ABC 45-1 (2022)

en J. Hatchwell, Univ. of Sheffield, UK

Dibuix de la coberta / Dibujo de la portada / Drawing of the cover Jordi Domènech Apis mellifera: abella de la mel, abella europea o abella occidental / Abeja europea, abeja doméstica o abeja melífera / Western honey bee or European honey bee Abella de la mel a prop del "cortín" o "alvariza", unes construccions de pedra i fang que han permès mantenir l'apicultura en zones amb óssos, conservant l'activitat econòmica i l'ecosistema natural / Abeja de la miel cerca del cortín o alvariza, unas construcciones de piedra y barro que han permito mantener la apicultura en zonas oseras, conservando la actividad económica y el ecosistema natural / Western honey bee near a 'cortín' or 'alvariza', two kind of stone and mud constructions that have made it possible to maintain beekeeping in bear areas, preserving economic activity and the natural ecosystem.

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 Alba Jiménez Assessorament lingüístic / Asesoramiento lingüístico / Linguistic advisers Carolyn Newey Pilar Nuñez

Animal Biodiversity and Conservation 45.1, 2022 © 2022 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: museucienciesjournals.cat/abc/

Animal Biodiversity and Conservation 45.1 (2022)

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 (AEET) Comportament / Comportamiento / Behaviour: Adolfo Cordero (SEEEE) Biologia Evolutiva / Biología Evolutiva / Evolutionary Biology: Santiago Merino (SESBE) Editors / Editores / Editors Pere Abelló Institut de Ciències del Mar CMIMA–CSIC, Barcelona, Spain Pelayo Acevedo Instituto de Investigación en Recursos Cinegéticos IREC–UCLM–CSIC–JCCM, Ciudad Real, Spain Javier Alba–Tercedor Universidad de Granada, Granada, Spain Russell Alpizar–Jara University of Évora, Évora, Portugal Marco Apollonio Università degli Studi di Sassari, Sassari, Italy Pedro Aragón Universidad Complutense de Madrid, Madrid, Spain Miquel Arnedo Universitat de Barcelona, Barcelona, Spain Beatriz Arroyo Instituto de Investigación en Recursos Cinegéticos IREC–UCLM–CSIC–JCCM, Ciudad Real, Spain Francisco Javier Aznar Institut Cavanilles de Biodiversidad y Biologia Evolutiva, Universitat de Valencia, Spain Xavier Bellés Institut de Biología Evolutiva UPF–CSIC, Barcelona, Spain Agustín Camacho Instituto de Biociências–USP, São Paulo, Brasil David Canal MTA Centre for Ecological Research, Vácrátót, Hungary Gonçalo C. Cardoso CIBIO–InBIO, Universidade do Porto, Portugal Salvador Carranza Institut de Biologia Evolutiva UPF–CSIC, Barcelona, Spain Luís Mª Carrascal Museo Nacional de Ciencias Naturales MNCN–CSIC, Madrid, Spain Martina Carrete Universidad Pablo de Olavide, Sevilla, Spain Pablo Castillo Institute for Sustainable Agriculture–CSIC, Córdoba, Spain Darío Díaz Cosín Universidad Complutense de Madrid, Madrid, Spain José A. Donazar Estación Biológica de Doñana EBD–CSIC, Sevilla, Spain Arnaud Faille Museum National histoire naturelle, Paris, France Jordi Figuerola Estación Biológica de Doñana EBD–CSIC, Sevilla, Spain Gonzalo Giribet Museum of Comparative Zoology, Harvard University, Cambridge, USA Susana González Universidad de la República–UdelaR, Montivideo, Uruguay Jacob González–Solís Universitat de Barcelona, Barcelona, Spain Iain Gordon Australian National University, Mysterton, Australia Sidney F. Gouveia Universidad Federal de Sergipe, Sergipe, Brasil Gary D. Grossman University of Georgia, Athens, USA Ben J. Hatchwell University of Sheffield, Sheffield, UK Joaquín Hortal Museo Nacional de Ciencias Naturales MNCN–CSIC, Madrid, Spain Jacob Höglund Uppsala University, Uppsala, Sweden Damià Jaume Institut Mediterrani d'Estudis Avançats IMEDEA–CSIC–UIB, Esporles, Spain José Jiménez Instituto de Investigación en Recursos Cinegéticos IREC–UCLM–CSIC–JCCM, Ciudad Real, Spain Miguel A. Jiménez–Clavero Centro de Investigación en Sanidad Animal–INIA, Madrid, Spain Jennifer A. Leonard Estación Biológica de Doñana EBD–CSIC, Sevilla, Spain Andras Liker University of Pannonia, Veszprém, Hungary Jordi Lleonart Institut de Ciències del Mar CMIMA–CSIC, Barcelona, Spain Josep Lloret Universitat de Girona, Girona, Spain Jorge Mª Lobo Museo Nacional de Ciencias Naturales MNCN–CSIC, Madrid, Spain Pablo J. López–González Universidad de Sevilla, Sevilla, Spain Ian MacGregor–Fors University of Helsinki, Lahti, Finland Jose Martin Museo Nacional de Ciencias Naturales MNCN–CSIC, Madrid, Spain Juan F. Masello Justus Liebig University Giessen, Giessen, Germany Manuel B. Morales CIBC–Universidad Autónoma de Madrid, Madrid Spain Joan Navarro Institut de Ciències del Mar, CMIMA–CSIC, Barcelona, Spain Juan J. Negro Estación Biológica de Doñana EBD–CSIC, Sevilla, Spain Daniel Oro Centre d'Estudis Avançats de Blanes CEAB–CSIC, Girona, Spain Vicente M. Ortuño Universidad de Alcalá de Henares, Alcalá de Henares, Spain Miquel Palmer Institut Mediterrani d'Estudis Avançats IMEDEA–CSIC–UIB, Esporles, Spain Per Jakob Palsbøll University of Groningen, Groningen, The Netherlands Reyes Peña Universidad de Jaén, Jaén, Spain Silvia Perea Universidad Nacional Autónoma de México UNAM, Ciudad de México, México Javier Perez–Barberia Estación Biológica de Doñana EBD–CSIC, Sevilla, Spain Silvia Pérez–Espona The University of Edinburgh, UK Juan M. Pleguezuelos Universidad de Granada, Granada, Spain Montserrat Ramón Institut de Ciències del Mar CMIMA­–CSIC, Barcelona, Spain Alex Richter–Boix CREAF, Universitat Autònoma de Barcelona, Bellaterra, Spain Diego San Mauro Universidad Complutense de Madrid, Madrid, Spain Ana Sanz–Aguilar Institut Mediterrani d'Estudis Avançats IMEDEA–CSIC–UIB, Esporles, Spain Rafael Sardà Centre d'Estudis Avançats de Blanes CEAB–CSIC, Girona, Spain Ramón C. Soriguer Estación Biológica de Doñana EBD–CSIC, Sevilla, Spain Constantí Stefanescu Museu de Ciències Naturals de Granollers, Granollers, Spain Diederik Strubbe University of Antwerp, Antwerp, Belgium José L. Tellería Universidad Complutense de Madrid, Madrid, Spain Simone Tenan Institute of Marine Sciences CNR–ISMAR, National Research Council, Venezia, Italy Francesc Uribe Museu de Ciències Naturals de Barcelona, Barcelona, Spain José Ramón Verdú CIBIO, Universidad de Alicante, Alicante, Spain Carles Vilà Estación Biológica de Doñana EBD–CSIC, Sevilla, Spain Rafael Villafuerte Instituto de Estudios Sociales Avanzados IESA–CSIC, Cordoba, Spain Rafael Zardoya Museo Nacional de Ciencias Naturales MNCN–CSIC, Madrid, Spain

Animal Biodiversity and Conservation 45.1 (2022)

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New data on the status and ecology of a galliform at risk of extinction: the Pyrenean grey partridge (Perdix perdix hispaniensis) in the Iberian System (Soria, Spain) J. C. Ceña, A. Ceña, V. Salvador–Vilariño, J. M. Meneses, C. Sánchez–García Ceña, J. C. Ceña, A., Salvador–Vilariño, V., Meneses, J. M., Sánchez–García, C., 2022. New data on the status and ecology of a galliform at risk of extinction: the Pyrenean grey partridge (Perdix perdix hispaniensis) in the Iberian System (Soria, Spain). Animal Biodiversity and Conservation, 45.1: 1–12; Doi: https://doi. org/10.32800/abc.2022.45.0001 Abstract New data on the status and ecology of a galliform at risk of extinction: the Pyrenean grey partridge (Perdix perdix hispaniensis) in the Iberian System (Soria, Spain). A study was conducted in 2008–2010 to gain knowledge on the status and ecology of the endangered subspecies of grey partridge (Perdix perdix hispaniensis), at its southernmost range edge. From an historic breeding range of 28,300 ha, 15 different coveys (adults with juveniles) were observed in an area comprising 5,550 ha, with an estimated minimum autumn population size of 103–113 birds and a maximum of 163–181 birds. Spring pair density was estimated at 2.3 pairs/1,000 ha, and when considering only coveys, 6.8 partridges/1,000 ha. The majority of birds were located at an altitude above 1,690 m a.s.l., mainly in mountain shrubland (especially Calluna vulgaris and Erica spp.). Habitat loss was the most important threat for the species' conservation. In conclusion, efforts should prioritize urgent habitat recovery and monitoring in order to change the fate of the species.

Key words: Conservation, Habitat selection, Management, Radio–tracking Resumen Nuevos datos sobre la situación y la ecología de un ave galliforme en peligro de extinción: la perdiz pardilla (Perdix perdix hispaniensis) en el sistema Ibérico (Soria, España). Se realizó un estudio entre 2008 y 2010 para conocer la situación y la ecología de la perdiz pardilla (Perdix perdix hispaniensis), que se encuentra en peligro de extinción, en el extremo meridional de su distribución geográfica. De una superficie de reproducción histórica de 28.300 ha, se confirmó la observación de 15 grupos familiares (adultos con juveniles) en una superficie de 5.550 ha con un tamaño de población estimado en otoño de entre 103 y 113 perdices como mínimo y de entre 163 y 181 como máximo. Se estimó una densidad de 2,3 parejas/1.000 ha en primavera y, al considerar únicamente grupos familiares, de 6,8 perdices/1.000 ha. La mayoría de las aves se encontraban a una altitud superior a 1.690 m s.n.m. y usaban principalmente matorrales de montaña (especialmente Calluna vulgaris y Erica spp.). La pérdida del hábitat fue el factor más perjudicial para la conservación de la especie; en este sentido, se concluyó que se debería dar prioridad a la recuperación urgente del hábitat y al seguimiento, lo que podría cambiar el destino de la especie.

Palabras clave: Conservación, Gestión, Selección de hábitat, Radioseguimiento Received: 10 III 21; Conditional acceptance: 17 V 21; Final acceptance: 11 X 21 J. C. Ceña, A. Ceña, Barrio del Medio 18, La Póveda, 42169 Soria (Spain).– V. Salvador–Vilariño, Servicio de Espacios Naturales, Flora y Fauna, Junta de Castilla y León, c/ Rigoberto Cortejoso 14, 47014 Valladolid (Spain).– J. M. Meneses, Servicio Territorial de Medio Ambiente de Soria, Junta de Castilla y León, c/ Linajes 1, 42001 Soria (Spain).– C. Sánchez–García, Fundación Artemisan, Avda. Rey Santo 8, 13001 Ciudad Real (Spain).

Corresponding author: C. Sánchez–García. E–mail: investigacion@fundacionartemisan.com ORCID ID: C. Sánchez–García: 0000-0002-0693-6411

ISSN: 1578–665 X eISSN: 2014–928 X

© [2022] Copyright belongs to the authors, who license the journal Animal Biodiversity and Conservation to publish the paper under a Creative Commons Attribution 4.0 License.

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Introduction The grey partridge subspecies Perdix perdix hispaniensis (hereafter referred to as Pyrenean grey partridge, PGP), is a galliform occurring in three mountain ranges, the Cantabrian Mountains and the Iberian System in Spain, and the Pyrenees between Andorra, France and Spain (Lucio et al., 1992). As these mountains are isolated, they represent the south–western limit of grey partridges occurrence within its entire range (Potts, 2012). The PGP shows genetic and phenotypic characteristics (Castroviejo, 1967; Lescourret et al., 1987; Martin et al., 2003; Bech et al., 2020) that differentiate it from other subspecies, and it inhabits uplands from 1,300 to 2,700 m a.s.l., depending on the time of the year. It selects habitats dominated by shrublands and steep slopes (Lucio et al., 1992), in contrast with the vast majority of the species’ range in Eurasia and North America, where grey partridges mainly occur in lowland and farmland habitats (Potts, 1986). PGP form pairs in winter–spring and family coveys from summer to winter, as do their lowland counterparts (Potts, 2012).

Research has covered aspects of the biology and ecology of PGP (Llamas and Lucio, 1988; Novoa et al., 1999, 2002, 2006), demographics (Junco Ruiz and Kilchenmann, 1998; Bro and Crosnier, 2012), controlled burning effects (Novoa et al., 1998), monitoring (Novoa, 1992; González et al., 2017) and hunting (Besnard et al., 2010). PGP is a game species in the French Pyrenees (Novoa et al., 2008) and Catalonia (Pagés, 2011) and both populations show favourable conservation status when compared to others (Martínez–Vidal, 2011).

While the population from the Pyrenees is relatively well–studied, there is a lack of research for PGP occurring in the Cantabrian Mountains and the Iberian System. According to the review conducted by Purroy and Purroy (2016), the majority of studies on PGP from these populations were conducted before the year 2000. Few studies have provided knowledge on the species in the last 20 years (Herrero et al., 2009) with the exception of monitoring conducted by regional wildlife departments.

In Spain, the PGP population size was estimated at 2,000 to 6,000 pairs in the late 1990s (Lucio and Sáenz de Buruaga, 1997). It is included in Annex I within the Directive 2009/147/EC of the European Parliament and of the Council on the conservation of wild birds and the species is categorized as 'Vulnerable' (VU C1), owing to the decline recorded during the last decades (Onrubia et al., 2004). This unfavourable trend seems to be especially severe in some peripheral areas of the Cantabrian Mountains and in the entire population of the Iberian System (Lucio et al., 1992), where local extinctions had occurred in recent times. In December 2020 the Iberian System population was declared 'at risk of extinction' by the Spanish Red List of Endangered Species (Orden TED/1126/2020).

Aiming to gain knowledge on the status and ecology of PGP in the Iberian System, we conducted field surveys and radio–tracked birds in the Soria province (where the species is known as 'Serreña'),

the southernmost range limit of the Iberian System population (Lucio et al., 1992). Our results may help to understand conservation problems and develop targeted management to improve the status of PGP populations.

Material and methods Study area This research was conducted in the Iberian System from 2008 to 2010, in the province of Soria (Castilla y León region, fig. 1). PGP had historically occurred in the study area in open mountainous areas dominated by shrublands, ranging from 1,400 to 2,000 m a.s.l.

The distribution of PGP in the Soria province is spread across six mountain ranges: Urbión, Cameros, Sierra Cebollera, up–River Tera, up–River Cidacos–Alhama and Sierra del Moncayo. Most of this area is included in the Special Protection Areas (SPA) of 'Sierra de Urbión' (ES4170013) and 'Sierra del Moncayo' (ES4170044).

PGP surveys We aimed to identify the areas where PGP occurred during the breeding season (from March to October), and to gain knowledge on the breeding success of the species, paying special attention to coveys (adults with juveniles, family groups). We were aware of the possible movements during autumn–winter to areas of lower altitude (especially after heavy snow) (Lucio et al., 1992) but it was not possible to cover all these areas; however, they were partially studied using radio–tracked birds.

We considered the areas of historic occurrence of PGP and created 283 UTM individual grids of 1 km2. A first visit was conducted to identify whether the habitat was suitable for PGP breeding, excluding grids where the only habitat was dense forest (Purroy and Purroy, 2016). At the same time, we conducted face–to–face surveys (n = 210) with wildlife rangers, game managers, hunters, shepherds, ornithologists and bird watchers in all the study area to evaluate past and current presence of PGP. After the first visits and face–to–face surveys, we estimated that the PGP potential breeding habitat area in Soria province comprised 12,500 ha (44 % of the historical range).

Field work was thus focused on this area. After considering the land that could be surveyed in one day of field work, we created 57 survey plots (size range: 150–250 ha) covering a total of 12,500 ha.

We combined several methods to detect PGP during 2008 and 2009: (1) Playback calls from March to June, following the methods of Novoa (1992) in a total area of 3,100 ha that was chosen after considering the first visits and information gathered from the face–to–face surveys; playback–calls from birds from the same area were used.

(2) Walked transects in July and August covering all survey plots. Each transect was surveyed at least

Animal Biodiversity and Conservation 45.1 (2022)

3

La Rioja

PGP distribution

Burgos

Soria Zaragoza 10

0

10

20

30

40 km

Fig. 1. Study area in the Iberian System, showing 10 x 10 km grids covering the former and current distribution of grey partridges (Perdix perdix hispaniensis) in the province of Soria (showing the neighbouring provinces) (Onrubia et al., 2003). As the species is 'at risk of extinction' it is not possible to show the current distribution of the species at a lower scale.

Fig. 1. Zona del estudio en el sistema Ibérico, donde se muestran las cuadrículas de 10 x 10 km que cubren el área de distribución histórica y la actual de la perdiz pardilla (Perdix perdix hispaniensis) en la provincia de Soria (se muestran las provincias limítrofes) (Onrubia et al., 2003). Dado que la especie se ha declarado como "en peligro de extinción", no es posible mostrar su distribución actual a menor escala.

once by three observers, aiming to detect signs of PGP presence (tracks, faeces, feathers), make direct observations, and gather data on habitat characteristics and possible threats, either natural or anthropogenic such as natural forest expansion, physical structures (tracks, roads and wind farms), and sources of disturbance from human activities; transects were conducted during the morning (9:00–12:00 a.m.).

(3) Pointing dogs in September and October. We hired a professional dog–trainer to cover those areas where coveys may occur (considering the information gathered from previous methods, and covered 2,200 ha using 4–5 dogs (English setter) and 3–4 observers during the morning (from 9:00–12:00 a.m.).

When PGP were seen or flushed, we attempted to distinguish between adults and juveniles, and estimate the age of juveniles.

And (4) transects in winter: this method was used in areas with at least 90 % of surface snow (23 transects in total, mainly up–River Tera and up–River Cidacos– Alhama). Access was difficult (González et al., 2017), and the PGP could be confounded with red–legged partridges (Alectoris rufa).

Survey plots were categorized into three categories: (1) 'confirmed breeding', when coveys were seen/ flushed; (2) 'probable breeding', when indirect evidence was found such as pairs; 'incubation faeces' and small faeces attributed to juveniles (but no juveniles observed), together with adult 'distraction displays'; and (3) 'not detected'.

The number of days per method invested in each mountain range varied according to surface and habitat characteristics. In total we conducted surveys on 248 days (table 1). The efficiency of each method was evaluated by conducting surveys in the plots with known presence of the radio–tagged birds and coveys (associated to these birds) (table 2). As red–legged partridges occurred in the areas, we also recorded birds seen or flushed.

Radio–tracking and home range analysis During May 2008 and 2009, we captured PGP for radio–tracking using a corvid cage–trap and a female PGP decoy from a French game farm (fig. 2). Once the presence of birds was confirmed through the play-

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Table 1. Grey partridge surveys conducted in the different mountain ranges, showing the number of survey plots, surface area covered and number of surveys conducted per method, together with density of spring pairs, coveys in autumn and size of the area where breeding was confirmed: T, total; Ur, Urbión; Ca, Cameros; Ce, Cebollera; Up–T, up–River Tera; Up–C; up–River Cidacos–Alhama; Mo, Moncayo; n.a., not available.

Tabla 1. Muestreos de perdiz pardilla realizados en los distintos macizos montañosos, donde se puede observar el número de sectores y la superficie de muestreo, el número y el tipo de muestreos, la densidad de parejas en primavera y de bandos en otoño y la superficie donde se había producido reproducción.

(Para las abreviaturas de los distintos macizos montañosos, véase arriba).

T

Ur

Ca

Ce

Up–T

Up–C

Mo

57

11

2

9

13

15

7

10,300

1,800

500

1,800

1,900

2,500

1,800

Face–to–face surveys (days)

21

3

2

3

4

6

3

Playback calls

72

8

4

10

20

25

5

Walked transects

76

10

6

14

20

21

5

Pointer dogs

56

12

5

11

12

12

4

Transects in winter

23

1

3

3

7

8

1

Total

248

34

20

41

63

72

18

Spring pairs/1,000 ha

2.3

3.8

2

2.7

3.1

1.6

0.5

Number coveys

15

5

1

3

2

4

0

5,550

1,800

300

1,650

1,100

500

200

6.8

14.14

6

7.2

4.2

8

n.a.

Survey plots Breeding habitat surveyed (ha)

Confirmed breeding (ha) Birds in coveys/1,000 ha (autumn)

Table 2. A, efficiency of the survey methods for grey partridges, showing the number of surveys conducted per method with the real presence of radio–tracked birds or coveys (n) and the percentage of surveys with confirmed detection (%): Rt birds, radio–tracked birds; n.c., not conducted. B, size of the PGP population in Soria province, considering the detectability percentages: Min, minimum; Max, maximum; D, detectability; * calculated considering the average number of juveniles per covey (3.5).

Tabla 2. A, eficiencia de los distintos métodos de muestreo de perdiz pardilla, donde se observa el número de muestreos realizados de cada método, la presencia real de aves marcadas con emisores de radiofrecuencia o de grupos familiares (n) y el porcentaje de muestreos con detecciones confirmadas (%): RT birds, aves marcadas; n.c., no realizado. B, tamaño de la población de perdiz pardilla en la provincia de Soria, considerando los porcentajes de detectabilidad: Min, mínimo; Max, máximo; D, detectabilidad; *calculado teniendo en cuenta el número medio de juveniles por grupo familiar (3,5).

A

B

Rt birds

Coveys

Min

n

%

n

%

Coveys

15

18–20 75–85

Face–to–face surveys 60

20

60

20

Breeding adults

17

20–23

Playback calls

4

75

n.c.

n.c.

Probable breeding adults* 6–8

Walked transects

20

5

8

38

Unsuccessful breeders 30–35

72–81 40–45

Pointer dogs

4

50

8

88

Juveniles

63–68*

Transects in winter

6

16

6

33

Total

50–53

Max

D(%)

8–9

103–113 163–181

Animal Biodiversity and Conservation 45.1 (2022)

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A

B

C

D

Fig. 2. A covey seen during the surveys in autumn (A), a single bird detected with the playback call (B), the cage–trap with the live decoy (C), and the first bird caught for radio–tracking (D).

Fig. 2. Un grupo familiar observado en otoño (A), una perdiz 'soltera' detectada mediante la emisión de cantos (B), la trampa de jaula con un reclamo vivo (C) y la primera ave capturada para marcarla con un emisor de radiofrecuencia (D).

back calls, cage–traps were set and checked daily (Besnard et al., 2010).

The age of the birds caught (n = 2) was determined by inspection of primary feathers and biometrics were taken (wing length, weight, tarsus width). Birds were ringed and fitted with collar radio–tags (8 g, RI–2D– M, Holohil Systems Ltd., Ontario, Canada), with an approximate life–span of 300 days and mortality switch.

Birds were released on the same day at the same location where they had been caught, and radio–tracking lasted until the battery ended. Radio–tracking was more frequent in May–November (from 1 to 5 fixes/ week). During the snow period (December–March) it was difficult to access the study areas and radio– tracking was conducted once every fortnight. While conducting radio–tracking, we tried to observe the bird when possible, especially during the breeding season to assess whether it was paired or with other adults.

Locations of radio–tracked birds were georeferenced after conducting triangulation, and home ranges were calculated using the minimum convex polygon (arithmetic mean algorithm excluding 5 % of fixes from the harmonic centre, MCP 95 %) (Harris et al., 1990) using

ArcView © (program version 3.2., Esri, Redlands, CA).

When possible, we calculated movements between consecutive days (daily inter–fix distance).

Habitat use We recorded the habitat characteristics of the location of PGP coveys (habitat used), considering a 200 m buffer from the point where they were seen/flushed, recording in situ the main habitats and the canopy cover fractions of the vegetation present. Assuming that birds could have moved while being approached, a further GIS analysis was conducted considering a 500 buffer to evaluate habitat availability in the context of each survey plot. Rather than using broad habitat categories, we decided to identify and evaluate the dominant (≥ 50 % of the surface) habitat in the 500m buffer as follows: Calluna vulgaris, Erica (arborea or australis ssp. aragonensis), Cytisus scoparius, Vaccinium myrtillus, Rubus ideaeus and Pinus spp., together with grassland (mainly Poaceae), and including 'screes' as a category. We also considered

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Table 3. Habitats used by untagged PGP coveys in Soria province (from the analysis of the 200 m buffer and dominant habitat 500 m buffer): N, covey number; BR, bilberry/raspberry (Vaccinium myrtillus/Rubus idaeus); C, Pyrenean broom Cytisus oromediterraneus; F, forest; Fp, forest plantation; G, grassland (mainly Poaceae); H, heather Calluna vulgaris; HR, heath Erica spp.; J, junipers Juniperus spp.; S, screes.

Tabla 3. Hábitats utilizados por grupos familiares sin marcar de perdiz pardilla en la provincia de Soria (a partir del análisis de la zona de influencia de 200 m) y el hábitat dominante (zona de influencia de 500 m): N, número degrupo familiar; BR, arándano y frambuesa Vaccinium myrtillus y Rubus idaeus); C, piorno Cytisus oromediterraneus; F, bosque; Fp, bosque de repoblación; G, pasto (principalmente Poaceae); H, biércol Calluna vulgaris; HR, brezo Erica spp.; J, junípero Juniperus spp.; S, canchal o pedregal.

N

200 m buffer

500 m buffer

N

200 m buffer

500 m buffer

1

HR–J–S

HR

9

HR–S–Fp

HR

2

HR–BR–S

HR

10

HR–G–Fp

HR

3

HR–J–S

HR

11

H–S–Fp

H

4

H–HR–BR

HR–H

12

HR–J–G

HR–C

5

HR–J–S

HR

13

H–C–F

H–C

6

H–HR–F–G

HR–H

14

HR–C–G

HR–C

7

H–G

H

15

C–HR

C

8

H–S–BR

H

combinations of habitats when they were present in similar proportions. The number of locations of birds seen during the surveys (either as individuals or in coveys) was limited (< 60 during the whole study, and sometimes at the same location). We therefore conducted a descriptive analysis, pooling all habitat locations of seen/flushed coveys, and compared the habitat where birds were seen with the dominant habitat (table 3). With regard to radio–tracked birds, we recorded the habitat characteristics as described for birds seen in the surveys. We also recorded the percentage of cover and the height of the dominant vegetation (cm) (Novoa et al., 2002), together with altitude and slope, calculated as percentages.

Owing to the small number of radio–tracked birds, we merged data and presented the main results for both birds.

Results PGP surveys From an initial potential breeding area of 12,500 ha, after surveys we estimated that 10,300 ha were suitable, with breeding confirmed in 5,550 ha, representing 54 % of the species’ suitable habitat in Soria province (table 1). Out of 57 plots surveyed, PGP occurred in 34, and possible breeding was detected in 24 (confirmed in 15 and probable in 9). Additionally, we documented evidence of breeding during the last 40 years in 21 plots, and during the last 10 years in 12 plots where no PGP were detected during the surveys (33 in total, fig. 3). The most efficient survey methodologies, after comparisons considering plots

with known presence of radio–tagged birds or coveys, were playback calls for unsuccessful breeders and pointer dogs for coveys (table 2). Combining methods, we estimated that the detectability of unsuccessful breeders was 40–45 % and for coveys 75–85 %.

During 2008 and 2009, we detected 15 coveys at the study area. A total of 103–113 birds were observed in autumn as follows; 17 breeding adults, 6–8 probable breeding adults, 30–35 unsuccessful breeders and 50–53 juveniles. Considering detectability, we calculated a maximum autumn population size of 163–181 PGP (and minimum of 53–60 in spring) (table 2). For the whole study, mean partridge density in the breeding habitat (considering plots with confirmed and probable breeding) was 2.3 pairs/1,000 ha, and when considering birds seen in autumn coveys it was 6.8 partridges/1,000 ha (table 1). The average number of juveniles per covey was 3.5 (range 2–7).

Using the number of birds seen during the counts with pointers, (53 juveniles and 17 adults), we found the age ratio was 3.1.

With regard to phenology, the observation of faeces from incubating birds and the size of the juveniles when seen/flushed allowed us to estimate that in 2008 the hatching period lasted from July 10th to the end of August. We observed chicks of small size at the beginning of September, probably from second or third nesting attempts. In 2009, first hatchings were estimated to occur by June 25th, with no evidence of late clutches.

Habitat use Overall, coveys were observed at a mean altitude of 1,781 m a.s.l. (range 1,550–2,010 m) and non–breeding birds at a mean altitude of 1,690 m a.s.l. (range

Animal Biodiversity and Conservation 45.1 (2022)

16 14 12

7

No detection Probable breeding Confirmed breeding

10 8 6 4 2

0

Ur

Ca

Ce Up–T Mountain ranges

Up–C

Mo

Fig. 3. Number of survey plots considering the mountain ranges where PGP breeding was confirmed, probable breeding occurred and plots with no detections: Ur, Urbión; Ca, Cameros; Ce, Cebollera; Up–T, up–River Tera; Up–C; up–River Cidacos–Alhama; Mo, Moncayo.

Fig. 3. Número de sectores considerando los macizos montañosos donde se había producido reproducción, donde la reproducción era probable y donde no se detectaron individuos. (Para las abreviaturas de los macizos montañosos, véase arriba).

1,525–2,000 m). The dominant habitat in the survey plots where coveys occurred is shown in figure 4. With the exception of one survey plot, the dominant habitat was heather Calluna vulgaris, heath Erica spp., their combinations or combinations of heather and Pyrenean broom (Cytisus oromediterraneus). All coveys showed use of heather, heath or both categories, with other categories in minor proportions: screes (40 % of the coveys), junipers (26 %), grassland (33 %), bilberry (Vaccinium myrtillus) and raspberry (Rubus idaeus), forest plantations and Pyrenean broom (20 %) and natural forest (13 %) (table 3, fig. 1s in supplementary material).

At the 24 plots where breeding birds occurred, we established different levels of extinction risk considering habitat quality, the number of threats identified and their impact, recording high risk of extinction in 6 plots (25 %), high–medium risk in 9 (37.5 %), and low–very low risk in the remaining 9 (37.5 %). The most important threats considering those of high impact within the 24 plots were: natural forest expansion (n = 9), pine tree plantations (n = 7), tracks and roads (n = 7), fragmented and reduced habitat (n = 7), wind farms (n = 5), red–legged partridge walked–up shooting at hunting grounds within (or very close to) plots where accidental/illegal hunting may occur (n = 4), human disturbances related to outdoor activities coming from tracks and roads (n = 4) and pigeon hunting (n = 3) (fig. 2s in supplementary material). We detected 21 different coveys of red–legged partridges, which were observed at a mean altitude of 1,610 m (range 1,450–2,050 m). We recorded a mean brood size of 5.5 (range 3–9). Red–legs showed a different pattern

of habitat use compared to PGP, as they were detected in open habitats such as grassland and crops, but also in shrublands and even forest.

Movements and home ranges The cage–traps were set in the mountains of Cameros, and we caught two adult male birds for radio–tracking: the first in May 2009 ('bird 1' after 32 days of trapping) and the second in May 2010 ('bird 2' after 7 days of trapping) (fig. 2).

For both birds, radio–tracking lasted until the battery ended (approximately 300 days), and 77 % of fixes were recorded within 100 m from bird's real location.

Neither of the birds paired up, though bird 1 joined a covey of 5 PGP in autumn. Both birds were located in areas where breeding was confirmed.

We obtained 164 locations for bird 1 from May– November, and during this period the bird had a total home range of 1,770 ha. However, bird 1 used mainly three areas and conducted three movements between them (4.5, 7.1, and 10.1 km), hence we calculated home range values (MCP 95 %) for each of the areas: 64.43, 38.57 and 0.12 ha respectively. For bird 2, we obtained 84 locations during the same period, and the bird used a main area of 60 ha, conducting two movements of 8.5 and 4 km within the same area, and staying in the upper parts of the mountains for brief periods of time (< 7 days) (table 4). Regarding shorter movements between consecutive days (daily inter–fix distance, n = 52), the average distance merging both birds was 220 m (range 0–1,105 m).

Birds were located at an altitude between 1,790 m

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14 Probable breeding Confirmed breeding

12 Survey plots

10 8 6 4 2

0

A

B

C D Dominant habitat

E

F

Fig. 4. Dominant habitat in the survey plots where confirmed and probable breeding occurred: A, Calluna vulgaris; B, Erica spp.; C, Calluna–Erica spp.; D, Calluna–Cytisus; E, Erica–Cytisus; F, Cytisus.

Fig. 4. Hábitat dominante en los sectores en los que se registró reproducción segura o probable. (Para las abreviaturas de hábitats dominantes, véase arriba).

and 2,000 m, and in winter both birds remained in areas covered by snow, including summits, and conducted small altitudinal movements. The exception was a location of bird 1 during a brief period of time at 1,500–1,600 m, after heavy snow. Radio–tracking continued during the winter months, but owing to the limited access, fixes were conducted once every fortnight. The habitat used by radio–tracked birds was in general mountain shrublands with a height higher than 50 cm and a percentage cover higher than 70 %. In 51 % of locations the habitat used was heath (mainly E. australis and but also E. arborea), followed by heather (36.5 %), Pyrenean brooms (8 %) and the remaining 4.5 % were locations in open habitats, i.e. grassland and screes. No locations were recorded in dense forest. Birds were mainly located on slopes ranging in incline from 25–50 % (44 % of locations) and 10–25 % (35 % of locations), with the remaining 21 % of locations on slopes shallower than 10 %, with no locations above 50 %.

Discussion The population of PGP in the Soria province has undergone a marked decline, as at the time this study was conducted, the species occurred in 20 % of the historic range. It was estimated that from 1998–2008, the species became extinct in 30% of the territory. The fact that breeding was confirmed in only 5,550 ha is concerning and supports the recent change of conservation status to 'at risk of extinction'. We do not know whether this negative trend has continued, but the last monitoring conducted in 2017 in the mountains of Urbión confirmed the occurrence of the species in half of the survey plots where birds were detected during the study presented here.

We agree with previous studies evaluating the PGP populations in Spain and in Castilla y León (Lucio et al., 1992; Robles et al., 2002), which also suggested that the conservation status of the Iberian System PGP population was worse than that of the populations in the Pyrenees and the Cantabrian Mountains. As is well known for other species at the limit of their range (Sexton et al., 2009), PGP in the Soria province would be especially sensitive to habitat changes and disturbance, and hence the conservation of these populations should be prioritized.

In the breeding habitat of Soria province where PGP still occur, the mean spring partridge density was 2.3 pairs / 1,000 ha, and the autumn density was 6.8 birds/1,000 ha. These values were lower than those recorded in the north–western part of the Iberian System (Burgos province) by Ansola et al. (1990), 7.5 pairs/1,000 ha, and far from the values for the Cantabrian Mountains and the Pyrenees, which are around 10–20 spring pairs/1,000 ha and above 100 birds/1,000 ha in autumn (Birkan and Jacob, 1988; Llamas and Lucio, 1988; Novoa et al., 2008). The age ratio was 3.1, the average number of juveniles per covey was 3.5, and both values were within the range of those recorded in the Pyrenees (Novoa, 1999; Novoa et al., 2008), where in recent decades the age ratio lay between 1.3 and 4.7 young per adult. Interestingly, around 50 % of adult birds detected did not reproduce, and breeding was not confirmed in the two radio–tracked birds. We do not know whether this was a year effect, but we cannot exclude that this population is affected by a demographic trait hampering breeding success, possibly attributed to isolation and the lack of exchange of individuals from other sub–populations (Eberhard, 1991).

The combination of survey methods proved effective at detecting birds in our study site, with the most

Animal Biodiversity and Conservation 45.1 (2022)

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Table 4. Main results from radio–tracked birds, including the characteristics of the sites where birds were located: * bird located outside site D, but it was not possible to calculate home range.

Tabla 4. Principales resultados obtenidos de las aves marcadas con emisores de radiofrecuencia y características de los lugares en los que fueron localizadas: * ave localizada fuera del lugar D y cuya área de campeo no se pudo calcular.

Bird 1 Bird 1 Catching date

Bird 2

May–2009 May–2010

A

B

Bird 2 C

D

Home range (ha) 64.63 38.57 0.12

64.63

300

300

Altitude (m)

1,801

Number fixes (May–Nov) 164

84

Period (days)

Survival (d)

1,790 1,897 1,839 211

82

7

300*

effective methods being playback calls and pointing dogs. This is not surprising as PGP is elusive and inhabits mountains in which walked transects may not be practical, though a study conducted on Cantabrian PGP showed that snow transects were efficient as long as paved roads existed (González et al., 2017).

Hence, those aiming to conduct accurate PGP surveys should prioritize playback calls in spring and pointing dog surveys in summer–autumn, while not discarding snow transects.

The majority of PGP detected during the surveys and the radio–tracked individuals were located above 1,690 m, with few locations in lower altitudes, except for movements during the breeding season and heavy snowfall (as recorded for the radio–tracked bird 1). The historic data from the Iberian System in this regard is limited, with observations from 1,100 to 1,960 m a.s.l., and the higher pair densities at a range of 1,700–1,800 m recorded 30–40 years ago (see the review in Lucio et al., 1992). It seems that PGP are now distributed in upper areas, especially coveys which may need the best possible habitat. This finding could be related to the habitat loss and fragmentation in lower areas.

As shown by Ansola et al. (1990), PGP mainly used mountain shrubland, i.e. Calluna vulgaris, Erica spp. Cytisus and their associations. Moreover, radio– tracked birds selected E. australis and E. arborea, followed by Calluna vulgaris, and these shrublands were at least 50 cm in height and the shrub canopy cover was > 70 %.

Neither the radio–tracked birds nor the PGP seen during the surveys used open habitats (grassland) and forest, despite sparse trees being present within the areas where they occurred. According to the review of Purroy and Purroy (2016), the dominant vegetation in the PGP habitats in the Cantabrian mountains are broom (Genista polygaliphylla and G. obtusirramea) and fern (Pteridium aquilinum), together with Erica australis ssp. aragonensis and Daboecia cantabrica, while in the eastern Pyrenees breeding PGP select open and dense shrublands, including Cytisus purgans and Juniperus communis with a canopy cover higher

than 40 % (Novoa et al., 2002). From our results, it is clear that PGP habitat requirements in the Iberian System differ slightly from those in other subpopulations.

Although only two birds were radio–tracked, the results provide valuable knowledge on PGP ecology, being the first published data of radio–tagged PGP outside the Pyrenees. In both birds, the radio–tracking was conducted until the battery ended, indicating 100 % survival during the study period. With just two birds it is difficult to draw conclusions as previous studies have shown that the species can be affected by high levels of natural and non–natural mortality (Besnard et al., 2010). Regarding the home range of our birds, rigorous comparisons with data from Pyrenean birds are difficult because of the different study periods, as Novoa et al. (2006) calculated a MCP for spring pairs ranging from 118 to 126 ha (March– September), and core areas of 6.2–14.4 ha, finding differences before and after hatching and between paired and 'unpaired' birds. However, our results on movements are similar to those calculated by Novoa et al. (2006), who reported daily interfix distances ranging from 126 to 249 m, and also recorded longer movements up to 20 km (Novoa, 1999). As suggested by the studies from the Pyrenees, we agree that the long movement distances in spring/summer could be explained not only by migration from the wintering to the breeding habitats, but also by weather changes, though we detected only one movement related to heavy snowfall. Interestingly, in winter and during the snow period (when radio–tracking was difficult to conduct), birds tended to remain on summits and conduct short altitudinal movements. To obtain a better understanding of their behaviour, more birds should be radio–tracked (including hens and juveniles).

During the study, we identified anthropogenic threats within the survey plots. These included forest tracks, wind farms, hunting grounds and pigeon hunting, all possibly leading to disturbance and direct mortality at certain times of the year. These threats were already identified decades ago for the Iberian System, and they have been also identified in other PGP populations (Lucio et al., 1992). The overlap with

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red–legged partridges at certain areas could have conservation implications for PGP as we cannot rule out the possibility that direct mortality and disturbance may occur on hunting grounds in which walked–up shooting is conducted. Hunters could perhaps be involved in the conservation of PGP through monitoring with pointers, as conducted in other PGP populations in France and Spain. Finally, the range overlap of these galliforms could favour inter–specific competition, as suggested in farmland habitats of France where red–legged and grey partridges occur and compete (Rinaud et al., 2020). However, this hypothesis would have to be further explored for the Iberian System.

Undoubtedly, the most important threat for PGP in Soria province was habitat loss and fragmentation.

We confirmed that the historic breeding habitat area of 28,300 ha has been reduced to 10,300 ha, and just a half of this area held PGP at the time this study was conducted. This might be explained by the forest plantation policy conducted in many areas 20–50 years ago, together with the natural growth of existing forest, which has dramatically reduced the optimal breeding habitat for PGP in its entire range in Soria province, which is similar to the situation in other parts of the Iberian System (Lucio et al., 1992).

Radio–tracked birds clearly used mountain shrubland, now restricted to 'patches' surrounded by vast areas of unfavourable habitats, dominated by pine tree forest and grasslands, with poor or absent shrub coverage. Partridges moved between these patches, sometimes covering several km, using summits of high altitude as 'stopovers' where they were frequently observed. The severe habitat changes in recent decades have dramatically reduced the former suitable habitat. Partridges are now restricted to sub–alpine areas, avoiding areas that were suitable in the past, such as valleys and hills, where the species rarely occurs nowadays. Additionally, we observed that in some locations, mountain shrubland had been recently cleared to increase available grass for grazing cattle (supported by public subsidies), ultimately reducing the key habitat for PGP in the Iberian System. However, it has been shown that in other PGP territories where dense shrubland has been cleared, partridges are favoured by a more diverse habitat (Lucio et al., 1992). It is true that among the mountain ranges considered, the conservation status of PGP was better in Urbión and Cameros, possibly because these areas hold more favourable habitats than the others, and in the Moncayo, the species could be on the verge of extinction (or is already extinct) owing to the fact that this mountain is completely isolated from the others.

From 2009 onwards, several habitat management interventions dedicated to PGP conservation at small scale (< 100 ha) and promoted by the regional government, have been conducted at the study site and could be replicated in the current and former PGP distribution area. These interventions have consisted of: (1) protection of mountain shrubland (mainly Calluna vulgaris and Erica spp.) where breeding PGP occur; (2) clearing of planted pine tree forest to restore the former optimal shrubland; (3) clearing of pine trees in summits which act as natural corridors between the

different breeding areas; (4) restricted track access to vehicles within the breeding areas during the breeding season; and (5) dissemination campaign targeting hunters and other stake–holders who conduct activities in the breeding areas (conferences, project brochure and poster) (fig. 3s in supplementary material).

Conclusion In summary, to avoid grey partridge extinction in the Iberian System, the priority should be to protect the mountain shrublands in those areas where they still occur. To increase optimal habitat, it would be possible to restore shrublands in those areas that have been planted with pine trees or changed to grassland for livestock in recent times. Because the regional government in Soria has already conducted targeted management actions for the species and as the former distribution area has been mapped and fully analysed, the key areas in which actions must be taken are already known. From a practical point of view, efforts to restore the habitat of the PGP remain worthwhile and may be the most cost–effective way to halt the population collapse and increase the chances of its recovery. Additionally, monitoring should be conducted to evaluate the short– and mid–term effects of these actions on the population dynamics of the species, together with further studies to gain knowledge on the biology and ecology of this endangered galliform.

Acknowledgements This study was promoted and fully funded by the regional government of Castilla y León (Consejería de Medio Ambiente, Junta de Castilla y León, Spain) who also provided permission for the study and an approved animal welfare protocol. The authors would like to thank all the people who collaborated in the face–to–face surveys and especially M. Barbero and J. Arrieta for their help with pointer dogs. We are indebted to C. Novoa for his help in several tasks of the study, and J. L. Guzmán for producing the map. Special thanks are given to the reviewers who provided valuable comments and suggestions, and R. Burrell for proofreading. This study is dedicated to the memory of the late G. R. 'Dick' Potts (1939–2017), who helped with the literature review and encouraged the writing of this article.

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Human disturbance modifies the identity and interaction strength of mammals that consume Attalea butyracea fruit in a neotropical forest C. M. Delgado–Martínez, E. Mendoza

Delgado–Martínez, C. M., Mendoza, E., 2022. Human disturbance modifies the identity and interaction strength of mammals that consume Attalea butyracea fruit in a neotropical forest. Animal Biodiversity and Conservation, 45.1: 13–21, Doi: https://doi.org/10.32800/abc.2022.45.0013 Abstract Human disturbance modifies the identity and interaction strength of mammals that consume Attalea butyracea fruit in a neotropical forest. Habitat loss and hunting are important drivers of mammal defaunation, affecting not only species presence but also their ecological roles. Frugivory is a key biotic interaction in the tropics due to its wide representation among mammals and its effects on forest dynamics. We assessed how human disturbance affects interactions between mammalian frugivores and Attalea butyracea fruit deposited on the forest floor by comparing visits to palms at two sites with contrasting levels of human disturbance (non–disturbed vs. disturbed sites) in the Lacandon rainforest in southern Mexico. Using camera traps, we recorded mammal species interacting with fruit and estimated their interaction strength. The frugivore ensemble was richer in the non–disturbed forest (nine species) than in the disturbed forest (four species), which lacked the largest body–sized mammals. Large–bodied mammals showed a stronger interaction with fruit in terms of the frequency and length of their visits. Our study highlights the need to undertake conservation actions not only to ensure that the species are maintained in disturbed forests but also to ensure that their biotic interactions remain unchanged.

Key words: Habitat fragmentation, Mammalian frugivory, Ground–dwelling mammals, Large–sized fruit Resumen La alteración antrópica modifica el tipo de mamíferos que consumen frutos de Attalea butyracea en una selva neotropical y la intensidad con que lo hacen. La pérdida del hábitat y la caza son dos de las principales causas de la disminución de mamíferos, que no solo afecta a la presencia de especies, sino también a sus funciones ecológicas. La frugivoría es una interacción clave en las zonas tropicales debido a que se halla muy extendida entre los mamíferos y los efectos que ejerce en la dinámica de la selva. Mediante la comparación de las visitas realizadas por mamíferos a la palma Attalea butyracea en dos sitios con grados bien diferenciados de alteración antrópica (con alteración y sin alteración) de la selva Lacandona, en el sureste de México, evaluamos cómo afecta la alteración antrópica a la interacción entre los mamíferos frugívoros y los frutos que se acumulan en el suelo de la selva. Usando cámaras trampa, registramos a los mamíferos que interactuaron con los frutos y estimamos la intensidad de la interacción. Registramos una mayor riqueza de especies de mamíferos frugívoros en el sitio sin alteración antrópica (nueve especies) que en el sitio con alteración (cuatro especies), donde no se registraron los mamíferos de mayor tamaño. Los mamíferos de talla grande mostraron una interacción más intensa con los frutos en cuanto a la frecuencia y la duración de sus visitas. Nuestro estudio hace hincapié en la necesidad de adoptar medidas de conservación que permitan asegurar la presencia de las especies en los sitios con alteración antrópica, así como sus interacciones bióticas.

Palabras clave: Fragmentación del hábitat, Mamíferos frugívoros, Mamíferos terrestres, Frutos de gran tamaño Received: 3 II 21; Conditional acceptance: 7 IV 21; Final acceptance: 17 X 21

ISSN: 1578–665 X eISSN: 2014–928 X

© [2022] Copyright belongs to the authors, who license the journal Animal Biodiversity and Conservation to publish the paper under a Creative Commons Attribution 4.0 License.

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Delgado–Martínez and Mendoza

Carlos M. Delgado–Martínez, Instituto de Investigaciones sobre los Recursos Naturales, Universidad Michoacana de San Nicolás de Hidalgo, Av. San Juanito Itzícuaro s/n., Col. Nueva Esperanza, C. P. 58337, Morelia, Michoacán, México. Posgrado en Ciencias Biológicas, Universidad Nacional Autónoma de México, Av. Ciudad Universitaria 3000, C. P. 04510, Coyoacán, Ciudad de México, México.– Eduardo Mendoza, Instituto de Investigaciones sobre los Recursos Naturales, Universidad Michoacana de San Nicolás de Hidalgo, Av. San Juanito Itzícuaro s/n., Col. Nueva Esperanza, C. P. 58337, Morelia, Michoacán, México.

Corresponding author: Eduardo Mendoza. E–mail: eduardo.mendoza@umich.mx ORCID ID: Carlos M. Delgado–Martínez: 0000-0002-0913-932X; Eduardo Mendoza: 0000-0001-6292-0900

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Introduction

Material and methods

Frugivory by mammals is a characteristic and widespread ecological interaction in the tropics (Jordano, 2014). The high proportion of tropical trees that produce fleshy fruits and rely on vertebrates for their dispersal reflects its relevance (Howe and Smallwood, 1982; Danell and Bergström, 2002). Unfortunately, an increasing amount of studies show that human disturbances such as land–use change and hunting are affecting tropical frugivore communities (Markl et al., 2012; Fontúrbel et al., 2015). Major negative impacts of these threats concentrate on medium and large–bodied/specialist mammalian species while small–bodied/generalist species seem to deal better with human impacts (Vidal et al., 2013; Carreira et al., 2020).

Such impacts affects various components of the frugivory interaction, such as visitation rates and the number of fruits and seeds removed (Markl et al., 2012; Fontúrbel et al., 2015). In the long term, these effects can have a profound impact on the structure and regenerative potential of the forests (Harrison et al., 2013; Kurten, 2013). However, most evidence comes from the study of interactions that occur in the canopy forest, while comparatively less attention has been paid to assessing the anthropogenic impact on the interactions between mammals and the fruit deposited on the forest floor.

Camera trapping has mainly been used to record vertebrate presence so as to to estimate ecological parameters such as abundance and community diversity (Burton et al., 2015), but it is also well–suited for providing information that is useful for studying frugivory interactions (Miura et al., 1997). Examples of this include its usefulness in identifying the main visitors to fruit tree species (Jayasekara et al., 2007) and quantifying fruit–removal rates (Prasad et al., 2010). Thus, in comparison with studies that mainly relied on direct observations (Moegenburg and Levey, 2003) and indirect evidence of mammal activity (e.g., teeth marks on the hard parts of the seeds; Wright and Duber, 2001), the use of camera trapping has great potential to document the anthropogenic impact on the interaction between mammalian frugivores and fruit on the forest ground in greater detail (Galetti et al., 2015; Carreira et al., 2020).

In this study we used camera trapping to assess the characteristics of the frugivory interaction between medium and large–bodied (> 500 g) mammalian frugivores and Attalea butyracea palm fruit deposited on the ground in two rainforest sites with contrasting levels of human disturbance (non–disturbed vs. disturbed) in the Lacandon rainforest in southern Mexico. We specifically addressed the following questions: (1) how does anthropogenic disturbance affect species richness and composition of the ensemble of mammalian frugivores that interact with A. butyracea fruit?; and (2) how does this disturbance modify the strength of the interaction that different mammalian species have with A. butyracea fruit? We expected to find a richer ensemble of mammals and more intense frugivory interactions (i.e., more frequent and involving more fruit) in the non–disturbed forest that is associated with the presence of larger–bodied mammals.

Palm species Attalea butyracea (henceforth Attalea) is a canopy palm found widely from southern Mexico to Bolivia (Govaerts and Dransfield, 2005). Its fruit consists of drupes grouped on an infructescence. With an average size of 5.5 cm long and 3 cm wide the fruit has a fleshy, sweet, and fatty mesocarp and a hard endocarp, containing from one to three seeds (Pennington and Sarukhán, 2005). In the Lacandon region, fruit fall from May to June, accumulating in a small area on the forest floor beneath the infructescences (C. Delgado–Martínez, pers. obs.). Attalea fruit is consumed by a wide variety of terrestrial mammals such as the agouti (Dasyprocta punctata) and the white–nosed coati (Nasua narica) (Jansen et al., 2014). Attalea is a good model to study frugivory interactions in the forest understory in view of its local abundance, widespread distribution across Mesoamerica, its high and regular fruit production favoring accumulations on the forest floor, and the attractiveness of its fruit for terrestrial mammals Study site Fieldwork was conducted in two areas, in the Montes Azules Biosphere Reserve (MABR) and in the Marques de Comillas region, both in the state of Chiapas in southern Mexico (fig. 1s in supplementary material). The MABR has a surface area of 3,312 km2 (16º 04' 55''–16º 57' 28'' N and 90º 45' 01''–91º 30' 24'' W). Its mean annual precipitation is 2,500 mm, with a dry season from December to April and a rainy season from May to November (Gómez–Pompa and Dirzo, 1995). The MABR is one of the protected areas with the highest mammalian diversity in the country (Medellín, 1994). Moreover, it hosts some of the largest populations of endangered mammals in the country, such as the tapir (Tapirus bairdii), and the white–lipped peccary (Tayassu pecari) (Naranjo et al., 2015). The Marques de Comillas region is located east of the MABR, along the Mexico–Guatemala border (fig. 1s in supplementary material). Nearly half of the land cover of this region has been lost due to deforestation caused by the impact of activities such as cattle ranching and the increasing establishment of oil palm plantations (Meli et al., 2015). Recent studies indicate that the few large forest fragments remaining in the area still support several species of mammals in the understory (Muench and Martínez–Ramos, 2016).

Monitoring of focal fruiting palms For seven days we traveled 15 km by boat along the Lacantun river and walked 10 km within the forest looking for fruiting Attalea palms in the southern portion of the MABR (henceforth the non–disturbed forest). Furthermore, over a period of 10 days, we walked 30 km in the Marques de Comillas region looking for Attalea palms in various forest fragments.

Delgado–Martínez and Mendoza

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We found Attalea fruiting palms only in one of these fragments (henceforth the disturbed forest), an area of 200 ha. This forest fragment was surrounded by cattle pastures to the north, by crops to the east, by a dirt road to the south, and by a plantation of African palm (Elaeis guineensis) in the west. We installed camera traps (Stealth Cam U838NXT) aimed at the natural fruit accumulations on the ground below five fruiting Attalea palms in the non–disturbed forest and five in the disturbed forest. These palms were selected based on two criteria: the abundance of fallen fruit accumulated at their base and a minimum distance of 100 m between them. The average (± SD) distance between monitored palms was 7,069.05 ± 4,361.87 m in the non–disturbed forest and 78.57 ± 52.24 m in the disturbed forest. Although a few of the monitored palms were located less than 100 m from the nearest tree they were not monitored simultaneously.

The cameras were set to take a 15–second video each time they were triggered and to have a 10–second delay before reactivation. Monitoring was conducted simultaneously in the two forests in May and June 2016. The cameras were checked weekly for 4.5 weeks (on average) to download videos and check battery levels. The time of camera deployment in the field depended on fruit availability; once fruit accumulations were no longer evident, the camera traps were retrieved.

Data analyses Based on the behavior depicted by the mammals in the videos, we classified them as those showing interaction with the fruit (i.e., mammals consuming or removing fruits) and those in which no interactions were evident. We recorded the number of fruits consumed or removed, the number of individuals observed during the interaction, and the length of the interaction in seconds. To avoid counting consecutive videos of the same species and camera trap as independent records, we grouped them using the protocol described in Camargo–Sanabria and Mendoza (2016). This protocol is based on grouping consecutive records using increasingly longer periods, until finding a species–specific minimum time of stabilization (i.e., when changes in the number of video groups with increasing time are minor, see table 1s in supplementary material). The resulting video groupings are referred to hereafter as events. We calculated a capture frequency (CF) for each mammalian frugivore that interacted with the fruit of each focal palm using the following equation: Number of events Sampling effort

x 100 camera trap days

To gain an understanding of how the spatial distribution of focal palms could have affected the recording of visiting mammal species and to know the extent to which focal palms, within each forest, could be treated as independent units, we tested for the existence of spatial autocorrelation among the mammalian ensembles recorded in Attalea palms.

To do this, we applied a Mantel test using the phy-

sical distance in meters between palm pairs and the compositional dissimilarity in the recorded mammalian faunas. The compositional dissimilarity was measured using the Canberra index (Lance and Williams, 1966) which calculates a sum of relative differences (based on CFs of each mammal species) between palm pairs (see supplementary material).

To compare the mammalian species richness of the ensembles interacting with fruit in each forest type, we generated sample–based species rarefaction curves.

Moreover, we calculated the first–order incidence– based estimator Jackknife (Jack1) to estimate the richness of mammal species associated with Attalea palms in each forest type. We conducted a non–metric multidimensional scaling (NMDS) using the Canberra index (based on CFs) to compare the composition of the ensembles of mammals that interacted with Attalea fruit. We obtained the stress value associated with the NMDS, which indicates the extent to which the two–dimensional ordination of focal palms accounted for the original distribution of palms in the multivariate space. The stress values range from 0 to 1, with values closer to 0 indicating a more effective representation (Borcard et al., 2018). We complemented the NMDS with an analysis of similarities (ANOSIM) to test for the existence of statistical differences between ensembles. These procedures were conducted using the vegan R package (Oksanen et al., 2019).

We calculated the interaction strength (IS) between each mammalian frugivore species and palms in both forest types using the following equation: IS =

PD * PI * LI * NI TF / MI / NE

where PD is the proportion of days each mammal species was recorded in focal palms, PI is the proportion of days in which interactions between mammals and focal palms occurred, LI is the mean duration of visits to focal palms, NI is the proportion of individuals of each mammal species that were interacting with fruit at each focal palm, TF is the total number of fruit consumed or removed by each mammal species at each focal palm, MI is the mean number of animals observed per interaction event, and NE is the total number of interaction events. This measurement is a modification of the approach applied by Camargo–Sanabria and Mendoza (2016). We standardized the IS values by dividing them by the overall maximum value, reaching values between zero and one without units.

Using these values, we calculated an average IS for each mammalian frugivore species in each forest type.

Results The total sampling effort was 180 camera trap days (126 in the non–disturbed forest and 54 in the disturbed forest), during which we recorded 107 events: 66 in the non–disturbed forest and 41 in the disturbed forest. Videos provided evidence of nine mammal species consuming or removing Attalea fruit in the non–disturbed forest, and four in the disturbed forest (table 1 and fig. 1s in supplementary

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Table 1. Mammalian species recorded interacting with Attalea fruit in two forests sites with contrasting levels of human disturbance: a Yes, the species interacted with fruit; No, the species did not interact with fruit; b Body mass taken from Aranda (2012); c EN, endangered; LC, least concern; VU, vulnerable; d P, endangered.

Tabla 1. Especies de mamíferos que se registraron interactuando con los frutos de Attalea en dos sitios con distinto grado de alteración antrópica: a Yes, la especie interactuó con el fruto; No, la especie no interactuó con el fruto; b Masa corporal obtenida de Aranda (2012); c EN, en peligro de extinción; LC, menor preocupación; VU, vulnerable; d P, en peligro de extinción.

Interactiona Non–disturbed

Disturbed

Body mass

IUCN

Species

forest

forest

(kg) b

status c

Mexican Norm

Cuniculus paca

Yes

Yes

5–13

LC

—

Dasyprocta punctata

Yes

—

2–5

LC

—

status

Dasypus novemcinctus

Yes

No

2.5–7

LC

—

Nasua narica

Yes

Yes

3–6

LC

—

Pecari tajacu

Yes

Yes

15–30

LC

—

Philander opossum

Yes

Yes

0.3–0.7

LC

—

Sciurus sp.

Yes

—

0.4–0.7

LC

—

Tapirus bairdii

Yes

—

150–300

EN

P

Tayassu pecari

Yes

—

25–42

VU

P

material). The spotted paca (Cuniculus paca) and the white–nosed coati were the most frequently recorded species, accounting for 27 % and 37 % of all events in the non–disturbed and disturbed forests, respectively (fig. 2s in supplementary material). The nine–banded armadillo (Dasypus novemcinctus) was the only mammal that did not interact with fruit in the disturbed forest. We did not detect evidence of spatial autocorrelation in the mammals interacting with Attalea fruit (non–disturbed forest: r = –0.1354, p = 0.625, n = 5; disturbed forest: r = –0.158, p = 0.5333, n = 5). This finding supports the use of palms as sampling units.

As expected, we found that human disturbance is reducing the species richness of mammals that interact with Attalea fruit in the disturbed forest (fig. 3s in supplementary material). The observed species richness in the non–disturbed forest was slightly lower than that estimated by Jack1 (9.99 ± 0.99 (SD)), whereas in the disturbed forest the observed species richness was the same as that estimated by Jack1.

The composition of the ensembles of frugivores that interacted with Attalea fruit contrasted between sites (ANOSIM R = 0.375, p = 0.028; fig. 1). This was particularly due to the absence of large–bodied mammals in the disturbed forest. The tapir had the strongest IS with Attalea fruit in the non–disturbed forest, whereas the squirrel (Sciurus sp.) had the weakest IS (fig. 2).

In contrast, in the disturbed forest, the white–nosed coati had the strongest interaction with Attalea fruit,

d

having an IS value of 1137.7 times higher than the corresponding value in the non–disturbed forest; the intensity of this interaction was reflected in coati bands, with up to 25 individuals, depleting fruit in three of the focal palms in only one week. Likewise, the IS of the collared peccary (Pecari tajacu) and the gray four–eyed opossum (Philander opossum) were 661 and 20 times greater in the disturbed forest than in the non–disturbed forest, respectively (fig. 2).

Discussion The documented differences in the frugivory interaction between mammals and Attalea palms in our study sites provide valuable insights regarding the potential impact of human disturbance on this biotic interaction. An important difference was that the two largest frugivore species in the region (i.e., the tapir and the white–lipped peccary) were not recorded interacting with Attalea fruit in the disturbed forest.

The absence of these species might be related to methodological issues, particularly an insufficiently long sampling effort in our focal forest fragment.

However, previous studies conducted in our study area, involving a larger sampling effort, have found that these species are rare in forest fragments (Garmendia et al., 2013; Muench and Martínez–Ramos, 2016; Porras et al., 2016). This evidence indicates that the absence of interactions between these lar-

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18

Dapu Cupa Peta

NMDS2

0

Nana

Phop

Dano Taba

–1

Sciu –2

Non–disturbed Disturbed

Tape –1

0 NMDS1

1

2

Fig. 1. Ordination of Attalea palms occurring in two forest sites with contrasting levels of human disturbance in the Lacandon rainforest in southern Mexico. The ordination is based on capture frequencies of mammalian frugivores. Stress = 0.0575. Species codes: Cupa, Cuniculus paca; Dapu, Dasyprocta punctata; Dano, Dasypus novemcinctus; Nana, Nasua narica; Peta, Pecari tajacu; Phop, Philander opossum; Sciu, Sciurus sp.; Taba, Tapirus bairdii; Tape, Tayassu pecari.

Fig. 1. Ordenación de las palmas de Attalea presentes en dos sitios con distinto grado de alteración antrópica de la selva Lacandona, en el sur de México. La ordenación está basada en la frecuencia de captura de los mamíferos frugívoros. Estrés = 0.0575. Código de especies: Cupa, Cuniculus paca; Dapu, Dasyprocta punctata; Dano, Dasypus novemcinctus; Nana, Nasua narica; Peta, Pecari tajacu; Phop, Philander opossum; Sciu, Sciurus sp.; Taba, Tapirus bairdii; Tape, Tayassu pecari.

ge–bodied mammals and Attalea fruit might not be an unusual situation in forest fragments in the Marques de Comillas region.

The absence of tapir interaction in our disturbed forest could reduce seed dispersal distances, which, in turn, would favor an increase in seedling and sapling aggregation near parent palms (Fragoso et al., 2003; Sica et al., 2014). On the other hand, the loss of the interaction between white–lipped peccary and Attalea fruit might produce a 'release' effect on seedling recruitment since this mammal's activity is an important cause of mortality among large–seeded palms (Beck, 2006). Therefore, the absence of white–lipped peccaries might exert an additional effect to increase spatial aggregation of Attalea seedlings around parent palms (Silman et al., 2003).

In contrast with what occurred with the tapir and the white–lipped peccary, the white–nosed coati

went from having the second–lowest IS value in the non–disturbed forest to being the species with the strongest interaction with Attalea fruit in the disturbed forest. It has been shown that a close relative of the white–nosed coati, the ring–tailed coati (Nasua nasua), patrols established circuits when looking for food (Hirsch et al., 2013). If white–nosed coatis have a similar foraging pattern, it would be possible for visitation rates to be higher in forest patches where resources are more limited. This behavior could explain the observed fruit depletion in three of the Attalea palms that were frequently visited by this mammal. It is not clear, however, how the increase in activity of white–nosed coatis could affect Attalea performance.

In some perturbated forests in Brazil, the ring–tailed coatis can disperse large seeds (Alves–Costa and Eterovick, 2007); however, more research is needed to know whether the activity of white–nosed coatis provides a dispersal service to Attalea seeds.

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Cupa (0.01 ± 0.02)

Dapu (0.001 ± 0.003) Sciu (0.00004 ± 0.0001)

Nana (0.0003 ± 0.0007)

Peta (0.0001 ± 0.0002)

Tape (0.003 ± 0.003)

Phop (0.0002 ± 0.0004)

Taba (0.029 ± 0.06)

0.341

IS values

Dano (0.06 ± 0.01)

0.029 < 0.001

Phop (0.004 ± 0.008)

Peta (0.066 ± 0.13)

Nana (0.341 ± 0.45)

Cupa (0.002 ± 0.004)

Fig. 2. Frugivore ensembles interacting with Attalea fruit in the non–disturbed forest (top) and the disturbed forest (bottom) in the Lacandon rainforest in southern Mexico. Arrow thickness indicates the interaction strength. (For the species code, see fig. 1).

Fig. 2. Ensamblajes de mamíferos frugívoros que interactuaron con los frutos de Attalea en el sitio sin alteración antrópica (arriba) y en el sitio con alteración antrópica (abajo) de la selva Lacandona, en el sur de México. El grosor de las fechas indica la intensidad de la interacción. (Para los códigos de especies, véase fig. 1).

We recorded the presence of the tapir and the agouti in the vicinity of our focal palms in the disturbed forest. However, as previously indicated, we did not find any evidence of these species interacting with the Attalea fruit. Previous studies have suggested that biotic interactions may be affected by anthropogenic impacts before the species involved in such interaction disappear (Valiente–Banuet et al., 2015).

In view of the difficulty in finding accessible plant species that synchronically produce abundant fruit and attract a variety of animal species our study had the following limitations: (1) limited replication, both in terms of number of sites and focal palms; (2) a more aggregated distribution of focal palms in the disturbed forest than in the non–disturbed

forest; and (3) differences in the sampling effort between sites. These limitations might have led to failure to detect a larger assemblage of mammals interacting with Attalea fruit in the disturbed forest.

However, we are confident that overall, our results provide a good approximation of the impact that human perturbation has on the characteristics of interaction between Attalea fruit and medium and large–bodied mammals inhabiting disturbed forests such as ours. Thus, our results highlight the need to design conservation strategies aimed not only at the maintenance of species in human–dominated landscapes but also at safeguarding the biotic interactions and ecosystem functions they promote (Soulé et al., 2003; Tylianakis et al., 2010).

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Acknowledgements We are very grateful to A. Camargo–Sanabria for her help during fieldwork and data analysis. We also thank I. López and R. Lombera for their help with the fieldwork. Thanks too to 'El Arca de Noé' and 'La Martucha' for providing housing facilities. Fieldwork was supported by funds awarded to EM through the Red PRODEP: 'Conservación de la biodiversidad en ambientes antropizados'.

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Prasad, S., Pittet, A., Sukumar, R., 2010. Who really ate the fruit? A novel approach to camera trapping for quantifying frugivory by ruminants. Ecological Research, 25(1): 225–231, Doi: 10.1007/s11284009-0650-1 Sica, Y. V., Bravo, S. P., Giombini, M. I., 2014. Spatial pattern of pindó palm (Syagrus romanzoffiana) recruitment in Argentinian atlantic forest: The importance of tapir and effects of defaunation.

Biotropica, 46(6): 696–703, Doi: https://doi. org/10.1111/btp.12152 Silman, M. R., Terborgh, J. W., Kiltie, R. A., 2003. Population Regulation of a Dominant Rain Forest Tree by a Major Seed Predator. Ecology, 84: 431–438, Doi: 10.1890/0012-9658(2003)084[0431:PROADR]2.0.CO;2 Soulé, M. E., Estes, J. A., Berger, J., Martinez Del Rio, C., 2003. Ecological Effectiveness: Conservation Goals for Interactive Species. Conservation Biology, 17(5): 1238–1250, https://www.jstor.org/ stable/3588949 Tylianakis, J. M., Laliberté, E., Nielsen, A., Bascompte, J., 2010. Conservation of species interaction networks. Biological Conservation, 143: 2270–2279, Doi: 10.1016/j.biocon.2009.12.004 Valiente–Banuet, A., Aizen, M. A., Alcántara, J. M., Arroyo, J., Cocucci, A., Galetti, M., García, M. B., García, D., Gómez, J. M., Jordano, P., Medel, R., Navarro, L., Obeso, J. R., Oviedo, R., Ramírez, N., Rey, P. J., Traveset, A., Verdú, M., Zamora, R., 2015.

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Nesting preferences of the green sea turtle (Chelonia mydas L.) and the hawksbill sea turtle (Eretmochelys imbricata L.) in the SW of Mahe Island in the Seychelles

F. Mata, P. Mata

Mata, F., Mata, P., 2022. Nesting preferences of the green sea turtle (Chelonia mydas L.) and the hawksbill sea turtle (Eretmochelys imbricata L.) in the SW of Mahe Island in the Seychelles. Animal Biodiversity and Conservation, 45.1: 23–31, Doi: https://doi.org/10.32800/abc.2022.45.0023 Abstract Nesting preferences of the green sea turtle (Chelonia mydas L.) and the hawksbill sea turtle (Eretmochelys imbricata L.) in the SW of Mahe Island in the Seychelles. Data concerning 212 turtles emerging on the southwest beaches of Mahe Island in the Seychelles were collected in 2017 and 2018. These data were used to model the probability of eggs being laid in relation to several variables. The probability of successful laying after emergence was highest on certain beaches and in areas of short vegetation, between open sand and trees. We found successful laying was related to the physical properties of the soil, indicating that survivability of embryos and hatchlings is higher in certain areas. The turtles appeared to choose zones where soil had low salinity, good drainage but ability to retain water, and absence of spring tides and extreme temperatures.

Key words: Nesting behaviour, Sea turtle, Seychelles, Soil properties Resumen Preferencias de la tortuga verde (Chelonia mydas L.) y la tortuga carey (Eretmochelys imbricata L.) con respecto al anidamiento en el suroeste de la isla Mahé en las Seychelles. Los datos relativos a las 212 tortugas que llegan a las playas del suroeste de la isla Mahé de las Seychelles se recopilaron en 2017 y 2018 y se utilizaron para determinar la probabilidad de que se pongan huevos en función de distintas variables. Se encontró que la probabilidad de que la puesta sea exitosa después de la llegada es mayor en ciertas playas y en la zona de vegetación baja entre la arena y los árboles. Se estableció una relación con las propiedades físicas del suelo y se dedujo que las zonas elegidas favorecen la supervivencia de los embriones y las crías. Aparentemente, las tortugas eligieron zonas donde el suelo tenía poca salinidad y buen drenaje, aunque mantenía cierta capacidad para retener la humedad, y que quedaban resguardadas de las mareas vivas y las temperaturas extremas.

Palabras clave: Comportamiento de anidación, Tortuga marina, Seychelles, Propiedades del suelo Received: 21 V 21; Conditional acceptance: 20 X 21; Final acceptance: 27 X 21 Fernando Mata, CISAS–Centre for Research and Development in Agri–food Systems and Sustainability, Instituto Politécnico de Viana do Castelo, Rua da Escola Industrial e Comercial Nun'Alvares 34, 4900–347 Viana do Castelo, Portugal.– Paula Mata, Ruskin Mill College, Millbottom, Stroud, GL6 0LA Gloucestershire, United Kingdom.

Corresponding author: Fernando Mata. E–mail: fernandomata@ipvc.pt ORCID ID: F. Mata: 0000-0002-5687-7114

ISSN: 1578–665 X eISSN: 2014–928 X

© [2022] Copyright belongs to the authors, who license the journal Animal Biodiversity and Conservation to publish the paper under a Creative Commons Attribution 4.0 License.

Mata and Mata

24

Introduction The Seychelles comprise around 150 granitic and coralline islands, believed to have been formed from the breakup of the supercontinent Gondwanaland in the Pre–cambric (Salm, 1978). Mahe is the main island, located in the centre of a granitic group and characterised by a very narrow coastal plateau of calcareous reef materials building up as sand dunes and pocket beaches known as 'anse'. The two types of prevailing soil are the ferrolithic of granitic weathering origin with clay rich A and B horizons and a fine texture, and the calcareous coarse sands areas close to the sea (Government of Seychelles, 2006). Mahe has many small, steep watercourses, most of which have ephemeral flow that depends on monsoons, and subterraneous water may contain traces of salt.

(FAO, 2005).

Overexploitation of turtle and chelonian products in the Seychelle Islands contributed to the decline of sea turtles. Since the early 1970s, positive measures of protection have led to the recovery of the main species of sea turtles in the archipelago: the green turtle (Chelonia mydas L.) and the hawksbill turtle (Eretmochelys imbricata L.) (Frazier, 1974). Numbers remained low in the 1980s and 90s (Mortimer et al., 1996) but appear to have increased in the recent decades (e.g. Mortimer et al., 2003; Allen et al., 2010).

Nevertheless, worldwide, green and hawksbill turtles are listed as endangered and critically endangered, respectively, by the International Union for Conservation of Nature Red List of Threatened Species since 1996 (IUCN, 2015).

Turtle reproduction has been affected by several environmental factors. Temperature, moisture and gaseous environment are of fundamental importance and are determinants of gender, embryo growth and hatchability (Júnior, 2009). The porosity of the turtle eggshell enables interaction between the embryo in the interior of the shell and the environment in the exterior.

This porosity allows exchanges of fundamental importance for successful incubation (Ackerman, 1997).

Ackerman (1997) reviewed the relationship between the embryonic development of the sea turtle and the nest environment in detail. He observed that possible survival of an egg clutch of sea turtles depended on the capacity of the female to select an optimal area and excavate a suitable nest chamber. The author considered climate, soil texture, drainage and salinity with water potential, gas exchange and temperature all played a role in hatchability and therefore in nest site selection. More recently, other authors, such as Miller et al. (2017), have revised the topic.

Together with a panel of worldwide experts on sea turtles, Hamann et al. (2010) established key priority topics of research for their management and conservation, the first (Q1) of which was 'factors that underpin nest site selection and behaviour of nesting turtles'.

In a review of the pertinent literature, these authors identified 'the factors (biotic and abiotic) driving where and when turtles lay their clutches', and 'management strategies that would help to protect or enhance the suitability of nesting habitat for sea turtles', as aspects

of particular interest to address this key priority area of research. In a study of nest site selection, Ditmer and Stapleton (2012) highlighted the need for future research to explore the roles of sand structure, nest moisture, and local weather conditions.

Rees et al. (2016) reviewed how these key priority areas were being addressed in relation to Q1. Several authors have published studies on this topic (e.g.

Neeman et al., 2015; Santos et al., 2016). However, they state that there remains a lack of understanding regarding how turtles select the nesting sites at intra and inter beach level.

In our study, we attempted to establish a relationship between nest site selection, the microenvironment of the site, and the edaphoclimatic variables of the southwest beaches of Mahe Island. The aim of was to contribute to the previously cited key priority area of research. Mahe is the largest of the Seychelle Islands, and it has the highest population density, especially around beaches, as tourism is an important local industry. Knowledge generated from this study could contribute to the establishment of priority areas of protection and conservation, stimulating the sustainability and coexistence of humans and turtles.

Material and methods Between 13 XI 17 and 11 I 18 we collected data regarding 212 turtles emerging on the southwest beaches of Mahe Island in order to evaluate the probability of eggs being laid.

The beaches considered were located (from north to south) in: Anse Intendance (INT) at 4º 47' 05'' S, 55º 30' 00'' E; Anse Cachee (CAC) at 4º 47' 44'' S, 55º 30' 20'' E; Anse Corail (COR) at 4º 47' 49'' S, 55º 30' 30'' E; Anse Bazarca (BZC) at 4º 48' 02'' S, 55º 30' 50'' E; Anse Petit Police (PPO) at 4º 48' 10'' S; 55º 31' 03'' E; Anse Grand Police (GPO) at 4º 48' 09'' E, 55º 31' 19'' E; and Anse Petit Boileau (APB) at 4º 48' 06'' S, 55º 31' 42'' E. Beaches were monitored every morning between 8 a.m. and 2 p.m. and the following variables were recorded for every turtle emerging: the each, progression inland (measured as distance to high tide line), eggs laid (yes, no), and zone of the beach at maximum progression and eventual lay (open sand, shrub, or trees). The exact location of the nest was also recorded by GPS, as were variables regarding the vegetation covering the soil: percentage of ground or canopy cover and dominant species.

A generalised linear model from the binomial family was used to model egg laying probability. The dependent variable considered was the success or failure of laying after each emergence. The independent variables investigated were the beach (INT, CAC, COR, BZC, PPO, GPO, and APB), the zone of the beach (open sand, bushes and trees), distance to high tide and the variables related with vegetation. Several link functions were tested (logit, log, log complement, negative log–log and complementary log–log) with the best fit after evaluation of Deviation and the Akaike’s Information Criterion being chosen as the model.

Animal Biodiversity and Conservation 45.1 (2022)

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Table 1. Parameters of model 1. Probability of a turtle laying eggs after an emergence depending on the beach and the distance to high tide. Deviance was 116 and AIC 132: SE, standard error; CI, confidence interval; OR, odds ratio; DHT, distance to high tide. Studied beaches: CAC: Anse Cachee; INT: Anse Intendance; GPO: Anse Grand Police; PPO: Anse Petite Police; APB: Anse Petite Boileau; BZC: Anse Bazarca; COR: Anse Corail.

Tabla 1. Parámetros del modelo 1. Probabilidad de que una tortuga ponga huevos después de llegar a una playa, en función de la playa y de la distancia a la marea alta. La desviación fue 116 y AIC 132.

(Para las abreviaturas, véase arriba).

Variables Beach

β

SE (β)

P–value

95% CI (β)

OR (eβ)

95% CI OR (eβ)

CAC

1.165

0.764

< 0.05

0.401; 1.929

3.206

1.493; 6.883

INT

0.849

0.765

0.084; 1.614

2.337

1.088; 5.023

GPO

0.816

0.579

0.237; 1.395

2.261

1.267; 4.035

PPO

DHT (m)

0

APB

0.064

0.790

–0.726; 0.854

1.066

0.484; 2.349

BZC

–0.211

0.547

–0.758; 0.336

0.810

0.469; 1.399

COR

–0.377

0.581

–0.958; 0.204

0.686

0.384; 1.226

0.738

0.041

< 0.05

0.697; 0.779

2.092

2.008; 2.179

Data were analysed using IBM® SPSS® Statistics for Windows, version 21.0. (IBM Corp., Armonk, NY, USA, 2012). Variables were selected for inclusion in the model through a forward stepwise procedure. The significance of the variables was evaluated using the Wald Chi–square test. The significance of the fitted model was evaluated by means of the likelihood ratio Chi–square test. All significance levels were set to P < 0.05.

Results A first model was successfully fit (P < 0.01) with a negative log–log link, with the variables 'beaches' (P < 0.05) and 'distance to high tide' (P < 0.05) being found significant. A higher distance to the high tide (higher progression into the land) determines a higher probability of laying success. A second model was also successfully fit (P < 0.01), again with a negative log–being significant. We observed that the more distant the zone of the beach from the water, the higher the probability of laying. Therefore, both models were similar correlating successful laying behaviour with higher progression inland.

Tables 1 and 2 present the details of the model parameter estimates. Figure 1 is the graphic representation of model 1 with equation: P (laying) = Exp (– Exp (– X1 + X2 x D))) where X1 is the parameter for 'beach', X2 the parameter for distance from the high tide and D the 'distance to high tide'. Figure 2 is the graphic representation of model 2 with equation:

P (laying) = Exp (– Exp (– X1 + X2))) where X1 is the parameter for 'beach' and X2 the parameter for 'zone of the beach'.

Discussion Incubation and hatchability of eggs is directly dependent on temperature, humidity, and the exchange of gases (Ackerman, 1997; Júnior, 2009). Refsnider and Janzen (2010) described oviposition site choice in several species as a major maternal effect factor affecting survival and phenotype of offspring. They considered that this maternal effect can be observed in fish, amphibians, reptiles and birds and described six hypotheses to explain the behaviour: maximization of embryo survival, maximization of maternal survival, modification of offspring phenotype, proximity to suitable habitat for the offspring, maintenance of natal philopatry, and factors related to mate choice. From these, only mate choice can be eliminated once turtles lack paternal care (Kamel and Mrosovsky, 2005).

In this discussion, we argue that the probability of successful laying after turtle emergence is positively correlated with the probability of the female turtle finding a place with a suitable microenvironment for laying, as hypothesised by Ackerman (1997) and suggested by Kamel and Mrosovsky (2005). In addition to the arguments made by these authors we relate site selection with the potential microenvironment of the clutch. The relation is established through the study of the physical properties of the soil with influence on the environmental variables, determining incubation and hatchability.

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Table 2. Parameters of model 2. Probability of a turtle laying eggs after an emergence in relation to the beach and zone of the beach. Deviance was 111 and AIC was 129. SE: standard error; CI: confidence interval; OR: odds ratio; HT: high tide. Zones A, B, C are respectively open sand, shrub and tree zones.

(For other abbrevitions, see table 1).

Tabla 2. Parámetros del modelo 2. Probabilidad de que una tortuga ponga huevos después de llegar a una playa, en función de la playa y de la zona de la playa. La desviación fue 111 y AIC 129. (Para las abreviaturas, véase arriba y la tabla 1).

Variables Beach

β

SE (β)

P–value

95% CI (β)

OR (eβ)

95% CI OR (eβ)

CAC

1.684

0.788

< 0.05

0.139; 3.229

5.387

3.188; 7.586

INT

1.176

0.793

–0.379; 2.731

3.241

1.031; 5.451

GPO

1.557

0.699

0.187; 2.928

4.745

2.733; 6.756

PPO

Zone

0

APB

0.199

0.785

–1.340; 1.737

1.220

–0.972; 3.413

BZC

0.075

0.581

–1.063; 1.213

1.078

–0.710; 2.866

COR

–0.142

0.670

–1.255; 1.171

0.868

–1.087; 2.822

A

–1.208

0.656

–2.493; 0.078

0.299

–1.628; 2.226

B

0.124

0.492

–0.841; 1.088

1.132

–0.504; 2.768

C

0

< 0.05

The relation between the variables in our models and clutch microenvironment Relatively to distance to the high tide, the first model relates the increase in laying probability with turtle progression inland. In the second model, however, with distance entered in the analysis as 'zones of the beach', this probability decreases slightly, but without a significant difference (P > 0.05) if the area with trees (longer distance) is reached, increasing from open sand (shorter distance) to the shrub zone (intermediary distance). While studying the nesting behaviour of the green turtle on a beach of the Guanahacabibes Peninsula, southwest of Cuba, Sánchez et al. (2007) found that the majority of successful laying after emergence also took place in the zone of vegetation after the open sand (75.3 % and 73 % respectively in 2002 and 2003). This was also the case in the study of Neeman et al. (2015), with 88.3 % in Tortuguero, Costa Rica. Neeman et al. (2015) also raised the hypothesis of choice due to optimal moisture and temperature.

Santos et al. (2016) studied similar models and found some variation, however, but they refer to consistency as their main findings. They studied the behaviour over eight seasons and found the same turtle tended to prefer the same beach ecosystem. While researching on the Sunshine Coast of Eastern Australia, Kelly et al. (2017) also concluded nesting preference close to vegetation, as did Hatase and Omuta (2018) in their studies on Yakushima Island, Japan.

The physical properties of the soil in these three different zones vary. In open sand, the soil has a coarse texture and may be within reach of the seawa-

ter, especially in spring tides. It is easy here for both turtle and hatchling to dig, respectively, in and out of the nest. The soil of the tree zone also has a sandy texture at the top, but is richer in organic matter and has some clay, as the typical ferrolitic soil, which is rich in clay, is closer to the surface. The tree zone has dense vegetation with strong rooting systems, making digging difficult. The shrub zone has intermediary characteristics, the transition between sand and sand richer in clay and organic matter, with clay in a deeper layer and sand with organic matter from the vegetation in the superficial layer. The different characteristics of the soils are responsible for different physical properties.

Moisture Sands have a coarse texture with macro porosity, allowing excellent drainage, but capacity for water retention is low. In consequence, nests may become exposed to long dry periods with the moisture levels dropping below optimum levels. On the other hand, if the nest is placed too near seawater, it may be inundated by waves, leaving the clutch exposed or washing it away (Diamond, 1976; Kraemer and Bell, 1980).

The fine texture of clay rich soils has a lower density and therefore greater porosity (micro pores are not normally perceived). This micro–porosity, where the capillarity force may overtake the gravity force, allows water to move vertically and horizontally, contributing to the maintenance of moisture levels. These soils have a higher capacity for water retention, but drainage tends to be relatively poor, and with intense

Probability of eggs being layed

Animal Biodiversity and Conservation 45.1 (2022)

27

1 0.9 0.8

CAC

0.7

INT

0.6

PPO

GPO

0.5

APB

0.4

COR

0.3

BZC

1 2 3 4 5 6 7 8 9 10 11 12 13 14 15 16 17 18 19 20 21 Distance to high tide line (m)

Fig. 1. Probability of a turtle laying eggs after an emergence depending on beach and distance to high tide (m). (For abbreviations of studied beaches see table 1).

Fig 1. Probabilidad de que una tortuga ponga huevos después de llegar a una playa, dependiendo de la playa y de la distancia a la marea alta (m). (Para las otras abreviaturas de las playas estudiadas, véase la tabla 1).

Probability of eggs being layed

rain they may easily be flooded and cause hatching failure due to embryo death (Wood, 1986). A nest in clay is also very difficult for a turtle to create and very difficult for a hatchling to dig out of.

The transition soils have a lower layer with clay and an upper level with sand and organic matter. A clutch located in this type of soil is covered by the sandy

horizon and is relatively close to a clay layer. Moisture can reach the upper horizons from below and levels above the wilting point and below saturation can be expected. The top soil horizon, rich in sand, allows water to flow, and the level that is relatively higher relatively to the beach allows drainage to occur, if the clay horizon reaches saturation.

0.9 0.8 0.7 CAC

0.6

INT

0.5

GPO

0.4

PPO

0.3

APB

0.2

BZC

0.1

COR

0

A

B Zone of the beach

C

Fig. 2. Probability of a turtle laying eggs after an emergence depending on beach and zone of the beach.

Zones A, B, C are respectively open sand, shrub and trees. (For abbreviations of studied beaches, see table 1).

Fig. 2. Probabilidad de que una tortuga ponga huevos después de llegar a la playa, en función de la playa y de la zona de la playa. Las zonas A, B y C son, respectivamente, de arena, de arbustos y de árboles. (Para las abreviaturas de las playas estudiadas, véase la tabla 1).

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Table 3. Typical values of thermal conductivity (TC) and thermal diffusivity (TD) (NERC, 2011).

Tabla 3. Valores típicos de conductividad térmica (TC) y difusividad térmica (TD) (NERC, 2011).

Soil

TC Wm–1K–1

TD m2day–1

Sand

0.77

0.04

Clay

1.11

0.05

Saturated sand

2.50

0.08

Saturated clay

1.67

0.06

Temperature Water has a higher specific heat than soil, and therefore a higher capacity to store heat; this property also prevents rapid changes in water temperatures (Reysa, 2012). The thermal conductivity (the rate at which heat is transferred to the surroundings) of soil is dependent on bulk porosity and the degree of soil saturation; the same occurs for thermal diffusivity (the rate at which heat is transmitted through the medium) (NERC, 2011). Table 3 shows the typical values for these thermal properties of soil in relation to texture and moisture.

As observed, without moisture, clay has a high capacity to gain and lose heat more quickly. However, with moisture (the required conditions for optimum incubation), this is not the case. In humid sandy soils, temperature fluctuates between high levels during the day and low levels at night. In clay rich soils with a higher moisture percentage, temperature fluctuation is minimized and a stable temperature is achieved.

Zones with vegetation can also protect the top soil with shade from extreme irradiation, therefore preventing scorching temperatures from reaching the clutch, as opposed to dry top layers of open sands. Incubation temperatures are profoundly influenced by the environment, such as air temperature, vegetation and sand albedo due to the colour (Esteban et al., 2016).

Clutch average depth was reported for the green turtle in the Lanyu Island of Taiwan, to be 66 cm (Cheng et al., 2009) or 85 cm in the bottom in Ascension Islands (Mortimer, 1990). For hawksbill turtles, Ditmer and Stapleton (2012) reported an average of 48 cm. From 50 cm depth, however, heat flow in the soil is minimal and soil temperature stabilizes around the daily average temperature at surface (Lavelle and Spain, 2003).

Exchange gases The soil atmosphere is determined by the space left by porosity and eventual presence of moisture occupying that space. Sandy soils provide good circulation of air and water, while clay rich soils have

a higher probability of having some of the porosity occupied by water. The presence of vegetation is responsible for the existence of macro–porosity in clay rich soils, caused by root penetration. Richer soils with clay and organic matter also have a higher degree of particle aggregation, responsible for a good soil structure that creates macro pores, promoting good atmospheric circulation. Sandy soils very close to seawater run the risk of flooding in spring tides, removing all the air from the soil porosity and creating conditions for egg drowning (Diamond, 1976; Kraemer and Bell 1980). The same can happen in soils with poor drainage in heavy rain conditions (Kraemer and Bell, 1980; Wood, 1986) Salinity Soil salinity is expected to decrease the greater the distance from high tide. Seawater loses influence and freshwater from rain and freshwater courses washes out the salt from the soil. Salinity in soils causes chemical deflocculation of clays, with destruction of soil aggregation of particles and overall structure (Goldberg and Glaubig, 1987). The clay is dispersed, the structure becomes cloddy, the pores are plugged and the soil becomes very hard. As a result, the circulation of air and water decreases and a sub–superficial impermeable layer of deflocculated clay may drown the clutch. On the other hand, if drainage is not a problem, salinity may also be responsible for an opposite effect; an increased osmotic potential surrounding the clutch may create difficulties for egg water absorption, and may promote the loss of water from the egg interior, which may result in desiccation (Ackerman, 1997).

The clutch microenvironment during incubation The nesting season of sea turtles in the Seychelles starts in August or September and continues until February or March (Diamond, 1976) or until April (Wood, 1986). During this time the climate in the Islands is temperate, between the cold season (mid–June to mid–May) and the warm season (mid–March to mid– May). Temperatures fluctuate between minimums of 25 to 26 ºC, and maximums of 29 to 30 ºC; precipitation has a high probability (88 %) of being moderate to light; relative humidity varies in the season below daily average minimums of 65 to 73 % and maximums of 90 to 95 % (WeatherSpark, 2014).

Beach temperatures vary depending on seasonal effects and the diurnal cycle. However, at 50 cm depth, nocturnal/diurnal effects are mitigated and the temperature tends to have lower variability and approach the average environmental temperature (Lavelle and Spain, 2003). This is especially true if moisture levels are maintained high (Ackerman,1997).

Incubation temperatures of sea turtles range between a minimum of 25–27 ºC and a maximum of 33–35 ºC (Ackerman, 1997); below 25 ºC and above 33 ºC embryos may die (Miller, 1982).

Like other reptiles, sea turtles are affected by environmental sex determination (Mrosovsky et

Animal Biodiversity and Conservation 45.1 (2022)

N

29

Riviére Intendance

INT

ére

Rivi

CAC COR BZC

nd Gra

GPO

e

c Poli

APB 4 km

Mahe Island

4º 45' S Seychelles 55º 27' E

20 km

INT: CAC: COR: BZC: PPO: GPO: APB:

Anse Intendance Anse Cachee Anse Corail Anse Bazarca Anse Petite Police Anse Grand Police Anse Petite Boileau Fresh watercourses

Fig. 3. Mahe Island, Seychelles. Expanded SW corner with beaches and local watercourses considered in the study.

Fig. 3. Isla Mahé, en las Seychelles. Esquina sudoccidental ampliada donde se pueden observar las playas consideradas en el estudio y los cursos de agua locales.

al., 1992). Pivotal temperatures (males above and females below) for hawksbill and green turtles were reported respectively as 29.32 ºC and 28.26 ºC (Ackerman, 1997).

Water exchange in turtle eggs is dependent on factors such as the structure of the eggshell, the water potential and temperature in the nest (Packard, 1999). An embryo with access to a good reserve grows to a larger size at hatching and has better survival conditions (Tracy et al., 1978; Packard, 1999).

Also, if the absorption of water is higher than water loss, shrinking of the original egg shape is avoided, thereby ensuring enough space for normal embryo development (Tracy et al., 1978). A higher level of nest humidity is therefore advantageous (Packard, 1999).

Sea turtle eggs depend on humidity uptake from the surrounding environment (Miller et al. 2017), and according to McGehee (1990), it is convenient for the turtle to lay at a minimum soil humidity of 25 %.

High levels of humidity may drown the embryo (Wood and Bjorndal, 2000) and low levels cause desiccation (McGehee, 1990). In the breeding season, it is therefore expected to find ideal conditions for incubation at nesting depth with a good moisture level and the temperature levels reported.

What is the turtle best place to lay?

It is important for the turtle to find a place that guarantees the right levels of humidity, temperature and gas exchange. According to the soil properties

discussed, a clutch laid close to the tree zone is ideal as the texture is thinner, sand is not so coarse, and organic matter and clay are present. This type of soil allows the nest to be dug deep enough for the walls to be firm and not collapse. This soil has also a fair water retention capacity without compromising drainage and allows air circulation. While studying nesting behaviour of green turtles at Ascension Island Mortimer (1990) found that a lower mean particle diameter of the soil with higher water potential was associated with higher laying and hatching success and lower mortalities.

Another important aspect is salinity. Water exchanges are especially sensitive to the substrate water potential (Tracy et al., 1978). Turtles would be looking for a place with low levels of salinity to facilitate a positive water exchange between eggs and soil.

This may explain the differences in the probabilities of successful laying after emergence in the different beaches. As can be observed in figure 3, Anse Itendance, Anse Grand Police and Anse Cachee are the only beaches that are fed by fresh watercourses, possibly accounting for the lower levels of salinity in the nesting areas of these beaches. While studying the green turtle, Johannes and Rimmer (1984) found that the mean salinity of the sand moisture in nesting beaches was half that of the non–nesting beaches in NW Cape Peninsula, Australia.

In favour of the theory that turtles actively evaluate a nesting place is the study by Hitchins et al. (2006) on hawksbill turtles in the Cousine Island of the

30

Seychelles. They found that when emerging, turtles not attempting to nest cover a shorter mean distance on land than those successfully or unsuccessfully attempting nesting. Also when crawling from the sea to the nest site, distances were longer and covered more slowly than those when crawling back to the sea. Studying the effect of experience on nest–site selection in loggerhead turtles in Queensland Australia, Pfaller et al. (2008) found that nesting–experienced females selected nest sites more successfully than those with little or no experience. The authors justify the behaviour with habituation to innocuous beach stimuli that encourage turtles to crawl farther from the sea in search of a nesting site. Sánchez et al. (2007) reported that the majority of the returns to the same beach happened with turtles laying in the vegetation area, which is the prime choice zone for laying in the beaches studied, and therefore turtles that previously identified a good laying place tended to return.

In conclusion, these finding to date allow an approach to the response to a key research question in sea turtles, as identified by Hamann et al. (2010) 'Why do turtles breed successfully at specific beaches and specific zones of the beach?'. They follow by saying this knowledge 'could lead to management strategies that maximise hatchling production in particular areas'.

The present study has the main limitation that environmental variables were not directly measured but deducted. Further research is needed to follow up on the results herein. Specifically, the environmental variables in nesting and non–nesting sites should be studied in greater depth. It is also evident that although the laying probability is higher at certain sites it is not null at others. Conservation efforts should centre attention on areas of higher laying probability but should not dismiss other areas, and eventual relocation of nests should be considered to improve hatchability.

Relocation to more convenient sites would also have the advantage of concentrating hatching, allowing the establishment of zones of protection against predation, and thereby improving survivability.

Acknowledgements The data used in this study were collected by the turtle team at the Marine Conservation Society Seychelles and used by kind permission of the Chairman Dr. Rowat.

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Small mammal sampling incidents related to wild boar (Sus scrofa) in natural peri–urban areas I. Torre, S. Cahill, J. Grajera, A. Raspall, M. Vilella

Torre, I., Cahill, S., Grajera, J., Raspall, A., Vilella, M., 2022. Small mammal sampling incidents related to wild boar (Sus scrofa) in natural peri–urban areas. Animal Biodiversity and Conservation, 45.1: 33–42, Doi: https:// doi.org/10.32800/abc.2022.45.0033 Abstract Small mammal sampling incidents related to wild boar (Sus scrofa) in natural peri–urban areas. The wild boar (Sus scrofa) has recently shown continuous population increases in many countries, leading to a rise in conflicts with human activities, including habituation to people and urban areas. Wild boar can disrupt the sampling of small mammals by reducing the number of potential captures. In this study we analysed whether sampling incidents recorded within a small mammal monitoring programme (SEMICE, www.semice.org) might be related to the density of wild boar in a network of protected parks. Our results suggested a peri–urban effect that was independent of wild boar densities in the protected parks; the number of damaged traps increased (rendering them inoperable for captures) and potentially resulted in underestimates of small mammals due to fewer functioning traps in the study area. We hypothesised that this high rate of damage to traps in a small and localised area in a peri–urban park could be related to wild boar associating human presence with greater opportunities to obtain food items of anthropogenic origin.

Key words: Live traps, Trap damage, Synurbization, Population, Underestimation, Collserola Park Resumen Incidentes en el muestreo de pequeños mamíferos relacionados con la presencia del jabalí (Sus scrofa) en zonas naturales perirubanas. Desde hace poco tiempo, las poblaciones de jabalí (Sus scrofa) de muchos países han venido aumentando de forma constante, lo que ha dado lugar a conflictos con las actividades humanas, como su habituación a las personas y a las zonas urbanas. Los jabalíes pueden interferir en el muestreo de pequeños mamíferos, ya que reducen las capturas potenciales. En este estudio analizamos si los incidentes registrados en el muestreo realizado en el marco de un programa de seguimiento de pequeños mamíferos (SEMICE, www.semice.org) podrían estar relacionados con la densidad de jabalíes en una red de parques protegidos. Nuestros resultados sugirieron que la ubicación periurbana ejercía un efecto independiente de la densidad de jabalíes en los parques, que hacía aumentar el número de trampas dañadas (no disponibles para la captura) y que podría producir la subestimación del tamaño de las poblaciones de micromamíferos debido a la menor disponibilidad de trampas. Presumimos que esta alta tasa de ataques a las trampas en una zona pequeña y localizada podría estar relacionada con la posibilidad de que, en el caso del parque periurbano, los jabalíes puedan asociar la presencia humana con más oportunidades de obtener alimentos de origen antrópico.

Palabras clave: Trampas de vivo, Daños a trampas, Sinurbización, Población, Subestimación, Parque de Collserola Rebut: 25 V 21; Conditional acceptance: 6 IX 21; Final acceptance: 8 XI 21 Ignasi Torre, Joan Grajera, Marc Vilella, BiBio Research Group, Natural Sciences Musem of Granollers, c/Francesc Macià 51, E–08402 Granollers, Spain.– Seán Cahill, Alfons Raspall, Consorci del Parc Natural de la Serra de Collserola, ctra. de l’Església 92, E–08017 Barcelona, Spain.

Corresponding author: I. Torre: itorre@mcng.cat ORCID ID: I. Torre: 0000-0002-4803-9524; M. Vilella: 0000-0002-0732-0985 ISSN: 1578–665 X eISSN: 2014–928 X

© [2022] Copyright belongs to the authors, who license the journal Animal Biodiversity and Conservation to publish the paper under a Creative Commons Attribution 4.0 License.

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Introduction Wild boar (Sus scrofa) populations are currently expanding both in distributional range and number in many countries (Keuling et al., 2017; Massei et al., 2015). In Europe, this increase is considered due to a variety of reasons, such as increasingly mild winters (Bieber and Ruf, 2005; Geisser and Reyer, 2005; Vetter et al., 2015), rural abandonment, reforestation of marginal land, and the expansion of crops, such as maize (Geisser and Reyer, 2005), that provide food and cover (Colino–Rabanal et al., 2012). Traditional recreational hunting is declining and is insufficient to restrain the population growth of wild boar (Massei et al., 2015). Furthermore, wild boar can display compensatory population responses when hunting pressure is high (Gamelon et al., 2011; Servanty et al., 2011).

Synurbization of wild boar is a recent phenomenon.

Their presence is increasing in ever–expanding [peri] urban environments where they can often thrive on a combination of both natural and anthropogenic food sources, and may become habituated to the presence of people (Cahill et al., 2012; Putman et al., 2014; Náhlik et al., 2017; González–Crespo, 2021).

While foraging, wild boars can have a negative impact on sampling protocols used to estimate small mammal abundance by interfering with traps (Focardi et al., 2000) and increasing the number of sprung traps (Torre et al., 2019). Although sprung traps reduce the sampling effort, this effect is usually ignored in calculations of population estimates for small mammals (Beauvais and Buskirk, 1999).

Consequently, the population size of small mammals could be underestimated if the number of effective traps available is reduced due to sampling interference by wild boar.

In this article, we analysed the influence of wild boar foraging activity on the sampling of small mammal populations through incidents related to trap disruption.

We speculated that sampling incidents recorded within a small mammal monitoring programme (SEMICE, www.semice.org) might be related to the abundance of wild boars, since foraging intensity (i.e., rooting) would presumably be higher with greater presence of wild boar (Mori et al., 2020). As a result, random trap encounters should increase. We also expected that higher sampling incidents would result in a lower number of captures of small mammals due to reduced trap availability. Furthermore, we analysed other sampling incidences, especially those related to the activity of the small mammals themselves, with regard to the number of captures recorded.

Material and methods We used the information generated by the SEMICE small mammals monitoring scheme (www.semice. org) at 21 sampling stations situated in seven protected parks located in the broader region surrounding the city of Barcelona. The SEMICE sampling programme has been operative since 2008 at more than 150 sampling stations situated in Spain and Andorra.

Torre et al.

The aim of the programme is to monitor common small mammal species with high detectability so as to compute reliable estimates of population change (Torre et al., 2018). As a citizen science monitoring programme (Torre et al., 2021), our collaborators are encouraged to select sampling areas close to their homes, located –whenever possible– in protected natural areas, and choosing habitats that are representative of those present in the surroundings.

The SEMICE live–trapping scheme consists of two annual trapping sessions, each one spanning three days. These sessions are based on grids of 36 traps (Longworth and Sherman) arranged in intercalated positions, separated by 15 m, and covering an approximate effective area of 1 ha. Longworth traps (Penlon Ltd., Oxford, UK) are aluminium box traps that consist of two separate sections that need to be joined when setting: a tunnel with a trigger, and a nest box with enough space for placing food and bedding material. Sherman traps are aluminium box traps (23 × 7.5 × 9 cm; H. B. Sherman Traps Inc., Tallahassee, FL, USA) that have a weight–sensitive trigger that closes the door when an animal steps on it. Both traps are widely used, and their performance has been tested (Sibbald et al., 2006; Torre et al., 2019). The traps were baited, provided with hydrophobic cotton for bedding, and concealed under vegetation to provide thermal insulation. We used the information available from 152 sampling sessions (trapping bouts of three days each) conducted between spring 2016 and spring 2020 at 21 different sampling stations (mean: 7.24 sessions/ sampling station). Each station was equipped with 36 trap positions.

We defined three kinds of trap incidents depending on the agent that caused them: those related to traps being moved or damaged by either (i) wild boar or by (ii) mesocarnivores, and (iii) those mostly related to small mammal activity. Incidents related to wild boar and mesocarnivores can be determined by the signs found at the trap location and the general condition of traps (fig. 1), and by using indirect clues of recent activity around the traps, such as rooting by wild boar and animal faeces. Incidents related to small mammal activity can be determined either from open traps with animal signs, i.e. traps that were not sprung but show clear evidence of use by small mammals (i.e. faeces, cotton removed, etc.), or from unmoved closed traps that were sprung but without any capture. The incidents were assigned to the most likely causative agent according to the observers’ criteria; almost all the information was recorded by professional biologists with at least 10 years of field experience. For the analyses, we considered five response variables: wild boar incidents, mesocarnivore incidents, and three incidents related to small mammals themselves; open traps with signs, closed traps without capture, and the sum of both open and closed traps. Calculations were based on each sampling session (three days) at each sampling station (36 traps), resulting in 108 trap–nights per sampling unit (n = 152). While open traps with signs could be unambiguously assigned to small mammals,

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A

A

A

B

B

Fig. 1. A, small mammal traps damaged by wild boar in Collserola Natural Park. Sherman traps were more prone to serious damage than Longworth traps. The marks of the two lower incisors can be seen in the dented Sherman traps. B, damage caused by foxes V. vulpes (and other mesocarnivores) can be differentiated by the marking of the canines and by the bending of the extremes of the traps to gain access to the interior (Garraf Park).

Fig. 1. A, trampas para pequeños mamíferos dañadas por jabalíes en el Parque Natural de Collserola. Las trampas Sherman eran más propensas a sufrir daños graves que las Longworth. Se pueden observar las marcas de los dos incisivos inferiores en las trampas Sherman deformadas. B, los daños provocados por zorros, V. vulpes (y otros mesocarnívoros) se pueden diferenciar por las marcas de los caninos y porque los extremos de las trampas están doblados para poder acceder al interior (Parque del Garraf).

closed traps were assigned to small mammals' activity when other evident signs of interference by large animals were absent, although their closure might possibly have been caused by other uncontrolled factors (i.e., wind, rain, etc.). For this reason, these two types of incidents were also analysed separately to account for this possible source of unknown variation. Each model had one fixed factor, namely the season when the sampling took place (autumn or spring), and a fixed covariate, the number of small mammal captures at each sampling unit. However, for incidents caused by wild boar, we added three

other explanatory variables: the protected park as a fixed factor (a nominal variable with seven levels), the campaign ordinal number (each station was sampled on a maximum of nine occasions during four consecutive years), and mean wild boar density in each park (a categorical variable with three levels) as fixed covariates. Data on the densities of wild boar for each protected park were obtained from the results of the seasonal hunting returns, reported by Rosell et al. (2019) for 2016–2019, from the official Catalan wild boar monitoring programme (Rosell et al., 2018). For our study, we defined

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Incidence rate (per 100 trap–nights) 0–0.1 0.1–0.2 0.2–0.3 2.8–2.9

MSY SLL

MGR LIT COL

MAR

Wild boar density High Medium Low

Spain

GAR 0

Barcelona

7.5 15 km

Fig. 2. Incidence rates of wild boar contact with small mammal traps according to average wild boar densities (individuals hunted/km2: low: < 5 indiv/km2; medium: 5–10 indiv/km2; high: > 10 indiv/km2) in seven protected parks in the region of Barcelona: COL, Serra de Collserola; LIT, Serralada litoral; MGR, Montnegre i el Corredor; SLL, Sant Llorenç del Munt i l'Obac; GAR, Garraf; MAR, Serralada de Marina; MSY, Montseny.

Fig. 2. Índices de incidencia del jabalí en las trampas para pequeños mamíferos en función de la densidad media de jabalí (individuos capturados/km2; baja: < 5 indv/km2; media: 5–10 indv/km2; alta: > 10 indv/km2) en siete parques protegidos en la región de Barcelona. (Para las abreviaturas de los parques, véase arriba).

three categories of density accordingly for each protected park (fig. 2) as either: 'low' (< 5 indiv/km2), 'medium' (5–10 indiv/km2) or 'high' (> 10 indiv/km2).

All five response variables were simple counts (i.e., number of traps affected by animal activity during each sampling session) and were considered to follow a Poisson error distribution (O'Hara and Kotze, 2010). However, a goodness–of–fit test performed with the goodfit function of the R package vcd (Meyer et al., 2021) showed a better fit to a negative binomial distribution. For each response variable, we built generalized linear mixed models (GLMM) by means of the glmer.nb function from the lme4 package (Bates et al., 2015), considering the 21 sampling stations as a random factor. We constructed models resulting from all combinations of explanatory variables using the dredge function of the MuMIn package for R (Bartoń, 2015), and we selected models with ∆AICc < 2 as being meaningful. These models were used to interpret which explanatory variables were most likely to influence small mammal trap incidents. Pseudo–R2 values (trigamma method) of the selected models were calculated with the same R package. Response variables were graphically represented as incidents per 100 trap–nights in order to be comparable with other methodologies.

Results During the study period, we recorded 872 incidents for 16,416 trap–nights (5.3 %). Most incidents (82.3 %) were related to the activity of small mammals themselves (traps not sprung or sprung without capture), while the remainder were attributed to larger animals (17.7 %) which had moved the traps. Of these, 63 incidents were assigned to mesocarnivores and 91 were assigned to wild boar. Of the latter incidents, 47 were major attacks (broken traps with irreparable damage, fig. 1A), and the remaining 44 were related to minor attacks (mostly reparable damage), showing knocks and dents to traps, perhaps merely caused unintentionally by wild boars while foraging. Most wild boar incidents (82.4 %) occurred in Collserola Natural Park, located on the outskirts of the city of Barcelona. The number of traps involved in wild boar incidents represented 2.88 ± 2.73 (SD, range 0–9.53) incidents per 100 trap–nights in Collserola, while such incidents were irrelevant in the other six protected parks (0.13 ± 0.43 traps per 100 trap–nights). The GLMM analysis confirmed that incidents caused by wild boar were mostly associated with the selected predictors (62 % of deviance), being higher in Collserola than in the other parks sampled in the province of Barcelona: the probability of incidence was twenty

Animal Biodiversity and Conservation 45.1 (2022)

37

10.0

Wild boar incidents

COL LIT 7.5

MGR SLL GAR

5.0

MAR MSY

2.5

Small mammal incidents

0.0 0

25

50 Small mammal captures

0

25

50 Small mammal captures

75

15

10

5

0 75

Fig. 3. A, sampling incidences related to wild boars in the seven protected parks; B, other sampling incidences not caused by large animals (traps sprung and not sprung with signs) were related to small mammal captures in all parks. (For abbreviations of the seven parks, see fig. 2) Fig. 3. A, incidencias en el muestreo relacionadas con jabalíes en los siete parques protegidos; B. en todos los parques se produjeron otros incidentes de muestreo que no fueron provocados por animales de gran tamaño (trampas cerradas y abiertas con signos) en relación con capturas de pequeños mamíferos. (Para las abreviaturas de los siete parques, véase fig. 2).

times higher, and out of 47 major attacks which resulted in irreparable damage to traps, 35 (74.4 %) occurred in Collserola (table 1, fig. 3A). The output of the analysis also showed that incidents caused by wild boar increased over time (from spring 2016 to spring 2020) in the study area, and that the number of small mammal captures was only negatively linked

to wild boar incidents in Collserola (table 1, fig. 3A).

However, the density of wild boar did not affect the frequency of trap incidents. In fact, parks with low densities of wild boar showed roughly similar incidence rates to those observed in parks with high densities, as well as to those in parks with intermediate densities, except for Collserola (fig. 2). In contrast, it was

38

not possible to interpret the effect of mesocarnivores on sampling incidents owing to the low explanatory power of the model (1 % deviance by fixed factors).

Sampling incidents –other than those caused by large animals– were positively associated with the number of captures of small mammals (table 1, fig. 3B). This positive association was observed in both types of incidents, especially among those reporting open traps with activity signs, but also when traps were sprung without capture, suggesting that most of the latter incidents were also related to the activity of the small mammals themselves..

Discussion A review by Barrios–Garcia and Ballari (2012) highlighted the fact that the foraging activity of wild boar (i.e., rooting) has an important impact on other animal communities. Amori et al. (2016) showed wild boar activity had strong negative effects on the abundance of small mammals, and especially so among ground–dwelling species (Mori et al., 2020). Foraging by wild boar could affect the population size of small mammals in two ways, either by direct effects, such as the destruction of their burrows and predation on individuals (Casula et al., 2017; Focardi et al., 2000), or by indirect effects, such as competition for food, alteration of rodent behaviour, and modification of vegetation that provides shelter for small mammals (Fagiani et al., 2014; Sunyer et al., 2016). In this study, we showed that the behaviour of wild boar, and specifically the sampling incidents they caused, could also affect our live–trapping estimates of the population size of small mammal communities by reducing the number of traps available to capture animals during field sampling. Our results indicate that the abundance of small mammals could potentially be underestimated if the incidence of trap disruption is high due to wild boar activity. Although trap disruption by wild boars has previously been mentioned by other authors (Focardi et al., 2000), no study has yet evaluated its possible impact on population estimates of small mammal communities.

Nevertheless, the indirect impact of large animals on these estimates was considered as negligible overall, owing to the fact that only a small number of traps were affected by wild boar (0.59 ± 1.57 incidents per 100 trap–nights), and even fewer by mesocarnivores (0.41 ± 1.08 incidents per 100 trap–nights). Indeed, sampling incidents related to the SEMICE monitoring programme in the same study area were deemed irrelevant (Torre et al., 2019).

When traps were found open but with evident signs of activity (cotton outside the trap, faeces inside, etc.), these incidents were generally attributed to the small mammals themselves, and our analysis confirms an association between small mammal abundance (capture rates) and rates of such incidents. However, incidents involving traps that are found sprung with no capture also seem to be positively linked to their abundance, as small mammals may also accidentally activate traps during external interactions with them

Torre et al.

(e.g. faeces are sometimes located on top of traps), or simply because the trap door has closed too soon after their entry into the trap.

Surprisingly, however, trap incidents caused by wild boar were unrelated to this species’ density, suggesting that in most protected parks wild boar either ignored or simply did not find traps while foraging. Similarly, some authors suggest that the impact of wild boar on natural habitats is unrelated to density (Adams et al., 2019). Moreover, in our study we show a 'Collserola effect', with a high incidence of damage caused by wild boar to small mammal traps in this park, independently of wild boar density, and contrasting with only occasional observations of such incidents in most other parks. Therefore, other reasons might explain why attacks by wild boar were significantly higher in Collserola Natural Park, an area where only moderate densities of wild boar were recorded (Rosell et al., 2019) during this study period. Collserola is located in the immediate vicinity of Barcelona, and wild boar in this park are generally accustomed to high human presence (Cahill et al., 2012). Used traps are clearly impregnated with the odour of the small mammals and of the baits used, possibly making them novel targets for wild boar while foraging. Thus, we hypothesised that this high rate of attacks on traps in a small and localised area could be related to the possibility that in Collserola, synurbic wild boars might already associate human presence with increased opportunities to obtain food items of anthropogenic origin. In effect, wild boar are known to frequently exploit such resources in (peri)urban areas of this park and are often fed, either directly or indirectly, by humans (Cahill et al., 2012; Castillo Contreras, 2019; González–Crespo, 2021). Also, recent studies on visitor use in Collserola Natural Park estimated a total of five million visits per year in 2019 (Farías–Torbidoni and Morera, 2020). Wild boar are thus accustomed to high human presence in this park, and animals that already associate humans with food might more readily examine objects displaying human odour. Currently, they readily explore and obtain food items from within closed structures such as rubbish bins (pers. obs.). On the other hand, in protected parks with a lower urban influence and less frequented by people, wild boar are, as yet, likely to be more wary of human odours. Also of note in Collserola Natural Park was the observation that disruption of small mammal traps by wild boar was negatively linked to capture success, thus potentially causing slight underestimates of population size for small mammals if high trap failure rates are not properly considered during calculations. Indeed, in one sampling session in Collserola, the number of disrupted traps accounted for almost 10 % loss of total traps available for capturing small mammals. Although competition among small mammals for available traps may be trivial during periods of low population abundance, trap failure might result in significant population underestimation during periods of high abundance due to lower trap availability (Beauvais and Buskirk, 1999; Torre et al., 2019). Therefore, in areas with high trap disruption caused by wild boar (or other animals), statistical

Animal Biodiversity and Conservation 45.1 (2022)

39

Table 1. Generalized linear mixed models (GLMM) with negative binomial error distribution selected according to ∆AICc < 2 for each response variable: wild boar, mesocarnivore and small mammal incidents (TO, trap open; TC, trap closed; TO + TC, trap open + closed) (n = 152 in all cases). Small mammal captures and season (spring as reference level) are two explanatory variables used in all models. For wild boar incidents, two other explanatory variables were added: the campaign ordinal number and a factor indicating the park to which the station belonged (Collserola as reference level).

Model β coefficients and their standard error in brackets, AICc, ∆AICc and pseudo-R2 (marginal and conditional) values are presented (*** p < 0.001, ** p < 0.01, * p < 0.05).

Tabla 1. Modelos mixtos lineales generalizados con distribución de los errores binomial negativa seleccionada de acuerdo con ∆AICc < 2 para cada variable de respuesta: jabalí, mesocarnívoros e incidentes relacionados con los pequeños mamíferos (TO, trampa abierta; TC, trampa cerrada; TO + TC, trampa abierta + cerrada) (n = 152 en todos los casos). Las capturas de pequeños mamíferos y la estación (la primavera se utiliza como nivel de referencia) son dos variables explicativas empleadas en todos los modelos. En el caso de los incidentes provocados por jabalíes, se añadieron otras dos variables explicativas: el número ordinal de la campaña y un factor que indicaba el parque al que pertenecía la estación (se tomó Collserola como referencia). Los valores de los coeficientes β del modelo y su error estándar se presentan entre corchetes, AICc, ∆AICc y pseudo-R2 (marginal y condicional) (*** p < 0,001, ** p < 0,01, * p < 0,05).

Type of incidents (Intercept) Captures Season

Wild boar

TO

TC

TO + TC

0.19

0.54

-0.89*

-1.18**

-1.50***

0.87***

0.80***

0.89*** 1.00***

(0.43)

(0.37)

(0.36)

(0.45)

(0.28)

(0.15)

(0.17)

(0.15) (0.13)

0.04*

-0.05

-

0.01

0.04

0.01

0.01

***

0.02*** 0.02***

(0.02)

(0.02)

(0.00) (0.00)

**

***

***

(0.01)

(0.00)

(0.00)

(0.00)

-1.01*

-0.90*

0.49*

-

0.12

(0.33) (0.41)

(0.42)

(0.21) (0.12)

0.56

Campaign

Mesocarnivores

Small mammals

0.16

*

(0.06)

-

0.18

-

(0.11)

0.15

*

(0.06)

Park Garraf Litoral

-1.90***

-1.75**

(0.54)

(0.52)

-33.11

-36.79

(34.99E5) (19.06E6) Marina Montnegre Montseny

-2.19***

-2.01***

(0.61)

(0.60)

-3.27

-3.23***

***

(0.63)

(0.63)

-38.37

-32.55

(13.70E6) (18.93E5) Sant Llorenç

-2.69***

-2.59***

(0.65)

(0.64)

216.57

217.10

247.90

249.06

391.70

693.46

694.72

0

0.54

0

1.15

0

0

1.26

0

0.57

Pseudo-R marg.

0.62

0.68

0.01

0.01

0.25

0.09

0.10

0.25

0.23

Pseudo-R2 cond.

0.62

0.68

0.04

0.04

0.36

0.27

0.28

0.44

0.42

AICc ∆AICc 2

744.78 745.35

Torre et al.

40

Fig. 4. Prototype protective guards against wild boar attacks made from galvanized steel. Their size allows insertion of small to medium–sized commercial traps (Longworth / Heslinga and medium–sized Sherman traps, the last shown in the photos).

Fig. 4. Prototipo de cubierta de protección frente a los ataques de jabalí hecha de acero galvanizado.

Su tamaño permite albergar trampas comerciales de tamaño pequeño y mediano (trampas Longworth o Heslinga y trampas Sherman medianas como las que muestran las imágenes).

techniques need to be applied to control for potential biases in the estimates of small mammal population sizes (Beauvais and Buskirk, 1999).

In such areas, traps should either be placed inside fenced areas (Focardi et al., 2000; Sunyer et al., 2016), or each trap should be individually protected by some form of cover, such as the prototypes now currently used in our study area. These comprise galvanized steel covers staked to the ground (fig. 4).

In the main, the damage that wild boar produced to traps was irreparable, making them inoperative after

major attacks. Damage caused by wild boar to small mammal traps in the Collserola Park represented a cost of € 1,650 over the four–year the period (roughly € 400/year), and an even higher total accumulated amount if we include damages incurred before this study and the added administrative time loss and costs due to procurement of new replacement traps.

Thus, if successful, the current use of fixed protective guards against attacks by wild boars (fig. 4) might be a significant financial saving if the small mammal monitoring programme continues in this region, con-

Animal Biodiversity and Conservation 45.1 (2022)

sidering that each trap protection unit currently costs € 23; an entire sampling grid of 36 traps could be protected through the investment of € 828. Nonetheless, the existing prototype design of these new protective guards will likely require further improvement and monitoring before their effectiveness at mitigating trap damage can be sufficiently guaranteed. In addition, given that incidents were only frequently observed in one protected park (Collserola), future research would ideally seek to confirm our hypothesis regarding the synurbization of wild boar as a driving factor behind increased damages to small mammal traps in highly frequented peri–urban natural areas.

Acknowledgements We are indebted to the local (Consorci del Parc Natural de la Serra de Collserola), provincial (Diputació de Barcelona) and Catalan Government (Departament de Territori i Sostenibilitat de la Generalitat de Catalunya) administrations of Catalonia for financial support to the Common Small Mammal Monitoring Programme (SEMICE) in the protected areas network in Catalonia since monitoring began in 2008. Also, we are grateful to Francesc Llimona and Daniel Díaz–Diethelm for assistance with logistic support to the SEMICE stations in Collserola Natural Park. We especially thank all the volunteers who collaborated in keeping some of the sampling stations active around Barcelona over the years: Dolors Escruela, James Manresa, Oriol Palau, Tomàs Pulido, Joan Manuel Riera, Cristina Terraza, Mar Unzeta. This study is a contribution to the BiBio Research Group (Bioindicators and Biodiversity) of the Natural Sciences Museum of Granollers. Professor Mario Díaz made valuable comments on a previous version of the article.

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New records of leeches of the genus Limnatis (Hirudinea, Praobdellidae) from the South Caucasus and Central Asia: phylogenetic relationships of Eurasian and African populations S. Utevsky, Y. Mabrouki, A. F. Taybi, M. Huseynov, A. Manafov, H. Morhun, O. Shahina, G. Utevsky, A. Khomenko, A. Utevsky Utevsky, S., Mabrouki, Y., Taybi, A. F., Huseynov, M., Manafov, A., Morhun, H., Shahina, O., Utevsky, G., Khomenko, A., Utevsky, A., 2022. New records of leeches of the genus Limnatis (Hirudinea, Praobdellidae) from the South Caucasus and Central Asia: phylogenetic relationships of Eurasian and African populations. Animal Biodiversity and Conservation, 45.1: 43–52, Doi: https://doi.org/10.32800/abc.2022.45.0043 Abstract New records of leeches of the genus Limnatis (Hirudinea, Praobdellidae) from the South Caucasus and Central Asia: phylogenetic relationships of Eurasian and African populations. Leeches of the genus Limnatis Moquin– Tandon, 1827 infest mucous membranes of various mammals, including humans and domestic ungulates. The type species of the genus L. nilotica (Savigny, 1822) was initially thought to occur throughout the Western Palaearctic, from North Africa to the Middle East and Central Asia. It was later found that L. paluda (Tennent, 1859) is a widespread Western Asian species. However, the South Caucasus and vast areas of Central Asia have not been explored sufficiently in terms of leeches of the genus Limnatis. We recorded L. paluda from Azerbaijan and Uzbekistan for the first time. We also carried out the first molecular characterisation of L. nilotica herein. We found a deep genetic differentiation (8 %) between the Western Asian L. paluda and North African (Moroccan) L. nilotica based on their COI sequences. This finding corroborates a previous morphology–based hypothesis on their separate species assignments. The low genetic diversity of L. paluda is explained by the recent colonisation of arid landscapes of Western Asia.

Key words: Annelida, Limnatis paluda, Limnatis nilotica, COI, Genetic diversity Resumen Nuevos registros de sanguijuelas del género Limnatis (Hirudinea, Praobdellidae) en el Cáucaso meridional y Asia central: relaciones filogenéticas de las poblaciones eurasiáticas y africanas. Las sanguijuelas del género Limnatis Moquin–Tandon, 1827 infestan las mucosas de varios mamíferos, incluidos los seres humanos y los ungulados domésticos. Se creía que la especie tipo del género, L. nilotica (Savigny, 1822), estaba presente en todo el paleártico occidental, desde África del norte hasta Oriente Medio y Asia central. Posteriormente, se observó que L. paluda (Tennent, 1859) es un especie ampliamente distribuida en Asia occidental. Sin embargo, no se han hecho estudios suficientes sobre el género Limnatis en la zona del Cáucaso meridional ni en buena parte de Asia central. Registramos L. paluda por primera vez en Azerbaiyán y Uzbekistán. La primera caracterización molecular de L. nilotica tuvo lugar allí. Las secuencias del gen citocromo oxidasa (COI) permitieron constatar que existe una profunda diferenciación genética (del 8 %) entre L. paluda, de Asia occidental, y L. nilotica, de África del norte (Marruecos). Ello corrobora la hipótesis basada en la morfología que se asignaba especies diferentes. La escasa diversidad genética de L. paluda se explica por la reciente colonización de territorios áridos de Asia occidental.

Palabras clave: Anélidos, Limnatis paluda, Limnatis nilotica, COI, Diversidad genética

ISSN: 1578–665 X eISSN: 2014–928 X

© [2022] Copyright belongs to the authors, who license the journal Animal Biodiversity and Conservation to publish the paper under a Creative Commons Attribution 4.0 License.

44

Utevsky et al.

Received: 02 VII 21; Conditional acceptance: 15 X 21; Final acceptance: 25 XI 21 Serge Utevsky, Halyna Morhun, Oleksandra Shahina, George Utevsky, Andrii Khomenko, Andriy Utevsky, Department of Zoology and Animal Ecology, Biological Faculty, V. N. Karazin Kharkiv National University, Kharkiv, Ukraine.– Youness Mabrouki, Biotechnology,Conservation and Valorisation of Natural Resources Laboratory, Faculty of Sciences Dhar El Mehraz, Sidi Mohamed Ben Abdellah University, Fez, Morocco.– Abdelkhaleq Fouzi Taybi, Équipe de Recherche en Biologie et Biotechnologie Appliquées, Faculté Pluridisciplinaire de Nador, Université Mohammed Premier, Nador, Morocco.– Mair Huseynov, Asif Manafov, Institute of Zoology, National Academy of Azerbaijan, Baku, Azerbaijan.– Halyna Morhun, Institute of Marine Biology, National Academy of Sciences of Ukraine, Pushkinska st. 37, 65148 Odesa, Ukraine.

Corresponding author: S. Utevsky. E–mail: serge.utevsky@karazin.ua

Animal Biodiversity and Conservation 45.1 (2022)

Introduction The genus Limnatis Moquin–Tandon, 1827 comprises bloodsucking leeches that occur in the south–western Palaearctic. They infest mucous membranes of various organs such as the pharynx, nasopharynx, oesophagus, larynx, trachea, bronchial tubes and female genital organs in humans and domestic mammals, including horses, cattle, camels, deer, and dogs (Kaburaki, 1921; Moore, 1927; Almallah, 1968; Arenas et al., 1993; Boye and Joshi, 1994; Al–Ani and Al–Shareefi, 1995; Ağin et al., 2008; Bahmani et al., 2012, 2014; Negm–Eldin et al., 2013; Rajaei et al., 2014; Raele et al., 2015). Moreover, leeches of the genus Limnatis can also be parasitic on amphibians (Lukin, 1976). The members of this genus are therefore important in terms of medicine, veterinary science, and parasitology. Taking global climate change into account, it is clear that they can pose a potential invasive threat by shifting their ranges, as has already happened with many species (Parmesan and Yohe, 2003).

According to the current view on the classification, the genus contains three species, Limnatis bacescui Manoleli, 1972, Limnatis nilotica (Savigny, 1822) and Limnatis paluda (Tennent, 1859) (Nakano et al., 2015). The type species of the genus, L. nilotica, was first described by the French zoologist Jules Cesar Savigny from Egypt under the name Bdella nilotica (Savigny, 1822). Later, in 1827, Moquin–Tandon renamed Savigny's Bdella nilotica to Limnatis nilotica in his seminal monograph (Moquin–Tandon, 1827). The second species is Limnatis bacescui Manoleli, 1972 from Romania (South–Eastern Europe), where it was described and is currently known only from its type locality (Manoleli, 1972). The third member of the genus, L. paluda, was described by Tennent (1859) as Haemopis paludum from Sri Lanka (= Ceylon).

Its taxonomic status was revised by Moore (1927) so that the species was transferred to the genus Limnatis and its name was changed to L. paluda (Tennent, 1859).

Traditionally, praobdellid leeches, found both in North Africa and Western Asia, were identified as L. nilotica. Representatives of this species had been recorded for Kazakhstan and Central Asia (Lukin, 1976), and southern Iran (Grosser and Pešić, 2006).

Subsequently, Phillips and Siddall (2009) and Nakano et al. (2015) found that the leeches of the genus Limnatis of Israel, Afghanistan and Kazakhstan should be assigned to L. paluda.

Despite the long history of previous studies, vast areas of Central Asia and the Caucasus have not been explored sufficiently in terms of their leeches of the genus Limnatis. There are no records based on molecular data concerning the species identity of those leeches in the South Caucasus and Central Asian countries except in Kazakhstan (Nakano et al., 2015), long known as regions of the Limnatis range (Lukin, 1976). Furthermore, North African leeches of the genus Limnatis have never been characterised based on their DNA sequences. The differentiation between the Western Asian L. paluda and the North

45

African L. nilotica has relied on morphological and geographical considerations (Moore, 1938). For this reason we aimed to identify leeches of the genus Limnatis collected in Uzbekistan (Central Asia) and Azerbaijan (the South Caucasus) and to clarify taxonomic and phylogenetic relationships between North African and Western Asian leech populations of the genus Limnatis using both morphological and molecular characters.

Material and methods Sample collection Leeches were collected during field trips in Uzbekistan, Azerbaijan, and Morocco (table 1). These samples were anesthetized in 10 % ethanol, fixed, and preserved in 96 % ethanol for further examination using both morphological and molecular methods.

The specimens are stored in the collection of invertebrate animals at the Department of Zoology and Animal Ecology, V. N. Karazin Kharkiv National University.

Morphological examination Identification was carried out using a stereomicroscope Konus Crystal–45. Photo documentation was done using a USB HDCE–50B camera. We relied on Moore (1938) to find morphological features distinguishing L. nilotica and L. paluda.

DNA extraction, amplification and sequencing Using molecular methods we analysed four specimens of L. paluda collected in Uzbekistan, one specimen from Azerbaijan and one from Morocco assigned to L. nilotica. A small piece of tissue from the posterior part of the body was taken for DNA extraction. Genomic DNA was isolated using a GeneElute Mammalian Genomic DNA Minprep Kits.

The mitochondrial cytochrome c oxidase subunit I (CO1) fragment was chosen as a standard animal DNA barcode gene region (Hebert et al., 2003) and amplified using following primers (Folmer et al., 1994): LCO1490, 5′–GGTCAACAAATCATAAAGATATTGG–3′ (forward) and HCO2198, 5′–TAAACTTCAGGGTGACCAAAAAATCA–3′ (reverse) by applying 5 cycles of 30 s at 94 ºC, 1 min 30 s at 45 ºC and 1 min at 72 ºC, 35 cycles of 30 s at 94 ºC, 45 s at 51 ºC and 1 min at 72 ºC, and 1 cycle of 5 min at 72 ºC after an initial 3 min denaturation step at 94 ºC. Alternatively, another PCR protocol (Utevsky et al., 2021) was implemented.

PCR products were cleaned using two enzymes, Exonuclease I and Shrimp alkaline phosphatase (SAP) (Fermentas, Thermo Fisher Scientific, USA).

Exonuclease I (0.2 μl) and SAP (1 μl) were added to 10 μl of the PCR product. The mixture was then incubated for 45 min at 37 ºC followed by 15 min incubation at 80 ºC. The cleaned PCR product was then sequenced in both directions by Macrogen

Utevsky et al.

46

Table 1. Information on specimens examined: COI, COI sequence code; N, number of specimens.

Tabla 1. Información sobre los ejemplares examinados: COI, código secuencias COI; N, número de especímenes.

Species L. paluda

Date

COI

N

Locality

Collector

15 VI 2005

LN

19

A mountain stream, Urgut District,

A. Abdullaev

L. paluda

7 VI 2007

m20,

4

m21

L. paluda

VIII 2017

T9

4

L. paluda

9 VIII 2016

T10

1

Samarqand Region, Uzbekistan Biological Station (University of Samarqand), Samarqand Region, Uzbekistan Boysun District, Surxondaryo 41 º 01' 19.48ʺ N, 48º 41' 45.50ʺ E, a warm spring, Hashi,

Quba District, Azerbaijan

22 IX 2019

U59

4

M. Asrorov

Region, Uzbekistan

L. nilotica

S. Utevsky

33 º 58' 01.4" N, 3º 02' 19.0" W,

A. Manafov

Y. Mabrouki

Ain Tafrent spring, Debdou, Morocco

Inc. (the Netherlands) using the same primers as at the amplification stage. The chromatograms of sequences were processed in ChromasPro 1.32 (Technelysium Pty., Queensland, Australia). The length of the newly generated COI sequences was 650–661 bp.

Phylogenetic analysis To reveal phylogenetic relationships of the Caucasian, Central Asian and North African leeches of the genus Limnatis, all available COI sequences of leeches assigned to that genus, sequences of Nearctic and Neotropic praobdellid leeches plus members of other families of Hirudiniformes and the erpobdellid leech Trocheta danastrica Stschegolew, 1938 were chosen for analysis and downloaded from GenBank (table 2).

The COI sequences were unambiguously aligned using MUSCLE algorithm in MEGA X. The final dataset contained a total of 1,302 positions. The alignment was checked for stop codons by translating it to amino acids using MEGA X (Kumar et al., 2018). Best–fit models of molecular evolution were determined for each codon position under the Bayesian information criterion using KAKUSAN4 (Tanabe, 2011): HKY85 with gamma distribution (+G) for the first codon position and GTR+G for the second and third positions.

Phylogenetic relationships were assessed by Bayesian inference using MrBayes v3.2.7a (Ronquist and Huelsenbeck, 2003) as implemented in the CIPRES Science Gateway (Miller et al., 2010, accessible at www.phylo.org). Searches were performed in two

parallel runs with eight chains each for ten million generations, sampled every 100th generation. After the first 25 % of the sampled trees were discarded, the final topologies were consented following the 50 % majority rule.

In addition, using MEGA X, we calculated the number of uncorrected base differences per site (based on p–distances) between species–level clades of the genus Limnatis.

Results Morphology The specimens collected in Azerbaijan and Uzbekistan were identified as L. paluda based on their morphological characters. The leeches range from 12.6 to 95.4 mm in length. The specimen from Azerbaijan is 48.9 mm in length. The body is flattened, indistinctly separated into the trachelosome and the urosome.

Some of the specimens have a well–defined clitellum.

The oral sucker is confluent with the trachelosome.

The caudal sucker is wide and constitutes 0.86 of the maximum width of the urosome. The oral sucker has a sulcus on its inner surface. The anus is inconspicuous. Mid–body segments are five–annulated. The gonopores are separated by five annuli. The leeches have five pairs of eyes that line up, creating a parabolic arc pattern. Papillae are weakly developed. Live leeches are greenish dorsally with no dark pattern.

Lateral margins of the body are orange. The venter is bluish black (fig. 1A, 1B).

Animal Biodiversity and Conservation 45.1 (2022)

47

Table 2. Collection sites and sequence accession data for COI sequences analysed: GenBank, GenBank accession number and specimen code. (Species names are presented according to this study).

Tabla 2. Lugares de recolección y datos sobre las accesiones de las secuencias COI analizadas: GenBank, número de acceso y código de la muestra. (Los nombres de las especies se indican de acuerdo con este estudio).

Taxon

GenBank

Country

Reference

MZ318072, U59

Morocco

This study

Limnatis sp.

GQ368754

Namibia

Phillips and Siddall (2009)

Limnatis sp.

AY763152

Croatia

Trontelj and Utevsky (2005)

Limnatis paluda (Tennent, 1859)

KY989474

Iran

Darabi–Darestani et al. (2021)

L. paluda

KY989473

Iran

Darabi–Darestani et al. (2021)

L. paluda

KY989472

Iran

Darabi–Darestani et al. (2021)

L. paluda

KY989471

Iran

Darabi–Darestani et al. (2021)

L. paluda

GQ368755

Afghanistan

Phillips and Siddall (2009)

L. paluda

AB981656

Kazakhstan

Nakano et al. (2015)

L. paluda

AB981654

Kazakhstan

Nakano et al. (2015)

L. paluda

AY425452

Israel

L. paluda

MZ318071, T9

Uzbekistan

This study

L. paluda

MZ318070, T10

Azerbaijan

This study

L. paluda

MZ318067, LN

Uzbekistan

This study

L. paluda

MZ318068, m20

Uzbekistan

This study

L. paluda

Limnatis nilotica (Savigny, 1822)

Borda and Siddal (2004)

MZ318069, m21

Uzbekistan

Limnobdella mexicana Blanchard, 1893

GQ368758

Mexico

Phillips and Siddall (2009)

This study

L. mexicana

GQ368756

Mexico

Phillips and Siddall (2009)

L. mexicana

GQ368757

Mexico

Phillips and Siddall (2009)

L. mexicana

GQ368759

Mexico

Phillips and Siddall (2009)

Myxobdella sinanensis Oka, 1925

LC192132

Japan

Nakano et al. (2017)

M. annandalei Oka, 1917

GU394014

India

Phillips et al. (2010)

Pintobdella chiapasensis (Caballero, 1957)

GU394015

Mexico

Phillips et al. (2010)

Tyrannobdella rex Phillips et al., 2010

GU394016

Peru

Phillips et al. (2010)

Hirudo orientalis Utevsky and Trontelj, 2005

EF405599

Uzbekistan

Utevsky et al. (2007)

Hirudo nipponia Whitman, 1886

AY763153

Korea

Trontelj and Utevsky (2005)

Whitmania laevis (Baird, 1869)

KT693113

India

Chatterjee et al. (2017)

Aliolimnatis michaelseni (Augener, 1936)

GQ368738

Guinea–Bissau

Phillips and Siddall (2009)

Hirudinaria manillensis (Lesson, 1842)

AY425449

Puerto Rico

Borda and Siddall (2004)

Haemopis sanguisuga (Linnaeus, 1758)

AF462021

Sweden

Semiscolex similis Oceguera–Figueroa, 2005

AY425457

Bolivia

Borda and Siddall (2004)

Patagoniobdella fraternal Ringuelet, 1976

AY425459

Chile

Borda and Siddall (2004)

Oxyptychus braziliensis (Pinto, 1920)

AY425455

Brazil

Borda and Siddall (2004)

Macrobdella decora (Say, 1824)

MH672573

North America

Müller et al (2019)

M. decora

EU100095

USA

Borda et al. (2008)

Philobdella gracilis Moore, 1901

DQ097218

USA

Phillips and Siddall (2005)

Philobdella floridana (Verrill, 1874)

DQ097219

USA

Phillips and Siddall (2005)

Haemadipsa sylvestris Blanchard, 1894

AF003266

Vietnam

Siddall and Burreson (1998)

Trocheta danastrica Stschegolew, 1938

MT013043

Ukraine

Khomenko et al. (2020)

Siddall (2002)

Utevsky et al.

48

A B S S C MG MG FG

FG MG

1 mm

10 mm

10 mm

FG

Fig. 1. External characters of leeches of the genus Limnatis: A, Limnatis paluda from Uzbekistan (Samarqand Region); B, Limnatis paluda from Azerbaijan; C, Limnatis nilotica from Morocco: FG, female gonopore; MG, male gonopore; S, sulcus. (Dorsal sides are depicted on the left and ventral sides, on the right for each specimen).

Fig. 1. Caracteres externos de las sanguijuelas del género Limnatis: A, Limnatis paluda procedente de Uzbekistán (región de Samarcanda), barra de escala de 10 mm; B, Limnatis paluda procedente de Azerbayán; C, Limnatis nilotica procedente de Marruecos: FG, gonoporo femenino; MG, gonoporo masculino; S, surco. (La imagen izquierda de cada ejemplar muestra la vista dorsal y la imagen derecha, la vista ventral).

The Moroccan leeches, which were assigned to L. nilotica, are 7.7–13.0 mm in length. The posterior sucker constitutes 0.76 of the maximum body width. The dorsal coloration pattern includes six longitudinal rows of black dots and short lines.

Pigmentation was largely bleached due to preservation in ethanol. The specimens do not have a well discernible sulcus. The gonopores are separated by five annuli (fig. 1C).

Phylogeny and genetic differentiation Phylogenetic analyses show the family Praobdellidae is a well–supported monophyletic group with a posterior probability of 1.00. The genus Limnatis is also a well– supported clade. All Middle Eastern, Caucasian and Central Asian samples joined a clade with a posterior probability of 0.78. This monophyletic group matches L. paluda. The phylogenetic structure of the clade is simple and shallow. The Namibian Limnatis sp. is sister to the clade of L. paluda with a posterior probability of 0.98. This group of Namibian and Western Asian leeches is sister to the clade consisting of the Moroccan L. nilotica and Croatian Limnatis sp. The latter clade is supported by a posterior probability of 0.62 (fig. 2).

The number of base differences per site from averaging sequence pairs within the clade of L. paluda is as low as 0.0030 ± 0.0012. Uncorrected distances between the Western Asian L. paluda, Balkan Limnatis sp., North

African Limnatis sp. and South African Limnatis sp. exceed 0.06 (table 3), suggesting species–level differences between those populations.

Discussion Morphological examination of the leeches of the genus Limnatis suggests the specimens collected in Azerbaijan and Uzbekistan should be assigned to the Middle Eastern and Central Asian L. paluda. This identification is based on the morphological features as follows: in contrast to the North African L. nilotica, Central Asian and South Caucasian leeches of the genus Limnatis are characterised by the monotonous green coloration with no dark dots or lines on the dorsum. The dorsal coloration pattern consisting of black stripes, lines and dots is characteristic of the North African L. nilotica (Moore, 1938). The Moroccan specimens have the typical coloration of L. nilotica.

The phylogenetic analysis corroborated the morphological identification and revealed a deep differentiation between North African and Western Asian leeches of the genus Limnatis. Molecular characterisation of the North African Limnatis, which is currently assigned to L. nilotica in the strict sense, was performed herein for the first time. All Middle Eastern, Caucasian, Central Asian and Afghan leeches joined the clade of L. paluda (fig. 2). This

Animal Biodiversity and Conservation 45.1 (2022)

0.59

49

Tyrannobdella rex GU394016 Pintobdella chiapasensis GU394015 Myxobdella sinanensis LC192132 Myxobdella annandale GU394474

0.96

0.89

Limnatis paluda KY989474 Iran Limnatis paluda AY425452 Israel Limnatis paluda KY989473 Iran Limnatis paluda KY989472 Iran 0.58 Limnatis paluda KY989471 Iran Limnatis paluda GQ368755 Afghanistan Limnatis paluda AB981656 Kazakhstan 0.78 Limnatis paluda AB981654 Kazakhstan Limnatis paluda LN MZ318067 Uzbekistan Limnatis paluda m20 MZ318068 Uzbekistan 0.98 Limnatis paluda m21 MZ318069 Uzbekistan Limnatis paluda T10 MZ318070 Azerbaijan 1.00 Limnatis paluda T9 MZ318071 Uzbekistan 0.60

Praobdellidae

1.00

Limnatis

1.00

Limnatis sp. GQ368754 Namibia Limnatis sp. AY763152 Croatia 0.62 Limnatis nilotica U59 MZ318072 Morocco Limnobdella mexicana GQ368758 1.00 Limnobdella mexicana GQ368756

0.76

0.58

1.00

Limnobdella mexicana GQ368757 Limnobdella mexicana GQ368759 Hirudo orientalis EF405599

1.00 1.00

0.63

Hirudo nipponia AY763153 Whitmania laevis KT693113

Aliolimnatis michaelseni GQ368738 Hirudinaria manillensis AY425449 Haemopis sanguisuga AF462021 Semiscolex similis AY425457 Patagoniobdella fraterna AY425459 Oxyptychus braziliensis AY425455 0.96

0.99 1.00 0.76 1.00 1.00 1.00 1.00

Macrobdella decora MH672573 Macrobdella decora EU100095 Philobdella gracilis DQ097218 Philobdella floridana DQ097219 Haemadipsa sylvestris AF003266

Trocheta danastrica MT013043 0.2

Fig. 2. Bayesian phylogenetic tree of arhynchobdellid taxa, including the genus Limnatis and other members of the family Praobdellidae, based on COI sequences. Posterior probabilities are shown for clades. The tree is rooted at Trocheta danastrica.

Fig. 2. Árbol filogenético bayesiano de los taxones de arrincobdélidos, incluidos el género Limnatis y otros miembros de la familia de los praobdélidos, basado en las secuencias de COI. Se muestra la probabilidad posterior de los clados. El árbol está enraizado en Trocheta danastrica.

corroborates independent species statuses of the North African and Western Asian populations, both of which had been assigned by some classical authors (Lukin, 1976) to L. nilotica. The Namibian Limnatis is sister to the L. paluda clade and belongs to an unidentified species, which may represent the little– known taxonomical diversity of the genus Limnatis in Africa (Moore, 1938). The Moroccan and Croatian leeches are in sister relationships. While the Moroccan specimen should be assigned to L. nilotica, the

Croatian leech is another unidentified species. The Moroccan–Croatian clade is sister to the clade of L. paluda plus the Namibian leech. The phylogenetic relationships imply that the taxonomy of the genus Limnatis is far from complete. Further studies of Southern European and South African populations of the genus Limnatis (fig. 3) are needed for proper identification. Moreover, the conspecific relationships of Sri Lankan and Western Asian leeches, which are currently assigned to L. paluda (Moore, 1927;

Utevsky et al.

50

20º 0' 0'' E

60º 0' 0'' E

S

10º 0' 0'' S

30º 0' 0'' S

N

Limnatis Limnatis Limnatis Limnatis

paluda sp. AY763152 nilotica sp. GQ368754

0

1,500 km

Fig. 3. Geographical distribution of leeches of the genus Limnatis: red, L. nilotica; green, L. paluda; grey, South European Limnatis sp.

Fig. 3. Distribución geográfica de las sanguijuelas del género Limnatis: en rojo, L. nilotica; en verde, L. paluda; en gris, Limnatis sp. del sur de Europa.

Phillips and Siddall, 2009; Nakano et al., 2015), should be substantiated using fresh specimens and molecular data. Biogeographically, the occurrence of the same leech species in Sri Lanka and in Western Asia appears questionable.

Despite the vast range of L. paluda, the species was found to have low genetic diversity and a shallow phylogenetic structure. This may be explained by the recent colonisation of arid landscapes in Western Asia. The range expansion could be attributed to the

Table 3. Estimates of evolutionary divergence based on p–distances between COI sequences of leeches of the genus Limnatis and their standard errors.

Tabla 3. Estimación de la divergencia evolutiva a partir de la distancia entre las secuencias (p–distance) de COI de las sanguijuelas del género Limnatis y su error estándar.

Limnatis sp. Namibia Limnatis sp. Croatia Limnatis nilotica Morocco

Limnatis paluda Limnatis sp. Namibia 0.0723 ± 0.0107 0.0947 ± 0.0115 0.1192 ± 0.0137 0.0795 ± 0.0103 0.1028 ± 0.0126

Limnatis sp. Croatia

0.0628 ± 0.0093

Animal Biodiversity and Conservation 45.1 (2022)

parasitism of these leeches on their ungulate hosts that appear to be able to transmit their parasites over long distances (Nakano et al., 2015). The eastern medicinal leech Hirudo orientalis Utevsky and Trontelj, 2005 is another instance of the rapid colonization of that area, which caused comparable genetic consequences (Trontelj and Utevsky, 2012). Migrations of nomads and their livestock throughout vast territories of the Middle East, Caucasus and Central Asia or other human activities in that area (as was discussed in Nakano et al., 2015) could contribute to shaping the genetic structure of L. paluda. Obviously, more studies are needed to clarify the evolutionary history and to elaborate a robust classification of the leeches of the genus Limnatis.

Acknowledgements Our sincere thanks to all the people who helped us in the field and collected leech specimens. Abdumalik Abdullaev kindly shared his experience in leech habitats and localities in Uzbekistan. Madamin Asrorov sampled other leeches from Uzbekistan.

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Native seed dispersal by rodents is negatively influenced by an invasive shrub A. F. Malo, A. Taylor, M. Díaz

Malo, A. F., Taylor, A., Díaz, M., 2022. Native seed dispersal by rodents is negatively influenced by an invasive shrub. Animal Biodiversity and Conservation, 45.1: 53–67, Doi: https://doi.org/10.32800/abc.2022.45.0053 Abstract Native seed dispersal by rodents is negatively influenced by an invasive shrub. Refuge–mediated apparent competition is the mechanism by which invasive plants increase pressure on native plants by providing refuge for generalist consumers. In the UK, the invasive Rhododendron ponticum does not provide food for generalist seed consumers like rodents, but evergreen canopy provides refuge from rodent predators, and predation and pilferage risk are key factors affecting rodent foraging and caching behaviour. Here we used a seed removal/ seed fate experiment to understand how invasion by an evergreen shrub can alter seed dispersal, seed fate and early recruitment of native trees. We used seeds of four species, small and wind–dispersed (sycamore maple Acer pseudoplatanus and European ash Fraxinus excelsior) and large and animal–dispersed (pedunculate oak Quercus robur and common hazel Corylus avellana), and monitored seed predation and caching in open woodland, edge habitats, and under Rhododendron. In the open woodland, wind–dispersed seeds had a higher probability of being eaten in situ than cached seeds, while the opposite occurred with animal–dispersed seeds.

The latter were removed from the open woodland and edge habitats and cached under Rhododendron. This pattern was expected if predation risk was the main factor influencing the decision to eat or to cach a seed.

Enhanced dispersal towards Rhododendron cover did not increase the prospects for seed survival, as density of hazel and oak saplings under its cover was close to zero as compared to open woodland, possibly due to increased cache pilferage or low seedling survival under dense shade, or both. Enhanced seed predation of ash and sycamore seeds close to Rhododendron cover also decreased recruitment of these trees. Rhododendron patches biased rodent foraging behaviour towards the negative (net predation) side of the conditional rodent / tree interaction. This effect will potentially impact native woodland regeneration and further facilitate Rhododendron spread due to refuge–mediated apparent competition.

Key words: Seed dispersal, Seed caching, Apodemus sylvaticus, Rhododendron ponticum, Refuge–mediated apparent competition, Plant–animal conditional mutualism Resumen Un arbusto invasor influye negativamente en la dispersión de semillas autóctonas por roedores. La competencia aparente mediada por refugio es el mecanismo por el que las plantas invasivas aumentan la presión sobre las autóctonas proporcionando un refugio para los consumidores generalistas. En el Reino Unido, la especie invasora Rhododendron ponticum no proporciona alimento a los consumidores generalistas de semillas, pero el dosel perenne ofrece refugio frente a roedores depredadores, y la depredación y el robo son los principales riesgos que afectan al comportamiento de alimentación y almacenamiento de los roedores. En el presente estudio, llevamos a cabo un experimento sobre la retirada y el destino de las semillas con objeto de entender cómo puede afectar la invasión de un arbusto perenne a la dispersión y el destino de las semillas y al reclutamiento temprano de árboles autóctonos. Utilizamos semillas de cuatro especies: semillas pequeñas y anemócoras (arce blanco Acer pseudoplatanus y fresno común Fraxinus excelsior) y semillas grandes y zoócoras (roble común Fraxinus excelsior y avellano común Corylus avellana) e hicimos el seguimiento de la predación y el almacenamiento de semillas en bosques abiertos, en hábitats de transición y debajo de Rhododendron. En los bosques abiertos, las semillas anemócoras tuvieron una mayor probabilidad de ser consumidas in situ que almacenadas, mientras que en las semillas zoócoras ocurrió lo contrario. Estas últimas ISSN: 1578–665 X eISSN: 2014–928 X

© [2022] Copyright belongs to the authors, who license the journal Animal Biodiversity and Conservation to publish the paper under a Creative Commons Attribution 4.0 License.

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se retiraron del bosque abierto y los hábitats de transición y se almacenaron debajo de Rhododendron. Esta pauta era la única esperable si el riesgo de predación era el factor más influyente en la decisión de consumir o almacenar las semillas. El aumento de la dispersión de las semillas hacia la cubierta de Rhododendron no hizo aumentar las perspectivas de supervivencia de las mismas, puesto que la densidad de las plántulas de avellano y roble debajo de la cubierta era prácticamente nula, a diferencia de lo que ocurre en el bosque abierto; un hecho se puede deberse al aumento del robo de las semillas almacenadas, a la baja supervivencia de las plántulas en condiciones de sombra densa o a ambas cosas. El aumento de la predación de semillas de fresno y arce cerca de la cubierta de Rhododendron también hizo disminuir el reclutamiento de estos árboles. Los parches de Rhododendron sesgaron el comportamiento de alimentación de los roedores hacia el lado negativo (predación neta) de la interacción condicional entre roedores y árboles. Es posible que este efecto influya en la regeneración de bosques autóctonos y facilite aún más la propagación de Rhododendron debido a la competencia aparente mediada por refugio.

Palabras clave: Dispersión de semillas, Almacenamiento de semillas, Apodemus sylvaticus, Rhododendron ponticum, Competición aparente mediada por refugio, Mutualismo condicional planta–animal Received: 11 X 21; Conditional acceptance: 3 XI 21; Final acceptance: 30 XI 21 Aurelio F. Malo, Global Change Ecology and Evolution Research Group, Departamento de Ciencias de la Vida, Universidad de Alcalá, GloCEE, 28805 Alcalà de Henares, Spain.– Aurelio F. Malo, Andrew Taylor, Department of Life Sciences, Imperial College London, Silwood Park, Ascot SL5 7PY, Berkshire, United Kingdom.– Mario Díaz, Department of Biogeography and Global Change, Museo Nacional de Ciencias Naturales (BGC–MNCN– CSIC), c/Serrano 115 bis, 28006 Madrid, Spain.

Corresponding author: Aurelio F. Malo: aurelio.malo@uah.es ORCID ID: A. F. Malo: 0000-0002-0846-2096; M. Díaz: 0000-0002-6384-6674

Animal Biodiversity and Conservation 45.1 (2022)

Introduction Invasive plants can cause dramatic and long–lasting changes in communities and ecosystems (Ehrenfeld and Scott, 2001; Callaway and Maron, 2006; Strayer et al., 2006). Understanding how invasive species interact with natives is crucial to detect and mitigate detrimental consequences to the native community.

Invasive species and natives can interact directly or indirectly. Indirect competition can occur because invasive plants can affect animal behaviour (Mattos and Orrock, 2010) and population abundance (Pearson, 2009). Apparent competition is a general type of indirect interaction between victims that operates through changes in density or foraging preferences of shared consumers (Holt and Bonsall, 2017; Noonburg and Byers, 2005; Orrock et al., 2008). In the context of interactions between plants and consumers, refuge–mediated apparent competition occurs when a plant provides refuge for a shared consumer and increases its local abundance or foraging efficiency (Berryman and Hawkins, 2006; Orrock et al., 2010), which impacts another non–refuge–providing plant.

This process has been documented in terrestrial (Caccia et al., 2006; Orrock et al., 2008; Orrock and Witter, 2010) and aquatic systems (Menge, 1995), and has been argued to play a role in determining the structure and invasibility of plant communities (for a review see Orrock et al., 2010). However, precise mechanisms of refuge–mediated apparent competition have not yet been thoroughly explored.

Here we test whether refuge–mediated dynamics involving rodents could be altering seed predation and dispersal patterns of native trees in Rhododendron ponticum–invaded woodland, influencing native tree spatial distribution and altering competition outcome between native tree species and the invasive plant.

Rodents have a significant role in seed predation and dispersal in woodlands (Jensen and Nielsen, 1986; Crawley, 1992; Hulme and Borelli, 1999). The rodent– tree interactions is a classical plant–animal conditional mutualism as the outcome of the interaction may be either mutualistic (dispersal) or antagonistic (predation). The balance between mutualism and antagonism is contingent on both intrinsic properties partners (seed size and nutritive value, rodent abundance) and on the ecological setting in which the interaction occurs (competition and predation risk; see Gómez et al., 2019; Morán–López et al., 2021 for recent reviews).

Three main hypotheses have been posed to explain what seeds to cache. First, the seed size/handling time hypothesis (Jacobs, 1992) poses that animals should decide to cache when food takes less time to cache than to eat. Second, the high–tannin hypothesis (Smallwood and Peters, 1986; Xiao et al., 2009) poses that animals should cache high–tannin seeds for later consumption, as seeds with higher tannins are less palatable and germinate later –less–perishable food and hence suitable to storage– than those with lower tannin levels (Smallwood and Peters, 1986; Smallwood et at., 2001). Third, the germination schedule hypothesis (Hadj–Chikh et al., 1996; Marti and Armario, 1998) poses that seed germination and perishability

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influence the decision of whether to eat or cache a seed. However, in the context of foraging decisions, the relative importance of these three hypotheses remains debated, as competition, both intra– and interspecific, and predation risk strongly modifies rodent foraging decisions (Morán–López et al., 2021). Mice select areas of high shrub/dense cover to reduce predation risk (Southern and Lowe, 1982; Brown, 1988; Kotler et al., 1991), which leads to higher foraging and seed dispersal rates under shrubs (Herrera, 1984; Hulme, 1996; Perea et al., 2011). In some situations, such as in high–insolation Mediterranean scrublands, this can facilitate seedling survival through the provision of beneficial shade and the reduction of herbivory (Hulme, 1997). In contrast, in other environments the presence of a dense cover–providing plant can reduce germination and recruitment (Wada, 1993), potentially leading to a change in the spatial distribution of native trees. Furthermore, concentration of rodents under safer shrub cover can lead to increased cache pilferage, thus reducing cache survival and recruitment under shrubs that may become ecological traps (Jordano and Herrera, 1995; Smit et al., 2009).

Rhododendron ponticum (hereafter Rhododendron) is probably the introduced plant with the greatest negative impact in the British Isles (Williamson, 2002).

During the eighteenth century it was introduced from Spain into Victorian country estates for its ornamental value. Its subsequent spread (Cross, 1975; Milne and Abbott, 2000) became a threat to native plants and biodiversity in the British isles (Dehnen–Schmutz et al., 2004; Tyler et al., 2006). The most serious effect on woodlands results from the lateral vegetative spread (Mejias et al., 2002) of large plants (> 12 years old) and the dense shade they cast, which excludes ground vegetation and prevents germination of native tree seedlings (Cross, 1981; Rotherham and Read, 1988).

Its competitive success also comes from its ability to escape natural predators, i.e. the seeds and vegetation are not a significant food source for herbivores in their invaded range (Keane and Crawley, 2002). Its invasion results in the disruption of woodland regeneration (Thomson et al., 1993) and in successional changes (Mitchell et al., 1997) that lead to declines in plant community diversity (Becker, 1988).

Rhododendron provides no food for generalist consumers such as rodents as neither its tiny seeds (average weight 0.063 g; Cross, 1975) nor its leaves are edible (Cross, 1981). However, its intertwined branches and dense evergreen cover provide refuge for them. Its structure is different from any native plant and it offers rodents protection from their main aerial predators. In fact, in our study site, mouse densities are two to five times higher under Rhododendron than on open woodland (Malo et al., 2013), leading to smaller home ranges under Rhododendron (Godsall et al., 2014). Such a response is not unusual; refuge–seeking behaviour of mammalian seed consumers is common (Clarke, 1983; Wolton, 1983; Spencer et al., 2005), including near exotic plants (Orrock et al., 2008). Thus, in Rhododendron–invaded woodlands, refuge–mediated apparent competition could lead to an increased negative impact of seed consumers

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on nearby native plants through seed caching and predation, indirectly benefiting the refuge–providing invasive plant, and potentially altering the outcome of direct competition. Understanding whether rodents can impact Rhododendron spread through refuge– mediated apparent competition, has implications for the conservation and regeneration of UK native woodlands.

Here we report on a spatially–explicit seed–fate experiment conducted in a well–studied rodent population in England to gain insight into the effects of an invasive shrub on native rodent–seed interactions and explore the behavioural mechanisms whereby granivore–mediated apparent competition may influence regeneration patterns in a multi–species context. We monitored two foraging decisions of rodents: seed selection and fate. First, rodents could choose between small, wind–dispersed seeds, and large and nutritive animal–dispersed seeds. Second, we recorded two foraging behaviours (eating and caching). We thus tested for variation in rodent foraging behaviour when encountering seeds of a different size that require different handling times and offer different energy rewards, and when encountering them in habitat types with different predation risk and competition levels.

Second, our approach allows testing how recruitment patterns of native trees with different seed dispersal modes can be influenced by Rhododendron presence.

We hypothesized that (1) seed contact probability would increase with Rhododendron presence due to its positive influence on mouse abundance (Malo et al., 2013); and (2) that the decision to cache or eat depends on relative handling cost, which covaries with seed size and habitat–specific predation and pilferage risk. More specifically, we predicted that lighter seeds (wind–dispersed) would be preferentially predated in the open and that larger nuts and acorns (animal– dispersed seeds) would be preferentially picked from the open and then cached under Rhododendron.

(3) Rhododendron areas, being safer foraging sites, would alter foraging behaviour, leading to asymmetric seed dispersal patterns, whereby dispersal from the open to the invasive–covered areas would be higher than the opposite possibility; (4) given that the shrub is known to impede seed germination of competitor native species and growth (Cross, 1981, 1982), we also predicted a deleterious impact on distribution and regeneration of saplings of native trees and a reduced density of samplings of native trees growing under Rhododendron cover.

Material and methods Study site and rodents The study site is located in a mixed deciduous woodland at Silwood Park, Ascot, England, UK (51º 24' 52.47'' N, 0º 38' 41.73'' W). This land is owned

by Imperial College London. The study area is 1.7 ha and is divided in 10 m x 10 m grids (100 m2 each, n = 170, fig. 1) and contains patches of the invasive Rhododendron (n = 6). We consider the different pat-

ches, to a large degree, independent spatial replicates because they are exposed mostly to different mice (71 % as inferred by the trapping data). However, as we could not determine how many rodents really interact with seeds at each patch, spatial independence is only expected. The total number of different individual rodents, wood mice (m, Apodemus sylvaticus) and bank voles (v, Myodes glareolus), trapped during July and August 2009 in each Rhododendron patch and in the open area was recorded. We consider the results of this study reflect seed foraging behaviour mostly of mice, first because voles are not seed dispersers and second, because diet studies have shown that during the summer seeds constitute 73 % of wood mice diet but only 19 % of bank voles diets (Watts, 1968). Each rodent had a unique fur clip mark for individual recognition.

Seed removal / fate experiments Dispersal by rodents of seeds and nuts is known to be influenced by seed type (Takahashi et al., 2007). We focused on caching and predation of two animal–dispersed (AD) seeds, pedunculate oak (Quercus robur), common hazel (Corylus avellana), and two wind–dispersed seed species (WD), sycamore maple (here forth sycamore, Acer pseudoplatanus) and European ash (Fraxinus excelsior). A total of 2,040 seeds (510 seeds per species; for more details see supplementary material) were individually tagged by piercing with a thin needle through the exocarp a 0.06 mm thick fluorocarbon coated string (80 mm) attached with a fluorescent orange tag (5 mm x 15 mm) labelled with a unique number. The fluorocarbon coating ensured there was no water absorption that could increase weight. The string length was minimised (80 mm) to reduce the chances of it getting entangled in the vegetation. Seed tagging has been shown to have a negligible effect on whether seeds were eaten or dispersed (Xiao et al., 2006; Morán–López et al., 2015, 2021). A group of 4 seeds was then placed in the centre of each 10 m x 10 m grid (N =170 grids x 3 temporal replicates = 510 seed stations in total). Each grid was classified as under Rhododendron cover (R), edge habitat (E) or open woodland (O) (fig. 1), which for rodents would represent low, intermediate and high predation risk, respectively. Seed fall for the four tree species occurs between mid–July and September (Malo, unpublished data; Gurnell, 1993).

To ensure that the experiment reflected natural rodent behaviour, the timing for experimental seed placement was selected to match start of the seed fall. Three seed trials were conducted: two in July (22nd and 29th) and one in August (5th) 2009. A 10–day pilot experiment was conducted in the field before the start of the experiment. The extent of seed removal conducted by species different to rodents was ascertained using 3 types of seed trays (accessible to invertebrates only, to invertebrates and rodents, and open to all). Removal rates by species different to rodents were negligible (see fig. 1s in supplementary material). Thus, open to all seed trays were used. In fact, further research on the study site testing habitat

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1

2

3

4 5 6

10 m4

Fig. 1. Study site located at Silwood Park. Irregular grey patches represent Rhododendron ponticum cover areas. Black, grey and white circles represent grids where seed stations were located under Rhododendron (n = 45), at the Rhododendron edge (n = 19) and in open woodland (n = 106), respectively. Seeds were located under the Rhododendron in areas where the Rhododendron does not cover the central area of a quadrant. Lines depict a seasonal 0.5–1.5 m wide water ditch. White numbers indicate patch identity.

Fig. 1. Zona del estudio situada en el parque Silwood. Las manchas irregulares de color gris representan las zonas cubiertas por Rhododendron ponticum. Los círculos negros, grises y blancos representan los cuadrantes en los que se colocaron estaciones de semillas debajo de Rhododendron (n = 45), en el margen de Rhododendron (n = 19) y en el bosque abierto (n = 106), respectivamente. En las zonas en que Rhododendron no cubría el centro de ningún cuadrante, las semillas se colocaron debajo de este arbusto. Las líneas representan un canal de agua estacional de 0,5 m–1,5 m de ancho. Los números de color blanco indican la identidad de los parches de Rhododendron.

differences in seed survival using motion–activated infrared cameras (Brouard and Malo, unpublished data) has shown that in a total of 1,215 hours of seed station footage recorded, not a single instance of seed predation by species other than rodents was recorded.

In each trial, 170 seed stations (R = 45, E = 19, O = 106) were set up, each one containing 4 seeds from each species (4 seeds x 170 grids = 680 seeds/trial).

These were arranged in a 10 cm x 10 cm square to ensure that all seeds were detected if the seed station was visited. One seed station was placed in each 100 m2 square. The plot was then left unvisited for three days (days 1, 2, and 3 after set up) to minimise disturbance. On day 4 each seed station was checked and the 100 m2 area per grid was checked for seeds. Grids where seeds had been removed had all microhabitat structures inspected. Grids with larger number of microhabitat features that hindered detection were searched for longer. This procedure was repeated on day 5. Seed stations were recorded as visited if they had at least one seed disturbed. Each disturbed seed was then classified as either eaten in situ (E, if the remains of the seed were found at the seed station), cached (C, if not found at the corresponding seed station but the tag and seed later retrieved) and two binomial variables, seed eaten (0–1) and seed cached (0–1), recorded. Regarding the

cached seeds after dispersal, we also recorded the new habitat type the seed was found in (R, E, O).

All seeds retrieved were assessed for teeth marks to ensure rodents were the consumers.

Data collection The following variables were recorded: (1) grid (n = 170): categorical variable naming each 100 m2 area where the seed–station was placed. (2) Rhododendron patch identity: a categorical fixed factor including six different patches to allow for spatial replication of Rhododendron effects (fig. 1). (3) Rhododendron patch size: a continuous variable to test for the effects of patch size (see supplementary material for patch size details). (4) volume of logs: defined as the sum of individual volume of logs (fallen trees) per grid using the cylinder formula (π · r · 2 · l; diameter = 2r and length = l); logs help protect rodents against aerial predation and aid silent travelling in the woodland, and this continuous variable (mean = 2.22, SD = 2.71 m3/grid) was included as a surrogate of habitat accessibility influencing seed encounter probability. (5) distance to Rhododendron: to characterize predation risk we calculated two variables: (i) the sum of the distances from each seed station to every other grid with Rhododendron presence (Rhodo dist) within a 20 m distance from the seed stations, and (ii) the total

58

number of grids with Rhododendron presence within a 20 m distance from the seed stations (Rhodo num).

(6) seed species: categorical variable (four levels: AP, FE, QR, CA). For some analyses, seeds were also ranked according to the total energy content of the endosperm (1, sycamore [lowest]; 2, ash; 3, oak; 4, hazel [highest]; table 3 from [Grodzinski and Sawicka–Kapusta, 1970]). And (7) seed encounter: a seed station was scored as visited (or encountered by mice) when at least one of the 4 seeds had been disturbed.

All the seedlings, saplings and adult trees of the four species present on the study area, were identified, measured and mapped. The number of ash (n = 7, WD) and oak individuals (n = 29, AD) was one order of magnitude smaller than that of sycamore (n = 370, WD) and hazel (n = 265, AD).

Given that the sparse distribution and small number of ash and oak trees could influence the absence of these species under Rhododendron, we discarded these two and focused on the more abundant and widespread sycamore and hazel. The number of saplings (diameter at breast height, dbh < 15 cm) of hazel and sycamore per grid (10 x 10 m) was recorded, and their frequency of occurrence and density per grid by habitat type were calculated and compared. In this analysis, grids were only considered as Rhododendron habitat if 100 % of their area was under the evergreen invasive canopy. Thus, some of the grids previously classified as R (because seeds were deployed well under Rhododendron cover), were coded as E (R = 23, E = 37, O = 110). Otherwise, individual trees located a couple of meters outside Rhododendron habitat, but falling within a grid classified as R (seed station placed under Rhododendron), would be analysed as if they were under cover.

Data analyses In all cases, generalized linear mixed effects models with binomial families were conducted and fitted by Laplace approximation in R (R Development Core Team, 2008). Stepwise model simplification was used to select the minimal adequate model (MAM) (Crawley, 2007). The MAM retained those terms that significantly increased deviance after removal. To account for spatial autocorrelation, grid was included as a random effect. However, because grid (random factor) and distance to Rhododendron (fixed covariate) were not expected to fully account for the spatial covariance inherent in all of the sampling points, the residuals of each final model were saved and used to test for spatial autocorrelation using variograms. For all models conducted, variogram methods confirmed the absence of spatial autocorrelation effects. Sample semivariogram, as derived from the residual variance in the model (Pinheiro and Bates, 2000), did not increase with increasing distance between sampled grids (fig. 3s in supplementary material). Initially, either patch identity or patch size were included in the same set of models to compare their effects. As there were no significant differences between models run with patch identity or size, we used patch identity in all subsequent models.

Malo et al.

Habitat effects on seed encounter probability A logistic regression model was constructed using the seed station as the sampling unit (lmer, binomial link function in R, N = 510): Seed–station visited (Y/N) ~ grid (random) + volume of logs + Rhodo num + Rhodo dist + habitat + trial + patch + log * habitat + patch * habitat + trial * habitat The non–significant effect of trial number in all of the models (P > 0.5) showed absence of temporal covariation effects (see fig. 4s in supplementary material).

Habitat and seed type effects on seed collection Once a seed station has been detected, seeds are either collected or not. Hence, to test for the determinants of seeds being collected, we used the following logistic regression models (N = 1,008): Collected (0–1) ~ grid (random) + volume of logs + Rhodo num + Rhodo dist + patch + trial + seed type + habitat + log * habitat + seed type * habitat + seed type * trial + seed type * patch Note that in these two models the binomial response captures two events per model: in the seed predation model a seed being eaten (1) vs. a seed not being eaten (0, cached or left untouched). In the seed caching model a seed being cached (1) vs. a seed not being cached (0, eaten or left untouched).

We also conducted a third model with a binomial response capturing the two actions conducted by mice at seed encounter (eating vs. caching). Thus, we included only seeds that were either eaten or cached and excluded those cases in which a rodent did not interact with the seed. Dispersal mode was included as a predictor at this stage of the analysis: Action (eaten/cached) ~ volume of logs + Rhodo num + Rhodo dist + patch + trial + habitat + dispersal type (AD/WD) + log * habitat + dispersal * trial + dispersal * patch dispersal * habitat + grid (random) Native seed dispersal asymmetry between habitats To test for the effect of habitat type at seed encounter on the habitat type where the seed was finally cached, we calculated the number and proportion of seeds that were cached in the same or different habitat type from where the seed was originally collected by rodents. In the cases in which caching or eating could not be determined in a contacted seed station (individual seeds = 758; seed stations = 213) these were excluded from the analysis. The expected and observed frequencies of seeds dispersed and cached between habitat types are reported. By expected we refer to the null hypothesis of no differences between dispersed and cached seeds by habitat type.

Animal Biodiversity and Conservation 45.1 (2022)

To address the effect of predation risk on caching behaviour (high in Open woodland, medium in Edge and low in Rhododendron) we used x2–tests (prop. trend.test function in R) to test for a significant trend in the proportion of seeds cached in O, E and R (by seed type).

Results During July and August 2009 we trapped 31 rodents, 23 wood mice and 8 bank voles in 5 Rhododendron patches (mean ± SD mouse number per patch = 7.33 ± 7.23; mean ± SD vole number per patch = 3 ± 2.89). Regarding mice, 23 appeared only in one patch, 6 appeared in two and 3 appeared in three patches. This suggests that the minimum spatial independence (between patches) achieved is at least of a 71 %, as 23 of the 32 mice only use seed stations from a single patch.

Habitat effects on seed encounter probability Overall, 49 % of the seed stations were visited by rodents (252/510). Habitat type strongly influenced seed encounter by rodents (table 1A). Rodents encountered 29 % of seed stations in the open habitat, 8 9 % on the edge and 77 % under Rhododendron (logistic regression model parameter estimates ± SE for Open = –1.24 ± 0.67; Edge = 2.87 ± 0.63; Rhododendron = 1.71 ± 0.71). Significantly fewer stations were detected in the Open than in the other two habitats (Open vs Edge, x2 = 86.95, df = 1, p < 0.0001 and Open vs Rhododendron x2 = 71.28, df = 1, p < 0.0001). There was a close–to–significant difference in seed encounter probability between the Rhododendron and the edge habitat (x2 = 3.59, df = 1, p = 0.06). There was also a marginally non–significant interaction between volume of logs and habitat (table 1A, fig. 5s in supplementary material). Further exploration of this interaction using habitat–specific linear regressions showed that the volume of logs had a strong negative effect on encounter probability under Rhododendron (linear model: param. estim. = –0.29 ± 0.13, p = 0.027), but not in the open and edge habitats. Neither trial nor patch identity had an effect on seed encounter probability (table 1A).

Habitat, seed type and seed dispersal type effects on seed predation and caching Overall, 1,032 seeds remained untouched (including those that had not been detected in the seed stations), 368 seeds were eaten, and 256 were cached (140 seeds were moved but not found, hence they were not scored as cached). Once a seed station was located, the probability that rodents would manipulate the seeds (eat or cache) was: ash (0.4) > hazel (0.39) > sycamore (0.23) > oak (0.19). Also, larger seeds (animal–dispersed) tended to be cached whereas the smaller, easier to handle seeds (wind– dispersed) tended to be eaten in situ (fig. 2). The first logistic regression model showed that the probability

59

of a seed being eaten was affected by habitat, seed type and their interaction (table 1B). A second model showed that the probability of being cached was influenced by the same factors (habitat, seed type and their interaction). Patch identity did not influence seed predation or seed caching (table 1B). The third model (binomial response: eaten or cached) showed that dispersal type and habitat had a significant effect on the probability of being eaten or cached (table 2).

There was also a significant interaction between habitat type and seed dispersal type on the probability of a seed being eaten or cached (table 2; fig. 3). Large seeds (hazel and oak) were cached more often than eaten in open and edge habitats and eaten more often than cached in Rhododendron habitat (for detailed results see fig. 6s in supplementary merial). For small seeds there was also a significant interaction (x2 = 49.6, df = 2, p < 0.00001) driven by the higher caching in open as compared to edge habitat, and higher eating in edge as compared to open habitat (x2 = 16.6, df = 1, p < 0.0001). Only the ash seeds were more frequently eaten than cached seeds in the open woodland habitat (fig. 6s in supplementary material).

Is native seed dispersal by rodents symmetrical between habitats?

Overall, the total number of seeds that were collected and cached in the same habitat type differed between habitats (Open = 66 %: 115 collected vs.

76 cached; Edge = 61 %: 54 collected vs. 33 cached; Rhododendron = 168 %: 87 collected vs. 147 cached; x2 = 28.42, df = 2, p < 0.0001). Significantly more seeds were cached in the Rhododendron than in the edge (x2 = 14.95, df = 1, p < 0.001) and in the open woodland habitats (x2 = 22.36, df = 1, p < 0.0001). No differences were observed between the edge and open habitat (p < 0.87). In the open woodland and edge, 33 % and 39 % of the collected seeds were cached elsewhere, respectively. The reverse pattern occurred in the Rhododendron, where all collected seeds appeared in the same habitat type, plus an extra 60 seeds not collected from Rhododendron habitat, representing a 69 % increase over the seeds originally deployed and cached in this habitat type. All seeds showed a higher caching probability than expected under Rhododendron, and lower than expected in the edge and open habitats (fig. 4A). There was a significant association between habitat type and caching probability for the four seed types (sycamore, x2 = 40.3; ash, x2 = 70.7; oak, x2 = 48.5; hazel, x2 = 61.6; in all cases df = 1, p < 0.0001). This shows that caching behaviour was asymmetrical across habitat types: the probability of a seed from open or edge being cached in Rhododendron was much higher than the probability of a seed from Rhododendron or edge being cached in open woodland.

Native tree sapling density differences between habitats We have shown above that hazel and sycamore seeds were consistently cached by rodents under

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Table 1. Results for three sets of generalized linear mixed effects models (binomial distribution) testing for the determinants of: A, seed station encounter (N = 510 observations); B, seed predation and seed caching (N = 1,138 observations [2040–902]); V.logs, volume of logs.

Tabla 1. Resultados de tres conjuntos de modelos mixtos lineales generalizados de los efectos (distribución binomial) en los que se analizan los factores determinantes de: A, modelo sobre el encuentro de las estaciones de semillas (N = 510 observaciones); B, modelos sobre la predación y el almacenamiento de semillas (N = 1.138 observaciones [2040-902]); V.logs, volumen de troncos.

A Seed station encounter model

x2

df

P

Volume of logs

1.13

1

0.28

Rhodo number

0.35

1

0.56

Rhodo distance

0.22

1

0.64

Habitat

84.96

2

< 0.0001

Patch

8.80

6

0.19

Trial

0.50

2

0.78

V.logs * habitat

5.79

2

0.055

Trial * habitat

< 0.5

4

0.9

Terms

B Seed predation and seed caching models Seed predation model

x

df

Volume of logs

0.39

1

Rhodo distance

0.71

1

Rhodo number

0.33

Terms

Seed Patch

Seed caching model

x

df

P

0.53

1.25

1

0.26

0.40

0.001

1

0.97

1

0.57

0.15

1

0.70

132.24

3

< 0.001

128.16

3

< 0.001

2

P

2

4.43

6

0.62

8.66

6

0.19

129.64

2

< 0.001

29.11

2

< 0.001

Trial

0.28

2

0.87

0.75

2

0.69

V.logs * habitat

3.89

3

0.27

1.65

2

0.44

Seed * patch

22.26

18

0.22

24.72

18

0.13

Seed * trial

3.49

6

0.75

2.47

6

0.87

Habitat * seed

42.97

6

< 0.001

51.75

6

< 0.001

Habitat

the evergreen Rhododendron cover. This leads to the expectation –which we test below– that hazel and sycamore seeds would have lower survival rates due to the dense shade of the invasive, and that there will be a lower density of saplings under Rhododendron.

Across the whole study site, a total of 643 hazel and 396 sycamore saplings were recorded. For both species, we found that the probability of sapling (< 15 cm) occurrence per grid was negatively related to

Rhododendron cover (test for a non–zero slope: hazel, Open = 64 %, Edge = 47 %, Rhododendron = 4 %, x2 = 7.78, df = 1, p = 0.005; sycamore, Open = 67 %, Edge = 55 %, Rhododendron = 4 %, x2 = 5.46, df = 1, p = 0.019). In the grids where saplings were present, their density by grid was also significantly lower in Rhododendron habitat than in the Edge and Open habitat (ANOVA test: hazel, F3,87 = 52.6, P < 0.0001; sycamore, F3,93= 30.02, P < 0.0001; fig. 4B).

Animal Biodiversity and Conservation 45.1 (2022)

A

0.36

61

C

0.30

FE

0.24 0.18 0.12 0.00

a

a CA

AP

1

Proportion of seed cached

Proportion of seed eaten

b

CA c

0.18 QR b

0.12

QR 0.00 2 3 4 Seed energy score B

d

0.24

a

FE

AP 1

2 3 4 Seed energy score

180 160 FE

Frequencies

140 CA

120 100

QR

80 60 40 AP

20

0

Eaten

Cached Seed fate

Fig. 2. Mice preference to eat less energy–rich wind–dispersed seeds on site, instead of caching them (A, B), and do the opposite with the more energy–rich animal–dispersed seeds (B, C): ash, highest eaten; sycamore, lowest cached; hazel, highest cached; Oak, lowest eaten). B, seed fate interaction between wind–dispersed (FE, ash, Fraxinus excelsior; AP, sycamore, Acer pseudoplatanus; light grey) and rodent– dispersed seeds (CA, hazel, Coryllus avellana; QR, oak, Quercus robur; dark grey). (Bars represent standard errors, N = 618, different lowercase letters indicate significant differences between factors).

Fig. 2. Preferencia de los ratones por consumir in situ las semillas anemócoras de menor contenido energético en lugar de almacenarlas (A, B) y hacer lo contrario con las semillas zoócoras de mayor contenido energético (B, C): fresno, la más consumida; arce, la menos almacenada; avellano, la más almacenada; roble: la menos consumida). B, destino de las semillas según su tipo de dispersión: anemócoras (FE, fresno, Fraxinus excelsior; AP, arce, Acer pseudoplatanus; en gris claro) y dispersadas por roedores (CA, avellano, Corylus avellana; QR, roble, Quercus robur; en gris oscuro). (Las barras representan el error estándar, N = 618, las letras en minúscula indican las diferencias significativas entre factores.)

Discussion Patches of invasive Rhododendron strongly influenced rodent foraging behaviour, as expected from refuge– mediated dynamics model. Habitat type influenced

the probability of rodents detecting seeds, with more seed stations being encountered under Rhododendron, probably due to its higher rodent density and lower risk of predation. Second, habitat type influenced whether different types of seeds were eaten or cached. Small,

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Table 2. Results from a set of generalized linear mixed effects models showing the effect of habitat, dispersal type and their interaction on the probability of seed eating and caching behaviour (binomial response): AD, animal– dispersed; WD, wind–dispersed.

Tabla 2. Resultados de un conjunto de modelos mixtos lineales generalizados de los efectos en los que se observa el efecto del hábitat, el tipo de dispersión y su interacción en la probabilidad de que se produzca el comportamiento de alimentación y de almacenamiento (respuesta binomial): AD, zoócora; WD, anemócora.

x2

df

P

Volume of logs

0.11

1

0.74

Rhodo distance

0.83

1

0.36

Terms

Rhodo number

0.74

1

0.39

136.34

1

< 0.001

Patch

4.59

6

0.59

Habitat

40.77

2

< 0.001

Trial

0.45

2

0.80

V.logs * habitat

2.93

2

0.23

Dispersal * patch

6.72

6

0.35

Dispersal * trial

0.10

2

0.95

Habitat * dispersal

8.34

2

0.015

Dispersal (AD/WD)

wind–dispersed seeds were mostly eaten in the open woodland, and proximity to Rhododendron patches increased predation rates, whereas large, animal– dispersed seeds were mainly cached. Large seeds were however 3–fold more likely to be eaten under Rhododendron, again probably due to higher rodent abundance (and hence cache pilferage) and lower perceived risk of predation (Morán–López et al., 2021).

Low predation risk under Rhododendron thus relaxed the potential effect of seed nutritive value, size or secondary compounds on seed preferences by allowing rodents enough safe time to circumvent potential seed defences by granivores (see Díaz 1996 for equivalent examples with other seed–eating animals). Third, large and small seeds were preferentially cached under the invasive Rhododendron. This generates anisotropic seed dispersal kernels, with potential implications for the spatial structure and dynamics of native trees.

Lower survival of seeds (this work) and seedlings (Cross 1981, 1982) under Rhododendron produced, however, lower densities of native tree saplings under Rhododendron cover, so that negative influences of the invasive plant extended well over their canopies though the foraging behaviour of rodents.

Native rodents play an important role in seed dispersal and native woodland regeneration (Jensen and

Nielsen, 1986), but this role can be modified by the biological integrity of the system where the interaction occurs (Morán–López et al., 2021). The prediction that consumer density increases with proximity to alien plants (Orrock et al., 2008) is supported by our results showing that mice present a 2–5 fold increased density under Rhododendron (Malo et al., 2013).

In our study, this should be related to a decreased risk of predation since food resources are not higher under the invasive plant (Holt, 1977). Rodents found seeds more frequently under Rhododendron and edge habitats than in the open woodland. The lack of an effect of Rhododendron patch identity suggests that different rodents behaved similarly in different Rhododendron patches, as of the 32 mice present at the time, 23 used only a single habitat patch.

The volume of logs per grid negatively affected seed encounter probability under Rhododendron, although not in the open woodland and edge habitats.

This result is consistent with previous results showing that under the invasive Rhododendron (which already provides cover from aerial predation) mice tend to avoid areas with a high abundance of fallen trees (Malo et al., 2013). We speculate that this avoidance behaviour of areas with high volume of logs under Rhododendron might be a strategy of rodents to minimize predation from terrestrial predators such as the least weasel (Mustela nivalis), their second most important predator after tawny owls (Strix aluco) (Korpimaki and Krebs, 1996; Godsall et al., 2014).

Under stressful conditions, seed size, germination schedule and tannins are the three main factors expected to influence rodent caching decisions (Marti and Armario, 1998). Although we did not collect data on handling times, our results support the seed size/ handling cost hypothesis (Jacobs, 1992). Overall, the trade–off between eating and caching was resolved towards eating the lighter, easier–to–open seeds of ash and sycamore, and towards caching the heavier and more nutritious animal–dispersed hazel and oak seeds in safer places. Preference for eating smaller seeds in situ and scatter–hoarding larger seed has also been shown in other rodents (Chang et al., 2009; Vander Wall, 2010), but these preferences vary under manipulated levels of competition and predation risk (Morán–López et al., 2021). Here we document how safer Rhododendron canopies interacted with seed choices and foraging decisions: preference for caching larger seed species in open habitat was reversed under Rhododendron cover, where predation risk by specialist aerial predators is minimized (Southern and Lowe, 1982), suggesting that predation risk influences the decision of eating or caching a seed: the cost of eating large seeds in the open was lower under safer Rhododendron cover. Under the invasive shrub, large seeds were more frequently eaten than cached, the opposite of what happens in the edge and open habitats.

Our study considered spatial replication at the Rhododendron patch level and temporal replication at the weekly level. However, two limitations of our experimental design are that it lacked between–site replication and between–year replication. Regarding

Probability of being eaten

Animal Biodiversity and Conservation 45.1 (2022)

1.0

WD

0.9

AD

0.8 0.7

63

b

b

a

a

0.6 0.5 0.4 0.3

c

c

0.2 0.1 0.0

Open

Edge

Rhodo

Fig. 3. Interaction between habitat and dispersal type on seed action (eaten vs. dispersal) (x2 = 8.34, df = 2, P = 0.01). Lines are left in the graph to highlight the change in slopes between habitats and seed types. Bars represent standard errors. Data includes only seed stations that were encountered by mice (Nseed station= 252). Different lower case letters indicate significant differences between factors: AD, animal–dispersed; WD, wind–dispersed.

Fig. 3. Destino de las semillas (consumidas o dispersadas) según la interacción entre el hábitat y el tipo de dispersión de las semillas (x2 = 8,34, df = 2, P = 0,01). Se dejan las líneas en el gráfico para destacar el cambio de las curvas entre hábitats de tipos de semillas. Las barras representan el error estándar. Los datos solo se refieren a las estaciones de semillas que fueron encontradas por ratones (Nestación de semillas = 252). Las letras en minúscula indican las diferencias significativas entre factore: AD, zoócora; WD, anemócora.

the lack of site replication, the lack of a patch effect albeit large variation in patch sizes, and the strength of all patterns observed makes us believe that these results are generalizable to other woodlands invaded

by Rhododendron. Regarding the between–year replicates, the strength of the observed patterns –seed detection probability and seed predation and caching drivers– and the lack of differences between within– year trials makes it unlikely for the results obtained to have substantially changed had the experiment been replicated between years. Background seed availability should not drastically alter our results. We have data showing that even after seed fall, rodents are absent from areas of the study site without shrub cover and fallen trees but completely loaded with beech seeds (under a >1 m diameter beech in the corner of my study site). This suggests that predation risk factors are more important drivers of foraging behaviour than seed abundance changes.

The preference for dispersing seeds under Rhododendron cover predictably has ecological consequences. Previous research suggests that the caching of seeds under the dense shade of Rhododendron would prevent them from germinating (Cross, 1981; Phillips and Murdy, 1985; Rotherham and Read, 1988). To test this possibility, we identified and measured every single tree sapling on the study site, comparing the differences in mean density/grid

of sycamore and hazel (the most abundant trees for each seed dispersal type). The results are striking; both native tree occurrence and density decreased from the open woodland to the Rhododendron, where the majority of the seeds were cached. Bamboo also provides refuge to small–mammal seed consumers, and has been shown elsewhere to reduce tree seed and seedling survival within bamboo dominated habitats (Wada, 1993; Caccia et al., 2006, 2009). This consistent asymmetric seed dispersal from open and edge habitats to Rhododendron, together with the absence of hazel and sycamore tree saplings under Rhododendron, suggests that seed survival is greatly reduced. This modifies the spatial structure of the native tree community and has potential implications for the dynamics and composition of communities (Tilman, 1994). Regarding the edge habitat, where competition for light between native and the invasive is more intense, we have shown, first that seeds were removed from the edge, transported and consistently cached under the invasive; second, that the edge also presents a significantly lower number of saplings than in the open woodland. This can have important implications; given that Rhododendron spreads vegetatively by lateral horizontal growth of the branches (Pierik and Steegmans, 1975; Cross, 1981; Mejias et al., 2002), native seed removal in the edge habitat could prevent future native seedlings from growing,

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64

40 30 20

B

AP FE QR CA

10 0 –10 –20 –30 –40

Open

6

Native tree samplins (< 15 cm) per 100 m2

Deviation from the expected proportion of seeds cached (%)

A

Edge

Rhodo

5

CA AP

a

4 3

b

b

b

2 1 0

c Open

Edge

c

Rhodo

Fig. 4. A, deviation from the expected proportion of seeds cached (%) (by species) in each habitat type.

Zero indicates the null expectation of no habitat effect on caching behaviour: seeds collected from a certain habitat type are expected to be cached in that same habitat. Mean percentage and standard error bars calculated from the three independent seed release trials conducted: AP, sycamore, Acer pseudoplatanus; FE, ash, Fraxinus excelsior; QR, oak, Quercus robur; CA, hazel, Corylus avellana. B, mean density per 100 m2 of hazel and sycamore saplings (< 15 cm) recorded in each habitat type (open woodland, edge habitat and Rhododendron).

Fig. 4. A, desviación de la proporción esperada de semillas almacenadas (%) (por especie) en cada tipo de hábitat. Cero significa que no se espera ningún efecto del hábitat en el comportamiento de almacenamiento: se prevé que las semillas recogidas en un determinado tipo de hábitat se almacenen en ese mismo hábitat. Barras de porcentaje medio y error estándar calculadas a partir de los tres estudios independientes realizados de introducción de semillas en el medio: AP, arce, Acer pseudoplatanus; FE, fresno, Fraxinus excelsior; QR, roble, Quercus robur; CA, avellano, Corylus avellana. B, densidad media por 100 m2 de plántulas de avellano y arce (< 15 cm) registrada en cada tipo de hábitat (bosque abierto, hábitat de transición y Rhododendron).

helping the invasive shrub by reducing competition for light and potentially increasing the rate of spread of the invasive shrub throughout the woodland understory. However, this possibility remains to be tested in the field.

By demonstrating Rhododendron's effect on mouse caching behaviour, our results suggest that the presence of native rodent communities in Rhododendron invaded areas might increase the invasibility of these habitats by (1) reducing the survival of native trees and (2) eliminating competition for light at the edge of Rhododendron patches, where more native seeds were removed thus preventing native tree recruitment. A second factor decreasing seed survival is their higher consumption under the invasive plant compared to the open woodland, which may be partly explained by conspecific pilferage (Leaver and Daly, 2001; Morán–López et al., 2021). However, our results cannot be extrapolated to very old or large Rhododendron patches. First, old patches that have precluded native seed regeneration should have lower seed densities as a result of having had no native tree regeneration for decades, and provide less food for mice. Likewise, very large patches over

the home range size of mice should make it difficult for them to forage in the edge and open habitat. It is likely that Rhododendron monocultures would result in declines in mouse population density due to the reduction in native tree seed supply. Previously we have shown that another important food resource during the spring and summer period, edible invertebrate biomass, is also reduced under Rhododendron (Malo et al., 2013). Hence, refuge–mediated interactions described here might be influential during initial and intermediate stages of Rhododendron vegetative spread under the woodland understory, and strictly to spread due to lateral growth, and not due to new invasions through colonisation of Rhododendron seeds (Stephenson et al., 2006).

Previous studies have shown how native species behaviour can mitigate the impact of invasive seaweeds (Wright et al., 2010). The present study is one of the few (Chaneton and Bonsall, 2000; Orrock et al., 2010) reporting the opposite; an ecological mechanism –unbalance of conditional mutualism towards its antagonistic side– through which native fauna could increase the deleterious impact of an invasive species by disrupting native woodland regeneration.

Animal Biodiversity and Conservation 45.1 (2022)

Acknowledgements We acknowledge two anonymous referees, Jacques Deere, Lochran Trail and Prof. Tim Coulson for providing constructive comments to a previous draft; Prof.

Mick Crawley for information about the history of the Rhododendron patches at the study site; Paul Beasley for help constructing the boxes for the foraging experiments conducted in the lab and Ben Godsall, Gavin Kingcome, and Lukasz Lukomski for help with habitat data collection. This research was funded by a ERC Marie Curie fellowship (PIEF–GA–2008–220322).

During the write up of this manuscript, AFM was partly supported by an ERC grant (249872) and by a Ramón y Cajal research contract from the MINECO (RYC–2016–21114).

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Ramírez–Arrieta et al.

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Recombinase polymerase amplification combined with fast DNA extraction for on–spot identification of Deinagkistrodon acutus, a threatened species

Y. Liang, H. Ye, X. Cai, E. Tian, F. Li, C. Li, G. Huang, Z. Chao Liang, Y., Ye, H., Cai, X., Tian, E., Li, F., Li, C., Huang, G., Chao, Z., 2022. Recombinase polymerase amplification combined with fast DNA extraction for on–spot identification of Deinagkistrodon acutus, a threatened species. Animal Biodiversity and Conservation, 45.1: 69–78, Doi: https://doi.org/10.32800/abc.2022.45.0069 Abstract Recombinase polymerase amplification combined with fast DNA extraction for on–spot identification of Deinagkistrodon acutus, a threatened species. This study addresses the use of recombinase polymerase amplification combined with fast DNA extraction for on–spot identification of Deinagkistrodon acutus, a snake species threatened due to over–exploitation and habitat destruction. For its conservation, an efficient species identification method is urgently neededto fight against illegal capture and trade. Fourteen individuals representing 12 snake species (including D. acutus and other snake species) were collected from mountainous regions in Southern China. Genomic DNA was extracted within five minutes by a modified alkaline lysis method. Species–specific primers for recombinase polymerase amplification (RPA) were designed based on the sequences of cytochrome C oxidase subunit I (COI) barcode region, and an optimized RPA assay system was set up. Specificity and sensitivity of the assay were checked, and the assay was validated by identifying 10 commercial Qi She crude drug samples derived from D. acutus. Under optimized RPA conditions, a distinct single band of 354 bp was amplified only for D. acutus but not for the related snake species. The entire procedure can be completed in 30 min at room temperature. Commercial Qi She crude drug identification validated effectiveness of the established assay system. Using a recombinase polymerase amplification (RPA) assay with rapid DNA extraction, we established an on–spot D. acutus identification method with good specificity and sensitivity. This method could become an efficient tool for rigorous supervision of illegal D. acutus capture and trade.

Key words: Deinagkistrodon acutus, Recombinase polymerase amplification, Rapid DNA extraction, On–spot identification, Conservation Resumen La amplificación por recombinasa y polimerasa combinada con la extracción rápida de ADN para la identificación sobre el terreno de Deinagkistrodon acutus, una especie en peligro de extinción. En este estudio se aborda la utilización de la amplificación por recombinasa y polimerasa (RPA en su sigla en inglés) en combinación con la extracción rápida de ADN para identificar sobre el terreno a Deinagkistrodon acutus, una especie de serpiente que se encuentra en peligro de extinción debido a la sobreexplotación y la destrucción del hábitat.

Con vistas a su conservación, es muy conveniente disponer de un método eficiente de identificación que permita combatir la captura y el comercio ilegales de la especie. Se capturaron 14 ejemplares de 12 especies de serpiente (incluida D. acutus) en las regiones montañosas del sureste de China. Se extrajo el ADN genómico con un método de lisis alcalina que llevó cinco minutos. Se diseñaron cebadores para la RPA, específicos de la especie, a partir de las secuencias de la región del código de barras de la subunidad I de la citocromo c oxidasa, y se estableció un sistema optimizado de análisis mediante RPA. Se comprobaron la especificidad y sensibilidad del ensayo, que se validó mediante la identificación de 10 muestras comerciales de la sustancia sin elaborar conocida como Qi She derivada de D. acutus. En condiciones de RPA optimizadas, se amplificó una banda única de 354 pb solo para D. acutus, pero no para las especies de serpiente relacionadas. El proceso se puede llevar a cabo en 30 minutos a temperatura ambiente. La identificación de la sustancia comercial sin tratar Qi She permitió validar la eficacia del sistema de análisis establecido. Al combinar el análisis mediante RPA con la extracción rápida de ADN, establecimos un método de identificación de D. acutus sobre el terreno ISSN: 1578–665 X eISSN: 2014–928 X

© [2022] Copyright belongs to the authors, who license the journal Animal Biodiversity and Conservation to publish the paper under a Creative Commons Attribution 4.0 License.

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con buena especificidad y sensibilidad. Podría convertirse en un instrumento eficiente para llevar a cabo la supervisión rigurosa de la captura y el comercio ilegales de D. acutus.

Palabras clave: Deinagkistrodon acutus, Amplificación por recombinasa y polimerasa, Extracción rápida de ADN, Identificación sobre el terreno, Conservación Received: 26 VI 21; Conditional acceptance: 30 IX 21; Final acceptance: 11 I 22 Yongshan Liang, Haoting Ye, Xuan Cai, Zhi Chao, Department of Pharmacy, Zhujiang Hospital, Southern Medical University, Guangzhou 510282, China.– Yongshan Liang, Enwei Tian, Fang Li, Chan Li, Guangsheng Huang, Zhi Chao, School of Traditional Chinese Medicine, Southern Medical University, Guangzhou 510515, China.– Zhi Chao, Guangdong Provincial Key Laboratory of Chinese Medicine Pharmaceutics, Guangzhou 510515, China.– Haoting Ye, Department of Pharmacy, Guangdong Second Provincial General Hospital, Guangdong Provincial Emergency Hospital, Guangzhou 510317, China.– Xuan Cai, Department of Pharmacy, Kweichow Moutai Staff Hospital, Zunyi 564501, China.

Corresponding author: Zhi Chao. E–mail: chaozhi@smu.edu.cn

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Introduction Deinagkistrodon acutus (Günther, 1888), the sharp– snouted pit viper, is the only representative of the monotypic genus Deinagkistrodon. More than 90 % of the populations of this species in the world occur in the Chinese provinces south of the Yangtse River, i.e., Zhejiang, Anhui, Jiangxi, Hubei, Hunan, Fujian, Taiwan, Guangdong, Guangxi, Chongqing, and Guizhou (Huang et al., 2007; Uetz et al., 2010; Roskov et al., 2019). This highly venomous species is well known in traditional chinese medicine for the effectiveness of its dried body (Agkistrodon, known as Qi She in Chinese, QS) in treating rheumatoid arthritis, stroke symptoms (such as limb numbness and spasm, eye and mouth drooping, facial paralysis, and hemiplegia), clonic convulsions (Chinese Pharmacopoeia Commission, 2015) and, of particular note, for treating hepatocellular carcinoma (Xu et al., 2020).

In recent years, there has been an ever–increasing demand for the QS crude drug, leading to over–exploitation of wild D. acutus (Zhou and Jiang, 2004). As the species has a narrow distribution, specialized habitat requirements, and low fecundity, D. acutus is facing both serious social–economical pressure and natural pressure. It has been estimated that the population was reduced by at least 30 % in the decade before 2004 (Huang et al., 2007). As a result, D. acutus was listed in China Species Red Book as ‘Endangered’ (Zhao, 1998). A later study further confirmed its vulnerability and assessed it as at high–risk (Zhou and Jiang, 2005). In the Biodiversity Red List of China: Vertebrates published in May 2015, it was again evaluated as 'Endangered' (Ministry of Environmental Protection and Academy of Sciences of China, 2015).

Conservation and protection of D. acutus is therefore of increasing concern. More recently, the National Forestry and Grassland Administration proposed to recognize D. acutus as a second class national–level protected animal (National Forestry and Grassland Administration of China, 2019). Efforts should be strengthened to combat its illegal capture, trafficking, and trade (Gong et al., 2020).

In the practice of law enforcement struggling against the illegal trade, officers often encounter lawbreakers who argue that what they are selling is not D. acutus but other snake species with similar body size and appearance, including related vipers such as Gloydius halys, G. intermedius, Ovophis monticola, Daboia siamensis, and even some rat snakes or elapid snakes with larger body size, such as Orthriophis moellendorffi, Euprepiophis mandarinus, Ptyas mucosus, and Naja atra (Su et al., 2016). It is especially difficult for the enforcement personnel to identify D. acutus once the snakes have been eviscerated and dried, and the spots and color pattern on the skin have disapeared. An accurate and fast on spot identification of species would provide the indispensable technical support for D. acutus conservation.

Conventionally, the identification of D. acutus is largely based on morphology or histology, which rely heavily on the experience of examiners and the

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result is often subjective (He, 1991; Xu and Gao, 1992; Wang, 1995). In recent years, DNA–based identification methods such as species–specific PCR (Tang and Feng, 2006) and DNA barcoding (Liao et al., 2013) have been developed for its identification, and the Chinese Pharmacopoeia has documented the use of a highly specific PCR method (State Pharmacopoeia Committee, 2010). A recent article

by Jiang et al. (2015) described a homogeneous fluorescent specific PCR method using cationic conjugated polymer for the identification of medicinal snakes including D. acutus. Nevertheless, all these molecular approaches for D. acutus identification do not allow efficient or convenient on–spot identification (Huang et al., 2014). Chen et al. (2014) recently reported a rapid PCR method to identify medicinal snake species, but this approach is still based on the classic PCR and requires the use of a PCR thermoclycler with high denaturation and annealing temperatures. Ideally, a rapid on–spot assay is expected to allow identification at room temperature within a short period of time and without requiring special equipment.

As one of the latest isothermal nucleic acid amplification techniques, recombinase polymerase amplification (RPA) shows much potential for application in on–spot identification of the species of medicinal snakes (Qin et al., 2017). RPA works with three core enzymes (a recombinase, a single–stranded DNA–binding protein (SSB) and a strand–displacing polymerase), and the reaction can take place over a wide range of ambient temperatures (Piepenburg et al., 2006). RPA shows some advantages over other isothermal nucleic acid amplification techniques (Liang et al., 2017). For example, loop–mediated isothermal amplification (LAMP) requires four primers aiming at six specific DNA binding sites in the target gene, so the primer design for LAMP is complicated; it is difficult to acquire species specific primers for differentiating close–related species with high sequence similarity (Notomi et al., 2000; Wu et al., 2016; Lee, 2017). In comparison, only one pair of primers is needed for RPA, and the design is almost as simple as conventional PCR. Like PCR amplification, RPA assay also has a high specificity, but it is much faster; in some cases, the result can be generated in 3–10 min (Daher et al., 2016). In addition, an RPA assay using a commercial kit does not require the use of an expensive thermocycler or any additional equipment or reagents (Fan et al., 2016). These advantages make RPA an ideal option for on–spot inspection in clinical diagnosis, laboratory medicine, forensic science, and the food industry (Gao et al., 2016; Ammour et al., 2017; Raja et al., 2017). Tian et al. (2017) have successfully extended its application to the field of medicinal plant identification.

To the best of our knowledge, no other report or publication has been available to describe the use of RPA for identification and conservation in threatened or endangered animal species. In this study, we established an RPA–based method combined with fast DNA extraction for on–spot D. acutus identification that can be completed in 30 min at 37 ºC.

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Material and methods Materials Fourteen specimens representing 12 snake species, including three specimens of Deinagkistrodon acutus (A1–A3) and 11 specimens of other snake species (A4–A14) (table 1), were collected from the mountainous regions in Guangdong, Hunan, Jinagxi and Guangxi in South China with the approval of the National Natural Science Foundation of China (NSFC No. 81573540). Vouchers were identified by Dr. Liang Zhang from Guangdong Institute of Applied Biological Resources and deposited in School of Traditional Chinese Medicine, Southern Medical University (Guangzhou, China). The snakes were euthanized through rapid cooling and freezing (Lillywhite et al., 2017). All these specimens were preserved in 75 % ethanol. In addition, 10 crude drug samples of QS (S1–S10) were purchased from local pharmacy or crude drug market in various cities of China (table 2).

The TwistAmp® Basic RPA kit, a product of TwistDx™ Limited (Cambridge, UK), was used in this study.

RPA primer design The species–specific RPA primers for D. acutus were designed with the aid of Primer Premier 5.0 software.

The primers targeted at cytochrome C oxidase subunit I (COI) barcode region, which provides abundant variation sites to distinguish D. acutus from other snake species. We had sequenced the COI region of each specimen, except for that of Daboia siamensis,

which was downloaded from GenBank. The accession numbers are listed in table 1. The primers were synthesized by Invitrogen Biotechnology (Shanghai) Co., Ltd. DNA extraction For each sample, about 30 mg of muscular tissue was dissected from the dorsal muscle. The total genomic DNA was extracted from 30 mg of tissue sample with a modified alkaline lysis method (Jiang et al., 2013). In brief, the homogenate tissues and 20 µL of extraction buffer (containing 0.5 mol/L NaOH, 1 % polyvinyl polypyrrolidone, and 1% Triton X 100) were added to a 200–µL PCR tube and vortexed for 10–15 s, followed by incubation in a boiling water bath for 10–15 s; 80 µL of Tris–HCl (0.1 mol/L, pH 8.0) was added to the cooled–down mixture, which was gently vortexed and then centrifuged for 5 min at 300xg. The supernatant containing the DNA was collected and the quality and concentration of the DNA solution were assessed using 1.5 % agarose gel electrophoresis and a NanoDrop 1000 UV/Vis spectrophotometer (Thermo Scientific, Wilmington, DE, USA), respectively.

Recombinase Polymerase Amplification (RPA) Following the instruction manual of TwistAmp® Basic (TwistDx Ltd., Babraham, UK) kit, RPA was performed with a total reaction volume of 50 µL containing the forward and reverse RPA primers (10 μM, 2.4 µL each), rehydration buffer (29.5 µL), magnesium acetate (280 mM, 2.5 µL), genomic DNA

Table. 1. Information about the species and sites where snake samples were collected: No, sample number; Genbank No, Genbank accession number.

Tabla 1. Información sobre la especie y el lugar de recogida de las muestras de las serpientes: No, número de la muestra; Genbank No, numero de acceso de Genbank.

No

Species

Voucher

A1 A2 A3 A4 A5 A6 A7 A8 A9 A10 A11 A12 A13 A14

Deinagkistrodon acutus TP–13 D. acutus TP–8 D. acutus TP–7 Protobothrops mucrosquamatus CH–1 Gloydius brevicaudus ZB–9 Trimeresurus stejnegeri ZS–13 Daboia siamensis DS–1609 Naja atra FC–2 Bungarus fasciatus TP–1 Orthriophis moellendorffi TP–9 Ptyas dhumnades TP–11 Ptyas mucosa DS–11 Lycodon rufozonatus ZS–3 Lycodon ruhstrati HB–Lr–1109171

Site of collection

Genbank No.

Hunan Guangxi Jiangxi Conghua, Guangdong Yongzhou, Hunan Zhongshan, Guangdong Zhongshan, Guangdong Taishan, Guangdong Guangxi Guangxi Conghua, Guangdong Zhongshan, Guangdong Zhongshan, Guangdong Conghua, Guangdong

JQ658433 JQ658432 JQ658431 JX233625 KC841453 JX233623 KP772294 JN833603 JN833615 JN833617 JX233651 JX233645 JN833598 JX233634

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Table. 2. Details of the commercial Qi She crude drug samples identification: + positive amplification; – negative amplification.

Tabla 2. Información detallada sobre la identificación de las muestras comerciales de la sustancia sin elaborar Qi She: + amplificación positiva; – amplificación negativa.

Sample No.

Location of acquisition

S1

Jian Qi Pharmacy

RPA results

BOLD identification results

–

Euprepiophis mandarinus

(= Elaphe mandarinus) S2

Oriental Pharmacy

+

Deinagkistrodon acutus

S3

Min Xin Pharmacy

+

Deinagkistrodon acutus

S4

Bao Zhi Lin Pharmacy

–

Euprepiophis mandarinus

S5

Ping Shan Tang Pharmacy

–

Naja kaouthia

S6

Zhangshu Crude Drug Market

+

Deinagkistrodon acutus

S7

Qingping Crude Drug Market

–

Lycodon rufozonatus

S8

Ji Shi Tang Dispensary

+

Deinagkistrodon acutus

S9

Bozhou Crude Drug Market

–

Daboia russelii

S10

Changcheng Pharmacy

–

Lycodon rufozonatus

(3.2 µL), and ddH2O (10.0 µL). With an initial incubation at 37º for 5 min, the reaction tube was taken out and shaken for 8–10 times at room temperature, followed by further incubation at 37º for 20 min. The RPA products were detected using 1.5 % agarose gel electrophoresis.

We tested different temperatures to optimize the amplification of the genomic DNA of a D. acutus sample (A1). Following the protocols for RPA reaction described above, temperature gradients of 37, 32, 27, and 22 ºC were tested for amplification for 20 min. The amplification time was optimized with a gradient of 10, 15, 20, 30, and 40 min at the optimal temperature using the same sample.

The specificity of RPA assay was evaluated using genomic DNA templates from D. acutus samples (A1–A3) and the other species (A4–A14). With the specific primers, the amplification protocols were identical to those described above. The primers were considered specific when the amplification signals were observed only in D. acutus samples but not in the related snake species.

For sensitivity assessment of the RPA assay, serially dilutions (100, 10, 1, 0.1, and 0.01 ng/µL) of the genomic DNA of D. acutus sample A1 were prepared to test the minimum detection limit of the assay with identical amplification protocols.

Identification of Qi She crude drug samples by RPA Ten commercial QS crude drug samples (S1–S10; table 2) were collected from the pharmacy stores or crude drug market in Guangzhou, Xingning (Guangdong), Zhangshu (Jiangxi), and Bozhou (Anhui).

RPA amplification was performed with the D. acutus –specific primers under the optimized temperature and reaction time.

To verify the accuracy of RPA assay, the zoological origin of the samples was identified by COI barcode.

The genomic DNA from those samples was amplified with COI barcode universal primers LCO1490 and HCO2198 (Folmer et al., 1994) or primers for Viperidae species DK1–CO1 and DK1–CO2 (Li et al., 2015; Cai et al., 2016). The PCR reaction mixture consisted of 12.5 µL of 2× Tap PCR MasterMix (with dye) (Tiangen, Beijing), 1 µL of each primer (10 μM), 2 µL of DNA template and 8.5 µL of ddH2O. Amplification was performed with an initial holding at 94 ºC for 5 min followed by 35 thermal cycles of 94 ºC for 30 s, 53 ºC for 1 min, and 72 ºC for 1 min, with a final extension at 72 ºC for 7 min. The PCR products were then sequenced. Species were identified using BOLD (Barcoding of Life Database) system V4.

Results Template DNA quality The concentrations of genomic DNA exracted from the samples were between 56 to 81 ng/µL, and the ratio of absorbance at 260 nm and 280 nm (A260/A280) ranged from 1.69 to 2.01.

Design and specificity assessment of RPA primers Successful specific amplification for D. acutus was obtained using the primer pair below:

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A

M A1 A2 A3 A4 A5 A6 A7 A8 A9 A10 A11 A12 A13 A14 NTC

B

M SD NTC S1 S2 S3 S4 S5 S6 S7 S8 S9 S10

400 bp

400 bp

Fig. 1. Electrophoretogram of RPA products: A: RPA of D. acutus and related snake species; A1–A3, D. acutus; A4–A14, other snake species; NTC, negative control (distilled water); M, 100 bp DNA ladder. B: RPA identification of 10 commercial QS samples; S1–S10, QS samples; M, 100 bp DNA Ladder I; SD, positive control; NTC, blank control.

Fig. 1. Electroforetograma de los productos de la RPA: A: RPA de D. acutus y las especies de serpiente relacionadas; A1–A3, D. acutus; A4–A14, otras especies de serpiente; NTC, control negativo (agua destilada); M, marcador de peso molecular de ADN de 100 pb. B: Identificación mediante RPA de 10 muestras comerciales de QS; S1–S10, muestras de QS; M, marcador de peso molecular de ADN de 100 pb; SD, control positivo; NTC, control.

QSCOAII: 5'–TTACTCCTATTACTATCCTCCTCCTACATC–3' QSCOF1130A: 5'–GCCCCCTCCGCTCGGATCAAAGAAGGTGGTGTTAAG–3' The binding sites of the forward and reverse primers were located at 12–50 nt and 335–355 nt of the COI region, respectively. Under optimized RPA conditions, a distinct single band of 354 bp was amplified using the primer pair only for D. acutus but not for the other snake species (fig. 1A), demonstrating the high specificity of the primers.

RPA optimization and sensitivity The results showed that the target fragment could be amplified at a temperature range of 27–42 ºC, and the optimal temperature range was 37 to 42 ºC (fig. 2A). From a practical point of view, we chose 37 ºC as the optimal reaction temperature for convenience of on–spot RPA. At this temperature, the reaction time was optimized and the results showed that amplifications for 15 to 40 min all yielded single and clear bands, whereas a reaction time of 10 min resulted in a rather weak signal (fig. 2B). We thus chose 15 min as the optimal reaction time for RPA.

The sensitivity of RPA was assessed using the optimized reaction temperature and time setting.

The target fragment could be successfully amplified from the serially diluted genomic DNA template with concentrations ranging from 100 to 0.1 ng/ µL; the target fragment failed to be amplified when the template was further tenfold diluted to 0.01 ng/ µL (fig. 2C). This result suggests that the lower detection limit of RPA was 0.1 ng/µL for D. acutus genomic DNA. Identification of commercial QS crude drug samples RPA of the 10 commercial QS samples with the optimized settings yielded a band of 354 bp for samples S2, S3, S6, and S8, and no amplification product was obtained for the other samples (table 2). These results suggested that samples S2, S3, S6, and S8 were authentic crude drug of D. acutus, while the others were the related species (fig. 1B).

DNA barcoding confirmed the results of diagnostic RPA with samples S2, S3, S6, and S8 derived from D. acutus. The identities of other samples were determined by DNA barcoding, and the results are listed in table 2.

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A M 1 2 3 4 5

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B M 1 2 3 4 5

C M 1 2 3 4 5

400 bp

Fig. 2. Agarose gel electrophoresis of the RPA product for amplification of the COI region: A, RPA reaction temperature optimization (1, 22 °C; 2, 27 °C; 3, 32 °C; 4, 37 °C; 5, 42 °C; M, 100 bp DNA ladder). B, optimization of RPA reaction time (1, 10 min; 2, 15 min; 3, 20 min; 4, 30 min; 5, 40 min; M, 100 bp DNA ladder). C, sensitivity of RPA assay for serially diluted genomic DNA of D. acutus. (Lanes 1–5, genomic DNA of 100, 10, 1, 0.1, 0.01 ng/µL, respectively; M, 100 bp DNA ladder).

Fig. 2. Electroforesis en gel de agarosa del producto de la RPA para la amplificación de la región COI. A, optimización de la temperatura de reacción de la RPA (1, 22 ºC; 2, 27 ºC; 3, 32 ºC; 4, 37 ºC; 5, 42 ºC; M, marcador de peso molecular de ADN de 100 pb). B, optimización del tiempo de reacción de la RPA (1, 10 min; 2, 15 min; 3, 20 min; 4, 30 min; 5, 40 min; M, marcador de peso molecular de ADN de 100 pb). C, sensibilidad del análisis mediante RPA del ADN genómico diluido en serie de D. acutus.

(Carriles 1–5, ADN genómico de 100, 10, 1, 0,1 y 0,01 ng/µl, respectivamente; M, marcador de peso molecular de ADN de 100 pb).

Discussion Over–exploitation led by increasing demands for D. acutus as a medicinal product has imposed great pressure on survival of this species (Yin et al., 2015).

In addition, habitat destruction has worsened its populations (Hu et al., 2013). In pace with its listing as a threatened species, more powerful and effective measures should be taken to ensure its conservation.

Species identification is a prerequisite for the enforcement of conservation and effective trade monitoring on regulated or protected animals and plants (Palumbi and Cipriano, 1998; Hsieh et al., 2001; Chapman et al., 2003; Wei and Szmidt, 2013). A rapid and accurate identification method will serve as a technical aid to help conserve D. acutus (Huang et al., 2014).

In this study, we established an RPA system combined with fast DNA extraction for rapid on–spot identification of the highly valued crude drug Qi She, the most common D. acutus product. The amplification results of the genomic DNA from D. acutus and other snake species showed high species specificity of the designed primers (fig. 2A). The optimal reaction temperature for RPA was 37 ºC, but successful amplification also occurred at a temperature as low as 27 ºC when the reaction time was extended, suggesting that an RPA assay could be conducted at room temperature or using only a heater when the environmental temperature is below 27 ºC. A visible target band could be generated after a reaction time of 10 min, and

amplification for 15 min resulted in a clearer band for effective detection. As the entire amplification can be completed in 15 min at room temperature without using any sophisticated instrument, this RPA assay provides an efficient and highly specific means for on–spot QS crude drug identification.

The manufacturer of the commercial RPA kit (TwistDx) provides three options for detection using agarose gel electrophoresis (TwistAmp® Basic), real–time fluorescent probes (TwistAmp® exo), or lateral flow strips (TwistAmp® nfo) (Piepenburg et al., 2013). Real–time fluorescent detection requires a probe and a real–time fluorescence analyzer as high–resolution melting (HRM) analysis does (Maslin et al., 2015; Wei et al., 2017). Lateral flow detection requires additional lateral flow strips, which increase the cost of the assay. The low cost and easy accessibility of gel electrophoresis is more appropriate for on–spot detection.

Jiang et al. (2014) reported a rapid PCR identification of Flos Lonicerae (flower bud of Lonicera japonica), in which fluorescent dye SYBR Green I was used to detect the amplified product. They added SYBR Green I dye after PCR amplification, and the genuine crude drugs generated a positive amplification signal by emitting green fluorescence when irradiated with 365 nm UV light, while the fake drugs did not produce any fluorescence due to non–amplification. Detection with the fluorescent dye was simple and fast, but our attempts at using SYBR Green I all failed, that is, of

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all the QS samples, either authentic or fake, the RPA products gave strong fluorescence signals. This might be attributed to the formation of dimers between the long RPA primers, which could be stained by the highly sensitive and nonspecific fluorescent dye (Li et al., 2019, 2020). In contrast, gel electrophoresis detection does not suffer so much from dimers.

A practical on–spot molecular identification method for crude drugs also needs to consider the time spent in DNA extraction. Conventional DNA extraction involves a complex procedure that includes several rounds of precipitation and centrifugation, which can be time–consuming. DNA extraction can take more than four hours with the CTAB method (Cota–Sánchez et al., 2006) and the SDS method (Yang et al., 2008), and 1–2 h with the commonly used silica gel column method (Abdel–Latif and Osman, 2017).

Chelex®100 resin–based extraction is a simple and rapid method for genomic DNA extraction from animal tissues. It can be performed in 30 min–1 hour, and shows excellent performance in difficult samples such as forensic samples (Li et al., 2019; Simon et al., 2020). However, alkaline lysis has also proved to be a quick and inexpensive single–tube method for extracting DNA (Klintschar and Neuhuber, 2000). The protocol developed in recent years contains only two steps: lysis and neutralization, and it can be completed in about five minutes (Jiang et al., 2013; Chen et al., 2014). For this reason we used the alkaline lysis method in this study for efficient DNA extraction. The extracted genomic DNA had a concentration above 50 ng/µL with the A260/A280 values ranging from 1.69–2.01, suggesting that both the DNA quantity and quality met RPA requirements (Lucena–Aguilar et al., 2016).

We tested 10 commercial QS crude drug samples following the established on–spot identification assay system. The results were confirmed by COI barcode identification, suggesting the validity of this RPA–based method for QS identification. The successful identification of commodities on the market suggests that the RPA method can be a practical means to combat illegal trade and to strengthen the protection and sustainable utilization of D. acutus (Liu and Ding, 2019).

Conclusions In conclusion, we established an efficient on–spot RPA assay coupled with rapid DNA extraction for QS identification. With a good detection specificity and sensitivity, this assay allows on–spot QS identification within 30 min at room temperature, and is thus a potential powerful means for D. acutus conservation and utilization.

Acknowledgements We are grateful to Dr. Liang Zhang for his help identifying voucher specimens. The work was financially supported by National Natural Science Foundation of

China (Grant number 81573540), Jiansheng Fresh Crude Drug Innovation and Research Foundation (Grant number JSJC-20190103-045, JSJC-20200102054) and Cultivation Planning Project on Science and Technology Development of Southern Medical University (Grant number X2016N004).

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Status of the common myna Acridotheres tristis Linnaeus, 1766 in Turkey E. Per

Per, E., 2022. Status of the common myna Acridotheres tristis Linnaeus,1766 in Turkey. Animal Biodiversity and Conservation, 45.1: 79–83, Doi: https://doi.org/10.32800/abc.2022.45.0079 Abstract Status of the common myna Acridotheres tristis Linnaeus, 1766 in Turkey. The common myna, a species native to central and southern Asia, has been introduced to many countries through the pet trade. The aim of this study was to determine the status of this myna in Turkey. We created a database of sightings of the species in Turkey. The total breeding population of myna birds is estimated to have reached 132–172 pairs. It was reported to have escaped from cages in 11 provinces and populations have become established in urban areas in three provinces. The pet trade has been the main pathway for the introduction of the species. The impact of the myna bird in Turkey remains unclear.

Key words: Bird trade, Cage–escape, Impact, Introduced bird, Non–native bird Resumen Situación del miná común, Acridotheres tristis Linnaeus, 1766 en Turquía. El miná común, una especie nativa de Asia central y meridional, ha sido introducido en muchos países a través del comercio de mascotas. Con este estudio se pretende determinar la situación de esta especie de miná en Turquía. Se ha creado un base de datos sobre la presencia de la especie en Turquía. Se estima que la población reproductora total de miná ha llegado a situarse entre las 132 y las 172 parejas. Asimismo, se ha notificado que algunos ejemplares han escapado de jaulas en 11 provincias y que se han establecido poblaciones en zonas urbanas de tres provincias. El comercio de mascotas es la principal vía de introducción de especies. El impacto de la presencia del miná siguen siendo poco conocidos.

Palabras clave: Comercio de aves, Escape de jaula, Impacto, Ave introducida, Ave exótica Received: 22 III 21; Conditional acceptance: 27 IV 21; Final acceptance: 12 I 22 Esra Per, Department of Biology, Faculty of Science, Gazi University, Ankara, Turkey. E–mail: esraper@gazi.edu.tr ORCID ID: 0000-0002-7764-1215

ISSN: 1578–665 X eISSN: 2014–928 X

© [2022] Copyright belongs to the authors, who license the journal Animal Biodiversity and Conservation to publish the paper under a Creative Commons Attribution 4.0 License.

Per

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Introduction Globalization is making it possible for Invasive Alien Species (IAS) to move into areas where they would not naturally occur. Changes in the environment, such as climate change, can make it easier for IAS to establish populations (Early et al., 2016). IAS are commonly introduced as a result of international trade (Levine and D'Antonio, 2003; Reino et al., 2017) such as pet trade. The Alexandrine parakeet (Psittacula eupatria) and the rose–ringed parakeet (Psittacula krameri) have both established breeding populations in the wild in Turkey, as recorded by citizens’ observations (Per, 2018).

The common myna (Acridotheres tristis) is native to central and southern Asia, including the Indian subcontinent (Hart et al., 2020). The intentional and accidental introduction of this bird into new areas, and the extensions of its range beyond the point of introduction have led to a significant increase in its global distribution. The common myna is one of the 100 worst invasive species in the world (Luque et al., 2014). It is omnivorous and opportunistic (Cramp and Perrins, 1994), thriving in urban areas with a little diversity and abundance of local species (Sol et al., 2012). Its impact has been recorded on many species and in many regions of the world (Evans et al., 2021).

Urbanization and the broad environmental tolerance of this species have likely facilitated its global success as an invasive species (Magory Cohen et al., 2019). For example, after being introduced to the Gulf Region of

the Middle East during the 1990s, the common myna quickly established breeding populations. It is currently distributed across most of the Middle East as a result of deliberate releases or unintentional cage–escapes (Holzapfel et al., 2006). Its distribution has been extending since ever since, and the expansion of its range areas may have allowed it to reach to Turkey and southern Russia (Craig and Feare, 2009).

In Turkey, a breeding population of the myna was first recorded in Ankara in 1996 (Bilgin, 1996). Since 2004, eKuşBank (eBird) has been recording birdwatching records of the myna in Turkey. However, IAS research in Turkey is at an early stage, and no nationwide review of the species’ distribution has been conducted to date. The purpose of this study was to determin current distribution and population size of the myna in Turkey.

Material and methods To establish the distribution of the myna in Turkey, I compiled a database of observations using various sources, scientific articles, books, a thesis, the online international bird watching databases, bird observation–based websites in Turkey, and social media posts. The database includes the observer’s name, observation date, location, province, number of individual birds, and notes. Using the database, the distribution and population of myna in Turkey was investigated.

Table 1. Data resources and records based on the common myna: Province, provinces in which the common myna has been observed; N, number od records observed; % percentage.

Tabla 1. Diferentes recursos y registros de datos sobre el miná común: Province, provincias en las que se ha observado el miná común; N, número de registros observados; % porcentaje.

Data type Literature

Data resource

Province

Article, book, Ankara, İstanbul, Trabzon reports and thesis Database eKuşBank/eBird Ankara, Antalya, Hatay, İstanbul, İzmir,

Mersin, Muğla, Rize, Samsun, Şanlıurfa GBIF Antalya, Hatay, İstanbul, İzmir,

Mersin, Muğla, Rize, Samsun Naturalist Hatay, İstanbul, İzmir Breeding Bird Atlas Ankara, Antalya, İstanbul, İzmir, Samsun Website Birdpx İstanbul, İzmir DogalHayat İstanbul, İzmir Ornito İstanbul, İzmir Trakuş Ankara, Antalya, Bursa, İstanbul, İzmir,

Kayseri, Mersin, Rize, Samsun Social media Facebook İstanbul, İzmir, Rize Instagram İstanbul, İzmir Twitter Çanakkale, İstanbul, İzmir

N

%

5

0.7

365

49.2

50

6.7

5 8 17 19 7 207

0.7 1.1 2.3 2.6 0.9 27.9

18 30 11

2.4 4.0 1.5

Animal Biodiversity and Conservation 45.1 (2022)

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Table 2. The breeding status of the common myna in various provinces: Date, date of first record; Fbd, first breeding date; Lbd, last breeding date (C, continue); Cbp, current breeding pairs; * Gülhane Parkı, İstinye–Yeniköy, Kartal Şehir Hastanesi, Anadolu Adalet Sarayı, Kartal Oto sanayi, Soğanlık Metro İstasyonu, Yıldız Parkı; ** Kültürpark, Bornovapark.

Tabla 2. La situación reproductiva del miná común en varias provincias: Date, fecha del primer registro; Fbd, primera fecha de cria; Lbd, última fecha de cría (C, contínua); Cbp, parejas reproductoras actuales; * Gülhane Parkı, İstinye–Yeniköy, Kartal Şehir Hastanesi, Anadolu Adalet Sarayı, Kartal Oto sanayi, Soğanlık Metro İstasyonu, Yıldız Parkı; ** Kültürpark, Bornovapark.

No

Provinces

Date

Status

1 Ankara 1996 Confirmed breeding 2 Antalya 2015 Cage escape

3 Bursa 2016 Possible breeding 4 Çanakkale 2021 Cage escape 5 Hatay 2018 Cage escape

6 İstanbul 1997 Confirmed breeding 7 İzmir 2005 Breeding 8 Kayseri 2009 Cage escape 9 Mersin 2016 Cage escape 10 Muğla 1994 Cage escape 11 Rize 2019 Cage escape 12 Samsun 2014 Cage escape 13 Şanlıurfa 1996 Cage escape 14 Trabzon 1975 Cage escape

Location

Fbd

Lbd

Cbp

METU Campus Akdeniz Kent Parkı, Atatürk Parkı Oto sanayi City center Ataürk Bulvarı, Sahil Yolu City center (*) City center (**) City center Göksu deltası Dalyan City center City center Birecik vadisi –

1996

2000

–

– 2016 –

– C –

– 2 –

– 2002 2013 – – – – –

– C C – – – – –

– 100–120 30–50 – – – – –

–

–

Total breeding pairs in Turkey

Results and discussion Various websites based on data from Turkey and from bird–watching platforms were evaluated. eKuşBank (49.2 %) and Trakuş (27.9 %) are the most powerful data sources. The contribution of social media data was low and limited (table 1).

The myna was recorded for the first time in Trabzon, Turkey in 1975, and it was first observed as a breeding bird in Ankara in 1996 where it bred at the Middle East Technical University campus until the year 2000 when this population became extinct. In 1997, it was seen in İstanbul and a breeding population was confirmed in the Kartal district in 2002. Breeding populations have since been observed in Gülhane Park, İstanbul Anadolu Adalet Sarayı, İstinye–Yeniköy, Kartal Şehir Hastanesi, Kartal Oto Sanayi, Soğanlık Metro İstasyonu and Yıldız Parkı in İstanbul. Flocks of more than 50 individuals have been observed in various breeding sites in İstanbul, and the breeding population there is estimated to be 100–120 pairs.

The second largest population so far is in İzmir, where four individuals were observed in 2013 but the current estimate is of about 30–50 pairs. The İzmir breeding

– 132–172

population is located in the Kültürpark and in the Bornova Park. There is a population that is thought to be possible breeding population in Bursa. It was observed for the first time in 2016. The number of individuals (four) has not increased but the individuals are seen in the same location throughout the whole year. Common myna have been occasionally reported from 11 of Turkey's 81 provinces, likely having escaped the cage (Ankara, Antalya, Çanakkale, Hatay, Kayseri, Mersin, Muğla, Rize, Samsun, Şanlıurfa, and Trabzon). Myna were mainly observed in urban parks (table 2). Myna populations seem to be rapidly increasing, but only rough estimates of the breeding population size can be provided for the time being as known observations are not based on a systematic breeding search. For the country’s breeding population, I propose therefore a conservative estimate would be 132–172 pairs.

Magory Cohen et al. (2019) used distribution modeling to predict the common myna's potential global distribution and found that the proximity of urbanized areas combined with the species’ broad environmental tolerance favor the range expansion of introduced populations. Common myna readily adapt to a variety

Per

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N

Disappeared population Escaped from captivity Possible breeding population Breeding population Geographic regions Border of Turkey

Marmara Black Sea

Aegean

Central Anatolia Mediterranean

Eastern Anatolia Southeastern Anatolia

0 155 310

620 km

Fig. 1. Distribution map of the common myna in Turkey.

Fig. 1. El mapa de distribución del miná común en Turquía.

of environments, especially in areas where humans are present. If the current established population keeps growing and starts expanding, and/or if cage escapes continue in the coming years, breeding populations can be expected to emerge in the provinces of the Mediterranean Region (Antalya, Hatay, Mersin, and Muğla). It should be noted that with the exception of Ankara, Kayseri and Şanlıurfa, all common myna observations were recorded in provinces with a coastline, and the species has never been seen in the wild in the Eastern Anatolia Region (fig. 1) where extreme climatic conditions prevail. This species' ecological, economic, and social impact across the country currently still is remains unclear.

The common myna is frequently imported as a caged bird by pet shops owners in Turkey (Bilgin, 1996), and current populations most likely stem from escapes from captivity (Kirwan et al., 2008). For example, according to news agencies, 75 myna birds smuggled into Van (eastern Turkey) in 2020 were identified and sent to Gaziantep Zoo (Anadolu Agency, 2020). The supports the assumption that of the origins of myna birds can thus be traced directly to the pet trade in Turkey. Informing the public through the media about the cases of wildlife smuggling and illegal trade is important to promote public awareness of cases and species.

Documenting and monitoring the establishment of new breeding populations of exotic bird species is of much importance. Such information is required to estimate future spread and develop management strategies. Birdwatchers and ornithologists should be aware of the importance of keeping records of observations of this and other introduced species into new locations. Citizen–science data are becoming increasingly important for early detection and monitoring of emerging invasive species, and birdwatchers should be encouraged to collect such data (Holzapfel et al., 2006). Since 2004, birdwatchers and bird photographers have already made valuable contributions to monitoring birds in Turkey. Yet, in terms of monitoring

IAS, much more extensive and rigorously collected data are needed for effective management.

Turkey's decision–maker institutions have been working on a trade ban in order to reduce the numbers of invasive alien species. If their trade is banned and smuggling is addressed efficiently, new cage escapes will likely be reduced in coming years. However, as breeding populations are already established, monitoring of the myna and studies on population dynamics are needed, as well as better understanding of the ecological factors driving its potential range expansion across Turkey. Ongoing urbanization, with the creation of parks and other recreational spaces –especially grassy areas– simultaneously creates habitats suitable for invasive birds such as the common myna. In order to manage the impact of alien species, a detailed prediction of areas at risk of invasion is needed.Such ecological assessments, however, need high–quality data concerning the current distribution of the species.

Acknowledgements I am very grateful to bird watchers. I would like to express my sincere thanks to Tom Evans, Diederik Strubbe, Ortaç Onmuş, Alper Tüydeş, Emre Per, Ufuk Kerman, Sezai Göksu, Ömer Yusuf Salman, Bahar Bilgen, Dilek Geçit and Gökçe Coşkun, for their contribution to this study.

References Anadolu Agency, 2020. Van’da ticareti yasak 75 çiğdeci kuşu ele geçirildi. Anadolu Agency, Turkey, https://www.aa.com.tr/tr/turkiye/vanda-ticareti-yasak-75-cigdeci-kusu-ele-gecirildi/1709187 [Accessed on 21 Jan 2020].

Bilgin, C. C., 1996. First record of the Common Myna (Acridotheres tristis) from Ankara. Turkey.

Zoology in the Middle East, 13(1): 25–26, Doi:

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10.1080/09397140.1996.10637703 Craig, A., Feare, C., 2009. Family Sturnidae (Starlings). In: Handbook of Birds of the World, vol 14 (J. Del Hoyo, A. Elliot, D. A. Christie, Eds.). Lynx Editions, Barcelona.

Cramp, S., Perrins, C. M. (Eds.), 1994. Handbook of the birds of Europe, the Middle East and Africa. The birds of the western Palearctic: crows to finches, vol VIII. Oxford University Press, Oxford.

Early, R., Bradley, B. A., Dukes, J. S., Lawler, J. J., Olden, J. D., Blumenthal, D. M., Gonzalez, P., Grosholz, E. D., Ibanez, I., Mıller, L. P., Sorte, C. J. B., Tatem, A. J., 2016. Global threats from invasive alien species in the twenty–first century and national response capacities. Nature Communications, 7: 12485, Doi: 10.1038/ncomms12485 Evans, T., Jeschke, J. M., Liu, C., Redding, D. W., Şekercioğlu, Ç. H., Blackburn, T. M., 2021. What factors increase the vulnerability of native birds to the impacts of alien birds? Ecography, 44.5: 727–739, Doi: 10.1111/ecog.05000 Hart, L. A., Rogers, A., van Rensburg, B. J., 2020.

Common Myna (Acridotheres tristis Linnaeus, 1766). In: Invasive birds: global trends and impacts: 25–32 (C. T. Downs, L. A. Hart, Eds.). CAB International. Wallingford, United Kingdom.

Holzapfel, C., Levin, N., Hatzofe, O., Kark, S., 2006.

Colonisation of the Middle East by the invasive Common Myna Acridotheres tristis L., with special reference to Israel. Sandgrouse, 28(1): 44–51.

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Kirwan, G. M., Boyla, K. A., Castell, P., Demirci, B., Özen, M., Welch, H., Marlow, T., 2008. The birds of Turkey: a study of the distribution, taxonomy and breeding of Turkish Birds. Christopher Helm, London.

Levine, J. M., D'Antonio, C. M., 2003. Forecasting biological invasions with increasing international trade. Conservation Biology, 17(1): 322–326, Doi: 10.1046/j.1523-1739.2003.02038.x Luque, G. M., Bellard, C., Bertelsmeier, C., Simberloff, D., Courchamp, F., 2014. The 100th of the world’s worst invasive alien species. Biological Invasion, 16: 981–985, Doi: 10.1007/s10530-013-0561-5 Magory Cohen, T., McKinney, M., Kark, S., Dor, R., 2019. Global invasion in progress: modeling the past, current and potential global distribution of the common myna. Biological Invasion, 21: 1295–1309, Doi: 10.1007/s10530-018-1900-3 Per, E., 2018. The status of the parrot trade from tropical forests to Turkey. Turkish Journal of Forestry, 19(3): 275–283.

Reino, L., Figueira, R., Beja, P., Araújo, M. B., Capinha, C., Strubbe, D., 2017. Networks of global bird invasion altered by regional trade ban. Science Advances, 3(11): e1700783, Doi: 10.1126/ sciadv.1700783 Sol, D., Bartomeus, I., Griffin, A. S., 2012. The paradox of invasion in birds: competitive superiority or ecological opportunism? Oecologia, 169: 553–564, Doi: 10.1007/s00442-011-2203-x

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Population increase of the invasive red–whiskered bulbul Pycnonotus jocosus in Valencia, Spain L. Domínguez–Pérez, J. A. Gil–Delgado

Domínguez–Pérez, L., Gil–Delgado, J. A., 2022. Population increase of the invasive red–whiskered bulbul Pycnonotus jocosus in Valencia, Spain. Animal Biodiversity and Conservation, 45.1: 85–95, Doi: https://doi. org/10.32800/abc.2022.45.0085 Abstract Population increase of the invasive red–whiskered bulbul Pycnonotus jocosus in Valencia, Spain. The red– whiskered bulbul Pycnonotus jocosus is a medium–sized passerine that has been classified as an invasive species because of its impact on native ecosystems. It was first reported in the Canary Islands of Spain in 1997. In March 2003, it was sighted in the province of Valencia, in eastern Spain, in a residential area called 'La Cañada'. From 2015 to 2020 we monitored its population in a suburban area close to La Cañada using point counts every spring. Since 2015, the population has shown a trend towards a significant increase in this area, with an estimate of (2,428 < 2,878 < 3,412) individuals in 2020. Its frequency of occurrence has also increased, and it appears to have a continuous distribution in the study area. In the last 17 years the red– whiskered bulbul has spread as far as 20 km from La Cañada, and it is expected to continue spreading and increasing in numbers, with consequences as yet unknown.

Key words: Abundance, Colonisation, Density, Dispersion, Invader Resumen Población en aumento del invasor bulbul orfeo Pycnonotus jocosus en Valencia, España. El bulbul orfeo Pycnonotus jocosus es un paseriforme de talla mediana que se ha clasificado como invasor debido a su impacto en los ecosistemas invadidos. En marzo de 2003, esta especie fue vista por primera vez en una urbanización llamada La Cañada, en la provincia de Valencia, al este de España. Hicimos un seguimiento de su población en una zona suburbana cercana a La Cañada utilizando puntos de conteo realizados todas las primaveras entre 2015 y 2020. Desde 2015, la población del bulbul orfeo ha mostrado una tendencia creciente significativa en el área muestreada y se estima que, en 2020, llegó a los (2.428 < 2.878 < 3.412) individuos.

Además, también aumentó su frecuencia de aparición, y se supone que sigue una distribución continua en el área de muestreo. La especie no se ha quedado restringida y se ha expandido hasta 20 km desde La Cañada en 17 años. Se prevé que la población de bulbul orfeo continuará aumentando y expandiéndose, con consecuencias aún desconocidas.

Palabras clave: Abundancia, Colonización, Densidad, Dispersión, Invasor Received: 9 III 21; Conditional acceptance: 15 VI 21; Final acceptance: 17 I 22 Laura Domínguez–Pérez, José A. Gil–Delgado, ICIBYBE/Dept. of Microbiology and Ecology, Univ. of Valencia, c/ Catedrático José Beltrán 2, 46980 Paterna, Valencia, Spain.

Corresponding author: L. Domínguez–Pérez: lauradop128@gmail.com ORCID ID: L. Domínguez–Pérez: 0000-0002-8038-212X; J. A. Gil–Delgado: 0000-0002-0244-0769

ISSN: 1578–665 X eISSN: 2014–928 X

© [2022] Copyright belongs to the authors, who license the journal Animal Biodiversity and Conservation to publish the paper under a Creative Commons Attribution 4.0 License.

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Introduction The population of invasive species is expected to fit a logistic curve, with a slow population growth in early stages of the invasion, followed by an exponential increase and a slowdown (Blackburn et al., 2009; Davis, 2009; Lockwood et al., 2013). Biological invasions are the result of intentional or unintentional translocation of species beyond their natural distribution range, with numerous well–known examples (Savidge, 1987; Ram and Palazzolo, 2008; da Silva et al., 2010; Standfuss et al., 2016; Linz et al., 2018).

There are generally four main stages in a biological invasion pathway: transport of a species out of its native range, deliberate or accidental introduction (i.e. release or escape from captivity), its successful establishment, and its expansion beyond the original area where it was introduced (Blackburn et al., 2009).

Although most introductions on a global scale took place during a period of major European expansion and settlement (Allen and Lee, 2006; Blackburn et al., 2009), the trends of invaders in Europe indicate an increase in the last two to three decades, with the beginning of globalisation, presumably as a response to the increase in and innovation of human transport over time (Kolar and Lodge, 2001; Hulme, 2009; Abellán et al., 2016). In addition, climate change might facilitate the arrival and establishment of new exotic species, and ease the expansion of those already established (Dullinger et al., 2017; Hulme, 2017; Meyerson et al., 2019). Although exotic species do not always have detrimental effects on the new ecosystem (Stromberg et al., 2009; Gleditsch and Carlo, 2011) they can damage human economy and natural systems in many ways (Pimentel, 2005; Scalera, 2010; Pyšek et al., 2020). Regarding exotic bird species, a high percentage come from temperate regions as the result of introduction during European colonisation, mainly for recreational purposes or as a food resource. In addition, it is estimated that more than half of the introductions occurred on islands, more specifically, on Pacific Islands and in Australasia. Invasive bird species often exhibit high productivity, social behaviour, and behavioural flexibility, which, together with other factors, have enabled their success (Blackburn et al., 2009; Sodhi, 2020).

The red–whiskered bulbul Pycnonotus jocosus is a medium–sized passerine whose natural distribution range includes part of the Indian subcontinent and South East Asia (del Hoyo et al., 2005). It is widespread in captivity around the world because of its popularity as a cage–bird for singing contests or simply as a pet. Its escape or deliberate release has resulted in its establishment on tropical and subtropical islands, and in continental areas of Africa, Asia, Europe, North America and Oceania (del Hoyo et al., 2005; Downs and Hart, 2020). Evidence suggests that climate and ecological similarities between its invaded habitat and its native habitat, together with its intrinsic species characteristics, such as its capacity for local adaptation, its habitat flexibility and its tolerance to human–altered habitats, have contributed to the species' success (Islam and Williams, 2000; Yap

Domínguez–Pérez and Gil–Delgado

and Sodhi, 2004; del Hoyo et al., 2005; Le Gros et al., 2016). In newly colonised territories, it first appears in anthropic habitats, where it is introduced due to release or escape, usually occupying public or private vegetated locations in suburban areas, such as parks and gardens, as it has done in Florida and California (Hardy, 1973; Carleton and Owre, 1975).

Furthermore, it may occupy other human–altered habitats such as fruit orchards. It is also known to venture into areas of native vegetation, and can even establish in these areas. This behaviour has been observed both on islands such as Réunion and Mauritius, and in continents, such as Australia (Mandon–Dalger et al., 1999; Linnebjerg et al., 2009, 2010; Mo, 2015). It seems to prefer lowlands and areas with plentiful resources due to alien plant species (Forys and Allen, 1999; Mandon–Dalger et al., 1999; Clergeau and Mandon–Dalger, 2001; Linnebjerg et al., 2009, 2010). However, it is not limited to such areas and has been found at high altitudes on Réunion (Mandon–Dalger et al., 1999; Clergeau and Mandon–Dalger, 2001) and in native arid habitats of Australia (Mo, 2015).

The red–whiskered bulbul has been classified as invasive because of its impact on its newly colonised ecosystems. Its most important on native habitats is its role as a seed disperser, which can change the composition of vegetation (Corlett, 2017). It can also facilitate the survival and spread of invasive plant species through the 'gut passage effect', which results in an invasional meltdown (Carleton and Owre, 1975; Simberloff and Von Holle, 1999; Mandon–Dalger et al., 2004; Linnebjerg et al., 2009). It is because of this interaction with other exotic species that some authors (e.g. Martin–Albarracin et al., 2015) have classified it as one of the species with the strongest local and global impact. Moreover, the red–whiskered bulbul is considered an agricultural pest for fruit trees in many countries (Carleton and Owre, 1975; van Riper et al., 1979; Mo, 2015). Other threats to native biodiversity are its competition with native and endangered birds species, including other bulbul species (Owre, 1973; Diamond, 1987; Sankaran, 1998; Lever, 2010), predation on native arthropods and reptiles (Diamond, 1987; Lever, 2010), predation on eggs of passerine species (Cheke, 1987; Roberts, 1988; Thibault et al., 2002), and its possible role as a reservoir for malaria (Shehata et al., 2001).

The red–whiskered bulbul was first reported in Spain in 1997 in Tenerife (Canary Islands, Spain) by Abellán et al. (2016), and its breeding was reported in 2001 by Lorenzo (2007). No further information about its progress and current status in Tenerife is available.

Between 2000 and 2021 there have been several sporadic observations of the species in Valencia, Alicante and Granada in Spain, and in Lisbon, Portugal, presumably as a result of independent introductions in the two countries (eBird, 2021; Ascensão et al., 2021). The earliest observation in Valencia was in March 2003, by A. Gil–Delgado and J. S. Monrós, who noted the species in a residential area (La Cañada, Paterna 39º 31' 45.84'' N; 0º 29' 08.26'' W).

Its population seems to have increased herein and

Animal Biodiversity and Conservation 45.1 (2022)

expanded to nearby areas up until 2010, when one of our team (J. A. Gil–Delgado) observed its presence in a residential area of l'Eliana 39º 33' 51.12'' N; 0º 33' 01.94'' W, where we decided to conduct our study. Its introduction was possibly due to its popularity as a cage–bird in Spain, with its subsequent release or escape from captivity. Since its inclusion in the Catalogue of Invasive Alien Species of Spain, the species has been submitted to all the measures listed in the corresponding laws, including a ban on its possession, transportation, trafficking, and trade.

The probability of new introductions into other parts of Spain is therefore low (Cardador et al., 2019). It is not included in the List of Invasive Alien Species of Union concern, but according to Carboneras et al. (2017) its assessment is of mid–priority. Such assessment, however, should have been conducted in 2020 in view of its high impact and the medium level of uncertainty we are facing regarding the outcome of this species in Europe. Despite some information gathered from sightings (Santos, 2016; T. Polo, M. Ferrís, M. Polo, pers. comm., 2021), no single study about the species has been conducted in Spain to check its population status and expansion in the Valencia province. In general, very little it is known about the species in this territory, despite the danger it represents (van Riper et al., 1979; del Hoyo et al., 2005; Lever, 2010).

Using data since 2015, we describe and discuss its population status and expansion process in the past few years. The monitored population has either stabilised or it is increasing, showing changes due to readjustments for carrying capacity.

Material and methods Study area The study area is located in Valencia (39º 28' 11.1'' N; 0º 22' 38.6'' W), a coastal province in eastern Spain.

Due to a demographic increase, municipalities have been transformed by urban growth and residential areas have proliferated. This area has been transformed into a mixture of natural and human–modified patches of Mediterranean vegetation, permanent and annual crops, tree groves, and urban and suburban areas. The red–whiskered bulbul is believed to have spread to suburban areas and artificial vegetated locations near La Cañada (fig. 1). Suburban areas and residential areas are herein used as synonyms, both understood as areas with low population density associated with artificially vegetated areas (e.g. private gardens and urban parks).

The point counts carried were conducted in the area where the red–whiskered bulbul was first observed outside La Cañada in 2010, a residential area of 1,052.25 ha that includes four municipalities of Valencia: l'Eliana, La Pobla de Vallbona, Riba–roja del Túria and San Antonio de Benagéber (fig. 1). This area from here on will be referred to as the sampled area. Other suburban areas outside the sampled area, except for La Cañada, are mentioned as extralimital areas for the purpose of simple reference.

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Density and population trend We determined the density of the red–whiskered bulbul using point counts (Blondel, 1969; García and Purroy, 1973; Sutherland, 2006; Buckland et al., 2015). To calculate density, frequency of occurrence (i.e. number of point counts with the presence of individuals) and population trend, we used a database from 2015–2020. This database consisted of 40 randomly distributed point counts that we repeated each year in spring (April and May). However, given some difficulties during counts (e.g. car traffic or construction site noise), it was not possible every year to perform all the points, but each one has a minimum of 30 and a maximum of 40 points. This means that even if the total number of points counted was 30 one year and 30 again the next, the location of these points could be different from year to year within the initial 40 points (fig. 1). Point counts should be separated by a minimum of 150 m according to Sutherland (2006); but in our case the distance was greater, since our minimum distance between points was 200 m. In addition, we controlled the maximum distance at which we could audibly detect singing individuals. For this environment (i.e. suburban areas) we developed a method to measure our hearing ability.

Hearing ability is defined as the distances at which individuals can be detected by participating observers.

It was measured by one observer (A) who stood next to a singing red–whiskered bulbul individual to ensure that the individual was in fact singing, while the other observer (B) walked to the maximum distance at which they (observer B) could still hear the individual.

Coordinates of the locations of both observers were taken and the distance between them was measured.

Moreover, the hearing ability of both observers was tested in order to rule out bias in the results, since two different observers conducted the point counts.

For each point count, bird abundance was obtained inside a space limited by a 50–metre radius, and was recorded by a single observer for 5 minutes, with a waiting time of at least 1 minute before the count.

Of the birds located within 50 m of the observer, 90 % were seen and heard at the same time, due to the species' habit of perching on high spots to sing.

Singing territorial males were counted as two to better estimate the whole population. Counting and waiting time were selected according to the species’ singing characteristics. Singing time was defined as the time it took a red–whiskered bulbul individual to sing once the observer arrived at the counting point.

It was measured by counting the time it took an individual to start singing at any one of 64 independent randomly selected points in the sampled area. After the observer had localised an individual, they waited at least 1 minute before starting to count. The count was stopped when the individual started singing. Time was recorded as seconds with a chronometer. Once again, these measurements were taken under the same conditions as the point counts to reduce any probability of disturbances, and they were used to set the waiting and counting times for the point counts.

Counts started at 8 a.m. and continued to 11 a.m.

Domínguez–Pérez and Gil–Delgado

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0 50 100 m 711000 E

712000 E

713000 E

0 50 100 m 710000 E

711000 E

712000 E

713000 E

4382000 N 712000 E

713000 E

2019

0 50 100 m

710000 E

2017

711000 E

711000 E

712000 E

713000 E

2017 4382000 N

710000 E

0 50 100 m

4382000 N

4382000 N

2016

730000 E

2018

710000 E

4382000 N

4383000 N 4381000 N

713000 E

725000 E

4383000 N

712000 E

4383000 N

711000 E

720000 E

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4382000 N 0 50 100 m

715000 E

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2015

710000 E

4381000 N

710000 E

4381000 N

705055 E

4383000 N

700000 E

4360000 N 4365000 N 4370000 N 4375000 N 4380000 N 4385000 N

10 km

4383000 N

5

605000 E

4381000 N

B

0

4381000 N

4360000 N 4365000 N 4370000 N 4375000 N 4380000 N 4385000 N

A

0 50 100 m

710000 E

711000 E

712000 E

713000 E

Fig. 1. Location of the range expansion of the red–whiskered bulbul in Spain: A, area in the Valencia Community, eastern Spain, where the red–whiskered bulbul is found. The sampled area (yellow), La Cañada (green), and extralimital locations where presence of the species has been confirmed by observations of the authors (black dots) or from the literature (Santos, 2016; T. Polo, M. Ferrís, M. Polo, pers. comm., 2021) (white dots). B, the 40 points used in the density analysis (2015 and 2020) in the sampled area; the points with presence each year (presence, red dots; absence, white dots).

Fig. 1. Localización del rango de expansión del bulbul orfeo en España. A, territorio en la Comunidad Valenciana, al este de España, donde el bulbul orfeo está presente. Se indican el área muestreada (en amarillo), La Cañada (en verde) y las zonas fuera del área muestreada donde la presencia de la especie se confirmó mediante observaciones de los autores de este estudio (puntos negros) o solo por la bibliografía (Santos, 2016; T. Polo, M. Ferrís, M. Polo, pers. comm., 2021) (puntos blancos). B, los 40 puntos utilizados para el análisis de la densidad (2015 y 2020) en el área de muestreo, con los puntos con presencia para cada año (presencia, puntos rojos; ausencia, puntos blancos).

Animal Biodiversity and Conservation 45.1 (2022)

89

25 20 15 10

360–390

330–360

300–330

270–300

Time (sec)

240–270

210–240

180–210

150–180

120–150

90–120

60–90

0

30–60

5

0–30

Number of observations

30

Fig. 2. Results of the time (sec) of each sample taken in the study area regarding the time each individual took to start singing, with the number of observations per category (R2 = 0.697, F1,11 = 11.53, p = 0.0025).

Fig. 2. Resultados del tiempo (en segundos) de cada muestra realizada en el área de estudio que tardan los individuos de esta especie en comenzar a cantar, con respecto al número de observaciones de cada categoría (R2 = 0,697; F1,11 = 11,53; p = 0,0025).

to increase the detection probability (Sotthibandhu, 2003). Counts were avoided on rainy and windy days, and were conducted at weekends or during holidays to reduce the probability of disturbances that could interfere with acoustic detection. Although the species was not actively monitored in the extralimital areas, its presence has been confirmed by independent observations since 2015.

Statistical analysis

<

<

The frequency of occurrence was tested using a binomial regression (glm) function (Hastle et al., 1992; Venables and Ripley, 2002). The population trend was analysed using a linear model (lm) function (Wilkinson and Rogers, 1973; Chambers, 1992), with a polynomial regression model, applying the (lm(y~x + I(x2))) function as the goodness–of–fit measure for the time at which individuals started singing. Hearing ability was tested using a two–sample t–test. All the previous statistical analyses were performed with RStudio 1.3.1093.

The density and detection probability of the red– whiskered bulbul was estimated using models based on distance sampling (Buckland et al., 2015) with the Distance 7.3 software (Thomas et al., 2010). Detection probability was modelled from a detection function (g(r)), which represents the probability of detecting

an object at a distance (x) from the sampling point (Buckland et al., 1993, 2015; Buckland, 2006). These functions are formed by a key function and its fitting terms (Buckland et al., 1993; Buckland, 2006; Thomas et al., 2010). In this case, the key functions tested were half–normal (hn) and hazard–rate (hr), finally choosing hn because of its lower AIC (indicated in the annexes). Thus we estimated density according to these premises: (i) the individuals at distance 0 were certainly detected, (ii) individuals were detected at their initial point; (iii) the distance between individuals and the observer was accurate between the two groups, considered to be 0–50 m and 50–100 m (Buckland et al., 1993; Buckland, 2006); (iv) detections were statistically independent (Groom et al., 2007). The result of its population density is given as A < N < B ind/ha, where A represents the minimum value adopted by the population, B the maximum value and N the best estimate. A transformation from density (ind/ha) to total individuals in the sampled area was applied for ease of interpretation based on Buckland (2006) and their distance sampling method, where individuals were calculated using each density applied to the whole sampling area (1,052.25 ha). We understood population growth rate as the intrinsic rate of increase (r) (see Margalef, 1974), where Nt is the final population, N0, the initial one, and t represents time: Nt = N0 ert.

Domínguez–Pérez and Gil–Delgado

90

90

40

35

35

80 31

Presence (%)

70 60

30

39

50 40 30 20 10 0

2015

2016

2017 2018 Year

2019 2020

Fig. 3. The point counts with presence of red–whiskered bulbul showed a significantly increasing tendency, i.e. frequency of occurrence (presence %), each year. The exact number of point counts conducted annually is shown at the top of each bar (p = 0.0001).

Fig. 3. Tendencia significativamente creciente del número de puntos de conteo con bulbul orfeo, es decir, frecuencia de aparición (presence %) de cada año desde 2015 en el área muestreada, con el número exacto de puntos de conteo realizados cada año en la parte superior de cada barra (p = 0,0001).

Results Hearing ability Thirty–six distances (half by each observer) were measured around the sampled area. The two–sample t test showed no difference between the means of each observer (O1= 89.4 m, SE = 1.98; O2 = 90.2 m, SE = 2.21; t = 0.28099; p = 0.7804). The maximum distance at which an individual could be heard by the observers was 105 m.

Time to start singing The number of individuals that started singing decreased significantly over time (R2 = 0.697; F1,11 = 11.53; p = 0.0025). Therefore, it took a short time for individuals to be detected, with 42 % being detected within 0–30 seconds. Furthermore, 16 % individuals began singing at 0 seconds. This means that individuals were already present during the waiting time or before sampling commenced (fig. 2). Moreover, only 10 % of individuals took more than 5 minutes to start singing.

The mean waiting time was 97 seconds, which validated a 5–minute sampling time (SE = 14.3; N = 64).

Density and population trend Since the first observation of the red–whiskered

bulbul in the sampled area in 2010, its frequency of occurrence has increased significantly (R2 = 0.8512; F1,4 = 29.6; p = 0.005). In 2015, just over 50 % of points were occupied, but this rose to 85 % in 2019 and remained the same in 2020. Therefore in five years, its frequency in the sampled area has increased by 35 % (fig. 3).

The red–whiskered bulbul population density in the sampled area also showed a marked increase. The highest recorded density was in 2020 (2.31 < 2.74 < 3.24 indiv/ha), 0.93 indiv/ha more than the density recorded in 2015 (table 1). Furthermore, its trend showed a significant increase (R2 = 0.8411; F1,4= 21.18; p = 0.01) (fig. 4A). We estimated that the population counted in 2015 with (1,581 < 1,900 < 2,282) individuals in the sampled area had reached (2,428 < 2,878 < 3,412) individuals in 2020. We estimated a total rate of growth of 15 %, a mean growth of 4.15 %, and a variance of ± 0.76 % in the six years we monitored the population in the sampled area (fig. 4B) Since 2015, independent first–hand observations have been made in different years, seasons and locations in extralimital areas within a 24–km radius of La Cañada. Due to variability in conditions, they were not considered in our study to calculate both population density and trend. However, this has allowed us to confirm the presence of this species in some extralimital areas (fig. 1).

Animal Biodiversity and Conservation 45.1 (2022)

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Discussion Table 1. Calculations of numbers of individuals in the sampled area (1,052.25 ha) and the corresponding density.

Many bulbul species are very vocal and their singing is sufficiently distinctive to be differentiated from sound–similar species (Lloyd et al., 1996; Woxvold et al., 2009; Kamtaeja et al., 2012). Vocalisations vary according to their purpose, and are more abundant during the breeding season, the period when we carried out our sampling (Islam and Williams, 2000; del Hoyo et al., 2005; Kamtaeja et al., 2012). Moreover, species of the genus Pycnonotus look for prominent perches to sing on (del Hoyo et al., 2005). They are therefore conspicuous and easy to detect, and the red–whiskered bulbul is no exception. In studies about this species as an invader, its presence is usually identified by sightings or by their song (Carleton and Owre, 1975; Mandon–Dalger et al., 1999; Mo, 2015).

Their song is quite particular and cannot be confused with that of other species. When their presence abounded, our results showed that this species was very easily seen or heard, meaning its presence and abundance can be easily controlled.

An increase in presence frequency and population density suggests that the population has increased as both these factors have grown significantly, not only regarding numbers of individuals but also regarding the extent of the area occupied over the years in the study area. Taking into account its presence in the nearby extralimital areas we can assume that its distribution will continue to extend in our study area.

2015 2016 2017 2018 2019 2020 Year

Year

Calculations

2015 2016 2017 2018 2019 2020

A

N

B

Density

1.50

1.81

2.17

Individuals

1,581

1,900

2,282

Density

1.91

2.29

2.75

Individuals

2,008

2,409

2,890

Density

1.52

1.87

2.30

Individuals

1,598

1,965

2,416

Density

1.87

2.20

2.58

Individuals

1,968

2,312

2,715

Density

1.94

2.30

2.73

Individuals

2,038

2,420

2,874

Density

2.31

2.74

3.24

Individuals

2,428

2,878

3,412

B 10

Increase (%)

Density (indiv/ha)

A 3.30 3.20 3.10 3.00 2.90 2.80 2.70 2.60 2.50 2.40 2.30 2.20 2.10 2.00 1.90 1.80 1.70 1.60 1.50

Tabla 1. Resultados de los cálculos para determinar la cantidad de individuos en el área muestreada (1.052,25 ha) y la densidad correspondiente.

5

0

–5

–10 2015

2016

2017 2018 Year

2019 2020

Fig. 4. A, trend in the population density of red–whiskered bulbul (2015–2020) in the study area using Distance Sampling (R2 = 0.8411; F1,4 = 21.18; p = 0.01). Values (ind/ha ± SE). B, annual growth rate (%) of the population density in the sampled area (2015–2020) (R2 = 0.5354, F1,4 = 6.762, p = 0.06).

Fig. 4. A, tendencia de la densidad poblacional del bulbul orfeo (2015–2020) en el área muestreada usando Distance Sampling (R2 = 0,8411; F1,4 = 21,18; p = 0,01). Valor (ind/ha ± EE). B, tasa de crecimiento anual (%) de la densidad poblacional del bulbul orfeo en el área muestreada (2015–2020) (R2 = 0,5354; F1,4 = 6,762; p = 0,06).

92

One aspect supporting our finding that the bulbul currently occupies the entire sampled area is that the number of point counts where it has been observed has reached 85 % in recent years, a figure similar to the number of individuals detected in 5 minutes.

Even when the species is present in the whole area, the result is not likely to be 100 % because some individuals may sing after 5 min, which was the unit of time used to calculate density.

Throughout the six years of this study the red–whiskered bulbul population not only increased in number but doubled in number, matching the expected development of an invasive population. As the percentage of increase from 2015 to 2018 was more pronounced than that in 2019–2020, the population may be approaching carrying capacity (i.e. the k–value) (see Margalef, 1974). This rapid increase is not an obstacle for the population to adjust to a stable pattern, a fluctuating pattern, or to collapse and disappearance (Simberloff and Gibbons, 2004; Simberloff and Rejmánek, 2011; Blackburn et al., 2009). However, between the fluctuating or stable alternatives, the first option seems more probable due to the species invasive nature.

The fluctuation can be partly explained by migration towards extralimital areas (Begon et al., 2006; Blackburn et al., 2009). Because the population has been monitored for a relatively short period, we must wait a few more years to see how it progresses, but based on its development in other invaded areas (Islam and Williams, 2000; Clergeau and Mandon–Dalger, 2001; Pranty, 2010; Mo, 2015), we suspect that this species, far from disappearing, will continue to extend.

In view of the current prohibitions of alien species in Spain, however, it is unlikely that new introductions will reinforce the established population in Valencia.

In 2017, we observed a decrease in the number of individuals. Although this was not significant, it was abrupt. The number of samples and presence frequency of the species were similar to those recorded in 2016. Other bulbul species show seasonal fluctuations due to local migration or other causes (Monadjem, 2002; Yamaguchi, 2005; Nakamura, 2007), and even invasive bulbuls display this behaviour in invaded areas (Brooks, 2013). In these locations, maximum density peaks are usually reached in spring–summer, when the number of individuals is larger due to seasonal migration or their behaviour fosters detection probability (Nakamura, 2007; Brooks, 2013). Due to the conditions of our sampling, the observed variation cannot be due to these seasonal changes. Therefore, we suspect that it may have been the result of some particularly adverse condition in 2017, or from a drop following exceptionally good conditions in 2016, which we are unaware of.

Our results are consistent with those obtained in other studies of invasive red–whiskered bulbul populations. Carleton and Owre (1975) estimated that the population in Florida in 1969 (nine years after its introduction) would have been about 250 individuals spread over 8.3 km2, with an approximate average annual population rate increment of 33–40 %. Pranty (2010) pointed out that the red–whiskered bulbul currently occupies an area of 41.7 km2 in Florida.

Domínguez–Pérez and Gil–Delgado

Although no exact or recent data are available, some authors indicate that its population could have exceeded 700 individuals in the 1980s (Rand, 1980; Pranty, 2010). The rising number of individuals in our study seems similar to that of the species in Florida, with similar population values, but with a smaller sampled area and lower average annual percentage in growth rate. In Australia, Wood (1995) calculated an average density of 0.4 indiv/ha of the red–whiskered bulbul in 37.8 ha in 1985. Considering that the red–whiskered bulbul first arrived in the area where this count was conducted in the 1950s (Mo, 2015), the average annual number of individuals reached 15 individuals in 26–35 years. This figure is lower than ours, although the area was larger. However, even with only 15 individuals, the species was able to continue spreading southwards until the 1990s (Mo, 2015), so its invasiveness should not be underestimated.

One study that calculated dispersion rates of red– whiskered bulbul (Clergeau and Mandon–Dalger, 2001) noticed spread was faster on islands than in continents, as supported by their observation that the rates after 3 years on Mauritius, Réunion and Oahu were higher than those after 10 years in Florida and Australia. Here we confirm the presence of the red–whiskered bulbul in several municipalities outside La Cañada. In 17 years it has expanded from its origin of release to residential areas within a 20–km radius. Our results can therefore be compared to the dispersal rates observed in Florida and Australia. Its different expansion on different continents could be attributed to species diversity, established interaction networks, predation, competition with other invasive species, and geographical and other ecological characteristics such as habitat suitability (van Riper et al., 1979; Williams and Giddings, 1984; Mandon–Dalger et al., 1999; Islam and Williams, 2000; Clergeau and Mandon–Dalger, 2001; Mo, 2015). It remains to be determined how these aspects may affect the species in the province of Valencia, Spain.

The expansion and colonisation of the red–whiskered bulbul in Valencia has been ignored for too long.

Human intervention is most effective in early stages of invasion when the number of individuals is still small.

We therefore need to know how much time there is before we reach the point of no return regarding this invasion (Clergeau and Mandon–Dalger, 2001; Blackburn et al., 2009; Davis, 2009; Lockwood et al., 2013). Eradication programmes in California have been unsuccessful as individuals continue to exist in the area (Owre, 1973; Islam and Williams, 2000; Lever, 2010). However, the programmes conducted on the Seychelles Islands by the Seychelles Island Foundation (SIF) in 2013–2014 seem to have worked as there has been no trace of the species since 2015 (Rimbault et al., 2017). It is of note that the characteristics of these two regions differ (eradication on an island versus eradication on a continent), and this could have influenced the different outcomes. Until more data become available concerning the ecology and impact of the red–whiskered bulbul the extent of the threat of the invasive species remains unknown.

Animal Biodiversity and Conservation 45.1 (2022)

Acknowledgements We thank E. Bernat–Ponce for his critical review and comments on the manuscript. We also thank A. Taberner and A. Bermudez for their help in field work, and T. Polo, M. Ferrís and M. Polo (Societat Valenciana d'Ornitologia) for providing us with a recent citation of the species in Picassent.

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Bold or shy? Examining the risk–taking behavior and neophobia of invasive and non–invasive house sparrows J. Quesada, C. A. Chávez–Zichinelli, M. García–Arroyo, P. J. Yeh, R. Guevara, J. Izquierdo–Palma, I. MacGregor–Fors Quesada, J., Chávez–Zichinelli, C. A., García–Arroyo, M., Yeh, P. J., Guevara, R., Izquierdo–Palma, J., MacGregor– Fors, I., 2022. Bold or shy? Examining the risk–taking behavior and neophobia of invasive and non–invasive house sparrows. Animal Biodiversity and Conservation, 45.1: 97–106, Doi: https://doi.org/10.32800/abc.2022.45.0097 Abstract Bold or shy? Examining the risk–taking behavior and neophobia of invasive and non–invasive house sparrows.

Behavior provides a useful framework for understanding specialization, with animal personality aiding our understanding of the invasiveness of birds. Invasions imply dispersion into unknown areas and could require changes in behavior or spatial clustering based on personality. Reduced neophobia and increased exploring behavior could allow individuals to colonize new areas as they test and use non–familiar resources. Here, we hypothesized that house sparrow (Passer domesticus) individuals from invasive populations would exhibit bolder behavior than in non–invasive populations. We assessed risk taking and neophobia in male house sparrows in Barcelona (where it is considered native) and in Mexico City (where it has become widely invasive), captured in two different habitats, urban and non–urban. We assessed latency to enter an experimental cage and to explore it, and latency to feed and feeding time in the presence of a novel object. We found that sparrows from Mexico City, both from urban and non–urban areas, were quicker to enter the experimental cage than the sparrows from Barcelona. The time it took the birds to start exploring the cage gave a similar result. We found no differences between cities or habitats in the latency to feed and feeding time while exposed to a novel object.

Our results partially support the view that the invader populations from Mexico City are bolder than those from Barcelona. Behavior is an important component of plasticity and its variability may have an important effect on adaptation to local situations. Future studies should disentangle the underlying mechanisms that explain the different personalities found in populations of different regions, contrasting populations of different densities, and taking different food availability scenarios into account.

Key words: Experimental, Exotic, Native, Passer domesticus, Personality, Urban ecology Resumen ¿Audaz o tímido? Examinando el comportamiento de toma de riesgo y neofobia de los gorriones comunes invasivos y no invasivos. Dado que el comportamiento proporciona un marco útil para comprender la especialización, la personalidad animal puede ayudar a explicar la capacidad invasiva de las aves. La invasión implica la dispersión por áreas desconocidas y podría requerir cambios en el comportamiento o agrupaciones espaciales basadas en la personalidad. La reducción de la neofobia y el aumento del comportamiento de exploración podrían permitir a los individuos colonizar nuevas áreas a medida que prueban y utilizan recursos que no les son familiares. En este trabajo suponemos que los individuos de gorrión común (Passer domesticus) mostrarán un comportamiento más audaz en las poblaciones invasivas que en las poblaciones no invasivas. En este estudio evaluamos la toma de riesgo y la neofobia en machos de gorrión común de Barcelona (donde se considera nativo) y de Ciudad de México (donde es invasivo) capturados en dos hábitats diferentes (urbano y no urbano). Evaluamos la latencia para entrar en la jaula experimental y para explorarla, así como la latencia para alimentarse y el tiempo de alimentación en presencia de un objeto extraño. Encontramos que los gorriones de Ciudad de México, tanto de hábitats urbanos como no urbanos, entraron más rápido en la jaula experimental que los gorriones de Barcelona. El resultado fue similar para el tiempo que les tomó comenzar a explorar la jaula. No encontramos diferencias entre ciudades y hábitats en cuanto a la latencia para alimentarse y el tiempo

ISSN: 1578–665 X eISSN: 2014–928 X

© [2022] Copyright belongs to the authors, who license the journal Animal Biodiversity and Conservation to publish the paper under a Creative Commons Attribution 4.0 License.

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que les tomó alimentarse en presencia de un objeto extraño. Nuestros resultados apoyan parcialmente la idea de que las poblaciones invasivas de Ciudad de México son más atrevidas que las de Barcelona. El comportamiento es un componente importante de la plasticidad y su variabilidad puede tener un efecto importante en la adaptación a situaciones locales. Se deberían llevar a cabo otros estudios para desentrañar los mecanismos que explican las diferencias de personalidad que se encuentran entre poblaciones de distintos orígenes, así como comparando poblaciones con diferente densidad demográfica y teniendo en cuenta diferentes contextos de disponibilidad de alimentos.

Palabras clave: Experimental, Exótico, Nativo, Passer domesticus, Personalidad, Ecología urbana Received: 18 VIII 21; Conditional acceptance: 1 X 21; Final acceptance: 11 II 22 Javier Quesada, Jaume Izquierdo–Palma, Departament de Vertebrats, Museu de Ciències Naturals de Barcelona, Barcelona, Catalonia, Spain.– Carlos A. Chávez–Zichinelli, El Colegio de Puebla, Puebla, Mexico.– Michelle García–Arroyo, Ian MacGregor–Fors, Ecosystems and Environment Research Programme, Faculty of Biological and Environmental Sciences, University of Helsinki, Niemenkatu 73, FI–15140, Lahti, Finland.– Pamela J. Yeh, Department of Ecology and Evolutionary Biology, University of California, Los Angeles, CA, USA; Santa Fe Institute, Santa Fe, NM, USA.– Roger Guevara, Red de Biología Evolutiva, Instituto de Ecología A.C., Xalapa, Veracruz, Mexico.– Jaume Izquierdo–Palma, Laboratorio de Ecología, UBIPRO, Facultad de Estudios Superiores de Iztacala, Universidad Nacional Autónoma de México, Tlanepantla de Baz, Estado de México, Mexico.

Corresponding author: Ian MacGregor–Fors. E–mail: ian.macgregor@helsinki.fi ORCID ID: J. Quesada: 0000-0002-6010-8473; M. García–Arroyo: 0000-0002-9167-4777; P. J. Yeh: 0000-0002-7264-9509; R. Guevara: 0000-0003-0768-3580; J. Izquierdo–Palma: 0000-0003-2420-3825; I. MacGregor–Fors: 0000-0003-3198-7322

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Introduction The success of exotic bird species may be explained by several traits (Candolin and Wong, 2012; Weis and Sol, 2016). Classical studies have focused on population dynamics while other important aspects such as behavior have received less attention (Sol and Maspons, 2016). Recently, behavior has been identified as a key element to explain successful invasion of new areas (Evans, 2010; González–Lagos and Quesada, 2017). Studies have highlighted the role of behavior in understanding specialization in species' use of movement and space with niche specialization (Spiegel et al., 2017; Schirmer et al., 2019). Because invasion of new environments implies dispersion into unknown areas and thus implies unknown dangers, one could expect that individuals with certain behavioral traits are more prone to successfully invade new areas (Chapple et al., 2012; Wolf and Weissing, 2012; Myles–Gonzalez et al., 2015). Not surprisingly, it has been shown that these behavioral traits can be subjected to selective pressures (Canestrelli et al., 2016).

Animals have systematic and structured within– population differences regarding their behavioral tendencies. Such tendencies have been shown to be stable over time when exposed to the same situation or context, a phenomenon that has been called personality (Wolf and Weissing, 2012). Personality traits may be identified and quantified through antagonistic behaviors that are observable among the individuals in a group or population, always under the same scenario (Réale et al., 2007). The personality traits most widely studied in the past are boldness–shyness, exploration, activity, aggressiveness, and sociability (Réale et al., 2007).

When an invasion process starts, individuals need a set of capabilities, including boldness, to explore new environments or use novel resources (Chapple et al., 2012; Weis and Sol, 2016). This topic has been studied in fish and reptile species, but evidence in bird species is still scarce (Dingemanse et al., 2007; Herczeg et al., 2009; Myles–Gonzalez et al., 2015; Lapiedra et al., 2017). Specifically, how personality might contribute to species invasiveness is a relatively unexplored area of research in birds. Here we hypothesized that within the same species (house sparrow Passer domesticus), individuals from populations in non–native areas where they are highly invasive would exhibit bolder behavior than individuals from a region where they are considered native. Since different personality traits are often correlated (Reàle et al., 2007), individuals from non–native and invasive populations could also present bolder, more exploratory and less neophobic behaviors.

The house sparrow is a good biological model to test this hypothesis as it represents a well–known example of a successful invasive species worldwide (Anderson, 2006). Specifically, the North American population was introduced into northeastern USA during the mid–18th century as the result of several independent events.

Afterwards, it invaded most of the USA and Mexico, presumably at the beginning of the 20th century (Wagner, 1959; Robbins, 1973; Anderson, 2006, Peña–Peniche et al., 2021). Unlike most bird species, house sparrows tend to increase their densities in urban areas when

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conditions are favorable, yet their body condition, which has shown to be stable across urban land–uses, can be compromised in heavily–industrialized sites, for instance (Chávez–Zichinelli et al., 2010; Bókony et al., 2012b; Bonier, 2012). This sparrow has shown high plasticity in an extensive set of behaviors. These include the use of a wide variety of nest cavities and nest substrates (Kimball, 1997; Nhlane, 2000; Peach et al., 2008; Hoi et al., 2011), an extensive array of foraging behaviors (Guillory and Deshotels, 1981; Kötél, 1984; Flux and Thompson, 1986; Anderson, 2006), and adaption to an omnivorous diet in urban environments, ranging from seeds to nectar, insects, and even discarded human–food leftovers (Stidolph, 1974; Gavett and Wakeley, 1986; Clergeau, 1990; Moulton and Ferris, 1991; Leveau, 2008). The species is not migrant (with the exception of a plesiomorphic population; i.e., Bactrianus; Sætre et al., 2012), and it has a small range in established populations (~5 km; Anderson, 2006). Its invasibility is thus not explained by population–based processes like migration and could perhaps be the result of an individual decision–making process based on personality. Thus, its behavioral flexibility (low levels of neophobia and high levels of exploratory behavior) could allow individuals to colonize new areas through their ability to recognize and use of non–familiar resources (Webster and Lefebvre 2001; Sol et al., 2002; Wright et al., 2010).

The house sparrow is a human commensal species. It has been hypothesized to have evolved along with human agriculture (Anderson, 2006), with only the Bactrianus population lacking association with human settlements (Sætre et al., 2012). Its relationship is so closely tied to humans across most of its distribution that local populations tend to become extinct when human settlements are abandoned (Anderson, 2006; Summers–Smith, 2010). Furthermore, this sparrow has learned to use humans as cues to find sites with large amounts of food resources in the form of human food waste (Fernández–Juricic, 2001), although this has been recorded only at intermediate levels of human presence (Fernández–Juricic et al., 2003).

Here, we studied risk taking and neophobia, assessed through boldness, in male house sparrows in two cities; one where the species is non–invasive and considered to be native (Barcelona, Spain), and another where it is invasive (Mexico City, Mexico). We acknowledge that the sparrow populations in Barcelona may have been the result of range–expansion together with that of the human species because the house sparrow originated in the Middle East (Anderson, 2006). We thus consider the Barcelona populations as well–settled (> 5,500 years; Ravinet et al., 2018) that arrived as a human commensal species through indirect human assistance (Anderson, 2006) in comparison with the more recently introduced (< 100 years) and settled populations in Mexico City (Peña–Peniche et al., 2021).

For this study, we focused on house sparrows from two habitats, urban and non–urban (mainly agricultural), from each city. We performed two experiments where we evaluated potential differences in (i) risk–taking (latency to enter the experimental

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Risk taking experiment Once the dark box was open we evaluated if the animal came out or not from the dark box, how much time it took each individual to go out (latency), and the time it took to explore the experimental cage.

Cage of metal mesh and two perches.

Neophobia experiment

60 cm

90

cm

Small dark box made of cardboard annexed to the experimental cage.

60 cm Food deprived for 2 hrs.

Food in familiar feeder.

Food deprived for 2 hrs.

Food in familiar feeder and a red ball close to the feeder.

Fig. 1. Experimental design of risk–taking and neophobia.

Fig. 1. Diseño experimental de toma de riesgos y neofobia.

cage and latency to explore the experimental cage) and (ii) neophobia (latency to feed and feeding time in the presence of a novel object). Given that previous studies have shown that house sparrows can be bolder when exotic and invasive (Martin and Fitzgerald, 2005; MacGregor–Fors et al., 2010, 2019), we predicted that sparrows from Mexico City would be bolder that those from Barcelona in both experiments.

Furthermore, given that urban birds have shown to be bolder than their non–urban relatives (Liker and Bókony, 2009; Bókony et al., 2012a; Riyahi et al., 2017), we expected urban individuals to be bolder than their non–urban counterparts, regardless of their native or non–native status.

Material and methods Study area and fieldwork We performed this study in two cities: Barcelona (Spain) and Mexico City (Mexico). Barcelona, located in the northeastern region of the Iberian Peninsula (41º 23' 30'' N, 2º 10' 25'' E), representing the second most populated urban center in Spain (1.6 million inhabitants), with its metropolitan area consisting of 3.2 million residents (AMB, 2021). Mexico City, located in the Valley of Mexico (19º 25' 56'' N, 99º 7' 59'' W), is the most populated urban center in Mexico (22 million residents; INEGI, 2020), and one of the most populated worldwide (United Nations, 2018).

We captured male adult house sparrows from September through December of 2014. We used mist nets that were open from dawn (6:00 h) to noon (12:00 h). In Mexico City, we captured 16 male house sparrows in the urban area (i.e., Ciudad Universitaria, UNAM) and 16 at the non–urban site (i.e., Milpa Alta).

In Barcelona, we captured 20 male house sparrows in the urban area (Parc de la Ciutadella) and 22 at the non–urban site (Parc Agrari del Baix Llobregat, Gavà). We released all sparrows at the sites of capture after the trials were completed.

Experimental trials Birds were housed in individual cages for at least 10 days in order to reduce the stress of captivity (Quesada et al., 2013). They were fed ad libitum with a mixture made for granivorous and insectivorous birds, complemented with vitamins (Moreno–Rueda and Soler, 2002; Bókony et al., 2012b). During the first three days we supplied an anti–parasite solution in the water so that a minimized condition was not a constraint that could mask the results of the experiment (Quesada et al., 2013).

We performed two personality tests: one of risk taking and one of neophobia. To evaluate risk–taking behavior, we put the sparrows individually inside a small dark box made of cardboard that was annexed to the experimental cage (made of metal mesh with two perches). After the observer was stationed behind the blinds, the door of the dark boxes was opened

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Table 1. Generalized linear models assessing the proportion of individuals, the time that male house sparrows from Mexico City and Barcelona took to come out of the dark box and to enter the experimental box, and the time to start exploring the experimental cage.

Tabla 1. Modelos lineales generalizados que muestran la proporción de individuos, el tiempo que tardaron los gorriones machos de Ciudad de México y de Barcelona en salir de la caja oscura y entrar en la jaula experimental y el tiempo que tardaron en comenzar a explorar la jaula experimental.

Deviance Residual Df Residuals Df Deviance

P

Number of individuals that entered the experimental cage City

1

2.98

72

43.35

0.08

City / habitat

2

1.43

70

41.91

0.48

City

1

18.57

65

18.95

< 0.001

City / habitat

2

0.38

63

18.56

0.50

City

1

0.39

64

4.02

0.01

City / habitat

2

0.03

62

4.00

0.79

Latency to enter the experimental cage

Latency to start exploring the cage

from a distance by means of a string system. We then evaluated whether the animal came out or not (as several animals did not leave the box) and how long each individual took to go out (latency to enter to the cage). Once the bird came out, we measured the time it took for the individual to begin exploring the experimental cage (latency to explore the cage).

For the neophobia experiment we examined the individual behavioral response to an unknown (novel) object. After the first experiment (described above), the birds that came out from the experimental box were maintained in the experimental cage for two hours and food–deprived in order to assure that they were motivated when food was presented in the cage (Quesada et al., 2013). After the deprivation period, we introduced a feeder for 30 min, similar to the one from which they had been fed in the captivity cages so that they could quickly recognize that there was food. We then deprived the sparrows of food again for two hours and repeated the same protocol, but in this case, we introduced a red ball (novel object) close to the feeder.

We recorded whether or not the birds approached the feeder, and measured the time it took them to approach and the length of time they fed (fig. 1).

Statistical analyses For the latency to enter the experimental cage, we performed two analyses. First, as some individuals did not enter the experimental cage in the 120 min of the trial, we assessed the number of birds that came out of the box using a generalized linear model (GLM; binomial distribution), considering a nested scenario of city and habitat as independent variables (i.e., city/ habitat). Second, for those individuals that did come out of the box, we performed two additional GLMs

(Gamma distribution, given the distribution of the dependent variables), one to assess the time they took to leave the box (latency to enter to the cage) and another to relate the time they took to explore the cage (i.e., latency to explore the cage).

For the latency to feed and feeding time in the presence of a novel object, we used a generalized linear mixed model (GLMM; Gamma distribution) for time to feed and a linear mixed model (LMM) for feeding time given the distribution of the data. For this test, using only the birds that came out of the box, we considered the nested scenario of city and habitat as independent variables (i.e., city/habitat), the experiment variables (foreign object, habitat) as a fixed factor, and the identity of the individual as random factor given that we used the same individual twice (with and without the novel object).

Results The proportion of male sparrows that came out of the box to explore the experimental cages was not related to habitat, but there was a non–significant trend for differences between cities (table 1). A higher proportion of house sparrows from Mexico City left the dark box (n = 31; 96.7%) than those from Barcelona (n = 35; 85.4 %). For those individuals that left the dark box, sparrows from Mexico City showed faster times than those from Barcelona, showing no interaction with habitat (Mxurb: 7.20 ± SE 0.97 secs; Mxagr: 30.06 ± SE 4.53 secs; Bcnurb: 967. 65 ± 55.45 secs; Bcnagr: 862.53 ± 41.75 secs, table 1). A similar pattern occurred for the time individuals took to start exploring the experimental cage (Mxurb: 32.93 ± SE 2.25 secs; Mxagr: 53.38 ± SE 5.32 secs; Bcnurb: 79.82 ± 6.18 secs;

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Table 2. Mixed models assessing the latency to feed and feeding time in the presence of a novel object (NO) of male house sparrows that came out of the box.

Tabla 2. Modelos mixtos que muestran la relación entre la latencia para alimentarse y el tiempo que tardaron los gorriones macho que salieron de la caja en alimentarse en presencia de un objeto extraño (NO).

x2

Df

P

Time to feed Intercept

118.23

1 < 0.001

NO

0.02

1

0.88

City

0.04

1

0.83

NO × City

0.05

1

0.82

NO × City / habitat

1.35

4

0.85

Feeding time Intercept

172.45

1 < 0.001

NO

1.22

1

0.26

City

0.07

1

0.79

NO × City

0.61

1

0.44

NO × City / habitat

4.10

4

0.39

Bcnagr: 74.47 ± 3.10 secs, table 1). For the neophobia trials, we found no differences in city or habitat regarding the time to feed and feeding time, or for the presence–absence of the novel object (table 2).

Discussion Invasiveness has traditionally been explained by life history traits, where some species have thrived in urban environments because they have a set of preexistent behavioral, morphological, or physiological traits that are the consequence of their evolutionary history (Partecke, 2014). However, behavior can also be a key factor related to the invasiveness of some species that a priori do not fit as candidates to thrive in urban habitats (González–Lagos and Quesada, 2017). Indeed, behavior is an important component of plasticity, whether as a consequence of genetic expression or plasticity–based learning processes (Snell–Rood, 2013), and it can have an important impact on population dynamics (Pelletier and Garant, 2012). This variability may have a notable effect on generating new strategies to thrive in a new environment as a product of phenotypic variance, beyond what would be expected by natural selection alone. However, this variance in behavior can be associated with behavioral syndromes (i.e., the way in which the personality traits are combined)

that result in different personalities (Drent et al., 2003; Sih et al., 2004).

In this study we tested the hypothesis that personality is associated with the invasiveness of the house sparrow from Mexico City when compared to those from Barcelona by analyzing two basic personality traits (risk taking and neophobia) in a bold–to–shy spectrum (Canestrelli et al., 2016). Our results partially support the view that invasive populations from Mexico City are bolder than those from Barcelona. The populations from Mexico City, urban and non–urban, took more risk in exploring new areas than those from Barcelona. However, none of the studied populations showed differences regarding fear to new objects. This is not surprising given the plasticity of a species that is so well adapted to urban and agricultural scenarios where they tend to be exposed to novel objects on a regular basis.

Personality may play an important role in spatial ecology (Spiegel et al., 2017) given that in certain ecological contexts a selective regime may favor some particular personalities (Myles–Gonzalez et al., 2015).

When the invasive process starts, the individuals need a set of capabilities for exploration and boldness to facilitate resource use, to cope with disturbances, and to enhance communication in new and unknown environments. Bolder personalities are thus candidates to thrive in areas outside their original range of distribution, given that explorers and bolder individuals tend to disperse farther (Canestrelli et al., 2016). Our results suggest that this variation in personality may be used to adapt to local situations. The house sparrows from the invasive population (Mexico) appeared to be bolder than the Barcelona population, at least in terms of risk taking (i.e, the time it took them to leave the experimental box and explore the cage).

Interestingly, dispersion ability, a key factor for invasiveness, is mediated by decision–taking processes that imply assuming the risk to explore (or not) new habitats, given that exploration implies the assumption of some cost in term of fitness (Chaine and Clobert, 2012; Gonzalez–Lagos and Quesada, 2017).

Our results agree with several experimental studies that have been carried out in similar approaches of risk–taking. For instance, several studies have shown how birds of the same species express different risk– taking behaviors (i.e., flight initiation distance) when they invade a new environment (Scales et al., 2011; Tryjanowski et al., 2016; Ducatez et al., 2017). One common example is the comparison of non–typical urban species invading urban areas. In most cases, urban populations take more risks than populations from the non–urban areas. This finding applies to the house sparrows; as Seress et al. (2011) showed.

They observed that young urban house sparrows were bolder than non–urban birds of any age group in Hungary, where the species is considered to be native, although older urban birds were less bold.

In addition, studies have demonstrated that more exploratory birds disperse larger distances than less exploratory birds according to theoretical (Spiegel et al., 2017) and empirical approaches (Dingemanse et al., 2003; Korsten et al., 2013; Botero–Delgadillo

Animal Biodiversity and Conservation 45.1 (2022)

et al., 2020; Jablonszky et al., 2020), supporting the hypothesis that personality may play a significant role in the success of bird invasions.

MacGregor–Fors et al. (2019) recently showed that house sparrows are bolder –in terms of alert distances– where it is exotic, invasive, and abundant, suggesting a density–dependent process in urban areas. Given that house sparrow densities in Mexico City are greater than those in Barcelona (MacGregor–Fors et al., 2017), our results could also be influenced by a density–dependent process. Thus, broadening the number of populations from a wider spectrum of the sparrow’s distribution could help to confirm whether our result is generalizable or not.

We recorded no differences for the neophobia experiment. Many studies have shown that birds adapted to new situations (i.e., urban birds) are less neophobic than those that remain in known conditions (Møller, 2008). A study carried out in in house sparrows in Hungary in different intensities of urbanization did not find significant differences between personality traits (e.g., neophobia, predatory risk–taking, level of activity) or behavioral syndromes (Bókony et al., 2012a).

In contrast, Cohen and Dor (2018) found that the southernmost range–expanding population of house sparrows in Israel had the fewest neophobic individuals, although this could be because the authors compared two different geographic populations (i.e., Biblicus, Indicus). Martin and Fitzgerald (2005) also compared neophobia in two invasive populations, one in Panama and the other in New Jersey (USA), at different stages of establishment. They found that the newer population (in Panama) approached and consumed novel food resources at faster rates. Altogether, these results suggest that differences in neophobia between invasive populations are detected when invasion is in its initial stages, but not when invasion has reached a temporal threshold as observed in the population of house sparrows in Mexico.

Regarding the selective mechanisms behind changes in the personality of invasive and non–invasive populations of one same species, we consider that a crucial question is: are these differences between populations a consequence of the selection factors of individuals with particular personalities (bold or shy personalities) or is this a consequence of learning processes, such as habituation or behavioral syndrome change? The time since colonization of both populations could shed some light on this question. Sparrows from Barcelona presumably range–expanded to the European Mediterranean region during the expansion of human agriculture, while they were intentionally translocated to North America (Anderson, 2006). Hence, the populations in Barcelona, generally considered native as they have been part of the avifauna of the region for millennia, appear to be the consequence of bold individuals, or groups that range–expanded to the European Mediterranean Basin as human commensals. The case of recent introductions in North America represents random samples of individuals (in terms of behavioral personality) that arrived to North America and underwent further selection processes to become bolder phenotypes, and then expanded

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their distribution down to Panama, becoming established in central Mexico less than 100 years ago (Peña–Peniche et al., 2021). Although differences in the selection pressure that occurred in both cases could explain differences in current behaviors, we consider that time since settling as viable populations is a key factor to explain this difference. The time of settlement has been important to explain differences in other personality traits related to boldness (Martin and Fitzgerald, 2005; but see Cohen and Dor, 2018).

Regardless of the mechanism behind the differences in risk–taking in invasive and non–invasive populations, our results reinforce recent studies that have highlighted the role of personality in understanding specialization in movement and use of space in avian species with niche specialization (Spiegel et al., 2017; Schirmer et al., 2019) and potential invasions (Sol and Maspons, 2016).

Another question yet to be answered is the role of group size in the interactions of populations (Liker and Bókony, 2009; Ducatez et al., 2017; MacGregor–Fors et al., 2019). In our experiments, we tested individuals separately; yet large groups may cope more effectively with unfamiliar situations through faster innovations of new solutions by some group members with favorable traits (Bókony et al., 2012a). This supposition agrees with a recent paper that emphasized the role of house sparrow density to explain the risk–taking of flight initiation distance (MacGregor–Fors et al., 2019).

Thus we consider that future studies should explore the mechanisms underlying the different personalities found in invasive and non–invasive populations by contrasting populations of different densities, considering a wider spectrum of populations across the species' distribution, and taking into account different food availability scenarios.

Acknowledgements We are deeply thankful to Lizeth Zavala for her help with field and lab work in Mexico City, as well as Gregorio Moreno–Rueda and an anonymous reviewer who made useful comments on an earlier version of the manuscript. Research funds were granted by UC MEXUS–CONACYT to P. J. Yeh and I. MacGregor–Fors (CN-13-587), National Geographic Society to P. J. Yeh and I. MacGregor–Fors, the Natural History Museum of Barcelona (PASSERCAT–2 project), Fundació Zoo de Barcelona and by the CGL2016-79568-C3-3-P and CGL2020 PID 2020-114907GB-C21 project from the Spanish Ministry of Economy, Industry, and Competitiveness to J. Quesada.

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Complete mitogenomes reveal limited genetic variability in the garden dormouse Eliomys quercinus of the Iberian Peninsula G. Forcina*, M. Camacho–Sanchez*, A. Cornellas, J. A. Leonard Forcina, G., Camacho–Sanchez, M., Cornellas, A., Leonard, J. A., 2022. Complete mitogenomes reveal limited genetic variability in the garden dormouse Eliomys quercinus of the Iberian Peninsula. Animal Biodiversity and Conservation, 45.1: 107–122, Doi: https://doi.org/10.32800/abc.2022.45.0107 Abstract Complete mitogenomes reveal limited genetic variability in the garden dormouse Eliomys quercinus of the Iberian Peninsula. The garden dormouse Eliomys quercinus is a poorly known Western Palearctic species experiencing a global decline. Even though the availability of genetic information is key to assess the drivers underlying demographic changes in wild populations and plan adequate management, data on E. quercinus are still scant. In this study, we reconstructed the complete mitogenomes of four E. quercinus individuals from southern Spain using in–solution enriched libraries, and found evidence of limited genetic variability. We then compared their cytochrome b sequences to those of conspecifics from other countries and supported the divergent but genetically depauperate position of this evolutionarily significant unit (ESU). The information produced will assist future conservation studies on this little–studied rodent.

Key words: Mammalia, Rodentia, Gliridae, Spain, Evolutionary significant unit, Rodent Resumen Los mitogenomas completos desvelan una limitada variabilidad genética del lirón careto Eliomys quercinus en la península ibérica. El lirón careto Eliomys quercinus es una especie paleártica occidental poco conocida cuyas poblaciones están experimentando un descenso a escala mundial. Aunque la información genética sea fundamental para determinar las causas de los cambios demográficos que se producen en las poblaciones salvajes y planificar adecuadamente su gestión, los datos sobre E. quercinus siguen siendo escasos. En este estudio reconstruimos el mitogenoma completo de cuatro individuos de E. quercinus del sur de España a través de librerías enriquecidas en solución y encontramos una variabilidad genética limitada. También comparamos las secuencias del citocromo b con las de conespecíficos de otros países y pudimos confirmar la divergencia, así como la baja diversidad genética de esta unidad evolutivamente significativa (ESU) en cuestión. Esta información será de ayuda para los futuros estudios de conservación de este roedor escasamente estudiado.

Palabras clave: Mammalia, Rodentia, Gliridae, España, Unidad evolutivamente significativa, Roedor Received: 16 VII 21; Conditional acceptance: 25 XI 21; Final acceptance: 22 II 22 Giovanni Forcina, Miguel Camacho–Sanchez, Anna Cornellas, Jennifer A. Leonard, Conservation and Evolutionary Genetics Group, Estación Biológica de Doñana (EBD–CSIC), 41092 Seville, Spain.– Giovanni Forcina, CIBIO/ InBIO, Centro de Investigação em Biodiversidade e Recursos Genéticos, Universidade do Porto, Campus Agrario de Vairão, 4485–661 Vairão, Portugal; BIOPOLIS Program in Genomics, Biodiversity and Land Planning, CIBIO, Campus Agrario de Vairão, 4485-661 Vairão, Portugal.

Corresponding author: G. Forcina, GloCEE, Global Change Ecology and Evolution Research Group, Departamento de Ciencias de la Vida, Universidad de Alcalá, 28805 Alcalà de Henares, Spain. E–mail: giovanni.forcina@cibio.up.pt * These authors contributed equally ORCID ID: G. Forcina: 0000-0001-5727-7770; M. Camacho–Sanchez: 0000-0002-6385-7963; J. A. Leonard: 0000-0003-0291-7819 ISSN: 1578–665 X eISSN: 2014–928 X

© [2022] Copyright belongs to the authors, who license the journal Animal Biodiversity and Conservation to publish the paper under a Creative Commons Attribution 4.0 License.

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Introduction Small mammals are a taxonomically diverse group that includes species of major ecological importance in terrestrial ecosystems. A variety of life history traits and their responsiveness to environmental variation make them ideal bioindicators (Talmage and Walton, 1991). Changes in morphology and genetic diversity in small mammal populations over time reflect alterations in their environment (Hadly et al., 1998; Larsen and Matoq, 2019; Forcina and Leonard, 2020; Parker et al., 2020), making them valuable for understanding ongoing local and global changes. For these reasons, attention has recently been devoted to developing a suite of increasingly informative molecular markers for monitoring small mammal communities (Maroja et al., 2003; Meyer et al., 2006; Loiseau et al., 2007; Kanitz et al., 2009; Barbosa et al., 2013; Forcina et al., 2021).

While some species like the 'lab rat' Rattus norvegicus and the black rat R. rattus have been the object of intense study (Aplin et al., 2011; Song et al., 2014) and benefit from well–developed genomic resources, many others, even in the same genus, are still scarcely known (Robins et al., 2010; Camacho–Sanchez et al., 2017, 2018; Liu et al., 2018; Thomson et al., 2018; Camacho–Sanchez and Leonard, 2020). This applies also to the house mouse Mus musculus, with some of its congeners largely understudied (Tong and Hoekstra, 2012). In the case of non–model species, the generation of basic genetic tools to be applied in a variety of systems would allow comparison and facilitate repeatability of the studies being, thus enabling important methodological improvements.

As a case in point, the garden dormouse (Eliomys quercinus Linnaeus, 1766) is a secretive representative of the family Gliridae (Rodentia) patchily distributed across Europe from the Urals to the Iberian Peninsula as well as on major western Mediterranean islands (Storch, 1978; Filippucci, 1999; Bertolino et al., 2008; fig. 1A). Other European members of this family are the forest dormouse (Dryomys nitedula), the edible dormouse (Glis glis), the hazel dormouse (Muscardinus avellanarius), and Roach's mouse–tailed dormouse (Myomimus roachi), with pairs of species often living in sympatry (Juškaitis and Šiožinytė, 2008). Even though a global demographic assessment has not been made, the garden dormouse is considered to be declining sharply in parts of its range over the last three decades (Bertolino, 2017), resulting in its inclusion among the species protected by Appendix III of the Bern Convention (The European Community, decision 82/72/EEC, Official Journal L of 10.02.1982), Annex IV, and 'Near Threatened' status by the International Union for the Conservation of Nature (IUCN, 2011).

The most severe declines have been those observed in eastern Europe (Bertolino, 2017), but the local disappearance of this species, which occurs in a range of forested habitats, scrublands and cultivated land (Storch, 1978; Bertolino et al., 2008), has also been recorded in Italy (Amori, 1993; Sarà, 2008) and Spain (Ruiz and Roman, 1999), where its reduction occurred even within protected areas (Moreno, 2002; Santoro et al., 2017). The progressive reduction of suitable

Forcina et al.

habitats was suggested as the leading likely cause for its decline in central Europe, where vast portions of forested areas have been cleared and replaced by bush–dominated vegetation (Anděra, 1986, 1994), a less suitable habitat for dormice as a whole (Fedyń et al., 2021). However, the same cause cannot be invoked for western Europe, where the extent of forested areas is increasing (Martín–Forés et al., 2020). A long–term study carried out in Doñana National Park (DNP), a UNESCO Heritage Site in Andalucía (southern Spain), at the southernmost edge of the species range, registered a dramatic drop of the garden dormouse over almost four decades. Once referred to as an abundant species in the local small mammal community inhabiting the core of the protected area (Camacho and Moreno, 1989), Doñana Biological Reserve (DBR), it is now virtually locally absent (Santoro et al., 2017). The shift in habits by local predators triggered by hemorrhagic diseases decimating the local European rabbit (Oryctolagus cuniculus) population and climate change are still being debated as plausible causes underlying the decline of the garden dormouse in southern Spain.

The increased predation rate of the garden dormouse – the second largest of the small mammals in the DBR community after lagomorphs– by means of a top–down control (Palomares et al., 1995) is a possibility. While this scenario looks plausible when considering the persistence of hemorrhagic diseases across Iberia (Abrantes et al., 2013) and a suite of new infections affecting the Iberian hare (Lepus granatensis) (Lopes et al., 2014; García–Bocanegra et al., 2019, 2021), the widespread use of pesticides along with changes in agroecosystem structure and function (Kuipers et al., 2012) –pressures that are also present in southern Spain (Santoro et al., 2017)– are generally invoked to explain the decreasing trend of garden dormouse observed at the global level (Bertolino, 2017). The spread of monoculture agriculture, in particular, has been suggested to be associated with a reduction of trophic resources causing their displacement by Rattus spp. (Cristaldi and Canipari, 1976; Macdonald and Barrett, 1993). DBR and DNP are largely bordered by agricultural land, but Rattus populations have also been observed to decline in the last decades in the reserve (Santoro et al., 2017), reducing the probability that they are an important factor.

Like the related edible dormouse, which is still consumed in Europe (Peršič, 1998; Werner, 2007), the garden dormouse was considered a pest by orchard owners and was eaten. This practice has a long history, as documented in the archeological record, such as the Bronze Age site of Cerro de la Encina (province of Granada) (Friesch, 1987). An extended network of typical warrens (gliraria) and numerous jars (dolia) used to keep and fatten dormice (Carpaneto and Cristaldi, 1995) were found in the Roman settlement (1st century B.C.E.) of Arucci (province of Huelva) (Bermejo et al., 2015). The Romans' great appreciation of dormice meat would explain the occurrence of garden dormice in archeological sites scattered across the Roman Empire, even where the species is not native like Great Britain (O'Connor, 1986), perhaps resulting in naturalized populations on some islands.

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Over the last two decades, scientific attention to this species has been spurred by its stunningly high karyotypic diversity, with the number of chromosomes ranging from 2N ¼ 48 to 2N ¼ 54 (Perez et al., 2013 and references therein). Likewise, garden dormice display remarkable range–wide morphological variation (Filippucci et al., 1988; Kryštufek and Kraft, 1997) which is unrelated to chromosomal variation, raising doubts about the taxonomic identity of several populations (Cristaldi and Canipari, 1976; Filippucci et al., 1988) and even leading to the proposal of new species within the genus (Miller, 1912). A specific investigation has been carried out to explore possible congruence between chromosomal information and population–level cytochrome b (cyt b) diversity (Perez et al., 2013). The study unveiled the existence of four distinct evolutionary significant units (ESUs: Moritz, 1994; Fraser and Bernatchez, 2001) –referred to as Iberian, Italian (including Corsica and Dalmatia), Western European (excluding Iberia) and Alpine– some of which host multiple chromosomal races and their hybrids. These data could suggest the persistence of gene flow between chromosomal races in spite of their different karyotype, in a similar way to what has been observed in the common shrew, Sorex araneus (Narain and Fredga, 1996).

The timing of divergence between the garden dormouse and its Middle Eastern relative, the Asian or large–eared dormouse (E. melanurus), and garden dormhouse intraspecific divergence was estimated, respectively, at 7.0 ± 0.9 mya (Montgelard et al., 2003) and 4.2 ± 1 mya (Perez et al., 2013), well before Quaternary glaciations. Interestingly, Perez et al.

(2013) found that the garden dormouse did not exhibit the typical pattern of northward postglacial recolonization from southern European refugia as observed in several other species (Hewitt, 1996, 1999; Weiss and Ferrand, 2007; Gentili et al., 2015; but see also Deffontaine et al., 2005; Kotlik et al., 2006; Feliner, 2011; Queirós et al., 2019), namely a decreasing trend of genetic diversity from the Iberian, Italian and Balkan peninsulas to the rest of the continent. However, the few genetic sequence based studies on the garden dormouse rely on a single mitochondrial DNA (mtDNA) locus, cyt b, and genetic data available for Spain are very limited, with only two genotyped individuals from the Balearics (Perez et al., 2013) and the Pyrenees (Barbosa et al., 2013). This knowledge gap is remarkable, especially considering previous evidence for multiple refugia within the Iberian Peninsula (i.e., Portugal: Perez et al., 2013) and the likely origin of the species there (Perez et al., 2013, Mansino et al., 2015), suggesting that it is a key region to unravel the evolutionary history of the species.

In order to provide new, useful molecular resources for this and related species, here we report complete mitochondrial genomes for four garden dormice from Andalucía (southern Spain, fig. 1). We also extracted and compared their cyt b sequences to those publicly available from conspecifics from across most of the species' range. The genetic information produced in this study paves the way for further studies addressing the conservation genetics of this declining rodent.

109

Material and methods Sample collection and library preparation We retrieved tissue (muscle and liver) from four garden dormouse specimens preserved at the biological collection of Doñana Biological Station (EBD–CSIC) in Seville, Spain. The specimens were collected in 2003–2013 from three sites in Andalucía (tables 1, 1s in supplementary material). DNA extraction, Illumina shotgun library preparation and sequencing were performed as in Forcina et al. (2021).

Mitogenome assembly and phylogenetic inferences Adapter trimming and minimal quality filtering on the 3'–end were performed with cutadapt 2.10 (Martin, 2011). A reference mitogenome was assembled de– novo with NOVOPlasty3.7 (Dierckxsens et al., 2017) by pooling the four samples and using the GenBank sequence of E. quercinus cyt b GQ453669 as a seed.

Sequences from each individual were mapped to this reference mitogenome using BWA mem 0.7.12–r1039 (Li, 2013). SAMtools 1.3 (Li et al., 2009), was used to discard reads with quality mapping below 40 and remove PCR duplicates. Mitogenomes were annotated in MITOS webserver (Bernt et al., 2013). The annotations were manually edited after comparing them in Geneious 8.1.5 (http://www.geneious.com: Kearse et al., 2012) with the NCBI staff–curated mitochondrial genomes of two closely related species: M. avellanarius NC_050264 and G. glis NC_001892. We used mafft v7.453 (Katoh et al., 2013) to produce an alignment including the newly sequenced individuals, E. quercinus GenBank reference MN935777, the above mentioned mitogenome GenBank references of other European dormice and GenBank reference HE978360 belonging to Kellen’s dormouse (Graphiurus kelleni) (Fabre et al., 2013). Codon positions 1 and 2 from protein coding genes were retained. Third codon positions were removed as they are mostly saturated at the subfamily level (Breinholt and Kawahara, 2013; Hinckley et al., 2021). The best partition scheme was determined using PartitionFinder 2.1.1 (Lanfear et al., 2012), and grouped into 6 partitions. Gblocks (Castresana, 2000) was used to remove positions in the alignment with little homology or gaps. The final alignment consisted of 6,432 positions. We produced a maximum likelihood (ML) tree in RAxML 8.0.0 (Stamatakis, 2014). The tree was rooted with the edible dormouse based on a mammal phylogeny recently built with genome–wide data (Upham et al., 2019). AMAS (Borowiec, 2016) was used to handle the multiple sequence alignments.

Indices of diversity and biogeographic reconstructions All garden dormouse cyt b sequences available in GenBank were retrieved with the rentrez package (Winter, 2017) in R 4.0 (R Core Team, 2020). They were aligned together with assembled cyt b sequences from the four new individuals using mafft. The resulting multiple sequence alignment was trimmed using trimAl v1.4.rev15 (Capella–Gutiérrez et al., 2009).

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We computed the number of polymorphic sites (S), number of haplotypes (h), haplotype diversity (Θ), and nucleotide diversity (π) of the four mitogenomes with DnaSP 5.10.1 (Librado and Rozas, 2009). A nexus cyt b alignment (partial sequence: 573 bp; nt. positions 386–959) including the newly generated sequences (n = 4) and all the E. quercinus references available in GenBank (n = 50: table 1s in supplementary material) was produced in Geneious and imported in DnaSP for computing the same indexes as above on a country basis as well as for each of the four ESUs identified

by Perez et al. (2013), assigning sequences retrieved in this and other studies on the basis of their genetic affiliation in the median–joining haplotype network (Bandelt et al., 1999) built by importing the cyt b alignment in PopART 1.7 (Leigh and Bryant, 2015).

DNA sequences from complete mitochondrial genomes are deposited in GenBank with accession numbers MZ130252–MZ130255, associated with the BioProject PRJNA727082. The data and code for data analysis are available at https://github.com/csmiguel/ eliomys and permanent repository in Zenodo (Doi: 10.5281/zenodo.6337428).

Results Mitogenome structure and genetic diversity The new garden dormouse mitogenomes (GenBank accession number MZ130252–MZ1302525) were 16,617 bp in length (10 bp shorter than GenBank reference MN935777 from Germany, the only E. quercinus entire mitogenome available at the time of manuscript writing), with a base composition of 33.6 % A, 28.8 % T, 24.3 % C, and 13.3 % G. They comprised 13 protein–coding genes, 2 ribosomal RNA (rRNA) genes, 22 transfer RNA (tRNA) genes, and the control region (D–loop). Coordinates for each feature are reported in table 2. We found three different mitogenome haplotypes, each one private to one sampling site (tables 1, 1s in supplementary material), but characterized by a limited variability (table 3) exemplified by the identification of only 19 segregating sites (S) in Spain, as opposed to 767 when they were compared with the homologous sequence from Germany (MN935777).

The phylogenetic reconstructions of European garden dormice (fig. 3) confirmed –despite a moderate bootstrap support – the same topology inferred on the basis of single mtDNA and nuclear loci, with Eliomys and Muscardinus being sister taxa (Montgelard et al., 2003) and Graphiurus 'in turn' sister to them, and with Glis being the most basal taxon, as confirmed by genome–wide data (Upham et al., 2019). Levels of genetic diversity within countries and ESUs were variable (table 3).

Biogeographic reconstructions Four distinct, previously described haplogroups emerged in the network (fig. 1B). One haplogroup consisted of individuals from the Iberian Peninsula and southern France, including the new haplotype

identified in this study. A second haplogroup included samples from the French and Italian Alps, and a third was distributed in central and western Europe (excluding Iberia), namely central and northern France, Belgium, Germany and Austria. The last haplogroup included individuals from the insular populations of Corsica, Sardinia and Sicily along with two continental representatives from Abruzzo (central Italy) and Croatia. Finally, individuals from Umbria (central Italy), just a few hundred kilometers north along the Italian Peninsula compared to the latter, were scattered in the middle of the four haplogroups.

Discussion Small mammals are key components of many terrestrial ecosystems. However, some species –even widespread species– are still under–investigated. In this study, we contribute to the scanty genetic resources available for the widespread garden dormouse, E. quercinus, by generating four entire mitochondrial genomes and examining the biogeographic affiliation of three populations from a region for which no mitochondrial sequence data have been available to date.

A look at the sampling effort of this and previous studies (Perez et al., 2013; Barbosa et al., 2013) in a spatial perspective (fig. 1A) shows that our knowledge of eastern populations of this species – which are highly fragmented and scattered over a large and remote area– is incomplete. Paradoxically, this gap might inhibit the understanding of the reason(s) underlying the fast–paced decline of the garden dormouse in this portion of its range, and should be tackled with the prompt adoption of a range–wide monitoring program (Bertolino, 2017).

Around the Mediterranean region, whole mitochondrial genome sequences will enable elucidation of the origin of the insular populations inhabiting Sardinia, Sicily, and the neighboring Lipari (where no records have been reported for the last three decades: Sarà, 2008). Both morphology and partial cyt b sequences have been used to evaluate the relationship of insular populations to mainland populations. Some studies have proposed each island being granted a distinct subspecific status, while others have suggested a human–mediated introduction (Vigne, 1992; Masseti, 2005; Ientile and Massa 2008; Angelici et al., 2009).

If garden dormice did arrive in the Balearics with the help of humans, based on fossil evidence, this introduction would have to date back to at least 2300–2200 B.C.E. (Traveset et al., 2009). This introduction may be supported by the genetic data, since the same cyt b haplotype was found in individuals from both Formentera (Balearics) and Andorra (fig. 1). Dating recent events using only cyt b can be very misleading, so the availability of whole mitogenome data is necessary to provide the resolution needed to reconstruct the routes and to date both recent human–mediated introductions and ancient natural dispersals. Mitochondrial genomes are particularly useful for molecular clock analyses due to the absence of recombination, the high mutation rate, and large numbers of previous studies providing

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A

B 26 23 24 39 38 25

16 14

15

17 11

2

1

7

37

9

20 21 19 32 13 12 18 33 34 10

3

35 36

40

22 5

30

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HE613982

34

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HE613983 18 HE611090

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29

18

HE611091

4,5 6 3

HE611092

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2

HE613988

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1

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HE613977 2

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1

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1

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HE614006 16 1 HE614003 HE613997–HE613998 24

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1

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HE614002 37

1

25

1

14,15 39 1

1 30

GQ453669

1

Iberian Italian Alpine Western European

26 HE614004

HE614005–MN935777

HE614007 17

Fig. 1. A, sampling localities used in this study (see table 1s in supplementary material for bibliographic references); B, distribution range (dark shading) of the garden dormouse E. quercinus based on Bertolino (2017); C, median–joining network based on partial mitochondrial cytochrome b sequences (573 bp): black numbers embedded in circles (haplotypes) refer to sampling localities of figure 1A; red numbers refer to mutational steps. The diameter of each haplotype is proportional to its frequency in the dataset.

Fig. 1. A, localidades de muestreo de este estudio (véase la tabla 1s en material suplementario para consultar las referencias bibliográficas); B, área de distribución (gris oscuro) del lirón careto E. quercinus según Bertolino (2017); C, red de haplotipos utilizando el método de unión por la mediana (Median–Joining) basada en secuencias parciales del citocromo b (573 pares de bases): los números en negro dentro de los círculos (esto es, los haplotipos) se refieren a las localidades de muestreo mostradas en la figura 1A; los números en rojo indican los saltos mutacionales. El diámetro de cada haplotipo es proporcional a su frecuencia en la base de datos.

Forcina et al.

nE

tr n V( ua c) 1

s–rRN A

op D–lo 16

I– rR N A

g) ug ) P( gu tr n T(u tr n kb 15

+

TB CY

tr

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2

(u

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)

3

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a)

ua

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tr n

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trnI(gau) trnQ(uug) trnM(cau) 4

ND5

ND2 12 ) ag ) L(ugcu ) n r ( t nS ug tr H(g tr n

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6

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uc

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tr

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1

X O

C

2

uu)

(u tr nK ATP8

COX

8 ATP6

a) ug c) S( tr n D(gu n 7 tr

MZ130252

Fig. 2. Graphical representation of garden dormouse E. quercinus mitogenome. Protein–coding genes, ribosomal RNA (rRNA) genes and transfer RNA (tRNA) genes are indicated in blue, orange and red, respectively, along the light (external part of the grey circle) and heavy (internal part of the circle) strand of the mitogenome. The blue vertical bars inside the circle are indicative of the sequence depth across the entire assembly.

Fig. 2. Representación gráfica del mitogenoma de lirón careto, E. quercinus. Los genes que codifican las proteínas y los que codifican los ARN ribosomales (ARNr) y de transferencia (ARNt) están indicados en azul, naranja y rojo, respectivamente, a lo largo de la cadena ligera (parte exterior del círculo gris) y la pesada (parte exterior del círculo gris). Las barras azules dentro del círculo indican la cobertura a lo largo de todo el ensamblaje genómico.

reference data and rates (Gilbert et al., 2008; Singh et al., 2009; Duminil and Besnard, 2021; Yu et al., 2021).

The characterization of mitochondrial genomes can also be useful to look at relationships between lineages within species (McGowen et al., 2009; Morin et al., 2010; Fabre et al., 2017; He et al., 2021), where single mitochondrial genes do not have sufficient power to resolve them (Galewski et al., 2005; Zardoya and Meyer, 1996; Sasaki et al., 2005; Viricel and Rosel, 2011; He et al., 2015). This is even more important when recent and rapid speciation events are not paralleled by the accumulation of morphological disparity, as happened

with the genus Rattus (Rowe et al., 2011). In the case of the garden dormice, the availability of complete mitogenomes will lay the ground for testing hypotheses regarding the species history and will contribute to resolving the contentious relationships between populations from the Maghreb and the eastern Mediterranean, now recognized as belonging to E. munbyanus (Holden, 2005) and E. melanurus (Montgelard et al., 2003), respectively. Moreover, complete mitogenomes might assist in pinpointing the populations of highest conservation value, as has occurred in other taxa for which the use of single mitochondrial genes had failed

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Table 1. List of garden dormice E. quercinus specimens used in this study: SV, specimen voucher; GenBank, GenBank accession number.

Tabla 1. Listado de ejemplares de lirón careto E. quercinus usados en este estudio: SV, código de la muestra; GenBank, número de acceso en GenBank.

SV

Collection site

EBD 32997M

Jerez de la Frontera

Collection date

Lat. / Long.

GenBank

21/11/2013

N 36.7; W –6.1

MZ130252

(Zoological Garden) EBD 32795M

Jerez de la Frontera

28/03/2011

N 36.6, W –6.2

MZ130253

EBD 32745M

Córdoba

01/01/2003

N 38.0, W –4.8

MZ130254

EBD 29942M

Montilla (Córdoba)

03/05/2012

N 37.6, W –4.6

MZ130255

Table 2. Features (total length in bp, start and end codon: left to right) of mitochondrial protein–coding genes and the D–loop in E. quercinus as compared with other European dormice. No mitogenome is currently available for D. nitedula or M. roachi. Animal photos are not to scale, for credits see the Acknowledgements.

Tabla 2. Características (longitud total en pares de bases, codón de inicio y codón de terminación: desde la izquierda hacia la derecha) de los genes codificantes de proteínas en la mitocondria y de la región control en E. quercinus en comparación con otros lirones europeos. Actualmente, no hay ningún genoma mitocondrial secuenciado de D. nitedula ni de M. roachi. Las fotos de los animales no están a escala, se pueden ver los agradecimientos en el apartado "Acknowledgements".

Eliomys quercinus MZ130252

Glis glis NC_001892

Muscardinus avellanarius NC_050264

Dryomys nitedula

Myomimus roachi

ND1

957, ATG, TAA

955, ATG, T– –

956, ATG, TA –

–

–

ND2

1,042, ATC, T– –

1,041, ATC, TAA

1,042, ATT, T– –

–

–

COI

1,554, ATG, AGG

1,542, ATG, AGG

1,548, ATG, AGA

–

–

COII

684, ATG, TAA

684, ATG, TAA

684, ATG, TAA

–

–

ATP8

192, ATG, TAA

204, ATG, TAA

192, ATG, TAA

–

–

ATP6

681, ATG, TAA

679, ATG, T– –

681, ATG, TAA

–

–

COIII

784, ATG, T– –

784, ATG, T– –

784, ATG, T– –

–

–

ND3

348, ATT, TAA

346, ATT, T– –

348, ATA, TAA

–

–

ND4L

297, ATG, TAA

297, ATG, TAA

297, ATG, TAA

–

–

ND4

1,378, ATG, T– –

1,378, ATG, T– –

1,378, ATG, T– –

–

–

ND5

1,812, ATA, TAA

1,812, ATA, TAA

1,809, ATA, TAA

–

–

ND6

525, ATG, TAG

525, ATG, TAG

525, ATG, TAG

–

–

Cyt b

1,140, ATG, AGA

1,140, ATG, AGA

1,140, ATG, TA –

–

–

1,146

1,157

1,272

–

–

D–loop

Forcina et al.

114

Table 3. Cyt b genetic diversity: N, number of sequences; S, number of polymorphic sites, h, number of haplotypes; Θ, haplotype diversity; and π, nucleotide diversity of the loci amplified across countries and ESUs as inferred by Perez et al. (2013) (n.d., not determined).

Tabla 3. Diversidad genética del citocromo b: N, número de secuencias; S, número de sitios polimórficos; h, número de haplotipos; Θ, diversidad haplotípica; y π, diversidad nucleotídica de los loci amplificados relativos a cada país y ESU según Pérez et al. (2013) (n.d., no determinado).

Locus

Country / ESU

N

S

h

Θ

π

Mitogenome

Spain

4

19

3

0.833

0.001

Cyt b

Portugal

5

7

5

1

0.005

Spain

6

7

3

0.6

0.006

Andorra

2

1

2

1

0.002

France

17

56

16

0.985

0.033

Belgium

4

27

3

0.83

0.024

Italy

16

36

14

0.983

0.022

Austria

1

n.d.

1

n.d

n.d

Germany

2

5

2

1

0.009

Croatia

1

n.d

1

n.d

n.d

Iberian

19

21

13

0.942

0.009

Italian

12

30

9

0.939

0.014

Western European

10

8

7

0.911

0.004

Alpine

13

16

13

1

0.008

to identify unique lineages (Knaus et al., 2011; Johri et al., 2020; Sun et al., 2021), an achievement which is particularly important in the case of endangered species for which new entire mitogenomes are being generated for this specific purpose (Ruiz et al., 2021; Wang et al., 2021; Skorupski, 2022).

The scarcity of genetic data for the garden dormouse from the Iberian Peninsula is a serious gap in our knowledge. Although southern Spain is the southernmost limit of the species' distribution, it appears to host a remarkable phenotypic diversity, and four subspecies have been proposed for this area: E. q. lusitanicus in the south–west, E. q. quercinus elsewhere on its mainland, E. q. gymnesicus in Mallorca and Menorca, and E. q. ophiusae in Formentera (Moreno, 2005). Iberia was an important boreal refugium for many species throughout Pleistocene climate fluctuations (Povoas et al., 1992; Kowalski, 2001; López Antoñanzas and Cuenca Bescós, 2002; Marks et al., 2002; Bicho et al., 2003; López–García et al., 2010; Horníková et al., 2021), but the population structure of the garden dormouse predates these environmental changes (Perez et al., 2013). A high level of morphological diversity is compatible with the fossil record that suggests that southern Spain is likely the center of origin for the species (Mansino et al., 2015), a hypothesis also supported by the molecular phylogeny (Perez et al., 2013). Our results, however, found limited genetic variability in the Iberian ESU in spite of its divergent

position, which is in line with the now consolidated view that the genetically most diverse populations are not those located at the southernmost latitude but at intermediate latitudes, as a consequence of the admixture of divergent lineages arose in separate refugia (Petit et al., 2003; Werner, 2007). In this respect, our results pointed to France, home to garden dormice falling into three different ESUs, as the country hosting the highest genetic diversity for this species (fig. 1; table 3). This outcome is in sharp contrast with that which emerged for the edible dormouse, with the entire continental lineage exhibiting a limited variability –presumably as the result of a recent expansion from a single refuge (Hürner et al., 2010). The garden dormouse pattern is more similar to that found in the hazel dormouse, which is characterized by a strong genetic structure of highly divergent lineages with low internal diversity (Mouton et al., 2012).

Structure within the Iberian Peninsula seems possible (Gómez and Lunt, 2007; Ferrero et al., 2011; Abellán and Svenning, 2014) and is suggested by these preliminary data (fig. 1), but only the genotyping of further individuals at multiple nuclear loci will clarify whether or not this is the case. Studies of other small mammals from this region have revealed a shallow mtDNA genetic structure despite high phenotypic variation (Lucas et al., 2015). In light of the high karyotypic diversity contained in this species (Perez et al., 2013), assessing the chromosomal races of

Animal Biodiversity and Conservation 45.1 (2022)

115

HE978360 G. kelleni MN935777 E. quercinus 100

MZ130254 E. quercinus 98

MZ130252 E. quercinus 100 77

MZ130255 E. quercinus 100

MZ130253 E. quercinus NC_050264 M. avellanarius NC_001892 G. glis 0.03 substitutions/site

Fig. 3. Maximum likelihood (ML) dormice phylogeny based on the whole mitogenome (14,906 bp).

Fig. 3. Filogenia de máxima verosimilitud (ML) de los lirones europeos basada en el mitogenoma entero (14.906 pares de bases).

the garden dormouse from southern Spain and other largely unexplored regions of the Iberian Peninsula would add key information to our understanding of the evolutionary history of this charismatic species.

The garden dormouse was historically widely distributed across the entire region but appears to have declined substantially in southern Spain. This local demographic trend has been inferred based on a limited number of trapping sites, and a rebound has been suggested in recent times (https://www. efe.com/efe/andalucia/huelva/constatan-aumento-de-lirones-careto-en-donana-indicador-mejora-forestal/50001127-4474704). Thus, more systematic genetic and ecological surveys for this species are needed to understand its occurrence outside protected areas, such as in orange and olive orchards and vineyards, where it is still recorded relatively frequently (Rey Benayas et al., 2017) according to GBIF records from the last 14 years (GBIF, 2021; Villares Muyo and Ruiz Franco, 2020) and from a recent study from central Italy based on common barn–owl (Tyto alba) pellets (Paniccia et al., 2022). Nevertheless, a paucity of information for central and southern Portugal (Cabral et al., 2005) points to a likely decline in Portugal as well. There are many possible drivers of this decline, and they are not mutually exclusive. The garden dormouse is generally regarded as a cold–adapted

species (Lanni et al., 1990; Perez et al., 2013; Giroud et al., 2018), even if for some authors this would likely apply to the northern chromosomal race as opposed to that from Iberia (Libois et al., 2012). Indeed, southern Spain is hot and dry, at the limit of the species' tolerance to environmental factors. Should the increasing demographic trend suggested in Andalusia be confirmed, climate warming as a leading driver of global decline in this species would be called into question.

The level and distribution of genetic diversity in Iberian garden dormice may be important for assessing the conservation priority of different populations in the context of the much advocated monitoring project for this species (Bertolino et al., 2014), and determine if they are able to survive these increasing challenges.

For this purpose, the sequencing of unlinked nuclear loci, such as introns (Igea et al., 2010; Rodríguez– Prieto et al., 2014), tested in the garden dormouse (Forcina et al., 2021), or other exonic markers widely used across rodent phylogenies (Schenk et al., 2013), is recommended as well.

In conclusion, we hope that the genetic resources reported here will stimulate further research aimed at understanding and preserving the genetic and morphological diversity of the garden dormouse and that of its poorly studied European relatives.

116

Acknowledgements The authors are thankful to Irene Quintanilla for her highly qualified laboratory work. Logistical support was provided by the Laboratory of Molecular Ecology (LEM–EBD), and by the infrastructures offered by Doñana’s Singular Scientific–Technical Infrastructure (ICTS–EBD). Materials were provided by the scientific collection at the Estación Biológica de Doñana.

The authors are also grateful to Carlos Urdiales for his assistance with material from the scientific collection, while nature photographers Pepe Garcia, Margitta Schuff Thomann, Klaus Rudloff and Nedko Nedyalkov (National Museum of Natural History, Bulgarian Academy of Sciences) are acknowledged for granting permission to use their pictures of E. quercinus (https://www.nikonistas.com/digital/foro/topic/409194–lir%C3%B3n–careto/), G. glis (https://www. flickr.com/photos/29698005@N04/42116504605), G. kelleni (https://www.biolib.cz/en/image/id318057/) and M. roachi (https://thehabitatfoundation.org/category/ species–groups/rodents/), respectively. The pictures of M. avellanarius (https://commons.wikimedia.org/ wiki/File:Muscardinus_avellanarius_–_MUSE.JPG) and D. nitedula (https://de.wikipedia.org/wiki/Baumschl%C3%A4fer#/media/Datei:Erdei_pele.jpg) were taken by Matteo De Stefano (Museo delle Scienze, Trento, Italy) and 'Attis'. Both photographs are freely usable under license CC BY–SA 3.0. The authors are also thankful to Sarah Viola Emser (Genomics Core Facility, Vetmeduni Vienna) for providing locality information on concerning the E. quercinus reference mitogenome deposited in GenBank (MN935777). This work was supported by the former Spanish Ministry of Science and Innovation (now Spanish Ministry of Economy, Industry and Competition) grants CGL2014–58793–P and CGL2017–86068–P, LifeWatch, and a 'Centre of Excellence Severo Ochoa' award to EBD–CSIC (SEV–2012–0262). Partial support was also granted by fellowships from the Portuguese Foundation for Science and Technology (FCT, PTDC/BAA–AGR/28866/2017) and the Spanish Government, Ministry of Universities (María Zambrano – Next Generation EU).

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An endangered new catfish species of the genus Cambeva (Cambeva gamabelardense n. sp.)

(Siluriformes, Trichomycteridae) from the Rio Chapecó drainage, southern Brazil W. J. E. M. Costa, C. R. M. Feltrin, A. M. Katz Costa, W. J. E. M., Feltrin, C. R. M., Katz, A. M., 2022. An endangered new catfish species of the genus Cambeva (Cambeva gamabelardense n. sp.) (Siluriformes, Trichomycteridae) from the Rio Chapecó drainage, southern Brazil. Animal Biodiversity and Conservation, 45.1: 123–129, Doi: https://doi.org/10.32800/abc.2022.45.0123 ZooBank LSID: http://zoobank.org/References/A5426669-DA97-44EE-A790-622AEC683854 Abstract An endangered new catfish species of the genus Cambeva (Cambeva gamabelardense n. sp.) (Siluriformes, Trichomycteridae) from the Rio Chapecó drainage, southern Brazil. Numerous species in fast–flowing streams of southern Brazil have not been described to date. As some of these species inhabit areas under pressure due to the ongoing, intense process of environmental degradation, formal descriptions are urgently needed so as to elaborate strategies for their conservation. We describe a new species, Cambeva gamabelardense n. sp., found in the middle Rio Chapecó drainage, Uruguay River basin, in an area where intense deforestation and soya plantation is endangering fish species. The new species is considered closely related to C. panthera, a species occurring in an isolated coastal basin about 380 km from the area inhabited by the new species, as the two species share a unique jaguar–like pattern on the flank. The new species differs from C. panthera by having shorter barbels, a different position of the origin of the dorsal–fin, more vertebrae, and osteological features that are unique among congeners.

Key words: Cambeva gamabelardense n. sp., Comparative morphology, Mountain biodiversity, Osteology, Uruguay River basin Resumen Una nueva especie de bagre en peligro de extinción del género Cambeva (Cambeva gamabelardense sp. n.)

(Siluriformes, Trichomycteridae) en la cuenca hidrográfica del río Chapecó, en el sur del Brasil. Se ha registrado una gran diversidad de especies no descritas para las corrientes rápidas del sur de Brasil, algunas de las cuales habitan zonas sometidas a un intenso proceso de degradación ambiental, lo que hace que sea urgente describir estas especies formalmente como primer paso para elaborar estrategias de conservación. La nueva especie descrita aquí, Cambeva gamabelardense sp. n., fue encontrada en el tramo medio del río Chapecó, en la cuenca del río Uruguay, en una zona sometida a una intensa deforestación y al cultivo de soja, lo que pone a las especies piscícolas en riesgo de extinción. Se considera que la nueva especie está estrechamente relacionada con C. panthera, que se encuentra en una cuenca costera aislada, a unos 380 km de la zona habitada por la nueva especie, ya que ambas especies comparten un patrón de color similar al del jaguar en el flanco. La nueva especie se diferencia de C. panthera en que tiene barbillas más cortas, una posición relativa diferente del origen de la aleta dorsal y más vértebras, además de tener características osteológicas únicas entre sus congéneres.

Palabras clave: Cambeva gamabelardense sp. n., Morfología comparada, Biodiversidad en zonas montañosas, Osteología, Cuenca del río Uruguay Received: 10 I 22; Final acceptance: 12 IV 22 Wilson J. E. M. Costa, Axel M. Katz, Laboratory of Systematics and Evolution of Teleost Fishes, Institute of Biology, Federal University of Rio de Janeiro, Caixa Postal 68049, CEP 21941–971, Rio de Janeiro, Brazil.– Caio R. M. Feltrin, Av. Municipal, 45, Siderópolis, CEP 88860–000, Santa Catarina, Brazil.

Corresponding author: W. J. E. M. Costa. E–mail: wcosta@acd.ufrj.br ORCID ID: W. J. E. M. Costa: 0000-0002-0428-638X; C. R. M. Feltrin: 0000-0002-1609-7295; A. M. Katz: 0000-0002-2933-7163 ISSN: 1578–665 X eISSN: 2014–928 X

© [2022] Copyright belongs to the authors, who license the journal Animal Biodiversity and Conservation to publish the paper under a Creative Commons Attribution 4.0 License.

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Introduction The fast–flowing streams draining mountain ranges of south–eastern and southern Brazil support a great diversity of trichomycterine catfishes (Costa, 2021; Costa et al., 2021a, 2021b). Trichomycterines comprise the most diverse subfamily of the Trichomycteridae (i.e. Trichomycterinae), and over 260 species are found in most mountain river drainages of South America (Katz et al., 2018; Costa et al., 2021b).

However, the rich diversity of trichomycterines in the upper–middle Rio Uruguai basin, Santa Catarina State, has only recently been revealed. Eight species of the genus Cambeva Katz, Barbosa, Mattos and Costa, 2018, endemic to this wide region, have been described in the last two years (Costa et al., 2020a, 2021b, 2022). Field studies have also shown the increasingly intense process of environmental degradation in the river drainages of the region, supporting the urgent need for descriptions of new species as a first step to elaborate strategies for their conservation.

We recently reported two new species from the middle section of the Rio Chapecó drainage, middle Rio Uruguai basin, that share a peculiar morphology (Costa et al., 2022). These species were collected during two expeditions (15 VII 2020 and 20 III 2021) to an inner plateau, about 640–900 m a.s.l. A single small specimen of a third, as yet non–described species, was collected in the second expedition. These species are endangered by the severe impact of intense deforestation and soya plantations (Costa et al., 2022). During a third expedition (5 VIII 2021), new specimens of this third species were collected. The objectives of this paper are to formally describe this new species and provide new data on environment factors threatening its survival.

Material and methods Morphometric and meristic data were taken following Costa (1992), with modifications proposed by Costa et al. (2020b). Measurements are presented as percent of standard length (SL) except for those related to head morphology which are expressed as percent of head length. Fin–ray counts include all elements and are expressed in lower case Roman numerals for procurrent unsegmented unbranched rays of unpaired fins, upper case Roman numerals for segmented unbranched rays of any fin, and Arabic numerals for segmented branched rays of any fin (Bockmann et al., 2004; Costa et al., 2020b). Vertebra counts include all vertebrae except those participating in the Weberian apparatus, with the compound caudal centrum counted as a single element.

Specimens were cleared and stained (C and S in list of specimens) for bone and cartilage following Taylor and Van Dyke (1985). In addition to morphological characters commonly used in taxonomical studies on trichomycterines, descriptions include some osteological structures that have informative variability for diagnosing species of Cambeva (Costa et al., 2020b, 2021a), including the mesethmoidal and cheek regions and the parurohyal morphology. Terminology for bones followed Costa (2021). Osteological illustrations were

made using a stereomicroscope Zeiss Stemi SV 6 with camera lucida. Cephalic laterosensory system terminology follows Arratia and Huaquin (1995), with modifications proposed by Bockmann et al. (2004).

Comparative material is listed in Costa (2021), with the addition of Cambeva balios (Ferrer and Malabarba, 2013) (UFRJ 7024, 2 C and S), Cambeva perkos (Datovo, Carvalho and Ferrer, 2012) (UFRJ 7025, 3 C and S), and Cambeva tropeira Ferrer and Malabarba, 2011) (UFRJ 6935, 2 C and S), besides specimens of congeners described more recently (Costa et al., 2021a, 2021b). Geographical names follow Portuguese terms used in the region. The material is deposited in the ichthyological collection of the Institute of Biology of the Federal University of Rio de Janeiro, Rio de Janeiro city, Brazil (UFRJ), and the ichthyological collection of the Centre of Agrarian and Environmental Sciences, Federal University of Maranhão, Campus Chapadinha, Brazil (CICCAA).

Results Phylum Chordata Haeckel, 1874 Class Actinopterygii Klein, 1885 Order Siluriformes Cuvier, 1817 Family Trichomycteridae Bleeker, 1858 Cambeva gamabelardense n. sp. (figs. 1–2) ZooBank LSID: http://zoobank.org/NomenclaturalActs/ 03ED64B-4EBD-4C1C-80EA-52B25F781730 Holotype UFRJ 7003, 107.2 mm SL; Brazil: Santa Catarina State: Abelardo Luz Municipality: stream tributary to the middle section of the Rio Chapecó, close to Parque Quedas do Chapecó, Rio Uruguai basin, 26º 33' 06'' S 52º 19' 17'' W, about 755 m a.s.l.; C. R. M. Feltrin, 5 VIII 2021.

Paratypes All from Brazil: Santa Catarina State: Abelardo Luz Municipality: middle Rio Chapecó drainage, Rio Uruguai basin: UFRJ 7004, 7, 35.4–111.2 mm SL; UFRJ 7005, 4 (C and S), 43.7–72.3 mm SL; CICCAA 02713, 3, 48.3–71.7 mm SL; all collected with holotype.

UFRJ 7017, 5, 28.8–40.9 mm SL; stream tributary to Rio Nova Aurora, 26º 28' 15'' S 52º 20' 12'' W, about 860 m a.s.l.; C.R.M. Feltrin, 5 VIII 2021.

Diagnosis Cambeva gamabelardense is distinguished from all other congeners, except C. panthera Costa, Feltrin and Katz, 2021, by having a jaguar–like colour pattern on the flank, consisting of irregularly shaped pale brown spots of variable size and shape, with their margins overlapped by small dark brown spots (fig. 1; vs. never a similar colour pattern). Cambeva gamabelardense differs from C. panthera by having shorter barbels, with the tip of the maxillary and rictal barbels reaching the middle portion of the interopercular patch of odontodes (vs. the tip of the maxillary barbel reaching the middle of the pectoral–fin base

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A

B

C

Fig. 1. Cambeva gamabelardense n. sp., holotype, UFRJ 7003, 107.2 mm SL: A, lateral view; B, dorsal view; C, ventral view.

Fig. 1. Cambeva gamabelardense sp. n., holotipo, UFRJ 7003, 107,2 mm SL: A, vista lateral; B, vista dorsal; C, vista ventral.

and the tip of the rictal barbel reaching between the interopercular patch of odontodes and the pectoral–fin base) and the tip of the nasal barbel reaching between the eye and the opercular patch of odontodes (vs. the tip of the nasal barbel reaching the middle of opercular patch of odontodes); the first pectoral–fin ray terminating in a rudimentary filament, weakly extending beyond the fin membrane (vs. filament about 10–15 % of the pectoral–fin length); the posterior extremity of the pelvic fin at a vertical through the anterior portion of the dorsal–fin base (vs. middle portion); the dorsal–fin origin at a vertical through the centrum of the 20th or 21st vertebra (vs. through the centrum of the 18th or 19th vertebra); and 38 or 39 vertebrae (vs. 37). Cambeva gamabelardense differs from all other trichomycterines examined by its long metapterygoid, its horizontal length longer than the horizontal length of hyomandibula anterior outgrowth (fig. 2B; vs. shorter) and a small lateral projection on the lateral margin of the lateral ethmoid, just posterior to the articular facet for the autopalatine (fig. 2A; vs. absence of a similar projection).

Description General morphology Morphometric data are shown in table 1. Body moderately slender, subcylindrical anteriorly to com-

pressed posteriorly. Greatest body depth between pectoral and pelvic fins. Dorsal and ventral profiles of head and trunk slightly convex. Skin papillae minute.

Anus and urogenital papilla at vertical just anterior to middle of dorsal–fin base. Head trapezoidal to sub–rectangular in dorsal view, anterior profile of snout slightly convex. Eye small, dorsally positioned on middle portion of head. Posterior nostril located nearer anterior nostril than orbital rim. Tip of maxillary and rictal barbels reaching middle portion of interopercular patch of odontodes; tip of nasal barbel reaching between eye and opercular patch of odontodes.

Mouth subterminal. Jaw teeth arranged in irregular transverse rows, slightly curved, pointed in internal to incisiform in external rows, 40–43 on premaxilla, 41–44 on dentary. Branchial membrane attached to isthmus only at its anterior point, in ventral midline.

Branchiostegal rays 8 or 9.

Dorsal and anal fins subtriangular, distal margin slightly convex; total dorsal–fin rays 11–12 (ii– iii + II + 7), total anal–fin rays 10 (iii + I–II + 5–6); anal–fin origin at vertical through posterior half of dorsal–fin base, base of first procurrent anal–fin ray in vertical through base of 6th principal dorsal–fin ray. Dorsal–fin origin at vertical through centrum of 20th or 21st vertebra; anal–fin origin at vertical through centrum of 24th or 25th vertebra. Pectoral

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A

B

Premaxilla

Metapterygoid Hyomandibula

Maxilla Antorbital Autopalatine Lateral ethmoid Sesamoid supraorbital

C

Opercle

Mesethmoid Quadrate Lep Frontal 1 mm

Preopercle

1 mm

1 mm

Interopercle

Fig. 2. Osteological features in Cambeva gamabelardense n. sp.: A, mesethmoidal region, middle and left portions, dorsal view; B, left suspensorium and opercular apparatus, lateral view; C, parurohyal, ventral view; Lep, lateral ethmoid projection. (Larger stippling represents cartilages).

Fig. 2. Características osteológicas de Cambeva gamabelardense sp. n.: A, vista dorsal de las porciones media e izquierda de la región mesetmoidal; B, vista lateral del suspensorio izquierdo y el aparato opercular; C, vista ventral del parurohial; Lep, proyección lateral del etmoides. (El punteado más claro representa los cartílagos).

fin subtriangular in dorsal view, posterior margin slightly convex, first pectoral–fin ray terminating in rudimentary filament, weakly extending beyond fin membrane; total pectoral–fin rays 7 (I + 6). Pelvic fin subtruncate, its extremity at vertical through anterior portion of dorsal–fin base; pelvic–fin bases medially in close proximity; total pelvic–fin rays 5 (I + 4).

Caudal fin subtruncate, postero–dorsal and postero–ventral extremities slightly rounded; total principal caudal–fin rays 13 (I + 11 + I), total dorsal procurrent rays 22–23 (xxi–xxii + I), total ventral procurrent rays 14–15 (xiii–xiv + I). Vertebrae 38–39. Ribs 14–15.

Two dorsal hypural plates, corresponding to hypurals 4 + 5 and 3, respectively; single ventral hypural plate corresponding to hypurals 1 and 2 and parhypural.

Supraorbital sensory canal continuous, connected to infraorbital sensory canal posteriorly. Supraorbital sensory canal with 3 pores: s1, adjacent to medial margin of anterior nostril; s3, adjacent to medial margin of posterior nostril; and s6, on middle part of dorsal surface of head, in transverse line just posterior to orbit; pore s6 slightly nearer orbit than its paired homologous pore. Anterior segment of infraorbital sensory canal absent; posterior segment with pore i10, adjacent to ventral margin of orbit, and pore i11, posterior to orbit. Postorbital canal with 2 pores: po1, at vertical line above posterior portion of interopercular patch of odontodes, and po2, at vertical line above posterior portion of opercular patch of odontodes. Lateral line of body short, with 2–3 pores, posterior–most pore at vertical just posterior to pectoral–fin base.

Mesethmoidal region (fig. 2A) Mesethmoid robust, its anterior margin about straight; mesethmoid cornu subtriangular in dorsal

view, extremity pointed; narrow lateral flap on intersection between cornu and main bone axis. Small lateral projection on thickened lateral margin of lateral ethmoid, just posterior to articular facet for autopalatine. Antorbital thin, drop–shaped, short, slightly longer than wide; sesamoid supraorbital slender, its length about three or four times antorbital length.

Premaxilla sub–trapezoidal in dorsal view, longer than maxilla. Maxilla boomerang–shaped, slender, slightly curved. Autopalatine sub–rectangular in dorsal view, broad, slightly longer than wide, with slightly sinuous medial margin and gently concave lateral margin; autopalatine posterolateral process subtriangular in dorsal view, short, its length about half the length of osseous portion of autopalatine, excluding posterolateral process.

Cheek region (fig. 2B) Metapterygoid thin, subtrapezoidal, large, its horizontal length longer than the horizontal length of the hyomandibula anterior outgrowth. Quadrate robust, dorsal process with constricted base and long posterior outgrowth, dorsoposterior margin in contact with the hyomandibula outgrowth. Hyomandibula relatively short, with well–developed anterior outgrowth, with weak to moderate dorsal concavity. Opercle relatively slender, opercular odontode patch depth shorter than dorsal hyomandibula articular facet, with 12 or 13 odontodes; odontodes pointed, straight to slightly curved, arranged in irregular transverse rows; dorsal process of opercle short; opercular articular facet for hyomandibula with small, rounded flat extension, articular facet for preopercle rounded, close to opercular facet. Interopercle moderate, about three fourths hyomandibula length, anterior margin slightly concave;

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interopercular odontode patch with 25–28 pointed odontodes, arranged in irregular longitudinal rows.

Preopercle compact, with short ventral extension.

Parurohyal (fig. 2C) Robust, lateral process slender, tip slightly pointed, posterior margin slightly convex; parurohyal head well–developed, with prominent anterolateral paired process; middle foramen minute; posterior process long, its length about equal to distance between anterior margin of parurohyal and anterior insertion of lateral process.

Colouration in alcohol (fig. 1) Flank and dorsum pale yellow; irregularly shaped pale brown spots of variable size and shape, irregularly arranged, smaller on ventral part of flank, with their margins overlapped by small dark brown spots, yielding jaguar–like pattern; similar marks but much smaller on dorsal and lateral surfaces of head. Dorsum with small black dots overlapping brown spots.

Venter and ventral part of head yellowish white.

Barbels pale brown. Fins pale grey to yellowish grey, with transverse rows of minute black dots, larger on distal portion of unpaired fins.

Distribution Cambeva gamabelardense n. sp. is only known from the type locality area, in the upper Rio Chapecó drainage, Rio Uruguai basin, at altitudes between about 750 and 860 m a.s.l. (fig. 3).

Habitat Cambeva gamabelardense n. sp. was commonly found below stones with small and medium size grain, diameter about 1–20 cm, with larger specimens above about 100 mm SL preferring larger stones, diameter about 45–70 cm. Some specimens however were also found below amphibious plants near banks, including excerpts with densely vegetated margins. A few small specimens were collected in a small tributary stream of the Rio Nova Aurora, demonstrating wide plasticity in using different habitats since in this latter location the substrate was clay, with a considerable contribution of organic matter from the riparian forest, and rhizomes and other fern structures along the bank of the stream.

Conservation The area is highly affected by soy plantations as the municipality of Abelardo Luz is considered the national capital of soybean seed, producing 50 types of cultivars and representing approximately 50 tons/ year. This agricultural activity in the area has severely reduced the extension of riparian forest, and measures required by law are frequently overlooked. Massive soy production has exacerbated the volume of siltation in rivers and especially in streams. In addition to silting, soy monoculture requires high loads of pesticides, which in turn affect the micro–basins in question. The Chapecó River has a high potential for hydroelectric power plants. Some stations have already been built and speculation into others is on–going. In short, natural habitats along the main

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Table 1. Morphometric data of Cambeva gamabelardense n. sp.: H, holotype; P, paratypes (n = 10).

Tabla 1. Datos morfométricos de Cambeva gamabelardense sp. n.: H, holtipo; P, paratipos (n = 10).

H Standard length (mm)

P

107.2 43.7–111.2

Percent of standard length Body depth

14.2

14.8–16.8

Caudal peduncle depth 11.6

11.6–13.3

Body width

10.7–12.6

12.1

Caudal peduncle width 4.8

3.8–5.2

Pre–dorsal length

61.7

60.7–63.1

Pre–pelvic length

58.1

56.1–60.6

Dorsal–fin base length 11.5

10.6–13.4

Anal–fin base length

8.7

7.5–10.0

Caudal–fin length

17.3

15.5–18.0

Pectoral–fin length

10.9

11.7–14.5

Pelvic–fin length

8.9

8.1–9.7

Head length

19.1

18.6–22.0

Percent of head length Head depth

55.0

49.6–55.3

Head width

88.2

74.1–86.3

Snout length

45.5

42.7–48.3

Interorbital length

24.1

23.0–27.1

Preorbital length

13.0

11.4–14.4

Eye diameter

7.1

7.6–12.3

river channel are gradually disappearing, and species are becoming restricted to peripheral stream environments, habitats that are also threatened by agricultural practices such as pesticides and siltation.

The type–locality is located in a tributary stream of the Chapecó River, which is in the damping zone of the Quedas do Chapecó Park, an important tourist spot in the city. The park plays the important role of displaying the natural beauty of Rio Chapecó and preserving the local environment, beneficial factors for the conservation of the type locality. However, in the areas surrounding the Park, the monoculture of soy has a tremendously negative impact on the riparian forest, often completely occupying upstream stretches, and even including their sources.

Etymology The name gamabelardense (gamma, the third letter of the Greek alphabet, and abelardense, a Portu-

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50º0'W

có

e ap

h

C io R 27º0'S

27º0'S Ri

o

Ur

ug

ua

i

0 25 50 75

100 km

50º0'W

Fig. 3. Geographical distribution of Cambeva gamabelardense n. sp. (stars) and C. panthera (lozenge).

Fig. 3. Distribución geográfica de Cambeva gamabelardense sp. n. (estrellas) y C. panthera (rombo).

guese word referring to people born in Abelardo Luz municipality) is an allusion to the third new species of Cambeva known to occur in this area. The other two species are described in Costa et al. (2022).

Discussion The wide diversity of colour patterns in Trichomycterines (Eigenmann, 1918) plays an important role in distinguishing species of Cambeva (Costa, 1992; de Pinna, 1992; Wosiacki and Garavello, 2004; Datovo et al., 2012; Ferrer and Malabarba, 2013; Costa et al., 2020a, 2021b, 2022; Reis et al., 2021).

Cambeva gamabelardense n. sp. has a peculiar colour pattern, resembling the typical colouration of the spotted jaguar (fig. 1). Among the approximately 120 species presently included in the eastern South American trichomycterine clade comprising the genera Cambeva, Scleronema Eigenmann, 1917 and Trichomycterus Valenciennes, 1832, only Cambeva panthera has a similar colour pattern, thus suggesting that this species is the closest relative to C. gamabelardense. Although it is not yet possible to infer osteological characters unambiguously supporting sister group relationships between C. gamabelardense and C. panthera, the high similarity of most osseus structures, including a broad autopalatine with a short latero–posterior process (compare fig. 2A with Costa et al., 2021b: fig. 11E), a derived condition occurring in a few intrageneric lineages (Costa et al., 2020, 2021b, 2022), reinforces this hypothesis. However, in contrast with C. gamabelardense,

here described from an inner plateau, about 640– 900 m a.s.l., drained by the middle section of the Rio Chapecó and its tributaries, C. panthera occurs in a lower altitude area, about 245 m a.s.l., in a small isolated coastal drainage, the Rio Tubarão basin, about 380 km from the type locality area of C. gamabelardense (Costa et al., 2021b).

Recent studies have reported a rich biodiversity, with numerous endemic species, in the inner plateau drained by the Rio Uruguai basin (Boldrini et al., 2009; Costa et al., 2021b), which is part of the biogeographical province known as the Araucaria plateau or Araucaria Forest (Ab'Saber, 1977; Morrone, 2006).

Cambeva gamabelardense and two other recently described species are only known from a small area of about 80 km2 originally covered by the Araucaria Forest, which is presently undergoing intense deforestation for soya monoculture (see above). The concentration of three new species in a relatively small area indicates the urgent need for additional field studies in the region to check their occurrence in neighbouring areas in order to allow consistent evaluations of the conservation status of these rare endemic species.

Acknowledgements CRMF is grateful to Alexandre Bianco e Ronaldo dos Santos Junior for help during collections. This work was partly supported by Conselho Nacional de Desenvolvimento Científico e Tecnológico (CNPq; grant 304755/2020–6 to WJEMC), and Fundação Carlos

Animal Biodiversity and Conservation 45.1 (2022)

Chagas Filho de Amparo à Pesquisa do Estado do Rio de Janeiro (FAPERJ; grant E–26/201.213/2021 to WJEMC and E–26/202.005/2020 to AMK). This study was also supported by CAPES (Coordenação de Aperfeiçoamento de Pessoal de Nível Superior, Finance Code 001) through Programa de Pós–Graduação em: Biodiversidade e Biologia Evolutiva/ UFRJ; Genética/UFRJ; and Zoologia, Museu Nacional/UFRJ. Instituto Chico Mendes de Conservação da Biodiversidade (38553–10) provided collecting permits.

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– 2022. Two new remarkable and endangered catfish species of the genus Cambeva (Siluriformes, Trichomycteridae) from southern Brazil. European Journal of Taxonomy, 794(1): 140–155, Doi: 10.5852/ejt.2022.794.1661 Costa, W. J. E. M., Katz, A. M., Mattos, J. L. O., Amorim, P. F., Mesquita, B. O., Vilardo, P. J., Barbosa, M. A., 2020b. Historical review and redescription of three poorly known species of the catfish genus Trichomycterus from south– eastern Brazil (Siluriformes: Trichomycteridae).

Journal of Natural History, 53: 2905–2928, Doi: 10.1080/00222933.2020.1752406 Datovo, A., Carvalho, M., Ferrer, J., 2012. A new species of the catfish genus Trichomycterus from the La Plata River basin, southern Brazil, with comments on its putative phylogenetic position (Siluriformes: Trichomycteridae). Zootaxa, 3327: 33–44, Doi: 10.11646/zootaxa.3327.1.3 de Pinna, M. C. C., 1992. Trichomycterus castroi, a new species of trichomycterid catfish from the Rio Iguaçu of Southeastern Brazil (Teleostei, Siluriformes). Ichthyological Exploration of Freshwaters, 3: 89–95.

Ferrer, J., Malabarba, L. R., 2013. Taxonomic review of the genus Trichomycterus Valenciennes (Siluriformes: Trichomycteridae) from the Laguna dos Patos system, Southern Brazil. Neotropical Ichthyology, 11: 217–246, Doi: /10.1590/S167962252013000200001 Katz, A. M., Barbosa, M. A., Mattos, J. L. O., Costa, W. J. E. M., 2018. Multigene analysis of the catfish genus Trichomycterus and description of a new South American trichomycterine genus (Siluriformes, Trichomycteridae). Zoosystematics and Evolution, 94: 557–566, Doi: 10.3897/zse.94.29872 Morrone, J. J., 2006. Biogeographic areas and transition zones of latin america and the caribbean islands based on panbiogeographic and cladistic analyses of the entomofauna. Annual Review of Entomology, 51: 467–494, Doi: 10.1146/annurev. ento.50.071803.130447 Reis, R. B. dos, Ferrer, J., Graça, W. J. da, 2021. A new species of Cambeva (Siluriformes, Trichomycteridae) from the Rio Iguaçu basin, Paraná State, Brazil and redescription of Cambeva stawiarski (Miranda Ribeiro 1968). Journal of Fish Biology, 96: 350–363, Doi: 10.1111/jfb.14947 Taylor, W. R., Van Dyke, O. C., 1985. Revised procedures for staining and clearing small fishes and other vertebrates for bone and cartilage study.

Cybium, 9: 107–109.

Wosiacki, W. B., Garavello, J. C., 2004. Five new species of Trichomycterus from the Iguaçu (rio Paraná Basin), southern Brazil (Siluriformes: Trichomycteridae). Ichthyological Exploration of Freshwaters, 15: 1–16.

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Costa et al.

Animal Biodiversity and Conservation 45.1 (2022)

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 (museucienciesjournals. cat/amz). 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 museucienciesjournals.cat/abc 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. Publicar a Animal Biodiversity and Conservation es gratuït per als autors, tot i que 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. © 2022 Museu de Ciències Naturals de Barcelona Papers are published under a Creative Commons Attribution 4.0 International License

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Resumen en castellà, traducció de l'Abstract. De la traducció se'n farà càrrec la revista per a aquells autors que no siguin castellano­parlants.

Palabras clave en castellà.

Adreça postal de l’autor o autors, 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-

ball serà establerta i s’ordenarà alfabè­ticament per autors i cronològicament per a un mateix autor, afegint les lletres a, b, c,... als treballs del mateix any. En el text, s’indi­caran en la forma usual: "... segons Wemmer (1998)...", "...ha estat definit per Robinson i Redford (1991)...", "...les prospeccions realitzades (Begon et al., 1999)...".

Taules. Es numeraran 1, 2, 3, etc. i han de ser sempre ressenyades en el text. Les taules grans seran més estretes i llargues que amples i curtes ja que s'han d'encaixar en l'amplada de la caixa de la revista.

Figures. Tota classe d’il·lustracions (gràfics, figures o fotografies) entraran amb el nom de figura i es numeraran 1, 2, 3, etc. i han de ser sempre ressenyades en el text. Es podran incloure fotografies si són imprescindibles. Si les fotografies són en color, el cost de la seva publicació anirà a càrrec dels autors. La mida màxima de les figures és de 15,5 cm d'amplada per 24 cm d'alçada. S'evitaran les figures tridimensionals. Tant els mapes com els dibuixos han d'incloure l'escala. Els ombreigs preferibles són blanc, negre o trama. S'evitaran els punteigs ja que no es repro­dueixen bé.

Peus de figura i capçaleres de taula. Seran clars, concisos i bilingües en la llengua de l’article i en anglès.

Grans quantitats de dades o taules numèriques molt llargues es publicaran com a Material suplementari. Aquest material suplementari només acompanyarà a la versió online de l'article, en cap cas a la versió impresa.

Els títols dels apartats generals de l’article (Introducción, Material y métodos, Resultados, Discusión, Conclusiones, Agradecimientos y Referencias) no aniran numerats. No es poden utilitzar més de tres nivells de títols.

Els autors procuraran que els seus treballs originals no passin de 20 pàgines (incloent–hi figures i taules).

Si a l'article es descriuen nous tàxons, caldrà que els tipus estiguin dipositats en una insti­tució pública.

Es recomana als autors la consulta de fascicles recents de la revista per tenir en compte les seves normes.

Comunicacions breus Les comunicacions breus seguiran el mateix procediment que els articles i tindran el mateix procés de revisió. No excediran de 2.300 paraules incloent–hi títol, resum, capçaleres de taula, peus de figura, agraïments i referències. El resum no ha de passar de 100 paraules i el nombre de referències ha de ser de 15 com a màxim. Que el text tingui apartats és opcional i el nombre de taules i/o figures admeses serà de dos de cada com a màxim. En qualsevol cas, el treball maquetat no podrà excedir les quatre pàgines.

Animal Biodiversity and Conservation 45.1 (2022)

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 (museucienciesjournals.cat/amz). 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.museucienciesjournals.cat/abc lo que permite una difusión mundial de sus artículos.

Todos los manuscritos son revisados por el editor ejecutivo, un editor y dos revisores independientes, elegidos de una lista internacional, a fin de garantizar su calidad. El proceso de revisión es rápido y constructivo, y se realiza vía correo electrónico siempre que es posible. La publicación de los trabajos aceptados se realiza con la mayor rapidez posible, normalmente dentro de los 12 meses siguientes a la recepción del trabajo.

Una vez aceptado, el trabajo pasará a ser propiedad de la revista. Ésta se reserva los derechos de autor, y ninguna parte del trabajo podrá ser reproducida sin citar su procedencia.

Los derechos de autor quedan reservados a los autores, quienes autorizan a la revista a publicar el artículo. Los artículos se publican con una Licencia Creative Commons Atribución 4.0 Internacional: no se podrá reproducir ni reutilizar ninguna de sus partes sin citar la procedencia.

Normas de publicación Los trabajos se enviarán preferentemente de forma electrónica (abc@bcn.cat). El formato preferido es un documento Rich Text Format (RTF) o DOC, que incluya las figuras y las tablas. Las figuras deberán enviarse también en archivos separados en formato TIFF, EPS o JPEG. Debe incluirse, con el artículo, una carta donde conste que el trabajo versa sobre inves­ tigaciones originales no publi­cadas an­te­rior­mente y que se somete en exclusiva a Animal Biodiversity and Conservation. En dicha carta también debe constar, para trabajos donde sea necesaria la manipulación de animales, que los autores disponen de los permiISSN: 1578–665X eISSN: 2014–928X

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sos necesarios y que han cumplido la normativa de protección animal vigente. Los autores pueden enviar también sugerencias para asesores.

Las pruebas de imprenta enviadas a los autores deberán remitirse corregidas al Consejo Editor en el plazo máximo de 10 días. Publicar en Animal Biodiversity and Conservation es gratuito para los autores, sin embargo los gastos debidos a modificaciones sustanciales en las pruebas de im­pren­­ta, introducidas por los autores, irán a ­cargo de los mismos.

El primer autor recibirá una copia electrónica del trabajo en formato PDF. Manuscritos Los trabajos se presentarán en formato DIN A–4 (30 líneas de 70 espacios cada una) a doble espacio y con las páginas numeradas. Los manuscritos deben estar completos, con tablas y figuras. No enviar las figuras originales hasta que el artículo haya sido aceptado.

El texto podrá redactarse en inglés, castellano o catalán. Se sugiere a los autores que envíen sus trabajos en inglés. La revista ofre­ce, sin cargo ninguno, un servicio de corrección por parte de una persona especializada en revistas científicas. En cualquier caso debe presentarse siempre de forma correcta y con un lenguaje claro y conciso.

Los caracteres en cursiva se utilizarán para los nombres científicos de géneros y especies y para los neologismos que no tengan traducción; las citas textuales, independientemente de la lengua en que estén, irán en letra redonda y entre comillas; el nombre del autor que sigue a un taxón se escribirá también en redonda. Se evitará el uso de términos extranjeros (p. ej.: latín, aleman,...).

Al citar por primera vez una especie en el trabajo, deberá especificarse siempre que sea posible su nombre común.

Los topónimos se escribirán bien en su forma original o bien en la lengua en que esté redactado el trabajo, siguiendo el mismo criterio a lo largo de todo el artículo.

Los números del uno al nueve se escribirán con letras, a excepción de cuando precedan una unidad de medida. Los números mayores de nueve se escribirán con cifras excepto al empezar una frase.

Las fechas se indicarán de la siguiente forma: 28 VI 99 (un único día); 28, 30 VI 99 (días 28 y 30); 28–30 VI 99 (días 28 al 30).

Se evitarán siempre las notas a pie de página.

Formato de los artículos Título. Será conciso pero suficientemente explicativo del contenido del trabajo. Los títulos con designaciones de series numéricas (I, II, III, etc.) serán aceptados excepcionalmente previo consentimiento del editor.

Nombre del autor o autores Abstract en inglés de 12 líneas mecanografiadas (860 espacios como máximo) y que exprese la esencia del manuscrito (introducción, material, métodos, resultados y discusión). Se evitarán las especulaciones y las citas bibliográficas. Irá enca© 2022 Museu de Ciències Naturals de Barcelona Papers are published under a Creative Commons Attribution 4.0 International License

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bezado por el título del trabajo en cursiva.

Key words en inglés (un máximo de seis) que especifiquen el contenido del trabajo por orden de importancia.

Resumen en castellano, traducción del abstract. Su traducción puede ser solicitada a la revista en el caso de autores que no sean castellano hablan­tes.

Palabras clave en castellano.

Direccion postal del autor o autores, se publicarán tal como se indique en el manuscrito recibido.

Identificadores de investigador (ORCID, ResearchID…, al menos del investigador principal y de quien asuma la correspondencia posterior.

(Título, Nombre de los autores, Abstract, Key words, Resumen, Palabras clave, Direcciones postalo e Identificadores de investigador conformarán la primera página.)

Introducción. En ella se dará una idea de los antecedentes del tema tratado, así como de los objetivos del trabajo.

Material y métodos. Incluirá la información referente a las especies estudiadas, aparatos utilizados, metodología de estudio y análisis de los datos y zona de estudio.

Resultados. En esta sección se presentarán únicamente los datos obtenidos que no hayan sido publicados previamente.

Discusión. Se discutirán los resultados y se compararán con otros trabajos relacionados. Las sugerencias sobre investigaciones futuras se podrán incluir al final de este apartado.

Agradecimientos (optativo).

Referencias. Cada trabajo irá acompañado de una bibliografía que incluirá únicamente las publicaciones citadas en el texto.

Las referencias deben presentarse según los modelos siguientes (método Harvard): * Artículos de revista: Conroy, M. J., Noon, B. R., 1996. Mapping of species richness for conservation of biological diversity: conceptual and methodological issues. Ecological Applications, 6: 763–773. * Libros y otras publicaciones no periódicas: Seber, G. A. F., 1982. The estimation of animal abundance. C. Griffin & Company, London. * Trabajos de contribución en libros: Macdonald, D. W., Johnson, D. P., 2001. Dispersal in theory and practice: consequences for conservation biology. In: Dispersal: 358–372 (T. J. Clober, E. Danchin, A. A. Dhondt, J. D. Nichols, Eds.). Oxford University Press, Oxford. * Tesis doctorales: Merilä, J., 1996. Genetic and quantitative trait variation in natural bird populations. Tesis doctoral, Uppsala University. * Los trabajos en prensa sólo se citarán si han sido aceptados para su publicación: Ripoll, M. (in press). The relevance of population

studies to conservation biology: a review. Animal Biodiversity and Conservation.

Las referencias se ordenarán alfabética­men­te por autores, cronológicamen­te para un mismo autor y con las letras a, b, c,... para los tra­bajos de un mismo autor y año. En el texto las referencias bibliográficas se indicarán en la forma usual: "...según Wemmer (1998)...", "...ha sido definido por Robinson y Redford (1991)...", "...las prospecciones realizadas (Begon et al., 1999)...".

Tablas. Se numerarán 1, 2, 3, etc. y se reseñarán todas en el texto. Las tablas grandes deben ser más estrechas y largas que anchas y cortas ya que deben ajustarse a la caja de la revista.

Figuras. Toda clase de ilustraciones (gráficas, figuras o fotografías) se considerarán figuras, se numerarán 1, 2, 3, etc. y se citarán todas en el texto. Pueden incluirse fotografías si son imprescindibles. Si las fotografías son en color, el coste de su publicación irá a cargo de los autores. El tamaño máximo de las figuras es de 15,5 cm de ancho y 24 cm de alto. Deben evitarse las figuras tridimen­sionales. Tanto los mapas como los dibujos deben incluir la escala. Los sombreados preferibles son blanco, negro o trama. Deben evitarse los punteados ya que no se reproducen bien.

Pies de figura y cabeceras de tabla. Serán claros, concisos y bilingües en castellano e inglés.

Grandes cantidades de datos o tablas numéricas muy largas se publicarán como material suplementario. Este material suplementario sólo acompañará a la versión online del artículo, en ningún caso a la versión impresa.

Los títulos de los apartados generales del artículo (Introducción, Material y métodos, Resultados, Discusión, Agradecimientos y Referencias) no se numerarán. No utilizar más de tres niveles de títulos.

Los autores procurarán que sus trabajos originales no excedan las 20 páginas incluidas figuras y tablas.

Si en el artículo se describen nuevos taxones, es imprescindible que los tipos estén depositados en alguna institución pública.

Se recomienda a los autores la consulta de fascículos recientes de la revista para seguir sus directrices.

Comunicaciones breves Las comunicaciones breves seguirán el mismo procedimiento que los artículos y serán sometidas al mismo proceso de revisión. No excederán las 2.300 palabras, incluidos título, resumen, cabeceras de tabla, pies de figura, agradecimientos y referencias.

El resumen no debe sobrepasar las 100 palabras y el número de referencias será de 15 como máximo.

Que el texto tenga apartados es opcional y el número de tablas y/o figuras admitidas será de dos de cada como máximo. En cualquier caso, el trabajo maquetado no podrá exceder las cuatro páginas.

Animal Biodiversity and Conservation 45.1 (2022)

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 (museucienciesjournals.cat/amz). 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 museucienciesjournals.cat/abc 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.

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ISSN: 1578–665X eISSN: 2014–928X

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(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.

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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.

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Animal Biodiversity and Conservation 45.1 (2022)

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Animal Biodiversity and Conservation 45.1 (2022)

107–122 Forcina, G., Camacho Sánchez, M., Cornella, A., Leonard, J. A. Complete mitogenomes of the garden dormouse (Eliomys quercinus): a new resource for the genetic monitoring of a fast–declining European small mammal

123–129 Costa, W. J. M., Feltrin, C. R. M., Katz, A. M. An endangered new catfish species of the genus Cambeva (Cambeva gamabelardense n. sp.) (Siluriformes, Trichomycteridae) from the Rio Chapecó drainage, southern Brazil

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

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

Índex / Índice / Contents Animal Biodiversity and Conservation 45.1 (2022) ISSN 1578–665 X eISSN 2014–928 X 1–12 Ceña, J. C., Ceña, A., Salvador–Vilariño, V., Meneses, J. M., Sánchez–García, C. New data on the status and ecology of a galliform at risk of extinction: the Pyrenean grey partridge (Perdix perdix hispaniensis) in the Iberian System (Soria, Spain) 13–21 Delgado–Martínez, C. M., Mendoza, E. Human disturbance modifies the identity and interaction strength of mammals that consume Attalea butyracea fruit in a neotropical forest 23–31 Mata, F., Mata, P. Nesting preferences of the green sea turtle (Chelonia mydas L.) and the hawksbill sea turtle (Eretmochelys imbricata L.) in the SW of Mahe Island in the Seychelles 33–42 Torre, I., Cahill, S., Grajera, J., Raspall, A., Vilella, M. Small mammal sampling incidents related to wild boar (Sus scrofa) in natural peri– urban areas 43–52 Utevsky, S., Mabrouki, Y., Taybi, A. F., Huseynov, M., Manafov, A., Morhun, H., Shashina, O., Utevsky, G., Khomenko, A., Utevsky, A. New records of leeches of the genus Limnatis (Hirudinea, Praobdellidae) from the South Caucasus and Central Asia: phylogenetic relationships of Eurasian and African populations

53–67 Malo, A. F., Taylor, A., Díaz, M. Native seed dispersal by rodents is negatively influenced by an invasive shrub 69–78 Liang, Y., Ye, H., Cai, X., Tian, E., Li, F., Li, C., Huang, G., Chao, Z. Recombinase polymerase amplification combined with fast DNA extraction for on–spot identification of Deinagkistrodon acutus, a threatened species 79–83 Brief communication Per, E. Status of the common myna Acridotheres tristis Linnaeus, 1766 in Turkey 85–95 Domínguez–Pérez, L., Gil–Delgado, J. A. Population increase of the invasive red– whiskered bulbul Pycnonotus jocosus in Valencia, Spain 97–106 Quesada, J., Chávez–Zichinelli, C. A., García–Arroyo, M., Yeh, P. J., Guevara, R., Izquierdo–Palma, J., MacGregor–Fors, I. Bold or shy? Examining the risk–taking behavior and neophobia of native and exotic house sparrows

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

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