for amphibians (fig. 4). Integration of CANs designed individually for each of the indicator groups consisted of 2,582 planning units (32.9 % of

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TS–10 TS–V

SCC–10 SCC–V

OS–10 AS–10

Selection frequebcy (%) < 49.99 50–59.99

60–69.99 70–79.99

80–89.99 90–99.99

100

Fig. 1. Conservation area networks designed with varying target levels for different subgroups of Mexican mammal species (i.e. scenarios, see abbreviations in table 1). Planning units of Marxan best solutions are shown by selection frequency intervals (degree of irreplaceability). Fig. 1. Redes de áreas de conservación diseñadas con diferentes valores de metas de conservación para distintos subgrupos de especies de mamíferos en México (es decir, escenarios; véanse las abreviaciones en la tabla 1). Las unidades de planificación de las mejores soluciones de Marxan se muestran por intervalos de sus frecuencias de selección (grado de irremplazabilidad).

the continental surface), while only about a third of the planning units were spatially congruent between two or more CANs (i.e. 47; 221 and 606 units were spatially coincident for four, three, and two of the CANs designed for indicator groups, respectively).

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Discussion Endangered species and endemic and rare species have commonly been used or proposed as surrogates in SCP exercises (e.g. Brooks et al., 2004; Diniz et al.,

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TS–20 SCC–20

OS–20 AS–20

2017). However, questions have arisen as to whether they are adequate indicators to guide the implementation of cost–efficient conservation approaches (see references in Drummond et al., 2010). This issue has been poorly explored with large datasets in the context of a real–world conservation planning process. Our assessment using a comprehensive terrestrial vertebrate dataset showed that threatened species provided coverage for 92 % of all mammal species examined, even though they constitute only 24 % of all species examined. Furthermore, nearly 30 % of all threatened species were missing in CANs not designed for these species. Very similar results were found by Drummond et al. (2010) for mammal species in Indonesia, another megadiverse country. Likewise, Lawler et al. (2003) in the Middle Atlantic region of the United States found that threatened species from several taxa performed well as surrogates as they covered on average of 84 % of all species examined, while threatened species were poorly represented (15–58 %) in CANs designed for other species. Even when using a limited number of species, rare and threatened species had the best surrogacy performance (Jones et al., 2016). The results of these studies and our own are similar even though the planning units differed widely in size(1–650 km2), which is not expected because larger planning units are more likely to represent more species. At the subcontinental scale, Tognelli (2005) also showed the effectiveness of threatened terrestrial mammal species but specifically noted that of geographically rare species as indicator groups for other South American mammal species.

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In the present work we highlight the ability of the set of threatened, endemic and rare species to act as effective surrogates in megadiverse Mexico for the coverage of all other species. The use of threatened species alone was not as adequate as it missed a large proportion of rare species. This is of particular relevance for conservation in a country with a high proportion of endemic and range restricted species (Llorente–Bousquets and Ocegueda, 2008). Like studies in other biogeographic regions (e.g. Lawler et al., 2003; Larsen et al., 2007; Jones et al., 2016), our findings illustrate that the efficacy of a surrogate group is related to a greater proportion of threatened and rare species regardless of the taxonomic group used, in particular when trying to cover other threatened and rare species. Nonetheless, this may not be always the case. Franco et al. (2009) found that threatened butterflies were not adequate to represent the non–threatened species of butterflies. The usefulness of taxonomic groups to act as surrogates in SCP is still an unresolved issue (see Rodrigues and Brooks, 2007). In particular, in high beta diversity countries, like Mexico, Koleff et al. (2008) anticipated that no vertebrate taxonomic group could be used to set priorities for other groups based on the general weak congruence of their diversity patterns. The results of the taxonomic surrogate assessment shown here indicate that at least half of the species in other taxonomic groups were covered by a single taxonomic group. Similar conclusions are found elsewhere (Lawler et al., 2003; Moore et al., 2003; Rondinini and Boitani, 2006; Larsen et al., 2009). While no conservation area system designed for a

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Distribution area (%)

a b

bc 20

TS–10 TS–V SCC–10 SCC–V OS–10 AS–10 R799 TS–20 SCC–20 OS–20 AS–20 R–1579 Scenarios

Fig. 2. Percentage of species distribution area covered by conservation area networks designed for different subsets of mammal species (i.e. scenarios; see table 1 for abbreviations). The letters indicate significant differences between scenarios (Kruskal–Wallis, n = 4,184, p < 0.001, K–W H test and pairwise comparisons conducted in SPSS® Statistics 19; highest significant pairwise p–value = 0.021). Fig. 2. Porcentaje de las áreas de distribución de las especies cubierta por redes de áreas de conservación diseñadas para diferentes subconjuntos de especies de mamíferos (es decir, escenarios; véanse las abreviaciones en la tabla 1). Las letras indican diferencias significativas entre los escenarios (Kruskal–Wallis, n = 4.184, p < 0,001, prueba K–W y comparaciones por pares realizadas en SPSS® Statistics 19; valor del par más significativo p = 0,021).

single taxon is able to represent all species of other taxonomic groups, the reptiles and amphibians were the most effective surrogates because they represented a very high percentage of non–target species, and performed better than expected at random. The efficacy of a surrogate group might vary between regions and taxonomic groups due to scale dependence in spatial patterns of species richness and species turnover (Hess et al., 2006; Franco et al., 2009). However, distributional congruence (Koleff et al., 2008) did not serve as predictor of the performance of indicators groups or of overall priorities for terrestrial vertebrate conservation. The number of species within an indicator group might influence their effectiveness (Larsen et al., 2012). In this case, however, it did not appear to be important, as 208 species of amphibians used as surrogates were as effective as 424 reptile species for the representation of other vertebrate species. On average, Mexican amphibians and reptiles, which were more effective surrogates, have smaller range sizes and higher species turnover rates than birds. Thus, efficacy of a surrogate group at this national scale of analysis appears to be related to the use of indicator species that have relatively non–overlapping ranges, collectively covering many environments (Lawler et al., 2003; Lewandowski et al., 2010). Moreover, taxa that contain

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many species with restricted ranges are less likely to be captured by CAN designed for other taxa with widespread distributions (Brooks et al., 2001; Moore et al., 2003). Likewise, ecological surrogates might be less effective for threatened taxa or of species of conservation concern than more widely distributed features (Grantham et al., 2010). A better insight could be gained by testing random species, and by controlling their distributional characteristics (e.g. widespread or restricted distributions, fragmented or continuous; see Sánchez–Fernández and Abellán, 2015). Concerns have arisen about the apparent arbitrariness of target–setting, and the often too low target values (Carwardine et al., 2009; Di Minin and Moilanen, 2012). For the gap analysis, a method was developed to establish species targets more objectively with the aid of expert input and easily measured characteristics related to the conservation status and the distribution of species (see Supplementary material). Here we did not try to thoroughly evaluate the influence of this target–setting method on species persistence per se, but rather to evaluate the appropriateness of this method in view of indicator effectiveness, and on the configuration of the CAN, which might ultimately influence species persistence. Results of the independent assessment indicate that variable target–setting, giving more weight to species

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100

Representation species (%)

90 80 70

Covered distribution area 60–100% 20–59.9% 10–19.9% < 10%

60 50 40

330 20 10 0

TS–10 TS–V SCC–10 SCC–V OS–10 AS–10 TS–20 SCC–20 OS–20 AS–20 Scenarios

Fig. 3. Percentage of rare Mexican mammal species represented in each scenario shown by the percentage of their distribution area covered by conservation area networks designed for different subgroups of mammal species (i.e. scenario, see table 1 for abbreviations). Fig. 3. Porcentaje de especies raras de mamíferos de México representadas en cada escenario según el porcentaje de su área de distribución cubierta por las redes de áreas de conservación diseñadas para diferentes subgrupos de especies de mamíferos (es decir, escenarios; véanse las abreviaciones en la tabla 1).

of mammals of conservation concern (i.e. set of threatened, endemic and rare species) than to conventional fixed targets of 10 % for all species (e.g. Urbina–Cardona and Flores–Villela, 2010), did not significantly

influence the number of represented species and average species distribution area covered by the CAN. However, it did generate a more compact, connected and more area efficient CAN with a higher number of

Table 2. Representation percentage of Mexican vertebrate species of conservation concern by taxonomic group (rows) in the conservation area network designed for amphibians, reptiles, birds, and mammals (columns): 1 average for non–target species. Tabla 2. Porcentaje de representación de especies de vertebrados de interés para la conservación por grupo taxonómico (filas) en la red de áreas de conservación diseñada para anfibios, reptiles, aves y mamíferos (columnas), en México: 1 promedio de las especies no objetivo.

Amphibians Reptiles Birds (n = 208) (n = 424) (n = 273)

# of units Amphibians Reptiles Birds Mammals Total Average1

Mammals (n = 241) Average1

477 967 1,510 795 100 82.7 75 63 73.5 78.1 100 77.8 74.5 76.8 95.6 99.3 100 98.5 97.8 66.8 75.1 69.3 100 70.4 83.9 91.5 80.8 83.5 80.1 85.7 74.0 78.7

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Amphibians Reptiles

Birds Mammals

Selection frequency (%)

Terrestrial vertebrates

< 49.99 50-59.99 60-69.99 70-79.99 80-89.99 90-99.99 100

Fig. 4. Conservation area networks designed for indicator species groups individually and collectively. Planning units of Marxan best solutions are shown by selection frequency intervals (degree of irreplaceability). Fig. 4. Redes de áreas de conservación diseñadas para cada grupo de especies indicadoras y para todos los grupos juntos. Las unidades de planificación de las mejores soluciones de Marxan se muestran por intervalos de sus frecuencias de selección (grado de irremplazabilidad).

irreplaceable sites, in contrast with expectations that more area is needed to meet higher targets (Justus et al., 2008), in particular in high beta diverse regions. Our results support the use of threat classifications and proxies for vulnerability to set larger targets for more threatened and vulnerable species (Lombard et al., 2003; Moore et al., 2003). The methods to select surrogate species in the gap analysis maximized species representation, and can be easily applied to other megadiverse regions and countries when faced

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with limited information about habitat requirements and minimum viable population sizes. Nonetheless, we also recommend adjustments to target–setting in order to assign higher target levels to rare species. On the other hand, other considerations should be included to guarantee that the CAN selects peak abundance locations where species are presumably more viable (Bonn et al., 2002). Furthermore, when very few areas are identified as irreplaceable, the conservation network is considered poorly defined

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Table 3. Estimation of the representation percentage of Mexican vertebrate species of conservation concern by taxonomic group (rows) in the conservation area network designed for amphibians, reptiles, birds and mammals (columns) considering an equal number of planning units (477): 1 average for non–target species. Tabla 3. Estimación del porcentaje de representación de las especies de vertebrados de interés para la conservación por grupo taxonómico (filas) en la red de áreas de conservación diseñada para anfibios, reptiles, aves y mamíferos (columnas), en México, considerando el mismo número de unidades de planificación (477): 1 promedio de las especies no objetivo.

Amphibians Reptiles Birds Mammals (n = 208) (n = 424) (n = 273) (n = 241) Average1

Amphibians 100 68.8 55.3 55.3 59.8 Reptiles 78.1 100 65.3 68.4 70.7 Birds

95.6 98.9 100 97.8 97.3

Mammals 66.8 67.6 62.2 100 65.6 Total

83.9 87.3 71.1 79.7

Average 80.1 78.4 60.9 73.8 1

100

Species represented (%)

90 80 70 60 50

Birds

Mammals

Reptiles

Amphibians Random (n = 1,500)

Random (n = 477)

10 0

200

400

600 800 1,000 Number of planning units

1,200

1,400

1,600

Fig. 5. Cumulative percentage of all non–target species represented in the conservation area network designed for Mexican amphibians, reptiles, birds and mammals, compared against random selected sites (n = 477 and n = 1,500). Dotted lines denote 95 % confidence intervals. Fig. 5. Porcentaje acumulado de todas las especies de otros grupos representadas en la red de áreas de conservación diseñada para los anfibios, reptiles, aves y mamíferos en México, comparado con sitios seleccionados aleatoriamente (n = 477 y n = 1.500). Las líneas discontinuas denotan intervalos de confianza del 95 %.

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and difficult to implement in real–world settings. As the opposite is also true when too many irreplaceable sites are identified. Levin et al. (2015) recommended performing a sensitivity analysis and provided a general guideline (i.e. 10–20 % of the study areas with sites of selection frequency values over 90 %) for a flexible conservation area network that could be more useful for managers and decision makers. Areas of high irreplaceability (90–100 % selection frequency, 16.6 % of the CAN) identified for all vertebrate species of conservation concern did not coincide with most conservation areas detected by Sarkar et al. (2009) in Mesoamerica for plant and vertebrate species at risk of extinction using a systematic approach —with targets set at 10 % and 20 % of species potential distribution area. This is most likely because the latter study did not include 'cost' information as it identifies large areas of transformed land (e.g. induced pasturelands in Veracruz) as important for protection. This highlights the need to incorporate information on the human impact at the beginning of the planning process —especially in the absence of species distribution maps that account for current habitat conditions— in order to come up with a more viable CAN for conservation action. Conclusions Designing conservation networks to guide biodiversity conservation actions is a particularly challenging task for governments and conservation practitioners in megadiverse countries. The available choice of data, biodiversity surrogates, and target–setting methods, for example, affect the accuracy of the desired conservation plan outcome (Rondinini et al., 2006). As such, the Mexican gap analysis project calls for the active participation and engagement of biodiversity scientists and national–level stakeholders. In this work, we contribute to the debate around the usefulness of surrogates for representing biodiversity at the species level. We also test the decisions taken with respect to the choice of surrogates and target setting methods used in the Mexican gap analysis project by means of two independent assessments. Chosen criteria were appropriate to select surrogates that maximized overall vertebrate species representation in a mega and high beta–diverse country such as Mexico (Koleff et al., 2008). This was a key step in the process, especially for stakeholders that wanted to know about the robustness of choice of surrogates. The use of a comprehensive dataset guards against bias and is of particular importance when implementating conservation action in real–world circumstances. In view of our results, we outline some practical recommendations to planners when selecting species–level surrogates: 1) conservation planning should include comprehensive suites of complementary measures of biodiversity, including as many species as possible from taxonomically diverse groups for which reasonable good quality distribution information is available; 2) surrogate species groups should be comprised a great number of rare species distributed

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across broad environmental gradients because as many of them are sensitive to environmental changes they might be good indicators for other species; 3) rare and threatened species need to be included in the planning process as they are most likely to be missed by other environmental or species indicators; and 4) there is a need for a better understanding of the distribution of rare, threatened and understudied species and their sensitivity to human impacts for dynamic conservation planning. The challenge in mega and beta diverse countries like Mexico with pressing environmental problems (Bradshaw et al., 2010) implies the need for urgent and effective conservation action. Mexico has partly paved the way by identifying areas where a significant proportion of its biodiversity can be maintained. Acknowledgements We acknowledge all the members of the Mexican Gap Analysis Project and all the workshop participants (see credits in Koleff et al. 2009), without whom this work would not have been possible. Special thanks to colleagues from CONANP and CONABIO. Funding was provided by CONANP. We also thank the two anonymous reviewers for their insightful comments and suggestions. References Ball, I. R., Possingham, H. P., 2000. Marxan (v 1.8.2): Marine Reserve Design Using Spatially Explicit Annealing, a Manual. The University of Queensland, Brisbane. Ball, I. R., Possingham, H. P., Watts, M. E., 2009. Marxan and relatives: Software for spatial conservation prioritization. In: Spatial conservation prioritization. Quantitative methods & computational tools: 185– 195 (A. Moilanen, K. A. Wilson, H. P. Possingham, Eds.). Oxford University Press, Oxford, UK. Bonn, A., Rodrigues, A. S. L., Gaston, K. J., 2002. Threatened and endemic species: Are they good indicators of patterns of biodiversity on a national scale? Ecology Letters, 5: 733–741. Bradshaw, C. J. A., Giam, X., Sodhi, N. S., 2010. Evaluating the Relative Environmental Impact of Countries. Plos One, 5: e10440. Brooks, T., Balmford, A., Burgess, N., Fjeldså, J., Hansen, L. A., Moore, J., Rahbek, C., Williams, P., 2001. Toward a blueprint for conservation in Africa. BioScience, 51: 613–624. Brooks, T., Da Fonseca, G. A. B., Rodrigues, A. S. L., 2004. Species, data, and conservation planning. Conservation Biology, 18: 1682–1688. Carwardine, J., Klein, C. J., Wilson, K. A., Pressey, R. L., Possingham, H. P., 2009. Hitting the target and missing the point: target–based conservation planning in context. Conservation Letters, 2: 4–11. Ceballos, G., Rodríguez, P., Medellín, R. A., 1998. Assessing conservation priorities in megadiverse Mexico: Mammalian diversity, endemicity, and

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Rondinini, C., Boitani, L., 2006. Differences in the umbrella effects of african amphibians and mammals based on two estimators of the area of occupancy. Conservation Biology, 20: 170–179. Rondinini, C., Wilson, K. A., Boitani, L., Grantham, H., Possingham, H. P., 2006. Tradeoffs of different types of species occurrence data for use in systematic conservation planning. Ecology Letters, 9: 1136–1145. Sánchez–Fernández, D., Abellán, P., 2015. Using null models to identify under–represented species in protected areas: A case study using European amphibians and reptiles. Biological Conservation, 184: 290–299. Sarkar, S., Pressey, R. L., Faith, D. P., Margules, C. R., Fuller, T., Stoms, D. M., Moffett, A., Wilson, K. A., Williams, K. J., Williams, P. H., Andelman, S., 2006. Biodiversity conservation planning tools: Present status and challenges for the future. Annual Review of Environment and Resources. Sarkar, S., Sánchez–Cordero, V., Londoño, M. C., Fuller, T., 2009. Systematic conservation assessment for the Mesoamerica, Chocó, and Tropical Andes biodiversity hotspots: A preliminary analysis. Biodiversity and Conservation, 18: 1793–1828. Sarukhán, J., Koleff, P., Carabias, J., Soberón, J., Dirzo, R., Llorente–Bousquets, J., Halffter, G., González, R., March, I., Mohar, A., Anta, S., De la Maza, J., Pisanty, I., Urquiza Haas, T., Ruiz González, S. P., García Méndez, G., 2017. Capital natural de México. Síntesis: evaluación del conocimiento y tendencias de cambio, perspectivas de sustentabilidad, capacidades humanas e institucionales. Comisión Nacional para el Conocimiento y Uso de la Biodiversidad, México. Soberón, J., Peterson, A. T., 2004. Biodiversity informatics: Managing and applying primary biodiversity data. Philosophical Transactions of the Royal Society B: Biological Sciences, 359: 689–698. Stafford–Smith, M., Gaffney, O., Brito, L., Ostrom, E., Seitzinger, S., 2012. Interconnected risks and solutions for a planet under pressure – overview and introduction. Current Opinion in Environmental Sustainability 4: 3–6. Tognelli, M. F, 2005. Assessing the utility of indicator groups for the conservation of South American terrestrial mammals. Biological Conservation, 121: 409–417. Urbina–Cardona, J. N., Flores–Villela, O., 2010. Ecological–niche modeling and prioritization of conservation–area networks for Mexican herpetofauna. Conservation Biology, 24: 1031–1041. Urquiza–Haas, T., Kolb, M., Koleff, P., Lira–Noriega, A., Alarcón, J., 2009. Methodological approach to identify Mexico’s terrestrial priority sites for conservation. Gap Bulletin, 16: 70–80.

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Animal Biodiversity and Conservation 42.1 (2019)

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

Normes de publicació Els treballs s'enviaran preferentment de forma electrònica (abc@bcn.cat). El format preferit és un document Rich Text Format (RTF) o DOC que inclogui les figures i les taules. Les figures s'hauran d'enviar també en arxius apart en format TIFF, EPS o JPEG. Cal incloure, juntament amb l'article, una carta on es faci constar que el treball està basat en investigacions originals no publicades anterior ment i que està sotmès a Animal Biodiversity and Conservation en exclusiva. A la carta també ha de constar, per a aquells treballs en que calgui manipular animals, que els autors disposen dels permisos necessaris i que compleixen la normativa de protecció animal vigent. També es poden suggerir possibles assessors. ISSN: 1578–665X eISSN: 2014–928X

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Les proves d'impremta enviades a l'autor per a la correcció, seran retornades al Consell Editor en el termini de 10 dies. Les despeses degudes a modificacions substancials introduïdes per ells en el text original acceptat aniran a càrrec dels autors. El primer autor rebrà una còpia electrònica del treball en format PDF. Manuscrits Els treballs seran presentats en format DIN A–4 (30 línies de 70 espais cada una) a doble espai i amb totes les pàgines numerades. Els manuscrits han de ser complets, amb taules i figures. No s'han d'enviar les figures originals fins que l'article no hagi estat acceptat. El text es podrà redactar en anglès, castellà o català. Se suggereix als autors que enviïn els seus treballs en anglès. La revista els ofereix, sense cap càrrec, un servei de correcció per part d'una persona especialitzada en revistes científiques. En tots els casos, els textos hauran de ser redactats correctament i amb un llenguatge clar i concís. Els caràcters cursius s’empraran per als noms científics de gèneres i d’espècies i per als neologismes intraduïbles; les cites textuals, independentment de la llengua, seran consignades en lletra rodona i entre cometes i els noms d’autor que segueixin un tàxon aniran en rodona. S'evitarà l'ús de termes extrangers (p. ex.: llatí, alemany,...). Quan se citi una espècie per primera vegada en el text, es ressenyarà, sempre que sigui possible, el seu nom comú. Els topònims s’escriuran o bé en la forma original o bé en la llengua en què estigui escrit el treball, seguint sempre el mateix criteri. Els nombres de l’u al nou, sempre que estiguin en el text, s’escriuran amb lletres, excepte quan precedeixin una unitat de mesura. Els nombres més grans s'escriuran amb xifres excepte quan comencin una frase. Les dates s’indicaran de la forma següent: 28 VI 99 (un únic dia); 28, 30 VI 99 (dies 28 i 30); 28–30 VI 99 (dies 28 a 30). S’evitaran sempre les notes a peu de pàgina. Format dels articles Títol. Serà concís, però suficientment indicador del contingut. Els títols amb desig nacions de sèries numèriques (I, II, III, etc.) seran acceptats previ acord amb l'editor. Nom de l’autor o els autors Abstract en anglès que no ultrapassi les 12 línies mecanografiades (860 espais) i que mostri l’essència del manuscrit (introducció, material, mètodes, resultats i discussió). S'evitaran les especulacions i les cites bibliogràfiques. Estarà encapçalat pel títol del treball en cursiva. Key words en anglès (sis com a màxim), que orientin sobre el contingut del treball en ordre d’importància. Resumen en castellà, traducció de l'Abstract. De la traducció se'n farà càrrec la revista per a aquells autors que no siguin castellanoparlants. Palabras clave en castellà. © 2019 Museu de Ciències Naturals de Barcelona Papers are published under a Creative Commons Attribution 4.0 International License

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

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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. Els títols dels apartats generals de l’article (Introducción, Material y métodos, Resultados, Discusión, Conclusiones, Agradecimientos y Referencias) no aniran numerats. No es poden utilitzar més de tres nivells de títols. Els autors procuraran que els seus treballs originals no passin de 20 pàgines (incloent–hi figures i taules). Si a l'article es descriuen nous tàxons, caldrà que els tipus estiguin dipositats en una institució pública. Es recomana als autors la consulta de fascicles recents de la revista per tenir en compte les seves normes. Comunicacions breus Les comunicacions breus seguiran el mateix procediment que els articles y tindran el mateix procés de revisió. No excediran de 2.300 paraules incloent–hi títol, resum, capçaleres de taula, peus de figura, agraïments i referències. El resum no ha de passar de 100 paraules i el nombre de referències ha de ser de 15 com a màxim. Que el text tingui apartats és opcional i el nombre de taules i/o figures admeses serà de dos de cada com a màxim. En qualsevol cas, el treball maquetat no podrà excedir de quatre pàgines.

16/09/2019 13:30:41

Animal Biodiversity and Conservation 42.1 (2019)

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

III

de animales, que los autores disponen de los permisos necesarios y que han cumplido la normativa de protección animal vigente. Los autores pueden enviar también sugerencias para asesores. Las pruebas de imprenta enviadas a los autores deberán remitirse corregidas al Consejo Editor en el plazo máximo de 10 días. Los gastos debidos a modificaciones sustanciales en las pruebas de im prenta, introducidas por los autores, irán a cargo de los mismos. El primer autor recibirá una copia electrónica del trabajo en formato PDF. Manuscritos Los trabajos se presentarán en formato DIN A–4 (30 líneas de 70 espacios cada una) a doble espacio y con las páginas numeradas. Los manuscritos deben estar completos, con tablas y figuras. No enviar las figuras originales hasta que el artículo haya sido aceptado. El texto podrá redactarse en inglés, castellano o catalán. Se sugiere a los autores que envíen sus trabajos en inglés. La revista ofrece, sin cargo ninguno, un servicio de corrección por parte de una persona especializada en revistas científicas. En cualquier caso debe presentarse siempre de forma correcta y con un lenguaje claro y conciso. Los caracteres en cursiva se utilizarán para los nombres científicos de géneros y especies y para los neologismos que no tengan traducción; las citas textuales, independientemente de la lengua en que estén, irán en letra redonda y entre comillas; el nombre del autor que sigue a un taxón se escribirá también en redonda. Se evitará el uso de términos extranjeros (p. ej.: latín, aleman,...). Al citar por primera vez una especie en el trabajo, deberá especificarse siempre que sea posible su nombre común. Los topónimos se escribirán bien en su forma original o bien en la lengua en que esté redactado el trabajo, siguiendo el mismo criterio a lo largo de todo el artículo. Los números del uno al nueve se escribirán con letras, a excepción de cuando precedan una unidad de medida. Los números mayores de nueve se escribirán con cifras excepto al empezar una frase. Las fechas se indicarán de la siguiente forma: 28 VI 99 (un único día); 28, 30 VI 99 (días 28 y 30); 28–30 VI 99 (días 28 al 30). Se evitarán siempre las notas a pie de página.

Normas de publicación

Formato de los artículos

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

Título. Será conciso pero suficientemente explicativo del contenido del trabajo. Los títulos con designaciones de series numéricas (I, II, III, etc.) serán aceptados excepcionalmente previo consentimiento del editor. Nombre del autor o autores Abstract en inglés de 12 líneas mecanografiadas (860 espacios como máximo) y que exprese la esencia del manuscrito (introducción, material, métodos, resultados y discusión). Se evitarán las especulaciones y las citas bibliográficas. Irá encabezado por el título del trabajo en cursiva.

ISSN: 1578–665X eISSN: 2014–928X

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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

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

16/09/2019 13:30:42

Animal Biodiversity and Conservation 42.1 (2019)

Animal Biodiversity and Conservation Animal Biodiversity and Conservation (formerly Miscel·lània Zoològica) is an interdisciplinary journal published by the Museu de Ciències Naturals de Barcelona since 1958. It includes empirical and theoretical research from around the world that examines any aspect of Zoology (Systematics, Taxonomy, Morphology, Biogeography, Ecology, Ethology, Physiology and Genetics). It gives special emphasis to studies that expose a new problem or introduces a new topic, presenting clear hypotheses and predictions, and to studies related to Cconservation Biology. Papers purely descriptive or faunal or chorological describing the distribution in space or time of zoological organisms will not be published. These works should be redirected to our sister magazine Arxius de Miscel·lània Zoològica (www.amz.museucienciesjournals.cat). Studies concerning rare or protected species will not be accepted unless the authors have been granted the relevant permits or authorisation. Each annual volume consists of two issues. Animal Biodiversity and Conservation is registered in all principal data bases and is freely available online at www.abc.museucienciesjournals.cat assuring world–wide access to articles published therein. All manuscripts are screened by the Executive Editor, an Editor and two independent reviewers so as to guarantee the quality of the papers. The review process aims to be rapid and constructive. Once accepted, papers are published as soon as is practicable. This is usually within 12 months of initial submission. Upon acceptance, manuscripts become the property of the journal, which reserves copyright, and no published material may be reproduced or cited without acknowledging the source of information. All rights are reserved by the authors, who authorise the journal to publish the article. Papers are published under a Creative Commons Attribution 4.0 International License: no part of the published paper may be reproduced or reused unless the source is cited.

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

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The first author will receive electronic version of the article in PDF format. Manuscripts Manuscripts must be presented in DIN A–4 format, 30 lines, 70 keystrokes per page. Maintain double spacing throughout. Number all pages. Manuscripts should be complete with figures and tables. Do not send original figures until the paper has been accepted. The text may be written in English, Spanish or Catalan, though English is preferred. The journal provides linguistic revision by an author’s editor. Care must be taken to use correct wording and the text should be written concisely and clearly. Scientific names of genera and species as well as untranslatable neologisms must be in italics. Quotations in whatever language used must be typed in ordinary print between quotation marks. The name of the author following a taxon should also be written in lower case letters. Foreing terms (e.g. Latin, German,...) should not be used. When referring to a species for the first time in the text, both common and scientific names should be given when possible. Do not capitalize common names of species unless they are proper nouns (e.g. Iberian rock lizard). Place names may appear either in their original form or in the language of the manuscript, but care should be taken to use the same criteria throughout the text. Numbers one to nine should be written in full within the text except when preceding a measure. Higher numbers should be written in numerals except at the beginning of a sentence. Specify dates as follows: 28 VI 99 (for a single day); 28, 30 VI 99 (referring to two days, e.g. 28th and 30th), 28–30 VI 99 (for more than two consecutive days, e.g. 28th to 30th). Footnotes should not be used. Formatting of articles Title. Must be concise but as informative as possible. Numbering of parts (I, II, III, etc.) should be avoided and will be subject to the Editor’s consent. Name of author or authors Abstract in English, no longer than 12 typewritten lines (840 spaces), covering the contents of the article (introduction, material, methods, results and discussion). Speculation and literature citation should be avoided. The abstract should begin with the title in italics. Key words in English (no more than six) should express the precise contents of the manuscript in order of relevance. Resumen in Spanish, translation of the Abstract. Summaries of articles by non–Spanish speaking authors will be translated by the journal on request. Palabras clave in Spanish. Author’s address will be published as they appear © 2019 Museu de Ciències Naturals de Barcelona Papers are published under a Creative Commons Attribution 4.0 International License

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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,...

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to papers of the same year. Bibliographic citations in the text must appear in the usual way: "...according to Wemmer (1998)...", "...has been defined by Robinson and Redford (1991)...", "...the prospections that have been carried out (Begon et al., 1999)..." Tables. Must be numbered in Arabic numerals with reference in the text. Large tables should be narrow (across the page) and long (down the page) rather than wide and short, so that they can be fitted into the column width of the journal. Figures. All illustrations (graphs, drawings, photographs) should be termed as figures, and numbered consecutively in Arabic numerals (1, 2, 3, etc.) with reference in the text. Glossy print photographs, if essential, may be included. The Journal will publish colour photographs but the author will be charged for the cost. Figures have a maximum size of 15.5 cm wide by 24 cm long. Figures should not be tridimensional. Any maps or drawings should include a scale. Shadings should be kept to a minimum and preferably with black, white or bold hatching. Stippling should be avoided as it may be lost in reproduction. Legends of tables and figures. Legends of tables and figures should be clear, concise, and written both in English and Spanish. Main headings (Introduction, Material and methods, Results, Discussion, Acknowledgements and References) should not be numbered. Do not use more than three levels of headings. Manuscripts should not exceed 20 pages including figures and tables. If the article describes new taxa, type material must be deposited in a public institution. Authors are advised to consult recent issues of the journal and follow its conventions. Brief communications Brief communications should follow the same procedure as other articles and they will undergo the same review process. They should not exceed 2,300 words including title, abstract, figure and table legends, acknowledgements and references. The abstract should not exceed 100 words, and the number of references should be limited to 15. Section headings within the text are optional. Brief communications may have up to two figures and/or two tables but the whole paper should not exceed four published pages.

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Animal Biodiversity and Conservation 42.1 (2019)

VII

Welcome to the electronic version of Animal Biodiversity and Conservation

Rec ele omme ctr nd o to you nic a c r li bra cess r y!

this

https://abc.journals.museuciencies.cat

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

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Animal Biodiversity and Conservation 42.1 (2019)

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113–126 Campista León, S., Beltrán Espinoza, J. A., Sosa Cornejo, I., Castillo Ureta, H., Martín del Campo Flores, J. R., Sánchez Zazueta, J. G., Peinado Guevara, L. I. Haplotypic characterization of the olive ridley turtle (Lepidochelys olivacea) in northwest Mexico: the northernmost limit of its distribution 127–134 García–De la Peña, C., García–De la Peña, C. Aguillón–Gutiérrez, D. R., Meza–Herrera, C. A., Vaca– Paniagua, F., Díaz–Velásquez, C. E., Valenzuela– Núñez, L. M., Ávila–Rodríguez, V. First insights into the fecal bacterial microbiota of the black–tailed prairie dog (Cynomys ludovicianus) in Janos, Mexico 135–142 Møller, A. P. Head size and personality in great tits Parus major 143–152 García–Grajales, J., Meraz Hernando, J. F., Arcos García, J. L., Ramírez Fuentes, E. Incubation temperatures, sex ratio and hatching success of leatherback turtles (Dermochelys coriacea) in two protected hatcheries on the central Mexican coast of the Eastern Tropical Pacific Ocean

153–161 Serna–Lagunes, R., Álvarez–Oseguera, L. R., Ávila– Nájera, D. M., Leyva–Ovalle, O. R., Andrés–Meza, P., Tigar, B. Temporal overlap in the activity of Lynx rufus and Canis latrans and their potential prey in the Pico de Orizaba National Park, Mexico 163–169 Gonçalves, M. S. S., Gil–Delgado, J. A., López–Iborra, G. M., dos Santos Pons, P. Wind effects on habitat use by wintering waders in an inland lake of the Iberian Peninsula 171–186 Varela–Romero, A., Ruiz–Campos, G., Findley, L. T., Grijalva–Chon, J. M., Gutiérrez– Millán, L. E. Mitochondrial evidence for a new evolutionary significant unit within the Gila eremica lineage (Teleostei, Cyprinidae) in Sonora, Northwest Mexico 187–202 Urquiza–Haas, T., Tobón, W., Kolb, M., Lira–Noriega, A., Contreras, V., Alarcón, J., Koleff, P. Assessing best practice for selecting surrogates and target–setting methods in a megadiverse country

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

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

Índex / Índice / Contents Animal Biodiversity and Conservation 42.1 (2019) ISSN 1578–665 X eISSN 2014–928 X 1–8 Pafilis, P., Kapsalas, G., Lymberakis, P., Protopappas, D., Sotiropoulos, K. Diet composition of the Karpathos marsh frog (Pelophylax cerigensis): what does the most endangered frog in Europe eat? 9–18 Cavalheiro, L. W., Fialho, C. B. Ontogeny of feeding by Astyanax paris in streams of the Uruguay River Basin, Brazil 19–29 Rund, D., Neves, V., Quillfeldt, P. Molecular survey of Hepatozoon infection of Teira dugesii in the Azores 31–37 Salehi, T., Sharifi, M. Comparing the predatory impact of captive–bred and free–living yellow spotted mountain newt (Neurergus microspilotus) on the larval green toad (Bufotes variabilis) 39–43 Fòrum Grossman, G. D. Writing promotion and tenure evaluations for life scientists: thoughts on structure and content 45–57 Zarzo–Arias, A. Romo, H., Moreno, J. C., Munguira, M. L. Distribution models of the Spanish argus and its food plant, the storksbill, suggest resilience to climate change

59–63 Castro, A., Drag, L., Cizek, L., Fernández, J. Rosalia alpina adults (Linnaeus, 1758) (Insecta, Coleoptera) avoid direct sunlight 65–68 Brief Communication Peñarrubia, L., Viñas, J., Sanz, N., Smith, B. L., Alvarado Bremer, J. R., Pla, C., Vidal, O. SNP identification and validation in two invasive species: zebra mussel (Dreissena polymorpha) and Asian clam (Corbicula fluminea) 69–77 Mendes Pontes, A. R., Mariz Beltrão, A. C., Melo Santos, A. M. Reconsidering mammal extinctions in the Pernambuco Endemism Center of the Brazilian Atlantic Forest: a critique 79–90 Maan, S. J., Chaudhry, P. People and protected areas: some issues from India 91–98 Sanz–Aguilar, A., Pradel, R., Tavecchia, G. Age–dependent capture–recapture models and unequal time intervals 99–112 Briones–Salas, M., Lavariega, M. C., Lira–Torres, I.† Mammal diversity before the construction of a hydroelectric power dam in southern Mexico

FUNDACIÓN ESPAÑOLA

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

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

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