Animal Biodiversity and Conservation issue 34.1 (2011) by Museu Ciències Naturals de Barcelona MCNB

Dibuix de la coberta: Asterias rubens, estrella de mar comuna, estrella de mar común, common starfish, de Jordi Domènech Editor executiu / Editor ejecutivo / Executive Editor Joan Carles Senar

Secretaria de redacció / Secretaría de redacción / Editorial Office

Secretària de redacció / Secretaria de redacción / Managing Editor Montserrat Ferrer

Museu de Ciències Naturals Passeig Picasso s/n 08003 Barcelona, Spain Tel. +34–93–3196912 Fax +34–93–3104999 E–mail abc@bcn.cat

Consell assessor / Consejo asesor / Advisory Board Eulàlia Garcia Anna Omedes Josep Piqué Francesc Uribe

Editors / Editores / Editors Pere Abelló Inst. de Ciències del Mar CMIMA–CSIC, Barcelona, Spain Javier Alba–Tercedor Univ. de Granada, Granada, Spain Russell Alpizar–Jara Univ. of Évora, Portugal Xavier Bellés Centre d' Investigació i Desenvolupament–CSIC, Barcelona, Spain Luís Mª Carrascal Museo Nacional de Ciencias Naturales–CSIC, Madrid, Spain Michael J. Conroy Univ. of Georgia, Athens, USA Adolfo Cordero Univ. de Vigo, Vigo, Spain Mario Díaz Univ. de Castilla–La Mancha, Toledo, Spain José Antonio Donazar Estación Biológica de Doñana–CSIC, Sevilla, Spain Jordi Figuerola Estación Biológica de Doñana–CSIC, Sevilla, Spain Gary D. Grossman Univ. of Georgia, Athens, USA Damià Jaume IMEDEA–CSIC, Univ. de les Illes Balears, Spain Jordi Lleonart Inst. de Ciències del Mar CMIMA–CSIC, Barcelona, Spain Jorge M. Lobo Museo Nacional de Ciencias Naturales–CSIC, Madrid, Spain Pablo J. López–González Univ de Sevilla, Sevilla, Spain Juan José Negro Estación Biológica de Doñana–CSIC, Sevilla, Spain Vicente M. Ortuño Univ. de Alcalá de Henares, Alcalá de Henares, Spain Miquel Palmer IMEDEA–CSIC, Univ. de les Illes Balears, Spain Javier Perez–Barberia The Macaulay Institute, Scotland, United Kingdom Oscar Ramírez Inst. de Biologia Evolutiva UPF–CSIC, Barcelona, Spain Montserrat Ramón Inst. de Ciències del Mar CMIMA–CSIC, Barcelona, Spain Ignacio Ribera Inst. de Biología Evolutiva CSIC–UPF, Barcelona, Spain Pedro Rincón Museo Nacional de Ciencias Naturales–CSIC, Madrid, Spain Alfredo Salvador Museo Nacional de Ciencias Naturales–CSIC, Madrid, Spain José Luís Tellería Univ. Complutense de Madrid, Madrid, Spain Francesc Uribe Museu de Ciències Naturals de Barcelona, Barcelona, Spain Carles Vilà Estación Biológica de Doñana–CSIC, Sevilla, Spain Rafael Zardoya Museo Nacional de Ciencias Naturales–CSIC, Madrid, Spain Consell Editor / Consejo editor / Editorial Board José A. Barrientos Univ. Autònoma de Barcelona, Bellaterra, Spain Jean C. Beaucournu Univ. de Rennes, Rennes, France David M. Bird McGill Univ., Québec, Canada Mats Björklund Uppsala Univ., Uppsala, Sweden Jean Bouillon Univ. Libre de Bruxelles, Brussels, Belgium Miguel Delibes Estación Biológica de Doñana–CSIC, Sevilla, Spain Dario J. Díaz Cosín Univ. Complutense de Madrid, Madrid, Spain Alain Dubois Museum national d’Histoire naturelle–CNRS, Paris, France John Fa Durrell Wildlife Conservation Trust, Jersey, United Kingdom Marco Festa–Bianchet Univ. de Sherbrooke, Québec, Canada Rosa Flos Univ. Politècnica de Catalunya, Barcelona, Spain Josep Mª Gili Inst. de Ciències del Mar CMIMA–CSIC, Barcelona, Spain Edmund Gittenberger Rijksmuseum van Natuurlijke Historie, Leiden, The Netherlands Fernando Hiraldo Estación Biológica de Doñana–CSIC, Sevilla, Spain Patrick Lavelle Inst. Français de recherche scient. pour le develop. en cooperation, Bondy, France Santiago Mas–Coma Univ. de Valencia, Valencia, Spain Joaquín Mateu Barcelona, Spain Neil Metcalfe Univ. of Glasgow, Glasgow, United Kingdom Jacint Nadal Univ. de Barcelona, Barcelona, Spain Stewart B. Peck Carleton Univ., Ottawa, Canada Eduard Petitpierre Univ. de les Illes Balears, Palma de Mallorca, Spain Taylor H. Ricketts Stanford Univ., Stanford, USA Joandomènec Ros Univ. de Barcelona, Barcelona, Spain Valentín Sans–Coma Univ. de Málaga, Málaga, Spain Tore Slagsvold Univ. of Oslo, Oslo, Norway Animal Biodiversity and Conservation 34.1, 2011 © 2011 Museu de Ciències Naturals, Institut de Cultura, Ajuntament de Barcelona Autoedició: Montserrat Ferrer Fotomecànica i impressió: ISSN: 1578–665X Dipòsit legal: B–16.278–58 The journal is freely available online at: http://www.bcn.cat/ABC

Animal Biodiversity and Conservation 34.1 (2011)

A new hygromiid for the Iberian malacofauna: Candidula corbellai n. sp. (Gastropoda, Pulmonata) A. Martínez–Ortí

Martínez–Ortí, A., 2011. A new hygromiid for the Iberian malacofauna: Candidula corbellai n. sp. (Gastropoda, Pulmonata). Animal Biodiversity and Conservation, 34.1: 1–10. Abstract A new hygromiid for the Iberian malacofauna: Candidula corbellai n. sp. (Gastropoda, Pulmonata).— We report a new Iberian hygromiid, Candidula corbellai n. sp., and describe its conchological and anatomical characteristics. This new species is compared with two other Iberian endemic species, Candidula camporroblensis and C. rocandioi, which present similarities in the reproductive system, such as the long flagellum. The shell of the new species is compared with specimens of the type series of these taxa. The reproductive system of C. corbellai n. sp. is distinguished from C. camporroblensis by its longer male part, although the flagellum is shorter than the penis and epiphallus together and it has a long bursa copulatrix with respect to its duct, which is shorter. The epiphallus and the bursa copulatrix duct are longer in C. rocandioi than in C. corbellai n. sp. A geographical distribution map of the three species in the Iberian peninsula is shown. Key words: Hygromiidae, Candidula corbellai n. sp., Candidula camporroblensis, Candidula rocandioi, Catalonia, Iberian peninsula. Resumen Un nuevo higrómido para la malacofauna ibérica: Candidula corbellai sp. n. (Gastropoda, Pulmonata).— Se describen las características conquiológicas y anatómicas de un nuevo higrómido ibérico, Candidula corbellai sp. n. Se compara con otros dos endemismos ibéricos, Candidula camporroblensis y C. rocandioi, especies con las que presenta similitud en el aparato reproductor, ya que ambos poseen el flagelo largo. La concha del nuevo taxón se ha comparado con ejemplares de la serie tipo de estos taxones. En cuanto al aparato reproductor C. corbellai sp. n. se distingue de C. camporroblensis porque la primera posee la porción masculina mucho más larga, aunque el flagelo no lo es tanto como el pene y epifalo juntos y por poseer la bursa copulatrix larga en relación a su conducto, que es corto. Con respecto a C. rocandioi el epifalo y el conducto de la bursa copulatrix son mucho más largos que en C. corbellai sp. n. Además se aporta un mapa de la distribución geográfica de las tres especies en la península Ibérica. Palabras clave: Hygromiidae, Candidula corbellai sp. n., Candidula camporroblensis, Candidula rocandioi, Cataluña, Península Ibérica. (Received: 17 IX 10; Conditional acceptance: 22 X 10; Final acceptance: 15 XI 10) Alberto Martínez–Ortí, Museu Valencià d'Història Natural and Dept. de Zoologia, Fac. de Ciències Biològiques, Univ. de València, c./ Dr. Moliner 50, 46100 Burjassot, Valencia, España (Spain). E–mail: amorti@uv.es

ISSN: 1578–665X

Introduction A series of specimens from a small terrestrial gastropod, collected recently in the north of Catalonia, could not be identified by their shell. These specimens present characteristics of their own and could not be definitively assigned to any of the known taxa in this geographical area. This is common in numerous hygromiid species in which conditions give rise to a morphological convergence that confers high conchological similarity between them (Martínez–Ortí et al., 2000). Samples of the new species described here may have been assigned to another hygromiid in Catalonia. To determine their exact taxonomical status it was necessary to study the anatomical characteristics of the reproductive system. The characteristics of the genitalia allowed us to assign it to the genus Candidula Kobelt, 1871. This genus is represented in the Iberian Peninsula by 13 species, ten of which are endemic, while the other three are widely dispersed in central and western Europe (Manga, 1983; Gittenberger, 1993a, 1993b; Altonaga et al., 1994; Puente, 1994; Bragado et al., 2010; Holyoak & Holyoak, 2010). Results The reproductive system of Candidula is characterized mainly by the presence in its stimulator system of a dart sac on one side with a dart inside, and a second rudimentary sac between this and the vagina that cannot be seen externally (Hausdorf, 1988). The shell morphology, the reproductive system anatomy, the radula, the jaw and their distribution area are described, drawn and compared with the species of Candidula that are most similar regarding the reproductive system, C. camporroblensis Fez, 1944 and C. rocandioi (Ortiz de Zárate, 1950), both of which also have a long flagellum. The studies used for morpho–anatomical comparison of the three taxa are Fez (1944), where C. camporroblensis is described, Ortiz de Zárate (1950, 1991) where the author describes C. rocandioi and studies the reproductive system of both species, Aparicio (1982), whose study details the differences between several Iberian hygromiids based on their anatomical characteristics, Manga (1983) who studies C. rocandioi, and Faci (1991) and Martínez–Ortí et al. (2000) whose study investigated the conchological and anatomical characteristics of the reproductive system of C. camporroblensis. Finally, we provide a map showing the distribution area of the three species in the Iberian Peninsula; it can be seen that C. corbellai n. sp. is found in an area away from the other two species. Family Hygromiidae Tryon, 1866 Genus Candidula Kobelt, 1871 Candidula corbellai n. sp. Type locality Sierra de Busa (Navés, Lleida), up to 1,375 m altitude, under stones in a calcareous meadow with southern

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orientation. Collected 11th October 2008 by Jordi Corbella (UTM 31TCG86). Type material The type series is constituted by eight specimens in ethanol and 10 shells. The holotype is deposited in the Museu Valencià d’Història Natural (MVHN) with the code MVHN–120609AB00a (ethanol 70%); 10 paratypes (6 shells; 4 in ethanol) deposited in the MVHN with the code 120609AB00b; 2 paratypes in the Museu de Ciències Naturals de Barcelona (Zoologia, MZB) with the codes MZB 2009–4021 (1 shell) and MZB 2010–1152 (1 in ethanol); 2 paratypes (1 shell; 1 in ethanol) in the Museo Nacional de Ciencias Naturales of Madrid (MNCN) with the code MNCN 15.05/53565; 2 paratypes (1 shell; 1 in ethanol) in the Natuurhistorich Museum–Naturalis of Leiden (The Netherlands) with the code RMNH– MOL.125987; 1 paratype (shell) in the Senckenberg Museum of Frankfurt (Germany) with the code SMF–335206. Etymology Dedicated to Jordi Corbella Alonso, collector of the specimens. Common name Small snail of Navés Diagnosis Small size shell (6.5–7.9 mm), homogeneous pale grey, it sometimes has a brown spiral band around the last whorl with an internal rib near the aperture, externally visible as a white cretaceous transverse band; sometimes there are two of them. The male portion is very long and the flagellum is also long, but not as long as the penis and the epiphallus together. Bursa copulatrix is well defined and long in relation to its conduct. Shell (figs. 1–9): for the conchological description 18 specimens of the type series have been used. It is dextral, small in size, from depressed to almost lenticular, low spire, convex, more flattened above and less convex below, solid, opaque, shiny, formed by 4¾ to 5¼ whorls, convex, slow growth, regular and well–marked sutures of grayish–white with brownish apex, highlighting the rest of the shell. In the first whorl, brownish longitudinal stripes can be seen rather close together, and two shells show a continuous brownish line crosses the periphery of the last whorl and one to 1½ whorls from the suture of the penultimate whorl. There are also some small patches of the same tone scattered in an irregular manner at the top of the shell. The umbilical area is white, occasionally with small isolated patches, but without spiral lines. The end of the final whorl, near the shell aperture, goes down slightly towards the middle, helping to give this an oval morphology, a little flattened, with maximum size between 2.65 and 3.5 mm in width and from 2.2 and 2.75 mm in height, the opening being larger than 3.5 mm wide x 2.75 mm high, raised in the upper area,

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70 μm

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Figs. 1–8. Candidula corbellai n. sp.: 1–3. Holotype (nº120609AB00a) (Ø = 6.85 mm); 4. Paratype (Ø = 6.9 mm). 5–7. Protoconch of other paratype: 5. General view; 6–7. Sculpture protoconch details; 8. Teleoconch sculpture. Figs. 1–8. Candidula corbellai sp. n.: 1–3. Holotipo (nº120609AB00a) (Ø = 6,85 mm); 4. Paratipo (Ø = 6,9 mm). 5–7. Protoconcha de otro paratipo: 5. Vista general; 6–7. Detalles de la ornamentación; 8. Ornamentación de la teloconcha.

with the peristome interrupted and sharp. The aperture has an internal callous rib on all specimens except for two youngest that can give the shell surface an easily detectable, well–marked white colour. The internal callous rib can occur several times on the shell, seen externally as white transverse bands. The umbilicus is small, deep, somewhat eccentric, and narrow, from 1.15 to 1.5 mm in diameter, with no completely visi-

ble internal spire, barely hidden by the reflection of the peristome. The protoconch is always brownish, consisting of 1⅛ to 1 ½ whorl spires, from 0.7 to 1.0 mm in width, and with a micro–sculpture formed by parallel spiral strips that are visible in the suture zone and with slightly marked ribs, also more or less parallel, possibly giving it a reticulate appearance (fig. 7). This protoconch shows malleolated marks (fig. 5) but there

are no hairs or signs to indicate their presence. The teleoconch has a slight costulation, regular in the first whorls, of wrinkled aspect, but becoming more apparent and irregular in the following whorls and seeming more pronounced in the last whorl that becomes quite angled. Ribs are also visible in the umbilical area; thinner but are well marked. The dimensions of all the shells of the type series range from 6.2 to 7.9 mm in diameter and from 3.7 to 4.7 mm in height. The holotype is 6.85 mm in diameter and 4.3 mm in height. Reproductive system (figs. 9–15): the reproductive system of three specimens was studied, including the holotype (figs. 9–10). The reproductive system pattern is similar to other species of the genus Candidula. The retractor muscle of the right ommatophore is free of penis and vagina. The penis retractor muscle is inserted into the diaphragm. The male portion reaches a maximum length of between 11.15 mm (holotype) and 13.0 mm. The penis has a length of between 3.35 (holotype) and 4.75 mm, the epiphallus between 3.25 (holotype) and 3.7 mm and the flagellum between 4.4 and 4.55 mm (holotype). The penis has a short penial papilla in its interior, measuring 1.05 mm in length, with subapical opening. The vagina is long, between 3.5 (holotype) and 4.75 mm, and the free oviduct is between 0.8 (holotype) and 1.0 mm. The length of the duct of the bursa copulatrix is shorter than the bursa copulatrix in all specimens, between 1.85 (holotype) and 2.6 mm. The bursa copulatrix, which is well defined, is elongated and somewhat widened in the distal zone, with maximum dimensions between 2.35 x 1.1 mm and 3.35 x 1.15 mm. The proximal portion of the larger dart sac, which lies outside the vagina, is short, measuring between 1.05 and 1.5 mm. The dart is curved. It is 4 mm long and has a circular section ending in a point, without edges. The eight glandulae mucosae are inserted around the vagina, united in 4 trunks that branch from the basal zone, two of which are located laterally, on opposite sides, while the other two are very close, in their middle zone. The atrium measures between 0.9 and 1.25 mm. Other characters (figs. 16–21): the body is whitish except in the most anterior dorsal zone where it is grey. The odontognate jaw has a few ribs in the central area (fig. 16). The radula of the holotype, 1.6 mm long and 0.55 mm wide, consists of 119 rows. The radular formula is: 6M + 14L + C + 14 + 6M. Geographical distribution, habitat and conservation Candidula corbellai n. sp. it is only known from a single locality, Sierra de Busa (Navés, Lleida) (fig. 38). It lives at a high altitude in steppe calcareous meadows, under stones and adhering to the base and /or of the stems of vegetation such as Santolina chamaecyparissus, Genista sp. and several gramineous plants. Due to ongoing taxonomical confusion among the several species of hygromiids in Catalonia, the limits of their geographical distribution need to be determined before some kind of protection for this new species can be proposed.

Martínez–Ortí

Discussion Checking only the shell could lead to errors in differentiating between different genus of hygromiids, such as Helicella Férussac, 1821 or Xerocrassa Monterosato, 1892, also present in Catalonia, or other Candidula species living in areas relatively close to the the Iberian peninsula. To avoid such error, it is necessary to study the reproductive system. The species of Candidula found in the Iberian peninsula are the following: C. arganica (Servain, 1880), C. camporroblensis (Fez, 1944), C. gigaxii (Pfeiffer, 1848), C. intersecta (Poiret, 1801), C. najerensis (Ortiz de Zárate, 1950), C. rocandioi (Ortiz de Zárate, 1950) and C. unifasciata (Poiret, 1801). Of all these species, those that are most similar to C. corbellai n. sp. in terms of reproductive anatomical characters are C. camporroblensis and C. rocandioi. Both have a long flagellum, among other characters, that differentiate them from the others whose flagellum is short (Ortiz de Zárate, 1950; Manga, 1983; Puente, 1994; Gittenberger, 1993a, 1993b). They also have a shell of approximately the same dimensions as C. corbellai n. sp., although this latter has a characteristic gray–white colour while the others show a more coloured pattern, with dark brown spiral bands on both sides and brownish spots all over, especially on the apical zone (figs. 22–24). The maximum diameter of the shell of C. rocandioi varies between 5.9 and 7.5 mm and height between 3.5 and 4 mm (Ortiz de Zárate, 1950), very similar to the new taxon. However, the umbilicus of the new taxon is wider and less deep, between 1.6 to 1.7 mm; it occupies about 1/3 or 1/4 of the umbilical area, is not eccentric, and shows all the internal spires in C. rocandioi. However, in C. corbellai n. sp. the umbilicus is smaller and narrower, and may not be seen around the inner spires, occupying 1/5 or 1/6 part of the total width of the shell. Although Ortiz de Zárate (1950) does not indicate the presence of hair or protoconch or teleoconch in juvenile specimens, these characteristics were later reported by Manga (1983) and Ortiz de Zárate (1991). No hairs were found in the paralectotypes of C. rocandioi deposited in MVHN (Martínez–Ortí & Uribe, 2008), although their marks are visible in the protoconch (figs 26–27), which can be confused with the malleolated marks present in C. corbellai n. sp. or C. camporroblensis. The aperture is larger in C. corbellai n. sp., whose dimensions vary between 2.2 mm high by 2.65 mm wide and 2.75 mm high by 3.5 mm wide, whereas in C. rocandioi they vary between 1.7 mm high by 2.2 mm wide and 1.85 mm high by 2.5 mm wide. The top of the aperture in C. rocandioi is somewhat more bowed downwards, descending well over the midline of the last whorl, and it has a more marked keel than in C. corbellai n. sp. Based on data about the reproductive system provided by Ortiz de Zárate (1950) and Aparicio (1982), C. rocandioi can be distinguished from C. corbellai n. sp. for its flagellum that is 1/3 of the total length of the epiphallus and the penis together, while in C. corbellai n. sp. the flagellum is double the third of the length of the penis and epiphallus together. The

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Figs. 9–15. Reproductive system of Candidula corbellai n. sp.: 9–10. Holotype; 11. Dart of a paratype; 12. Point detail of dart of one paratype; 13. Penis distal portion; 14. Papilla penial; 15. Genitalia of a paratype. Figs. 9–15. Aparato reproductor de Candidula corbellai sp. n.: 9–10. Holotipo; 11. Dardo de un paratipo; 12. Detalle de la punta del dardo de un paratipo; 13. Porción distal del pene; 14. Papila penial; 15. Genitalia del paratipo.

epiphallus in C. rocandioi is very long, almost twice the flagellum, while in C. corbellai n. sp. the epiphallus is always smaller than the flagellum. Furthermore, in C. rocandioi the penis and epiphallus together can reach 13.5 mm while in C. corbellai n. sp. they do not exceed 7.45 mm. Another character mentioned by Ortiz de Zárate (1950: fig. 5), Aparicio (1982: fig. 6), Manga (1983) and Puente (1994: fig. 63) is that in C. rocandioi there is no clear difference between the end of the bursa copulatrix duct and the start of the bursa copulatrix itself due to the similar thickness of both. In C. corbellai n. sp. there is a clear separation between the two structures (figs. 9–10). Furthermore, the whole duct and bursa copulatrix together reaches 11.5 mm, while this is almost only half as long in C. corbellai n. sp., 5.95 mm. In respect to C. camporroblensis, species with great similarities with C. corbellai n. sp., Fez (1944)

describes shells of 4.0 to 5.0 mm in diameter and 3.0 mm in height, Faci (1991) reports maximum dimensions of 5.25 mm in diameter and 3.0 mm in height for Aragonese populations, Martínez–Ortí (1999) and Martínez–Ortí et al. (2000) mention dimensions of 5.25 mm in diameter and 3.5 mm in height for Valencian populations, and finally, Bragado et al. (2010) describe a diameter of 4.7 and 6.5 mm for the populations in Castilla–La Mancha. In the 18 shells of the new species only one specimen showed a diameter (6.2 mm) less than the maximum diameter known for the species (7.9 mm), while two specimens –somewhat fragmented– showed a larger diameter than that known for C. camporroblensis, generally more than 7.0 mm and nearly 8.0 mm C. corbellai n. sp. In C. camporroblensis the ombilicus has a width of between 0.9 mm (Faci, 1991) and 1.0 mm (Fez, 1944) while in C. corbellai n. sp. it is somewhat wider, up to 1.5 mm.

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17 16

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Figs. 16–21. Jaw and radula of C. corbellai n. sp.: 16. Jaw; 17. General view of the central and first lateral teeth; 18. Central tooth view; 19. Detail of the transition area between the last lateral teeth and the first marginal teeth; 20. Last marginal teeth; 21. Detail of the marginal teeth. Figs. 16–21. Mandíbula y rádula de C. corbellai sp. n.: 16. Mandíbula; 17. Vista general de los dientes central y primeros laterales; 18. Vista del diente central; 19. Detalle de la zona de transición de los últimos dientes laterales y primeros dientes marginales; 20. Últimos dientes marginales; 21. Detalle de los dientes marginales.

The colour of the apex in C. corbellai n. sp. is dark and it stands out from the rest of the shell (fig. 1). In C. camporroblensis (fig. 18), on the other hand, the colouring is less varied and is similar to the rest of the shell; in two of 40 shells examined, however, the apex was rather more colourful as was the rest of the shell in both cases. This was not observed in any C. corbellai shells which were all a whitish grey.

The teleoconch in C. corbellai n. sp. is less wrinkled (figs. 1–5, 8) than in C. camporroblensis (figs. 18–21, 25–26) and than in C. rocandioi (fig. 28). With regard to the reproductive system the differences between the two taxa are also significant. Aparicio (1982) indicates that the main characteristics of the genitalia of C. camporroblensis are the ratio of the length of the bursa copulatrix duct, which is

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Figs. 22–28. Paralectotypes of C. rocandioi (MVHN nº597): 22–24. Shell (Ø = 6.2 mm); 25. Protoconch; 26–27. Protoconch sculpture details; 28. Teleoconch sculpture of the first laps. (The photographs 25–28 were taken without a gold–palladium layer.) Figs. 22–28. Paralectotipos de C. rocandioi (MVHN nº597): 22–24. Concha (Ø = 6,2 mm); 25. Protoconcha; 26–27. Detalles de la ornamentación; 28. Ornamentación de la teloconcha. (Las fotografías 25–28 se han realizado sin haber sido cubiertas por una capa de oro–paladio.)

twice the length of the bursa, and the flagellum has approximately the same length or is slightly longer than half the penis and epiphallus together. Regarding the ratio of the length of the bursa copulatrix duct in C. corbellai n. sp., this never exceeds the length of the duct to the bursa copulatrix, rather the contrary, the bursa copulatrix is always a little longer than the duct, allowing us to differentiate it from C. camporroblensis. Moreover, the morphology of the bursa in C. corbellai n. sp. is elongated (1.1 to 1.15 mm in width), while

in C. camporroblensis it is ovoid (Ortiz de Zárate, 1950; Aparicio, 1982; Faci, 1991; Martínez–Ortí, 1999; Martínez–Ortí et al., 2000). Respect to the length of the flagellum, that of C. corbellai n. sp. is longer, between 4.4 and 4.55 mm, while that of C. camporroblensis is shorter, reaching a maximum 3.8 mm. Despite being longer, in C. corbellai n. sp. the length of the flagellum is always less than the whole of the length of the penis and epiphallus, whereas in C. camporroblensis it is equal to or slightly greater than the other two organs

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Figs. 29–37. Candidula camporroblensis: 29–31. Lectotype (MVHN nº321A) (Ø = 5.1 mm); 32. Protoconch of a paralectotype; 33–35. Protoconch sculpture details; 36–37. Teleoconch sculpture details. Figs. 29–37. Candidula camporroblensis: 29–31. Lectotipo (MVHN nº321A) (Ø = 5,1 mm); 32. Protoconcha de un paralectotipo; 33–35. Detalles de la protoconcha; 36–37. Detalles de la ornamentación de la teloconcha.

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T S

Fig. 38. Geographical distribution of Candidula corbellai n. sp. (locus typicus, circle), C. camporroblensis (area, grey coloured; locus typicus, large dot; others localities, small dots) and C. rocandioi (area, inside the line; locus typicus, square; other localities, small squares) in the Iberian peninsula. Fig. 38. Distribución geográfica de Candidula corbellai sp. n. (locus typicus, círculo), C. camporroblensis (área, sombreada en gris; locus typicus, punto grande; otras localidades, puntos pequeños) y C. rocandioi (área, delimitada por la línea; locus typicus, cuadrado; otras localidades, cuadrados pequeños) en la península Ibérica.

together. Other characters of the genitalia clearly differentiate the two taxa. The male portion of C. corbellai n. sp. is longer than in C. camporroblensis; while the former is between 11.35 and 13.0 mm, the latter only reaches 7.3 mm (Ortiz de Zárate, 1950), 7.95 mm for 17 specimens (Martínez–Ortí et al., 2000), 8.2 mm (Aparicio, 1982) and 7.9 mm (Faci, 1991). In C. camporroblensis the penis is between 1.45 and 1.7 mm long (Martínez–Ortí et al., 2000); Faci (1991) indicates 3.2 mm for a single specimen and Aparicio (1982) reports length between 1.7 and 2.5 mm, whereas in C. corbellai n. sp. it is longer size, ranging between 3.35 mm in the holotype and 4.75 mm in one of the two paratypes studied. The total length of the penis and epiphallus together for C. camporroblensis ranges between 5.0 mm (Ortiz de Zárate, 1950), 5.4 mm (Aparicio, 1986), 4.3 mm (Faci, 1991) and 3.95 (Martínez–Ortí et al., 2000), while in C. corbellai n. sp. it ranges between 6.55 mm in the holotype and 8.45 mm in one of the paratypes. The dart in C. corbellai n. sp. is longer

than in C. camporroblensis, reaching 4 mm, while the latter ranges between 1.8 and 3.30 mm. The dart sac, defined as the part of the free sac outside the vagina, is shorter in C. corbellai n. sp., ranging between 1.05 and 1.5 mm, while Faci (1991) points to C. camporroblensis 2.5 mm and Aparicio (1982) between 3.1 and 4.4 mm. The vagina is short in C. camporroblensis (1 mm in Faci, 1991) while in C. corbellai n. sp. it is longer, from 3.5 mm of the holotype to 4.75 mm in one of the paratypes examined. About the distribution of the three species, C. corbellai n. sp. is known only from one locality, well away from the distribution area of C. camporroblensis and C. rocandioi, which are more abundant in the Iberian peninsula, coexisting even in the same geographical area (fig. 38). C. camporroblensis has been cited in Catalonia (Tarragona) since the 1950s (Altimira, 1959; Bech, 1990). Puente (1994) notes that the description and the citations of these authors do not correspond to this species and should be considered erroneous.

New samples of C. camporroblensis in these localities should validate or not its presence in Catalonia after study of its reproductive system. C. camporroblensis is present in a large area, occupying almost the entire Iberian System, from Soria to Cuenca in Castilla–La Mancha (Altonaga et al., 1994; Puente, 1994; Talaván & Talaván, 2004; Bragado et al., 2010) provinces of Valencia and Castellón in Valencian Community (Martínez–Ortí, 1999; Martínez–Ortí et al., 2000), the Community of Aragón with the provinces of Zaragoza and Teruel (Faci, 1991; Puente, 1994) (fig. 38), while in Catalonia, Puente (1994) considers erroneous the quotes of Altimira (1959) and Bech (1990) in the province of Tarragona, as noted above. Alonso (1975) also indicated that C. camporroblensis could be found in Andalusia, but Puente (1994) has questioned this. Indeed, Ruiz et al. (2006) have not included this species in their guide to land snails of Andalusia. C. rocandioi, which has never been cited in Catalonia or in its neighboring provinces, has been found in León, north of Palencia, Burgos and Soria in Castilla–León, La Rioja, Cuenca and Guadalajara in Castilla–La Mancha and in Aragón where it is known from a single locality (Manga, 1993; Altonaga et al., 1994; Puente, 1994; Bragado et al., 2010). Acknowledgements We thank the Sección de Microscopía Electrónica del S.C.S.I.E. of the Universitat de València for their help using the SEM Hitachi S–4100. This work received financial support from the project of the Spanish Ministerio de Investigación, Ciencia e Innovación (CGL2008–01131/BOS). References Alonso, M. R., 1975. Fauna malacológica terrestre de la depresión de Granada (España). II. El género Helicella Férussac, 1821. Cuadernos de Ciencias Biológicas, 4(1): 11–28. Altimira, C., 1959. Contribución al conocimiento de la fauna malacológica de la provincia de Tarragona. Miscel.lània Zoològica, 1: 89–95. Altonaga, K., Gómez, B., Martín, R., Prieto, C. E., Puente, A. I. & Rallo, A., 1994. Estudio faunístico y biogeográfico de los Moluscos terrestres del norte de la Península Ibérica. Parlamento Vasco, Vitoria. Aparicio, M. T., 1982. Observations on the antomy of some Helicidae from Central Spain. Malacologia, 22(1–2): 621–626. Bech, M., 1990. Fauna malacològica de Cataluña. Mol·luscs terrestres i d’aigua dolça. Treballs de la Institució Catalana d’Història Natural, 12: 1–229. Bragado, M. D., Araujo, R. & Aparicio, M. T., 2010. Atlas y Libro Rojo de los Moluscos de Castilla–La Mancha. Organismo Autónomo Espacios Naturales

Martínez–Ortí

de Castilla–La Mancha, Junta de Comunidades de Castilla–La Mancha, Guadalajara. Faci, G., 1991. Contribución al conocimiento de diversos moluscos terrestres y su distribución en la Comunidad Autónoma Aragonesa. Ph. D. Thesis, Univ. de Zaragoza. Fez, S. de, 1944. Contribución a la malacología de la provincia de Valencia. Boletín de la Real Sociedad Española de Historia Natural, 42: 211–224. Gittenberger, E., 1993a. Digging in the graveyard of synonymy, in search of Portuguese species of Candidula Kobelt, 1871 (Mollusca: Gastropoda Pulmonata: Hygromiidae). Zoologische Mededelingen, 67: 283–293. – 1993b. On Trochoidea geyeri (Soós, 1926) and some conchologically similar taxa (Mollusca: Gastropoda Pulmonata: Hygromiidae). Zoologische Mededelingen, 67: 303–320. Hausford, B., 1988. Zur Kenntnis der systematischen Bezieehungen einiger Taxa der Helicellinae Ihering, 1909 (Gastropoda, Hygromiidae). Archiv für Molluskenkunde, 119: 9–37. Holyoak, G. A. & Holyoak, D. T., 2010. A new species of Candidula (Gastropoda, Hygromiidae) from central Portugal. Iberus, 28(1): 67–72. Manga, Y., 1983. Los Helicidae (Gastropoda, Pulmonata) de la provincia de León. Ed. Diputación Provincial de León. Institución 'Fray Bernardino de Sahagún'. Martínez–Ortí, A., 1999. Moluscos terrestres testáceos de la Comunidad Valenciana. Ph. D. Thesis, Univ. of València. Martínez–Ortí, A., Faci, G. & Robles, F., 2000. Taxonomical revision of Trochoidea (Xerocrassa) llopisi Gasull, 1891 (Gastropoda, Pulmonata, Hygromiidae, Geomitrinae), from the province of Castellón, Spain. Basteria, 64: 7–14. Martínez–Ortí, A. & Uribe, F., 2008. Los ejemplares tipo de las colecciones malacológicas del Museu de Ciències Naturals de Barcelona y del Museu Valencià d’Història Natural. Arxius de Miscel·lànea Zoològica, 6: 1–156. Ortiz de Zárate, A., 1950. Observaciones anatómicas y posición sistemática de varios helícidos españoles. Boletín de la Real Sociedad Española de Historia Natural, 48: 21–85. – 1991. Descripción de los moluscos terrestres del Valle del Najerilla. Gobierno de La Rioja, Consejería de Educación, Cultura y Deportes, Logroño. Puente, A. I., 1994. Estudio taxonómico y biogeográfico de la superfamilia Helicoidea Rafinesque, 1815 (Gastropoda: Pulmonata: Stylommatophora) de la Península Ibérica e Islas Baleares. Ph. D. Thesis, Univ. of País Vasco. Ruiz, A., Cárcaba, A., Porras, A. & Arrébola, J. R., 2006. Caracoles Terrestres de Andalucía. Guía y manual de identificación. Junta de Andalucía, Fundación Gypaetus, Sevilla. Talaván Gómez, J. & Talaván Serna, P. J., 2004. Contribución a la malacología de la Sierra de Cuenca. Spira, 1(4): 11–21.

Animal Biodiversity and Conservation 34.1 (2011)

Bird community patterns in sub–Mediterranean pastures: the effects of shrub cover and grazing intensity S. C. Nikolov, D. A. Demerdzhiev, G. S. Popgeorgiev & D. G. Plachiyski

Nikolov, S. C., Demerdzhiev, D. A., Popgeorgiev, G. S. & Plachiyski, D. G., 2011. Bird community patterns in sub–Mediterranean pastures: the effects of shrub cover and grazing intensity. Animal Biodiversity and Conservation, 34.1: 11–21. Abstract Bird community patterns in sub–Mediterranean pastures: the effects of shrub cover and grazing intensity.— Shrubs are widely considered a threat to grassland biodiversity. We investigated the effects of shrub cover and grazing intensity on bird communities in sub–Mediterranean pastures in Bulgaria. The point–count method was used on 80 plots in open (< 10% shrub cover) and shrubby (approx. 20% cover) pastures under either intensive or extensive management (grazing intensity) from 2008 to 2009. We recorded a total of 1,956 observations of birds from 53 species. Main environmental gradients accounting for the bird community pattern were related to vegetation succession and land productivity. Bird species richness was higher in shrubby pastures than in open sites, while no effect was found in respect to total bird abundance. Bird species diversity (i.e. H’ index) was highest in extensive shrubby pastures. Shrubland specialists were positively affected by shrub cover and extensive management of pastures while grassland and woodland specialists showed no significant response to these factors. We conclude that a small proportion of shrubs within pastures may be beneficial for farmland birds and sustainable management of pastures could be achieved by greater flexibility of national agri–environmental schemes. Key words: Agri–environmental scheme, Farmland birds, Grassland management, Semi–natural habitats, Shrubby vegetation. Resumen Patrones de las comunidades de aves en los pastos submediterráneos: el efecto de la cubierta arbustiva y la intensidad de pastoreo.— Se suele considerar a los arbustos como una amenaza a la biodiversidad de los pastos. Investigamos los efectos de la cubierta arbustiva y la intensidad del pastoreo sobre las comunidades de aves en los pastos submediterráneos de Bulgaria. Se utilizó el método de estaciones de escucha en 80 puntos de registro en pastos abiertos (cubierta arbustiva < 10%) y arbustivos (aproximadamente un 20% de la superficie cubierta), con una gestión de pastoreo tanto intensiva como extensiva desde 2008 a 2009. Registramos un total de 1.956 observaciones de aves pertenecientes a 53 especies distintas. Los gradientes ambientales principales responsables de los patrones de las comunidades de aves se relacionaron con la sucesión de la vegetación y la productividad de la tierra. La riqueza de especies de aves era mayor en los pastos arbustivos que en los lugares abiertos, aunque no se observó efecto alguno con respecto a la abundancia total de aves. La mayor diversidad de especies de aves (índice H’) se daba en los pastos arbustivos con gestión extensiva. Los especialistas en zonas arbustivas se veían afectados positivamente por la cubierta arbustiva y la gestión extensiva de los pastos, mientras que los especialistas de praderas y bosques no presentaron ninguna respuesta positiva a dichos factores. Nuestra conclusión es que una pequeña proporción de arbustos dentro de los pastos puede ser beneficiosa para las aves de tierras de labrantío, y la gestión sostenible de los pastos podría alcanzarse mediante una mayor flexibilidad de los esquemas agroambientales nacionales. Palabras clave: Esquema agroambiental, Aves de labrantío, Gestión de prados, Hábitats seminaturales, Vegetación arbustiva. (Received: 14 VI 10; Conditional acceptance: 6 IX 10; Final acceptance: 10 II 11) Stoyan C. Nikolov, Inst. of Biodiversity and Ecosystem Research–BAS, 2 Yurii Gagarin Str., 1113 Sofia, Bulgaria.– Dimitar A. Demerdzhiev, Georgi S. Popgeorgiev & Dimitar G. Plachiyski, Bulgarian Society for the Protection of Birds–BirdLife Bulgaria, 27A P. Todorov Str., 4000 Plovdiv, Bulgaria. Corresponding author: S. C. Nikolov. E–mail: nikolov100yan@abv.bg ISSN: 1578–665X

Nikolov et al.

Introduction

Methods

Semi–natural grasslands are among the high nature value farming systems of conservation concern as they are biodiversity–rich and provide agricultural benefits through stock grazing and haymaking (Henle et al., 2008). These habitats were created under traditional agricultural practices but currently, due to agricultural intensification or land abandonment, they have now become significantly reduced in area in northern European (Pärt & Söderström, 1999), western European (Tucker & Heath, 1994; Fuller et al., 1995) and eastern European countries (Meshinev et al., 2005). In the European Union, semi–natural grasslands are under the regulation of the Common Agricultural Policy (CAP) which was adopted in 1957 to increase agricultural production by ensuring sufficient food for all inhabitants and a fair standard of living for farmers (Verhulst et al., 2004). Implementation of the CAP resulted in a polarization of production areas by stimulating land use intensification in some areas (Donald et al., 2002) and leading to abandonment of other, marginally profitable areas (Bignal, 1998). It was found that intensification or abandonment of land management can greatly reduce biodiversity by threatening the survival of many species adapted to the diversity of structures and resources of high nature value farmlands (Sirami et al., 2007; Kleijn et al., 2009; Nikolov, 2010). Grassland bird populations for instance, declined sharply due to agricultural intensification in Europe over the past half century (Gregory et al., 2004; Donald et al., 2006). On the other hand, abandonment of land management benefits vegetation succession through the development of woody vegetation, providing benefits to shrubland and woodland birds whilst negatively affecting open–habitat specialists (Preiss et al., 1997; Suárez–Seoane et al., 2002; Pons et al., 2003; Verhulst et al., 2004). As a result, the development of shrubby and woody vegetation was considered a potential threat to grassland biodiversity, and the CAP strongly advised removal of these habitat features as a management recommendation (Boccaccio et al., 2009). In many countries (e.g. France, Sweden, Greece and Bulgaria), this measure was not tested but applied directly in the national agri–environmental schemes (Lefranc, 1997; Pärt & Söderström, 1999; Söderström et al., 2001; Kati & Sekercioglu, 2006; Nikolov, 2010). Indeed, in northern and southeastern Europe small covers of woody vegetation (≤ 20%) were found to increase avian species richness and diversity by favouring some threatened species (Pärt & Söderström, 1999; Söderström et al., 2001; Nikolov, 2010). The main objective of this study was to test the effects of shrub cover and grazing intensity on farmland birds in sub–Mediterranean pastures. The obtained results may serve as a basis for more sustainable and regionally–oriented pasture management aiming to maintain species rich, diverse bird communities.

Study area The study area covers the territory of the Special Protected Area (SPA) Besaparski Hills (147.7 km2) in southern Bulgaria (42° 7' N–24° 23' E; fig. 1). The landscape represents sub–Mediterranean limestone hills with an average altitude of 350 m a.s.l. (ranging from 184 m a.s.l. to 536 m a.s.l.) (Demerdzhiev, 2007). Most of the area is covered by arable land (about 50% of the territory) and by dry grasslands with some shrub heath (about 33% of the territory) and the rest of the territory is covered by vineyards and orchards (6%), wetlands (3%), stone pits (3%), urban areas and roads (3%) and small forests (1%). Grasslands are not fertilized and most of them are used for pastures (mainly for sheep and cattle). Study design Based on a digital map of the area (Bulgarian Society for the Protection of Birds, unpublished data), a total of 80 point–count stations were equally distributed and located randomly within two categories of pastures (open pastures with up to 10% shrub cover and shrubby pastures with more than 10% shrub cover) with the restriction that any two adjacent point–count stations should be a minimum of 250 m apart (Ralph et al., 1995). All study plots with difficult accessibility to the field were replaced using a second random selection. As a result, an aggregation of study plots in the eastern part of the study area appeared, but as the study plots were equally distributed between the studied pastureland categories (25 vs. 25 study plots in shrubby and open pastures, respectively) within the area of aggregation, we assumed that our data were not biased by spatial autocorrelation effects. After a pilot visit to the study area, we found that 41 point–count stations were located within open pastures and 39 in shrubby pastures. Supplementary data on grazing intensity within the studied areas was collected from the local agricultural authorities and studied plots were classified according to their grazing regime as intensively grazed (0.8 AU ha–1; n = 30 study plots) and extensively grazed pastures (0.2 AU ha–1; n = 50 study plots). Finally, we used 31 study plots in open and extensive pastures, 10 in open and intensive pastures, 19 in shrubby and extensive pastures and 20 in shrubby and intensive pastures. Fieldwork was carried out during the breeding seasons of 2008 and 2009. Birds were sampled twice per year (in May and June), in the mornings (6:00– 10:00 a.m.), under appropriate weather conditions and by the same observer (D. D.). The point count method (Gibbons & Gregory, 2006) was applied, with a counting period of 5 min and a radial distance of 100 m. All birds seen or heard were recorded. Individuals simply flying over the point–count stations and not foraging in flight were excluded from the analysis (Batáry et al., 2007). To investigate how different ecological groups of birds respond to vegetation composition within pastures in respect to their habitat specialization we

Animal Biodiversity and Conservation 34.1 (2011)

N Bulgaria

1.5

6 km Krichim

SPA boundaries Shrubby pastures Open pastures

Study plots in shrubby pastures Study plots in open pastures Town

Fig. 1. Location of the study area (SPA Besaparski Hills) in Bulgaria and distribution of point–count stations within the semi–natural grasslands. Fig. 1. Localización del area de estudio (área protegida especial de las colinas de Besaparski) en Bulgaria y distribución de las estaciones de escucha en el interior de los prados seminaturales.

classified birds as grassland, shrubland or woodland specialists, following Iankov (2007). The conservation status of birds at the European and national levels was described following BirdLife International (2004) and Spassov (2007), respectively (appendix). Data on habitat composition were collected within a radius of 50 m centred on each point–count station. The relative cover of rocks and stones, arable land and vegetation layers was estimated visually and recorded in percentages (%). The following vegetation layers were recognized in the field: (1) grass, consisted mainly of Medicago spp., Trifolium spp., Sideritis montana, Chrysopogon gryllus, Dichnthium ischaemum, Eryngium campestre and Stipa capillata; (2) shrubs, consisted of woody vegetation up to 2 m height and dominated mainly by Paliurus spina–christi, Rubus sp., Jasminum fruticans, Juniperus oxycedrus; and (3) trees, consisted of woody vegetation above 2 m height and dominated mainly by Quercus pubescens. Elevation was recorded using a Global Positioning Systems unit (Etrex Summit). Studied pasture cat-

egories differed significantly only in their grass cover and shrub cover (table 1). Data analyses Bird data were square root transformed and habitat variables were arcsine transformed to approach normal distributions. For comparisons of environmental variables between open and shrubby pastures, t–test for unpaired samples was performed using STATISTICA version 7.0 software package (StatSoft, 2004). To analyze bird species richness and overall bird abundance we used the mean values of the maximum numbers of species and individuals recorded at each point–count station during both visits in both years. Bird species diversity was calculated using the Shannon–Wiener diversity index H’. Relationships between bird species and habitat characteristics were determined by Canonical Correspondence Analysis (CCA) computed in CANOCO 4.5 software (Ter Braak & Smilauer, 2002). Length of

Nikolov et al.

Table 1. Habitat characteristics (mean ± SE) and their comparisons (t–test for independent samples; StatSoft, 2004) in shrubby and open pastures in SPA Besaparski Hills, S Bulgaria: N. Sample size; * Significant P–values are in bold. Table 1. Características del hábitat (media ± EE) y sus comparaciones (test t para muestras independientes; StatSoft, 2004) en pastos arbustivos y abiertos del área protegida especial de las colinas de Besaparski, S de Bulgaria: N. Tamaño de la muestra; * Los valores significativos de P se dan en negrita. Environmental variables Altitude

Shrubby pastures

Open pastures

(N = 39)

(N = 41)

t78

326.49 ± 12.00

300.15 ± 10.57

1.65

0.103

Cover of grass

64.10 ± 2.82

85.82 ± 3.09

–6.35

< 0.001

Cover of shrubs

21.09 ± 1.75

2.12 ± 0.40

10.60

< 0.001

Cover of trees

1.92 ± 1.31

0.57 ± 0.30

1.03

0.305

Cover of rocks

8.72 ± 2.54

8.41 ± 2.66

0.09

0.931

Cover of arable land

2.76 ± 1.41

3.07 ± 1.75

–0.17

0.868

bird data gradient was checked by preliminary detrended correspondence analysis (DCA) and unimodal ordination was applied even though the gradient was relatively short (i.e. 2.82 for the first canonical axis), because this model better explained data variability and because the length of the gradient was close to the range for which both linear and unimodal methods work well (Lepš & Šmilauer, 2003). The Monte–Carlo permutation test was used to assess the statistical significance of canonical axes (Lepš & Šmilauer, 2003). The effects of shrub cover, grazing intensity and their interaction on birds at community and ecological group levels were analysed by General Linear Models (GLM) in STATISTICA version 7.0 software package (StatSoft, 2004). For each of the studied dependent variables (i.e. bird species richness, H’ diversity index and abundance) separate GLM was conducted, where pasture categories (in respect of shrub cover and grazing intensity) were categorical factors and studied habitat characteristics (see table 1) were continuous predictors. In the GLM, a Tukey HSD post–hoc test was used to determine significant differences between groups (α = 0.05). Results

Lullula arborea, showed a positive association with the open semi–natural grasslands (fig. 2). Shrubland and woodland species displayed the opposite pattern, being associated with semi–natural grasslands with an increased cover of shrubs and trees. However, these species were more widely spread along the gradient. Apart from grassland species, the open pastures also sheltered aerial feeders (e.g. European bee–eater Merops apiaster and barn swallow Hirundo rustica) or birds that forage in open landscapes (e.g. Spanish sparrows Passer hispaniolensis). The second environmental gradient was related to land productivity, (represented by land conversion: higher cover of arable lands at the one extremity of the gradient and the less productive rocky fields at the other extremity, fig. 2) and was represented by the second CCA axis accounting for 8.1% of bird data variability (species–environment correlation = 0.736). Most birds associated with this gradient were shrubland and woodland species: some of them benefited from arable mosaics (e.g. common cuckoo Cuculus canorus, common starling Sturnus vulgaris and European roller Coracias garrulus), while others were associated mainly with the less productive grasslands (e.g. ortolan bunting Emberiza hortulana, blackcap Sylvia atricapilla and European greenfinch Carduelis chloris).

Habitat composition and bird community pattern

Effects of shrubby vegetation

A total of 1,956 individuals from 53 species were recorded in the semi–natural grasslands of Besaparski Hills SPA (appendix). The main environmental gradient in the studied habitat was related to vegetation succession (representing the transition of open to shrubby pastures, fig. 2) and was represented by the first CCA axis accounting for 15.6% of bird data variability (species–environment correlation = 0.743). All grassland specialists, excluding the woodlark

Species richness was positively affected by the cover of shrubby vegetation while the effect of grazing intensity was not significant (table 2). Number of species ranged from 3–33 species/point–count station (mean ± SE = 7.56 ± 0.89, n = 39 point–count stations) in shrubby pastures and 1–15 species/ point–count station (mean ± SE = 4.95 ± 0.43, n = 41 point–count stations) in open pastures. Bird species diversity (i.e. H’ index) was influenced

Animal Biodiversity and Conservation 34.1 (2011)

0.8

Arable land

CucCan

AntCam

StuVul

Axis 2

BurOed

CorGar UpuEpo HirRus Altitude EmbCir HipPal MerApi PicPic TurMer OenIsa SylCur PasHis Shrubs LusMeg EmbMel LanCol Sward MelCal AlaArv Trees MilCal CalBra CarCar PasMon ApuApu OenOen GalCri LanSen CarChl SylAtr EmbHor

–0.6

LulArb

Rock –0.6

Axis 1

1.0

Bird species

∆

Grassland specialist ¨ Shrubland specialist

¡ Woodland specialists + Other species

Fig. 2. Two–dimensional ordination by CCA relating bird abundances to habitat characteristics in the sub–Mediterranean lowland pastures in southern Bulgaria. The first two canonical axes account for 23.9% of bird data variability and all axes are statistically significant (Monte–Carlo test based on 499 random permutations, F = 2.06, p = 0.02). Environmental variables are indicated by arrows. (Only brid species with fit > 5% in the model are shown; for bird species acronyms see the appendix.) Fig. 2. Ordenación bidimensional por análisis canónico de correspondencias (CCA) que relaciona las abudancias de aves con las características del hábitat en los pastos submediterráneos de tierras bajas del sur de Bulgaria. Los dos primeros ejes canónicos responden del 23,9% de la variabilidad de los datos de las aves, y todos los ejes son estadísticamente significativos (test de Monte–Carlo, basado en 499 permutaciones al azar, F = 2,06, p = 0,02). Las variables ambientales se indican por medio de flechas. (Sólo se muestran las especies de aves que se ajustan > 5% al modelo; para los acrónimos de las especies de aves, ver el apéndice.)

by the interaction between the effects of shrubby vegetation cover and grazing intensity (table 2), and H’ index had the highest values in extensive shrubby pastures (extensive shrubby pastures: mean H’ ± SE = 1.58 ± 0.005, n = 19; extensive open pastures: mean H’ ± SE = 1.55 ± 0.004, n = 31; intensive shrubby pastures: mean H’ ± SE = 1.56 ± 0.006, n = 20; intensive open pastures: mean H’ ± SE = 1.56 ± 0.009, n = 10). Total bird abundance showed no significant response to shrub cover or grazing intensity (table 2). Grassland and woodland birds did not show significant response to the cover of shrubs within pastures or grazing intensity, while the low proportion of shrub cover and extensive grazing of pastures were found

to increase the species richness and abundance of shrubland birds (table 3). Regarding conservation status, open and shrubby pastures sheltered similar numbers but different species from conservation priority. Of the 15 species associated with open pastures (left sector of the biplot in fig. 2), three are included in Annex 1 of the Directive on the conservation of wild birds (Directive 2009/147/EC), 10 have an unfavourable status in Europe and three are known to be decreasing in Bulgaria. Of the 17 species associated with shrubby pastures (right sector of the biplot in fig. 2), five are included in Annex 1 of the Directive 2009/147/ EC, nine have an unfavourable status in Europe and three are known to be decreasing in Bulgaria.

Nikolov et al.

Table 2. The effects of grazing intensity (GI), shrubby cover (SC) and their interaction (INT) on bird community parameters in semi–natural grasslands of SPA Besaparski Hills (GLM, StatSoft, 2004): * Significant P–values are in bold; ** Between–group comparisons were determined by applying Tukey HSD post–hoc test, only significant differences (P < 0.05) are shown. Table 2. Los efectos de la intensidad de pastoreo (GI), la cubierta arbustiva (SC) y su interacción (INT) con los parámetros de las comunidades de aves en las praderas seminaturales del área protegida especial de las colinas de Besaparski (modelos lineales generales, GLM, StatSoft, 2004): * Los valores de P significativos están en negrita; ** Las comparaciones entre grupos se determinaron aplicando el test HSD de Tukey post–hoc, sólo se muestran las diferencias significativas (P < 0,05). Parameter

F1, 76

P *

0.91

0.342

5.68

0.019

INT

0.98

0.326

< 0.01

0.703

2.0

0.167

INT

5.0

0.031

Species richness

Species diversity

0.13

0.11

Effect

Interpretation**

Total abundance

0.05

Shrubby pastures > open pastures

Extensive shrubby pastures > extensive open pastures

2.76

0.101

0.13

0.721

INT

0.21

0.885

Discussion Birds and shrubby vegetation cover Our results demonstrate that a small proportion of shrubby vegetation (ca. 20%) within semi–natural grasslands may increase the species richness of bird communities in the sub–Mediterranean pastures of southern Bulgaria. This finding is consistent with the results from dry pastures in northern Europe (Pärt & Söderström, 1999) and upland pastures in south– eastern Europe (Nikolov, 2010) where retention of 10–15% shrub cover within pastures is advised as beneficial for farmland birds dependent on shrubs. This phenomenon could be explained by the increased habitat complexity within shrubby pastures, which provides more varied resources to bird species for nesting, searching for food, displaying (Verhulst et al., 2004) or escaping from predators (Shaefer & Vogel, 2000). This finding contributes to the concept that habitat heterogeneity is a key predictor for species richness within farmlands (Benton et al., 2003; Kati et al., 2009). Most of the positive effects of shrubby vegetation upon the structure of bird communities could be attributed to shrubland birds. Often this is a post–factum effect observed after land abandonment and the resulting secondary succession of the vegetation (Preiss et al., 1997; Suárez–Seoane et al., 2002; Verhulst et al., 2004; but see Batáry et al., 2007). However,

a limited presence of shrubby vegetation (ca. 20% cover) had no significant effects on grassland specialists. This was not expected as the presence of shrubs and trees within pastures reduces the overall area of the prime habitat for this group of birds. In western and central Europe, grassland bird abundance was observed to decrease as a consequence of the reduction of open grassland habitats (Preiss et al., 1997; Brotons et al., 2005) and it has also been found that increase in habitat heterogeneity may suppress the abundance of grassland specialists (Batáry et al., 2007). Possible ecological mechanisms explaining this pattern include an increasing predation risk for some grassland specialists due to the high vegetation cover in the surroundings (Shaefer & Vogel, 2000) or increased nest predation (Suárez & Manrique, 1992). In our study, the lack of negative effects of shrubby vegetation cover on grassland specialists may be explained by the relatively low cover of this habitat feature within studied pastures (about 20% mean cover for shrubby pastures; see table 1). Conversion of pastures into arable land Although several species (e.g. common cuckoo, common starling and European roller) were positively affected by the presence of arable lands, this should not be misinterpreted as a good reason for the conversion of grassland habitats into arable lands. In our study, the only threatened species associated with

Animal Biodiversity and Conservation 34.1 (2011)

Table 3. The effects of grazing intensity (GI), shrub cover (SC) and their interaction (INT) on species richness and abundance of ecological groups of birds in respect of their habitat specialization in pastures of Besaparski Hills, Bulgaria (GLM, StatSoft, 2004): * Significant P–values are in bold; ** Between– group comparisons were determined by applying Tukey HSD post–hoc test; only significant differences (P < 0.05) are shown. Tabla 3. Efectos de la intensidad de pastoreo (GI), la cubierta arbustiva (SC) y su interacción (INT) con la riqueza de especies y la abundancia de grupos ecológicos de aves respecto a su especialización en cuanto al hábitat en los pastos de las colinas de Besaparski, Bulgaria (modelos lineales generales–GLM, StatSoft, 2004): * Los valores significativos de P se dan en negrita; ** Las comparaciones entre grupos se determinan por medio del test HSD de Tukey post–hoc; sólo se muestran las diferencias significativas (P < 0,05). Bird group

Parameter

Effect

F1,76

P *

Interpretation **

Grassland birds Species richness 0.03

0.79

0.377

0.33

0.565

INT

0.66

0.420

Abundance

2.41

0.125

1.73

0.192

INT

0.10

0.748

4.3

0.041

Extensive pastures > intensive pastures

9.26

0.003

Shrubby pastures > open pastures

INT

0.68

0.412

Abundance

11.14

0.001

Extensive pastures > intensive pastures

4.17

0.045

Shrubby pastures > open pastures

INT

0.66

0.418

Species richness

0.99

0.321

2.19

0.143

INT < 0.01

0.965

Abundance

1.91

0.171

0.37

0.544

INT

0.18

0.676

0.08

Shrubland birds Species richness 0.22

0.23

Woodland birds

arable lands was the European roller, but it is known that highly intensified agricultural practices could have deleterious effects on its populations (Avilés & Parejo, 2004). Furthermore, some grassland specialists, including calandra lark Melanocorypha calandra and short–toed lark Calandrella brachydactyla, which are from high conservation priority within Natura 2000 network, are negatively affected by the presence of arable lands in grassland–dominated landscapes (Brotons et al., 2005). Therefore, our results suggest that it may be possible to support small parcels of arable land as a part of the rural mosaic within SPAs,

but this practice should be adopted with caution and strictly controlled, as it is recognized as a major cause of the loss of the semi–natural grassland habitats (Robertson et al., 1990) and one of the main threats to the local avifauna (Demerdzhiev, 2007). Bird conservation in sub–Mediterranean pastures The CAP was implemented rapidly in many countries of the European Union, and most agri–environment schemes were applied without sufficient testing at national scales (Wrbka et al., 2008; Stoate et al., 2009;

Nikolov, 2010). It was expected that the fast process of CAP implementation and the resulting changes in agricultural land use (i.e. agricultural intensification and abandonment) would cause alterations to traditional extensive exploitation systems and structure of grassland habitats (Tucker & Evans, 1997). Some of these effects on birds have been investigated (e.g. Batáry et al., 2007; Herzon et al., 2008; Nagy et al., 2009), but they are regionally dependent and should not be directly extrapolated to represent other locations with different cultural, economical and landscape characteristics (Nikolov, 2010). For instance, in many countries the removal of shrubby vegetation from pastures was promoted as an agricultural practice by CAP (Boccaccio et al., 2009), being considered as a threat to grassland biodiversity (Preiss et al., 1997; Stefanović et al., 2008). However, several studies from northern and south–eastern Europe have provided sound evidence that the availability of shrubby vegetation within semi–natural grasslands may be beneficial for the local avifauna (Pärt & Söderström, 1999; Söderström et al., 2001; Kati & Sekercioglu, 2006; Nikolov, 2010). Our results support this statement, demonstrating that from a conservation viewpoint open and shrubby pastures benefit a similar number of species of high conservation priority. We found that the small proportion of shrubby vegetation within pastures should not be considered as a threat, but as a potential factor increasing the conservation value of the protected area through the addition of some non–grassland threatened species to the existing typical grassland avifauna. Therefore, a possible way to counteract the negative effects of CAP on avian diversity in sub–Mediterranean areas could be at the level of national agri–environmental schemes (Verhulst et al., 2004) by providing higher flexibility of national standards at the regional scale (Wrbka et al., 2008). Particularly, regarding protected areas in the ecological networks this could be done by elaboration of management plans and zoning of agricultural activities at local scale. Finally, to ensure the effective long–term conservation of birds that are dependent on pastoral landscapes, it is crucial to assess the potential for resulting conflicts between intended outcomes for farmers and biodiversity (Henle et al., 2008). Therefore, we strongly advise further investigation into the stock holder’s production losses in relation to the availability of shrubs within pastures, and the potential opportunities for the compensation of these losses. Acknowledgements The study was funded by the project 'Conservation of globally important biodiversity in high nature value semi–natural grasslands through support for the traditional local economy' supported by the UNDP (Project No. 43595) and GEF (Project ID 2730) and executed by the Bulgarian Society for the Protection of Birds / BirdLife Bulgaria. The authors are grateful to Kiril Metodiev for providing the floristic data and to Hristo Ivanov for assistance with bird counts. The

Nikolov et al.

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Appendix. List of bird species found in pastures of SPA Besaparski Hills. Ecological groups (EG) of birds according to their habitat specialization (Iankov, 2007): G. Grassland birds; S. Shrubland birds; W. Woodland birds; O. Other birds. The status in Europe and Bulgaria follow BirdLife International (2004) and Spassov (2007), respectively. Apéndice. Lista de especies de aves encontradas en los pastos del área de protección especial (SPA) de las colinas de Besaparski. Grupos ecológicos (EG) de aves en función de su especialización en cuanto al hábitat (Iankov, 2007): G. Aves de prados; S. Aves de zonas arbustivas; W. Aves de zonas arboladas; O. Otras aves. El estatus en Europa y en Bulgaria según BirdLife International (2004) y Spassov (2007), respectivamente. Acronym

Europe

Eurasian Hobby

Species Falco subbuteo

FalSub

Secure

Bulgaria

Common Kestrel

F. tinnunculus

FalTin

Declining

Uncertain

Common Quail

Coturnix coturnix

CotCot

Depleted

Decreasing

Stone–curlew

Burhinus oedicnemus

BurOed

Vulnerable

Common Cuckoo

Cuculus canorus

CucCan

Secure

Uncertain

Common Swift

Apus apus

ApuApu

Secure

Uncertain

Eurasian Hoopoe

Upupa epops

UpuEpo

Declining

Decreasing

European Bee–eater

Merops apiaster

MerApi

Depleted

European Roller

Coracias garrulus

CorGar

Vulnerable

Green Woodpecker

Picus viridis

PicVir

Depleted

Great Spotted Woodpecker Dendrocopos major

DenMaj

Secure

Uncertain

Common Skylark

Alauda arvensis

AlaArv

Depleted

Decreasing

Crested Lark

Galerida cristata

GalCri

Depleted

Decreasing

Woodlark

Lullua arborea

LulArb

Depleted

Greater Short–toed Lark

Calandrella brachydactyla CalBra

Declining

Calandra Lark

Melanocorypha calandra

MelCal

Declining

Barn Swallow

Hirundo rustica

HirRus

Depleted

Red–rumped Swallow

H. daurica

HirDau

Secure

Common House Martin

Delichon urbica

DelUrb

Declining

Tawny Pipit

Anthus campestris

AntCam

Declining

Common Nightingale

Luscinia megarhynchos

LusMeg

Secure

Northern Wheatear

Oenanthe oenanthe

OenOen

Declining

Isabelline Wheatear

O. isabellina

OenIsa

Secure

Black–eared Wheatear

O. hispanica

OenHis

Depleted

Song Thrush

Turdus philomelos

TurPhi

Secure

Common Blackbird

T. merula

TurMer

Secure

Blackcap

Sylvia atricapilla

SylAtr

Secure

Lesser Whitethroat

S. curruca

SylCur

Secure

Common Whitethroat

S. communis

SylCom

Secure

Olivaceous Warbler

Hippolais pallida

HipPal

Secure

Chiffchaff

Phylloscopus collybita

PhyCol

Secure

Great Tit

Parus major

ParMaj

Secure

Long–tailed Tit

Aegithalos caudatus

AegCau

Secure

Red–backed Shrike

Lanius collurio

LanCol

Depleted

Uncertain Uncertain Uncertain

Uncertain Uncertain Increasing

Uncertain Decreasing

Animal Biodiversity and Conservation 34.1 (2011)

Appendix. (Cont.)

Species

Acronym

Europe

Bulgaria

Woodchat Shrike

L. senator

LanSen

Declining

Lesser Grey Shrike

L. minor

LanMin

Declining

Common Magpie

Pica pica

PicPic

Secure

Uncertain

Eurasian Jay

Garrulus glandarius

GarGla

Secure

Decreasing

Common Raven

Corvus corax

CorCor

Secure

Common Starling

Sturnus vulgaris

StuVul

Declining

Roseâ&#x20AC;&#x201C;coloured Starling

S. roseus

StuRos

Secure

Eurasian Golden Oriole

Oriolus oriolus

OriOri

Secure

Uncertain

House Sparrow

Passer domesticus

PasDom

Declining

Uncertain

Spanish Sparrow

Passer hispaniolensis

PasHis

Secure

Eurasian Tree Sparrow

P. montanus

PasMon

Declining

Common Linnet

Carduelis cannabina

CarCan

Declining

European Goldfinch

C. carduelis

CarCar

Secure

Uncertain

European Greenfinch

C. chloris

CarChl

Secure

Uncertain

Ortolan Bunting

Emberiza hortulana

EmbHor

Depleted

Uncertain

Yellowhammer

E. citrinella

EmbCit

Secure

Cirl Bunting

E. cirlus

EmbCir

Secure

Blackâ&#x20AC;&#x201C;headed Bunting

E. melanocephala

EmbMel

Depleted

Uncertain

Corn Bunting

Miliaria calandra

MilCal

Declining

Decreasing

Uncertain

"La tortue greque" Oeuvres du Comte de Lacépède comprenant L'Histoire Naturelle des Quadrupèdes Ovipares, des Serpents, des Poissons et des Cétacés; Nouvelle édition avec planches coloriées dirigée par M. A. G. Desmarest; Bruxelles: Th. Lejeuné, Éditeur des oeuvres de Buffon, 1836. Pl. 7

Editor executiu / Editor ejecutivo / Executive Editor Joan Carles Senar

Secretaria de Redacció / Secretaría de Redacción / Editorial Office

Secretària de Redacció / Secretaria de Redacción / Managing Editor Montserrat Ferrer

Museu de Zoologia Passeig Picasso s/n 08003 Barcelona, Spain Tel. +34–93–3196912 Fax +34–93–3104999 E–mail mzbpubli@intercom.es

Consell Assessor / Consejo asesor / Advisory Board Oleguer Escolà Eulàlia Garcia Anna Omedes Josep Piqué Francesc Uribe

Editors / Editores / Editors Antonio Barbadilla Univ. Autònoma de Barcelona, Bellaterra, Spain Xavier Bellés Centre d' Investigació i Desenvolupament CSIC, Barcelona, Spain Juan Carranza Univ. de Extremadura, Cáceres, Spain Luís Mª Carrascal Museo Nacional de Ciencias Naturales CSIC, Madrid, Spain Adolfo Cordero Univ. de Vigo, Vigo, Spain Mario Díaz Univ. de Castilla–La Mancha, Toledo, Spain Xavier Domingo Univ. Pompeu Fabra, Barcelona, Spain Francisco Palomares Estación Biológica de Doñana, Sevilla, Spain Francesc Piferrer Inst. de Ciències del Mar CSIC, Barcelona, Spain Ignacio Ribera The Natural History Museum, London, United Kingdom Alfredo Salvador Museo Nacional de Ciencias Naturales, Madrid, Spain José Luís Tellería Univ. Complutense de Madrid, Madrid, Spain Francesc Uribe Museu de Zoologia de Barcelona, Barcelona, Spain Consell Editor / Consejo editor / Editorial Board José A. Barrientos Univ. Autònoma de Barcelona, Bellaterra, Spain Jean C. Beaucournu Univ. de Rennes, Rennes, France David M. Bird McGill Univ., Québec, Canada Mats Björklund Uppsala Univ., Uppsala, Sweden Jean Bouillon Univ. Libre de Bruxelles, Brussels, Belgium Miguel Delibes Estación Biológica de Doñana CSIC, Sevilla, Spain Dario J. Díaz Cosín Univ. Complutense de Madrid, Madrid, Spain Alain Dubois Museum national d’Histoire naturelle CNRS, Paris, France John Fa Durrell Wildlife Conservation Trust, Trinity, United Kingdom Marco Festa–Bianchet Univ. de Sherbrooke, Québec, Canada Rosa Flos Univ. Politècnica de Catalunya, Barcelona, Spain Josep Mª Gili Inst. de Ciències del Mar CMIMA–CSIC, Barcelona, Spain Edmund Gittenberger Rijksmuseum van Natuurlijke Historie, Leiden, The Netherlands Fernando Hiraldo Estación Biológica de Doñana CSIC, Sevilla, Spain Patrick Lavelle Inst. Français de recherche scient. pour le develop. en cooperation, Bondy, France Santiago Mas–Coma Univ. de Valencia, Valencia, Spain Joaquín Mateu Estación Experimental de Zonas Áridas CSIC, Almería, Spain Neil Metcalfe Univ. of Glasgow, Glasgow, United Kingdom Jacint Nadal Univ. de Barcelona, Barcelona, Spain Stewart B. Peck Carleton Univ., Ottawa, Canada Eduard Petitpierre Univ. de les Illes Balears, Palma de Mallorca, Spain Taylor H. Ricketts Stanford Univ., Stanford, USA Joandomènec Ros Univ. de Barcelona, Barcelona, Spain Valentín Sans–Coma Univ. de Málaga, Málaga, Spain Tore Slagsvold Univ. of Oslo, Oslo, Norway

Animal Biodiversity and Conservation 24.1, 2001 © 2001 Museu de Zoologia, Institut de Cultura, Ajuntament de Barcelona Autoedició: Montserrat Ferrer Fotomecànica i impressió: Sociedad Cooperativa Librería General ISSN: 1578–665X Dipòsit legal: B–16.278–58

Animal Biodiversity and Conservation 34.1 (2011)

Rare insights into intraspecific brood parasitism and apparent quasi–parasitism in black–capped chickadees K. A. Otter, B. W. Murray, C. I. Holschuh & K. T. Fort

Otter, K. A., Murray, B. W., Holschuh, C. I. & Fort, K. T., 2011. Rare insights into intraspecific brood parasitism and apparent quasi–parasitism in black–capped chickadees. Animal Biodiversity and Conservation, 34.1: 23–29. Abstract Rare insights into intraspecific brood parasitism and apparent quasi–parasitism in black–capped chickadees.— Genetic analysis of passerine birds often finds evidence of extra–pair copulations within species, but genetic evidence of intraspecific brood parasitism (IBP) and quasi–parasitism (Q–P) are relatively rare. Further, it is even rarer for genetic patterns that might indicate quasi–parasitism (resident male sires offspring through extra–pair copulations, and allows the female to lay these within the male’s nest) to be coupled with observational evidence of this behavior. In this paper, we report behavioral observations surrounding the nest of black–capped chickadee, one of the few species in which both IBP and Q–P have been detected through a genetic analysis. These were later confirmed to have young genetically mismatched with both attending parents, as well as mismatched with the attending female but sired by the attending male. The behavioral patterns associated with this nest are contrasted with the two previously reported cases of IPB/Q–P in this species, and suggest that rare ‘detection’ of quasi–parasitism may be explained by converging patterns of extra–pair behavior and the rarer strategy of intraspecific brood parasitism. Key words: Interspecific brood parasitism, Quasi–parasitism, Black–capped chickadees. Resumen Ideas poco frecuentes del parasitismo de puesta intraespecífico y el cuasiparasitismo aparente del carbonero cabecinegro.— El análisis genético de los paseriformes a menudo se tropieza con evidencias de cópulas fuera de pareja ocurridas dentro de la misma especie, sin embargo las evidencias genéticas del parasitismo de puesta intraespecífico (IBP) y el cuasiparasitismo (Q–P) son relativamente raras. Además, es incluso más raro que los patrones genéticos que podrían indicar el cuasiparasitismo (un macho residente engendra hijos mediante una cópula fuera de su pareja, y permite que la hembra ponga los huevos dentro del nido masculino) estén respaldados por evidencias observadas de esta conducta. En este artículo, informamos de las observaciones etológicas que tuvieron lugar en torno a un nido de carbonero cabecinegro, una de las pocas especies en las que se ha detectado tanto el IBP como el Q–P mediante análisis genético. Más adelante se confrimó que los jóvenes no coincidían genéticamente con ambos padres cuidadores, así como tampoco coincidían con la hembra cuidadora, pero si con el macho cuidador. Los patrones conductuales asociados a este nido se comparan con los otros dos casos conocidos con anterioridad de IPB/Q–P en esta especie, y se sugiere que la "detección” poco frecuente del cuasiparasitismo puede explicarse mediante los patrones convergentes de las conductas extra pareja y la estrategia aún más rara del parasitismo de puesta intraespecífico. Palabras clave: Parasitismo de puesta intraespecífico, Cuasiparasitismo, Carbonero cabecinegro. (Received: 20 XII 10; Final acceptance: 7 III 11) Ken A. Otter, Brent W. Murray, Carmen I. Holschuh & Kevin T. Fort, Ecosystem Science & Management Program, Natural Resources and Environmental Studies, Univ. of Northern British Columbia, Prince George, BC, V2N 4Z9, Canada.– Carmen I. Holschuh, Westland Resource Group, #203 – 830 Shamrock Street, Victoria, BC, Canada V8X 2V1.– Kevin T. Fort, Pacific & Yukon Canadian Wildlife Service, 5421 Robertson Rd., Delta, BC, Canada V4K 3N2. Corresponding author: Ken A. Otter. E–mail: otterk@unbc.ca ISSN: 1578–665X

Introduction Intraspecific brood parasitism (IBP) –eggs/nestlings appearing within a nest though they are not genetically related to either the resident male and female– commonly occurs in some bird taxa that have precocial young, such as ducks and coots. The prevalence of IBP (or 'egg dumps') within these groups increases with nesting densities (Eadie et al., 1998). Conservation programs to increase population sizes, such as in wood ducks (Aix sponsa), find that rates of IBP increase if nestboxes are placed either too close together or in too exposed a location, attracting parasitic females and giving them the opportunity for egg dumping (Eadie et al., 1998). As IBP increases, clutch sizes increase, and hatching success of parasitized females declines due to inefficient incubation. Yet IBP is not a byproduct of management but has evolved as a strategy by which some females can increase their reproductive success through parasitizing the efforts of others (Eadie et al., 1998; Slagsvold, 1998) IBP as a reproductive strategy, however, appears to be rarer among socially monogamous passerines with atricial young (Slagsvold, 1998). Despite numerous studies in the past decade, mixed parentage in most passerines results from extra–pair copulations (young sired by the resident female and an extra–pair male) rather than from brood parasitism by conspecific parents not associated with the nest (IBP), or through quasi–parasitism (young sired by the resident male and an extra–pair female and subsequently laid in the male’s nest) (Griffith et al., 2004). Where IBP has been noted, these strategies may be adopted by females whose ability to nest independently is either limited or uncertain (a bet– hedging strategy to increase the chances of leaving some offspring; Otter et al., 1998; Hughes et al., 2003; Blackmore & Hinsohn, 2008). Alternately, IBP may be a mixed, sequential strategy, such as occurs in Starlings (Sturnus vulgaris) where unmated, recently–arriving females parasitize conspecifics during the period in which they themselves are settling, and then lay their own clutches normally once they have acquired a nest site (Sandell & Diemer, 1999). In this manner, the female supplements the fecundity of her own nests by adding offspring in other nests, ecologically equivalent to the gains made by males pursuing extra–pair copulations. Finally, IBP may be both a strategic and opportunistic behaviour, such as when females in colonial species in close proximity of each others’ nests occasionally egg–dump to take advantage of both mixed strategies and bet–hedging strategies (Alves & Bryant, 1998). Quasi–parasitism is an even rarer strategy than IBP among socially–monogamous passerines (Griffiths et al., 2004). In the few studies where it has been noted, it is difficult to distinguish whether the pattern of mixed parentage is the result of a quasi–parasitic strategy –males seeking EPCs, and then allowing the females to lay their sired egg within the male’s nest– or whether quasi–parasitism is the occasional result of converging patterns of EPCs and IBPs within a population. If females seek EPCs with males

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with specific attributes as extra–pair partners, and the same social pairs are the occasional targets of females engaging in IBP, then occasionally the parentage of nests may reflect patterns that appear to invoke explanations of quasi–parasitism (Griffiths et al., 2004). True quasi–parasitism, though, suggests intentionality on the part of the male to allow laying access to the nest by the secondary female, and thus is partially a male behavioural strategy. Young in the nest mismatched to the resident female, but not the resident male (apparent quasi–parasitism), that result from female–initiated patterns of EPC and IBP may not, however, involve the resident male intentionally granting laying access to the nest. Detailed behavioural observations are required to distinguish between these, but which are typically lacking due to the surreptitious nature of many of the strategies under discussion. In this paper, we report on the behaviour observed at a nest of black–capped chickadees Poecile atricapillus (one of the few species in which genetic patterns of quasi–parasitism have been reported (Otter et al., 1998)) in which IBP was suspected, and later genetically confirmed. These observations, coupled with observations from our previous studies, provide potential insight into the rare occurrence of maternally–mismatched young within chickadee nests, and suggest quasi–parasitic parentage patterns may result from overlapping strategies of EPC and IBP in female chickadees. We then discuss the potential implications for land management scenarios that may increase the particular circumstances that appear to promote these alternate strategies in chickadees, and perhaps other species. Methods Banding Birds were captured in winter flocks during January through February 2000 at temporary feeders with either potters traps or mistnets. Upon capture, each bird was fitted with a CWS numbered legband and a unique combination of three additional colored legbands to facilitate individual identification. A 75 mL blood sample was extracted at the time of winter capture for the majority of birds. For others, banding and blood samples were collected from adults by catching them at nests in June 2000 as they fed nestlings. Nest monitoring Individually–marked birds were tracked every three to four days throughout late April, May and June 2000 to document the breakup of winter flocks and individual territorial establishment (defined as the exclusive defense of habitat used for nesting and foraging). Territory boundaries were delineated by mapping the locations of song contests and fights between neighboring males following the period of flock–breakup, which occurs within our population around mid–April.

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Sample collection In the study nest, the attending male and female and the nine chicks were all captured, banded and blood sampled in 2000. The secondary pair associated with this nest consisted of a banded female captured and blood sampled in 2000, and an unbanded male with whom she associated. This male was not captured in 2000, but may have been among the returning adult birds captured in winter 2001, so all males from the subsequent winter were screened for paternity in analysis. Blood samples were collected from 59 additional territorial males that were captured and banded in 2000 through 2002. Approximately 90% of the adult territorial male birds in our study area were sampled over these three years. Whole blood was stored in 95% ethanol and DNA was extracted following a standard phenol/chloroform isolation procedure (Sambrook & Russell, 2001) Microsatellite typing Microsatellite alleles were typed at three highly informative loci, Pocc6 (Bensch et al., 1997), Pdou5 (Griffith et al., 1999) and PAT–MP243 (Otter et al., 1998). One µL of DNA was added to 14 µL of master mix containing the following: 1X PCR buffer, 1.67–3.67 mM MgCl2 (3.67 mM for Pocc6 and 1.67 mM for Pdou5 and PAT–MP243), 100 µM of each dNTP, 0.5 µM of each primer, 0.5 mg/mL BSA, and 0.5 units of Taq DNA polymerase (Invitrogen, Carlsbad, CA). Amplifications were carried out on a DYAD or PTC– 100 thermal cycler (MJ Research Inc., Waltham, MA). The thermal cycle for PAT–MP243 and Pdou5 loci was 94°C for 4 minutes, followed by 5 cycles of 94°C for 1 minute, 57°C for 1 minute (decreased by 2°C per cycle), and 72°C for 1.5 minutes; and 33 cycles of 94°C for 30 seconds, 59°C for 30 seconds, and 72°C for 1.5 minutes, and a final extension phase of 72°C for 4 minutes. The thermal cycle for Pocc6 samples was 94°C for 3 minutes, followed by 40 cycles of 94°C for 30 seconds, 56°C for 30 seconds, and 72°C for 30 seconds, and a final extension phase of 72°C for 4 minutes. Products of the separate amplification reactions were pooled. Alleles were sized using the CEQ8000 Genetic Analysis system (Beckman Coulter Inc., Fullerton, CA) and viewed using the Fragment Analysis Module (400 bp size standard; cubic model; PA ver.1 dye mobility calibration). To ensure accuracy of scoring all nestlings, putative mothers and immediate resident males were typed at least three times. The remaining males were typed multiple times as needed to obtain genotype information for all loci examined. Paternity analysis Maternal identification was established based on non–exclusion. As the maternity of all nestlings was consistent with one of the two females associated with the nest, a paternity analysis on each chick was performed using the program CERVUS, version 3.0 (Kalinowski et al., 2007) assuming known maternity.

The set of 60 resident males was used to assess the loci for Hardy–Weinberg equilibrium and to estimate allele frequency data upon which to build likelihood estimates and run paternity simulations for significance testing. Simulation parameters were 10,000 offspring, 60 candidate fathers, 85% of the candidate fathers sampled, all 3 loci completely typed, 1% of loci mistyped and a 1% error rate. An 85% chance of sampling the candidate father was used, based on the average number of males within the neighborhood of the nest that were blood–sampled and geneotyped in the analysis. The delta LOD score was used to assess confidence in the assigned male using relaxed (80%) and strict (95%) criteria. Results Observational data Following flock–breakup for the rest of the population in late April 2000, we observed two males and two female chickadees traveling together and collectively defending a single breeding territory. These birds constituted what appeared to be a non–disbanded wintering flock made up of a dominant male (YA/ RB) and female (UB) along with a subordinate male (UB) and female (BA/MG). The dominant male and subordinate female were banded prior to the breeding season (in January and February), but the dominant female and subordinate male were unbanded at the beginning of the spring. These two birds were still distinguishable due to their clear associations within this territorial flock, and the fact that all but one neighboring male in the area was color–banded in 2000 (this lone unbanded neighbor was mated to a banded female, allowing for identification through pair–wise association). The relative dominance status of the four birds was determined by repeated observations of interactions during tracking periods, with the unbanded female supplanting the banded female, and the banded male supplanting the unbanded male. Chickadee pairs can be distinguished even within flocks, as mated pairs usually travel in closer proximity to each other during foraging than non–mated pairs (Smith, 1991). Using these criteria, we determined that the dominant banded male and dominant unbanded female constituted one 'pair', and the subordinate unbanded male and subordinate banded female appeared to constitute the other 'pair' within this tetrad. Between 23 April 2000 and 19 May 2000, when the dominant female began incubating at the sole nest in the territory, this tetrad of birds was consistently seen in the same actively–defended territory on seven separate occasions. Apart from the distinction of being a tetrad, the birds acted as typical territorial breeding black–capped chickadees. The dominant male actively engaged in territorial contests with the six neighboring males that bordered this territory, often with the subordinate male from the tetrad taking a minor reinforcing role. However, this does not appear to have constituted a case of 'tolerating floaters' as

the subordinate male and female actively engaged with the dominant pair in the excavation of at least three separate nest cavities (witnessed 23, 30 April and 3, 4 May). At each of the three nest sites, all four birds were seen simultaneously excavating in alternating fashion typical of black–capped chickadees, where each individual waited on a nearby branch while another bird was in the nest, entering only when the excavating bird had exited to dispose of the excavated wood. Just prior to egg laying, female chickadees begin producing a distinctive food–solicitation call, the broken dee (see Ficken et al., 1978 for description). As the two females in the tetrad began producing this call (14 May), the subordinate male disappeared from the territory. The dominant male was witnessed giving courtship feedings to both the dominant and subordinate female during the broken dee calling period (14 May). However, once incubation of the nest began (19 May), the subordinate banded female was never seen entering the cavity to incubate, and disappeared from the territory early in this period (additional nest observations on 23, 25, 29 May and 1, 5 June). For the remainder of the breeding season, only the unbanded dominant female and the dominant male were seen attending the nest and feeding the nestlings. The subordinate female was seen foraging alone on a neighboring territory on the 23 May, but the banded female associated with this territory was already incubating and her banded mate was witnessed feeding her at their cavity within the same hour as the subordinate female was seen. During banding of nestlings (10 June), the dominant female was captured at the nest and banded (MA/ GO). A blood sample was extracted at this time for parentage analysis. Within this nest, nine nestlings were sampled and two unhatched eggs observed; this clutch size of 11 was over double that of the average clutch size for this population (mean 5.0 ± 1.6 SD based on 16 other nests that fledge in 2000 where clutch size could be determined). Allele frequency data All three loci were found to be highly informative markers for paternity analysis (table 1). Due to the large number of alleles at the Pat–MP243 and Pocc–6 loci, many genotypes had sample sizes less than 5, and analysis of Hardy–Weinberg equilibrium could not be performed. Observed and expected heterozygosity values for these loci are similar. Analysis of the Pdou–5 locus shows no significant difference from Hardy– Weinberg expectations. The average polymorphism information content for these loci is very high and the combined non–exclusion probability of a second parent, with the first known is below 0.05 (0.016). Paternity analysis Of the nine offspring, four were genetically consistent with maternity of the dominant female (MA/GO) and five for the subordinate female (BA/MG). In the paternity analysis, the identity of the female was

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assumed as known. For eight of the offspring, the banded, dominant male (YA/RB) associated with this nest was identified as the most likely father (four at 95%, two at 80% confidence, and the remaining two at just under 80% confidence). Other candidate males were identified for offspring who had 80% confidence or below; however, only the assumed father was common to all candidate lists, and was classified as the most likely father in all cases. Further, many of the other candidate males did not have territories neighboring the focal nest. One nestling with subordinate female maternity was not the offspring of the assumed father (a case of IBP). In this case two candidate males were identified but one of these was banded in distant territories while the most likely father (80% confidence) was the resident male of a territory adjacent to the nest (BA/GY). This male was a subordinate male within his own flock, and subordinate to YA/ RB, but his territorial status would suggest he was dominant to the unbanded, subordinate male of the tetrad (relative age and territorial status would infer this relationship –Smith, 1991). Among the four identified parents within this nest (YA/RB, MA/GO, BA/MG and the neighbor BA/GY), only BA/GY returned in 2001 to breed. As dispersal between breeding seasons is very low in black–capped chickadees (Smith, 1991), failure to return to the breeding population in subsequent years is typically an indicator of failure to survive the winter. Discussion If assessed strictly on genetic classification of nestlings, without insight into the behavioural patterns surrounding this particular nest, we would have identified this as a case of both quasi– and Intraspecific brood parasitism; five of the young in the nest were sired by a female not incubating the nest (IBP) and of these, four young were sired by the resident attending male (quasi–parasitism). Further, the identified parasitic female was known to be subordinate to the attending pair in the preceding winter flock, and the sole other identified father of the one pure IBP nestling was a male subordinate to the attending male. This pattern closely matches the one other documented case of genetically–identified quasi–parasitism in black–capped chickadees (Otter et al., 1998), where the eggs of a subordinate female were found in the nest of a dominant pair from the same flock. A number of those eggs were sired by the dominant male through extra–pair copulations (as these were truly separate breeding pairs), as well as by the parasitic female’s social mate —himself subordinate to the dominant male in whose nest his mate had dumped her eggs. In this nest, we had not been able to clarify whether mis–matched maternity may have been the result of a rapid nest switch. However, in that same study a second nest with pure IBP was also found among chickadees in which rapid nest switching could be ruled out. In all three nests (this study and those in Otter et al., 1998), however, the pattern of mis–matched

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Table 1. Summary statistics of allele frequency analysis of 60 resident males. For each locus the sample size (N), number of alleles detected (k), observed (Hobs)) and expected (Hexp) heterozygosity, polymorphism information content (PIC) and the probability of excluding father when the mother is known (NE–pat) is shown: * Product of individual loci. Tabla 1. Resumen estadístico del análisis de frecuencia alélica de 60 machos residentes. Para cada locus se incluye el tamaño de la muestra (N), el número de alelos detectados (k), la heterozigosidad observada (Hobs) y esperada (Hexp), el contenido de información del polimorfismo (PIC) y la probabilidad de excluir al padre cuando se conoce la madre (NE–pat): * Producto de locus individuales. Locus

Hobs

Hexp

PIC

Pat–MP243

0.950

0.900

0.884

0.211

Pocc–6

0.950

0.921

0.907

0.175

Pdou–5

0.750

0.786

0.748

0.424

Average (product*)

15.67

0.883

0.869

0.846

0.016*

maternity, regardless of the siring male, is consistent. The eggs mismatched to the attending female came from a subordinate female that occupied the same winter flock, and that subordinate female was unsuccessful in nesting in the year that dumping occurred. In this regard, IBP in chickadees may be a rare strategy following the ‘bet–hedging’ pattern suggested by Hughes et al. (2003) when population densities are relatively high and rates of nest failure are fairly high. The latter is certainly the pattern that occurs in black–capped chickadees, where nesting success is rank–dependent . Subordinate pairs are more likely to suffer nest loss than are dominant pairs (Otter et al., 1999), an effect that is amplified in marginal habitats (Fort & Otter, 2004; Otter et al., 2007). If subordinate females have low potential to successfully rear a brood to fledging, it may be strategic for them to lay eggs in the nests of known dominant neighbours, as dominant birds are more likely to successful fledge young. Where this has potential conservation implications is when land–use practices affect the perceived quality of the habitats to the birds –within our region, birds that nest in young forests that are regenerating from recent logging experience low success rates, especially among subordinate pairs (such as early– seral forests; Otter et al., 2007). Such scenarios could lead to increased use of intraspecific brood parasitism as a strategy by subordinate females to ensure reproductive success. This, however, does not account for two of the three nests reported between this study and Otter et al. (1998) having offspring that mis–matched the attending female, but not the attending male of the focal nest. Such genetic patterns could lead one to conclude that quasi–parasitism was an active, if rare, strategy in black–capped chickadees. Griffith et al. (2004) point out that quasi–parasitism in its truest sense is at least partially a male strategy; the male

NE–pat

engages in extra–pair copulations and then allows the extra–pair female to lay these eggs within the nest he attends with his social mate. We do not feel that this to be a likely scenario in explaining the patterns of nesting observed in chickadees. First, quasi–parasitism would assume a similar surreptitious nature to that associated with extra–pair copulations. It is not in the male’s interest to advertise to his social mate that a number of the young within their nest are not her genetic offspring, lest she diminish parental care. Neither does it necessarily benefit the male to have the extra young of a second female within his brood without securing the parental care from both females. Studies involving experimental increases of clutch sizes clearly indicate that condition and survival of the attending parents are compromised with the increased parental effort involved in feeding large broods (e.g. Gustaffson et al., 1995; Yamaguchi, 1997). It is perhaps not a coincidence that the male and female attending the nest in this study, whose brood size was substantively higher than the average for this population, did not return to breed in 2001; by comparison, the neighboring male that sired the single purely–IBP nestling, but did not attend the nest, did survive. However, if this does not constitute true quasi–parasitism in a strategic sense, why does the genetic pattern of this nest –and that reported in Otter et al. (1998)– suggest this pattern? Griffith et al. (2004) have likely identified the explanation: where females within a species seek males with certain attributes for extra–pair copulations, and the nests of the same pairs are also the targets for intraspecific brood parasitism, occasionally the genetic patterns of these nests will collide to suggest quasi–parasitism. This is the likely scenario to explain these rare occurrences in black–capped chickadees. Female chickadees are the sex that actively seeks extra–pair copulations (Smith, 1988), and their selection of males

is not arbitrary. Females prefer extra–pair sires that are socially dominant to the soliciting female’s mate (Smith, 1988; Otter et al., 1994, 1998; Mennill et al., 2004). The only three known cases of nests containing young genetically–mismatched from the attending female were also deposited by subordinate females into the nests of higher–ranking flockmates. This may be strategic due to the differential survival of nests of dominant pairs (Otter et al., 1999; Fort & Otter, 2004). Extra–pair copulations in black–capped chickadees are relatively common, occurring in about 30% of nests (Otter et al., 1998; Mennill et al., 2004), and they typically involve the same males (higher–ranked flockmates) that were found to be the target nests of parasitic females in this and Otter et al. (1998) study. Therefore, it is probabilistic that parasitic females may have engaged in EPCs with the attending male prior to parasitizing the nest. Although the currently– reported case is unusual in the failure of the flock to disband and the early–season breeding affiliation of the resident male with both females, other cases of apparent quasi–parasitism may easily be explained by overlapping, but independent, female strategies that simply have a common target male phenotype. This is already known to be the case for the female strategies of divorce and EPCs in Black–capped Chickadees (Ramsay et al., 2000). Why, then, is IBP so rare in chickadees? More than a decade of analysis of paternity on an Ontario, Canada population (115 nests combined between Otter et al., 1998; Mennill et al., 2004) found only the two cases of mismatched–maternity. This may arise from the costs of intraspecific nest parasitism evolving when conspecifics have non–asymmetry in nestling size between within–pair and parasitic young (Slagsvold, 1998). Such a scenario merely increases broodsize, stretching parental provisioning efforts and decreasing both the resources to and condition of the individual nestlings. Young from such enlarged broods are likely to have decreased survival, which would tend to diminish the evolutionary potential to pass on parasitic genes. Simultaneously, non–discriminating parents who accept enlarged broods would suffer higher mortality, strongly selecting for behaviors such as abandonment. This has been suggested as an explanation as to why intraspecific brood parasitism is more common in species with precocial young, where increased brood size does not necessarily have the same debilitating effect on parents or offspring survival (Slagsvold, 1998). Of concern, then, is habitat changes that decrease chances for subordinate females to breed (Otter et al., 2007) as this could lead to increased IBP, and in turn result in higher mortality among targeted dominant pairs that suffer from increased parental care costs. Acknowledgements This work was funded by NSERC Discovery Grants (K. O. and B. M.), the University of Northern BC, the Northern Land Use Institute (K. O.), and the Canadian Foundation for Innovation (B. M.). Finally, we would

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also like to thank the many patient students of Biol 425 –Applied Genetics and Biotechnology– at UNBC over the past five years, who have been given the DNA for individuals associated with this nest as an advanced exercise in genetoyping patterns of parentage; we felt it was time to explain how this might have arisen. References Alves, M. A. S. & Bryant, D. M., 1998. Brood parasitism in the sand martin, Riparia riparia: evidence for two parasitic strategies in a colonial passerine. Animal Behaviour, 56: 1323–1331. Bensch, S., Price, T. & Kohn, J., 1997. Isolation and characterization of microsatellite loci in a Phylloscopus warbler. Molecular Ecology, 6: 91–92. Blackmore, C. J. & Heinsohn, R., 2008. Variable mating strategies and incest avoidance in cooperatively breeding grey–crowned babblers. Animal Behaviour, 75: 63–70. Eadie, J., Sherman, P. & Semel, B., 1998. Conspecific brood parasitism, population dynamics and the conservation of cavity–nesting birds. In: Behavioral ecology and conservation biology: 306–340 (T. Caro, Ed.). Oxford Univ. Press. Ficken, M. S., Ficken, R. W. & Witkin, S. R., 1978. Vocal repertoire of the black–capped chickadee. Auk, 95: 34–48. Fort, K. T. & Otter, K. A., 2004. Effects of habitat disturbance on reproduction in black–capped chickadees (Poecile atricapillus) in northern British Columbia. Auk, 121: 1070–1080. Griffith, S. C., Lyon, B. E. & Montgomerie, R., 2004. Quasi–parasitism in birds. Behavioral Ecology and Sociobiology, 56: 191–200. Griffith, S. C., Stewart, I. R. K., Dawson, D. A., Owens, I. P. F. & Burke, T., 1999. Contrasting levels of extra–pair paternity in mainland and island populations of the house sparrow (Passer domesticus): is there an ‘island effect’? Biological Journal of the Linnaean Society, 68: 303–316. Gustafsson, L., Qvarnström, A. & Sheldon, B. C., 1995. Trade–offs between life–history traits and a secondary sexual character in male collared flycatchers. Nature, 375: 311–313. Hughes, J. M., Mather, P. B., Toon, A., Ma, J., Rowley, I. & Russell, E., 2003. High levels of extra–group paternity in a population of Australian magpies Gymnorhina tibicen: evidence from microsatellite analysis. Molecular Ecology, 12: 3441–3450. Kalinowski, S. T., Taper, M. L. & Marshall, T. C., 2007. Revising how the computer program CERVUS accommodates genotyping error increases success in paternity assignment. Molecular Ecology, 16: 1099–1106. Mennill, D. J., Ramsay, S. M., Boag, P. & Ratcliffe, L. M., 2004. Patterns of extrapair mating in relation to male dominance status and female nest placement in black–capped chickadees. Behavioral Ecology, 15: 757–765. Otter, K., Ratcliffe, L. & Boag, P., 1994. Extra–pair paternity in the black–capped chickadee. Condor,

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96: 218–222. Otter, K., Ratcliffe, L., Michaud, D. & Boag, P., 1998. Do female black–capped chickadees prefer high– ranking males as extra–pair partners? Behavioral Ecology and Sociobiology, 43: 25–36. Otter, K., Ramsay, S. M. & Ratcliffe, L., 1999. Enhanced reproductive success of female black–capped chickadees mated to high–ranking males. Auk, 116: 345–354. Otter, K. A., Van Oort, H. & Fort, K. T., 2007. Habitat quality and reproductive behavior in chickadees and tits: potential for habitat matrix use in forest generalists. In: The ecology and behavior of chickadees and titmice: an integrated approach: 277–291 (K. A. Otter, Ed.). Oxford Univ. Press. Ramsay, S. M., Otter, K. A., Mennill, D. J., Ratcliffe, L. M. & Boag, P. T., 2000. Divorce and extrapair mating in female black–capped chickadees (Parus atricapillus): separate strategies with a common target. Behavioral Ecology and Sociobiology, 49:

18–23. Sambrook, J. & Russell, D. W., 2001. Molecular cloning: a laboratory manual. Third Edition. Cold Spring Harbor Laboratory Press, New York. Sandell, M. I. & Diemer, M., 1999. Intraspecific brood parasitism: a strategy for floating females in the European starling. Animal Behaviour, 57: 197–202. Slagsvold, T., 1998. On the origin and rarity of interspecific nest parasitism in birds. American Naturalist, 152: 264–272. Smith, S. M., 1988. Extra–pair copulations in black– capped chickadees: the role of the female. Behaviour, 107: 15–23. – 1991. The black–capped chickadee; behavioural ecology and natural history. Comstock Publishing Associates, New York. Yamaguchi, Y., 1997. Intraspecific nest parasitism and anti–parasite behavior in the grey starling, Sturnus cineraceus. Journal of Ethology, 15: 61–68.

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Museu de Zoologia Passeig Picasso s/n 08003 Barcelona, Spain Tel. +34–93–3196912 Fax +34–93–3104999 E–mail mzbpubli@intercom.es

Consell Assessor / Consejo asesor / Advisory Board Oleguer Escolà Eulàlia Garcia Anna Omedes Josep Piqué Francesc Uribe

Animal Biodiversity and Conservation 34.1 (2011)

Loggerhead sea turtle bycatch data in artisanal fisheries within a marine protected area: fishermen surveys versus scientific observations M. Lozano, J. Baro, T. García, A. Frías, J. Rey & J. C. Báez

Lozano, M., Baro, J., García, T., Frías, A. Rey, A. & Báez, J. C., 2011. Loggerhead sea turtle bycatch data in artisanal fisheries within a marine protected area: fishermen surveys versus scientific observations. Animal Biodiversity and Conservation, 34.1: 31–34. Abstract Loggerhead sea turtle bycatch data in artisanal fisheries within a marine protected area: fishermen surveys versus scientific observations.— Loggerhead sea turtles can be incidentally captured by artisanal gears but information about the impact of this fishing is inconsistent and scarce. Recent studies have observed that the bycatch, or incidental catch rate, in fishermen surveys is irregular. The aim of this study was to compare direct data (onboard observers) concerning the incidental catch of loggerhead sea turtles by the artisanal vessels versus data from fishermen surveys. The study area was the Cabo de Gata–Níjar marine protected area, situated in the western Mediterranean (southeast of the Iberian peninsula). We observed two loggerhead turtles that were incidentally caught in a total of 165 fishing operations. According to fishermen surveys, a total of nine loggerheads were incidentally caught in 861 fishing operations. The differences between the loggerhead sea turtle bycatch reported by fishermen surveys and scientific observations versus random distribution (x2 = 0.3146, P = 0.575, df = 1) were not significant. We conclude that the surveys are useful but that findings should be interpreted with caution. Key words: Fishermen surveys, Marine protected area, Mediterranean, Sea turtle. Resumen Análisis de los datos de capturas accidentales de tortugas bobas por la pesca artesanal en una área marina protegida: notificaciones de pescadores encuestados frente a observaciones científicas.— La tortuga boba puede ser capturada accidentalmente por la flota de pesca artesanal, pero la información del impacto de estas pesquerías sobre esta especie es escasa e inconsistente. En trabajos recientes, se ha observado que las capturas fortuitas, o tasas de captura accidental, según las cuestas realizadas a los pescadores, son irregulares. El objetivo de este estudio fue comparar los datos directos (de observadores a bordo) de capturas fortuitas de tortugas bobas, frente a los datos recogidos mediante encuestas en pesquerías artesanales. La zona de estudio fue el área marina protegida de Cabo de Gata–Níjar, situada en el Mediterráneo occidental (sureste de la península ibérica). Observamos dos capturas accidentales de tortuga boba en 165 operaciones de pesca. El total de capturas fortuitas de tortugas bobas controlado mediante las encuestas fue de nueve de un total de 861 operaciones de pesca. No se observaron diferencias significativas entre las capturas accidentales de tortugas bobas reportadas por los pescadores mediante encuesta, con respecto a las observadas de forma científica (x2 = 0,3146, P = 0,575, gl = 1). Llegamos a la conclusión de que las encuestas podrían ser consideradas útiles, pero deberían interpretarse con prudencia. Palabras clave: Encuestas a pescadores, Áreas marinas protegidas, Mediterráneo, Tortuga marina. (Received: 26 XI 10; Acceptació condicional: 2 II 11; Acceptació definitiva: 21 III 11) Matías Lozano, Jorge Baro, Teresa García, Javier Rey & José C. Báez, Inst. Español de Oceanografía, Centro Oceanográfico de Málaga, Málaga, España (Spain).– Antonio Frías, Reservas Marinas de Cabo de Gata–Níjar e Isla de Alborán, TRAGSATEC. ISSN: 1578–665X

Lozano et al.

Introduction

Material and methods

Fisheries bycatch has been identified as a primary driver of population decline in marine turtles (Lewison et al., 2004). For this reason, studies on the impact of marine fishing have increased. Many of these studies are not scientific surveys but are based on fishermen bycatch–reports. Artisanal fishing makes up the largest sector of the industry worldwide, employing more than 90%, of the fishermen in the world. Moreover, on a worldwide level, almost half the landings are estimated to originate from artisanal fisheries (FAO, 2003). Artisanal fisheries are characterized by a large number of small vessels that exploit a wide variety of species using multiple fishing gears. Loggerhead sea turtles can be incidentally captured by artisanal gears, but information on the impact of this fishing is inconsistent and scarce, as shown by Báez et al., (2006). Interviews surveys have been widely used in the characterization of sea turtle capture (e.g. Álvarez de Quevedo et al., 2010; Báez et al., 2006). Moore et al. (2010) proposed a protocol that consists of in–depth interaction with fishermen to collect data and to use this to map the sea turtle bycatch of artisanal fishing , concluding that interview surveys are an inexpensive and fast means to achieve coarse–level information for large areas. However, Báez et al. (2006) found that the rates of incidental turtle catch in fishermen surveys was irregular. Studies to verify data from fishermen surveys are lacking. The aim of this study was to compare direct data (onboard observers) concerning loggerhead sea turtles incidentally caught by the artisanal vessels versus data from fishermen surveys in these vessels within a marine protected area. This study was part of a Spanish Research Project (PARCGA, Monitoring Artisanal and Recreational Fisheries in the Marine Reserve of Cabo de Gata–Níjar) whose objective is to obtain a reliable picture of fishing activities in the marine reserve through the description of the fishing fleet and the estimation of catches.

The study area is located within Cabo de Gata–Níjar Marine Reserve (southwestern Mediterranean, Spain). This protected area covers 16,853 ha. Waters are mainly shallow (< 50 m) and characterized by a typical Mediterranean ecosystem with rocky reefs, sandy bottoms and seagrass beds (Posidonia oceanica). Fishing grounds are exploited by a small artisanal fleet of 14 units. As boats are generally small (< 12 m length, two fisherman per boat), fishing grounds are located in shore areas. In this artisanal fishing Sepia sp. and Mullus sp. trammelnets are the most common used fishing gear. This study was carried out from March 2008 to April 2010. We sampled eight of the trammel net vessels. Catch and effort data (fishing trips), which are shown in table 1, were collected in two ways: through weekly phone calls, and through observers on board for five days each month. A total of 576 telephone interviews were made and 165 boarding trips were done. In each interview, fishermen were asked about number of fishing days, haul location and catches, and during the surveys, observers accompanied fishermen for one full–day fishing trip. Net length, haul location, depth and fishing time data were collected on every trip. The observer identified, measured (total length) and weighed all retained individuals (commercial catches and discards). We tested the differences in loggerhead sea turtle bycatch data reported by fishermen surveys versus scientific observations using the chi–squared (x2) test. Expected values in the x2 tests were calculated according to the number of fishing operations controlled for each case (data from fishermen surveys and from onboard observations). Thus, we tested the observed turtle bycatch distribution per surveys and scientific observations versus weighted random distribution total of turtle bycatch.

Table 1. Turtle bycatch taken from data collected from fishing trips with onboard observers and from telephone surveys during the study period: T. Total for the year; TBr. Turtle bycatch report. Table 1. Capturas accidentals de tortugas a partir de datos recogidos durante los viajes de pesca con observadores a bordo y mediante informes telefónicos llevados a cabo durante el periodo de estudio: T. Total del año; TBr. Capturas fortuitas.

2008

2009

2010

TBr

Onboard observations

102

165

Fishermen surveys

289

411

161

861

Total

342

513

171

1026

Animal Biodiversity and Conservation 34.1 (2011)

2º 20' 52'' W 2º 17 '40'' W 2º 14' 28'' W 2º 11' 16'' W 2º 8' 4'' W 2º 4' 52'' W 2º 1' 40'' W 1º 58' 28'' W 1º 55' 16'' W 1º 52' 4'' W 36º 59' 9'' N

Carboneras

MPA

36º 59' 9'' N

Incidental catch (surveys) Incidental catch (observed on board)

Aguamarga

36º 55' 57'' N

Las Negras

36º 52' 45'' N

Retamar 36º 49' 33'' N

36º 49' 33'' N

La Isleta del Moro S. Miguel de Cabo de Gata

36º 46' 21'' N

San José Spain

36º 43' 9'' N

4,200

8,400

36º 43' 9'' N Mediterranean Sea

16,800 m

Atlantic Ocean

36º 39' 57'' N

2º 20' 52'' W 2º 17 '40'' W 2º 14' 28'' W 2º 11' 16'' W 2º 8' 4'' W 2º 4' 52'' W 2º 1' 40'' W 1º 58' 28'' W 1º 55' 16'' W 1º 52' 4'' W

Fig. 1. Map of the marine reserve of Cabo de Gata–Níjar. The black dots show the loggerheads bycatches observed on board, the polygons are defining the areas where the fishermen surveyed say they’ve catched of loggerheads bycatches: MPA. Marine protected area. Fig. 1. Mapa de la reserva marina de Cabo de Gata–Níjar. Los puntos negros representan las capturas fortuitas de tortuga boba observadas a bordo, los polígonos delimitan el área donde los pescadores encuestados afirman haber realizado las capturas fortuitas: MPA. Área marina protegida.

Results and discussion

Acknowledgements

Eleven loggerhead turtles were incidentally caught according to data from fishermen surveys plus the figures from direct observations in a total of 1,026 fishing operations per eight boats controlled in the study period (table 1). We observed directly onboard two loggerheads sea turtles incidentally caught out of a total of 165 fishing operations (fig. 1). In general, we observed a low frequency of the incidental catches of sea turtle in the study area. We did not observe significant differences between the loggerhead sea turtle bycatch reported by fishermen surveys versus scientific observations (x2 = 0.3146, P = 0.575, df = 1). In accordance with Moore et al. (2010), in function of our results we conclude that the interview surveys should be considered useful to obtained rapid information about sea turtle bycatch, but data should be interpreted with caution as interview surveys should be based on previous in–depth interaction with fishermen.

This study was partially funded by the Spanish Institute of Oceanography and Sea General Secretary (Ministry of Environment, Rural and Marine). We are grateful to the skippers and fishermen for providing the data from the boats. References Alvarez de Quevedo, I., Cardona, L., De Haro, A., Pubill, E. & Aguilar, A., 2010. Sources of bycatch of loggerhead sea turtles in the western Mediterranean other than drifting longlines. ICES Journal of Marine Science, 67: 677–685. Báez, J. C., Camiñas, J. A. & Rueda, L., 2006. Incidental capture of marine turtles fisheries of South Spain. Marine Turtle Newsletter, 111: 11–12. FAO, 2003. Strategies for Increasing the Sustainable Contribution of Small–scale Fisheries to Food Security and Poverty Alleviation. FAO, Rome: 1–14.

Lewison, R. L., Crowder, L. B., Read, A. J. & Freeman, S. A., 2004. Understanding impacts of fisheries by–catch on marine megafauna. Trends in Ecology and Evolution, 19: 598–604. Moore, J. E., Cox, T. M., Lewison, R. L., Read, A. J., Bjorkland, R., McDonald, S. L., Crowder, L.

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B., Aruna, E., Ayissi, I., Espeut, P., Joynson– Hicks, C., Pilcher, N., Poonian, C., Solarin, B. & Kiszka, J., 2010. An interview–based approach to assess marine mammal and sea turtle captures in artisanal fisheries. Biological Conservation, 143: 795–805.

Animal Biodiversity and Conservation 34.1 (2011)

¿Existe isomorfía en los huevos de las especies de la familia Ardeidae (Aves, Ciconiiformes)? D. Denis & U. Olavarrieta

Denis, D. & Olavarrieta, U., 2011. ¿Existe isomorfía en los huevos de las especies de la familia Ardeidae (Aves, Ciconiiformes). Animal Biodiversity and Conservation, 34.1: 35–45. Abstract Is there isomorphy in eggs belonging to the family Ardeidae (Aves, Ciconiiformes)?— Egg shapes in birds reflect many anatomical, biophysical and ecological aspects. In previous literature it has been assumed that a similarity in volumetric indexes from external dimensions is an indicator of constancy in shape of egret eggs (Aves, Ardeidae), but results are not consistent. Previous researchers have used lineal dimension rates to estimate shape, but these can distort the results because both aspects are orthogonal by definition. In the current research we analyze differences in egg shape between eight species of egrets and herons using elliptic Fourier descriptors and landmarks over 203 digital pictures of eggs kept in oological collections. Comparison between species and a discriminate function analysis show that shape is insufficient to distinguish species. The elongation index and breadth of eggs were significantly correlated. Our results suggest that egg shape can discriminate ecological groups but not species, indicating there is no general isomorphy in Ardeidae. Key words: Egg shape, Egrets, Herons, Fourier, Landmarks. Resumen ¿Existe la isomorfía en los huevos de la familia Ardeidae (Aves, Ciconiiformes)?— La forma de los huevos de las aves responde a numerosos factores desde anatómicos, biofísicos o ecológicos. Se ha mencionado que la ausencia de diferencias en los índices volumétricos derivados de dimensiones es un indicador de estabilidad en las formas de los huevos de las garzas (Aves, Ardeidae) pero los resultados no son consistentes. Investigaciones previas se han basado en tasas de dimensiones lineales como estimadores de la forma, lo cual puede distorsionar los resultados, al ser ambas categorías ortogonales por definición. En este trabajo se analizan las diferencias en la forma de los huevos de ocho especies de garzas empleando los descriptores elípticos de Fourier y puntos clave sobre fotografías digitales de 203 huevos depositados en colecciones. Un análisis discriminante con los coeficientes de los descriptores de Fourier y un análisis de curvaturas principales mostraron la imposibilidad de determinar la especie a partir de la forma de los huevos. El índice de elongación y la anchura de los huevos están significativamente correlacionados. Nuestros resultados sugieren que la forma del huevo permite diferenciar grupos ecológicos pero no especies, lo que indica que no hay isomorfismo en los huevos de la familia Ardeidae. Palabras claves: Forma del huevo, Garzas, Fourier, Puntos clave. (Received: 12 I 10; Conditional acceptance: 19 VII 10; Final acceptance: 10 IX 10) Dennis Denis & Ulises Olavarrieta, Depto. Biología Animal y Humana, Fac. de Biología, Calle 25 entre J e I, Vedado, Ciudad Habana, Cuba.

ISSN: 1578–665X

Introducción El estudio de los huevos de las aves ha tenido siempre un importante lugar dentro de la biología de la reproducción de las aves, por su papel central en ecología reproductiva del grupo. Los huevos son extraordinariamente variables en tamaños, formas y patrones de coloración, sin embargo, sus descripciones y en particular la de sus formas, han sido históricamente narrativas y altamente subjetivas. Los estudios oológicos se han centrado históricamente en el análisis de las variaciones en el tamaño y el grosor de la cáscara (ej. Coulson et al., 1969; Ricklefs, 1984; Jover et al., 1993; Martin et al., 2006), ignorándose las formas cuya descripción es generalmente esporádica y narrativa, con pocas excepciones (ej.: Todd & Smart, 1984; Barta & Székely, 1997). En relación a las formas, suelen clasificarse en elípticos, cónicos, semicónicos, ovalados, etc. (Palmer, 1962; Harrison, 1978). Esto se debe a que, a pesar de tener los huevos formas relativamente simples, estas son muy difíciles de cuantificar o comparar. Se han seguido varios enfoques matemáticos, sobre todo aproximándolas a esferoides de revolución, pero no han recibido buena acogida (ej. Gemperle & Preston, 1955; Preston, 1968). Preston (1969) en un trabajo clásico de obligada consulta y cita en cualquier libro general de Ornitología, a partir del cálculo de tres índices básicos que reflejan diferencias supuestamente invariantes de la escala entre huevos, sugiere diferencias importantes entre grupos de aves a amplias escalas taxonómicas. Algunos autores han mencionado la posibilidad de que la forma del huevo fuese un subproducto de otras tendencias evolutivas asociadas al peso corporal o la anatomía interna del conducto reproductor (Gill, 1990), sin embargo, es muy probable que también sea el resultado de la selección natural, dada su relación con importantes parámetros reproductivos (Hoyt, 1976). La forma del huevo en cada especie aparece como un compromiso entre el volumen necesario para la producción de una cría, el tamaño de la nidada y la capacidad o área de incubación de los adultos. Así, la forma redondeada presenta la mayor relación superficie / volumen y es característica de aves de gran tamaño o robustas en las que, aunque el tamaño de la nidada sea grande, el volumen del huevo respecto al de los adultos es menor (ej: Galliformes). Esta forma, además, maximiza la conservación del calor y aumenta la resistencia estructural de la cáscara (Gill, 1990). Los huevos con forma alargada o semicónica son característicos de aves relativamente pequeñas, o sea, que ponen huevos relativamente más grandes y numerosos (ej. Charadriiformes). Los huevos puntiagudos tienden a rodar menos, en un arco menor dentro del nido, disminuyendo así la probabilidad de caer del nido. Generalmente, los huevos de las especies que nidifican alto, o cuyos nidos son menos cóncavos, presentan estas formas. En las aves de la familia Ardeidae, la forma de los huevos no ha sido suficientemente explorada, como igualmente sucede en la mayoría de los grupos

Denis & Olavarrieta

de aves. Un notable aporte lo hace Telfair (1983), quien describe cuantitativamente la variación de las formas de huevos de la Garza Ganadera (Bubulcus ibis), empleando categorías nominales (subelípticos, elípticos, ovales, fusiformes...) y con una detallada discusión de sus frecuencias de aparición en una muestra de casi 700 huevos de esta especie en Texas. Para esta familia, posteriormente, Ruiz et al. (1992) sugieren un alto nivel de isomorfía en los huevos, al analizar la constante K que relaciona el volumen con el producto de las dimensiones lineales. Esta isomorfía, referida a la constancia de las formas de los huevos dentro de la familia, apoyaría la idea de que este parámetro no refleja tendencias adaptativas diferenciables a este nivel taxonómico a pesar de las diferencias en tallas y estrategias ecológicas generales descritas entre sus miembros. La diferencia entre el valor empírico obtenido y la constante teórica, K = Pi/6, definida geométricamente por Preston (1968), es un indicador de otros parámetros de forma no definidos por los diámetros mayor o menor de los huevos. Posteriormente, Denis (2002), al comparar la razón entre los diámetros, como un índice de “forma”, entre siete especies, detecta diferencias estadísticas que atribuye a las diferencias en riesgos de caída en los nidos. Sin embargo, todos los estudios desarrollados hasta el presente se basan en las dimensiones lineales de los huevos, o índices derivados de estas, que no son indicadores adecuados de forma sino de proporción, y que se correlacionan con las dimensiones, lo cual puede distorsionar las interpretaciones biológicas. El desarrollo de la morfometría geométrica (Rohlf & Marcus, 1993) han sacado al discurso científico los problemas teóricos y metodológicos de emplear las dimensiones lineales clásicas para la caracterización de las formas, por medio de índices relativos. Todos los estudios existentes que describen las formas de los huevos han utilizado las proporciones entre diámetros mayores y menores del huevo, como indicadores de la forma, pero esto puede distorsionar o enmascarar las diferencias, ya que el propio concepto de "forma" en si mismo, excluye las dimensiones. La persistente correlación entre los índices de elongación y las medidas asociadas al tamaño evidencian estas limitaciones. Hoyt (1976) analizando el posible efecto de la forma sobre la superficie del huevo detecta una correlación significativa entre el índice superficie–volumen y los índices de forma: elongación y asimetría. Las tablas con los estadísticos descriptivos de dimensiones pueden mostrar similitudes numéricas enmascarando profundas diferencias de forma, y viceversa. Las variabilidades obtenidas para conjuntos de dimensiones no pueden ser analizadas en conjunto por los altos niveles de covarianzas internas. Los métodos de la morfometría geométrica analizan las formas basándose en el espacio de forma de Kendall (Kendall, 1977, 1984), o sus aproximaciones en el espacio tangente, en el cual las diferencias entre formas son evaluadas por las distancias procrustes. Un creciente número de trabajos demuestran el uso de este "aparato" matemático logra mayor potencia

Animal Biodiversity and Conservation 34.1 (2011)

estadística e impone menos restricciones en los patrones de variación que pueden ser detectados (ej.: Rohlf, 1999; 2000a, 2000b). Sin embargo, no existen antecedentes de su utilización al estudio de los huevos en las aves que, desde el punto de vista metodológico, se ha visto reducido a la determinación de dimensiones lineales y proporciones. El clásico trabajo de Preston (1969), en un notable ejemplo de previsión científica, menciona que posiblemente en el futuro, con el desarrollo de los métodos digitales se aceleraría considerablemente la captura de datos eliminándose la tediosa y muy trabajosa tarea de medir el enorme volumen de ejemplares existente. También, que las computadoras permitirían un tratamiento matemático más intenso y revelador que el que podían hacer a mediados del pasado siglo. Según sus propias palabras 'it is just possible that some day it may seem worthwhile to understand egg shapes more completely'. Su excelente trabajo, aunque careció de un profundo procesamiento estadístico por las limitaciones de la época, fue suficiente indicador de las posibilidades y vacíos de información que sobre el tema de los huevos de las aves existía en aquel momento. Sin embargo, medio siglo más tarde no se han hecho notables adelantos a pesar de que ya se han cumplido todas sus predicciones. En el presente trabajo se analizan las diferencias en formas puras de los huevos de ocho especies de la familia Ardeidae, partiendo de la hipótesis de que, por su potencia, los métodos propios de la morfometría geométrica pueden validar las diferencias interespecíficas detectadas por los índices lineales y así invalidar la isomorfía sugerida, reabriendo una posible línea de análisis ecomorfológicos casi abandonada. Para ellos se emplean dos de estos métodos relativamente novedosos: un método de contorneo, utilizando descriptores elípticos de Fourier, y puntos morfológicos claves (landmarks). Ambos métodos, aunque parten del mismo tipo de dato —ubicación relativa de puntos del contorno del huevo— se complementan, ya que emplean diferentes cantidades de puntos y se les da una importancia relativa diferente a cada uno. Los descriptores elípticos de Fourier son los coeficientes que se obtienen al calcular la combinación lineal de funciones senos o cosenos que mejor describen la función cíclica formada por los ángulos de salto entre un punto y otro de un contorno cerrado (funciones de Fourier). Para este cálculo se toman en cuenta todos los puntos de un contorno, sin discriminación ni información adicional relativa a su ubicación espacial, con lo cual hace énfasis en la magnitud de las diferencias pero no en su ubicación. Sin embargo, para el análisis por puntos morfológicos se utilizaron solo ocho puntos del contorno, ubicados en posiciones específicas (puntos extremos y diagonales), y que permiten una interpretación directa de la ubicación y sentido de las variaciones en forma. Material y métodos Se tomaron las dimensiones lineales (diámetro mayor y menor) y fotografías digitales de 167

huevos de siete especies de garzas depositados en las colecciones científicas del Instituto cubano de Ecología y Sistemática. Las especies incluidas, con las siglas empleadas para identificarlas y los tamaños de muestra, fueron: Butorides virescens (Bv, n = 20), Ixobrychus exilis (Ie, n = 14), Egretta thula (Eth, n = 36), Egretta tricolor (Etr, n = 24), Egretta caerulea (Ec, n = 16), Ardea alba (Aa, n = 11) y Nyctanassa violacea (Nv, n = 16). Se emplearon además, fotografías de 36 huevos de Bubulcus ibis (Bi) colectados en la colonia del Rincón de Guanabo, Ciudad de La Habana, en julio de 2008. Las fotos se tomaron con una cámara digital de 8 Mp sobre un fondo oscuro, para maximizar el contraste de los bordes del huevo. Las medidas lineales se tomaron con un pie de rey de 0,05 mm de precisión y con ellas se calculó el índice de elongación (mal llamado "índice de forma" por Denis, 2002), como la razón entre los diámetros menores y mayores de cada huevo. Para este cálculo se tomaron, además, las dimensiones de otros 283 huevos medidos durante estudios de campo (lo que aumentó la muestra de B. virescens, A. alba, E. tricolor, E. thula y B. ibis a 82 huevos). Como otra forma de ver la constancia de la forma en cada especie se determinó la magnitud de la correlación entre ambas variables: variaciones exactamente isométricas generarían máximas correlaciones. Las imágenes de cada huevo se procesaron digitalmente en el programa ImageJV1.34s (htp:// rsb.info/nih/gov/ij/), desaturándolas para eliminar la información de color pero manteniendo el formato RGB de ocho bits y manipulando el brillo y contraste para eliminar todo el ruido de la imagen y reducirla solo a la silueta. Cada imagen resultante fue descrita por un código de cadena (Freeman, 1974) y se determinaron los coeficientes de los descriptores elípticos de Fourier (DEF) (Lestrel, 1997) para ocho armónicos. Estos descriptores fueron normalizados sobre la elipse del primer armónico según el procedimiento de Kuhl & Giardina (1982) para eliminar las posibles diferencias relacionadas con la talla, rotación y puntos iniciales de la traza del contorno. A partir de este procedimiento, la forma fue descrita por 32 coeficientes normalizados de Fourier, cuya información fue resumida por un análisis de componentes principales basado en la matriz de varianza–covarianza de los coeficientes y los puntajes de los primeros componentes fueron empleados para las comparaciones interespecíficas subsiguientes. La variación en la forma en cada especie fue representada utilizando la transformación de Fourier inversa (Rohlf & Archie, 1984; Furuta et al., 1995). Todo este procesamiento y análisis se realizó con el programa SHAPE v.5. (Iwata & Ukai, 2002). Además del análisis del contorno se superpuso a cada imagen un sistema de radios equiangulares (cada 45o) con el origen en la mitad del eje mayor del huevo y se digitalizaron, por una misma persona para evitar sesgos, puntos clave en las ocho intersecciones con el contorno utilizando el programa APS–Dig. Las configuraciones de puntos fueron superpuestas con un registro de Bookstein,

Denis & Olavarrieta

Tabla 1. Dimensiones de los huevos de siete especies de garzas (Aves, Ardeidae) empleados en el análisis de la forma de los huevos (media ± EE, entre paréntesis min.–máx.). Table 1. Measurements of eggs of eight species of egrets and herons (Aves, Ardeidae) used in the egg shape analysis (mean ± SE, in brackets min.–max.).

Especie

Butorides virescens

Egretta thula

E. tricolor

E. caerulea

Ardea alba

Ixobrychus exilis

Nyctanassa violacea

Bubulcus ibis

Diámetro mayor (mm)

Diámetro menor (mm)

Diferencia (%)

37,65 ± 0,45

29,03 ± 0,21

22,7

(33,50–40,60)

(26,70–30,40)

(14,7–33,3)

45,10 ± 0,32

32,96 ± 0,15

26,8

(42,00–50,10)

(31,20–34,90)

(22,62–33,73)

45,00 ± 0,42

32,11 ± 0,16

28,5

(40,55–49,50)

(30,30–33,55)

(20,59–34,84)

44,45 ± 0,42

32,49 ± 0,27

26,8

(42,00–47,30)

(30,20–33,90)

(20,48–33,40)

53,27 ± 0,39

39,89 ± 0,25

25,1

(51,40–55,30)

(38,90–41,80)

(22,57–28,00)

30,08 ± 0,48

23,15 ± 0,12

22,8

(27,40–33,40)

(22,50–23,90)

(17,02–28,44)

51,12 ± 0,50

38,45 ± 0,35

24,7

(46,20–54,30)

(35,70–40,50)

(18,35–30,57)

44,5 ± 0,35

32,1 ± 0,17

38,6

(43,8–45,2)

(31,7–32,4)

(23,22–51,13)

empleando como línea base al diámetro mayor. Con los consensos por especie se realizó un análisis de curvaturas principales en el programa PAST v1,75 y se calculó la energía de curvatura. Los diagramas de distorsiones, el tamaño del centroide y las distancias Procrustes entre especies se obtuvieron en el programa TPSSpline, tomando a una elipse perfecta como configuración de referencia para todos los casos. Procesamiento estadístico Se utilizaron pruebas de Kruskal–Wallis o ANOVAs de clasificación simples para las comparaciones entre especies, en dependencia del cumplimiento de las asunciones de normalidad y homogeneidad de varianzas. Se empleó la prueba de Tukey o una prueba de rangos múltiples para los contrastes a posteriori en caso necesario. De esta manera se compararon entre especies los índices de elongación, los puntajes de los componentes principales de los coeficientes de Fourier, las energías de curvaturas y los tamaños del centroide entre especies. Con los puntajes del ACP, además, se realizó un análisis de función discriminante, utilizando la distancia Euclidiana, empleando cada vez un componente más hasta que se incluyeran los que describían el 100% de la varianza.

Se hicieron correlaciones de Pearson entre los índices de elongación de los huevos y el diámetro mayor en cada especie y las distancias procrustes se utilizaron en un análisis de agrupamientos utilizando el método de Ward como estrategia de agrupamiento. Todo el procesamiento estadístico se realizó en el programa Statistica v8.0. Resultados Las dimensiones de los huevos en las especies analizadas (tabla 1) se encuentran dentro del rango previamente reportado en poblaciones cubanas (Denis et al., 2001; Denis, 2002). Los huevos son diferentes entre especies, en relación directa con el tamaño corporal, tal y como ha sido descrito en general para las aves. El índice de elongación "clásico" mostró diferencias entre especies (fig. 1), lo cual apuntaría a diferencias en las formas de los huevos de estas especies. Estos índices permiten separar las especies en tres grupos generales, aunque la alta variabilidad en algunas especies refleja una posible falta de potencia estadística para alcanzar la significación en algunos casos. El grupo de mayor índice está formado por las dos especies de hábitos más solitarios y de menor talla, B. virescens e I. exilis, y por E. thula,

Animal Biodiversity and Conservation 34.1 (2011)

0,80

Índice de elongación

0,79

(F = 9,18; p < 0,001)

± ES ± LC 95%

0,78 0,77 0,76

0,75 0,74

Media

0,73 0,72

Ie Bv Nv Aa Etr Eth Ec Bi (n = 14) (n = 82) (n = 16) (n = 82) (n = 82) (n = 82) (n = 16) (n = 82) Especies

Fig. 1. Índices de elongación del huevo de ocho especies de garzas (Aves, Ardeidae) (las letras indican diferencias significativas según la prueba de Tukey). (Para las abreviaturas de las especies ver Material y métodos.) Fig. 1. Elongation indexes of eggs in eight egret and heron species (Aves, Ardeidae) (letters indicate statistical differences according to Tukey’s post hoc test). (For species abbreviations see Material y métodos.)

con valores medios superiores a 0,745. Un grupo intermedio, "indiferenciado" estadísticamente, formado por N. violacea, E. tricolor y B. ibis, y un tercer grupo con los mínimos valores medios del índice, es decir, con los huevos más alargados, formado por A. alba y E. caerulea. En todas las especies el índice de elongación mantiene una correlación significativa con el diámetro mayor del huevo, quien a su vez está fuertemente correlacionado con el diámetro menor (fig. 2). Para analizar la forma, eliminando los efectos del tamaño se emplearon los descriptores elípticos de Fourier, con ocho armónicos, y los 32 coeficientes obtenidos se redujeron por un análisis de componentes principales, cuyos primeros cuatro componentes explicaron más del 70% de la variación (tabla 2). Como todos los coeficientes de Fourier tienen la misma importancia y reflejan sutiles diferencias en secciones de arco de la silueta del huevo se requieren 13 componentes para llegar a explicar el 95% de la variabilidad total. Los puntajes de estos componentes para cada ejemplar de huevo pueden ser empleados como si fueran una variable morfométrica "tradicional". Al comparar entre especies los puntajes de los tres primeros componentes, se obtuvieron diferencias significativas (fig. 3). La prueba de Kruskal–Wallis y la prueba no paramétrica de comparación múltiple

de tendencias centrales permiten establecer algunas diferencias interespecíficas. Sin embargo, el análisis en general evidencia baja potencia, reflejada en los grupos pareados. El primer componente, que explica cerca del 30% de la variación de los DEF, separa sin dificultad a E. tricolor, del grupo formado por I. exilis, B. virescens y E. thula. Los puntajes del segundo componente, que explica un 23% de la variación, ortogonal al primer componente, solo permite diferenciar a E. thula de B. ibis. En las demás especies los puntajes son similares entre si. Los puntajes del tercer componente, que explica un 10,8% de la variación, sirven para diferenciar a E. tricolor de B. virescens. A partir del cuarto componente, los puntajes no muestran diferencias estadísticas entre especies (Kruskal–Wallis, H(7, N = 174) = 13,65; p = 0,06). Al realizar un análisis discriminante empleando simultáneamente los puntajes de los componentes principales como variables, se verifica que, en general, la forma de los huevos no es suficiente para diferenciar efectivamente todas las especies de garzas. Si se emplean los primeros tres componentes principales, aquellos que mostraron diferencias interespecíficas en el análisis univariado, se obtienen porcentajes de clasificación correcta muy bajos. Incorporando secuencialmente los demás componentes principales, el error de clasificación va mejorando ligeramente, pero aún

Denis & Olavarrieta

0,88

Bv Ie Ec Eth Etr Aa Nv Bi

0,86 Índice de elongación

0,84 0,82 0,80 0,78 0,76

r = –0,87, p < 0,001 r = –0,96, p < 0,001 r = –0,67, p = 0,004 r = –0,78, p < 0,001 r = –0,88, p < 0,001 r = –0,73, p = 0,011 r = –0,59, p = 0,016 r = –0,75, p = 0,001

0,74 0,72 0,70 0,68 0,66 0,64 0,62

40 45 50 55 60 Diámetro mayor (mm)

Fig. 2. Relaciones entre el índice de elongación y el diámetro mayor del huevo en ocho especies de garzas (Aves, Ardeidae). (Para las abreviaturas de las especies ver Material y métodos.) Fig. 2. Relationship between elongation index and egg breadth in eight species of egret and heron (Aves, Ardeidae). (For species abbreviations see Material y métodos.)

con los componentes que explican el 100% de la variación, el error de clasificación es de más de un 40%. El método de los DEF parte de la selección de armónicos a partir de un punto aleatorio de la silueta, dándole la misma importancia a todos los puntos, por lo que no permite una interpretación directa del lugar donde se acumulan las variabilidades. Por esta razón, se empleó la variante de utilizar puntos claves equiangulares como otra forma de describir mejor la forma. Cuando se superponen los puntos obtenidos con las imágenes se observa una elevada constancia en la forma promedio de los huevos, evidenciada con las semejanzas en la energía de curvatura al registrarlas contra un círculo perfecto (fig. 4). A pesar de ello, es posible detectar diferencias significativas entre las garzas medianas (B. ibis, E. tricolor y E. thula) contra las garzas chicas (B. virescens e I. exilis). El tamaño del centroide, definido como la sumatoria de las distancias cuadradas entre el punto central y cada punto de la conformación, también es un indicador de la forma, al haberse eliminado las diferencias en tallas por una superposición de Bookstein. A mayor tamaño del centroide, más cercana es la forma a una circunferencia, dando además la posibilidad de analizar la variabilidad en las formas. Un análisis de varianzas detectó diferencias significativas entre los tamaños del centroide entre especies (F = 7,99; p < 0,0001) (fig. 5) y la prueba de Tukey identificó nuevamente las diferencias entre

Tabla 2. Resultados del Análisis de Componentes Principales (ACP) con los coeficientes de Fourier que describen la forma de los huevos de ocho especies de garzas (Aves, Ardeidae): E. Eigenvalue; P. Proporción (%); PA. Proporción acumulada (%). Table 2. Results of the Principal Component Analysis ACP with Fourier coefficient describing egg shape in eight species of egret and heron (Aves, Ardeidae): E. Valores propios; P. Proporción (%); PA. Proporción acumulada (%). ACP

–5

CP1

8,76 x 10

29,28

CP2

6,81 x 10

–5

22,74

52,02

CP3

3,23 x 10

10,78

62,81

CP4

2,55 x 10–5

8,52

71,33

CP5

1,93 x 10–5

6,44

77,73

CP6

1,31 x 10

4,37

82,13

CP7

1,15 x 10

3,86

85,99

CP8

8,50 x 10–5

2,84

88,83

CP9

5,14 x 10

1,72

90,55

–5

–5 –5

–5

Animal Biodiversity and Conservation 34.1 (2011)

Kruskal–Wallis: H(7, N=174) = 42,74; p < 0,001

PC1

0,010 0,008 0,006 abc 0,004 0,002 0,000 bc –0,002 –0,004 –0,006 –0,008 –0,010 –0,012

ab ab a

Etr

abc Aa

ab bc Bv, Eth Bi, Nv, Ec

c Ie

Kruskal–Wallis: H(7, N=174) = 15,10; p = 0,04

PC2

0,010 0,008 0,006 ab 0,004 0,002 0,000 –0,002 –0,004 –0,006 –0,008 –0,010 –0,012

ab ab

Eth

Bv, Etr Ie, Nv, Ec

PC3

Kruskal–Wallis: H(7, N=174) = 15,81; p = 0,03 0,010 0,008 ab 0,006 0,004 0,002 0,000 –0,002 –0,004 –0,006 –0,008 –0,010 –0,012 Aa

a b

Media

Ie Etr Eth Nv Bi Especies ± ES

± LC

Etr

Eth, Bi Ie, Nv, Ec

Reconstrucción de la silueta con los puntajes de cada componente

Fig. 3. Comparación entre especies de los puntajes medios calculados por los tres componentes principales que más variabilidad explican en la forma de los huevos de ocho especies de garzas (Aves, Ardeidae), a través de los DEF (las letras indican diferencias significativas según la prueba no paramétrica de comparación múltiple de medias). (Para las abreviaturas de las especies ver Material y métodos.) Fig. 3. Interspecific comparison of mean scores calculated using the three main components that account for highest variability in egg shape of eight species of heron and egret (Aves, Ardeidae), through EFD (letters indicate statistical differences according to multiple comparison of mean ranks). (For species abbreviations see Material y métodos.)

Denis & Olavarrieta

Butorides virescens (Ec = 2,773E–02, CV = 66,87)

Egretta caerulea (Ec = 4,388E–02, CV = 28,33)

Bubulcus ibis (Ec = 4,240E–02, CV = 28,45)

Ardea alba (Ec = 4,387E–02, CV = 19,79)

Egretta tricolor (Ec = 5,140E–02, CV = 26,77)

Egretta thula (Ec = 4,561E–02, CV = 24,56)

Nyctanassa violacea (Ec = 3,746E–02, CV = 24,86)

Ixobrychus exilis (Ec = 3,033E–02, CV = 28,86)

Fig. 4. Diagramas de distorsión y energías de curvaturas de las siluetas consensos de los huevos de ocho especies de garzas (Aves, Ardeidae) (las letras indican las diferencias según la prueba de Tukey). Fig. 4. Distortion grid of egg consensus silhouettes and curvature energy in eight species of egret and heron (Aves, Ardeidae) (letters indicate statistical differences according to Tukey’s test).

las garzas más pequeñas y las medianas. El patrón de semejanzas se pone de manifiesto de una forma evidente en un análisis de agrupamiento empleando las distancias procrustes entre todas las configuraciones de puntos (fig. 5). Discusión La forma de los huevos es un elemento más de la amplia diversidad funcional que presentan los huevos de las aves. La forma debe ser independiente de las dimensiones y para lograr esa independencia es que históricamente se han empleado los índices relativos, usualmente la razón entre diámetros mayores y menores —índice de elongación— o entre los ángulos de curvaturas de los extremos, para su descripción. Sin embargo, la correlación existente entre las dimensiones y el índice de elongación sugiere que este último indicador no es ortogonal con las dimensiones lineales. Por esta razón, el empleo de este índice puede distorsionar las diferencias entre especies ya que el propio concepto de "forma" en si mismo, es totalmente independiente de las dimensiones. Por ello, las diferencias detectadas entre especies pueden

simplemente estar reflejando las diferencias en tamaño de los huevos, lo cual si depende directamente del tamaño corporal adulto. Estas correlaciones fueron identificadas y llamadas "contaminación matemática" por Preston (1969). Los valores de índices de elongación en los huevos de estas especies, registrados en la literatura, varían ligeramente entre localidades: Preston (en Palmer, 1962) estima una elongación de 0,75 en E. caerulae y de 0,72 en A. alba. Hoyt (1976) determinó para Ardea herodias una elongación de 0,68. Para la familia Ardeidae completa, Preston (1969) determinó, empleando 19 especies o subespecies, un índice de elongación promedio de 0,73 (0,70–0,77). Estas diferencias geográficas indican una determinación ambiental de la forma, al menos parcialmente. En la Paloma (Columba livia) se han encontrado variaciones en la forma de los huevos a lo largo de amplio rango de localidades geográficas (Janiga, 1996, 1997). Ruiz et al. (1992) sugieren un alto nivel de isomorfía en los huevos de las especies de la familia Ardeidae, al comparar entre especies la constante K, que relaciona el volumen con el producto de las dimensiones lineales, por medio de la comparación de las pendientes de la regresión del volumen con las dimensiones lineales.

Distancia procrustes (método de Ward)

Animal Biodiversity and Conservation 34.1 (2011)

0,045 0,040 0,035 0,030 0,025 0,020 0,015 0,010 0,005 0,000 1,26

Media

Tamaño del centroide

1,25

± ES

1,24

± LC 95%

1,23

1,22 1,21

b b

1,20 1,19 1,18 1,17

Aa Etr Especies

Eth

Fig. 5. Comparación entre los tamaños del centroide en los huevos de ocho especies de garzas (Aves, Ardeidae) y diagrama de agrupamientos formados con las distancias procrustes entre las conformaciones de puntos consensos por especie (las letras indican las diferencias según la prueba de Tukey). (Para las abreviaturas de las especies ver Material y métodos.) Fig. 5. Comparison between centroid size in eggs from eight species of egrets and herons (Aves, Ardeidae) and cluster diagram with procrustes distances between consensus point configurations by species (letters indicate statistical differences according to Tukey’s test). (For species abbreviations see Material y métodos.)

Denis (2002) detecta diferencias interespecíficas al emplear un índice de proporción, el cual a partir de una muestra considerable (entre 39 y 563 huevos por especie) y con baja variabilidad (CV entre 4,2 y 7,8%) sugiere que Butorides y Egretta thula ponen los huevos más redondeados y Egretta rufescens y Bubulcus más alargados. Nycticorax, Ardea y Egretta tricolor presentan proporciones muy similares entre sí. Basándose en estos resultados, sugiere que los huevos en esta familia no son de igual forma y explica las variaciones en función de las características de los sitios de nidificación. Posteriormente, Denis et al. (2009) recalculan constantes volumétricas especie–específicas para siete especies de esta familia, y aunque no comparan estadísticamente las constantes propuestas entre especies sus resultados también sugieren diferencias entre especies.

Los resultados del presente trabajo demuestran que, si bien no puede asegurarse que los huevos de todas las especies de la familia Ardeidae sean exactamente isomórficos, tampoco hay evidencias para que la forma por si sola pueda discriminar entre especies. A pesar de su potencia, los métodos de morfometría geométrica no pueden separar exactamente las especies en función de la forma de los huevos como se pensaba inicialmente, y se mantiene un alto nivel de superposición entre ellas en las variables de forma empleadas. La forma de los huevos puede servir para separar con mayor o menor eficiencia las garzas solitarias pequeñas (Butorides e Ixobrychus) del conjunto de las especies de garzas coloniales medianas y de las garzas grandes. Las variables de forma pura muestran un cambio gradual entre especies con un ordenamiento un tanto diferente al sugerido por

los índices lineales. La especie con huevos menos redondeados es Egretta tricolor a diferencia de lo que indica el índice lineal que apunta hacia Bubulcus ibis. Los huevos de esta especie han sido descritos como "Ovate to oval; elliptical ovate or elliptical oval" mientras que los de las otras especies han sido descritos como "Elliptical to subelliptical" (Bent, 1926; Sprunt, 1954; Palmer, 1962; Frederick, 1997). Esta diferencia refleja lo mencionado acerca de las imprecisiones del uso de índices lineales de proporción como indicadores de forma. En este caso, las proporciones son menores en Bubulcus pero la forma no, por la ubicación relativa más centrada y simétrica de la región más ancha del huevo, que en E. tricolor está más desplazada hacia un polo, alejando más su forma de la de un círculo. Ahora bien, estas diferencias interespecíficas dejan abiertas numerosas interrogantes en relación a la función posible de las formas en los huevos. Se ha mencionado que los huevos con forma esférica maximizan la conservación del calor, la resistencia de la cáscara y el ahorro de los materiales en la formación de la misma (Gill, 1990). Según Grant (1982), si la forma del huevo es adaptativa, su significación debe relacionarse con la incubación pero Hoyt (1976), al analizar el posible efecto de la forma sobre la superficie del huevo, había notado que es poco probable que la forma del huevo se relacione con ningún intercambio con el ambiente. Este autor se basa en que, ya que la densidad de poros, su área y el grosor de la cáscara son independientes de la forma, no debe existir un efecto significativo sobre el intercambio de gases. Y el efecto de la variabilidad del área superficial sobre el intercambio térmico también es mínimo, en relación a los efectos de las características del microhábitat de nidificación, del nido o de los patrones de incubación. Este autor concluye que aún permanece como un acto de fe la creencia de que la forma del huevo es derivada de selección natural en base a aspectos funcionales de la misma. Su posible efecto sobre el volumen de la puesta y su posible ventaja energética también parece ser poco probables, por su pequeña magnitud en relación a la influencia de los patrones en variación en la talla. Las diferencias geográficas intraespecíficas en la forma de los huevos de otras especies, reflejadas en la literatura, apuntan a algún tipo de efecto ambiental sobre esta variable, por lo cual pueden buscarse los factores selectivos en aspectos ecológicos relacionados con el hábitat de nidificación y las probabilidades de caída del nido, como se ha observado en comparaciones interespecíficas. Los huevos más puntiagudos tienden a rodar menos, en un arco menor que los huevos que tienen forma redondeada, disminuyendo así la probabilidad de caer o alejarse del nido. Generalmente, las especies que nidifican alto, o cuyos nidos son menos cóncavos, tienden a presentar estas formas. Permanecen por demostrar los posibles efectos de las diferencias en forma sobre las ventajas estructurales (resistencia a la tensión), que tal vez ayuden a explicar diferencias entre grupos taxonómicos superiores. El estudio de la variación en las formas de los huevos dentro de una nidada también ha mostrado la existencia de patrones como, por ejemplo, que en el Galleguito (Larus atricilla) el radio de curvatura del

Denis & Olavarrieta

polo mayor y la asimetría del huevo, tienden a ser más pequeños en el último huevo de la nidada (Preston & Preston, 1953). Esto es poco probable partiendo de efectos ambientales y debe relacionarse con las estrategias reproductivas de alguna otra forma. Por tanto, a pesar de los esfuerzos realizados, se mantiene el señalamiento de Hoyt (1976) de que aún dista de estar clara la significación ecológica o evolutiva de estos patrones encontrados en los huevos. Agradecimientos Los autores desean reconocer al Dr. Xavier Ruiz, Catedrático de Zoología de la Universidad de Barcelona, recientemente fallecido y autor de la comunicación original sobre la isomorfía en los huevos de las garzas, por haber generado la idea básica de este trabajo y por su continuo apoyo a las investigaciones sobre aves acuáticas en Cuba. Referencias Bent, A. C., 1926. Life histories of North American marsh birds. U.S. Natl. Mus. Bull., 135: 1–392. Coulson, J. C., Potts, G. R. & Horobin, J., 1969. Variation in the eggs of the Shag (Phalacrocorax aristotelis). Auk, 86(2): 232–245. Barta, Z. & Székely, T., 1997. The optimal shape of avian eggs. Functional Ecology, 11: 656–662. Denis, D., 2002. Ecología reproductiva de siete especies de garzas (Aves: Ardeidae) en la ciénaga de Birma, Cuba. Tesis doctoral, Univ. de La Habana, Cuba. Denis, D., Olavarrieta, U. & Andraca, L., 2009. Constantes de Hoyt para estimar el volumen de los huevos en garzas cubanas (Aves: Ciconiiformes). Biología, 22(1–2): 75–77. Denis, D., Rodríguez, P., Rodríguez, A. & Torrella, L., 2001. Ecología reproductiva de tres especies de la familia Ardeidae. Biología, 15(1): 27–36. Frederick, P. C., 1997. Tricolored Heron. In: The Birds of North America, 306: 1–28 (A. Poole & F. Gill, Eds.). Acad. Nat. Sci., Philadelphia, PA, and Am. Ornithol. Union, Washington, D.C. Freeman, H., 1974. Computer processing of line drawing images. Comp. Surv., 6: 57–97. Furuta, N., Ninomiya, S., Takahashi, S., Ohmori, H. & Ukai, Y., 1995. Quantitative evaluation of soybean (Glycine max L. Merr.) leaflet shape by principal component scores based on elliptic Fourier descriptor. Breed. Sci., 45: 315–320. Gemperle, M. E. & Preston, F. W., 1955. Variation of shape in the eggs of the common tern in the clutch–sequence. Auk, 72: 184–198. Gill, F. B., 1990. Ornithology. W. H. Freeman y Company, New York. Grant, P. R., 1982. Variation in the size and shape of darwin’s finch eggs. Auk, 99: 15–23 Harrison, C., 1978. A field guide to the nests, eggs, and nestlings of North American birds. Collins, London, United Kingdom.

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Hoyt, D. F., 1976. The effect of shape on the surface–volume relationships of birds’ eggs. Condor, 78: 343–349. Iwata, H. & Ukai, Y., 2002. SHAPE: A computer program package for quantitative evaluation of biological shapes based on elliptic Fourier descriptors. Journal of Heredity, 93: 384–385. Janiga, M., 1996. Variation in size and shape of eggs of the Feral Pigeon (Columba livia). Folia Zoológica, 45(4): 301–310. – 1997. Effects of geographic variation and hatching asynchrony on size and shape of eggs of the feral pigeon (Columbia livia). Folia Zoológica, 46(1): 23–32. Jover, L., Ruiz, X. & González–Martín, M., 1993. Significance of intraclutch egg size variation in the Purple Heron. Ornis Scandinavica, 24(2): 127–134. Kendall, D. G., 1977. The diffusion of shape. Advances in Applied Probability, 9: 428–430. – 1984. Shape manifolds, Procrustean metrics and complex projective spaces. Bulletin of the London Mathematical Society, 16: 81–121. Kent, J. T., 1997. Data analysis for shapes and images. Journal of Statistical Inference and Planning, 57: 181–193. Kuhl, F. P. & Giardina, C. R., 1982. Elliptic Fourier features of a closed contour. Computer Graphics and Image Processing, 18: 236–258. Lestrel, P. E., 1997. Fourier descriptors and their applications in biology. Cambridge Univ. Press. Martin, T. E., Bassar, R. D., Bassar, S. K., Fontaine, J. J., Lloyd, P., Mathewson, H. A., Niklison, A. M. & Chalfoun, A., 2006. Life–history and ecological correlates of geographic variation in egg and clutch mass among passerine species. Evolution, 60(2): 390–398. Palmer, R. S., 1962. Handbook of North American Birds. I. Yale Univ. Press, New Haven, Connecticut.

Preston, F. W., 1953. The shapes of birds’ eggs. Auk, 70: 160–182. – 1968. The shapes of birds’ eggs: mathematical aspects. Auk, 85: 454–463. – 1969. Shapes of birds’ eggs: extant North American families. Auk, 86: 246–264. Preston, F. W., & Preston, E. J., 1953. Variation of birds’ eggs within the clutch. Ann. Carnegie Mus., 33: 129–139. Ricklefs, R. E., 1984. Variation in the size and composition of eggs of the European Starling. Condor, 86: 1–6. Rohlf, F. J., 1999. Shape statistics: Procrustes superimpositions and tangent spaces. Journal of Classification, 16: 197–223. – 2000a. On the use of shape spaces to compare morphometric methods. Hystrix, 11(1): 8–24. – 2000b. Statistical power comparisions among alternative morphometric methods. American Journal of Physical Anthropology, 111: 463–478. Rohlf, F. J. & Archie, J. W., 1984. A comparison of Fourier methods for the description of wing shape in mosquitoes (Diptera: Culicidae). Syst. Zool., 33: 302–317. Rohlf, F. J. & Marcus, L. F., 1993. A revolution in morphometrics. Trends Ecol. Evol., 8: 129–132. Ruiz, X., Petriz, J. & Jover, L., 1992. Estimating egg volumes from linear dimensions: isomorphy in eggs belonging to the family Ardeidae. Misc. Zool., 16: 254–257. Sprunt, A., Jr., 1954. Florida birdlife. Coward–McCann, New York. Telfair, R. C. II., 1983. The Cattle Egret: a Texas focus and world view. Kleberg Stud. Nat. Resour. Tex. Agric. Exp. Stn., Texas A. & M. Univ. Todd, P. H. & Smart, I. H. M., 1984. The shape of birds’ eggs. Journal of Theoretical Biology, 106: 239–243.

Editor executiu / Editor ejecutivo / Executive Editor Joan Carles Senar

Secretaria de Redacció / Secretaría de Redacción / Editorial Office

Secretària de Redacció / Secretaria de Redacción / Managing Editor Montserrat Ferrer

Museu de Zoologia Passeig Picasso s/n 08003 Barcelona, Spain Tel. +34–93–3196912 Fax +34–93–3104999 E–mail mzbpubli@intercom.es

Consell Assessor / Consejo asesor / Advisory Board Oleguer Escolà Eulàlia Garcia Anna Omedes Josep Piqué Francesc Uribe

Animal Biodiversity and Conservation 34.1 (2011)

Variación morfológica de las especies de Astyanax, subgénero Zygogaster (Teleostei, Characidae) R. I. Ruiz–C., C. Román–Valencia, B. E. Herrera–M., O. E. Peláez & A. Ermakova–A.

Ruiz–C., R. I., Román–Valencia, C., Herrera–M., B. E., Peláez, O. E. & Ermakova–A., A., 2011. Variación morfológica de las especies de Astyanax, subgénero Zygogaster (Teleostei, Characidae). Animal Biodiversity and Conservation, 34.1: 47–66. Abstract Morphological variation of Astyanax species, subgenus Zygogaster (Teleostei, Characidae).— The diverse Neotropical fish genus Astyanax inhabits a variety of aquatic environments. As with other species in this genus, the taxonomic status and phylogenetic relationships of species of this subgenus remain largely undetermined. Based on 354 individuals, we analyzed the morphological variation of four species of the subgenus Zygogaster (A. atratoensis, A. caucanus, A. filiferus, and A. magdalenae) using procrustes analysis and compared findings with two species of the sister group: subgenus Poecilurichthys (A. orthodus y A. superbus). The PCA (Principal Component Analysis) and CVA (Canonical Variates Analysis) showed morphological affinity between the subgenera and indicated variance in body depth, anterior trend of dorsal fin origin and humeral spot, depression on the dorsal surface of the skull, and ventral displacement of the orbit and snout. The variation in these structures may provide evidence supporting adaptive speciation as an alternative to speciation driven by geographical isolation. Key words: Astyanax, Characid fish, Morphogeometry, Disparity, Colombia. Resumen Variación morfológica de las especies de Astyanax, subgénero Zygogaster (Teleostei, Characidae).— Astyanax es un género diverso de peces neotropicales, cuyas especies habitan gran variedad de ambientes acuáticos. La situación taxonómica de los subgéneros y de sus especies, no difiere de la problemática que presenta el género Astyanax. Basándonos en 354 individuos, se analizó la variación morfológica de cuatro especies del subgénero Zygogaster (A. atratoensis, A. caucanus, A. filiferus y A. magdalenae) mediante un análisis morfogeométrico comparado con dos especies del grupo hermano Poecilurichthys (A. orthodus y A. superbus). El ACP (Análisis de Componentes Principales) y AVC (Análisis de Variables Canónicas) evidenciaron afinidad morfológica entre los subgéneros e indicaron varianza en la profundidad del cuerpo, tendencia anterior del origen de la aleta dorsal y mancha humeral, depresión sobre la superficie dorsal del cráneo, y desplazamiento ventral de la orbita y el hocico. La variación entre las especies indicó aislamiento del tercer infraorbital del preopérculo y protrusión del extremo ventral del maxilar. La variación en estas estructuras evidencia especiación adaptativa como posible alternativa a la especiación por aislamiento geográfico. Palabras clave: Astyanax, Characido, Morfogeometría, Disparidad, Colombia. (Received: 24 V 10; Conditional acceptance: 10 X 10; Final acceptance: 21 XII 10) R. I. Ruiz–C., C. Román–Valencia, B. E. Herrera–M., O. E. Peláez, Lab. de Ictiología, Univ. del Quindío, A. A. 2639, Armenia, Quindío, Colombia.– A. Ermakova–A., Dept. of Biology, Univ. of Edinburgh, UK. Correponding author: R. I. Ruiz–C. E–mail: zutana_1@yahoo.com

ISSN: 1578–665X

Introducción Astyanax es un género diverso de peces neotropicales, cuyas especies habitan gran variedad de ambientes acuáticos (Román–Valencia & Ruiz–C., 2005; Ruiz–C. & Cipriani, 2006; Ruiz–C., 2010). El género incluye alrededor de 130 especies válidas (Eschmeyer, 2010), distribuidas desde el sur de los Estados Unidos hasta el río Negro en el norte de la Patagonia en Argentina (Almirón et al., 1997). Astyanax fue discutido por Eigenmann (1917) como un grupo con morfología generalizada en Characidae, e identificado como el núcleo en la clasificación radial propuesta. En la redescripción del género Astyanax (Eigenmann, 1921) se plantea tres subgéneros: 1) Astyanax en el sentido estricto, caracterizado por presentar una serie continua de escamas sobre el área predorsal, incluye la especie tipo del género: Astyanax mexicanus (De Filippi, 1853); 2) Poecilurichthys con el área predorsal sin serie continua de escamas; y 3) Zygogaster cuya área preventral es comprimida, comparte con Poecilurichthys la ausencia de una serie continua de escamas sobre el área predorsal. La situación taxonómica de estos subgéneros y de sus especies, no difiere de la problemática que presenta en general el género Astyanax (Ruiz–C., 2010). A partir de descripciones iníciales para especies que conforman el grupo, se ha resaltado su difícil identificación, en especial entre las más relacionadas por su similaridad morfológica y solapamiento de caracteres útiles en su reconocimiento; dado el rampante paralelismo de caracteres externos y osteológicos observados en Characidae (Mirande, 2010). Cinco especies son reconocidas en el subgénero Zygogaster (Eigenmann, 1921), descritas para drenajes trasandinos: A. stilbe (Cope, 1870), A. caucanus (Steindachner, 1879), A. atratoensis Eigenmann, 1907, A. magdalenae Eigenmann & Henn 1916 y A. filiferus (Eigenmann, 1913). De éstas especies cuatro (A. magdalenae, A. atratoensis, A. caucanus y A. filiferus) presentan localidad tipo en Colombia, mientras sólo A. stilbe fue descrita para el río Pará en Brazil. Las especies de Zygogaster se caracterizan, al igual que las especies de Poecilurichthys, por presentar un modelo de pigmentación críptico (Ruiz–C., 2010). La semejanza morfológica puede ser producto del aislamiento geográfico reciente y de sus relaciones de parentesco (González–Díaz et al., 2005). Sin embargo, existen evidencias de que las formas se encuentran asociadas a condiciones dadas por el ambiente en que habitan; así, especies gregarias presentes en ambientes lénticos poseen mayor profundidad del cuerpo y engrosamiento en el pedúnculo caudal, en relación a especies no gregarias, que habitan la corriente del cauce (Román–Valencia & Ruiz––C., 2005; Ruiz–C., 2010) La morfometría geométrica es una herramienta para distinguir entre forma y talla (Bookstein, 1989; Zelditch et al., 2004). Permite comparar organismos mediante estructuras homólogas, cuantificar la variabilidad de la forma en múltiples escalas espaciales y estudiar las posibles correlaciones entre la variabilidad y otros parámetros morfológicos y/o ambientales

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(Bookstein et al., 1999; Ruiz–C. & Cipriani, 2006; García–Alzate et al., 2010). Conocer la información obtenida del análisis de la forma en las especies de Zygogaster, así como la existencia de alguna relación morfológica con otras especies, posiblemente relacionadas por caracteres externos, incluidas en Poecilurichthys, es útil en la medida que permite identificar probables estrategias evolutivas relacionadas con la forma, en grupos exitosos como los subgéneros de Astyanax, dada la alta diversidad y riqueza descrita para éste género. Evaluar la disparidad morfológica expresada en especies distribuidas en simpatría, como sucede con en la mayoría de las especies de Zygogaster (A. caucanus, A. filiferus y A. magdalenae), respecto a la morfología de A. atratoensis, distribuida en alopatría respecto a éstas, podría proporcionar una idea sobre su evolución en dichas condiciones. Con base en lo anterior, el objetivo del presente estudio fue realizar un análisis morfológico comparado, mediante morfometría geométrica, entre especies del subgénero Zygogaster y algunas del subgénero Poecilurichthys, presentes en Colombia. Material y métodos Los datos analizados en éste trabajo incluyen cuatro especies del subgénero Zygogaster: A. magdalenae, A. atratoensis, A. caucanus y A. filiferus y dos del subgénero Poecilurichthys: A. orthodus y A. superbus (fig. 1), con localidad tipo en Colombia, excepto A. superbus (fig. 2); las observaciones corresponden a diversas poblaciones para las especies analizadas (tabla 1). Las acronimias de los museos siguieron a Sabaj–Perez (2010), con adición de la Colección de Ictiología, Instituto de Biología de la Universidad de Antioquia, Medellín, Colombia (CIUA). Un método generalizado de superposición fue usado en la obtención de una matriz de coordenadas procrustes, requerida en el análisis morfo geométrico de las especies de Poecilurichthys y Zygogaster, sobre ejemplares adultos. Se registró la imagen del lado izquierdo de cada ejemplar con una cámara digital Cyber Shot w360 Sony instalada sobre un banco de fotografía; se utilizó la opción macro del lente para adquirir imágenes sin deformaciones en los márgenes; se empleó papel milimetrado como fondo en cada imagen para verificar la existencia de dicha deformación, y determinar la escala en cada imagen (Ruiz–C. & Cipriani, 2006). Se creó un archivo tps por medio del programa tpsUtil (Rohlf, 2004a). Se digitalizaron los hitos morfológicos sobre la imagen de cada individuo en el programa tpsDig2 (Rohlf, 2004b). Los hitos corresponden a estructuras homólogas que permitieron describir la forma general del cuerpo; se registraron 26 hitos tipo I y II (tabla 2). La talla de cada configuración fue estimada al usar la "talla del centroide" (TC) que es la raíz cuadrada de la suma de los cuadrados de las distancias entre cada hito y su centroide (Bookstein, 1991). Una vez que todas las imágenes fueron digitalizadas se eliminaron los efectos de traslación, escala y rotación

Animal Biodiversity and Conservation 34.1 (2011)

1 cm 1 cm A. atratoensis

1 cm

A. caucanus

1 cm A. filiferus

A. magdalenae

1 cm

1 cm A. orthodus

A. superbus

Fig. 1. Especies del género Astyanax: A. Subgénero Zygogaster; B. Subgénero Poecilurichthys. Fig. 1. Species of Astyanax genus: A. Zygogaster subgenus; B. Poecilurichthys subgenus.

del conjunto de configuraciones, mediante un análisis ortogonal de mínimos cuadrados generalizados de procrustes (AGP) (Rohlf & Slice, 1990), con en el programa tpsSmall (Rohlf, 2003); en éste procedimiento todas las configuraciones fueron escaladas a TC = 1. La matriz de coordenadas generada a partir del programa tpsDig2 fue analizada con el programa MorphoJ en el que se realizó: 1) un Análisis de Componentes Principales (ACP) para evaluar la variación morfológica entre individuos de las especies analizadas, sobre datos estandarizados; 2) un análisis de variables canónicas (AVC), realizado sobre los datos estandarizados, éste análisis multivariado maximiza la varianza entre grupos, en contraste con la varianza entre individuos (ACP), fue útil para discriminar la variación entre los subgéneros, entre las especies de los subgéneros, así como entre las poblaciones de cada especie; ésta exploración evalúa la significancia de las distancias de Mahalanobis entre los grupos; los cambios en la forma asociados a cada variable canónica (VC) son reflejados en rejillas de deformación que describen las diferencias entre grupos.

Material examinado Subgénero Zygogaster Astyanax atratoensis Colombia: IAvHP–7140, 21 (55,02–117,97 mm LE), río Atrato, S.F.; IAvHP–7187, 2 (78,42–82,32 mm LE), caño Muerto, última Cienaga, vereda el Cuarenta, Turbo, Antioquia, cuenca del río Atrato, S.F.; IAvHP–7206, 3 (71,25–91,87 mm LE), ciénaga Tumarado, vereda El cuarenta, Turbo, Antioquia, cuenca del río Atrato, S.F.; IUQ–84, 2 (85,73–88,20 mm LE), 2 D&T (83,60– 84,95 mm LE), río Atrato en Malecón de Quibdó, II 1993; IUQ– 696, 3 (90,21–110,72 mm LE), río Negro, vereda El Ganado, río Atrato, 10 II 1988; IUQ–745, 2 (68,93–76,09 mm LE), 1 D&T (64,85 mm LE), ciénaga de Achuara, en el medio Atrato, Chocó, 7 I 1988. Astyanax caucanus Colombia: ICN–9580, 3 (76,23–84,05 mm LE), río Magdalena, Purificación, Tolima, 28 XII 2005.

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m 12º N

10º N

5.000 4.000 3.000 2.000 1.000 0

8º N

6º N

4º N

A. atratoensis A. caucanus

2º N

A. filiferus A. magdalenae A. orthodus

0º N

80º W

78º W

76º W

200 74º W

400 km 72º W

70º W

A. superbus

68º W

Fig. 2. Distribución de las especies de los subgéneros Zygogaster y Poecilurichthys: A. atratoensis, A. caucanus, A. filiferus, A. magdalenae, A. orthodus y A. superbus. Fig. 2. Distribution of species of subgenera Zygogaster y Poecilurichthys: A. atratoensis, A. caucanus, A. filiferus, A. magdalenae, A. orthodus, and A. superbus.

IAvHP–8439, 23 (56.51–76,52 mm LE), 1 C&T (73,11 mm LE), Quebrada. La Arenosa, afluente del río La Miel, sistema del Magdalena medio, La Dorada, Caldas, S.F.; IAvHP–10568, 1 (79,32 mm LE), río Chicamocha, área de Pescadero, Pie de cuesta Santander, Magdalena, S.F.; ICN–10968, 3 (71,47– 94,15 mm LE), quebrada Bocorná, en confluencia con el río Guarinó, sistema del Magdalena, La Dorada, Caldas, 1 VIII 2003; ICN–11490, 2 (68,25–69,06 mm LE), río Guarinó, sistema del Magdalena, corregimiento La Victoria, La Dorada, Caldas, VII–VIII 2004; ICN–11522, 1 (75,91 mm LE), Quebrada Casanquilla, afluente del río Guarinó, sistema del Magdalena, en la vía La Dorada– Victoria, Caldas, VII–VIII 2004; IUQ–171, 2 (81,35–90,91 mm LE), Canal del Dique, Atlántico, 5 VIII 1991; IUQ–1025, 1 (52,03 mm LE), 1 C&T (71,45 mm LE), Suan, Atlántico, 28 IV 1990; IUQ–1296, 3 (61,16–61,98 mm LE), El Jaguey, Dique Viejo, Canal del Dique, Atlántico, 28 IV 1990; IUQ–2249, 1 (89,64 mm LE), río La Miel, cuenca del Magdalena, corregimiento San Miguel, Sonsón, Antioquia, 8 IX 2006; IUQ–2250, 1 (113,89 mm LE), río La Miel, corregimiento San Miguel, Sonson, Antio-

quia, 1 II 2006; IUQ–2807, 10 (59,68–97,58), 2 C&T (72,28–74,57 mm LE), quebrada Casanquilla, puente vía a La Victoria, Caldas, 13 II 2010; IUQ–2808, 10 (65,53–90 mm LE), 2 C&T (74,32–78,55 mm LE), Quebrada Casanquilla, afluente río Guarinó, La Victoria, Caldas, 13 II 2010; ICN–11728, 1 (68,58 mm LE), La Dorada, Caldas. S.F.; ICN–11538, 5 (60,24– 102,37 mm LE, La Dorada, Caldas, S.F.; ICN–15074, 3 (70,21–70,65 mm LE), La Dorada, Caldas, S.F. Astyanax filiferus Colombia: ICN–15618, 1 (71,38 mm LE), río Samaná, Caldas. S.F.; CIUA–692, 8 (59,43–89,21 mm LE), lago Los Deseos, La Dagua de Ibérico, Puerto Cesár, cuenca del río Magdalena, S.F.; IUQ–2230, 2 (71,49–83,96 mm LE), río La Miel en La Cachaza, Norcasia, Caldas; 8 XI 2006; IUQ–2232, 1 (79,13 mm LE), río La Miel en La Cachaza, Norcasia, Caldas, 30 I 2006; IAvHP–3969, 3 (74,32– 91,15 mm LE), 2 C&T (67,35–70,08 mm LE), Quebrada Palagua, Puerto Boyacá, Boyacá, cuenca del río Magdalena medio, 27 VI 1999; IUQ–988, 4 (18,98– 84,17 mm. LE), 2 D&T (55,5– 60,53 mm. LE), canal del Dique,

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Tabla 1. Especies y poblaciones de los subgéneros Poecilurichthys y Zygogaster examinadas en este estudio: N. Número de individuos. Table 1. Species and populations of Poecilurichthys and Zygogaster subgenera examined in this study: N. Number of individuals.. Especie

Localidad

Zygogaster A. atratoensis 25 Río Atrato, Quibdo, medio Atrato

3 Cienaga Achuara, medio Atrato

A. caucanus

Total 33

Última cienaga, bajo Atrato

3 Purificación, medio magdalena

31 Río Guarino, medio Magdalena

Río La Miel, medio Magdalena Canal El Dique, bajo Magdalena

A. filiferus

8 Puerto Cesár, bajo Magdalena

5 Puerto Boyacá, medio Magdalena

3 12

Total 67

Total 28

Río La Miel, Caldas Canal El Dique, bajo Magdalena

A. magdalenae 10 Bajo Magdalena

Caño Juan esteban, medio

2 La Dorada, medio Magdalena

Río Rancheria, caribe colombiano

1 Río Tigre, caribe colombiano

Río Cachirí, lago de Maracaibo

8 Río Guasare, lago de Maracaibo

Río Palmar, lago de Maracaibo

5 Río Catatumbo, lago de Maracaibo

Río Escalante, lago Maracaibo

Total 89

Poecilurichthys A. orthodus

2 Truandó, bajo Atrato

2 Bajo Atrato

Total 114

Río Yuto, alto Atrato 4 Rio San Juan

51 Río Telembi A. superbus

6 Río Guache, medio Orinoco

Río Apure, medio Orinoco

5 Río Tua, alto Orinoco

Río Tacuya, alto Orinoco

en Soplaviento, Bolívar, 1 VI 2003; IUQ 1019, 4 (78,87– 66,66 mm. LE), 2 D&T (39,66–78,41 mm LE), canal del Dique, en Soplaviento, Bolívar, 01 VI 2003. Astyanax magdalenae Colombia: CIUA–470, 7 (53,54– 66,72 mm LE), 2 D&T (47,64–51,09 mm LE), complejo lagunar del bajo Sinú, Lorica, Momil, Córdoba, IX 2006; IAvHP–8185, 6 (48,50–59,32 mm LE), Caño Juan Esteban, sistema del Magdalena, Santander, 2 II 1998; IAvHP–8186, 8 (40,08–52,62 mm LE), Caño Juan Esteban, sistema del Magdalena, Santander, 25 IX 2008; IUQ–1413, 1 (49,22 mm LE), Arroyo frente a Santa Lucia, At-

Total 23

Total 354

lántico 12 XI 1999; MPUJ–4221, 2 (50,97–52,69 mm LE) río Magdalena, La Dorada, Caldas, S.F.; ICN–16626, 4 (54,92–66,45), 1 C&T (58.19 mm LE), S.F.; ICN–9678, 2 (75,98–77,59 mm LE), Ranchería, 1 V 1992; IMCN–3040, 1 (102,32 mm LE), río Tigre, Córdoba, 1 IX 2004; IUQ–1004, 5 (60,57–86,02 mm LE), 2 D&T (73.17–75.38 mm LE), río Ranchería, Guajira, 6 I 1988; ICN–9619, 7 (71,15–100,20 mm LE), río Ranchería, Guajira, 1 III 006; IUQ–938, 1 (101,24 mm LE), río Ranchería, 16 VI 1981; IUQ–715, 2 (108,11–110,20 mm LE), río Ranchería, Guajira, 16 VI 1981; IUQ–1390, 1 (69,19 mm LE), río Ranchería, III 1988.

Venezuela: MBUCV–24776, 3 (73,24–82,87 mm LE), río Cachirí, lago de Maracaibo, Zulia, 14 XII 1982; MBUCV–18265, 4 (51,86–77,49 mm LE), río Cachirí, lago de Maracaibo, Zulia, 4 XII 1982; MBUCV–17027, 5 (60,28–102,76 mm LE), Guasare, lago de Maracaibo, Zulia, 13 XII 1982; MBUCV–23828, 8 (85,20–103,55 mm LE), río Palmar, Sierra de Perijá, lago de Maracaibo, Zulia, 7 XII 1982; MBUCV–23790, 5 (51,63–81,37 mm LE), río Catatumbo, lago de Maracaibo, Zulia, 11 XII 1982; MBUCV–23827, 6 (60,61–86,22 mm LE), río Cachirí, lago de Maracaibo, Zulia, 4 XII 1982; UF–25431, 3 (51,75–94,40 mm LE), río Guasare, hacienda Pamplona, municipio de Guajira, Estado de Zulia, 4 IV 1977; INHS–55460, 6 (63,51–82,47 mm LE), 2 D&T (83,89–87,45 mm LE), caño El Padre, río Onia, río Escalante, lago Maracaibo. Sobre la carretera de Hwy. 2 a town del km 35, Zulia, S.F. Subgénero Poecilurichthys Astyanax orthodus Colombia: AMNH–5370, 2 (66,28–72,04 mm LE), Truando, 1913; IAvHP– 6494, 1 (85,32 mm LE), río Yuto, Yuto–Chocó, cuenca del Atrato, 26 VI 2008, IAvHP– 7146, 19 (39,95–97,73), Atrato, río Sucio, S.F.; IAvHP–7208, 11 (38,68–51,37 mm LE), quebrada Tendal al lado de la motobomba PNN Katios, vereda Sautata, Riosucio, Chocó, cuenca del Atrato, S.F.; IAvHP–7209, 11 (32,50–77,07 mm LE), quebrada tendal, vereda Sautata PNN Katios, Riosucio, Chocó, cuenca del Atrato, S.F.; IAvHP–7210, 1 (82,46 mm LE), quebrada Tendal al lado de la motobomba PNN Katios, vereda Sautata, Riosucio, Chocó, cuenca del Atrato, S.F.; ICN–207, 6 (48,88–86,95 mm LE), San Juan, 1 XI 2002; INCIVA (IMCN) 1594, 2 (87,08–90,69 mm LE), quebrada Nalde–río San Juan, Itsmina, 17 I 003; INCIVA (IMCN) 1822, 4 (67,92–82,24 mm LE), Chocó, 1 I 002; INCIVA (IMCN) 1834, 6 (51,58–85,36 mm LE), San Juan, Chocó, 7 VIII 2002; IUQ 2252, 47 (38,07–76,65 mm LE), 4 D&T (39,88–69,11 mm LE), Quebrada finca La Hacienda, afluente río Telembi, 16 VII 2008. Astyanax superbus Venezuela: INHS–28666, 1 (43,62 mm LE), 2 D&T (37,21– 44,55 mm LE), río Guache, afluente del río Portuguesa, sistema del Orinoco, Portuguesa, S.F.; MCNG–6350, 4 (64,09– 89,86 mm LE), 2 D&T (56,98– 65,70 mm LE), caño Musao, afluente del río Apure, Barinas, 30 V 2005; MCNG–36349, 3 (68,72–70,97 mm LS), río Guache, en Garabote, afluente del río Portuguesa, sistema del Orinoco, Portuguesa, 26 III 1993. Colombia: IavHP–7910, 1 (62,46 mm LE), río Túa en el puente de la vía Villao, Yopal, Monterrey–Casanare, S.F.; IavHP–7911, 2 (71,99–73,05 mm LE), quebrada afluente del río Túa, Monterrey, Casanare, S.F.; IAvHP 3539, 2 D&T (73,21–76,80 mm LE), río Cravo Sur, río Túa y Unete, Orinoco–Meta, S.F.; IAvHP 3547, 4 (66,16–75,34 mm LE), 2 D&T (68,51–74,53 mm LE), río Tacuya, Orinoco–Meta, S.F.

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Resultados Aunque el ACP indicó que la morfología de las especies de Zygogaster y Poecilurichthys es generalizada y presenta solapamiento, existe cierto nivel de divergencia entre ambos (fig. 3A), el primer componente principal explica la mayor divergencia entre ambos (23%), indica cambio en la profundidad del cuerpo, desplazamiento anterior del origen de la aleta dorsal y de la mancha humeral, la cual fue representada como proyecciones verticales de ésta sobre el área predorsal, transformación con tendencia ventral de la superficie dorsal del cráneo, orbita y hocico, el cual, es probable, se encuentre relacionado con la menor altura del cráneo en las especies de Zygogaster. Se indica aislamiento del tercer infraorbital del preopérculo, carácter presente y variable de las especies de Zygogaster y protrusión del extremo ventral del maxilar, el cual es más corto en las especies de Zygogaster respecto a las especies de Poecilurichthys (fig. 3B). La variación morfológica de las especies del subgénero Zygogaster, representada en los componentes principales reveló ambigüedad relacionada con la semejanza existente entre especies, reconocidas como crípticas; aunque indica que la forma de A. caucanus es generalizada entre las especies del subgénero Zygogaster, A. filiferus y A. magdalenae, distribuidas en la región trans–interandina junto con A. caucanus, exhiben tendencias opuestas entre ambas; mientras A. atratoensis, en alopatría con relación a otras especies de Zygogaster, presenta forma centralizada entre especies de éste subgénero (fig. 4A). El ACP generó una mayor discriminación entre las especies de Poecilurichthys (fig. 4B), indicó contracciones y expansiones generales en diferentes regiones del cuerpo (cefálica, pectoral, dorsal, pélvica y caudal), comunes entre especies y subgéneros, que en algunos casos, presentaron direcciones contrarias, así que en consenso éstas deformaciones son contrastadas y no revelan una forma común para los subgéneros mencionados. El AVC reveló diferencias entre subgéneros (fig. 5A), sin embargo se confirma el resultado obtenido por el ACP, en el sentido de señalar formas generalizadas entre ambos, así la variación fue significativa (Distancia de Mahalanobis: 4,7; P < 0,0001), en una prueba de permutación con 1.000 replicaciones, de la cual se obtuvo un único autovalor (valor propio: 3,2), que explicó la totalidad de la variación de la forma entre ambos subgéneros (fig. 5B), describe: contracción en el hocico (hito 1), depresión en la superficie dorsal del cráneo (hitos 2 y 3), contracción entre el origen de la aleta dorsal (hitos 4, 5 y 6) y el área predorsal (hito 17), reducción en la longitud del borde posterior del tercer infraorbital (hitos 22, 23), protrusión anterior dorsal del extremo ventral del maxilar (hito 18), leve disminución en la profundidad del cráneo (hitos 25, 26), incremento en la profundidad ventral del cuerpo (hitos 13 y 14) y contracción anterior del pedúnculo caudal.

Animal Biodiversity and Conservation 34.1 (2011)

Tabla 2. Hitos morfológicos incluidos en el análisis: H. Hito; T. Tipo de hito. Table 2. Morphological landmarks included in the analysis: H. Landmark; T Landmark type. H T Rasgo morfológico

H T Rasgo morfológico

1 I Extremo anterior del hocico

14 I Origen de la aleta pelvica

2 II Proyección vertical del borde anterior de 15 I la orbita sobre la superficie dorsal del cráneo 3 II Proyección vertical del borde posterior de 16 II la orbita sobre la superficie dorsal del cráneo 4 II Proyección vertical del borde anterior de la 17 II mancha humeral sobre área predorsal sobre el área prepectoral 5 II Proyección vertical del borde posterior de la 18 I mancha humeral sobre área predorsal 6 I Origen de la aleta dorsal 19 II 7 I Origen de la aleta adiposa 20 II 8 II Proyección vertical del extremo posterior de 21 I la aleta anal sobre la superficie dorsal del pedunculo caudal 9 I Extremo posterior del penultimo radio 22 II procurrente en el lobulo dorsal de la aleta caudal 10 I Extremo posterior del penultimo radio 23 II procurrente en el lobulo ventral de la aleta caudal

Proyección horizontal del hito 21 sobre el borde posterior del tercer infraorbitalç Eje transversal entre los hitos 2 y 15 sobre el borde posterior del tercer infraorbital

11 I Extremo posterior de la aleta anal 24 II 12 II Proyección vertical del origen de la aleta 25 II adiposa sobre la base de la aleta anal 13 I Origen de la aleta anal 26 II

Eje transversal entre los hitos 1 y 15 sobre sobre el borde posterior del tercer infraorbital Proyección vertical del borde posterior de la orbita sobre la superficie ventral del cráneo Proyección vertical del borde anterior de la orbita sobre la superficie ventral del cráneo

Subgénero Zygogaster Diferencias morfológicas entre las especies de Zygogaster fueron establecidas como significativas a partir del AVC (tabla 3); descritas por tres autovalores (valores propios: 4,7; 1,1; 0,4), que aportan a la explicación de la variación de la forma entre especies (tabla 4). La variación de la forma en Zygogaster es afectada por la distribución de las especies que lo conforman, así: A. atratoensis en la cuenca del Pacifico, difiere en forma con otras especies del subgénero distribuidas en la región trans inter–andina (A. caucanus, A. filiferus y A. magdalenae) (fig. 6A). A partir del AVC se determinó que A. atratoensis es más parecida a

Origen de la aleta pectoral Proyección vertical del origen de la aleta pectoral sobre el área preventral Proyección horizontal anterior del origen de la aleta pectoral Extremo ventral posterior del maxilar Borde anterior de la orbita trazado desde una horizontal respecto al hito Borde posterior de la orbita trazado desde una horizontal respecto al hito 1 Punto de union dorsal de los infraorbitales 2 y 3

A. caucanus, mientras ésta última presenta forma generalizada entre las especies trans–interandinas de Zygogaster (fig. 6A, tabla 3). La variación morfológica entre especies de Zygogaster es descrita por tres VC (fig. 6B). La VC1 explica el 74.8% de la variación morfológica, indica una extensión anterior de la orbita (hitos 2, 3, 19, 20), reducción en la longitud del maxilar (hito 18), proyección anterior dorsal de la unión de los infraorbitales dos y tres (hito 21), extensión dorsal del borde posterior del tercer infraorbital, expansión anterior ventral de la región prepectoral (hitos 17, 25), proyección posterior del origen de la aleta pectoral y de su prolongación vertical sobre el área ventral (hitos 15 y 16), contracción en sentido dorsal de la región

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Zygogaster Poecilurichthys

Componente principal 2

0,04 0,02 0,00 –0,02 –0,04

–0,06 –0,06

–0,04 –0,02 0,00 0,02 0,04 Componente principal 1

0,06

B 4

7 2 1 19

21 18 26

22 23 24 17

12 15 16

Fig. 3. A. Análisis de componentes principales que explica la variación morfológica entre las especies de los subgéneros Poecilurichthys y Zygogaster; B. Deformación expresada por el primer componente, representa el 23% de la divergencia entre ambos subgéneros. Fig. 3. A. Principal component analysis explaining the morphological variation among species of the subgenera Poecilurichthys and Zygogaster; B. Deformation expressed by the first component, representing 23% of the difference between the two subgenera.

ventral en el origen de la aleta pélvica y anal (hitos 13, 14), contracción anterior de la región precaudal (hitos 7, 8, 11, 12), y encogimiento sobre el origen de la aleta dorsal (fig. 6B–a). La VC2 revela el 18% de variación morfológica, se observa una contracción anterior del borde posterior de la orbita (hitos 3, 20), contorsión anterior del borde posterior del tercer infraorbital, enseña un aislamiento de ésta estructura del preopérculo, encogimiento general en la profundidad del cuerpo (hitos 4, 5, 6, 13, 14, 16), expansión en la base del pedúnculo caudal (hitos 7, 8, 9, 10, 11, 12), contorsión en el área prepectoral (hitos 15, 17) (fig. 6B–b). La VC3 explica el 7% de la variación morfológica entre las especies de Zygogaster, señala deformaciones congruentes con las expresa-

das por las dos primeras VC: contracción anterior del tercer infraorbital, expansión del área prepectoral y encogimiento de la región precaudal. Sin embargo, expresa información independiente en relación a la contracción en la longitud del hocico y en la depresión de la superficie dorsal del cráneo (fig. 6B–c). Astyanax caucanus A. caucanus está ampliamente distribuida en la cuenca del río Magdalena. El AVC permitió identificar variación morfológica entre poblaciones de A. caucanus, significativa entre algunas poblaciones (tabla 5), aunque la diferencia entre poblaciones de los ríos Magdalena–Cauca no son significativas, indica

Animal Biodiversity and Conservation 34.1 (2011)

A Componente principal 2

0,06 0,04

A. A. A. A.

atratoensis caucanus filiferus magdalenae

0,02 0,00 –0,02 –0,04

–0.06 –0,08 –0,06 –0,04 –0,02 0,00 0,02 0,04 0,06 Componente principal 1 B Componente principal 2

0,06 0,04

A. orthodus A. superbus

0,02 0,00 –0,02

–0,04 –0,06 –0,04 –0,02 0,00 0,02 0,04 Componente principal 1

0,06

0,08

Fig. 4. A. Análisis de componentes principales que explica la variación morfológica entre las especies de: A. Subgénero Zygogaster; B. Subgénero Poecilurichthys. Fig. 4. Principal component analysis explaining the morphological variation among species of: A. Subgenus Zygogaster; B. Subgenus Poecilurichthys.

formas diferentes (fig. 7A). Fueron obtenidos cuatro autovalores (valores propios: 19,9; 12,1; 11,1; 2,2) y las contribuciones de cada uno en la explicación de la variación de la forma entre las poblaciones de A. caucanus (tabla 4). La variación morfológica entre éstas poblaciones es descrita por cuatro VC (fig. 7B). La VC1 explica el 43,8% de la variación morfológica entre poblaciones de A. caucanus, describe un aumento en la superficie del cráneo, se indica por la extensión anterior del hocico (hito 1), leve prolongación dorsal de la superficie del cráneo (hito 2, 3) y pronunciada proyección de la superficie ventral del cráneo en sentido ventral (hitos 25, 26). El borde

posterior del tercer infraorbital se proyecta en igual sentido, con aumento del contacto en el preopérculo, sin embargo, el borde ventral del tercer infraorbital no presenta desplazamientos, una contracción hacia la parte anterior de la región pectoral (hitos 15, 16, 17), al igual que en el cuerpo (hitos 4, 5, 6 13, 14), encogimiento entre el origen de las aletas pélvica y anal, y en la región precaudal (hitos 7, 8, 10, 11) (fig. 7B–a). La VC2 explica el 26,6% de la variación morfológica entre poblaciones, presenta una deformación generalizada en todos los hitos con fuerte tendencia de expansión, excepto una contracción evidente entre el origen de las aletas pélvica, anal y en el extremo distal del lóbulo

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Tabla 3. Distancias de Mahalanobis entre las especies de Zygogaster: Aa. A. atratoensis; Ac. A. caucanus; Af. A. filiferus; Am. A. magdalenae. Table 3. Mahalanobis distances between species of Zygogaster. (For abbreviations see above.)

4,7*

5,3*

2,3*

5,0*

2,2*

3,2*

Tabla 4. Variación entre subgéneros Zygogaster y Poecilurichthys, como entre poblaciones de las especies incluidas en el análisis: Vc. Variable canónica; Vp. Valores propios. V. Varianza (%); A. Acumulado (%). Table 4. Variation among subgenus Zygogaster and Poecilurichthys, and among populations of the species included in the analysis. (For abbreviations see above.)

Taxon 0

dorsal del pedúnculo caudal (hito 9), el cual se retrae con notable energía en relación al lóbulo ventral de ésta estructura (fig. 7B–b). La VC3 explica el 24,5% de la deformación entre las poblaciones de A. caucanus, describe una contracción generalizada del cráneo, con tendencia ventral del segundo y tercer infraorbital, sin embargo, a diferencia de otras variables canónicas, indica aislamiento del tercer infraorbital y el preopérculo (fig. 7B–c). La VC4 explica 5% de la variación entre poblaciones de A. caucanus. Revela acortamiento del hocico, contracción en la longitud de la orbita; otras deformaciones descritas son congruentes con lo expuesto por las tres primeras VC. Sin embargo, a diferencia de lo descrito en la VC1, el tercer infraorbital tiene contacto ventral con el preopérculo. En contraste con la VC2, el extremo distal del lóbulo ventral del pedúnculo caudal contrae hacia la parte anterior, mientras el extremo dorsal se extiende en sentido ventral. Otra novedad de ésta VC consiste en que es la única que describe una expansión entre el origen de la aleta pélvica y anal (fig. 7B–d). Astyanax filiferus El AVC permitió identificar variación morfológica entre poblaciones de A. filiferus, y es significativa (tabla 6), e indica formas diferentes entre poblaciones reconocidas como A. filiferus (fig. 8A). Fueron obtenidos dos autovalores (valores propios: 7,3; 5,8) y las contribuciones de cada uno en la explicación de la variación de la forma entre las poblaciones de A. filiferus (tabla 4). La variación morfológica entre éstas poblaciones es descrita por dos VC (fig. 8B). La VC1 explica el 55,5% de la variación morfológica entre poblaciones de A. filiferus, describe un incremento en la profundidad ventral del cráneo (hitos 25, 26), leve aumento con tendencia anterior dorsal del hocico, proyección anterior del extremo ventral posterior del maxilar, ligera tendencia dorso posterior de la orbita y proyectada sobre la superficie dorsal del cráneo, hitos relacionados con la unión del segundo y tercer infraorbital como del borde posterior del tercer

Zygogaster

4,7

74,8

1,1

92,9

0,4

100

A. caucanus

19,9

43,8

12,1

26,6

70,4

11,1

24,5

94,9

2,2

100

A. filiferus

7,3

55,5

5,8

44,5

100

A. magdalenae

37,8

65,9

20,9

86,8

4,8

8,4

95,2

2,7

4,7

100

Poecilurichthys

16,4

100

A. orthodus

7,8

69,3

2,1

19,3

88,7

1,21

1,2

100

A. superbus

100

Tabla 5. Distancias de Mahalanobis entre poblaciones de A. caucanus: Cs. Casanquilla; Cc. Cauca; Dq. El Dique; Ml. La Miel; Mg. Magdalena. Table 5. Mahalanobis distances between populations of A. caucanus. (For abbreviations see above.) Cs Cs

12*

15,1*

13,8

8,8*

12,8*

15,1*

10,1*

15,6

12*

10,6*

Animal Biodiversity and Conservation 34.1 (2011)

Frecuencia

Zygogaster Poecilurichthys

0 –6

–4

–2 0 Variable canónica

B 4 2

20 21 22 23 18 24 26 15 25 17 16

1 19

Fig. 5. A. Análisis de variables canónicas que explica la variación morfológica entre las especies de los subgéneros Poecilurichthys y Zygogaster; B. Deformación morfológica entre los subgéneros Poecilurichthys y Zygogaster, a partir de la variable canónica obtenida por datos morfogeométricos. Fig. 5. A. Canonical variates analysis explaining the morphological variation among species of the subgenera Poecilurichthys and Zygogaster; B. Morphological deformation between subgenera Poecilurichthys and Zygogaster, from the canonical variable data obtained by procrustes.

infraorbital son estáticos, sin embargo, señala leve tendencia anterior dorsal del extremo ventral del tercer infraorbital, lo que representa un aislamiento de ésta estructura del preopérculo (hito 24), una expansión de la región prepectoral (hitos 15, 16, 17) revela una expansión general del cuerpo y una contracción anterior sobre la región precaudal (fig. 8B–a). La VC2 explica el 44,5% de la variación morfológica entre poblaciones de A. filiferus, describe una depresión en la superficie dorsal del cráneo (hitos 2, 3) respecto a la expansión dorsal observada sobre el área predorsal, contracción del hocico, leve extensión del extremo posterior ventral del maxilar en igual sentido. A diferencia de lo observado en la VC1, el tercer infraorbital demuestra un mayor contacto con el preopérculo sobre el borde posterior (hitos 22, 23) en relación al extremo ventral (hito 24), y expansión anterior de la región prepectoral (hitos 15, 16, 17). Otras deformaciones son congruentes con lo descrito para la VC1 (fig. 8B–b).

Astyanax magdalenae La variación morfológica identificada a partir AVC entre las poblaciones de A. magdalenae, fue significativa (tabla 7), indica formas diferentes entre las reconocidas como A. magdalenae (fig. 9A). Fueron obtenidos cuatro autovalores (valores propios: 37,8; 12,0; 4,8; 2,7) y las contribuciones de cada uno en la explicación de la variación de la forma entre las poblaciones de A. magdalenae (tabla 4). La variación morfológica entre éstas poblaciones es descrita por cuatro VC (fig. 9B). La VC1 explica el 65,9% de la variación morfológica entre poblaciones de A. magdalenae, describe una depresión en la superficie dorsal del cráneo, contracción en la longitud del hocico (hito 1), y en la profundidad del cráneo (hitos 25, 26), proyección posterior del extremo ventral del maxilar, el tercer infraorbital describe una contracción en longitud horizontal, se observa un

Ruiz–C. et al.

A. A. A. A.

Variable canónica 2

–2

–4 –6

atratoensis caucanus filiferus magdalenae

–4

–2 0 2 Variable canónica 1

B 6 4 5

7 8 9 3 2 a 20 1 19 1110 12 21 22 23 18 26 24 25 15 17 16

2 1 19 20 22 23 24

25 17

1821 26

7 8

11 10 12

1 19 20 22 18 21 23 26 2415 25 17 16

4 5 3

9 10

15 16

Fig. 6. A. Análisis de variables canónicas que explica la variación morfológica entre las especies del subgénero Zygogaster; B. Deformación morfológica entre las especies de Zygogaster, expresadas por las variables canónicas obtenidas por datos morfogeométricos: a. Variable canónica 1; b. Variable canónica 2; c. Variable canónica 3. Fig. 6. A. Canonical variates analysis explaining the morphological variation among species of the subgenus Zygogaster; B. Morphological deformation among Zygogaster species, expressed by the canonical variables obtained by procrustes data: a. Canonical variable 1; b. Canonical variable 2; c. Canonical variable 3.

aislamiento del borde posterior del tercer infraorbital con el preopérculo, mientras el extremo ventral posterior se pliega al preopérculo. Proyección anterior de la región pectoral, leve expansión anterior de la superficie ventral del cuerpo (hitos 13, 14) y prolongación posterior de la región precaudal (hitos 7, 12) (fig. 9B–a). La VC2 explica el 20,9% de la variación entre poblaciones de A. magdalenae, describe, en general, las deformaciones expresadas en la VC1, sin embargo,

a diferencia de ésta plantea un aislamiento del tercer infraorbital del preopérculo y definida proyección dorsal anterior del extremo posterior del pedúnculo (hitos 9, 10) (fig. 9B–b). La VC3 representa el 8,4% de la variación morfológica entre poblaciones de A. magdalenae. Revela una expansión anterior del hocico, proyección de la superficie dorsal del cráneo en igual sentido, incremento en la longitud horizontal del tercer infraorbital

Animal Biodiversity and Conservation 34.1 (2011)

Río Guarinó Río La Miel Medio Magdalena Bajo Magdalena Canal El Dique

Variable canónica 2

6 3 0

–3 –6 –9 –12

–9

B 4

–6 –3 0 Variable canónica 1

7 8

2 1 1920 21 22 18 23 26 24 15 2517 16

11 12 14

9 10

2 3 11 10 1 20 19 22 a 21 23 12 26 18 24 25 17 15 16 13 14

4 5

6 7

3 2 20 22 1 1921 23 18 26 24 15 25 17 16

2 20 1 19 21 22 18 23 26 24 25 17 15 16

4 5

11 10 12 13

6 7 8 11 12 14

Fig. 7. A. Análisis de variables canónicas que explica la variación morfológica entre las poblaciones de Astyanax caucanus; B. Deformaciones morfológica entre las poblaciones de A. caucanus, representan variables canónicas obtenidas por datos morfogeométricos: a. Variable canónica 1; b. Variable canónica 2; c. Variable canónica 3; d. Variable canónica 4. Fig. 7. A. Canonical variates analysis explaining the morphological variation among populations of Astyanax caucanus; B. Morphological deformation among the populations of A. caucanus, represent canonical variables obtained from procrustes data: a. Canonical variable 1; b. Canonical variable 2; c. Canonical variable 3; d. Canonical variable 4.

(hito 21), prolongación antero ventral en el extremo del maxilar, y posterior de la región pectoral, contracción entre el área predorsal y el origen de la aleta dorsal, contracción dorso posterior del origen de las aletas pélvica y anal que describen concavidad en la región pélvica (fig. 9B–c).

La VC4 explica el 4,7% de la variación morfológica, es congruente con las deformaciones descritas en las VC anteriores, sin embargo, en ésta se plantea un contracción en la longitud del borde posterior del tercer infraorbital (hitos 22, 23) y una expansión en la superficie ventral del cráneo (fig. 9B–d).

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Tabla 6. Distancias de Mahalanobis entre poblaciones de A. filiferus: Dq. EL Dique; Ml. La Miel; Mg. Magdalena. Table 6. Mahalanobis distances between populations of A. filiferus. (For abbreviations see above).

Tabla 7. Distancias de Mahalanobis entre poblaciones de A. magdalenae: C. Cienaga; Mg. Magdalena; Mr. Maracaibo; Rn. Rancheria; Sn. Sinu. Table 7. Mahalanobis distances between populations of A. magdalenae. (For abbreviations see above.)

5,7*

7,7*

7,4*

Subgénero Poecilurichthys Diferencias morfológicas entre especies de Poecilurichthys fueron establecidas a partir del análisis de variables canónicas (AVC), y son significativas (fig. 10A); fue obtenido un autovalor (valor propio: 16,4) y explica en total la variación de forma entre especies (tabla 4); describe la diversidad de formas entre especies de Poecilurichthys, indica una leve expansión horizontal de la orbita; protrusión anterior del extremo ventral del maxilar, extensión en la longitud horizontal del tercer infraorbital (hito 21, 22) e incremento del contacto de ésta estructura con el preopérculo (hitos 22, 23, 24), contracción entre la superficie ventral del cráneo y la proyección anterior del origen de la aleta pectoral (hitos 17, 25) y expansión general en la profundidad del cuerpo (fig. 10B). Además, el AVC determinó la variación entre poblaciones de las especies de Poecilurichthys incluidas en éste análisis. Astyanax orthodus La variación morfológica entre poblaciones de A. orthodus fue significativa (tabla 9), indicó formas diferentes entre las reconocidas como A. orthodus (fig. 11A). Fueron obtenidos tres autovalores (valores propios: 7,8; 2,1; 1,2) y las contribuciones de cada uno en la explicación de la variación de la forma entre poblaciones de A. orthodus (tabla 4). La variación morfológica entre éstas poblaciones es descrita por tres VC (fig. 11B). La VC1 explica el 69,3% de la variación entre poblaciones de A. orthodus, demuestra la extensión anterior del hocico, expansión de la orbita, protrusión posterior ventral del extremo ventral del maxilar, incremento en el contacto del tercer infraorbital con el preopérculo, contracción entre la superficie ventral posterior del cráneo con la proyección horizontal anterior del origen de la aleta pectoral. Leve contracción entre el área predorsal y el origen de la aleta dorsal como entre el origen de la aleta pélvica y anal, y leve encogimiento de la región precaudal (fig. 11B–a). La VC2, representa el 19,3% de la variación morfológica entre poblaciones de A. orthodus,

Rn Sn

9,5*

11,3*

7,1*

15,3*

9,8*

11,3

14,8*

16,9

18,5*

constituye deformaciones generales descritas en la VC1, sin embargo, la protrusión del extremo ventral del maxilar, contracción posterior de la superficie ventral del maxilar sobre la región pectoral, la cual se proyecta en sentido ventral, mientras el área predorsal y el origen de la aleta dorsal se proyectan en sentido dorsal (fig. 11B–b). La VC3 explica el 11,2% de la disparidad morfológica entre poblaciones de A. orthodus, describe ondulaciones a lo largo del cuerpo, y con mayor magnitud deformaciones planteadas en la VC1: un fuerte protrusión del extremo ventral del maxilar, y contracción entre la superficie ventral posterior del cráneo con la proyección horizontal anterior del origen de la aleta pectoral. Representa contracción dorsal del la región pélvica del cuerpo entre el origen de la aleta pélvica y anal, un contraste entre la extensión dorsal del área predorsal y depresión del origen de la aleta dorsal (fig. 11B–c). A. superbus La variación morfológica entre poblaciones de A. superbus no fue significativa (tabla 10), aunque indicó formas diferentes (fig. 12A). Se obtuvo un autovalor (valor propio: 5) que explica la totalidad de la variación en forma entre poblaciones de A. superbus (tabla 4). La variación morfológica entre estas poblaciones es descrita por una VC (fig. 12B), reveló un perfil dorsal de cráneo convexo, proyección ventral del hocico y de la orbita (hitos 19 y 20). Protrusión posterior ventral del extremo del maxilar, superficie ventral del cráneo con proyección anterior ventral, leve reducción en la longitud horizontal del tercer infraorbital e incremento del contacto de ésta estructura con el preopérculo. Contracción entre la superficie ventral posterior del cráneo con la proyección horizontal anterior del origen de la aleta pectoral, expansión posterior de la región pectoral, proyección dorsal del área predorsal y del origen de la aleta dorsal.

Animal Biodiversity and Conservation 34.1 (2011)

Río La Miel Río Magdalena Canal El Dique

Variable canónica 2

4 2 0 –2 –4 –6 –8

–6

–4 –2 0 2 Variable canónica 1

4 3

6 7 8

2 11 19 20 22 126 a 12 10 21 23 18 25 2415 17 13 14 16

3 2 1 20 19 21 22 18 23 2624 25

7 8 12

15 16

9 10

Fig. 8. A. Análisis de variables canónicas que explica la variación morfológica entre las poblaciones de Astyanax filiferus; B. Deformaciones morfológica entre las poblaciones de A. filiferus, representan variables canónicas obtenidas por datos morfogeométricos: a. Variable canónica 1; b. Variable canónica 2. Fig. 8. A. Canonical variates analysis explaining the morphological variation among populations of Astyanax filiferus; B. Morphological deformation between the populations of A. filiferus represent canonical variables obtained by procrustes data: a. Canonical variable 1; b. Canonical variable 2.

Discusión Los resultados obtenidos en éste trabajo, confirman observaciones previas sobre la morfología de las especies del subgénero Zygogaster (Eigenmann, 1921); indican una forma generalizada entre las especies de éste subgénero, por lo que fue reconocido como un grupo críptico, de baja resolución taxonómica. Se observó similaridad morfológica entre los subgéneros Zygogaster y Poecilurichthys, con tendencia a incrementar la profundidad del cuerpo en la parte anterior de éste; lo cual representa una estrategia para evadir la depredación, dado que los depredadores buscaran presas pequeñas o de cuerpos más fusiformes (Chizinski et al., 2010); este factor explica porque éstos grupos se encuentran entre los más exitosos de la íctiofauna de la región

neotropical. Sin embargo, a pesar del similaridad morfológica, se identificó disparidad morfológica entre los subgéneros Zygogaster y Poecilurichthys, relacionado con estructuras del cráneo, indican diferencias en la posición de la orbita, extremo ventral del maxilar, longitud del maxilar, longitud del hocico, tercer infraorbital respecto al preopérculo y profundidad del cráneo, éstas diferencias también fueron observadas entre poblaciones de las especies que los representan en éste trabajo. La variación en estas estructuras evidencia especiación adaptativa como posible alternativa a la especiación por aislamiento, como es el caso de A. caucanus, A. magdalenae y A. filiferus, las cuales se encuentran en simpatría en diversas localidades del sistema del río Magdalena. En éste aspecto, modelos de especiación han cambiado del énfasis tradicional

Ruiz–C. et al.

Variable canónica 2

0 La Ciénaga Lago de Maracibo Río Magdalena Río Renchería Río Sinú

–5

–10 –15

–12

–9 –6 –3 0 3 Variable canónica 1

B 4 5

3 8 2 a 119 20 11 10 22 21 23 12 18 24 26 25 15 17 13 16 14

4 5

2 19 20 22 18 21 23 24 26 25 15 17 16

2 3 1 1920 22 1821 23 26 25 24 1715 16

7 8

11 12

3 1

11 12

4 5

7 8

23 119 2022 18 21 23 26 24 15 25 17 16

1110 12

Fig. 9. A. Análisis de variables canónicas que explica la variación morfológica entre las poblaciones de Astyanax magdalenae; B. Deformaciones morfológicas entre las poblaciones de A. magdalenae, representan variables canónicas obtenidas por datos morfogeométricos: a. Variable canónica 1; b. Variable canónica 2; c. Variable canónica 3; d. Variable canónica 4. Fig. 9. A. Canonical variates analysis explaining the morphological variation among populations of Astyanax magdalenaae; B. Morphological deformation among populations of A. magdalenae, represent canonical variables obtained from procrustes data: a. Canonical variable 1; b. Canonical variable 2; c. Canonical variable 3; d. Canonical variable 4.

en la distribución geográfica, a una perspectiva más amplia con hincapié en los mecanismos y diversificación evolutiva (Doebeli et al., 2005; Chizinski et al, 2010). La especiación bajo condiciones de contacto ecológico entre especies divergentes, podría ser generada por fuerzas de selección disruptiva que causan una discontinuidad en la variación intra poblacional

y genera dos o más fenotipos distintos. Fenotipos extremos se adaptan mejor a nichos alternos, dicha especialización ecológica ha sido documentada en diversos sistemas (Martin & Pfennig, 2009). Mientras la alta riqueza de especies y la uniformidad en estructura de hábitats complejos, han estado asociados con nichos fragmentados (Willis et al., 2005).

Animal Biodiversity and Conservation 34.1 (2011)

Variable canónica 2

A. orthodus A. superbus

8 6 4 2 0

3 0 Variable canónica 1

4 2

1211

1 19 2022 21 1823 26 24 25 1715

Fig. 10. A. Análisis de variables canónicas que explica la variación morfológica entre las especies del subgénero Poecilurichthys; B. Deformación morfológica entre las especies de Poecilurichthys, representa la única variable canónica obtenida por datos morfogeométricos. Fig. 10. A. Canonical variates analysis explaining the morphological variation among species of the subgenus Poecilurichthys; B. Morphological deformation among Poecilurichthys species is the only canonical variable obtained BY procrustes data.

Tabla 8 Distancias de Mahalanobis entre especies de Poecilurichtys: Ao. A. orthodus; As. A. superbus. Table 8. Mahalanobis distances between species of Poecilurichtys. (For abbreviations see above.)

8,4

As 0

Tabla 9. Distancias de Mahalanobis entre poblaciones de A. orthodus: At. Atrato; Rs. Río sucio. Sj. San Juan. Tl. Telembi. Table 9. Mahalanobis distances between populations of A. orthodus. (For abbreviations see above.)

3,7

6,5*

5,7

6,9

5,4*

6,8

Ruiz–C. et al.

A Variable canónica 2

2 0 –2 –4

Río Río Río Río

–6 –8 –4

–2 0 2 4 Variable canónica 1

Atrato Sucio San Juan Telembi

4 5

a 3 2 1 1920 2122 18 23 24 2625 17 15 16

7 8

4 5

2 119 20 2122 18 23 26 24 25 15 17 16

11 10

7 8

1 2 20 1921 22 182324 15 2625 17 16

12 14

11 12

9 10

Fig. 11. A. Análisis de variables canónicas que explica la variación morfológica entre las poblaciones de Astyanax orthodus; B. Deformación morfológica entre las poblaciones de A. orthodus, representa la única variable canónica obtenida por datos morfogeométricos. Fig. 11. A. Canonical variates analysis explaining the morphological variation among populations of Astyanax orthodus; B. Morphological deformation among A. orthodus populations is the only canonical variable obtained from procrustes data.

La semejanza morfológica de las especies del subgénero Zygogaster; podría estar relacionada al aislamiento geográfico reciente y a las relaciones de parentesco (González–Díaz et al., 2005); además, indican una probable relación entre especies y las áreas de distribución de éstas, aunque se encontraron diferencias en su forma, que son identificadas en este trabajo como evidencia de polifenismo; es decir, fenotipos alternativos que muestran un uso diferente de los recursos (Pfennig & McGee, 2010). Sin embargo, la forma de una de las especies del subgénero Zygogaster, se explica por aislamiento geográfico: A. atratoensis presenta una forma diferente a otras especies de Zygogaster distribuida a lo largo de

Tabla 10. Distancias de Mahalanobis entre poblaciones de A. superbus: Ap. Río Apure; Mt. Río Meta. Table 10. Mahalanobis distances between populations of A. superbus. (For abbreviations see above.) Ap Mt

Ap 0 4,1

Mt 0

Animal Biodiversity and Conservation 34.1 (2011)

2,0

Río Portuguesa Río Meta

Frecuencia

1,5

1,0

0,5

0,0 –6

–4

–2 0 Variable canónica 1

4 5 2 3 1 19 20 21 22 23 2618 24 25 15 1716

6 7 12

9 8 11

Fig. 12. A. Análisis de variables canónicas que explica la variación morfológica entre las poblaciones de A. superbus; B. Deformación morfológica entre las poblaciones de A. superbus, representa la única variable canónica obtenida por datos morfogeométricos. Fig. 12. A. Canonical variates analysis explaining the morphological variation among populations of A. superbus; B. Morphological deformation among A. superbus populations is the only canonical variable obtained by procrustes data.

la cuenca del río Atrato, y en la cual, a diferencia de lo observado en otras especies de Zygogaster, no se encontraron diferencias significativas entre sus poblaciones. Las especies de Poecilurichthys, en alopatría, presentan formas divergentes entre si, aunque indican caracteres estables que difieren de lo observado en Zygogaster, como es el evidente incremento en el contacto del tercer infraorbital con el preopérculo y el aumento en la longitud del maxilar. Un caso similar se observó en peces del género Hemibrycon (Román–Valencia et al., 2009). Las especies de éste grupo analizadas aquí divergen en la forma general de la superficie dorsal del cráneo, mientras A. orthodus muestra una expansión dorsal continua en ésta área, A. superbus presenta una superficie convexa. También se encontraron diferencias generales relacionadas con la orbita, mientras A. orthodus presenta

un incremento en su longitud, el análisis indica un desplazamiento ventral de éste junto con el hocico en A. superbus. También fueron encontradas diferencias significativas entre la forma descrita en las poblaciones de ambas especies. A. orthodus, con localidad tipo en el río Atrato, vertiente del Atlántico, posee forma diferente a las descritas para otras poblaciones aisladas como las de los ríos San Juan y Telembi en la vertiente del Pacifico. Igual sucede con la forma de A. superbus, para la cuenca del río Portuguesa, Orinoco medio en Venezuela, la cual difiere de la población del alto Orinoco. En especies del género Astyanax, la disparidad morfológica corresponde a diferencias en la altura del cuerpo y deformaciones en la región cefálica y caudal (Ruiz–C. & Cipriani, 2006); éstos resultados concuerdan con el tipo de variación morfológica que se esta-

blece entre especies de los subgéneros Poecilurichthys y Zygogaster. Estos también son reportados para otro género de Characidae como es Hyphessobrycon (García–Alzate et al., 2010); por lo que podría evaluarse si este tipo de variaciones morfológicas hacen parte de la estrategia de diversidad de Characidae, en respuesta a los diversos ambientes que ocupan. Para tal propósito, se recomienda la implementación inicial de la morfometría geométrica en el análisis de la variación morfológica a nivel de especie, y de ésta manera, llegar a inferencias válidas que expliquen la diversificación de éstos grupos. Agradecimientos Ana Ermakova recibió financiación, para una pasantía de investigación en IUQ, de la Universidad del Quindío y Associattion for the Exchange of Students of Technical Experience (IAEST). También se obtuvo apoyo de la Universidad del Quindío–vicerrectoria de investigaciones. Donald C. Taphorn leyó una versión final de este articulo y ofreció valiosas sugerencias. Referencias Almirón, A. E., Azpelicueta M., Casciotta J. R. & López Cazorla, A., 1997. Icthyogeographic boundary between the Brazilian and Austral subregions in South America, Argentina. Biogeographica, 73: 23–30. Bookstein, F. L., 1989. "Size and shape": A comment on semantics. Systematic Zoology, 38: 173–180. – 1991. Morphometric tools for landmark data: geometry and biology. Cambridge Univ. Press. Cambridge. Bookstein, F. L., Schafer, K., Prossinger, H., Seidler, H., Fieder, M., Stringer, C., Weber, G. W., Arsuaga, J. L., Slice, D., Rohlf, F. J., Recheis, W., Mariam, A. J. & Marcus, L. F., 1999. Comparing frontal cranial profiles in archaic and modern Homo by morphometric analysis. The Anatomical Record, 257: 217–224. Chizinski, C. J., Pope K. L., Wilde G. R. & Strauss R. E., 2010. Implications of stunting on morphology of freshwater fishes. Journal of Fish Biology, 76: 564–579. Doebeli, M., Dieckmann, U., Metz, J. A. J. & Tautz, D., 2005. What we have also learned: adaptative speciation is theoretically plausible. Evolution, 59(3): 691–695. Eigenmann, C. H., 1917. The American Characidae. Memoirs of the Museum of Comparative Zoology, 43: 1–102. – 1921. The American Characidae. Part 3. Memoirs of the Museum of Comparative Zoology, 43: 209–310. Eschmeyer, W., 2010. CAS–Ichthyology–Catalog of Fishes, California academy of sciences, San Francisco, CA, USA. Disponible http: www. calacademy.org/research/ichthyology/catalog/ fishcatsearch.html. García–Alzate, C., Román–Valencia, C. & González, M., 2010. Morfogeometria de los peces del género

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Hyphesssobrycon (Characiformes: Characidae), grupo heterohabdus, en Venezuela. Revista de Biología Tropical, 58: 801–811. González–Díaz, A., Díaz–Pardo, E., Soria–Barreto, M. & Rodiles–Hernández, R., 2005. Análisis morfométrico de los peces del grupo labialis, género Profundulus (Cyprinodontiformes: Profundulidae), en Chiapas, México. Revista Mexicana de Biodiversidad, 76: 55–61. Martin, R. A. & Pfennig, D. W., 2009. Disruptive selection in natural populations: The roles of ecological specialization and resource competition. The American Naturalist, 174: 268–281. Mirande, J. M., 2010. Phylogeny of the family Characidae (Teleostei:Characiformes):from characters to taxonomy. Neotropical Ichthyology, 8: 385–568. Pfennig, D. W. & McGee, M., 2010. Resource polyphenism increases species richness: a test of the hypothesis. Philosophical Transactions of Royal Society, 365: 577–591. Rohlf, F. J., 2003. TpsSmall, version 1.20. Department of Ecology and Evolution, State Univ. of New York at Stony Brook. – 2004a. tpsUtil, file utility program. version 1.26. Department of Ecology and Evolution, State Univ. of New York at Stony Brook. – 2004b. tpsDig, digitize landmarks and outlines, version 2.0. Department of Ecology and Evolution, State Univ. of New York at Stony Brook. Rohlf, F. J. & Slice, D. E., 1990. Extensions of the procruster method for the optimal superimposition of landmarks. Systematic Zoology, 39: 40–50. Román–Valencia, C., Arcila-Mesa D. & Hurtado, H., 2009. Variación morfológica de los peces Hemibrycon boquiae y Hemibrycon rafaelense (Characiformes: Characidae) en el Río Cauca, Colombia. Revista de Biología Tropical, 57(3): 541–556. Román–Valencia, C. & Ruiz–C., R. I., 2005. Diet and reproduction aspects of Astyanax aurocaudatus (Teleostei: Characidae) from the upper part of the río Cauca, Colombia. Dahlia (Revista de la Asociación Colombiana de Ictiólogos), 8: 9–17. Ruiz–C., R. I., 2010. Filogenia y biogeografía de Astyanax subgénero Poecilurichthys (Pisces: Characidae). Tesis doctoral, Universidad Central de Venezuela–Instituto de Zoología Tropical, Caracas. Ruiz–C., R. I. & Cipriani, R., 2006. Análisis morfogeométrico de Astyanax siapae. Dahlia (Revista Asociación Colombiana de Ictiólogos), 9: 63–75. Sabaj–Pérez, N. H. (Ed.), 2010. Standard symbolic codes institutions resource collections in Herpetology and Ichthyology: an on line reference, version 1.5 (4 Oct.2010). Electronically accessible at http.// www.asih.org/, American Society Ichthyologist and Herpetologist, Washington, D.C. Willis, S. C., Winemiller, K. O. & López–Fernández, H., 2005. Habitat structural complexity and morphological diversity of fish assemblages in a floodplain river. Oecologia, 142: 284–295. Zelditch, M. L., Swiderski, D. L., Sheets, H. D. & Fink, W. D., 2004. Geometric morphometrics for biologists: A primer. Elsevier Acad. Press, New York.

Animal Biodiversity and Conservation 34.1 (2011)

Actes del XVI Simposi Ibèric d'Estudis de Biologia Marina, Alacant, Espanya Actas del XVI Simposio Ibérico de Estudios de Biología Marina, Alicante, España Proceedings of the XVIth Iberian Symposium of Studies on Marine Biology, Alicante, Spain

Comitè organitzador / Comité organizador / Organizing Committee Just T. Bayle Sempere Univ. de Alicante, Alicante, Spain Pablo Sánchez Jerez Univ. de Alicante, Alicante, Spain José Luis Sánchez Lizaso Univ. de Alicante, Alicante, Spain José A. González Pérez Inst. Canario de Ciencias Marinas, Islas Canarias, Spain Alfonso A. Ramos Esplá Univ. de Alicante, Alicante, Spain Francisca Giménez Casalduero Univ. de Alicante, Alicante, Spain Manuel Biscoito Estaçao de Biologia Marihna do Funchal, Islas Madeira, Portugal Yolanda Fernández Torquemada Univ. de Alicante, Alicante, Spain Carlos Valle Pérez Univ. de Alicante, Alicante, Spain Angel Borja AZTI–Tecnalia, Bizkaia, Spain Mafalda Freitas Estaçao de Biologia Marihna do Funchal, Islas Madeira, Portugal José Miguel González Correa Univ. de Alicante, Alicante, Spain Juan Eduardo Guillén Nieto Inst. d’Ecologia Litoral, Alicante, Spain Sofia Oliveira Pires Univ. de Alicante, Alicante, Spain Aitor Forcada Almarcha Univ. de Alicante, Alicante, Spain

Comitè científic / Comité científico / Scientífic Committee Alberto Brito Univ. de La Laguna, Islas Canarias, Spain Alberto Martín Zazo Univ. Simon Bolivar, Venezuela Alexandra Cunha Univ. do Algarve, Faro, Portugal Angel Borja AZTI–Tecnalia, Bizkaia, Spain Anna Sabatés CMIMA–CSIC, Barcelona, Spain Celia Olabarria Univ. de Vigo, Pontevedra, Spain Elsa Vázquez Otero Univ. de Vigo, Pontevedra, Spain Eva García Univ. de Oviedo, Oviedo, Spain Fernando Tuya Centro Interdisciplinar de Investigação Marinha e Ambiental, Porto, Portugal Joao Jose Castro Univ. de Évora, Évora, Portugal Jorge Baro Inst. Español de Oceanografía, Málaga, Spain José A. García Charton Univ. de Murcia, Murcia Spain José A. González Pérez Inst. Canario de Ciencias Marinas, Islas Canarias, Spain Juan M. Ruiz Inst. Español de Oceanografía, Pontevedra, Spain Karim Erzini Univ. de Algarve, Faro, Portugal Manuel Maldonado Centro de Estudios Avanzados de Blanes–CSIC, Girona, Spain Manuel Biscoito Estaçao de Biologia Marinha do Funchal, Islas Madeira, Portugal Miguel Neves dos Santos Inst. de Investigação das Pescas e do Mar, Olhao, Portugal Mikel Becerro Centro de Estudios Avanzados de Blanes–CSIC, Girona, Spain Montserrat Demestre CMIMA–CSIC, Barcelona, Spain Ricardo Haroun Univ. de Las Palmas de Gran Canaria, Islas Canarias, Spain Teresa Cruz Univ. de Évora, Évora, Portugal Xavier Turón Centre d'Estudis Avançats de Blanes–CSIC, Girona, Spain Assessors dels articles / Asesores de los artículos / Referees of papers Ángel Fernández Inst. Español de Oceanografía, Murcia, Spain Carmen Barberá Cebrián Inst. Español de Oceanografía, Palma de Mallorca, Spain Damián Jover Univ. de Alicante, Alicante, Spain Elsa Vázquez Univ. de Vigo, Pontevedra, Spain Enrique Ballesteros Centre d'Estudis Avançats de Blanes–CSIC, Girona, Spain Fernando Tuya Univ. de las Palmas de Gran Canaria, Islas Canarias, Spain Francisca Giménez Casalduero Univ. de Alicante, Alicante, Spain Javier Romero Univ. de Barcelona, Barcelona, Spain Jesús Salinas Calvete Univ. de Alicante, Alicante, Spain Jesús Troncoso Univ. de Vigo, Pontevedra, Spain José Antonio García Charton Univ. de Murcia, Murcia, Spain José Enrique García Raso Univ. de Málaga, Málaga, Spain José Jacobo Zubcoff Univ. de Alicante, Alicante, Spain José Luis Sánchez Lizaso Univ. de Alicante, Alicante, Spain José Templado Museo Nacional de Ciencias Naturales-CSIC, Madrid, Spain Juan Junoy Univ. de Alcalá de Henares, Madrid, Spain Juan Manuel Ruiz Fernández Inst. Español de Oceanografía, Murcia, Spain Just Bayle Sempere Univ. de Alicante, Alicante, Spain Luis Vicente López Llorca Univ. de Alicante, Alicante, Spain Mercedes González Wangüemert Univ. do Algarve, Faro, Portugal Montse Demestre CMIMA–CSIC, Barcelona, Spain Oscar Ocaña Museo del Mar de Ceuta, Ceuta, Spain Pablo Arechavala Univ. de Alicante, Alicante, Spain Pablo Sánchez Jerez Univ. de Alicante, Alicante, Spain Rodrigo Riera Centro de Investigaciones Medioambientales del Atlantico, Islas Canarias, Spain Romana Capaccioni Univ. de Valencia, Valencia, Spain Yolanda Fernández Torquemada Univ. de Alicante, Alicante, Spain

Animal Biodiversity and Conservation 34.1 (2011)

Índex / Índice / Contents 71

J. L. Sánchez Lizaso, J. Bayle Sempere & P. Sánchez Jerez Marine biology, biodiversity and conservation: foreword to the SIEBM 2010 Conference

141–150 E. Cacabelos, J. Moreira, A. Lourido & J. S. Troncoso Ecological features of Terebellida fauna (Annelida, Polychaeta) from Ensenada de San Simón (NW Spain)

73–82 J. M. Ruiz, L. Marín–Guirao, J. Bernardeau–Esteller, A. Ramos–Segura,R. García–Muñoz & J. M. Sandoval–Gil Spread of the invasive alga Caulerpa racemosa var. cylindracea (Caulerpales, Chlorophyta) along the Mediterranean Coast of the Murcia region (SE Spain)

151–163 E. Rojo–Nieto, P. D. Álvarez–Díaz, E. Morote, M. Burgos–Martín, T. Montoto–Martínez, J. Sáez– Jiménez & F. Toledano Strandings of cetaceans and sea turtles in the Alboran Sea and Strait of Gibraltar: a long–term glimpse at the north coast (Spain) and the south coast (Morocco)

165–177 M. Samy, J. L. Sánchez Lizaso & A. Forcada Status of marine protected areas in Egypt

101–111 L. M. Ferrero–Vicente, Á. Loya–Fernández, C. Marco– Méndez, E. Martínez–García & J. L. Sánchez–Lizaso Soft–bottom sipunculans from San Pedro del Pinatar (Western Mediterranean): influence of anthropogenic impacts and sediment characteristics on their distribution 113–122 M. García–Rodríguez, A. Fernández &. A. Esteban Biomass response to environmental factors in two congeneric Mullus, M. barbatus and M. surmuletus, off Catalano–Mediterranean coast of Spain: a preliminary approach 123–132 M. G. Pennino, J. M. Bellido, D. Conesa & A. López– Quílez Trophic indicators to measure the impact of fishing on an exploited ecosystem 133–140 A. Garcia, S. Cecchetti, M. N. Santos, S. Mattiucci, G. Nascetti & R. Cimmaruta Population structure of Atlantic swordfish (Xiphias gladius L. 1758) (Teleostea, Xiphiidae) using mitochondrial DNA analysis: implications for fisheries management

179–190 V. Fernandez–Gonzalez & P. Sanchez–Jerez Effects of sea bass and sea bream farming (Western Mediterranean Sea) on peracarid crustacean assemblages 191–203 C. Ojeda–Martínez, J. T. Bayle–Sempere, P. Sánchez– Jerez, F. Salas, B. Stobart, R. Goñi, J. M. Falcón, M. Graziano, I. Guala, R. Higgins, F. Vandeperre, L. Le Direach, P. Martín–Sosa & S. Vaselli Review of the effects of protection in marine protected areas: current knowledge and gaps 205–215 E. Azevedo, M. F. Caeiro, R. Rebelo & M. Barata Biodiversity and characterization of marine mycota from Portuguese waters 217–227 F. Giménez–Casalduero, C. Muniain, M. González– Wangüemert & A. Garrote–Moreno Elysia timida (Risso, 1818) three decades of research

Animal Biodiversity and Conservation 34.1 (2011)

Marine biology, biodiversity and conservation: foreword to the SIEBM 2010 Conference J. L. Sánchez Lizaso, J. Bayle Sempere & P. Sánchez Jerez

Sánchez Lizaso, J. L., Bayle Sempere, J. & Sánchez Jerez, P., 2011. Marine biology, biodiversity and conservation: foreword to the SIEBM 2010 Conference. Animal Biodiversity and Conservation, 34.1: 71. The SIEBM (Iberian Symposium of Studies on Marine Biology) has a long tradition. The conference was first convened in 1979 and it has since been held approximately biannually. It was originally focused on marine benthos, but two editions ago it was decided to widen its objectives to include other related subjects. The 15th edition was held at the University of Alicante from 6th to 10th September 2010. One of the strong points of SIEBM is that it provides the opportunity for interaction between well–known researchers and young marine scientists, mainly from Spain and Portugal. In the 2010 edition, however, it was attended by 268 participants from 15 countries. SIEBM 2010 focused on the human impact on marine ecosystems and the conservation of marine biodiversity to coincide with the United Nations declaring 2010 the International Year of Biodiversity. From a total number of 330 communications received, the Scientific Committee selected 70 for oral presentations and the rest were presented as posters. Each day a plenary conference covered what we considered as the fastest growing areas in marine biology: landscape ecology. The impact of global warming on marine ecosystems, approaches to fisheries management and the study and conservation of deep ecosystems. The oral presentations were divided into 15 sessions that included both basic research on biodiversity, phylogeny or marine ecology, and applied research in management, fisheries, conservation and the human impact on marine ecosystems. Abstracts, posters and presentations are available at http://www.siebm.org. The studies included in this volume show the variety of subjects and approaches addressed during the conference. These papers were selected and edited by a group of referees and we wish to thank them for their efforts and conscientious work. A list of their names is included. Of the 28 studies received, 13 were finally accepted for this volume. We also would like to thank all the people who helped in one way or another to make the 15th SIEBM Conference and the proceedings such a success. We thank the Scientific Committee for revising the abstracts and selecting the oral presentations. Most members of this committee also chaired sessions and undertook this task with commitment and dedication. The planning, logistics and running of the conference were made possible thanks to the enthusiasm of the organising committee and a group of student helpers. We are profoundly indebted to Carlos Valle and Aitor Forcada for their invaluable contribution to the organisation. Finally, we thank all the sponsors who contributed to the meeting, and especially the Fundación Biodiversidad and the Natural History Museum of Barcelona for their support in publishing this special issue of Animal Biodiversity and Conservation.

José Luis Sánchez Lizaso1, Just Bayle Sempere2 & Pablo Sánchez Jerez3, Dept. de Cièncias del Mar i Biologia Aplicada, Univ. d'Alacant, Ap 99, 03080 Alacant, Espanya (Spain). 1

JL.Sanchez@ua.es

ISSN: 1578–665X

Bayle@ua.es

Editor executiu / Editor ejecutivo / Executive Editor Joan Carles Senar

Secretaria de Redacció / Secretaría de Redacción / Editorial Office

Secretària de Redacció / Secretaria de Redacción / Managing Editor Montserrat Ferrer

Museu de Zoologia Passeig Picasso s/n 08003 Barcelona, Spain Tel. +34–93–3196912 Fax +34–93–3104999 E–mail mzbpubli@intercom.es

Consell Assessor / Consejo asesor / Advisory Board Oleguer Escolà Eulàlia Garcia Anna Omedes Josep Piqué Francesc Uribe

Animal Biodiversity and Conservation 34.1 (2011)

Spread of the invasive alga Caulerpa racemosa var. cylindracea (Caulerpales, Chlorophyta) along the Mediterranean coast of the Murcia region (SE Spain) J. M. Ruiz, L. Marín–Guirao, J. Bernardeau–Esteller, A. Ramos–Segura, R. García–Muñoz & J. M. Sandoval–Gil Ruiz, J. M., Marín–Guirao, L., Bernardeau–Esteller, J., Ramos–Segura, A., García–Muñoz, R. & Sandoval–Gil, J. M., 2011. Spread of the invasive alga Caulerpa racemosa var. cylindracea (Caulerpales, Chlorophyta) along the Mediterranean Coast of the Murcia region (SE Spain). Animal Biodiversity and Conservation, 34.1: 73–82. Abstract Spread of the invasive alga Caulerpa racemosa var. cylindracea (Caulerpales, Chlorophyta) along the Mediterranean Coast of the Murcia region (SE Spain).— The aim of this paper was to document the appearance and spread of the green alga Caulerpa racemosa along the coast of Murcia in south–eastern Spain. It was found for the first time in the area in 2005 and over the next two years the number of new sightings increased almost exponentially. In the period 2005–2007 the total surface area colonised by the alga in the region was estimated to be at least 265 ha. Benthic assemblages colonised by the alga were rocky bottoms with photophilic algae, dead P. oceanica rhizomes, infralittoral and circalittoral soft bottoms and maerl beds. No penetration of the alga was observed in P. oceanica meadows, except in one locality. Biometric analysis indicated high vegetative development in the established colonies in comparison to those described in other Mediterranean areas. Rapid spreading dynamics observed in the Murcia region is a potential threat for native benthic communities. Key words: Biological invasions, Caulerpa racemosa var. cylindracea, Colonised surface area, Distribution, Mediterranean Sea, Spain. Resumen Introducción y expansión del alga invasora Caulerpa racemosa var. cylindracea en el litoral de la región de Murcia (SE España).— En el presente trabajo se documenta la aparición y dispersión del alga verde Caulerpa racemosa a lo largo de la costa de Murcia, región situada en el sureste español. El alga fue detectada por primera vez en el año 2005 y durante los dos años consecutivos se observó un crecimiento casi exponencial en el número de áreas colonizadas. La superficie total colonizada por el alga en Murcia durante el periodo 2005–2007 ha sido estimada en 265 ha., siendo las comunidades bentónicas afectadas algas fotófilas sobre sustrato rocoso, "mata muerta" de P. oceanica, fondos blandos infralitorales y circalitorales y fondos con comunidades de maërl. La presencia del alga dentro de praderas de P. oceanica solamente fue detectada en una localidad. Los estudios biométricos realizados muestran un elevado desarrollo vegetativo de las poblaciones de C. racemosa en Murcia en comparación con colonias de otras áreas del Mediterráneo, siendo esta rápida dinámica de expansión una amenaza potencial para las comunidades bentonicas nativas. Palabras clave: Invasiones biológicas, Caulerpa racemosa var. cylindracea, Superficie colonizada, Mar Mediterraneo, España. Juan M. Ruiz, Lázaro Marín–Guirao, Aranzazu Ramos–Segura & Rocío García Muñoz, Inst. Español de Oceanografía, Centro Oceanográfico de Murcia, Seagrass Ecology Group, c./ Varadero 1, 30740 San Pedro del Pinatar, Murcia, España (Spain).– Jaime Bernardeau–Esteller & Jose Miguel Sandoval–Gil, Depto. de Ciencias del Mar y Biología Aplicada, Univ. de Alicante, Alicante, España (Spain). Corresponding author: Juan M. Ruiz. E–mail: juanm.ruiz@mu.ieo.es

ISSN: 1578–665X

Ruiz et al.

Introduction

Material and methods

A total of 745 alien species have been reported in the Mediterranean Sea. Of these, 385 (52%) are already well established (Zenetos et al., 2005). Within this group, the green alga Caulerpa racemosa (Forsskål) J. Agardh var. cylindracea (Sonder) Verlaque, Huisman et Boudouresque (hereinafter C. racemosa) has been described as one of the most notorious invaders and has been included in the '100 worst invaders' list for the Mediterranean Sea (Streftaris & Zenetos, 2006). The biological characteristics of C. racemosa (high rates of vegetative dispersal, production of allelopathic substances, etc.) determine its high colonisation potential and its extraordinary ability to outcompete and alter native benthic assemblages, which make this species a particular potential threat for the Mediterranean coastal ecosystem (Piazzi et al., 2005). C. racemosa was observed in the Eastern Mediterranean Sea for the first time along the coast of Libya in 1990 (Nizamuddin, 1991), the origin of this invasive variety as yet unknown (Verlaque et al., 2003; Durand et al., 2002; Panayotidis, 2006). Since then, the species has spread rapidly, gradually invading the Mediterranean Sea. This has been well documented in the western basin along the coasts of Italy, France and North Africa (Piazzi et al., 2005). Along the Mediterranean coast of Spain, the species was first sighted in the Balearic Islands in 1998 (Ballesteros et al., 1999). It reached the east coast of the Iberian peninsula (Castellón) in 1999 (Aranda et al., 1999) and began to spread quickly southward, being sighted in Alicante (SE Spain) in 2000 (Aranda et al., 2003). At that point, the algal spread seemed to stabilise (fig. 1A), but its presence was confirmed in the Murcia region in 2005, indicating that the colonising process was spreading southward. Precise studies documenting the presence of the alga in newly colonised areas (i.e. colony size, depth range, substrate type, morphometric data and invaded native communities) are crucial to elucidate its colonising potential, spreading dynamics and mechanisms (vectors) at local and large spatial scales (Klein & Verlaque, 2008). Cartographic methods make it possible to measure the extent of the spread and can assist in helping to predict potential impacts and future scenarios (Meinesz, 2007). Detailed information on the spreading dynamics and extent of C. racemosa is available for a limited number of Mediterranean regions (Piazzi et al., 1997; Ruitton et al., 2005a). The goals of the present study were: (1) to document the spreading dynamics of C. racemosa along the coast of Murcia (SE Spain) from its appearance in 2005 to 2007, both at regional and local scales, and (2) to provide some quantitative estimate of the colonised surface area. Furthermore, the work includes several characteristics of the invaded sites (colonised assemblages, colonization depth) together with the vegetative development of several colonies in this geographical area.

Study area and field sampling programme This study was carried out on the Mediterranean coast of Murcia, SE Spain (fig. 1A). After C. racemosa was first sighted in the region in 2005, an active detection programme was established to map the distribution of the alga and its spreading dynamics over time (Meinesz, 2007; Ruitton et al., 2005a). To this end, we initially selected 42 sampling stations unevenly distributed along 224 km of the Murcia coastline through a depth range from 2 to 30 m (fig. 1B, 1C). These stations were selected from different long–term sampling programmes that had already been initiated in the region for different purposes (scientific monitoring of P. oceanica meadows, environmental impact assessments and other scientific projects), but since they were visited at least once a year by specialised divers, this ensured reliable information about the date of appearance of the alga. Of course, this sampling strategy resulted in a non–systematic sampling design, but it provided insight into the colonisation process in a representative area of the Murcia coast. The period covered by this sampling programme was 2005–2007 i.e. the first three years of the colonisation process of C. racemosa in the Murcia region. Once the alga was detected at a given station, divers surveyed a total surface area of 0.5 ha to characterise the colonised area (depth range, types of colonised substrate and benthic assemblages) and to estimate its surface area (i.e. colonisation levels sensu Ruitton et al., 2005a). Based on the field data obtained, invaded localities were assigned to one of the following five categories of colonisation level: (I) one or a few small colonies covering a surface area of less than 10 m2; (II) colonies of varying sizes covering a total surface area between 10 and 104 m2; and (III) meadows covering surface areas between 104 and 105 m2, (IV) 105 and 106 m2 and (V) greater than 106 m2. For categories I and II, the surface area was estimated in a single survey within the sampling station using quadrats and transects. For cases belonging to categories III–V, where the colonised area extended beyond the area surveyed by divers at a single station, additional dives were necessary to determine the limits of the total colonised area. These additional dives were performed at neighbouring points separated from the sampling station by several hundreds of metres and at different depths and directions (a specific sampling design was established in each case). Once the limits of the invaded area were identified, these were determined by GPS and input into a Geographic Information System (ArcView microcomputer programme Version 9.0, Esri ©) to estimate the surface area of the polygon thus generated. Biometric analysis Biometric analysis of C. racemosa colonies was performed using data from summer 2007 (June 28th to August 9th), a season in which the vegetative development of the alga was close to its annual

Animal Biodiversity and Conservation 34.1 (2011)

Level of colonisation

Marsella 1997

Castellón 1999 Alicante 2000

Palma de Mallorca 1998

an Sea diterrane

Murcia 2005

1–10 m2

10–10.000 m2

III

10.000–100.000 m2

100.000 –1.000.000 m2

Argel 206

> 1.000.000 m2

B 2

1 4

Cabo de Palos Cartagena 31 32 33 34 35 36

29 30 28 26,27

23 22

15,16,17,18 20

39,40,41,42

20 km

C N 9

Marine Reserve Cabo de Palos–Islas Hormigas 7 6

5 12 13

10 11

2 km

Fig. 1. A. Recent spread of C. racemosa in the Western Mediterranean basin; B. Distribution of sampling stations on the Murcia coast; C. Distribution of sampling stations in the Marine Reserve Cabo de Palos– Islas Hormigas. Invaded stations are indicated by black circles, the size of which corresponds to one of the five categories of colonisation level. Information related to the 42 sampling stations is included in appendix 1. Fig. 1. A. Expansión reciente de C. racemosa en la cuenca mediterránea occidental; B. Distribución de las estaciones de muestreo en la costa de Murcia; C. Distribución de las estaciones de muestreo en la reserva marina del Cabo de Palos–Islas Hormigas. Las estaciones invadidas se indican mediante círculos negros, cuyo tamaño corresponde a una de las cinco categorías de nivel de colonización. En el apéndice 1 se incluye la información relacionada con las 42 estaciones de muestreo.

Ruiz et al.

Table 1. Sampling stations colonised by Caulerpa racemosa along the coast of Murcia. For number of stations see figure 1 and appendix 1 and for level of colonisation see figure 1: S. Station; Y. Year of colonisation; L. Level of colonisation; N/A. Data not available. Tabla 1. Estaciones de muestreo colonizadas por Caulerpa racemosa a lo largo de la costa murciana. Para el número de las estaciones ver la figura 1 y el apéndice 1 y para el nivel de colonización ver la figura 1: S. Estación; Y. Año de colonización; L. Nivel de colonización; N/A. Datos no disponibles. S

Locality

2005 (m2)

2006 (m2)

2007 (m2)

La Manga

2007

Isla Grosa

2006

221

104

Piles I

2007

Isla Hormiga

2007

La Barra

2007

104

Los Punchosos

2007

4.5 x 10

II III

Calblanque

2005

300

N/A

2.5 x 10

Cabo Negrete

2007

104–105

III

C. Tiñoso

2006

13,724

9 x 105

maximum in this (unpublished data) and other regions of the Western Mediterranean (Klein & Verlaque, 2008; but see Cebrian & Ballesteros, 2009). Samples were collected at three of the most invaded stations: station 1 (–10 m), station 14 (–25 m), and station 25 (–22 m) (fig. 1B). Fronds, stolons and rhizoids were carefully collected by hand within six replicated 1,600 cm2 quadrats that were randomly placed within fully colonised areas (i.e. 100% cover) along a 50 m transect. Samples were processed in the laboratory to determine the following biometric variables as described by Capiomont et al. (2005) and Ruitton et al. (2005b): the total length of stolons (m/m2), number of stolon apices (no. apices/m2), number of fronds

(no. fronds/m2) and frond height (cm). Total biomass (g dw/m2) was determined by drying the samples at 70ºC until constant weight. Results Distribution and estimated colonised area Field data obtained at the invaded localities are summarised in figure 1 (B and C) and table 1. C. racemosa was first detected in station 14 (locality of Cablanque) in 2005 as dispersed patches covering a total surface area of more than 104 m2. By 2007, the

Table 2. Biometric analysis of the Caulerpa racemosa populations studied (mean values ± SD): S. Station number; D. Depth (m); B. Total biomass (g dw/m2); A. Number of apices (No. apices/m2); F. Number of fronds (No. fronds/m2); Sl. Total stolon length (m/m2); Fh. Frond height (cm). Tabla 2. Análisis biométrico de las poblaciones de Caulerpa racemosa estudiadas (valores medios ± DE): S. Número de estación; D. Profundidad (m); B. Biomasa total (g ps/m2); A. Número de ápices (nº de ápices/m2); F. Número de frondas (nº frondas/m2); Sl. Longitud total de los estolones (m/m2); Fh. Altura de las frondas (cm).

Locality

5,401 ± 2,631 3,487 ± 1,073

Isla Grosa

62.7 ± 42.7

1,133 ± 958

Calblanque

16.9 ± 7.3

323 ± 187

1,260 ± 589

3,913 ± 1,133

3.1 ± 0.3

1.6 ± 0.4

C. Tiñoso

49.7 ± 21.0

2,238 ± 832

6,256 ± 1,316

4,528 ± 798

1.9 ± 0.7

Animal Biodiversity and Conservation 34.1 (2011)

Station 4 (Isla Grosa) 2006

9,721 m2

221 m2

150

2007

300 m

Station 25 (Cabo Tiñoso) 2006

13,724 m2

El Muellecico

89,187 m2

El Muellecico El Arco

2007

El Arco

150 300 m

Caulerpa racemosa Fig. 2. Distribution and estimated surface area colonised by Caulerpa racemosa in: A. Station 4 (Isla Grosa); B. Station 25 (Cabo Tiñoso) in 2006 and 2007. Black arrows indicate the presence of new small patches of the alga. Fig. 2. Distribución y superficie del área colonizada por Caulerpa racemosa en: A. Estación 4 (Isla Grosa); B. Estación 25 (Cabo Tiñoso) en 2006 y 2007. Las flechas negras indican la presencia de nuevas pequeñas zonas del alga.

colony had formed a more homogenous meadow of at least 2.5·106 m2. After 2005, the number of newly invaded localities increased almost exponentially: two in 2006 and six in 2007 (table 1). In 2006, the population of station 4 (locality of Isla Grosa) was first found as a few small patches over a total surface area of 221 m2 that increased to 104 m2 in 2007 (figs. 1B, 2B). In station 25 (locality of Cabo Tiñoso), the initial surface area in 2006 was estimated as 13,724 m2 and this increased to 89,187 m2 in 2007. In 2007, all new sightings were concentrated along the easternmost coast of the region (stations 1, 6, 9, 12, 13 and 19) with very different colonisation levels, ranging between categories I and III (fig. 1B, 1C; table 1). The cumulated field data gave a gross estimation of the total invaded area of 265 ha in 2007, which is probably an underestimation of the

real colonised area since information on areas deeper than 30 m was not available and some coastal zones were excluded from the survey. Characteristics of the colonised areas The depth of invaded areas ranged from 2–30 m, but the maximum colonised depth was greater than 30 m since deeper stands continued further to this isobath (annex I). Shallow colonies (< 10 m) were the least frequent while most of the studied colonies fell within 10–30 m. The alga colonised a wide range of substrates and native assemblages: rocky photophilic algae (boulders and vertical walls), infralittoral and circalittoral soft–bottoms, dead mats of P. oceanica, and maerl beds (annex 1). In most localities, C. racemosa formed compact multilayered

mats up to 12 cm thick over the substrate. Cabo Tiñoso was the only locality where the P. oceanica meadow was partially invaded by the alga, but no penetration of the seagrass canopy was observed at the other localities. Biometric characterization of the colonies Table 2 summarises the biometric characteristics of C. racemosa colonies at three selected stations: 4, 14 and 25. The total biomass varied between 9.4 and 13.9 g dw/m2. The number of stolon apices ranged from 150 to 3,756/m2, while the total number of fronds ranged from 937 to 9,018.7/m2. In addition, the total length of stolons ranged from 1,684 to 5,777 m/m2, while the height of fronds varied between 0.3 and 9.5 cm. Discussion C. racemosa was observed for the first time on the coast of Murcia as an isolated colony at the locality of Calblanque (station 14) in 2005. The origin and the introducing vector of the alga in the region is unknown, but two hypotheses can be given: (1) dispersion from the nearest colonies, located in the province of Alicante 90 km to the north; and (2) introduction through the nearby harbour at Cartagena, which is a crucial point for the very dense maritime traffic supported by this part of the Mediterranean Sea (fig. 1B). Further regional dispersion in subsequent years occurred in an almost exponential manner and new colonies appeared without a clear spatial pattern in localities separated by hundreds of metres to tens of kilometres. This rapid and discontinuous regional spread is similar to that described by Langar et al. (2002) on the Tunisian coast and by Ruitton et al. (2005a) along the French Mediterranean coast. This pattern of spread has been attributed to the efficient reproductive mechanisms reported for C. racemosa, both sexual (Panayotidis & Žuljević, 2001) and vegetative (Renoncourt & Meinesz, 2002), that determine its higher colonisation potential relative to other invasive Caulerpales (e.g. C. taxifolia, Meinesz, 2007). In 2007 the most widespread population of C. racemosa was found in station 14 at the locality of Calblanque, the site where it was first sighted. However, there was no relationship between the actual colony size and the time elapsed since it was first observed, as indicated by the large variation in the estimated surface area between new colonies detected in 2006 and 2007 (1 to 105 m2). This is because at some localities the alga appeared before the date of its first sighting, but was not detected in the preceding year. The alga was probably already present as one or a few small inconspicuous patches that would be difficult to find, even by trained divers, but this implies that C. racemosa is able to spread over a surface area of at least 1 ha in a 1–year period, which represents a very fast colonisation rate. Station 25 at the locality of Cabo Tiñoso is a good example: the invaded area increased 44–fold in 1 year (i.e. 7.5 ha/year). Similarly, the alga colonised a surface area of almost 1 ha in

Ruiz et al.

2007 in station 4 from only few small patches found one year earlier (fig. 2). Similar spreading dynamics have been reported from other Mediterranean localities (Piazzi et al., 1997; Piazzi & Cinelli, 1999; Piazzi et al., 2001; Ruitton et al., 2005b), showing that once the alga arrives at a locality, substrate colonisation can be a very rapid process. This could be due to the high stolon elongation rate (up to 2 cm/day) and reproductive capacity of the alga (Panayotidis & Žuljević, 2001; Ceccherelli & Piazzi, 2001; Renoncourt & Meinesz, 2002; Ruitton et al., 2005b), but habitat characteristics such as substrate type and the complexity of native communities are also a determinant. As described for other Mediterranean localities (Ruitton et al., 2005a; Žuljević et al., 2003; Piazzi et al., 2005; Klein & Verlaque, 2008; Piazzi & Balatta, 2009), the alga fully colonised a variety of substrates and biocenoses present throughout its depth distributional range (i.e. rocky photophilic algae, dead rhizomes of P. oceanica, detritic soft bottoms and maerl beds), with the exception of P. oceanica meadows and unstable soft–bottoms. Accordingly, other studies have shown that P. oceanica meadows are among the most resistant habitats to invasion (Occhipinti–Ambrogi & Savini, 2003; Piazzi & Cinelli, 1999; Katsanevakis et al., 2010; Infante et al., 2011). Penetration of the algae in P. oceanica meadows only occurs in low–density canopies (Ceccherelli et al., 2000; Montefalcone et al., 2007.) Biometric analysis of C. racemosa meadows indicated a high degree of vegetative development in the studied colonies. The mean values of biometric descriptors were within the ranges observed in other invaded localities of the Western Mediterranean at a similar depth range (Buia et al., 2001; Ruitton et al., 2005b; Capiomont et al., 2005; Klein & Verlaque, 2008, Cebrian & Ballesteros, 2009). The mean values of total stolon length found in this study (3–4.5 km/m2) were higher than the maximum values reported in these studies for a similar season (summer–autumn: 0.3–1.2 km/m2; Capiomont et al., 2005; Ruitton et al., 2005b; Ivesa & Devescovi, 2006; Cebrian & Ballesteros. 2009). This extensive stolon development indicates overgrowth over the colonised substrate (Capiomont et al., 2005) and hence, a high potential impact on the native assemblages, particularly in those with lower vertical stratification (Balata et al., 2004; Piazzi & Balata, 2009) such as the maerl beds observed in deeper stations (e.g. station 14: Calblanque, 25 m). Nonetheless, the clear reduction of algal biomass with the depth of the sampling station (table 2) suggests that the potential future impact on these deep native assemblages could be lower than that on shallower ones. Of course, there is not a general, consistent vertical pattern of algal abundance across the different invaded localities of the Mediterranean coast, due to local differences in a great variety of abiotic and biotic factors (Cebrian & Ballesteros, 2009). However, the vertical pattern of biomass reported in this and other studies (e.g. De Biasi et al., 1999) suggests that changes of environmental factors associated to depth gradients such as light and temperature (which are also primary factors

Animal Biodiversity and Conservation 34.1 (2011)

of algal growth) could be responsible for the observed reduction of algal biomass with depth. The enlargement of C. racemosa fronds observed in the deepest station Calblanque is consistent with results obtained in other deep colonized areas (Cebrian & Ballesteros, 2009) and is a typical photoacclimatory response of these benthic macrophytes to light limitation (e.g. Ohba & Enomoto, 1987; Kirk, 1994), supporting the hypothesis that light reduction could contribute to explain (at least in part) the vertical pattern of algal biomass reported in this study. After this study, new sightings of the alga has been reported since 2008 confirming its presence along the west coast of Murcia and in the neighbouring province of Almeria, indicating that its geographical propagation is still continuing in a westerly direction. Accordingly, C. racemosa appeared in Algeria in 2006 (Ould–Ahmed & Meinesz, 2007) and in Ceuta in 2007 (Rivera–Ingraham et al., 2010). The estimated rate of spread of the alga in the Murcia region (at least 265 ha in three years) is higher than that reported in the Marseille region (France) over six years (350% increase; Ruitton et al., 2005a), but is similar to the highest values reported for southern latitudes (coast of Tuscany, Italy) for a similar time period (Piazzi et al., 1997). This evidence, together with the high degree of stolon development reported in this study, suggests that the highly invasive nature of C. racemosa on the SE coast of Spain may be favoured by the higher temperatures and irradiance characteristic of these southern Mediterranean latitudes. However, these regional variations should be interpreted with caution due to the low number of studied cases, and the use of different methodologies and experimental conditions (Klein & Verlaque, 2008). It is clear that the invasion by this and other introduced species represents a serious potential threat for native marine communities, but the real ecological consequences and their economic impact (local fisheries, tourism) are subjects that need to be addressed further. Acknowledgements This article is part of the research project Seguimiento de las praderas de Posidonia oceanica de la Región de Murcia funded by the Servicio de Pesca y Acuicultura de la Comunidad Autónoma de Murcia, Fondo Europeo de la Pesca (FEP) and by the Instituto Español de Oceanografía (IEO). We are grateful to Elena Fernández Aracil, Isabel M. Pujante Rodríguez (Universidad de Cádiz) and Noelia Salazar (Universidad Autónoma de Madrid) for their help in processing the biological samples in the laboratory. We are especially grateful to the diving centres of Murcia, volunteer and professional divers and other research groups who have collaborated in this project and have alerted us to the presence of C. racemosa. We thank also the Servicio de Reservas Marinas (Ministerio de Medio Ambiente y Medio Rural y Marino) and the Servicio de Protección y Conservación de la Naturaleza (Comunidad Autónoma de la Región de Murcia).

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Ivesa, L. & Devescovi, M., 2006. Seasonal vegetation patterns of the introduced Caulerpa racemosa (Caulerpales, Chlorophyta) in the northern Adriatic Sea (Vrsar, Croatia). Periodicum Biologorum, 108: 111–116. Katsanevakis, S., Issaris, Y., Poursanidis, D. & Thessalou–Legaki, M., 2010. Vulnerability of marine habitats to the invasive green alga Caulerpa racemosa var cylindracea within a marine protected area. Marine Environmental Research, 70: 210–218. Kirk, J. T. O., 1994. Light and photosynthesis in aquatic ecosystems, 2nd edition. Cambridge Univ. Press, Cambridge. Klein, J. & Verlaque, M., 2008. The Caulerpa racemosa invasion: A critical review. Marine Pollution Bulletin, 56: 205–225. Langar, H., Djellouli, A. S., Sellem, F. & El Abed, A., 2002. Extension of two Caulerpa species along the Tunisian coast. Journal of Coastal Conservation, 8: 163–167. Meinesz, A., 2007. Methods for identifying and tracking seaweed invasions. Botanica Marina, 50: 373–384. Montefalcone, M., Morri, C., Peirano, A., Albertelli, G. & Bianchi, C. N., 2007. Substitution and phase shift within Posidonia oceanica seagrass meadows of NW Mediterranean Sea. Estuarine, Coastal and Shelf Science, 75: 63–71. Nizamuddin, M., 1991. The Green Marine Algae of Libya. Elga Publisher, Bern. Occhipinti–Ambrogi, A. & Savini, D., 2003. Biological invasions as a component of global change in stressed marine ecosystems. Marine Pollution Bulletin, 46: 542–551. Ohba, H. & Enomoto, S., 1987. Culture studies on Caulerpa (Caulerpales, Chlorophyceae) II. Morphological variation of C. racemosa var. laetevirens under various culture conditions. Japanese Journal of Phycology, 25: 178–188. Ould–Ahmed, N. & Meinesz, A., 2007. First record of the invasive algae Caulerpa racemosa (Caulerpales, Chlorophyta) on the coast of Algeria. Cryptogamie, Algologie, 28(3): 303–305. Panayotidis, P., 2006. On the enigmatic origin of the Mediterranean invasive Caulerpa racemosa (Caulerpales, Chlorophyta). Mediterranean Marine Science, 7: 119–121. Panayotidis, P. & Žuljević, A., 2001. Sexual reproduction of the invasive green alga Caulerpa racemosa var. occidentalis in the Mediterranean Sea. Oceanologica Acta, 24(2): 199–203. Piazzi, L. & Balata, D., 2009. Invasion of alien macroalgae in different Mediterranean habitats. Biological Invasions, 11: 193–204. Piazzi, L., Balata, D., Ceccherelli, G. & Cinelli, F., 2001. Comparative study of the growth of the two co–occurring introduced green alga Caulerpa taxifolia and Caulerpa racemosa along the Tuscan coast (Italy, western Mediterranean). Cryptogamie,

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Appendix 1. Summary of the information related to the 42 sampling stations: localities, coordinates UTM, surveyed depth (colonized depth) and habitat/biocenosis present (RF. Rocky photophilous; RS. Rocky sciaphylous; IS. Infralittoral sands; D. Detritic; DM. Maerl; P. P. oceanica meadows; PU. P. oceanica meadows, upper limit; PD. P. oceanica meadows, deep limit; PM. P. oceanica–death 'matt'): S. Station number: L. Locality; D. Depth (m); C. Colonised; NC. Not colonised. Apéndice 1. Resumen de la información relacionada con las 42 estaciones de muestreo: localidades, coordenadas UTM, profundidad estudiada (profundidad colonizada) y hábitat/biocenosis presente (RF. Rocoso fotófilo; RS. Rocoso esciófilo; IS. Arenas infralitorales; D. Detrítico; DM. Maërl; P. Praderas de P. oceanica; PU. Praderas de P. oceanica, límite superior; PD. Praderas de P. oceanica, límite de profundidad; PM. "Matas muertas" de P. oceanica): S. Número de la estación; L. Localidad; D. Profundidad (m); C. Colonizado; NC. No colonizado.

UTM X

Habitat/biocenosis

La Manga

703019 4182444

Tomás Maestre

700822 4179442

5–7

Isla Grosa

701766 4178400

4–8

Isla Grosa

701985 4177946

4–12

Cala Túnez

703513 4168161

5–8

Piles I

704710 4168591 18–22 (20)

Piles II

704991 4168756

10–20

Bajo de en medio

706008 4168409

10–20

Isla Hormiga

707082 4170103

10–25

NC NC

10 C. Escalera–somera 703946 4167629

10–12

11 C. Escalera–profunda 703966 4167543

20–24

26–27

D DM

P PU PD PM

C NC

NC NC

12 La Barra

702954 4167457

2–4

13 Los Punchosos

702866 4166766

5–13 (7)

14 Calblanque

700052 4161847

25–26

C NC

15 Calblanque

693773 4160728

NC NC

16 Calblanque

694689 4160208

NC NC

17 Calblanque

695176 4159498

NC NC

18 Calblanque

697283 4159339

19 Cabo Negrete

697404 4159360

C NC

20 Cabo Negrete

697569 4159519

NC NC

21 El Gorguel

687416 4160406

5–17

22 Cabo de Agua

683908 4158528

5–17

23 Punta del Aguilón

682494 4158927

5–17

NC NC

24 Isla de las Palomas 673129 4160784

19–22

25 C. Tiñoso

664394 4156520

5–30

26 C. Tiñoso

663125 4156704

0–12

27 C. Tiñoso

663172 4156648

21–25

28 La Azohía

661074 4157999

17–21

29 Mazarrón

656472 4159510

17–21

NC NC

30 Mazarrón

655306 4159294

17–21

NC NC

31 Bolnuevo

646505 4157273

NC NC

32 Bolnuevo

645186

24–25

NC NC

4156116

NC NC

NC NC C

NC NC NC NC

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Appendix 1. (Cont.)

UTM X

Habitat/biocenosis RF

DM P PU PD PM

33 Bolnuevo

644435 4155026

17–18

NC NC

34 Calnegre

643447 4154594

16–17

NC NC

35 Calnegre

642439 4153362

19–20

NC NC

36 Calnegre

641492 4151921

24–25

NC NC

37 Calabardina

632933 4142986

7–15

38 Isla del Fraile

629651 4141582

14–16

NC NC NC NC

39 Punta Parda

622786 4137131

NC NC

40 Punta Parda

622625 4136981

NC NC

41 Punta Parda

621972 4136136

42 Punta Parda

621615 4136042

NC NC

NC NC NC

Animal Biodiversity and Conservation 34.1 (2011)

Conservation status and updated census of Patella ferruginea (Gastropoda, Patellidae) in Ceuta: distribution patterns and new evidence of the effects of environmental parameters on population structure G. A. Rivera–Ingraham, F. Espinosa & J. C. García–Gómez Rivera–Ingraham, G. A., Espinosa, F. & García–Gómez, J. C., 2011. Conservation status and updated census of Patella ferruginea (Gastropoda, Patellidae) in Ceuta: distribution patterns and new evidence of the effects of environmental parameters on population structure. Animal Biodiversity and Conservation, 34.1: 83–99. Abstract Conservation status and updated census of Patella ferruginea (Gastropoda, Patellidae) in Ceuta: distribution patterns and new evidence of the effects of environmental parameters on population structure.— The Strait of Gibraltar has important populations of the highly endangered patellid limpet Patella ferruginea. Between 2006 and 2010, an exhaustive census was carried out in Ceuta. The total coastline was divided into 17 sectors. The coast of each sector was examined by using 10 m transects. For the case of those sectors composed of breakwaters, jetties or islets, no transects were used, and instead, the total number of individuals was recorded. Each individual was measured to the nearest millimetre using a calliper. Moreover, the complete rocky shore length where the species could potentially be present was calculated, and an estimation of the total number of individuals that each sector could host was made. Results indicate that Ceuta could be home to around 44,000 individuals. The species found in Point Benzú, its westernmost limit of distribution on the North African coasts. The largest populations were recorded on the South Bay, with higher Mediterranean influence. Our results indicate that substrate roughness (topographic heterogeneity) and the area’s accessibility highly influence the abundance and population structure. Those populations located on high topographic heterogeneity substrates show higher recruitment rates and lower percentages of larger individuals, while medium to low rugosity surfaces presented the opposite pattern. Additionally, easily accessible areas (and frequented by humans) presented smaller average shell sizes. Implications of the results for conservation purposes are discussed. Key words: Limpet, Patella ferruginea, Endangered species, Conservation, Substrate roughness, Strait of Gibraltar, Ceuta. Resumen Estado de conservación y censo actualizado de Patella ferruginea (Gastropoda, Patellidae) en Ceuta: patrones de dsitribución y nueva evidencia de los efectos de los parámetros ambientales en la estructura de la población.— El Estrecho de Gibraltar presenta importantes poblaciones de la lapa Patella ferruginea altamente amenazada. Entre 2006 y 2010, se llevó a cabo un censo exhaustivo en Ceuta. La totalidad de la costa fue dividida en un total de 17 sectores. Se examinó el litoral de cada uno de estos sectores empleando transectos de 10 m. Para el caso de aquellos sectores compuestos por escolleras, espigones o islotes, no se emplearon estos transectos, y se registró el número total de individuos. Cada individuo fue medido por medio de un pie de rey con precisión de un milímetro. Además, se estimó la longitud total de costa rocosa que podría potencialmente presentar individuos y se realizó una estimación del número total de individuos que podría albergar cada sector. Los resultados indican que Ceuta podría presentar en torno a 44.000 individuos. La especie demostró tener en Punta Benzú su límite occidental de distribución en el Norte de África. Las poblaciones más importantes fueron registradas en la Bahía Sur, con importante influencia mediterránea. Nuestros resultados indican que la rugosidad del sustrato (heterogeneidad topográfica) y la accesibilidad de la zona influencian en gran medida la abundancia y la estructura de las poblaciones. Aquellas que se localizan sobre sustratos de alta heterogeneidad topográfica muestran mayores tasas de reclutamiento y menores porcentajes de individuos de gran tamaño, mientras que los sustratos de media a baja rugosidad muestran el patrón contrario. Además, aquellas áreas

ISSN: 1578–665X

Rivera–Ingraham et al.

fácilmente accesibles (y frecuentadas por el ser humano) presentan poblaciones con tallas medias menores. Se discuten las posibles implicaciones de los resultados de cara a la conservación de la especie. Palabras clave: Lapa, Patella ferruginea, Especies amenazadas, Conservación, Rugosidad del sustrato, Estrecho de Gibraltar, Ceuta. Georgina Alexandra Rivera–Ingraham, Free Espinosa & Jose Carlos García–Gómez, Lab. de Biología Marina, Depto. de Fisiología y Zoología, Univ. de Sevilla, Av. Reina Mercedes 6, 41012 Sevilla, España (Spain). Corresponding author: G. A. Rivera–Ingraham. E–mail: grivera@us.es

Animal Biodiversity and Conservation 34.1 (2011)

Introduction The patellid limpet Patella ferruginea is an intertidal species endemic to the Mediterranean. It was widely distributed throughout the Western Mediterranean basin during the Pleistocene (Caton–Thompson, 1946; Laborel–Deguen & Laborel, 1991a) and its presence on European and North African coasts was recorded until the end of the 19th century. But it was in the beginning of the 20th century when the species started to suffer clear regression (Laborel–Deguen & Laborel, 1991a; Templado, 2001). Nowadays, the species has almost completely disappeared from European waters (Templado & Moreno, 1997). Its presence has been recorded in Southern Spain (Espinosa, 2006; Moreno & Arroyo, 2008) (where 700 individuals have been located, most of them present in Algeciras Bay). It can also be found on the coasts of Alborán Island (Paracuellos et al., 2003; Templado et al., 2006), North and Western Corsica (Laborel–Deguen & Laborel, 1991a; Curini–Galletti, pers. com.; Rivera–Ingraham et al., pers. obs.), Sardinia (Porcheddu & Milella, 1991; Doneddu & Manunza, 1992; Cristo et al., 2007; Cristo & Caronni, 2008; Rivera–Ingraham et al., pers. obs.), Pantellaria and Egadi Islands (Laborel–Deguen & Laborel, 1991a) and Tuscany (Italian peninsula) (Curini–Galletti, 1979; Biagi & Poli, 1986). In any case, the most important populations are currently located on the North African Coasts: Ceuta (Guerra–García et al., 2004; Espinosa, 2006; Espinosa et al., 2009), Melilla (González García et al., 2006), Chafarinas Islands (Guallart et al., 2006), Morocco (Bazairi et al., 2004), the Algerian isles of Rachgoun (Frenkiel, 1975) and Habibas (Boumaza & Semroud, 2001; Espinosa, 2009) and reaching Cape Bon (Espinosa, 2006) and Zembra Island (Boudouresque & Laborel–Deguen, 1986) (Tunisia). This regression has been mainly caused by human exploitation (Aversano, 1986; Guerra–García et al., 2004; Moreno, 2004) which takes place in the highly accessible intertidal habitat (Raffaelli & Hawkins, 1996; Haedrich & Barnes, 1997; Rochet & Trenkel, 2003) where the species lives. It is commonly collected to be used as food, as fishing bait because of its muscular foot (Pombo & Escofet, 1996) or even for ornamental purposes, as happens with other big intertidal limpets such as Lottia gigantea (Lindberg et al., 1998; Kido & Murray, 2003). The species is presently considered as the most endangered marine invertebrate species on the Western Mediterranean rocky coasts (Laborel–Deguen & Laborel, 1991a; Ramos, 1998), and is protected at a European level (Annex II of Berne Convention; Annex IV of the Habitats Directive; Annex II of Barcelona Convention) as well as by Spanish laws (National Catalogue of Threatened Species; Andalusian Catalogue of Endangered Species). Little is known regarding the biology of the species (Guerra–García et al., 2004) although many studies have been recently conducted in order to provide more information (e.g. Espinosa, 2006; Rivera–Ingraham, 2010). In this sense, the Spanish National Strategy for the Conservation of P. ferruginea has also been recently approved (MMAMRM, 2008) with the objective of pro-

moting activities that could contribute to the knowledge of the species and obtain the necessary information to propose adequate conservation measures. One of these objectives is to obtain updated quantitative data regarding the distribution and the conservational status of the species, by developing detailed censuses. Some other authors like Paracuellos et al. (2003) have also pointed out the interest of quantifying the remaining P. ferruginea populations. The aim of the present study was to achieve a complete description of the metapopulation of P. ferruginea in Ceuta, and to estimate the total number of individuals present in the area. Special attention is paid to the influence of substrate roughness and the area’s accessibility on the species distribution and population structures. Material and methods Study site The study was conducted in Ceuta, located on the African coast of the Strait of Gibraltar (fig. 1A), which is known to have important P. ferruginea populations (see Guerra–García et al., 2004; Espinosa, 2006; Espinosa et al., 2009). The coasts of this city are composed of natural rocky shores, beaches and small rocky islets. We also find many artificial jetties (especially on the South Bay). The city’s commercial port has the peculiarity of being connected with the North Bay (through the port’s main entrance) and with the South Bay through a moat. Sampling methods After the inspection of Ceuta’s coastline, all areas presenting natural or artificial rocky shores that could potentially present P. ferruginea individuals were recorded. In order to estimate the total number of P. ferruginea individuals present in the area, the complete coastline was divided into 17 sectors (fig. 1B, table 1). However, within some of these sectors, some clearly differentiated structures (islets, breakwaters, jetties, etc.) were separately taken into account (refer to table 2). Later, for each of the sectors considered and only for rocky shores, a minimum of five transects (each composed of 10 m) running parallel to the shore were randomly established on the coastline with the help of a metric tape. Each P. ferruginea individual located within these transects was measured to the nearest millimetre using a calliper (Guerra–García et al., 2004; Espinosa, 2009; Rivera–Ingraham, 2010). Small individuals are often difficult to detect (Guallart et al., 2006), so special attention was paid to avoid missing this fraction of the population. For those sectors where the census was carried out using transects, the total length of rocky shoreline where the species could potentially be present was calculated by using 1:9,000 maps obtained using Google Earth ©. With this value and the average density recorded, an estimation of the total number of individuals within each sector was obtained. However, for the specific case of walls,

Rivera–Ingraham et al.

A 14º 45º

9º

4º

1º

6º

40º

40º Mediterranean Sea

Atlantic Ocean 35º

35º

14º B

5.23º Q

9º

5.22º

35.54º

4º

1º

5.21º 5.20º 5.19º 5.18º North Bay

O Ceuta

2.5 km

L C

35.53º

500 km 5.17º 35.55º

M N

Morocco

6º

I H

K E

G 35.53º

A B South Bay 5.20º

5.19º 5.18º

35.52º 5.17º

Fig. 1. A. Location of Ceuta; B. Sectors the study site was divided into to estimate the total number of individuals of P. ferruginea. Fig. 1. A. Localización de Ceuta; B. Sectores en los que fue dividida la zona de estudio para la estimación del número total de individuos de P. ferruginea.

jetties, breakwaters and those islets that are not completely submerged during high tide, the total perimeter was measured, again using a metric tape, and a complete census was carried out when possible. The complete coastline was inspected from 2006 to 2010. Additionally, and for each of the prospected locations, some physical parameters were recorded: i) type of substrate (natural/artificial), considering as an artificial substrate any man–made structure, regardless of the origin of the substrate itself (e.g. Fauvelot et al., 2009; Bulleri & Chapman, 2010); ii) accessibility by humans, dividing areas into three categories: (a) high, easily accessible areas where it is common to find people collecting intertidal organisms; (b) medium, areas relatively easy to access and where it is not frequent find people at intertidal levels;

and (c) low, areas which are difficult or cannot be accessed by land, and where no collection presumably occurs; and iii) substrate roughness or topographical heterogeneity index (THI), calculated as in Blanchard & Bourget (1999): THI = Tr/Ts, where Tr is the actual distance between two points (measured following the substrate’s irregularities) and Ts the linear distance between the two same points. Substrates were again classified into three categories: (a) high roughness (THI >1.30); (b) medium roughness (THI = 1.30–1.15); and (c) low roughness (THI < 1.15). It is also important to note that in the present study, the total number of individuals present in Ceuta is considered to constitute a metapopulation, which would be composed of different sub–populations that are genetically connected (see review by Badii & Abreu, 2006).

Animal Biodiversity and Conservation 34.1 (2011)

Table 1. Data corresponding to each of the considered sectors: TR.Total rocky shore length (m); N. Number of individuals recorded; SL. Shore length inspected; T. Total number of individuals Tabla 1. Datos correspondientes a cada uno de los sectores considerados: TR. Longitud total de la orilla rocosa (m); N. Número de individuos registrados; SL. Longitud de costa inspeccionada; T. Número total de individuos.

Sector A B C D E F G H I J K L M M N O P Q

Sector Name Frontera and P. Pineo P. Gordas and P. Brazo Chorrillo and Foso jetties Fuentecaballos Mellizos Sarchal Desnarigado Desnarigado–Point Almina Point Almina–Point Sirena San Amaro Dique de Levante Inner port Dique de Poniente (concrete cube section) Dique de Poniente (limestone section) Benítez Desaladora–Point Bermeja Point Bermeja–Point Blanca Point Blanca–Benzú

Total

TR 300 311 712 350 1,142 350 1,710 2,201 1,271 959 555 3,754 1,156

N 195 2,260 3,518 2,262 104 106 297 479 440 19 709 4,111 23

SL 100 311 712 350 100 100 100 258 100 100 100 3,754 30

T 585 2,260 3,518 2,262 1,187 371 5,079 4,086 5,592 182 3,934 4,111 886

1,115

244

5,441

316 1,217 1,138 900

273 351 42 19

100 143 100 100

863 2,987 477 171

19,457 m

15,452 ind.

6,608 m

43,992 ind.

Statistical analyses Univariate analyses were carried out using the statistical package SPSS 15.0. One and two–way ANOVA tests were carried out between several physical parameters (nature of substrate (natural/ artificial), sector location (North Bay, South Bay or Port) and substrate roughness) and the most common population parameters (density of individuals per size class, adult average shell size, maximum shell size, adult density and total density). Finally, and taking into account that the reduction of data to summary statistics (such as means, medians, etc.) for comparisons can significantly reduce the amount of available information (Sagarin et al., 2007), multivariate analyses were also conducted to compare size distributions among populations, as has been satisfactorily used by other authors (e.g. Sagarin et al., 2007; Espinosa, 2009). In order to do this, the total number of individuals for each size class (1 cm intervals) and sector was used. But considering that

the total length of shoreline inspected varied considerably between locations, these frequency values were standardized by transforming them to percentages (over the total number of recorded individuals in the sector). Additionally, these data were later transformed to log (x+1) to homogenize variances. An MDS (Multi–dimensional scaling) analysis was carried out using PRIMER–E v.6.0 and based on the UPGMA (Unweigh Pair–Group Method using arithmetic means) method and the Bray–Curtis similarity index (Bray & Curtis, 1957). Moreover, the Kruskal stress coefficient was used to determine ordination (Kruskal & Wish, 1978). Results P. ferruginea distribution and abundance Throughout the study, a total of 6,608 m of Ceuta’s rocky shores was surveyed (33.96% of the total rocky shore that could potentially house P. ferruginea

Rivera–Ingraham et al.

North Bay

Ceuta Morocco Fig. 3 South Bay 2.5 km

0 ind. < 2 ind./m 2–4 ind./m 4–6 ind./m > 6 ind./m

Fig. 2. General distribution of P. ferruginea in Ceuta, general view. Circle diameter corresponds to the density of individuals. Coastline sections plotted with thicker lines indicate the location of non–suitable areas (e.g. beaches) for the species. Light grey circles indicate the density recorded for each transect established (except for sectors H and I where only five transects are represented due to lack of space), while darker circles represent the density of individuals in an area where a complete census of the coastline was carried out. The discontinuous line indicates the inaccesssible areas of the port. Fig. 2. Distribución general de P. ferruginea en Ceuta, vista general. El diámetro de los círculos se corresponde con la densidad de individuos registrados. Aquellas secciones de costa marcadas con trama más gruesa indican la localización de zonas (p. ej. playas) no adecuadas para la especie. Los círculos claros representan la densidad registrada en cada uno de los transectos realizados (excepto para los sectores H e I para los que sólo se han representado cinco por falta de espacio), mientras que los oscuros representan la densidad media total en la zona tras la realización de un censo completo. La línea discontinua indica las zonas inaccesibles del puerto.

individuals), finding 15,452 individuals. Therefore, we estimated that Ceuta’s metapopulation is composed of 43,992 individuals (table 1). Figure 2 shows how the South Bay of Ceuta is home to more important P. ferruginea subpopulations. Although it is frequent to find P. ferruginea subpopulations in the North Bay, they became scarcer as we moved towards the Atlantic. In fact, the density of adult individuals (> 25 mm) in the South Bay (4.90 ± 7.46 ind./m) was significantly higher than the average values recorded for subapopulations located in the North Bay (1.44 ± 1.64 ind./m) and inside the commercial port (1.45 ± 2.02 ind./m) (x2 = 7.51; p = 0.023). It is also interesting to note the spatial distribution of individuals in certain areas such as the artificial breakwaters located in the southern area of the commercial port as well as on the jetties located in the South Bay (fig. 3). It was evident that the maximum densities of individuals were found within the inaccessible areas of Parque del Mediterráneo and the Guardia Civil’s base. Jetties also showed

important densities, especially in the inner sides of these structures (those oriented towards the moat exit) and preferentially in their endings. Effect of physical parameters on population structure Information regarding the subpopulations’ size structures is represented in figure 4 and table 2. When taking into account these values and the physical parameters recorded in each sector, the following results were obtained: Effect of the nature of substrate and sector location The nature of substrate (natural vs. artificial) did not seem to influence any of the above mentioned population parameters. However, a one–way ANOVA test showed that the areas inside the port presented the highest maximum shell size values (8.89 ± 1.16 cm), higher than the values recorded in the North Bay (7.18 ± 1.19 cm) and the South Bay (8.12 ± 1.06 cm) (F = 4.87; p = 0.016). In fact, it was inside the port

Animal Biodiversity and Conservation 34.1 (2011)

Foso jetty

Chorrillo eastern jetty

Fuentecaballos jetty

Chorrillo western jetty

Chorrillo main jetty

0 ind. < 2 ind./m 2–4 ind./m 4–6 ind./m > 6 ind./m

Fig. 3. Density of P. ferruginea individuals on the breakwaters of the southern area of the commercial port and on the jetties of the South Bay. The area represented has been completely censused. Circles represent density values grouped in 30 m stretches to visualize the spatial distribution of individuals. The inaccessible areas of the port (Parque del Mediterráneo and the Guardia Civil’s base) are surrounded by a discontinuous line. The small arrow indicates the location of a wastewater outfall. Note the important influence of substrate inaccessibility and the presence of the outfall on the density of individuals. The two large arrows indicate the direction of water currents with east winds (light grey arrow) and west winds (right dark grey arrow). Fig. 3. Densidad de individuos de P. ferruginea en las escolleras del sur del puerto comercial así como en los principales espigones de la Bahía Sur. Toda el área mostrada ha sido completamente censada. Los círculos representan valores de densidad agrupados en tramos de 30 m para la visualización de la distribución de individuos. Las zonas inaccesibles del puerto (Parque del Mediterráneo y la base del servicio marítimo de la Guardia Civil) estan enmarcadas con una línea discontinua. La flecha indica la localización de un emisario aguas. Nótese la importante influencia de la inaccesibilidad de la zona y la presencia del emisario sobre la densidad de individuos. Las dos flechas grandes indican la dirección de las corrientes de agua con vientos de levante (flecha clara) y vientos de poniente (flecha oscura).

(within the Guardia Civil’s military base) where the largest individual (10.7 cm) was located. No significant differences were recorded between the North and South Bays regarding this parameter. Effect of substrate roughness and accessibility on population structure After conducting a multivariate analysis of classification (based on population size structures) figure 5 was obtained. Figure 5A shows how for an 80% similarity, a total of four groups of subpopulations can be differenced: first, sectors P, F, G, K, N, C and D were found, all of them characterized for presenting a clear predominance of small (< 25 mm) and medium (25–50 mm) size individuals, and where the largest individuals (> 50 mm) represented less than 20% of the subpopulation. These subpopulations were subject to medium to high impact by collection, and were also located on medium to high

THI substrates (fig. 5B). The second group was composed of sectors L, H, I, A, B, E, M y Q. Subpopulations in these areas were mainly composed of medium and large size individuals, while small individuals (< 25 mm) only represented up to 25% of the subpopulation. Returning to figure 5B, it can be observed how individuals in these areas presented low impact by collection, and most of them were located on medium THI substrates. The cluster analysis detected sector J as significantly different from the rest of the areas considered, which was subject to medium collection impact. But the main characteristic of this area was the atypical substrates found in the area, composed of natural rocks of very low roughness. This was the only area where the largest individuals clearly dominated over the rest (47% of the total subpopulation). Finally, sector O was also clearly segregated from the rest of sectors, and it was the only area where juveniles (< 25 mm) clearly dominated the subpopulation (71%).

Rivera–Ingraham et al.

Table 2. Summary statistics for Patella ferruginea sizes (cm) for each subpopulation considered in the study: Sc. Sector; Sp. Subpopulation; TS. Type of substrate (N. Natural; Ar. Artificial riprap; Asw. Artificial sea wall); D. Density (ind./m); R. Recruit density (ind./m, < 25 mm); As. Average shell size (cm); Ms. maximum shell size (cm); K. Kurtosis († kurtosis is considered significative when its absolute value is greater than 2*SE kurtosis); Kt. Kurtosis type (Pk. Platikurtic; Lk. Leptokurtic ); Sk. Skewness (Skew is considered significative when its absolute value is greater than 2*SE Skew); Am. Asymmetry (P. Positive; N. Negative). Tabla 2. Resumen de las estadísticas de los tamaños de Patella ferruginea (en cm) para cada subpoblación considerada en el estudio. Sc. Sector; Sp. Subpoblación; TS. Tipo de sustrato (N. Natural; Ar. Escollera artificial; Asw. Muro marino artificial); D. Densidad (ind./m); R. Densidad de reclutamiento (ind./m, < 25 mm); As. Tamaño medio de la valva (cm); Ms. Tamaño máximo de la valva (cm); K. Curtosis († la curtosis se considera significativa cuando su valor absoluto es mayor que el doble de su EE); Kt. Tipo de curtosis (Pk. Platicúrtica; Lk. Leptocúrtica); Sk. Coeficiente de asimetría (la desviación se considera significativa cuando su valor absoluto es mayor que el doble de su EE); Am. Asimetría (P. Positiva; N. Negativa). Sc

A A B

Frontier shoreline 'Pineo' islet 'Piedras Gordas' islets

N N N

2.68 6.77 6.71

0.28 0.18 1.44

4.58 4.25 4.12

8.20 7.50 9.20

–0.796 –0.643 –0.866†

B C C C C D E F G H I J K L L L L L L L M M

'Brazo' islet Chorrillo main jetty Chorrillo western jetty Chorrillo eastern jetty Foso jetty Fuentecaballos jetty Mellizos Sarchal Desnarigado Desnarigado–Point Almina Point Almina–Point Sirena San Amaro Dique de Levante Parque del Mediterráneo Dique del Comercio S. M. Guardia Civil Muelle de España Muelle de Poniente Muelle de Babor Interior Foso D. Poniente (concrete cube section) D. Poniente

N 35.14 Ar 6.86 Ar 6.36 Ar 1.87 Ar 4.31 Ar 6.46 N 1.04 N 1.06 N 2.97 N 1.86 N 4.40 N 0.19 Ar 7.09 Ar 6.81 Ar 2.61 Ar 2.92 Asw 0.10 Asw 0.13 Asw 0.08 Asw 0.04 Ar 0.77

5.71 0.73 4.03 0.91 1.08 2.18 0.01 0.20 0.75 0.07 0.84 0.04 2.81 1.71 0.83 0.35 0.02 0.01 0.01 0.01 0.10

4.68 4.28 2.44 2.99 3.65 3.47 4.42 3.47 3.34 4.36 5.09 4.40 3.10 3.67 3.20 6.30 3.84 4.84 4.85 4.58 4.81

8.80 8.60 6.10 7.50 8.70 9.30 8.40 6.40 7.50 9.70 9.40 6.90 7.80 10.00 7.60 10.7 8.00 9.50 7.70 8.90 9.70

–1.157† –0.198 0.554† –0.247 –0.290 –0.406† 1.684† –0.326 –0.027 –0.842† –0.715† –1.032 –0.534† 0.881† 0.614† –0.571 –0.268 –0.399 –0.546 –0.695 –0.846

Pk 0.092 – Pk –0.246 – – –0.082 – Lk 0.899* P – 0.695* P – 0.407* P Pk 0.521* P Lk 1.091* P – 0.581 – – 0.252 – Pk 0.254* P Pk 0.212 – – –0.500 – Pk 0.372* P Lk 1.125* P Lk 0.664* P – –0.648* N – 0.573* P – 0.309 – – –0.113 – – 0.422 – – –0.186 –

4.88

1.18

3.58

6.60

–0.636†

0.198*

N O O P

(limestone section) Benítez Ar Desaladora–Point Bermeja N (natural substrate section) Desaladora–Point Bermeja Ar (breakwater section) Point Bermeja–Point Blanca N

2.73 0.95

1.02 0.26

3.00 3.31

6.60 6.90

–0.903† –0.641

Pk –

0.167 0.306

– –

4.36

3.62

1.68

5.60

2.169†

1.673*

0.42

0.05

3.69

5.90

0.914

–

–0.299

–

Point Blanca–Benzú

0.19

0.01

4.13

7.70

2.122

0.627

–

†

– –

Sk 0.132 –0.281

Am – –

Animal Biodiversity and Conservation 34.1 (2011)

Frequency

1,200 1,000 800 600 400 200 0

Frequency

1,200 1,000 800 600 400 200 0

Frequency

1,200 1,000 800 600 400 200 0

1,200

Frequency

1,000 800 600 400 200 0

Q Frequency

1,200

1 2 3 4 5 6 7 8 9 10 11 Size class (cm)

1,000 800 600 400 200 0

1 2 3 4 5 6 7 8 9 10 11 Size class (cm)

Fig. 4. Size frequencies for each of the sectors considered. Letters correspond to the code used in table 1 and figure 1: 1(0–1), 2 (1–2), 3 (2–3), 4 (3–4), 5 (4–5), 6 (5–6), 7 (6–7), 8 (7–8), 9 (8–9), 10 (9–10), 11 (10–11). Fig. 4. Frecuencias de tamaños para cada uno de los sectores considerados. Las letras corresponden al código usado en la tabla 1 y la figura 1. (Para las abreviaturas de los tamaños de clase, ver arriba.)

Two–way ANOVA tests were conducted to define the effects of the substrate’s accessibility and topographical heterogeneity on population structure and density. Results indicate that adult average shell size was influenced by accessibility (table 3). However, overall adult density was not apparently influenced by

either of these two factors. However, figure 6 shows how substrate roughness alone may be influencing subpopulation density and structure. In this figure, the differences in density of small individuals (< 25 mm) among substrates are noticeable, the maximum being observed in high THI surfaces (fig. 6). Substrate

Rivera–Ingraham et al.

Table 3. Two–way ANOVA results for the influence of the substrate’s accessibility and THI (topographical heterogeneity index) on adult average shell size (SS): n.s. Non significance; * P < 0.05. Tabla 3. Resultados del ANOVA de dos factores para la influencia de la accesibilidad y THI (índice de heterogeneidad topográfica) sobre la talla media de los adultos (SS): n.s. No significativo; * P < 0,05.

Source of variation

Mean ± SD (cm)

Accessibility

SS 4.043

F 5.439

High

4.04 ± 0.29

Medium

4.29 ± 0.61

Low

4.93 ± 0.71

Substrate THI

0.114

0.153

High

4.30 ± 0.87

Medium

4.65 ± 0.51

Low

4.81 ± 0.53

Acc x THI

roughness may also influence adult average size. In high THI substrates, adult individuals showed an average size of 4.30 ± 0.87 cm, while average values of 4.65 ± 0.51 cm and 4.81 ± 0.53 cm were recorded for subpopulations on medium and low THI surfaces, respectively. To corroborate the abovementioned results, the specific case of 'Dique de Poniente' was taken into consideration. This sector presents the same physical parameters throughout the area, except for the fact that half of the breakwater is composed of high roughness rocks while the second half was constructed using smooth 3 x 3 m concrete cubes. The density of individuals recorded for the transects in each of these two subareas was compared using a one–way ANOVA, which showed that the area with the highest THI presented significantly higher densities of P. ferruginea individuals (4.88 ± 1.82 ind./m) than the cube area (0.77 ± 0.76 ind./m) (F = 13.24; p = 0.01). Moreover, adult average size in the concrete cube area (5.08 ± 0.40 cm) was significantly higher than that recorded for adults located on the other section of the mole (4.09 ± 0.20 cm) (F = 22.16; p = 0.005). Discussion After inspecting 33.96% of the coast of Ceuta that could potentially present P. ferruginea individuals, we estimated that the city is home to around 44,000 individuals. The methodology used in the study has been used successfully by other authors to estimate the total number of individuals in metapopulations, such as that in Habibas Islands (Espinosa, 2009) or Zembra Island (Boudouresque & Laborel–Deguen, 1986). Our estimation is considerably higher than those obtained by other authors who determined

1.546

1.040

P 0.012*

n.s.

that Ceuta holds between 12,000 (Guallart et al., 2006) and 3,704 individuals (Ocaña et al., 2010). Although in the latter cases the sampling effort was considerably lower, during the present study we were able to physically count more than 15,000 individuals, supporting the fact that the aforementioned estimations clearly underestimated the total number of specimens in Ceuta. We can additionally highlight that our estimations can be considered conservative. It is known that the smaller the step (or scale) used in calculating a perimeter, the higher the values and better the estimations obtained (Mandelbrot, 1967). Taking into account that the perimeter of the coastline potentially used by the species has been calculated using 1:9,000 maps, we can suggest that we clearly underestimated this value, and in consequence, the total number of individuals. An additional focus of variability is found in the quantification of the smallest fraction of the population (< 25 mm), which is often difficult to detect (e.g. Guallart et al., 2006; Espinosa, 2009). Although censuses were always carried out by the same experienced team (reducing the probability of missing these individuals), a considerable fraction of this variability is induced by the inter–annual differences in recruitment rates that have been recorded in the area (Rivera–Ingraham, 2010) and by the time of the year in which the censuses were carried out (in relation to the species’ reproduction period). The most important P. ferruginea subpopulations (regarding density) were, in general terms, located in the South Bay of Ceuta, which is mainly influenced by Mediterranean waters. On one hand, this result may seem surprising, specially taking into account that one of the main problems that P. ferruginea encounters is human collection (Laborel–Deguen & Laborel, 1991a, 1991b; Templado & Moreno, 1997; Ramos, 1998), and that the South Bay’s coastline is generally very acces-

Animal Biodiversity and Conservation 34.1 (2011)

Transform Log (x+1) Resemblance: S17 Bray–Curtis similarity Bray–Curtis Similarity

A 70 75 80 85 90 95

100 P F G K N B

L H I A Sector s

2D stress 0.11 I

B L

C D G

N K

F P

Q J

Substrate rugosity

A M

High

Medium

Low

Collection impact

High Medium Low

Fig. 5. Multivariate analyses: A. Cluster analysis. Continuous lines indicate significantly different groups (SIMPROF analysis). B. Spatial representation of centroids (MDS) for each sector. Circle diameter is positively correlated with the area's substrate roughness (THI) (Blanchard & Bourget, 1999): High, > 1.30; Medium, 1.15–1.30; Low, < 1.15. Colours are associated with the different grades of impact by collection suffered by individuals in the area (High. Easily accessible areas, where it is common to find people recollecting intertidal macroinvertebrates; Medium. Areas of relatively easy access, although they do not present high impact by collection; Low. Areas with difficult or no access by land to the intertidal fringe, and where no people have been seen). Fig. 5. Análisis multivariante: A. Análisis de grupos. Las líneas continuas indican diferencias significativas entre grupos (análisis SIMPROF). B. Representación especial de centroides (MDS) para cada sector. El diámetro de los círculos está positivamente correlacionado con el coeficiente de rugosidad del área (THI) (Blanchard & Bourget, 1999): High, rugosidad alta > 1,30; Medium, rugosidad media, 1,15–1,30; Low, rugosidad baja, < 1,15. Los colores están asociados con los diferentes grados de impacto por recolección sufrido por los individuos de cada área (High. Áreas fácilmente accesibles, donde es habitual encontrar personas recogiendo macroinvertebrados intermareales; Medium. Áreas de relativamente fácil acceso, aunque no presentan altas tasas de impacto por recolección; Low. Áreas de difícil acceso o inaccesibles por tierra, donde no se han detectado personas durante todo el estudio).

sible and consequently highly frequented by fishermen. However, these subpopulations became scarcer as we moved towards the Northwestern coasts (with higher Atlantic influence). This seems a predictable pattern, taking into account that we are dealing with a Mediterranean endemic species. This type of distribution pattern has previously been pointed out by several

authors (Templado, 2001; Guerra–García et al., 2004; Espinosa, 2006). Guallart et al. (2006) considered that the western limit of distribution of this limpet on the North African coasts could be in Point Blanca, although we recorded considerably important subpopulations reaching Point Benzú (sector Q) (0.19 ind./m) (table 2). Indeed, no individuals have been found in the near

Point Leona or Point Cires, although the former is separated from our sector Q by a long beach, meaning that Benzú could actually constitute the western limit of distribution of this species in northern Africa. On the other hand, the above mentioned authors also indicated that the highest densities were recorded inside the port. However, our results indicate that the densities recorded in the South Bay were significantly higher than those recorded inside the port. In any case, it should be taken into account that the aforementioned authors only inspected the breakwaters inside the port of Ceuta (Parque del Mediterráneo), while the present study considered its complete extension. Vertical and smooth substrates (e.g. sea walls) are clearly dominant in the port, and this seriously affected the average density values recorded in the area as the species is not frequent on this type of substrate (pers. obs.). It is known that P. ferruginea is commonly present on artificial rocks (Guerra–García et al., 2004). This is corroborated by the observations and results obtained from the present study. However, one of the main differences we observed between natural rocky shores and artificial structures was related to substrate roughness and the presence of microstructures, which is supported by other authors (see review by Bulleri & Chapman, 2010). Our observations and results indicate that substrate THI may be playing an important role in recruitment processes and in consequence, also on the distribution of P. ferruginea. The substrates presenting the highest irregularities showed coefficients higher than 1.30 (using the equation provided by Blanchard & Bourget, 1999) and around 1.017 (Rivera–Ingraham, 2010) using fractal dimensions (Mandelbrot, 1967) [following an adapted method from Beck (2000) and using profile gauges to obtain rock profiles as in Frost et al. (2005)]. Statistical analyses showed that these surfaces presented the highest recruitment rates. It is known that the presence of irregularities in substrates is associated with high recruitment rates (as these structures can enhance settlement and provide shelter for juvenile limpets) (e.g. Creese, 1982) and may determine the distribution patterns of intertidal species such as barnacles (Crisp & Barnes, 1954) and gastropods (including limpets) (Beck, 2000; Underwood, 2004). However, it should be taken into consideration that high THI substrates also showed low density of large individuals. This would considerably reduce competition events on these surfaces and could also explain why such important recruitment rates have been recorded. An experimental study approaching the effect of substrate roughness and the density of adult conspecifics on P. ferruginea recruitment rates is highly recommended in order to support our observations and results. Based on this and the abovementioned studies (which support the hypothesis that high numbers of recruits are associated with substrate topographic heterogeneity), an ethological component could also be considered: larvae of many invertebrate species can actively select substrates during the settlement process (see reviews by Woodin, 1986; Butman, 1987; and results obtained by Wilson, 1990). Additionally, similar evidence of the influence of larvae behaviour on

Rivera–Ingraham et al.

settlement has also been pointed out for other limpet species like Cellana grata (Williams & Morritt, 1995), Patelloida pygmaea (Nakai et al., 2006) P. caerulea or Cymbula nigra (Rivera–Ingraham et al., 2011b). For P. ferruginea, previous studies indicate that settlement may additionally be mediated by chemical cues that are presumably produced by adult conspecifics (Rivera–Ingraham et al., 2011b), which could attract larvae and produce the natural aggregation pattern that has been observed in the species (Espinosa et al., 2006a; Rivera–Ingraham, 2010). Surprisingly, the substrate’s topography also influences the average adult shell size, and those subpopulations located on high THI surfaces had considerably lower average shell sizes than medium and low roughness substrates. P. ferruginea is a homing limpet (Espinosa et al., 2008a), and it has already been observed that when individuals cannot continue growing in their home scar because of the presence of unavoidable irregularities, they can move to smoother areas (Espinosa et al., 2008a). However, the impossibility of individuals to move to smoother surfaces may prevent the normal growth of the individuals, which has already been observed in other areas such as Tarifa Island (obs. pers.), explaining our results. On the other hand, smoother substrates (some natural substrates like in 'San Amaro') presented old and low density subpopulations, with the lowest recruitment rates recorded, and where individuals of more than 50 mm of shell length represented high percentages of the subpopulations. The low density values could be the result of low recruitment rates, although individuals could reach large sizes as the substrate’s surface lacks relevant irregularities. These patterns have already been observed in other invertebrates (Taniguchi & Tokeshi, 2004) and giant limpet species such as L. gigantea (Kido & Murray, 2003). Moreover, these conclusions are additionally supported by the case of 'Dique de Poniente': while the area composed of 3 x 3 m concrete cubes (smooth surfaces) is clearly dominated by other patellid limpets such as Cymbula nigra (Rivera–Ingraham et al., 2011a), and presents very low densities of P. ferruginea individuals (0.77 ind./m), the limestone area shows an average density of 4.88 ind./m. Moreover, results indicate that the area’s accessibility to humans also plays an important role in determining population density and size structures (fig. 3). The maximum densities were recorded on the artificial breakwaters inside the port (Parque del Mediterráneo and within the Guardia Civil’s military base) and at the locally known as 'Piedra del Brazo' islet. The first two sites are restricted access areas composed of high roughness substrates. The latter showed the highest density value recorded in the present study (an also the highest ever recorded for the species), which was established at 35.14 ind./m. It is a 4 x 7 m islet only some meters away from the shore in sector B, only visible during low tide. These areas were also home to some of the largest individuals found. It has been frequently suggested that human collection of organisms reduces population density and preferentially affects the largest fraction of populations. This

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Size class

4,00

< 25 mm

Average density (ind./m)

25–50 mm a

3,00

2,00

1,00

0,00

> 50 mm

c High

Medium Low Substrate roughness

Fig. 6. Average density values recorded for populations located on substrates of high (THI > 1.30), medium (THI = 1.30–1.15) and low (THI < 1.15) roughness (measured using equation provided by Blanchard & Bourget, 1999). Results have been divided into three size classes (< 25 mm, 25–50 mm, > 50 mm). Those values associated with the same letter (a, b, c) and colour belong to the same roughness subset based on a one–way ANOVA and a a posteriori multiple comparison test Student–Neuman–Keuls. Fig. 6. Valores de densidad medios registrados para las poblaciones localizadas sobre sustratos de rugosidad alta (THI > 1,30), media (THI = 1,30–1,15) y baja (THI < 1,15) (medida usando la ecuación descrita por Blanchard & Bourget, 1999). Los resultados han sido divididos en tres clases de tamaños (individuos < 25 mm, 25–50 mm y > 50 mm). Aquellos valores asociados con la misma letra (a, b, c) y el mismo color pertenecen al mismo subgrupo de rugosidad en base a los resultados de un ANOVA de un factor y un test a posteriori de comparación múltiple Student–Newman–Keuls.

is true for giant limpets such as Lottia gigantea (Kido & Murray, 2003), and has already been proven for P. ferruginea (Espinosa et al., 2009; Rivera–Ingraham, 2010). Our observations support these results. On the other hand, the largest individuals were again located inside the port, in the Guardia Civil base and 'Parque del Mediterráneo'. The areas’ protection status and the fact that people are usually reluctant to collect and consume organisms from port areas (presumably living in more polluted waters) (Doneddu & Manunza, 1992) would determine a low collection rate and favour the survivorship of large individuals (Espinosa et al., 2009). The subpopulation in Piedra del Brazo can be considered as very mature, in the sense that 49.6% of the individuals recorded exceeded 50 millimetres in maximum shell length, while recruits only constituted 16% of the total subpopulation. The area was surveyed in July 2009, when most of the recruits resulting from the previous reproduction process should be evident (see Rivera–Ingraham, 2010). Additionally, it was quite frequent to find recruits established on the

adult shells (foresy), clearly at a considerably higher frequency than the rest of the areas surveyed. It is interesting to comment that after a later inspection in 2010, no growth in the number of small individuals was detected. This suggests that the area has reached its carrying capacity, and that the juvenile fraction of the subpopulation recorded the previous year did not survive, probably because of the lack of available space and the competition for food, which are the limiting factors in intertidal habitats (Branch, 1975a). The distribution pattern observed in figure 3 also deserves further examination. Individuals were mainly distributed on the sides of the jetties facing the moat's exit. This could be explained by considering the direction of water currents (represented by the large arrows in figure 3) and assuming that the populations inside the port are the main larvae suppliers. Hydrodynamic currents can determine the local distribution and genetic patterns of intertidal organisms (Baus et al., 2005; Casu et al., 2006). As commented above, 20% of the subpopulation located inside the port exceeds 50 mm

in shell length, partly thanks to the presence of two inaccessible areas (Parque del Mediterráneo and the Guardia Civil base). As the species delays the timing at which sex change occurs when the density of individuals larger than 50 mm is high (Rivera–Ingraham et al., in press), the port holds some of the largest males and females. It is also known that for P. ferruginea fecundity increases with shell size (Espinosa et al., 2006b). In consequence, these 'protected' subpopulations located inside the port may be producing an important number of larvae, which would be exported by tidal currents through the North and South exits of the port. This may be yet another example of the importance of protecting key populations (like the abovementioned) and creating marine reserves for the repopulation of nearby areas, as has been described extensively for many species (e.g. Hastings & Botsford, 1999; McClanahan & Mangi, 2000). Furthermore, a large concentration of individuals was detected in all the jetty endings, which could be explained if we consider that the marine currents responsible for larvae distribution, at a microscale level, run parallel to the shore. In this case, the jetty endings could act as larvae traps, and such a heterogeneous coastline may also determine a heterogeneous distribution of individuals. Finally, it is interesting to comment on the two recently created artificial structures: Fuentecaballos (sector D) and the concrete cube area of Dique de Poniente (Sector M). For the former, a density of 6.46 ind./m was recorded, and the maximum shell length was established at 9.3 cm. Taking into account that this jetty was finished in April 2005 (J. L. Ruiz, pers. com.), and that the census was carried out in March 2010, we could estimate that individuals may present an average growth rate of 1.86 cm/year. On the other hand, the cube area in Sector M, which presents one of the most important C. nigra subpopulations in Ceuta (Rivera–Ingraham, 2010), presented P. ferruginea individuals with a maximum shell length of 9.7 cm (February 2010). Its construction ended in the early months of 2004 (J. Medina, pers. com.), so we can estimate a growth rate of 1.62 cm/year. It has been found that growth rates in the species highly depend on the age of the individuals, in the sense that smaller/younger individuals show higher growth rates that larger/ older individuals (Espinosa et al., 2008b; Rivera–Ingraham, 2010). Previous studies indicate that individuals located in the North Bay of Ceuta present growth rates of 1.15 and 0.73 cm/year, for initial shell lengths of 2 and 8 cm, respectively (Rivera– Ingraham, 2010), while for individuals in the South Bay, values reach 1.64 and 0.38 cm/year. Our calculations establish that individuals located both in Sector D and the concrete–cube section of Sector M should have developed growth rates significantly higher than the aforementioned values. These differences could be due to the fact that the recorded individuals were probably the first to colonize these newly constructed structures (with abundant microalgae biofilm and almost no fauna). If this was the case, newly settled individuals would have abundant trophic resources at their disposal and these could result in higher growth rates than those for individuals that

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settle in areas with well developed subpopulations (where individuals would have to compete for space and food) (Branch, 1975b). Furthermore, a shore may contain individuals in very different conditions of wave exposure or emersion. It may be inappropriate to average the population dynamics of such assemblages at the shore scale when demographic rates are likely to vary greatly within shores (Johnson, 2006). Implications for conservation In the present study, two main parameters seem to highly determine distribution and population structure in P. ferruginea: substrate roughness and the area’s accessibility by humans. The creation of coastal infrastructures has grown in the last decades as a response to the increasing density of humans living by the sea (see Hinrichsen, 1999) to meet port needs and security requirements. The construction of these types of structures can have an important impact on communities settled in nearby areas. Additionally, invader organisms can easily colonize recently created structures, and affect natural/endemic species (see review by Bulleri & Chapman, 2010). Notwithstanding, the results of this study indicate that medium to high roughness substrates (as some materials that are frequently used in the creation of breakwaters and jetties) can favour settlement in P. ferruginea. The population structure recorded on newly created structures (such as Fuentecaballos) suggests that artificial moles and jetties created with this type of substrate can constitute a good habitat that may be collecting larvae from nearby areas. It is important to note that these results and evidence cannot be considered as an excuse to promote the creation of civil engineering constructions (which frequently produce fatal effects on species and communities). However, if these type of structures need to be developed we suggest using dolomitic rocks with medium roughness surfaces (1.15–1.30 roughness coefficients) instead of the concrete cubes, which are starting to be more frequently used. This would allow good recruitment rates and the obtainment of large individuals. Additionally, it has been found that these port areas are avoided by fishermen (Doneddu & Manunza, 1992), as organisms in ports are usually associated with polluted waters. Moreover, these breakwaters are usually difficult to access and would be easy to patrol and guard, which could boost the conservation of the species with the collaboration of port authorities with very little cost (García–Gómez et al., 2011). It has been shown that the inaccessibility and guarding of certain breakwaters is a good way to enhance the obtainment of larger individuals (e.g. Parque del Mediterráneo, Guardia Civil base, etc.), which additionally contribute in large measure to the reproduction event (Espinosa et al., 2006b). Acknowledgements The authors would like to express their gratitude to Natalia Márquez Martínez and Jorge Francisco Marín

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Lora for their help in the sampling process. Thanks also go to the Consejería de Medio Ambiente de Ceuta (OBIMASA) staff for their support and to Dr. José Templado for his comments on the original manuscript. The present study was financed by a F. P. U. grant from the Spanish Ministry of Science and Innovation awarded to G. A. Rivera–Ingraham (AP–2006–04220). References Aversano, F. R., 1986. Esperimento di insediamento artificiale di Patella ferruginea Gmelin, 1791 nelle acque del Golfo di Arzachena (Sardegna settentrionale). Bollettino Malacologico, 22: 169–170. Badii, M. H. & Abreu, J. L., 2006. Metapoblación, conservación de recursos y sustentabilidad. Daena: International Journal of Good Science, 1: 37–51. Baus, E., Darrock, D. J. & Bruford, M. W., 2005. Gene–flow patterns in Atlantic and Mediterranean populations of the Lusitanian sea star Asterina gibbosa. Molecular Ecology, 14: 3373–3382. Bazairi, H., Salvati, E., Benhissoune, S., Tunesi, L., Rais, C., Agnesi, S., Benhamza, A., Franzosini, C., Limam, A., Mo, G., Molinari, A., Nachite, D. & Sadki, I., 2004. Considerations on a population of the endangered marine mollusc Patella ferruginea Gmelin, 1791 (Gastropoda, Patellidae) in the Cala Iris islet (National Park of Al Hoceima–Morocco, Alboran Sea). Bollettino Malacologico, 40: 95–100. Beck, M. W., 2000. Separating the elements of habitat structure: independent effects of habitat complexity and structural components on rocky intertidal gastropods. Journal of the Experimental Marine Biology and Ecology, 249: 29–49. Biagi, V. & Poli, D., 1986. Considerazioni su una popolazione di Patella ferruginea Gmelin, 1971 per le acque del promontorio di Piombino. Bollettino Malacologico, 22: 171–174. Blanchard, D. & Bourget, E., 1999. Scales in coastal heterogeneity: influence on intertidal community structure. Marine Ecology Progress Series, 179: 163–173. Boudouresque, C. F. & Laborel–Deguen, F., 1986. Patella ferruginea. In: Le benthos marin de l’ile de Zembra (Parc National, Tunisie): 105–110 (C. F. Boudouresque, J. G. Harmelin & A. Jeudy de Grissac, Eds.). GIS Posidonie Publishers, Marseille. Boumaza, S. & Semroud, S., 2001. Inventaire de la population de Patella ferruginea Gmelin, 1791 des Iles Habibas (ouest Algerien). Rapport du Congres de la Commission Internationale pour l’Exploration Scientifique de la Mer Mediterranée, 36: 361. Branch, G. M., 1975a. Notes on the ecology of Patella concolor and Cellana capensis, and the effects of human consumption on limpet populations. Zoologie Africaine, 10: 75–85. – 1975b. Intraspecific competition in Patella cochlear Born. Journal of Animal Ecology, 44: 263–282. Bray, J. R. & Curtis, J. T., 1957. An ordination of the upload forest communities of southern Wisconsin. Ecological Monographs, 27: 325–349. Bulleri, F. & Chapman, M. G., 2010. The introduction

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on the limpet Lottia gigantea: a field study in Baja California (USA). Pacific Science, 50: 393–403. Porcheddu, A. & Milella, I., 1991. Aperçu sur l’ecologie et sur la distribution de Patella ferruginea (L.) Gmelin, 1791 en mers italiennes. In: Les Espèces marines à protéger en Méditerranée: 119–128 (C. F. Boudouresque, M. Avon & V. Gravez, Eds.). GIS Posidonie Publishers, Marseille. Raffaelli, D. & Hawkins, S., 1996. Intertidal Ecology. Chapman and Hall, London. Ramos, M. A., 1998. Implementing the habitats directive for mollusc species in Spain. Journal of Conchology, Special Publication, 2: 125–132. Rivera–Ingraham, G. A., 2010. Biología de la conservación de especies de patélidos en el umbral Atlántico–Mediterráneo. Ph. D. Thesis, Univ. of Seville. Rivera–Ingraham, G. A., Espinosa, F. & García–Gómez, J. C., 2011a. Ecological considerations and niche differentiation between juvenile and adult black limpets (Cymbula nigra). Journal of the Marine Biological Association of the U.K., 91: 191–198. – 2011b. Effect of γ–amino butyric acid on limpet populations: towards the future management and conservation of endangered patellid species. Journal of Chemical Ecology, 37: 1–9. – In press. Environmentally–mediated sex change in the endangered limpet Patella ferruginea (Gastropoda: Patellidae). Journal of Molluscan Studies. Rochet, M. J. & Trenkel, V. M., 2003. Which community indicators can measure the impact of fishing? A review and proposals. Canadian Journal of Fisheries and Aquatic Sciences, 60: 86–99. Sagarin, R. D., Ambrose, R. F., Becker, B. J., Engle, J. M., Kido, J., Lee, S. F., Miner, C. F., Murray, S. N., Raimondi, P. T., Richards, D. V. & Roe, C., 2007.

Ecological impacts on the limpet Lottia gigantea populations: human pressure over a broad scale on island and mainland intertidal zones. Marine Biology, 150. Taniguchi, H. & Tokeshi, M., 2004. Effects of habitat complexity on benthic assemblages in a variable environment. Freshwater Biology, 49: 1164–1178. Templado, J., 2001. Patella ferruginea Gmelin, 1791. In: Los invertebrados no insectos de la Directiva Hábitats en España: 41–49 (M. A. Ramos, D. Bragado & J. Fernádez, Eds.). Organismo Autónomo de Parques Nacionales. Dirección General de Conservación de la Naturaleza, Ministerio de Medio Ambiente, Madrid, Spain. Templado, J., Calvo, M., Moreno, D., Flores, A., Conde, F., Abad, R., Rubio, J., López–Fe, C. M. & Ortiz, M., 2006. Flora y fauna de la Reserva Marina de Pesca de la Isla de Alborán. Secretaría General de Pesca Martítima, MAPA, Madrid, Spain. Templado, J. & Moreno, D., 1997. La lapa ferrugínea. Biológica, 6: 80–81. Underwood, A. J., 2004. Landing on one’s foot: small–scale topographic features of habitat and the dispersion of juvenile intertidal gastropods. Marine Ecology Progress Series, 268: 173–182. Williams, G. A. & Morritt, D., 1995. Habitat partitioning and thermal tolerance in a tropical limpet, Cellana grata. Marine Ecology Progress Series, 124: 89–103. Wilson, F. S., 1990. Temporal and spatial patterns of settlement: a field study of molluscs in Bogue Sound, North Carolina. Journal of Experimental Marine Biology and Ecology, 139: 201–220. Woodin, M. C., 1986. Settlement of infauna: larval choice? Bulletin of Marine Science, 39: 401–407.

Editor executiu / Editor ejecutivo / Executive Editor Joan Carles Senar

Secretaria de Redacció / Secretaría de Redacción / Editorial Office

Secretària de Redacció / Secretaria de Redacción / Managing Editor Montserrat Ferrer

Museu de Zoologia Passeig Picasso s/n 08003 Barcelona, Spain Tel. +34–93–3196912 Fax +34–93–3104999 E–mail mzbpubli@intercom.es

Consell Assessor / Consejo asesor / Advisory Board Oleguer Escolà Eulàlia Garcia Anna Omedes Josep Piqué Francesc Uribe

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Soft–bottom sipunculans from San Pedro del Pinatar (Western Mediterranean): influence of anthropogenic impacts and sediment characteristics on their distribution L. M. Ferrero–Vicente, Á. Loya–Fernández, C. Marco–Méndez, E. Martínez–García & J. L. Sánchez–Lizaso Ferrero–Vicente, L. M., Loya–Fernández, Á., Marco–Méndez, C., Martínez–García, E. & Sánchez–Lizaso, J. L., 2011. Soft–bottom sipunculans from San Pedro del Pinatar (Western Mediterranean): influence of anthropogenic impacts and sediment characteristics on their distribution. Animal Biodiversity and Conservation, 34.1: 101–111. Abstract Soft–bottom sipunculans from San Pedro del Pinatar (Western Mediterranean): influence of anthropogenic impacts and sediment characteristics on their distribution.— We analysed the distribution of soft bottom sipunculans from San Pedro del Pinatar (Western Mediterranean). This study was carried out from December 2005 to June 2010, sampling with biannual periodicity (June and December). Physical and chemical parameters of the sediment were analysed (granulometry, organic matter content, pH, bottom salinity and shelter availability). Nine different species and subspecies were identified, belonging to five families. Aspidosiphon muelleri muelleri was the dominant species, accumulating 89.06% of the total abundance of sipunculans. Higher sipunculan abundances were correlated with stations of higher percentage of coarse sand, empty mollusc shells and empty tubes of the serpulid polychaete Ditrupa arietina, where some of the recorded species live. Sediment characteristics played the main role controlling the sipunculans distribution. Anthropogenic impacts could be indirectly affecting their distribution, changing the sediment characteristics. Key words: Sipuncula, Aspidosiphon muelleri, Mediterranean, Anthropogenic impact, Soft–bottom. Resumen Sipuncúlidos de fondos blandos de San Pedro del Pinatar (Mediterráneo occidental): influencia de los impactos antropogénicos y las características del sedimento sobre su distribución.— Se analizó la distribución de los sipuncúlidos de fondos blandos de San Pedro del Pinatar (Mediterráneo occidental). Este estudio se llevó a cabo entre diciembre de 2005 y junio de 2010, muestreando con periodicidad semestral (junio y diciembre). Se analizaron parámetros físicos y químicos del sedimento (granulometría, contenido de materia orgánica, pH, salinidad de fondo y disponibilidad de refugio). Nueve especies y subespecies diferentes fueron identificadas, pertenecientes a cinco familias. Aspidosiphon muelleri muelleri fue la especie dominante, acumulando el 89,06% de la abundancia total de sipuncúlidos. Las mayores abundancias de sipuncúlidos se correlacionaron con las estaciones con mayores porcentajes de arena gruesa, conchas de moluscos vacías y tubos vacíos del poliqueto serpúlido Ditrupa arietina, donde viven algunas de las especies registradas. Las características del sedimento jugaron el papel principal en el control de la distribución de sipuncúlidos. Los impactos antropogénicos podrían estar afectando indirectamente su distribución, cambiando las características del sedimento. Palabras clave: Sipuncúlidos, Aspidosiphon muelleri, Mediterráneo, Impacto antropogénico, Fondos blandos. Luis Miguel Ferrero–Vicente, Ángel Loya–Fernández, Candela Marco–Méndez, Elena Martínez–García & José Luis Sánchez–Lizaso, Dept of Marine Sciences and Applied Biology, Univ. of Alicante, P. O. Box 99, 03080 Alicante, España (Spain). Corresponding author: L. M. Ferrero–Vicente. E–mail: lmferrero@ua.es

ISSN: 1578–665X

102

Introduction The phylum Sipuncula is composed of about 150 species and subspecies (Cutler, 1994). Sipunculans are exclusively marine benthic organisms. Almost 75% of the sipunculan species are concentrated in shallow waters (< 200 m), and half of them are confined to the upper photic zone of the shelf (Murina, 1984). The phylum Sipuncula has been overlooked and was barely studied for many years. In the Mediterranean Sea, its distribution has been studied widely since the 1990s (Saiz–Salinas, 1993; Murina et al., 1999; Pancucci–Papadopoulou et al., 1999; Açik et al., 2005; Açik, 2007, 2008a, 2008b, 2009). Most of these studies, however, deal with the Eastern Mediterranean and only Saiz–Salinas gives concise information about sipunculans in the Western Mediterranean (Saiz–Salinas, 1982, 1986, 1993; Saiz–Salinas & Murina, 1982; Saiz–Salinas & Villafranca–Urchegui, 1990). Most sipunculan worms are deposit feeders and many of them live in soft substrata. Soft–bottom sipunculans live buried inside the sediment and they obtain a substantial part of their food through the ingestion of sediment. Some species of sipunculans have been described as important bioturbators in soft sediments (Murina, 1984; Kędra & Wlodarska–Kowalczuk, 2008; Shields & Kędra, 2009), playing a main role in benthic ecosystems. It is well known that some of these species often find shelter inside the empty shells of certain molluscs or empty tubes of polychaetes (Gibbs, 1985; Troncoso & Urgorri, 1992; Saiz–Salinas, 1993; Murina et al., 1999; Troncoso et al., 2000; Açik et al., 2005; Schulze, 2005; Wanninger et al., 2005), usually empty tubes of the serpulid Ditrupa arietina (Solís–Weiss, 1982; Morton & Salvador, 2009) and other cavities inside hard structures buried into the sediment. Those facts promote a strong relation between sipunculans and characteristics of the sediment such as granulometry, pH, organic matter content and shelter availability. The aim of this study was to analyse the distribution of sipunculan worms in this area and assess the possible effect of abiotic factors of the sediment or anthropogenic impacts on their distribution. Material and methods Study area and sampling design The area studied is located near San Pedro del Pinatar coast (SE Spain). Different types of anthropogenic disturbances merge in this area: a sewage outfall, a brine discharge and fish farm cages (fig. 1). We compared a grid of 12 stations covering the influence area of the three anthropogenic impacts. We sampled twice a year (December and June) over a five–year period, from December 2005 to June 2010. We established three transects (A, B and C), separated by approximately 2 km. Four stations at each transect were established (1, 2, 3 and 4), separated from 250 to 500 m according to depth. The distance between C3 and C4 was shorter due to sampling problems. The initial sampling point for C3 was not soft–bottom and was relocated.

Ferrero–Vicente et al.

Four samples of each station were taken using a Van Veen grab and covering a surface area of 0.04 m2 for each sample. The sewage outfall has been in place for decades and produces a flow of 5,000 m3 per day in winter and 20,000 m3 per day in summer, with wastewater secondary treatment (Del–Pilar–Ruso et al., 2009). The sewage discharge point is located between stations A1 and B1. The desalination plant began operations in January 2006 with a discharge of 80,000 m3 per day, but in October 2006 the production was doubled with the start up of a new plant. The discharge takes place through a shared outfall at 33 m depth over soft sediment, and it is located at the B2 station. The brine presents a high salinity, ranging between 60–68 psu. In March 2010, a diffuser was connected to the end of the pipeline, and this has notably reduced the concentration of brine 10 m away from the discharge point (from 68 to 40 psu, unpublished data). A complete description about how the brine plume flows along the sea bottom can be found in Fernández–Torquemada et al. (2009). The fish farm complex has operated since 1998, with an annual production of 6,197 tons of blue fin tuna (Thunnus thynnus), sea bream (Spaurus aurata), meagre (Argyrosomus regius) and sea bass (Dicentrarchus labrax) (Ruiz et al., 2010). Some of the fish farm cages are less than 200 m away from stations C3 and C4. Laboratory analysis Three samples were used for the faunistic analysis, and in June 2010 one of these samples was also used to count the total shelters available in each station. The samples used for the faunistic analysis were sieved through a 0.5 mm mesh screen to separate the macrofauna and they were fixed in 10% buffered formalin. Later the fauna was preserved in ethanol 70% and sorted in different taxa. Sipunculans were identified to the species level through analysis of their internal and external anatomy using a binocular scope and observing typical structures with taxonomic value, such as presence and shape of hooks and papillaes, using a microscope. Length and thickness of the trunk of each specimen were taken. The fourth sample was used for the sediment characterization (granulometric analysis, pH and amount of organic matter). We measured the pH of the sediment immediately after collection, using a pH–meter Crisom with a sensor 52–00. Furthermore, a sub–sample of approximately 30 g of the shallow layer of the sediment was separated to subsequently obtain the percentage of organic matter. We determined the organic matter content from the sub–samples as weight loss on ignition after 4 hr at 400ºC inside the muffle furnace. We carried out the granulometric analysis following the methodology described by Buchanan (1984), sorting the samples into six categories: gravel, coarse sand, medium sand, fine sand and silt and clays. Depth and bottom salinity were recorded using a CTD sensor (RBR–XR–420/620). We calculated the shelter availability from one of the samples used for the fauna analysis in June 2010.

Animal Biodiversity and Conservation 34.1 (2011)

698000

103

700000

702000

4192000

A1 A2

4192000

4190000

B1 B2B3

4190000

BD FF 4188000

C1 C2

C3 C4

4188000

4186000

N 698000 1,860

700000 830

702000 1,860 m

Fig. 1. Map of the study area showing the sampling stations: SO. Sewage outfall; BD. Brine discharge; FF. Fish farm cages. (Adapted from Del–Pilar–Ruso et al., 2009). Fig. 1. Mapa del área de estudio mostrando las estaciones de muestreo: SO. Emisario de aguas residuales: BD. Vertido de salmuera; FF. Jaulas de piscifactoría. (Adaptado de Del–Pilar–Ruso et al., 2009).

We took three sub–samples and we counted the shelters in each sub–sample to calculate average of shelters/m2. Available shelters, able to be used by sipunculids according to observations since 2005, were sorted into three categories according to their shape: spiral shape shells (gastropod shells such as Turritella sp.), cylindrical shape (scaphopod shells and serpulid tubes), and undefined shape (pieces of shell or calcareous debris among others). We also sorted the shelters into three size classes: S1, S2 and S3, corresponding to the length ranges < 3.5 mm, 3.5–7.0 mm and > 7.0 mm respectively. Statistical analysis Univariate and multivariate techniques were used to detect possible changes in the distribution of sipunculan species, and to define their possible relation both with the abiotic factors of the sediment and

with the different anthropogenic impacts. Pearson product–moment correlation coefficient (r) was used to detect a possible linear correlation between abundance of sipunculans, species richness and abiotic factors of the sediment (granulometry, pH, organic matter content, and salinity). Pearson coefficient was also used to detect a possible linear dependence relation between shelter availability and sipunculans abundance. Multivariate analysis is considered a sensitive tool for detecting changes in the structure of marine faunal community (Clarke & Warwick, 1994). Multivariate analysis of data was carried out using the PRIMER statistical package. We used the square root of abundance of sipunculans, and Bray–Curtis similarity coefficient was chosen to calculate the triangular similarity matrix. From this similarity matrix, non–metric multidimensional scaling techniques (nMDS) were applied.

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Table 1. Total abundance for each species and for each station: N. Total individuals. Tabla 1. Abundancia total para cada especie y para cada estación: N. Total de individuos.

Stations

A1 A2 A3 A4 B1 B2 B3 B4 Aspidosiphon muelleri muelleri 9 7 9 35 9 1 8 42 Phascolion cf. caupo 0 0 0 2 0 0 3 1 Thysanocardia procera 0 1 2 4 4 0 1 4 Onchnesoma steenstruppi steenstruppi 0 0 1 0 0 0 0 0 Aspidosiphon muelleri kowalevski 1 0 2 0 0 0 0 0 Phascolion strombus 0 0 0 0 0 0 0 1 Phascolosoma granulatum 0 0 0 0 0 0 0 0 Sipunculus nudus 0 0 0 0 0 0 1 0 Golfingia vulgaris vulgaris 0 0 1 0 0 0 0 0 Unidentified 0 0 2 1 0 0 1 2 N 10 8 17 42 13 1 14 50 % 1.26 1.00 2.14 5.28 1.64 0.13 1.76 6.29

385

138

708 89.06

39 4.91

22 2.77

5 0.63

3 0.38

2 0.25

1 0.13

3 0 0 2 10 394 164 72 1.26 49.56 20.63 9.06

11 1.38 795 100 100 ─

Table 2. N. Number of individuals; TL. Average trunk length (mean ± SE); MaxTL. Maximum trunk length; MinTL. Minimum trunk length; TW. Trunk width; Shelter. % individuals inhabiting some kind of shelter (empty shells and empty tubes, among others). Tabla 2. N. Número de individuos; TL. Longitud promedio del tronco (media ± EE); MaxTL. Longitud máxima del tronco; MinTL. Longitud mínima del tronco; TW. Anchura del tronco; Shelter. % de individuos viviendo en algún tipo de refugio (conchas y tubos vacíos entre otros).

TL (mm)

MaxTL MinTL (mm) (mm)

Aspidosiphon muelleri muelleri

708

4.95 ± 0.01

22.0

0.5

Phascolion cf. caupo

3.29 ± 0.44

12.0

Thysanocardia procera

3.87 ± 0.57

11.0

Onchnesoma steenstruppi steenstruppi 5

1.50 ± 0.22

Phascolion strombus strombus

1.50 ± 0.29

Aspidosiphon muelleri kovalevskii

Phascolosoma granulatum

Sipunculus nudus Golfingia vulgaris vulgaris

TW (mm) Shelter 0.63 ± 0.00

94.17

0.5

0.74 ± 0.11

81.82

1.0

0.87 ± 0.13

0.00

2.0

1.0

0.60 ± 0.10

0.00

2.0

1.0

0.10 ± 0.00

66.67

4.83 ± 2,12

9.0

2.0

0.83 ± 0.17

33.33

3.00 ± 1.41

4.0

2.0

0.75 ± 0.35

0.00

13.00

─

3.00

0.00

1.50

─

0.50

0.00

Animal Biodiversity and Conservation 34.1 (2011)

100%

105

80% 60% 40% 20% 0%

D J D J D J D J D J 05 06 06 07 07 08 08 09 09 10

100%

D J D J D J D J D J 05 06 06 07 07 08 08 09 09 10

80% 60% 40% 20% 0%

D J D J D J D J D J 05 06 06 07 07 08 08 09 09 10

100%

D J D J D J D J D J 05 06 06 07 07 08 08 09 09 10

80% 60% 40% 20% 0%

D J D J D J D J D J 05 06 06 07 07 08 08 09 09 10

100%

D J D J D J D J D J 05 06 06 07 07 08 08 09 09 10

80% 60% 40% 20% 0%

D J D J D J D J D J 05 06 06 07 07 08 08 09 09 10

Gravel

D J D J D J D J D J 05 06 06 07 07 08 08 09 09 10

Coarse sand

Medium sand

D J D J D J D J D J 05 06 06 07 07 08 08 09 09 10

Fine sand

Silt and clays

Fig. 2. Granulometric characteristics at each station from 2005 to 2010: J. June; D. December. Fig. 2. Caracteristicas granulométricas para cada estación desde 2005 hasta 2010: J. Junio; D. Diciembre.

Results A total of 360 benthic samples, from December 2005 to June 2010, were collected for the fauna analysis. We analysed 795 specimens, finding 9 different species (table 1), belonging to the families: Sipunculidae (1 sp.), Golfingiidae (2 spp.), Phascolionidae (3 spp.),

Phascolosomatidae (1 sp.) and Aspidosiphonidae (2 spp.). The dominant species was Aspidosiphon (Aspidosiphon) muelleri muelleri (Diesing, 1851), with 708 individuals (89.06% of the total abundance). Eleven specimens (1.38%) could not be identified due to their small size or their state of deterioration. According to the different stations, higher abundances

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Ferrero窶天icente et al.

Table 3. Depth (m), organic matter (OM, %), pH and bottom salinity (psu) at the different stations, and dates sampled: S. Station and depth; Y. Year; M. Month; BS. Bottom salinity. S

2005

33.5 m

2006

Dec

9.10

6.98

Jun

13.13

7.51

37.5

2005

Dec

38.0

32.8 m

2006

Jun

2006

Dec

3.19

7.69

37.5

2006

Dec

2007

Jun

4.58

7.46

38.0

2007

Jun

2007

Dec

2008

Jun

2.13

7.38

37.2

2007

Dec

5.11

7.75

37.6

2008

Jun

2008

Dec

6.48

7.79

37.6

2008

Dec

2009

Jun

5.2

7.38

37.4

2009

Jun

2009

Dec

5.00

7.62

37.8

2009

Dec

2010

Jun

7.59

7.13

37.9

2010

Jun

2005

Dec

12.61

6.80

37.5

2005

Dec

34.1 m

2006

Jun

7.35

7.54

37.5

33.6 m

2006

Jun

2006

Dec

3.52

7.71

37.5

2006

Dec

2007

Jun

13.23

7.39

37.0

2007

Jun

2007

Dec

3.52

7.59

37.2

2007

Dec

2008

Jun

10.58

7.66

37.2

2008

Jun

2008

Dec

5.46

7.66

37.4

2008

Dec

2009

Jun

5.92

7.22

37.3

2009

Jun

2009

Dec

4.26

7.46

37.8

2009

Dec

2010

Jun

8.20

7.14

37.8

2010

Jun

2005

Dec

6.47

7.04

37.5

2005

Dec

35.6 m

2006

Jun

15.85

7.04

37.5

33.7 m

2006

Jun

2006

Dec

3.62

7.82

38.0

2006

Dec

2007

Jun

13.27

7.48

37.0

2007

Jun

2007

Dec

3.48

7.48

36.8

2007

Dec

2008

Jun

10.04

7.73

37.5

2008

Jun

2008

Dec

4.41

7.89

37.4

2008

Dec

2009

Jun

4.56

7.43

37.4

2009

Jun

2009

Dec

4.23

7.44

37.6

2009

Dec

2010

Jun

5.05

7.19

37.7

2010

Jun

2005

Dec

2.55

7.34

37.5

2005

Dec

36.8 m

2006

Jun

6.97

7.70

37.5

37.2 m

2006

Jun

2006

Dec

1.72

7.85

37.5

2006

Dec

2007

Jun

1.41

7.80

37.0

2007

Jun

2007

Dec

3.29

7.84

36.8

2007

Dec

2008

Jun

0.98

7.77

37.8

2008

Jun

2008

Dec

2.26

7.71

37.5

2008

Dec

2009

Jun

0.95

7.83

37.4

2009

Jun

2009

Dec

3.09

7.71

38.7

2009

Dec

2010

Jun

4.82

7.28

38.1

2010

Jun

Animal Biodiversity and Conservation 34.1 (2011)

107

Tabla 3. Profundidad (m), materia orgánica (OM, %), pH y salinidad del fondo (ups) para las diferentes estaciones y fechas muestreadas: S. Estación y profundidad; Y. Año; M. Mes; BS. Salinidad del fondo. OM

18.43

7.29

37.5

11.06

7.29

38.0

2.98

7.88

8.95

2005

Dec

4.94

7.5

38.0

32.6 m

2006

Jun

4.93

6.8

38.0

37.5

2006

Dec

1.28

7.9

38.0

7.43

37.5

2007

Jun

3.37

7.7

37.0

2.27

7.76

37.2

2007

Dec

1.92

7.8

37.6

7.80

38.1

2008

Jun

1.29

7.8

37.9

4.39

7.65

38.3

2008

Dec

2.18

7.9

37.7

2.16

7.27

39.4

2009

Jun

1.99

7.4

37.4

3.23

7.39

41.6

2009

Dec

1.69

7.6

37.7

8.38

7.19

37.7

2010

Jun

2.30

7.5

37.5

9.15

7.47

37.5

2005

Dec

4.81

7.3

38.0

3.82

7.47

42.0

33.4 m

2006

Jun

9.67

7.4

37.5

1.85

7.35

39.0

2006

Dec

1.77

7.7

38.0

4.54

7.20

45.0

2007

Jun

1.27

7.8

37.0

1.55

7.80

46.0

2007

Dec

1.79

7.6

37.6

3.60

7.83

52.9

2008

Jun

1.84

7.7

37.7

1.92

7.70

44.4

2008

Dec

1.92

7.8

37.7

4.99

7.55

45.1

2009

Jun

1.01

7.7

37.4

2.46

7.42

48.1

2009

Dec

1.54

7.7

37.7

4.68

7.19

38.4

2010

Jun

2.86

7.6

37.6

3.91

7.12

37.5

2005

Dec

1.89

7.4

38.0

14.50

7.63

50.0

34.9 m

2006

Jun

9.05

7.5

38.0

1.62

7.65

41.5

2006

Dec

1.89

7.8

37.5

2.56

7.52

42.0

2007

Jun

4.66

7.4

37.0

1.99

7.45

41.4

2007

Dec

1.66

7.8

37.6

2.37

7.60

47.5

2008

Jun

4.27

7.7

37.8

1.62

7.80

43.1

2008

Dec

1.62

7.8

37.6

1.55

7.40

41.3

2009

Jun

2.00

7.7

37.5

2.15

7.77

43.4

2009

Dec

2.80

7.7

37.6

4.70

7.23

38.2

2010

Jun

2.80

7.2

37.7

6.61

7.06

37.5

2005

Dec

14.14

7.3

38.0

19.96

7.86

38.5

35.6 m

2006

Jun

6.59

7.6

37.5

2.16

7.74

38.0

2006

Dec

1.87

7.8

37.5

3.48

7.43

37.5

2007

Jun

2.95

7.5

37.0

2.34

7.38

38.8

2007

Dec

1.29

7.7

37.6

4.50

7.71

37.6

2008

Jun

3.41

7.7

37.7

4.69

7.82

37.9

2008

Dec

1.60

7.7

37.6

6.76

7.42

37.5

2009

Jun

2.45

7.6

37.4

3.96

7.74

37.8

2009

Dec

1.94

7.5

37.8

7.79

7.12

38.0

2010

Jun

3.42

7.5

37.8

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Table 4. Pearson correlation coefficient (r) of the most abundant species vs. sediment characteristics: OM. Organic matter (%); D. Depth; S. Salinity; G. Gravel (%); CS. Coarse sand (%); MS. Medium sand (%); FS. Fine sand (%); SC. Silt and clays (%) (* p < 0.001). Tabla 4. Coeficiente de correlación de Pearson (r) de las especies más abundantes en relación a las características del sedimento: OM. Materia orgánica (%); D. Profundidad; S. Salinidad; G. Gravas (%); CS. Arena gruesa (%); MS. Arena media (%); FS. Arena fina (%); SC. Limos y arcillas (%) (* p < 0,001).

Aspidosiphon muelleri muelleri –0.200* 0.192* –0.030 –0,131 –0.120 0.041 0.497*** 0.170 –0.287*** Phascolion cf. caupo

–0.100 0.054 –0.025 –0.057 0.012 0.262** 0.280*** –0.041 –0.245**

Thysanocardia procera

–0.071 0.127 0.214* –0.116 –0.07 –0.039 –0.053 0.027

0.054

Table 5. Pearson correlation coefficient (r) of Aspidosiphon (A.) muelleri muelleri vs. shelters availability: S1 < 0.5 mm; S2 = 0.5–1.0 mm; S3 ≥ 1.00 mm; *** p < 0.001. Tabla 5. Coeficiente de correlación de Pearson (r) de Aspidosiphon (A.) muelleri muelleri en relación a la disponibilidad de refugio: S1 < 0,5 mm; S2 = 0,5–1,0 mm; S3 ≥ 1,00 mm; *** p < 0,001.

Cylindrical shape

A. muelleri muelleri

Spiral shape S1

Undefined shape S3

0.593*** 0.844*** 0.802*** –0.042 –0.308 –0.268 –0.150 0.086

were found in C2, C3 and C4, with 49.56%, 20.63% and 9.06% of the total individuals respectively. On the other hand, lower contributions, with less than 2% of the individuals for each station, were recorded in B3 (1.76%), B1 (1.64%), C1 (1.26%), A1 (1.26%), A2 (1.00%) and B2 (0.13%). Four species inhabited empty mollusc shells, empty tubes of Ditrupa arietina or other less common shelters (calcareous debris or crevices in chunks of rock) (table 2). The remaining species were found bare in the sediment. The granulometry of the sediment was heterogeneous among the different sites studied (fig. 2). We detected high levels of silt and clays in the stations close to the sewage discharge (A1, A2, A3 and B1). Despite the variation over seasons and years, higher values of organic matter content were usually found in these stations. In the southern stations (C1, C2, C3 and C4), the granulometric analysis showed lower values of the finest fraction, and a more equitable distribution of the grain size, with higher values of gravel and sand, and decrease of silt and clays. A similar granulometric pattern was observed in B2, the brine discharge point. The bottom salinity records showed how the station B2 and, to a lesser extent B1 and B3, were affected

S3 ─

by the brine discharge (table 3), reaching maximum salinity values of 52.9 psu (VI 2008; B2). While the desalination plant was working (XII 2006–VI 2010) the salinity average in B1, B2 and B3 was 38.4 ± 0.5 psu (mean ± SE), 44.6 ± 1.5 psu and 43.2 ± 0.9 psu respectively. We analysed the correlation between the three most abundant species and several characteristics of the sediment (table 4). The number of specimens of the remaining species was too low (< 20 individuals) to analyse a possible correlation. A. muelleri muelleri was positively correlated with the percentage of medium sands (r = 0.497, p = 0.000) and pH (r = 0.192, p = 0.018), whereas it was negatively correlated with the percentage of organic matter (r = –0.200, p = 0.014) and the percentage of silt and clays (r = –0.287, p = 0.001). A. muelleri muelleri was also correlated with the availability of shelter, showing a strong positive correlation with the three different sizes corresponding to the cylindrical shape (table 5). Phascolion. cf. caupo showed a positive correlation, with the percentage of coarse sand and the percentage of medium sand (r = 0.262 and r = 0.280 respectively both p < 0.01). T. procera only showed significant correlation with the depth (r = 0.214, p < 0.05).

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Stress: 0.03

I B2

II A1 A2

B1 A3

III C1

B3 C4 A4 B4

Fig. 3. MDS plot using Bray–Curtis similarity indices from square root of average abundances of sipunculan species from 2005 to 2010. Group I. Station matching with the brine discharge point; Group II. Stations closest to the sewage spill; Group III. Stations closest to fish farm cages. Fig. 3. Gráfico MDS usando índices de similaridad de Bray–Curtis a partir de la raíz cuadrada de las abundancias medias de sipuncúlidos desde 2005 hasta 2010. Grupo I. Estación coincidente con el punto de vertido de la salmuera; Grupo II. Estaciones cercanas al emisario de aguas residuales; Grupo III. Estaciones cercanas a las jaulas de piscifactoría.

Stations nearest the different impacts were grouped by the nMDS procedure (fig. 3). Group I consisted only of B2 station, which matched with the brine discharge point. The second group marked (group II) was made up of the closest stations to the sewage outfall (A1, A2, A3 and B1), the muddiest stations, with a higher fraction of silt and clays. Group III was made up of the stations near the fish farms cages (C2, C3 and C4). In addition, these stations also had the highest percentages in coarser fractions of the sediment. Discussion Sipunculans have a worldwide distribution because of their high tolerance to a wide range of temperatures and depths, inhabiting different habitats ranging from shallow waters to the abyssal zone (Cutler, 1965, 1977; Murina, 1984). However, the sensitivity of the phylum to environmental changes has barely been investigated, perhaps because there is a lack of specialists for its identification throughout the world (Pancucci–Papadopoulou et al., 1999) and their abundances are usually low, although on some occasions they can become dominant species (Klaoudatos et al., 2006; Solís–Weiss, 1982). A. muelleri muelleri was the prevalent species at the sites studied. It was found mainly inside empty shells or tubes, and this shelter was rarely shared with another individual or another kind of organism,

such as nematodes, bivalves or polychaetes. This behaviour has been described previously (Gage, 1968; Murina, 1984; Solís–Weiss, 1982). In our case, the shared shelter always had a cylindrical shape (almost always D. arietina tubes). The Pearson correlation coefficient showed some correlation between sediment characteristics and abundance of sipunculan worms. We found a negative correlation between A. muelleri muelleri, P. cf. caupo and a high proportion of the fine fraction of the sediment. Solís–Weiss (1982) described how the abundance of A. muelleri can decrease in muddy sediments. These two species also have a good positive correlation with medium grain size sand. The strongest correlation was established with the availability of shelters with cylindrical shape, particularly with the size classes S2 and S3, and there was no correlation with spiral shells. The reason is probably that spiral shells, usually Turritella sp., were found in abundance in the muddy stations. These merged factors make it difficult to be conclusive establishing a relation between abundances of sipunculans and shelter availability. On the other hand, species that do not usually inhabit empty shells, such as T. procera, did not show any trend toward muddy sediments or toward a particular grain size. T. procera has been recorded in different types of sediment and several different habitats, even parasiting the polychaete Aphrodite aculeata (Stephen & Edmonds, 1972 in Saiz–Salinas, 1993).

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Group II formed in the MDS includes the stations nearest to the sewage outfall. These stations are characterized by their high sedimentation load and their muddy bottom, resulting in a decline in abundance of A. muelleri muelleri. There are limited studies about the response of sipunculans to a salinity increase (Oglesby, 1982), and lack of information about the effect that a sudden fluctuation in salinity, either temporary or constant, could have over the survival of these animals and their behaviour. Some species studied seem to be osmoconformers, with limited power of ion regulation, during an event of salinity change (Adolph, 1936; Oglesby, 1982; Ferraris et al., 1994; Chew et al., 1994). However, studies are limited to only few particular species and do not explain how a salinity change could influence over their behaviour or their long–term survival ability within these altered conditions. Only Phascolosoma arcuatum (Gray, 1828) has been described as an active ion regulator with good response and resistance to sudden fluctuations in salinity (Chew et al., 1994), making it capable of survival in mangroves with freshwater inputs. The nMDS separates the B2 station (the brine discharge point) far away from each other. Although we did not find any sipunculan in the discharge area until 2010, after the diffuser was installed and the salinity concentrations near the discharge point dropped drastically, a direct relation with the brine discharge is difficult to establish. Salinity could be affecting sipunculans assemblage but the absence of individuals at the discharge point before the implementation of the desalination plant and the heterogeneity of the bottoms studied do not allow results to be conclusive. In addition, only one sipunculan appeared in June 2010, and its presence can be considered anecdotic. Additional studies are necessary, including field and laboratory experiments, to accurately determine the effect that salinity fluctuations could have on the sipunculan distribution. Group III established by the nMDS procedure (stations C2, C3 and C4) is characterized by its higher sipunculan abundance. These three stations are the closest to fish cages. A decrease in abundance of A. muelleri from fish farm bottoms has been reported (Klaoudatos et al., 2006) but muddy sites in fish farms are more likely to be identified as impacted than coarse sediment sites (Papageorgiou et al., 2010). Stations C2, C3 and C4 are distinguished by their low percentage of silt and clays and high percentages of medium sand, which was the fraction best correlated with A. muelleri muelleri abundance. This area also presented the highest amount of scaphopod shells and Ditrupa arietina tubes. The main factor explaining the distribution of sipunculans in San Pedro del Pinatar seems to be the abiotic characteristics of the bottom, especially the granulometry of the sediment. It seems that anthropogenic impacts could indirectly play a role by changing the sediment characteristics, as is probable in the case of sewage discharge that is increasing the silt and clay fractions at the closest stations.

Ferrero–Vicente et al.

Acknowledgements This study was funded by 'Mancomunidad Canales del Taibilla'. We thank Cristina Celdrán for technical assistance in the laboratory work. References Açik, S., 2007. Observations on the population characteristics of Apionsoma (Apionsoma) misakianum (Sipuncula: Phascolosomatidae), a new species for the Mediterranean fauna. Scientia Marina, 71(3): 571–577. – 2008a. Occurrence of the alien species Aspidosiphon (Aspidosiphon) elegans (Sipuncula) on the Levantine and Aegean coasts of Turkey. Turkish Journal of Zoology, 32(4): 443–448. – 2008b. Sipunculans along the Aegean coast of Turkey. Zootaxa, 1852: 21–36. – 2009. Soft–bottom sipunculans in Izmir Bay (Aegean Sea, eastern Mediterranean). Zootaxa, 2136: 40–48. Açik, S., Murina, G. V., Çinar, M. E. & Ergen, Z., 2005. Sipunculans from the coast of northern Cyprus (eastern Mediterranean Sea). Zootaxa, 1077: 1–23. Adolph, E. F., 1936. Differential permeability to water and osmotic exchanges in the marine worm Phascolosoma. Journal of Cellular and Comparative Physiology, 9(1): 117–135. Buchanan, J. B., 1984. Sediment analysis. In: Methods for the Study of Marine Benthos, 2nd edition: 41–46 (N. A. Holme & A. D. McIntyre, Eds.). Blackwell Scientific Publications, Oxford. Chew, S. F., Peng, K. W., Low, W. P. & Ip, Y. K., 1994. Differences in the responses between tissues of the body wall and the internal organs of Phascolosoma arcuatum (Sipuncula) to changes in salinity. Comparative Biochemistry and Physiology, 107A: 141–147. Clarke, K. R. & Warwick, R. M., 1994. Change in Marine Communities: An Approach to Statistical Analysis and Interpretation. Plymouth Marine Laboratory, Plymouth. Cutler, E. B., 1965. Sipunculids of Madagascar. Extrait des Cahiers ORSTOM. Oceanographie, 3(4): 51–63. – 1994. The Sipuncula: Their Systematics, Biology, and Evolution. Cornell Univ. Press, Ithaca, New York. – 1977. The bathyal and abyssal Sipuncula. Galathea Report, 14: 135–156. Del–Pilar–Ruso, Y., De–la–Ossa–Carretero, J. A., Loya–Fernández, A., Ferrero–Vicente, L. M., Giménez–Casalduero, F. & Sánchez–Lizaso, J. L., 2009. Assessment of soft–bottom Polychaeta assemblage affected by a spatial confluence of impacts: sewage and brine discharges. Marine Pollution Bulletin, 58: 765–786. Fernández–Torquemada, Y., Gónzalez–Correa, J. M., Loya, A., Ferrero, L. M., Díaz–Valdés, M. & Sánchez–Lizaso, J. L., 2009. Dispersion of brine discharge from seawater reverse osmosis desalination plants. Desalination and Water Treatment,

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Editor executiu / Editor ejecutivo / Executive Editor Joan Carles Senar

Secretaria de Redacció / Secretaría de Redacción / Editorial Office

Secretària de Redacció / Secretaria de Redacción / Managing Editor Montserrat Ferrer

Museu de Zoologia Passeig Picasso s/n 08003 Barcelona, Spain Tel. +34–93–3196912 Fax +34–93–3104999 E–mail mzbpubli@intercom.es

Consell Assessor / Consejo asesor / Advisory Board Oleguer Escolà Eulàlia Garcia Anna Omedes Josep Piqué Francesc Uribe

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Biomass response to environmental factors in two congeneric species of Mullus, M. barbatus and M. surmuletus, off Catalano–Levantine Mediterranean coast of Spain: a preliminary approach M. García–Rodríguez, A. Fernández & A. Esteban García–Rodríguez, M., Fernández, A. & Esteban, A., 2011. Biomass response to environmental factors in two congeneric species of Mullus, M. barbatus and M. surmuletus, off Catalano–Levantine Mediterranean coast of Spain: a preliminary approach. Animal Biodiversity and Conservation, 34.1: 113–122. Abstract Biomass response to environmental factors in two congeneric species of Mullus, M. barbatus and M. surmuletus, off Catalano–Levantine Mediterranean coast of Spain: a preliminary approach.— We analyzed the influence of some abiotic variables in the biomass distribution of these species using survey data collected over four years (2006–2009) in the Catalano–Levantine coast of Spain. The preliminary results show that variables such as time (year) and latitude feebly affect the biomass distribution of these species. Depth, by itself, is not as significant as believed, masking the influence of other variables. M. barbatus biomass distribution seems to be especially influenced by salinity and, to a lesser extent, by temperature, while only temperature seems to have a significant effect on the M. surmuletus biomass distribution. These results are consistent with the bathymetric distribution of both species, with M. barbatus showing affinity for low salinity waters and M. surmuletus for warmer waters, which may contribute to the segregation of the species. Key words: Mullus, Biomass, Abiotic factors, Western Mediterranean. Resumen Respuesta de la biomasa a factores ambientales en dos especies congenéricas, M. barbatus y M. surmuletus, en aguas catalano–levantinas de la costa Mediterránea española: planteamiento preliminar.— Se ha analizado la influencia de algunos factores abióticos en la distribución de las biomasas de las especies utilizando datos de campaña recogidos a lo largo de cuatro años (2006–2009) en la costa catalano–levantina de España. Los resultados preliminares muestran que las distribuciones de la biomasa de las dos especies se ven afectadas débilmente por la época del año y la latitud. La profundidad no resulta tan significativa como se esperaba, enmascarando la influencia de otras variables. La distribución de la biomasa de M. barbatus está especialmente afectada por la salinidad y, en menor medida, por la temperatura; y en el caso de M. surmulentus, parece que únicamente la temperatura tiene un efecto significativo en la distribución de la biomasa. Estos resultados son consistentes con la distribución batimétrica de ambas especies, mostrando una afinidad por aguas de salinidad reducida en el caso de M. barbatus, y por aguas cálidas en el caso de M. surmuletus, que puede contribuir a la segregación de las especies. Palabras clave: Mullus, Biomasa, Factores abióticos, Mediterráneo occidental. Mariano García Rodríguez, Inst. Español de Oceanografía, Servicios Centrales, c./ Corazón de María 8, 28002, Madrid, España (Spain).– Ángel Fernández & Antonio Esteban, Inst. Español de Oceanografía, Centro Oceanográfico de Murcia, c./ Varadero 1, 30740 San Pedro del Pinatar, Murcia, España (Spain). Corresponding author: M. García–Rodríguez. E–mail: mariano.garcia@md.ieo.es

ISSN: 1578–665X

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Introduction The red mullet (Mullus barbatus L., 1758) and the striped red mullet (Mullus surmuletus L., 1758) are common demersal fishes of the Mediterranean Sea that appear distributed all around the Mediterranean basin and the North–Western Atlantic, mostly at depths less than 200 m in the shelf. M. barbatus inhabits sandy and muddy bottoms, while M. surmuletus is generally found on bottoms with heterogeneous granulometry and often on Posidonia beds. They show bathymetric habitat partitioning and clear niche segregation in relation to the bottom type that constitutes their habitat (Margalef, 1980; Hureau, 1986; Lombarte et al, 2000). Both species are among the most valuable resources for fisheries, being fished simultaneously or sequentially using a number of gears that vary over the year (Martin et al., 1999). In the Spanish Mediterranean, the trawl fleets generate 80% of the Mullus landings, with M. barbatus representing ≈ 70% of this fraction. However, in small–scale fisheries that account for the remaining 20% of the total landings, M. surmuletus represents 75% of the catch, and M. barbatus accounts for the remaining 25%. The mullet trammel nets are preferably used in areas where M. surmuletus concentrates, such as the coastal rocky bottoms and, more generally, at depths over 50 m or at the limit of the meadows of Posidonia oceanica (L.) Delile (1813), thus attaining higher yield in the bottoms (Baino et al., 1998), while avoiding any interference with trawl fishing (García–Rodríguez et al., 2006). Taking both fisheries together, the proportion of species in total landings is almost balanced, with a slight dominance (60/40) of M. barbatus. Inter–annual fluctuations in volume are high (Fernández, pers. com.) and are present despite fishing efforts remaining almost constant. This suggests that fluctuations do not depend only on fishing activities, but also on the environmental conditions. In this sense, some recent studies related sea–surface temperature with recruitment success for M. barbatus in the strait of Sicily (Levi et al., 2003), and Machias et al. (1998) established the ranges of bottom depth, temperature and salinity over which M. surmuletus is distributed in the Cretan shelf. In addition, generalised additive models (GAMs) have been applied to test the hypothesis that M. barbatus abundance is related to the bathymetry, spatial location and temperature variability of the NE Mediterranean (Maravelias et al., 2007). To shed some light on this topic, we developed an exploratory study on the influence of several abiotic variables (year, latitude, depth, temperature and salinity) in the distribution of these two species in the Catalano–Levantine coast of Spain. Material and methods Sampling took place in the Catalano–Levantine coast of Spain (FAO–GFCM Geographic Sub Area 06, GSA 6). All samples were collected during the course of four consecutive MEDITS_ES International Spring

Table 1. Main statistical values of the considered abiotic variables (depth, temperature and salinity) of total samples recorded over the 2006–2009 period in the Catalano–Levantine coast of Spain: D. Depth (m); T. Temperature (ºC); S. Salinity (0/00). Tabla 1. Principales valores estadísticos de las variables abióticas consideradas (profundidad, temperatura y salinidad) del total de las muestras recogidas en el periodo 2006–2009 en las costas catalano–levantinas de España: D. Profundidad (m); T. Temperatura (ºC); S. Salinidad (0/00).

Min.

12.81182

37.79290

Max.

816

16.72884

38.54560

Average

193.55

13.5469

38.2414

(± SD)

(± 188.26)

Range

783

(± 0.71355) (± 0.1727) 3.917018

0.752704

Trawl Surveys (from 2006 to 2009) according to the international standard methodology (Relini et al., 2008). Sea depth, temperature and salinity were recorded using a CTD SBE–37 probe located in the mouth of the gear and represented in situ observations of the hydrological conditions associated with each catch. For each of the above variables, individual haul averages were estimated from the data recorded during the effective trawl (when the gear is in contact with the bottom) and included in the analyses as variables. Another variable included was latitude, while year of survey (2006–2009) was considered as a factor. Fish biomass per haul was calculated as the catch in weight by sweep area and expressed in kg/km2. Some cartographic depictions of sea temperature, salinity and fish abundance (expressed as biomass captured per square kilometre) were obtained applying a geostatistical kriging model over the cumulated data collected in the study. An exploratory scrutiny of the data was carried out by means of covariance analysis, linear regression and correlation (Pearson’s correlation coefficient) to elucidate whether the above variables had any relationship to the biomass distribution of the two mullets during the study period. To clarify whether species had any 'preference' in their appearance, a t–test, an analysis of variance (ANOVA) and a Tukey test were performed over the data distribution, previously normalised by means of a logarithmic transformation, to test for significant differences in the mean values of the variables between samples with presence and samples without presence of each species.

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Table 2. Main statistical values of abiotic variables considered (depth, temperature and salinity) in samples with Mullus occurrence. Tabla 2. Principales valores estadísticos de las variables abióticas consideradas (profundidad, temperatura y salinidad) en las muestras con presencia de las especies de Mullus.

Depth (m)

Temperature (ºC)

Salinity (0/00)

Biomass (kg/km2)

M. barbatus Min.

12.8118

37.7929

0.36

Max.

535

16.7288

38.3913

658.53

102.93

13.6990

38.1569

43.68

(± SD)

(± 68.34)

(± 0.8362)

(± 0.1176)

(± 86.73)

Range

502

3.9170

0.5984

658.16

12.8118

37.9048

0.50

Average

M. surmuletus Min. Max.

535

16.7288

38.5334

547.49

126.71

13.7364

38.1922

19.04

(± SD)

(± 103.51)

(± 0.8792)

(± 0.1492)

(± 62.96)

Range

499

3.9170

Average

Finally, a generalized linear model (GLM) was also performed. Data were normalised by transforming biomass to Ln, and the relationship between the different factors and the species biomass was analysed by means of multiple regressions, applying a simple model without interactions and identity as a link: Ln (biomass jklm) = μ + Yj + Lj + Dk + Tl + Sm + εijklm where: μ is overall mean; Yi, effect of year i; Lj, effect of latitude j; Dk, effect of depth stratum k; Tl, effect of temperature l, Sm, effect of salinity m; ε, error term assumed to be distributed normally. A deviation analysis was carried out to evaluate the significance of the factor and variables in the model. Deviance represents the variation present in the data and its analysis results in a table that summarises the information related to the sources of variation of the data, in a similar way to an ANOVA. In this table, each variable copes with an amount of deviance that represents the amount of variation of the response explained by the variable. Statistical analysis was performed with the S–PLUS software (MathSoft, Seattle, WA, USA). Results Analyses comprised data from 293 hauls (33–816 m depth) collected over a period of four years (2006– 2009). Table 1 shows the average and range of each of the studied abiotic variables. Depth showed negative and positive significant correlations with

0.6286

546.99

temperature (r2 = 0.54; t = –0.916; p = 0.000) and salinity (r2 = 0.75; t = 19.525; p = 0.000), respectively. Correlation between temperature and salinity was negative and significant (r2 = 0.44; t = –8.335; p = 0.000). Thus, both salinity and temperature were highly correlated with depth, showing a gradient on both the continental shelf and the upper slope, with colder and saltier waters in deeper zones (fig. 1). The average biomass of each species seemed to covariate positively with latitude and temperature and negatively with depth and salinity. Despite r2 values being low (< 0.27), in the case of M. barbatus all relations were significant (latitude: t = 2.0658, p = 0.041; depth: t = –2.424, p = 0.0165; temperature: t = 3.564, p = 0.0005; and salinity: t = –3.248, p = 0.0014). In M. surmuletus, temperature was the only variable showing a significant relation with the fish biomass (r2 = 0.218; t = 2.362; p = 0.01996). M. barbatus appeared in 161 hauls (55% of the total) comprising a depth range 502 m wide, with maximum biomass values in the 0–200 m depth interval decreasing abruptly thereafter. M. surmuletus appeared in 114 samples (39%), showing a depth range 499 m wide, with maximum values in the 0–50 m interval decreasing gradually with depth. Ranges and averages of variables in samples with presence of each species were moderately similar. Only average biomass was different, with M. barbatus being more abundant (twofold) than M. surmuletus (table 2). Mean values and variance of variables in samples with presence of any of the mullet species were similar to those without it excepting depth, which showed evident differences (fig. 2). Latitude had no significance;

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García–Rodríguez et al.

Temperature 42.5 42.0

Iberian peninsula

ºC 15.5 15.25 15 14.75 14.50 14.25 14 13.75 13.50 13.25 13 12.75 12.50

41.5 41.0 40.5 40.0 39.5 39.0 38.5 38.0 37.5

42.5

Mediterranean Sea –1 –0.5 0

0.5

1.5

2.5

42.0

Iberian peninsula

38.55 38.50 38.45 38.40 38.35 38.30 38.25 38.20 38.15 38.10 38.05 38.00 37.95 37.90 37.85 37.80 37.75

41.0 40.5 40.0 39.5 39.0 38.5 38.0

Mediterranean Sea –1 –0.5

Mullus barbatus 42.5

0/00

41.5

37.5

3.5

Salinity

0.5

1.5

2.5

3.5

Mullus surmulentus

Iberian peninsula

kg/km2

42.0

42.5

Iberian peninsula

kg/km2

42.0

41.5

100

41.5

100

41.0

40.5

40.0

39.5

39.0

38.5

38.0 37.5

Mediterranean Sea –1 –0.5

0.5

1.5

2.5

3.5

30 20 10 5 1

38.5

38.0 37.5

Mediterranean Sea –1 –0.5

0.5

1.5

2.5

3.5

Fig. 1. Representation of the spring distribution of temperature and salinity, and M. barbatus and M. surmuletus biomasses in the studied area. Cumulated data for 2006–2009. Fig. 1. Representación de la distribución en primavera de la temperatura y salinidad, así como las biomasas de M. barbatus y M. surmuletus, en el área estudiada. Datos acumulados para el periodo 2006–2009.

however, sea depth, temperature and salinity resulted in significant differences in all tests (p < 0.05), suggesting that biomass distribution of the species shows some 'preferences' regarding the selected variables. The mean values of depth of samples associated with the presence of biomass of both species were lower than means in samples where they were not found, in a similar way to salinity. On the other hand, temperature exhibited upper mean values related with species appearance. In a first interpretation, both species seem to 'prefer' shallow waters (well–known circumstance), characterised by low salinity and warmer temperature. Ranges (minimum and maximum) of variables in samples with

species’ appearance could be considered as ranges of distribution of the species in the sampled area. Preliminary GLM analysis resulted in a strongly asymmetrical biomass distribution, with numerous extreme data. Logarithmic transformation of the data (biomass) partially solves this imbalance, reducing part of the extreme values for M. barbatus (fig. 3) as well as for M. surmuletus (fig. 4). The modelled biomasses exhibit moderate linearity, with some scattered data, and the total deviance explained by the models is scarce. The partial residuals of the variables, in relation to the response, indicates that the model corresponds with the data (fig. 5). The model explained 17% of

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3.80

7 6

3.75

3.70

Depht

Latitude

2.8

3.67

3.66

2.6

3.65

3.80

7 6

3.75

Depht

Latitude

3 3.68

Salinity

Temperature

3.65 2.9

2.7

3.70

4 3 3.68

2.7

2.6

Salinity

Temperature

3.65 2.9

2.8

3.67

3.66

3.65 0 1 Presence

0 1 Presence

Fig. 2. Box plots for each of the studied abiotic variables (latitude, depth, temperature and salinity) in samples with absence (0) and occurrence (1) of: A. M. barbatus; B. M. surmuletus. (Means of depth, temperature and salinity in samples with species presence were significantly different from those without presence.) Fig. 2. Diagramas de caja para cada variable estudiada (latitud, profundidad, temperatura y salinidad) en muestras con ausencia (0) o con presencia (1) de: A. M. barbatus; B. M. surmuletus. (Las medias de profundidad, temperatura y salinidad de las muestras con presencia de la especie resultaron ser significativamente diferentes de las muestras sin presencia.)

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García–Rodríguez et al.

Histogram biomass

Histogram logarithm biomass 40

120

80 60 40 20 0

20 10 0

0 200 400 600 Biomass

Box plot biomass

200

2 1 0 -1 -2 -3 -4

Pearson Residuals

400

Logarithm biomass

0 2 4 Logarithm biomass

Box plot logarithm biomass

600

–2

0 2 4 6 Fitted: year + latitude + depth + + temperature + salinity

2 1 0 1 2 Quantiles of standard normal

Fig. 3. Histograms and box plots of the M. barbatus biomass in the positive observations: raw data, top and middle left; and transformed data, top and middle right. Response of the adjusted variable (bottom left) and normal probability graphic of the Pearson residuals (bottom right) for the applied model are also included. Fig. 3. Histogramas y diagramas de caja de la biomasa de M. barbatus en las observaciones positivas: arriba y centro a la izquierda datos brutos; arriba y centro a la derecha datos transformados. Respuesta de la variable ajustada (abajo izquierda) y gráfico de probabilidad normal de los residuales de Pearson (abajo derecha) para el modelo aplicado.

variance in M. barbatus, with salinity (7%) as the most explanatory factor, followed by temperature (3%). In the case of M. surmuletus, the model explained 8% of variance, with temperature (5%) as the main explanatory factor (table 3). Although for both species the biomass decreased with increased depth, salinity and rise of temperature and the ranges of the analysed variables were quite similar, we found remarkable differences between the two in the specific effect of each variable. These results suggest that mullets have environmental

preferences in their distribution, with an important and negative effect of salinity in M. barbatus biomass and a positive and less intense effect of temperature in the M. surmuletus case. Discussion Our results suggest that, in the studied area, the biomass contribution of each species is different, with lower values for M. surmuletus. This may be because

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Histogram biomass

Histogram logarithm biomass 40

120

80 60 40 20 0

20 10 0

0 200 400 600 Biomass

Box plot biomass

4 2

200

2 1 0 -1 -2 -3 -4

Pearson Residuals

6 Logarithm biomass

400

0 2 4 Logarithm biomass

Box plot logarithm biomass

600

â&#x20AC;&#x201C;2

0 0 2 4 6 Fitted: year + latitude + depth + + temperature + salinity

2 1 0 1 2 Quantiles of standard normal

Fig. 4. Histograms and box plots of the M. surmuletus biomass in the positive observations: raw data, top and middle left; and transformed data, top and middle right. Response of the adjusted variable (bottom left) and normal probability graphic of the Pearson residuals (bottom right) for the applied model are also included. Fig. 4. Histogramas y diagramas de caja de la biomasa de M. surmuletus en las observaciones positivas: arriba y centro a la izquierda datos brutos; arriba y centro a la derecha datos transformados. Respuesta de la variable ajustada (abajo izquierda) y grĂĄfico de probabilidad normal de los residuales de Pearson (abajo derecha) para el modelo aplicado.

the gear used as sampler (bottom trawl net) can only be applied in smooth bottoms. Thus, M. surmuletus showed a clear preference for rough bottoms, while M. barbatus had a greater abundance on soft bottoms, being more accessible to the sampler. This difference in substrate preference was especially marked in young individuals (Lombarte et al., 2000), which inhabit very close to the shoreline. In addition, these results coincide with findings of Tserpes et al. (2002) for the Mediterranean shelf. With respect to the effect of time

(year) on the biomass distribution, no significance was found, and only a diminishing trend could be identified in M. barbatus (fig. 5). Both species had a relatively wellâ&#x20AC;&#x201C;balanced distribution along the sampling area. Averaged biomass values for M. barbatus diminished slowly toward the north, while M. surmuletus increased slightly, but not in a significant way in any case (fig. 5). Depth is assumed to have an important role in species distribution. In this study, this factor correlated

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García–Rodríguez et al.

Table 3. Results of the GLM analysis of the considered variables, for Mullus biomass as response, showing the explained deviance for each variable and its significance. Df. Degree od freedom; % Dev. Percentage of total deviance explained by the variable; F. Values of Fisher's test; Pr(F). Probability at a 0.05 significance level. Tabla 3. Resultados del análisis GLM de las variables consideradas, con la biomasa de Mullus como respuesta, y mostrando la deviación explicada para cada variable y su significación: Df. Grados de libertad; % Dev. Porcentaje de la desviación total explicada por la variable; F. Valores del test de Fisher; Pr(F). Probalbilidad a un nivel de significación de 0,05.

Deviance

160

464.7119

Year

18.33183

157

446.3800

Latitude

9.25879

156

Depth

4.97387

155

Temperature

14.83192

Salinity

33.40117

Null

M. barbatus

Deviance

Pr(F)

3.94

2.43524

0.0669743

437.1212

1.99

3.68987

0.0566035

432.1474

1.07

1.98222

0.1611854

154

417.3155

3.19

5.91091

0.0162028

153

383.9143

7.19

13.31125

0.0003613

Total % Null

M. surmuletus

113

179.436

% Dev

17.39

Year

4.918596

110

174.5174

2.74

1.057042

0.3706481

Latitude

1.411066

109

173.1063

0.79

0.909745

0.3423521

Depth

0.260992

108

172.8453

0.15

0.168267

0.6824842

Temperature

8.153103

107

164.6922

4.54

5.256483

0.0238397

Salinity

0.280186

106

164.412

0.16

0.180642

0.6716835

Total %

8.37

negatively with the biomass of both species, and the correlation was significant for M. barbatus. In the case of M. surmuletus, previous studies have shown a significant relationship between biomass and depth over the year in the Iraklion Gulf (Machias et al., 1998). However, our results were far from significant in this species. Although both species concentrates in the first 200 m depth of the shelf, with highest biomass values in the first 50 m and decreasing thereafter, M. surmuletus increases its average biomass in the 200–500 m depth interval, recovering the values showed in the 50–100 m depth interval (fig. 5). Until now, M. surmuletus was thought to have a wider bathymetric range than M. barbatus which, in contrast to M. surmuletus, never appeared below the 200 m depth in the Spanish Mediterranean (Lombarte et al., 2000), but reached the 328 m depth in the Ionian Sea (Mytilineou et al, 2005). In addition, the observed bathymetric distribution of M. surmuletus has increased with time. Thus, Hureau (1986) reported that M. surmuletus inhabits depths of less than 100 m and Macpherson & Duarte (1991) found a depth range of 12 to 182 m. More recently, Machias et al. (1998) found the species between 28 and 310 m depths in Crete, Mytilineou et al. (2005) expanded the range from 5 to 409 m depth in the Ionian Sea, and García–Rodríguez et al. (2007) found M. surmuletus down to 716 m off Castellón. In the present study, both species appeared in similar depth

ranges, with M. barbatus achieving 535 m depth and M. surmuletus biomass recovered in the 200–500 m interval, mainly due to the occurrence of a small amount of big–sized individuals. Depth by itself only means barometric pressure, with the water masses being characterised by their physical characteristics and chemical composition, and has no significance in the GLM results (table 3). Consequently, we consider that bathymetric segregation is not as clear as believed to date as a function of depth, and could be attributed to other abiotic variables, highly correlated with depth, but possibly masked in their influence by depth. Temperature is the most important physical characteristic of water masses. In this seasonal study (spring), temperature decreased with depth, and the observed range for temperature (3.92ºC) (table 1) was slightly wider than that observed by Machias et al. (2000) in the Cretan spring for M. surmuletus (2.8ºC). In both studies, temperature had a positive correlation with biomass, being significant only in the present study. Temperature was also one of the explanatory variables in the GLM models. In the case of M. barbatus, and despite temperature not being the most explanatory variable, it explains ≈ 3% of the observed deviance, while in M. surmuletus, temperature is the most important variable (table 3). Maravelias et al. (2007) found that the mean M. barbatus abundance in the Aegean Sea was consistently higher in areas with shallower

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2 0

–2

2 1 0 –1 –2

Partial for year

3 2 2 0

–1 –2 –3 –4

100 200 300 400 500 Depth

4 2 0

14 15 16 Temperature

37.8

38.0 38.2 Salinity

38.4

3 2 1 0

Partial for salinity

4 2 0

2 1 0

–2

2006 2007 2008 2009 Year

–1

–2

Partial for temperature

M. surmuletus

–1

39 40 41 42 Latitude

100 200 300 400 500 Depth

3 2 1 0

–1 –2

–2

–4

39 40 41 42 Latitude

Partial for depth

–4

–2

–3

Partial for latitude

Partial for temperature

2006 2007 2008 2009 Year

2 1 0 –1 –2 –3 –4

Partial for salinity

–4

Partial for depth

Partial for latitude

Partial for year

M. barbatus

14 15 16 Temperature

37.8

38.0 38.2 Salinity

38.4

Fig. 5. Graphs showing partial residuals of variables in the model of M. barbatus and M. surmuletus. Points (●) represent the biomass values plotted against mean (year) or the adjusted regression line in variables. Bar width at the bottom of each figure is proportional to observations. Fig. 5. Gráficas de los residuales parciales de las variables en el modelo para M. barbatus y M. surmuletus. Los puntos (●) representan los valores de biomasa con respecto a la media (año) o a la recta de regresión ajustada en las variables. La anchura de las barras de la parte inferior de cada figura es proporcional a las observaciones.

122

depths (35–60 m) and warmer bottom waters (19ºC) than in the central area, with deeper and colder waters. This species seems to avoid the cold bottom waters (< 16ºC) of the deeper regions (Maravelias et al., 2007). In the present study, M. barbatus biomass seems to be related to temperatures comprised in the range 12.81–16.73ºC and, although the relationship was positive, the temperatures we recorded were lower than those reported by Machias et al. (2000) in spring and by Maravelias et al. (2007) in summer. Besides, our results suggest that temperature is an explanatory variable for both species distribution, especially in the case of M. surmuletus, for whom it represents the main source of biomass variation (table 3, fig. 5). Salinity is the main chemical characteristic of marine water and, in combination with temperature, it clearly defines the different water masses of any specific area. Some authors (Tsimenides et al., 1991; Machias et al., 2000) hold that salinity shows very small variation and it is considered not to have an effect on the fish distribution on the Cretan shelf. In this study, salinity increases with depth, showing a range of variation of 0.7527 psu, and is negatively correlated with the biomass of the studied species, a correlation that is significant in the case of M. barbatus. GLM results suggest that for M. barbatus, salinity is the main explanatory variable for the biomass distribution, followed in importance by temperature. In the case of M. surmuletus, the influence of salinity variations in its biomass distribution is negligible (table 3, fig. 5). In conclusion, we observed that the biomass of both mullet species varied minimally over time, showing a uniform distribution in the studied area, with M. barbatus being more abundant. Depth seems to have a moderate influence by itself in the biomass distribution, and the observed variations can be attributed to other water characteristics highly correlated with depth. Thus, in the case of M. barbatus, biomass distribution is related with the T–S of the water masses, but this relation is mostly due to the water salinity. In contrast, M. surmuletus biomass seems to be significantly affected by the temperature of the water. These preliminary results are consistent with the similar bathymetric distribution shown by both species, with an affinity for waters with low salinity in the case of M. barbatus, and for warmer waters in the case of M. surmuletus, which can contribute, more clearly than depth, to the segregation of the species. Further analysis over an extended database, to improve accuracy, may support these interesting preliminary results. References Baino, R., Righini, P. & Silvestri, R., 1998. Target species and CPUE of trammel, gillnet and combined net in sandy and rocky bottoms. Rapport Commission Internationale pour l’exploration Scientifique de la Mer. Mediterraneé, 35(2): 516–517. García–Rodríguez, M., Fernández, A. M. & Esteban, A., 2006. Characterisation, analysis and catch rates

García–Rodríguez et al.

of the small–scale fisheries of the Alicante Gulf (S. E. Spain) over a ten–year series. Fisheries Research, 77(2): 226–238. García–Rodríguez, M., Pérez Gil, J. L., Peña, J. & Sáez, R., 2007. Resultados de una prospección comercial al arrastre de fondo en una zona no explotada del talud continental, junto a las Islas Columbretes (Castellón, este de la península ibérica). Informes Técnicos del Instituto Español de Oceanografía., 187: 1–52. Hureau, J. C., 1986. Mullidae. In: Fishes of the North– eastern Atlantic and the Mediterranean, Vol. II: 877– 882 (P. J. P. Whitehead, M. L. Bauchot, J. C. Hureau, J. Nielsen & E. Tortonese, Eds.). UNESCO, Paris. Levi, D., Andreoli, M. G., Bonanno, A., Fiorentino, F., Garofalo, G., Mazzola, S., Norrito, G., Patti, B., Pernice, G., Ragonese, S., Giusto, G. B. & Rizzo, P., 2003. Embedding sea surface temperature anomalies into the stock recruitment relationship of red mullet (Mullus barbatus L. 1758) in the Strait of Sicily. Scientia Marina, 67(1): 259–268. Lombarte, A., Recasens, L., González, M. & Gil de Sola, L., 2000. Spatial segregation of two species of Mullidae (Mullus surmuletus and M. barbatus) in relation to habitat. Marine Ecology Progress Series, 206: 239–249. Machias, A., Somarakis, S. & Tsimenides, N., 1998. Bathymetric distribution and movements of red mullet Mullus surmuletus. Marine Ecology Progress Series, 166: 247–257. Macpherson, E. & Duarte, C. M., 1991. Bathymetric trends in demersal fish size: is there a general relationship? Marine Ecology Progress Series, 71: 103–112. Maravelias, C. D., Tsitsika, E. V. & Papaconstantinou, C., 2007. Environmental influences on the spatial distribution of European hake (Merluccius merluccius) and red mullet (Mullus barbatus) in the Mediterranean. Ecological Research, 22: 678–685. Margalef, R., 1980. Ecología. Ed. Omega, Barcelona. Martin, P., Sartor, P. & García–Rodríguez, M., 1999. Comparative analysis of the exploitation strategy of the European hake (Merluccius merluccius), Red mullet (Mullus barbatus) and striped red mullet (Mullus surmuletus) in the Western Mediterranean. Journal of Applied Ichthyology, 15: 24–28. Mytilineou, C., Politou, C.–Y. Papaconstantinou, C., Kavadas, S., D’onghia, G. & Sion, L., 2005. Deep– water fish fauna in the Eastern Ionian Sea. Belgian Journal of Zoology, 135(2): 229–233. Relini, G., Carpenteri, P. & Murenu, M. (Eds.), 2008. Manuale di Instruzioni Medits. Biologia Marina Mediterranea, 15(2): 1–78. Tserpes, G., Fiorentino, F., Levi, D., Cau, A., Murenu, M., Zamboni, A. & Papaconstantinou, C., 2002. Distribution of Mullus barbatus and M. surmuletus (Osteichthyes: Perciformes) in the Mediterranean continental shelf: implications for management. Scientia Marina, 66(2): 39–54. Tsimenides, N., Tserpes, G., Machias, A. & Kallianiotis, A., 1991. Distribution of fishes on the Cretan shelf. Journal of Fish Biology, 39: 661–672.

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Trophic indicators to measure the impact of fishing on an exploited ecosystem M. G. Pennino, J. M. Bellido, D. Conesa & A. López–Quílez

Pennino, M. G., Bellido, J. M., Conesa, D. & López–Quílez, A., 2011. Trophic indicators to measure the impact of fishing on an exploited ecosystem. Animal Biodiversity and Conservation, 34.1: 123–131. Abstract Trophic indicators to measure the impact of fishing on an exploited ecosystem.— There is currently a global call for more use of an ecosystem approach to fisheries management (EAFM) to provide a holistic view of ecosystem–fisheries interactions. Trophic indicators could therefore be used to support the implementation of an EAFM by providing information on the state of the ecosystem. In this paper we propose a set of indicators such as the marine trophic index (MTI), the fishing in balance (FiB), the cutmarine trophic index (cutMTI) and the pelagic/demersal index (P/D) to assess the dynamics and the trophic changes in the Black Sea large marine ecosytem from 1970 to 2005. Our analysis shows a heavily exploited ecosystem where overfishing and anthropogenic eutrophication are probably responsible. The decline of the MTI, cutMTI and FiB together with the rising trend of the P/D index could be interpreted as a fishing down marine food web process with a strong decrease in abundance of high trophic level species and a considerable increase of low trophic level species. Key words: Trophic indicators, Ecosystem approach to fisheries management, Black Sea. Resumen Indicadores tróficos para medir el impacto de la pesca en un ecosistema explotado.— Cada vez son más frecuentes las iniciativas para la aplicación de un enfoque del ecosistema para la gestión de la pesca (ecosystem approach to fisheries management, EAFM) que proporciona una visión holística de las interacciones entre la pesca y el ecosistema. Los indicadores tróficos pueden ser herramientas ideales para la aplicación de EAFM proporcionando información sobre el estado del ecosistema. En este trabajo se propone un conjunto de indicadores tales como el índice trófico marino (MTI), el índice pesca en equilibrio (FiB), el índice trófico marino rebajado (cutMTI) y el índice pelágicos/demersales (P/D) para evaluar la dinámica y los cambios tróficos en el Mar Negro desde 1970 hasta 2005. Nuestro análisis muestra un ecosistema muy explotado, donde la pesca excesiva y la eutrofización antropogénica son probablemente los responsables. El decremento en la tendencia del MTI, el cutMTI y el FiB conjuntamente con el incremento de la proporción P/D, se podría interpretar como una situación de la red trófica marina debida a la sobrepesca (fishing down marine food weeb situación). Este fenómeno implica una fuerte disminución en la abundancia de las especies de alto nivel trófico y un considerable aumento de las especies de bajo nivel trófico. Palabras clave: Indicadores tróficos, Enfoque del ecosistema para la gestión de la pesca, Mar Negro. Maria G. Pennino & José M. Bellido, Inst. Español de Oceanografía, Centro Oceanográfico de Murcia, c./ Varadero 1, San Pedro del Pinatar, 30740 Murcia, España (Spain).– David Conesa & Antonio López–Quílez, Dept. Estadística e Investigación Operativa, Univ. de València, c./ Dr. Moliner 50, Burjassot, 46100 València, España (Spain). Corresponding author: Maria G. Pennino. E–mail: grazia.pennino@mu.ieo.es

ISSN: 1578–665X

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Introduction The sustainability of the present fishery system is being questioned and failure of the traditional stock assessment and management is generally recognized (Garcia & De Levia Moreno, 2003). Sources of failure are many and varied for different types of fishing. In general, the failure can be attributed to the type of approach used, inadequate to capture multispecies and ecosystem effects (Cury et al., 2004). Considering factors that impact marine resource populations in a context beyond just the species level is one of the principal problems (Christensen et al., 1996). It has led to greatly simplify the nature and dynamics of resources, neglected the socio–economic dimensions of fisheries and the trophic relationships between species. The main consequence has been a series of ecological and economic collapses in numerous important fisheries in many parts of the world (Pauly et al., 2000a). These considerations led to the new ecosystem approach to fisheries management (EAFM) which is a more holistic approach to resource allocation and management (Larkin, 1996). The FAO defined the EAFM as the extension and integration of conventional methods of management of marine resources, stressing the close interdependence between the welfare of humanity, the preservation of the environment and the need to maintain productivity of ecosystems for present and future generations (Ward et al., 2002; García et al., 2003). The aim of an EAFM is to sustain an ecosystem in a healthy, productive and resilient condition so that it can continue to provide the services humans want and need (Link, 2002). To make this new type of ecosystem management possible, new viewpoints and tools must be developed to facilitate communication between managers, stake holders and scientists. Acknowledging ecological interactions is a key point for an EAFM (Cury et al., 2003). The power of ecological processes such as trophodynamic interactions, i.e. predation and competition, has been identified as being of enormous importance in fish population dynamics (Bax, 1998). This involves two major problems in fisheries management. The first is the decrease in food resources on which some components of the ecosystem survive, necessitating the elimination of other localities or cause depletion (Link, 2002). The second is the indirect effect of decreasing fish biomass on the functioning of ecosystems (Cury et al., 2003). Therefore, there is a need for descriptive indicators (Murawski, 2000) that reflect and describe the complex interactions between fisheries and marine ecosystems (Pauly & Watson, 2004). Trophic indicators can help describe these in simpler terms that can be understood and used by non– scientists to make management decisions. Indicators could therefore be used to support the implementation of an ecosystem approach to fisheries management (EAFM) by providing information on the state of the ecosystem, the extent and intensity of effort or mortality and the progress of management in relation to objectives. Indicators should guide the management of fishing activities that have led to, or are most likely

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to lead to, unsustainable impacts on ecosystem components or attributes (Cury et al., 2005). The potential danger of interpreting a single indicator may be that it does not analyze the causes of the observed trajectory or understand the dynamics of fisheries (Pennino et al., 2011). The application of a set of trophic indices could possibly resolve this problem. The weaknesses of one index are overcome by others that allow us to understand what has occured in past years in this ecosystem on a macro scale (Pennino et al., 2011). In this paper we suggest a selection of indicators concerning TL to measure the impact of fishing on the Black Sea fish community, focusing on their meaning and sensitivity to fishing impacts. Material and methods The study area The Black Sea is the world’s most isolated sea, connected to the oceans via the Mediterranean Sea through the Bosphorus, Dardanelle and Gibraltar straits, and linked to the Sea of Azov in the northeast through the Kerch Strait (http://www.icpdr.org). The Black Sea is a highly productive ecosystem (> 300 g / cm2 year1) with a continental climate. The fluvial discharge (Balkas et al., 1990), the natural winter production (Sur et al., 1994; Nezlin et al., 1999) the presence in summer of upwelling and a strong density stratification, make the Black Sea the largest anoxic basin of the global ocean (Sea Around Us, 2007). The deep waters do not mix with the upper layers of water that receive oxygen from the atmosphere. As a result, over 90% of the deeper Black Sea volume is anoxic water (Oguz & Ducklow, 1999). The most peculiar feature of the Black Sea is the absence of marine life at depths beyond 150–200 m, except for a few anaerobic bacteria (The Encyclopedia of Ukraine, www.encyclopediaofukraine.com). Living organisms are concentrated in the shallow waters of the continental shelf and river mouths along the northwestern coast. The number of alien species recorded at the regional level amounts to 217 (Sea Around Us, 2007). This number, together with the high level of pollution, suggests a serious impact on the native biological diversity in the he Black Sea and negative consequences for human activities. The data set In this study we used the fishery landings of the Black Sea large marine ecosystem (see www.lme.noaa.gov for more details) for the years 1970–2005. Fishery data and TLs of the species were extracted from the database in http://www.seaaroundus.org. The data set consists of recorded nominal catches and does not include discarded species. Specific landings were grouped into 11 trophic groups taking their trophic level into account (table 1). All indexes described were applied to this database to obtain an integrated vision of the environmental

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Table 1. Trophic groups with trophic levels (TL) and standard errors (SE). Tabla 1. Los grupos troficos con los respectivos niveles troficos (TL) y errores estándar (SE). Groups

Bivalvia

2.02

0.03

Crustaceans

2.53

0.7

Molluscs

2.82

0.8

Engraulidae

3.11

0.45

Flatfishes

3.13

0.32

Cupleidae

3.15

0.19

Scorpionfishes

3.56

0.45

Carangidae

3.64

0.32

Percidae

3.66

0.19

Gadidae

4.05

0.68

Sharks and rays

4.15

0.79

and fishery problems of the Black Sea ecosystem and to test their performance. Also, each series of indicators was smoothed using a locally weighted scatterplot smoothing, LOESS, which estimates a polynomial regression curve using local fitting (Cleveland, 1979). Bootstrap methodology was applied to assess variability. A 95% confidence band obtained from the bootstrapped samples was calculated for the original LOESS fit. Bootstrap (Efron, 1979) is a computer–intensive method that quantifies uncertainty for a large range of problems. It is based on resampling from an original sample of data to create replicate datasets and it provides inferences on the quantities of interest. Finally, the standard error of the TL was included in the bootstrap procedure for the indicators derived from the TL, such as MTI, FiB, and the cutMTI. All calculations were made with R (R Development Core Team, 2008). Fishery ecosytem indicators The marine trophic index – MTI In February 2004, the Conference of the Parties to the Convention on Biological Diversity (CBD) recognized a number of indicators to observe the decrease in the current rate of biodiversity loss (CBD, 2004). The MTI is one of these indicators. MTI is calculated from a combination of fisheries landings and diet composition data of the landed fish species. It is computed, for each year k from: MTIk = ∑i (TLi )(Yi ) / ∑ i Y i where MTI is the mean trophic level of landing in year

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k, Yi refers to the landings of trophic group i and TL is the trophic level of trophic group i. Fishing often targets the highest predators, allowing individuals at the next TL to expand in numbers, leading to excessive grazing on the level below, reduced predation on the level below that, and so on, alternately, down to the base of the food chain. These 'trophic cascades' can, in extreme circumstances, be disastrous for marine ecosystems (Daskalov et al., 2007). Pauly et al. (1998) showed the global declining trend of mean TL of catches from 1950 to 1994 based on the FAO dataset. The proposed explanation for this phenomenon, now widely known as 'fishing down marine food webs' (FDFW) is that the fishery catches are shifting from large, high–TL species to the small, low–TL species in response to their relative abundance in the ecosystem. The fishing down marine food web effect has also been shown in Thailand (Christensen, 1998), Canada (Pauly et al., 2001), Greece (Stergiou et al., 2000), Iceland (Valtỳsson & Pauly, 2003) and in many others countries (Pauly & Watson, 2004). This phenomenon is widespread because the high–TL species (e.g., large piscivorous fishes such as sharks, etc.), which are long–lived species with a low reproductive rate, are less resilient to overfishing and tend to be depleted quickly as compared to low–TL species, which are short lived and fast growing (Froese et al., 2004). The cutmarine trophic index – cutMTI The use of mean TL as a measure of impact of fisheries on marine ecosystem was questioned by Caddy et al. (1998) who had in mind processes such as the eutrophication. This phenomenon of coastal areas may result in increasing abundance of planktivores, thus lowering mean TL. As a diffuse and general problem eutrophication can modify the ratio between predator and prey abundances, which could then be confused with effects of fisheries (Caddy, 1993). The analysis of this increase would lead by decreasing average calculated TL at a high inference TL fish depletion, although this cannot be reduced in absolute terms. Pauly et al. (1998) noted a related problem due to fluctuations in the abundance of Peruvian anchoveta (Engraulis ringens), whose enormous catches strongly influence the mean TL of global catches. To obviate this problem Pauly & Watson (2004) suggested that the MTI should in fact be based on time series that exclude low–TL organisms. This would lead to an indicator labeled cutMTI, with the ‘cut’ referring to the lowest (cut–off) TL value used in the computation. Pauly proposes the cut–off value of 3.25 (3.25MTI). With a cut–off value of 3.25 all TL–levels lower than 3.25 are removed from the computation of the MTI, to eliminate the herbivores, detrivores and the planktivores whose biomass tends to vary widely in response to environmental factors (Pauly & Watson, 2004). Fishing in balance index – FiB Marine ecosystems operate as pyramids wherein the primary production generated at one TL is moved up toward the higher TL, with a huge fraction of that production being wasted in the process for the

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Landings 9e+05

Tonnes

7e+05

5e+05

3e+05

1970 1975 1980 1985 1990 1995 2000 2005 Year

Fig. 1. Black Sea landings from 1970 to 2005. The continuous line represents the fit of the catch and the dotted lines represent the 95% confidence bands. Fig. 1. Capturas en el Mar Negro desde 1970 hasta 2005. La linea contínua representa el ajuste de la tendencia de las capturas, mientras las lineas de puntos representan las bandas de confianza del 95%.

maintenance, reproduction and other activities of the animals in the systems (Pauly & Christensen, 1995). Thus, notwithstanding our preference for catching and consuming large predators, deliberately fishing down should enable more of an ecosystem’s biological production to be captured by fishing. However, to avoid waste here as well, any decline in the mean TL of the fisheries catches should, in this case, be matched by an ecologically appropriate increase in these catches, the appropriateness of that increase being determined by the transfer efficiency (TE) between TL. The average transfer efficiency between trophic levels in marine systems is c. 10% (Pauly & Christensen, 1995), Pauly et al. (2000b) predicted that a fall of 1 in the level at which a fishery operates would lead to a 10–fold increase in potential catches. To study this effect Pauly et al. (2000b) and Christensen (2000) introduced the fishing in balance (FiB) index as following: FiBk = log [Yk · (1/TE)MTIk] – log [Y0 (1/TE)

MTL

]

where Y corresponds to landings in year k, TL is the mean TL of the landings in year k, TE is the transfer efficiency (here set at 0.1 following Pauly et al., 2000a), and 0 refers to any year used as a baseline to normalize the index (Pauly et al., 2000; Christensen, 2000; Cury et al., 2005). This index computes whether the increase in landings due to focusing on lower TL matches the ecological appropriate increase (determined by the transfer efficiencies between TL’s). The FiB index remains constant if the TL–changes match ‘ecological appropriate’ changes in landings. When the FiB index

decrease this may indicate that fisheries withdraw so much biomass from the ecosystem that its functioning is impaired (Pauly & Watson, 2005). A decrease in FiB will also be observed if discarding that is not reflected in the reported catches takes place (Pauly & Watson, 2005). FiB requires the assumption that transfer efficiency is constant (and known sufficiently well) across trophic levels (Pauly et al., 2000c). Nevertheless, FiB is believed to provide a better indicator of ecosystem change than catch or catch composition, because of its integrative nature (Garcia & Staples, 2000). Pelagic/demersal index – P/D Changes in the trophic composition of marine communities can be tested in terms of large trophic groups such as planktivorous, benthivorous, or piscivorous animals (Caddy & Garibaldi, 2000). The expected effect of fishing (although not exclusive) is a decrease in the proportion of piscivorous fish. This is an easily understood indicator that can be estimated based on the knowledge of the biology of the species present in the community rather than on extensive diet data. A related index that has been proposed as an indicator for marine environments is the pelagic (P) to demersal (D) fish biomass ratio in fishery landings (Caddy, 2000; De Leiva Moreno et al., 2000). However, the P/D ratio in fisheries catches is not exclusive in that it might be an indicator of eutrophication rather than exploitation (De Leiva Moreno et al., 2000).The pelagic fish are positively influenced by nutrient enrichment when it stimulates the plankton production (Caddy, 1993), while the demersal fish are influenced by the dynamics of benthic community,

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Marine trophic index 3.40 3.55 3.30 3.25 3.20

1970 1975 1980 1985 1990 1995 2000 2005 Year

Fig. 2. Black Sea marine trophic index (MTI) between 1970 and 2005. Fig. 2. Indice trofico marino (MTI) del las capturas en el Mar Negro entre 1970 y 2005.

which generally responds negatively to the conditions of excessive enrichment. It follows that a positive trend over time in the P/D index may depend both on the eutrophication both from the overexploitation of resources (Libralato et al., 2004). In addition, like other catch–based indicators, it will be sensitive to changes in the fishing targets and methods.

Results Total reported landings in the Black Sea showed critical peaks and troughs, driven primarily by the fluctuation in the landings of European anchovy (Engraulis encrasicolus) with a peak landing of 790,000 tones recorded in 1984 (fig. 1). The landings increased

Fishing in balance 0.0

–0.5

–1.0

–1.5

1970 1975 1980 1985 1990 1995 2000 2005 Year

Fig. 3. Black Sea fishing in balance index (FiB) between 1970 and 2005. Fig. 3. Indice de pesca en equilibrio (FiB) del las capturas en el Mar Negro entre 1970 y 2005.

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Marine trophic index

3.76 3.74 3.72 3.70 3.68

1970 1975 1980 1985 1990 1995 2000 2005 Year

Fig. 4. Black Sea 3.25 marine trophic index (3.25MTI) from 1970 to 2005. The cut–off value of 3.25 removed from the computation of the index for all species whose biomass tend to vary widely in response to environmental factors. Fig. 4. Indice trofico marino rebajado de 3,25 (3.25MTI) del las capturas en el Mar Negro desde 1970 hasta 2005. El corte del nivel trofico de 3,25 elimina del calculo del indice todas las especies cuya biomasa varia mucho en funcion de los factores ambientales.

following a precipitous decline from 1989 to 1991. However, they have not returned to the level achieved in the mid 1980s. MTI showed an increase of 0.2 in the first two decades. In fact the values of MTI grew from 3.22 to 3.42 from 1970 to 1990 (fig. 2). In contrast, from 1990 to 2000 the MTI index showed an abrupt decline, with a decrement of 0.22 (fig. 2). Only in the last five years of the time series did the MTI index show a slight increase from a value of 3.20 to 3.25 (fig. 2). The FiB index showed negative values in all 35 years of the series (fig. 3). The increase in the FiB index from the 1970s to the mid–1980s was driven by the increased reported landings during this period (fig. 3). In contrast, the decrease in the MTI values since 1990 was not countered by an increase in landings; thus the FiB index also declined in the early 1990 (fig. 3). After reaching the minimum value of –1.8, the FIB index increased by 0.5 from 1995 to 2005 (fig. 3). The cutMTI showed a drastic decline in 1984, the same year that the landings of European anchovy peaked (fig. ). In the mid–1990s the index increased from 3.69 to 3.77 in 1995 (fig. 4). In contrast, in the last five years the cutMTI index showed a strong decrease from 3.77 in 1995 to 3.69 in 2005 (fig. 4). The P/D index showed positive values for all the times series (fig. 5). This trend showed a decreasing trend from the early 1970s until the minimum values recorded in 1990 (fig. 5). In the following years the index started to increase slowly, showing a maximum

value of 9.7 in 2002 (fig. 5). In contrast, the last three years of the time series showed a decrease in the index from 9.2 to 4 (fig. 5). Discussion The indices show a heavily exploited ecosystem where overfishing and the anthropogenic eutrophication are probably responsible (Daskalov, 2002). The decline of the MTI and of the FiB show a fishing down the marine food web situation in this ecosystem. Fishing down has been tested on Mediterranean large marine ecosystems (Pennino et al., 2011), but the speed with which changes occur in Black Sea fish communities is much higher. Our results support previous studies that show strong changes in the fish community of the Black Sea in recent years (Lleonart, 2005; Daskalov, 2002). The fishery eliminated the top predator during the 1970s (Daskalov et al., 2007); this led to reduced predation on planktivores, causing them to increase in the 1980s. Intense and unregulated fishing pressure in these years led to severe overexploitation of most major fish stocks (Black Sea Commission, 2002; UNEP, 2002). The MTI and FiB decreases may indicate that fisheries withdraw so much biomass from the ecosystem that its functioning could be impaired (Pauly & Watson, 2005). The technological growth and demand

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P/D index

1970 1975 1980 1985 1990 1995 2000 2005 Year

Fig. 5. Black Sea ratio of pelagic/demersal landings (P/D) from 1970 to 2005 (P/D index). Fig. 5. Relación entre las capturas pelagicas y demersales (índice P/D) en el Mar Negro desde 1970 hasta 2005.

of fisheries in the 1980s has led to over–exploitation of marine resources, particularly for highest trophic demersal resources. This suggests that the changes in the food web could be influenced by the impact of advances in fishing technologies and changes in market–driven exploitation (Caddy & Garibaldi, 2000; Stergiou, 2002). Indeed, the top predators, such as swordfish and tuna, were heavily exploited with the introduction of purse seining and through large scale surface longline and gill–net fisheries in the 1980s in this large marine ecosystem (Caddy, 1993). Some demersal species have pratically disappeared (Daskalov, 2002). This has been exacerbated by destructive fishing practices such as catching of under–sized fish (UNEP, 2002). Furthermore, the invasion of the Mnemiopsis leidyi contributed to a catastrophic decline of fish stocks in the mid–1980s (Shiganova, 1998). The trend of the cutMTI and P/D index confirmed this shift in the mid–1980s. The M. leidyi eats eggs and larvae of pelagic fish (Kube et al., 2007) and it caused a dramatic drop in fish populations. The trend of the P/D index showed that the dramatic fall of the black sea fish catch was most pronounced for small pelagic species with a four–fold reduction in catches between 1988 and 1991, although the landings of these species have partially recovered over the past decade, as shown by the trend of the MTI. The P/D index shows this increasing trend in recent years for the small pelagic species, which are r species with a higher turnover, especially in a highly eutrophic ecosystem such as the Black Sea. Also, this

increase in small planktivorous species might have been a result of the transition of the ecosystem from an oligotrophic to eutrophic stae caused by nutrient enrichment (Caddy, 1993). This phenomenon is confirmed by the decrease in recent years of the cutMTI index. This trend indicates that the resources with high TL, which correspond to the demersal fishes, are running out while the low TL species are increasing. Low TL species are usually small pelagic species. The interpretation of trophic indicators is still very subjective. References points or limit values with which to unambiguously assess the results obtained with these indicators have not yet been established. These values cannot be established and standardised due to the sheer complexity of ecosystems. The tropic structure of each ecosystem is specific and unique. However, we consider that jointly these indices are a good way to compare the dynamics of different ecosystems, since the important aspect for study of ecosystems is the trends in indices over times rather than the values they assume. Sustainability, however defined, must imply some notion of permanence in at least some of the entities or processes being evaluated. Thus, if, in a given ecosystem, there is a clear trend of the relative abundance of high–TL to low–TL fishes, as indicated by declining MTI values, then this indicates the absence of sustainability and the need for intervention. Multispecies fishery can safely be assumed to be unsustainable if the mean TL of the species it exploits keeps going down (Pauly & Watson, 2005).

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Furthermore, the indicators used are easily measurable and provide a clear understanding of complex processes that occur in an ecosystem. Like many other indicators, they do not detect changes induced by exclusive fishing, but they are, however, very sensitive to the dynamics of the fishes community. In conclusion we recommend the use of these indicators to analyze an ecosystem with a macro–scale approach and to obtain an overview of the ecosystem as a whole. The causes and the drivers that led to the changes highlighted by the analysis must be verified at the micro–scale, when and where better data are available. We consider that ecosystem indicators are promising tools to assess ecosystem conditions because they are easy to standardize and to estimate with commonly available data. Application of the selected indicators to other marine ecosystems is encouraged so as to fully evaluate their usefulness for a broad selection of large marine ecosystems (LMEs), to assess their usefulness for an ecosystem approach to management and to establish international comparability. Acknowledgements David Conesa and Antonio López–Quílez would like to acknowledge financial support from the Ministerio de Educación y Ciencia (the Spanish Ministry of Education and Science) via research grants MTM2007– 61554 (jointly financed with the European Regional Development Fund). The authors are grateful to the reviewers for their truly helpful comments. References Bax, N. J., 1998. The significance and prediction of predation in marine fisheries. ICES Journal of Marine Science, 55: 997–1030. Balkas, T., Decheo, G., Mihnes, R., Serbanescu, O. & Unluata, U., 1990. Review of the State of the Marine Environmente of the Black Sea. United Nations Environment Programme, Regional Seas Report and Studies, 124: 201–218. Black Sea Commission, 2002. State of the Environment of the Black Sea–Pressures and Trends 1996–2000. Istanbul, Turkey. Caddy, J., 1993. Toward a comparative evaluation of human impacts on fisheries ecosystems of enclosed and semi enclosed seas. Reviews Fisheries Science, 1: 57–95. Caddy, J. F., 2000. Marine Catchment Basin effects versus impacts of fisheries on semi–enclosed seas. ICES Journal Marine Science, 57: 628–640. Caddy, J. F., Csirke, J., Garcia, S. M. & Grainger, R. J. R., 1998. How pervasive is 'Fishing down marine food webs?' Science (Washington, D.C.), 282: 1383. Caddy, J. F. & Garibaldi, L., 2000. Apparent changes in the trophic composition of world marine haverests: the perspective from FAO capture database. Ocean Coastal Management. Fisheries Research, 72: 241–252. CBD, 2004. Annex I, decision VII/30. The 2020

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Editor executiu / Editor ejecutivo / Executive Editor Joan Carles Senar

Secretaria de Redacció / Secretaría de Redacción / Editorial Office

Secretària de Redacció / Secretaria de Redacción / Managing Editor Montserrat Ferrer

Museu de Zoologia Passeig Picasso s/n 08003 Barcelona, Spain Tel. +34–93–3196912 Fax +34–93–3104999 E–mail mzbpubli@intercom.es

Consell Assessor / Consejo asesor / Advisory Board Oleguer Escolà Eulàlia Garcia Anna Omedes Josep Piqué Francesc Uribe

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Population structure of Atlantic swordfish (Xiphias gladius L. 1758) (Teleostea, Xiphiidae) using mitochondrial DNA analysis: implications for fisheries management

A. Garcia, S. Cecchetti, M. N. Santos, S. Mattiucci, G. Nascetti & R. Cimmaruta Garcia, A. Cecchetti, S., Santos, M. N., Mattiucci, S., Nascetti, G. & Cimmaruta, R., 2011. Population structure of Atlantic swordfish (Xiphias gladius L. 1758) (Teleostea, Xiphiidae) using mitochondrial DNA analysis: implications for fisheries management. Animal Biodiversity and Conservation, 34.1: 133–140. Abstract Population structure of Atlantic swordfish (Xiphias gladius L. 1758) (Teleostea, Xiphiidae) using mitochondrial DNA analysis: implications for fisheries management.— Recent studies on Atlantic swordfish (Xiphias gladius L. 1758) genetic structure have demonstrated significant heterogeneity but the precise boundary between populations remains to be identified. In this context, genetic diversity was investigated by PCR–RFLP analysis at the control region of mitochondrial DNA (D–loop) from 274 swordfish specimens collected from five different areas of the Atlantic Ocean. The analysis of molecular variance (AMOVA) showed that genetic variation was mainly due to differences within rather than between the studied areas. Additionally, the phylogenetic analysis did not show evident relationships among haplotypes from all areas. However, low but significant FST values were recorded when comparing Equatorial samples with those from the north central and north tropical Atlantic. These results do not support a need for changing the current management boundary for the Atlantic fishery. Key words: Xiphiidae, Swordfish, Xiphias gladius, Mitochondrial DNA, Genetic variability, Atlantic Ocean. Resumen Estructura poblacional del pez espada del Atlántico (Xiphias gladius L. 1758) (Teleostea, Xiphiidae) usando análisis de ADN mitocondrial: implicaciones para la gestión de pesquerías.— Estudios recientes sobre la estructura genética del pez espada del Atlántico (Xiphias gladius L. 1758) han demostrado una heterogeneidad significativa, pero los límites precisos entre poblaciones no han sido identificados. En este contexto, la diversidad genética se ha investigado mediante análisis PCR–RFLP en la región control de ADN mitocondrial (bucle D) de 274 peces espada recolectados en cinco zonas diferentes del océano Atlántico. El análisis de la varianza molecular (AMOVA) mostró que la variación genética se debía a diferencias en cada zona y no entre las zonas estudiadas. Además, los análisis filogenéticos no muestran relaciones evidentes entre los haplotipos de todas las zonas. A pesar de ello, al comparar las muestras ecuatoriales con las de zonas más al norte, se obtienen valores de FST bajos pero significativos. Estos resultados indican que no es necesario cambiar los límites de las zonas de gestión para la pesquería del Atlántico. Palabras clave: Xiphiidae, Pez espada, Xiphias gladius, ADN mitocondrial, Variabilidad genética, Océano Atlántico. A. Garcia & M. N. Santos, INRB, I.P./IPIMAR, Av. 5 de Outubro s/n., 8700-305 Olhão, Portugal.– S. Cecchetti, G. Nascetti & R. Cimmaruta, Dept. of Ecology and Sustainable Economic Development (DECOS), Tuscia–Univ. of Viterbo, Largo dell’Università s/n., 01100 Viterbo, Italy.– S. Mattiucci, Dept. of Public Health Sciences, Section of Parasitology, Sapienza–Univ. of Rome, P.le Aldo Moro, 5, 00185 Rome, Italy. Corresponding author: A. Garcia. E–mail: agarcia@ipimar.pt

ISSN: 1578–665X

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Introduction The swordfish (Xiphias gladius L. 1758) is a pelagic and highly migratory species inhabiting tropical and temperate waters in a geographical range extending from 45º N to 45º S (Nakamura, 1985). Up to two decades ago, the oceanic environment was considered free from pronounced barriers, suggesting a lack of population genetic structure in species with high migratory ability, such as swordfish (Waples, 1998). However, more recent studies on the genetic structure of swordfish populations (i.e. biological stocks) using both mitochondrial (mtDNA) and nuclear DNA (nDNA) have shown significant inter–oceanic differentiation of Atlantic, Indo–Pacific and Mediterranean populations (Alvarado Bremer et al., 1995, 1996, 2005; Kotoulas et al., 1995; Rosel & Block, 1996; Chow et al., 1997; Chow & Takeyama, 2000; Greig et al., 2000; Jean et al., 2006; Muths et al., 2009). Moreover, a significant northwest versus south Atlantic genetic distinction has been detected as sampling efforts increased both using mtDNA (Alvarado Bremer et al., 1996, 2005; Chow et al., 1997) and nDNA markers (Greig et al., 1999, 2000). In particular, significant heterogeneity in mtDNA diversity was found between the northwest and south Atlantic regions as well as by means of both nuclear loci aldolase B (aldB) and lactate dehydrogenase A (IdhA) (Greig et al., 1999, 2000). Also, a large survey based on nDNA calmodulin gene intron 4 (CaM) showed a sharp differentiation between northwest Atlantic swordfish and those from southern regions (Chow & Takeyama, 2000; Chow et al., 2007). Despite the considerable advances in our understanding of Atlantic swordfish population structure that suggest the existence of genetic heterogeneity within the Atlantic, the data so far acquired still leave two questions unanswered. The first concerns the genetic homogeneity within the North Atlantic (West versus East), which has not yet been thoroughly assessed because of the limited sampling effort in the Eastern region. As a further consequence, the degree of differentiation among swordfish from the South and NE Atlantic has been poorly evaluated. The second question regards the geographic boundary and the possible mixing of North and South Atlantic populations. ICCAT sets the management boundary between the Atlantic populations at 5° N but scientists and managers are still questioning the validity of this hypothesis (Chow et al., 2007; Viñas et al., 2007; Smith & Alvarado Bremer, 2010). This is due to the fact that the genetic studies carried out so far share several limitations concerning the sampling scheme or the marker used (Alvarado Bremer et al., 2006). For example, Kasapidis et al. (2007) increased the sampling effort in the North Atlantic region and, by means of microsatellites, suggested a north–south reduction in gene flow within the Atlantic but they recommended further studies using a higher number of microsatellites. More recently, Smith & Alvarado Bremer (2010) applied different nuclear markers (SNPs and RFLPs of nuclear genes CaM, ARP, Mlc2, ActA2) to assess the extent of population admixture along the north–south boundary: the markers were

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promising but the low number of specimens assayed made the results preliminary. The present study deals with this latter question, concerning the location and the extent of admixture of Northwestern and Southern Atlantic swordfish stocks. Management implications regarding this question are obvious since the Atlantic swordfish is subject to intense harvest and, to date, it is managed as two stocks that are regulated by quota assignments. Thus, unequivocal data supporting a sustainable management of the Atlantic swordfish are urgently needed as this is widely acknowledged as crucial for their conservation and management (Hilborn et al., 2003). To this end, the sampling effort in the Eastern Atlantic region and within the boundary location between the Atlantic populations were increased in order to give a more comprehensive analysis of the Atlantic swordfish stock structure trying to fill the gaps that so far exist. For this purpose, the highly polymorphic control region of the mtDNA (D–loop) was used to characterise the genetic diversity within and among swordfishes from different Atlantic areas and to examine their level of genetic divergence through: i) investigating the population structure of North and Central Atlantic swordfish; and ii) testing the heterogeneity among swordfish specimens collected from distinct geographic areas. Material and methods Sampling Muscle tissue samples were collected from 274 swordfish specimens onboard commercial long line fishing vessels between May 2003 and November 2007 in five different areas throughout the Atlantic Ocean, namely: Northwest (NW, N = 39, July 2006), North Central (NC, N = 53, May 2003), Northeast (NE, N = 51, May 2003), Northern Tropical (TR, N = 48, June–August 2007) and Equatorial (EQ, N = 83, June–November 2007) (fig. 1). A small portion of muscle tissue was taken from each specimen using a sterilized scalpel, stored in an Eppendorf tube and kept frozen at –80ºC until assayed. DNA extraction, PCR amplification and RFLP analysis DNA extraction was based on the CTAB (Cetyltrimethyl Ammonium Bromide) method slightly adapted from Murray & Thompson’s (1980) protocol. The dried DNA pellet was re–suspended in 100 μl of TE–buffer (EDTA 1 mM and tris 10 mM, pH 8). DNA was then diluted (1:50) with distilled water and so suitable for polymerase chain reaction (PCR). The DNA amplification was achieved using the primers specifically designed for the control region of swordfish (Alvarado Bremer et al., 1995). The L–strand primer L15998 (5'–TAC CCC AAA CTC CCA AAG CTA–3') was used in combination with the H–strand primer H235 (5'–CGT GTG CAC TCT GAA ATG TCA–3') to amplify a DNA fragment composed by c. 525 bp using a Perkin Elmer thermal cycler apparatus (GeneAmp 2400). The PCR reaction was then cycled

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50 40

Latitude (º)

10 EQ 0

–10 –20 –30 –40 –70 –60 –50 –40 –30 –20 –10 Longitude (º)

Fig. 1. Swordfish sampling areas in the Atlantic Ocean: Northwest (NW), North Central (NC), Northeast (NE), Northern Tropical (TR) and Equatorial (EQ). Fig. 1. Áreas de muestreo del pez espada en el océano Atlántico: noroeste (NW), norte central (NC), nordeste (NE), norte tropical (TR) y ecuatorial (EQ).

for 4 min at 94ºC followed by 35 cycles of 1 sec. at 94ºC, 1 sec at 50ºC, 1 sec at 72ºC and the final cycle at 72ºC for 10 sec. Negative controls without DNA template were prepared in every series of amplification to exclude the possibility of contamination of reagents or reaction buffers. Five restriction enzymes (Alu I, Dra I, Vsp I, Hpa II and Dde I) were used to digest the amplified fragments. Digests were performed in 7.5 μl final reaction volume. After centrifuging, the reaction mixture digestions were incubated for 3 h at 37ºC for all the enzymes used in the study. The resulting restricted fragments were analysed by electrophoresis on 3.0% agarose gel using TBE–buffer (0.045M tris–borate; 0.001M EDTA, pH 8), stained with 0.01% ethidium bromide and finally photographed. A molecular weight marker and a reference PCR product was run along with the digested PCR products to estimate the sizes of the resulting mtDNA fragments. Restriction patterns generated from each restriction endonuclease were labelled with letters (A, B, C and D) indicating variant digestion patterns (table 1). Composite mtDNA haplotypes were constructed from all the enzymes used and arranged in the following order: Alu I, Dra I, Vsp I, Hpa II and Dde I. Thus, each fish was assigned a code of five letters that described its composition in terms of multi–enzyme haplotype.

Genetic data analysis Chromatographic curves of forward and reverse sequences were edited in Chromas v.1.6. All sequences were then aligned by eye using Clustal X v.1.83 (Thompson et al., 1997). The amount of sequence divergence for each geographical population was assessed by estimating the number of polymorphic sites (S), haplotype diversity (h; Nei, 1987), nucleotide diversity (π, Nei, 1987) and average number of pairwise nucleotide differences (k; Tajima, 1983) all performed using Arlequin v.3.1 (Excoffier et al., 2005). The level of genetic diversity within and among sampled areas was hierarchically evaluated by analysis of molecular variance (AMOVA, Excoffier et al., 1992). Significance of pairwise comparison was tested with 10,000 permutations. Samples were hierarchically divided into two groups (1: NW, NC, NE and TR; 2: EQ) to test the accuracy of the current management boundary set at 5º N (see fig. 1). Significance of the pairwise p–values achieved for the FST comparisons between the sampled areas was corrected applying a B–Y FDR method (Narum, 2006). As mentioned by Narum (2006), this method provides the most important critical value for evaluating significance of population differentiation in conservation genetics. Tajima (1989) D–test and Fu

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Table 1. Restriction patterns of the mtDNA D–loop haplotypes recorded in 274 swordfishes from the Atlantic Ocean (values show restriction fragments length in bp): RE. Restriction enzymes. Tabla 1. Patrones de restricción de los haplotipos del bucle D del ADN mitocondrial, registrados en 274 peces espada del océano Atlántico (los valores muestran la longitud de los fragmentos de restricción en pares de bases): RE. Enzimas de restricción. RE Alu I

Haplotype

Recognition sequence

AG–CT

525

273

177

241

252

348

177

TTT–AAA

254

227

294

298

500

206

Dra I

AT–TAAT

525

112

230

111

291

295

414

122

Vsp I

C–CGG

Hpa II

490

262

224

C–TNAG

Dde I

275 250

(1997) FS–test were used to test the deviation from neutral molecular evolution in relation to mtDNA sequences. Significance was assessed in both tests by generating random samples (1,000 simulated

525

39 192

525

274

155

252

samples) under the hypothesis of selective neutrality and population equilibrium. Both AMOVA and neutrality tests were performed with Arlequin v.3.1 (Excoffier et al., 2005).

Table 2. Genetic variability of Atlantic swordfish within the five sampled areas: SA. Sampling area; N. Sample size; H. Number of haplotypes; S. Number of polymorphic sites; h. Haplotype diversity; k. Mean pair–wise nucleotide differences; π. Nucleotide diversity. Tabla 2. Variabilidad genética del pez espada del Atlántico dentro de las cinco zonas muestreadas: SA. Área de muestreo; N. Tamaño de la muestra; H. Número de haplotipos; S. Número de sitios polimórficos; h. Diversidad del haplotipo; k. Media de las diferencias de los pares de nucleótidos; π. Diversidad nucleotídica. SA N H S h ± s.d. k ± s.d. π ± s.d. NW

Tajima’s D test (p–value)

Fu’s FS test (p–value)

58 0.852 ± 0.056 11.178 ± 5.187 0.021 ± 0.011 –0.742 (0.234) –1.988 (0.244)

63 0.755 ± 0.063 11.144 ± 5.142 0.021 ± 0.011 –0.702 (0.255) 0.700 (0.663)

65 0.868 ± 0.043 10.534 ± 4.880 0.020 ± 0.010 –0.914 (0.195) –1.877 (0.282)

61 0.781 ± 0.060 8.843 ± 4.150

77 0.935 ± 0.017 13.748 ± 6.236 0.026 ± 0.013 –0.394 (0.412) –6.050 (0.090)

0.017 ± 0.009 –1.288 (0.068) –1.456 (0.336)

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Table 3. Hierarchical AMOVA analysis on the Atlantic swordfish molecular data: Group 1 (NW, NC, NE and TR), and Group 2 (EQ): SV. Source of variation; VC. Variance components: P. Percentage of variation; Fi. Fixation indices. (* significant at 0.05 level). Tabla 3. Análisis AMOVA jerárquico sobre los datos moleculares del pez espada del Atlántico: Grupo 1 (NW, NC, NE y TR) y Grupo 2 (EQ): SV. Fuente de variación; VC. Componentes de la Varianza; P. Porcentaje de variación; Fi. Índices de fijación (* significativo al nivel 0,05). SV

Among groups 0.012 2.95

Fi FCT = 0.030

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Table 4. Pairwise F ST values among the five sampled swordfish areas in the Atlantic Ocean. FST values were calculated with 110 permutations (significant by B–Y FDR method). (For areas abbreviations, see figure 1) Tabla 4. Valores FST representados por parejas, entre las cinco áreas del océano Atlántico muestreadas para el estudio del pez espada. Los valores de FST se calcularon con 110 permutaciones (significativas mediante el método B–Y FDR). (Para las abreviaturas de las áreas, ver figura 1).

NW NC –0.00329

Among areas Within groups

0.000 0.02

FSC = 0.000

NE –0.00213 0.00714

Within areas

0.424 97.03 FST = 0.030*

TR –0.00183 –0.00484 0.00938 EQ 0.01689 0.04195* 0.01633 0.03336*

The phylogenetic relationships among haplotypes were graphically arranged in Mega v.4 (Tamura et al., 2007) with unrooted neighbour–joining dendrogram using the gamma corrected Tamura–Nei distance matrix. The statistical robustness of neighbour–joining distances was determined by 1,000 bootstrap replicates (Felsenstein, 1985). Results Molecular attributes A single fragment of approximately 525 bp was amplified from each specimen and no apparent size differences among them were observed. Restriction profiles obtained by each of the five endonucleases showed five patterns in Alu I, four in Dra I, three in Vsp I, three in Hpa II and four in Dde I (table 1). Most samples shared the most common restriction pattern for each endonuclease digestion, despite the difference in their frequencies. Pattern A in Alu I digestion was the most common in EQ, while pattern B prevailed in the remaining samples. Pattern A in Dra I, Vsp I, Hpa II and Dde I was the most frequent in all the sampled areas. A total of 68 composite haplotypes were recorded but only one (BAAAA) was distributed across the five areas with relatively high frequency (NW = 38.5%; NC = 49.1%; NE = 35.3%; TR = 45.8%; EQ = 20.5%) and 38 were limited to one area, namely: 8 in NW, 2 in NC, 6 in NE, 7 in TR and 15 in EQ. However, it must be noted that a larger number of individuals were sampled in the Equatorial region, which could influence the observed results. A total of

84 polymorphic nucleotide sites were observed, of which 19 were singleton variable sites and 65 were parsimony informative. The substitution bias favoured transitions over transversions, with their ratio being 10.9. The overall relative nucleotide frequencies were: C = 20.7%, T = 31.2%, A = 31.7% and G = 16.5%. Genetic variability was high for the pooled samples, displaying a value of 0.860 ± 0.019 for haplotype diversity (h) and 0.022 ± 0.011 for nucleotide diversity (π). The h ranged between 0.935 ± 0.017 in EQ and 0.755 ± 0.063 in NC, while π from 0.026 ± 0.013 (EQ) to 0.017 ± 0.009 (TR) (table 2). Population structure and phylogeny The overall hierarchical AMOVA showed that the greatest genetic differentiation (97.03%, p = 0.005) in the swordfish control region was found within distinct geographical areas, and a very small amount (0.02%, p = 0.354) was due to divergence among distinct geographical areas within groups (table 3). The amount of variance among groups was 2.95% and did not display statistically significant differences (p = 0.191). If a single AMOVA group was considered in the analysis, the results showed again that most variance was explained by differences within areas (98.10%, p = 0.001) rather than among them (1.90%). The highest pairwise FST values were found between EQ and NC (0.042, p < 0.05) and the lowest between TR and NW (–0.002, p > 0.05) (table 4). It is worth noting that statistically significant differences among pairwise FST values occurred between EQ vs. NC and EQ vs. TR (table 4).

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0.01 88

96 80

93 99

71 91

Northwest North central Northeast Northern tropical Equatorial

79 74 95

81 95

78 90

Fig. 2. Neighbor–joining tree of the 68 haplotypes from Atlantic swordfish. Bootstrap value support above 70% is shown by the branches. Symbols are related with the geographic origin of each haplotype. Fig. 2. Árbol por agrupación de vecinos de los 68 haplotipos del pez espada del Atlántico. El soporte del valor bootstrap por encima del 70% se muestra mediante las ramas. Los símbolos están relacionados con el origen geográfico de cada haplotipo.

The indices of neutral evolution (Tajima’s D and Fu’s FS tests) yielded moderately low negative values in the large part of the sampled areas, but failed to detect statistically significant differences (table 2). The topology of the gene tree obtained using neighbour–joining analysis showed that there is no evident phylogeographic relationship, as they are randomly allocated across the tree diagram (fig. 2). No apparent latitudinal gradient of the proportion of each haplotype was observed, excepting the most common haplotype (BAAAA) that decreased in the Equatorial area. Indeed, statistically significant differences (x2–test: p < 0.05) were observed when comparing EQ with NC and TR areas. Moreover, only 19 branches received values higher than 70%. Discussion The mitochondrial control region sequence of swordfish showed moderately high levels of variation, with

16% of the nucleotide positions being polymorphic. However, the substitution bias ratio (r = 10.9) herein achieved for swordfish was higher than those recorded by Alvarado Bremer et al. (1997) in albacore (r = 9.0) mitochondrial genome. All the samples were characterized by high levels of haplotype diversity, with 56% of the haplotypes as private of a single area. Such a pattern has been previously reported for other scombrid species, such as albacore (Chow & Ushiama, 1995; Viñas et al., 2004). The nucleotide diversity found was within the range reported for Atlantic swordfish by Alvarado Bremer et al. (2005), but lower than the values found for other highly migratory species, such as albacore (Viñas et al., 2004) bigeye tuna (Martinez et al., 2006) and bluefin tuna (Carlsson et al., 2004). High genetic diversity within geographic areas and low genetic differentiation among areas within the same ocean basin are commonly observed in large pelagic marine fishes and could be explained by their wide distribution range that favour gene flow, and their large population sizes (Avise, 1998).

Animal Biodiversity and Conservation 34.1 (2011)

In the present study, no appreciable variance was attributable to variation among groups (north versus south, divided at 5º N), despite the significant heterogeneity found within the sampling areas. The lack of evident structure among the North Atlantic samples (NW, NC, NE and TR) was also confirmed by the moderately low FST statistics, which pointed out that the genetic exchange rate between them is sufficient to prevent genetic divergence. However, it is worth noting that FST values were significantly different when the Equatorial samples (EQ) were compared with NC and TR samples. This observation could reinforce the findings that the Equatorial area (between 5º N and 10º S) may represent a zone of intergradation within the Atlantic Ocean. Such results may also be supported by the finding that the higher levels of genetic divergence were recorded exactly within the Equatorial samples. However, these results must be interpreted with caution due to the higher sample size in the Equatorial area that could influence the higher number of haplotypes recovered there. Moreover, the reduced sampling effort in the South Atlantic area probably did not allow the detection of north–south stock differences. Therefore, the present results do not support a need for changing the current management boundary at 5º N. The absence of an evident genetic differentiation among the NW, NC, NE and TR samples did not exclude the two–stock hypothesis (east to west) considered for managing the Atlantic swordfish fisheries (Miyake & Rey, 1989), as the sample coverage was limited to the East of 40ºW. Kasapidis et al. (2007) also failed to detect a east–west difference. However, as in the present study, the west area was limited to longitude 47º W. Moreover, for the same reason, the possible subdivision between NW Atlantic (west of 40º W) and South Atlantic swordfish could not be excluded. In fact, such a division was strongly supported by concordant results for both mtDNA and scnDNA data (Alvarado Bremer et al., 2005). Furthermore, the extent of mixing between Atlantic swordfish samples should cover a large area southern to the current boundary location, but the present study does not suggest evidence for stock mixing north of 5º N. Indeed, the results presented here did not reveal significant heterogeneity in mtDNA diversity between the samples collected more southern and northern than 30º N, as suggested by Chow & Takeyama (2000) analysing CaM locus. Kasapidis et al. (2007) also failed to detect genetic divergences among northern and mid–Atlantic areas. At present, the main swordfish fisheries in the Atlantic Ocean are managed as two different units, North and South. The management unit definition recognizes genetically structured populations connected by limited gene flow. The results achieved with the present study did not provide evidence to change the current management boundary for the Atlantic swordfish fishery, as slight genetic structuring was observed only for the Equatorial area. Further examination of a larger number of samples from the Southern Atlantic is needed to confirm and possibly quantify the extent of genetic differentiation reported herein.

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Acknowledgements The authors express their gratitude to PESCARADE and to the crews of the fishing vessels 'Príncipe das Marés' and 'Paula Filipa' for their assistance in swordfish sampling. The first author holds a Ph D grant from the Portuguese Foundation for Science and Technology (SFRH/BD/25391/2005) in co–operation with the University of Rome 'La Sapienza' and University of Viterbo 'Tuscia' in Italy. References Alvarado Bremer, J. R., Hinton, M. G. & Greig, T. W., 2006. Evidence of spatial genetic heterogeneity in Pacific swordfish (Xiphias gladius) revealed by the analysis of LDH–A sequences. Bulletin of Marine Science, 79(3): 493–503. Alvarado Bremer, J. R., Mejuto, J. & Baker, A. J., 1995. Mitochondrial DNA control region sequences indicate extensive mixing of swordfish (Xiphias gladius L.) populations in the Atlantic Ocean. Canadian Journal Fisheries Aquatic Sciences, 52: 1720–1732. Alvarado Bremer, J. R., Mejuto, J., Gómez–Márquez, J., Boán, F., Carpintero, P., Rodríguez, J. M., Viñas, J., Greig, T. W. & Ely, B., 2005. Hierarchical analyses of genetic variation of samples from breeding and feeding grounds confirm the genetic partitioning of northwest Atlantic and South Atlantic populations of swordfish (Xiphias gladius L.). Journal of Experimental Marine Biology and Ecology, 327: 167–182. – 1996. Global population structure of the swordfish (Xiphias gladius L.) as revealed by the analysis of the mitochondrial DNA control region. Journal of Experimental Marine Biology and Ecology, 197: 295–310. Alvarado Bremer, J. R., Naseri, I. & Ely, B., 1997. Orthodox and unorthodox phylogenetic relationships among tunas revealed by the nucleotide sequence analysis of the mitochondrial control region. Journal of Fish Biology, 50: 540–554. Avise, J. C., 1998. The history and purview of phylogeography: a personal reflection. Molecular Ecology, 7: 371–379. Carlsson, J., McDowell, J. R., Díaz–Jaimes, P., Carlsson, J. E. L., Boles, B. B., Gold, J. R. & Graves, J. E., 2004. Microsatellite and mitochondrial DNA analyses of Atlantic bluefin tuna (Thunnus thynnus thynnus) population structure in the Mediterranean Sea. Molecular Ecology, 13: 3345–3356. Chow, S., Clarke, S., Nakadate, M. & Okazaki, M., 2007. Boundary between the north and south Atlantic populations of the swordfish (Xiphias gladius) inferred by a single nucleotide polymorphism at calmodulin gene intron. Marine Biology, 152: 87–93. Chow, S., Okamoto, H., Uozumi, Y., Takeuchi, Y. & Takeyama, H., 1997. Genetic stock structure of the swordfish (Xiphias gladius) inferred by PCR–RFLP analysis of the mitochondrial DNA control region. Marine Biology, 127: 359–367.

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Chow, S. & Takeyama, H., 2000. Nuclear and mitochondrial DNA analyses reveal four genetically separated breeding units of the swordfish. Journal of Fish Biology, 56: 1087–1098. Chow, S. & Ushiama, H., 1995. Global population structure of albacore (Thunnus alalunga) inferred by RFLP analysis of the mitochondrial ATPase gene. Marine Biology, 123: 39–45. Excoffier, L., Laval, G. & Schneider, S., 2005. Arlequin ver. 3.1: an integrated software package for population genetics data analysis. Evolutionary Bioinformatics Online, 1: 47–50. Excoffier, L., Smouse, P. E. & Quattro, J. M., 1992. Analysis of molecular variance inferred from metric distances among DNA haplotypes: application to human mitochondrial DNA restriction data. Genetics, 131: 479–491. Felsenstein, J., 1985. Confidence limits on phylogenies: an approach using the bootstrap. Evolution, 39: 783–791. Fu, Y. X., 1997. Statistical tests of neutrality of mutations against population growth, hitchhiking and background selection. Genetics, 147: 915–925. Greig, T. W., Alvarado Bremer, J. R. & Ely, B., 1999. Preliminary results from genetic analyses of nuclear markers in swordfish, Xiphias gladius, reveals concordance with mitochondrial DNA analyses. Collective Volume Scientific Papers ICCAT, 49(1): 476–482. – 2000. Nuclear markers provide additional evidence for population subdivision among Atlantic swordfish. Collective Volume Scientific Papers. ICCAT, 51: 1637–1641. Hilborn, R., Quinn, T. P., Schindler, D. E. & Rogers, D. E., 2003. Biocomplexity and fisheries sustainability. Proceedings of the National Academy of Sciences of the USA, 100: 6564–6568. Jean, C., Bourjea, J., Jouen, E. & Taquet, M., 2006. Stock structure of the swordfish (Xiphias gladius) in the southwest Indian Ocean: A preliminary study. Bulletin of Marine Science, 79(3): 521–526. Kasapidis, P., Mejuto, J., Tserpes, G., Antoniou, A., Garcia–Cortes, B., Peristeraki, P., Oikonomaki, K., Kotoulas, G. & Magoulas, A., 2007. Genetic structure of the swordfish (Xiphias gladius) stocks in the Atlantic using microsatellite DNA analysis. Collective Volume Scientific Papers ICCAT, 61: 89–98. Kotoulas, G., Magoulas, A., Tsimenides, N. & Zouros, E., 1995. Marked mitochondrial–DNA differences between Mediterranean and Atlantic populations of the swordfish, Xiphias gladius. Molecular Ecology, 4: 473–481. Martinez, P., González, E.G., Castilho, R. & Zardoya, R., 2006. Genetic diversity and historical demography of Atlantic bigeye tuna (Thunnus obesus). Molecular Phylogenetics and Evolution, 39: 404–416. Miyake, P. M. & Rey, J. C., 1989. Status of Atlantic broadbill swordfish stocks. In: Planning the future of billfish, research and management in the 90's and

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beyond (R. H. Stroud, Ed.). Proceedings 2nd Int. Billfish Symp., Kailua–Kona, Hawaii, 1–5 August 1988. National Coalition for Marine Conservation, Inc., Savannah, Ga. Murray, M. G. & Thompson, W. F., 1980. Rapid isolation of high molecular weight plant DNA. Nucleic Acids Research, 8: 4321–4325. Muths, D., Grewe, P., Jean, C. & Bourjea, J., 2009. Genetic population structure of the swordfish (Xiphias gladius) in the southwest Indian Ocean: Sex–biased differentiation, congruency between markers and its incidence in a way of stock assessment. Fisheries Research, 97: 263–269. Nakamura, I., 1985. Billfishes of the world. An annotated and illustrated catalogue of marlins, sailfishes, spearfishes and swordfishes known to date. In: FAO species catalogue, vol. 5. FAO Fisheries Synopsis 125. FAO, Rome. Narum, S., 2006. Beyond Bonferroni: Less conservative analyses for conservation genetics. Conservation Genetics, 7: 783–787. Nei, M., 1987. Molecular Evolutionary Genetics. Columbia Univ. Press, New York, NY, USA. Rosel, P. E. & Block, B. A., 1996. Mitochondrial control region variability and global population structure in the swordfish, Xiphias gladius. Marine Biology, 125: 11–22. Smith, B. L. & Alvarado–Bremer, J. R., 2010. Inferring population admixture with multiple nuclear genetic markers and bayesian genetic clustering in Atlantic swordfish (Xiphias gladius). Collective Volume Scientific Papers ICCAT, 65(1): 185–190. Tajima, F., 1983. Evolutionary relationship of DNA sequences in finite populations. Genetics, 105: 437–460. – 1989. Statistical methods to test for nucleotide mutation hypothesis by DNA polymorphism. Genetics, 123: 585–595. Tamura, K., Dudley, J., Nei, M. & Kumar, S., 2007. MEGA4: Molecular Evolutionary Genetics Analysis (MEGA) software version 4.0. Molecular Biology and Evolution, 24: 1596–1599. Thompson, J. D., Gibson, T. J., Plewniak, F., Jeanmougin, F. & Higgins, D. G., 1997. The CLUSTAL_X windows interface: flexible strategies for multiple sequence alignment aided by quality analysis tools. Nucleic Acids Research, 25: 4876–4882. Viñas, J., Alvarado Bremer, J. R., Mejuto, J., de la Serna, J. M., Garcia–Cortez, B. & Pla, C., 2007. Swordfish genetic population structure in the North Atlantic and Mediterranean. Collective Volume Scientific Papers ICCAT, 6(11): 99–106. Viñas, J., Alvarado Bremer, J. R. & Pla, C., 2004. Inter–oceanic genetic differentiation among albacore (Thunnus alalunga) populations. Marine Biology, 145: 225–232. Waples, R. S., 1998. Separating wheat from the chaff: patterns of genetic differentiation in high gene flow species. Journal of Heredity, 89: 438–450.

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Ecological features of Terebellida fauna (Annelida, Polychaeta) from Ensenada de San Simón (NW Spain) E. Cacabelos, J. Moreira, A. Lourido & J. S. Troncoso

Cacabelos, E., Moreira, J., Lourido, A. & Troncoso, J. S., 2011. Ecological features of Terebellida fauna (Annelida, Polychaeta) from Ensenada de San Simón (NW Spain). Animal Biodiversity and Conservation, 34.1: 141–150. Abstract Ecological features of Terebellida fauna (Annelida, Polychaeta) from Ensenada de San Simón (NW Spain).— Ecological features of Terebellida (Annelida, Polychaeta) inhabiting the intertidal and subtidal soft–bottoms of Ensenada de San Simón (NW Spain) were analysed by means of quantitative sampling. A total of 4,814 specimens belonging to five families (Ampharetidae, Pectinariidae, Terebellidae, Trichobranchidae and Sabellariidae) and ten species were collected in a variety of substrata and depths. Ampharetidae was the numerically dominant family mostly due to the abundance of Ampharete finmarchica and Melinna palmata; these species accounted for up to 94% of the total Terebellida abundance. Intertidal areas colonised by the seagrasses Zostera marina L. and Z. noltii Hornem. One thousand eight hundred and thirty–two harboured low densities of Terebellida, whereas the deeper subtidal muddy bottoms showed high abundances of ampharetids. Multivariate analyses suggested that Terebellida assemblages are highly correlated with sediment composition. Key words: Terebellida, Polychaeta, Biodiversity, Soft bottoms, Ensenada de San Simón, Atlantic Ocean. Resumen Características ecológicas de los Terebellida (Annelida, Polychaeta) de la Ensenada de San Simón (NO de España).— Las características ecológicas de los Terebellida (Annelida, Polychaeta) presentes en los fondos blandos intermareales y sublitorales de la Ensenada de San Simón (NW España) son analizadas por medio de muestreos cuantitativos. Un total of 4.814 individuos pertenecientes a cinco familias (Ampharetidae, Pectinariidae, Terebellidae, Trichobranchidae y Sabellariidae) y diez especies fueron recolectados en distintos sustratos y profundidades. Los Ampharetidae fueron la familia dominante en términos numéricos debido a la abundancia de Ampharete finmarchica y Melinna palmata; estas especies constituyeron hasta el 94% del total de los Terebellida. Las áreas intermareales estaban colonizadas por las fanerógamas Zostera marina L. y Z. noltii Hornem. Mil ochocientos treinta y dos presentaron bajas densidades de Terebellida; por el contrario, los fondos fangosos sublitorales más profundos mostraron una gran abundancia de anfarétidos. Los análisis multivariantes indicaron que las agrupaciones de Terebellida estaban altamente correlacionadas con la composición del sedimento. Palabras clave: Terebellida, Polychaeta, Biodiversidad, Fondos blandos, Ensenada de San Simón, Océano Atlántico. Eva Cacabelos, Antía Lourido & Jesús S. Troncoso, Área de Biología Animal, Fac. de Ciencias del Mar, Univ. de Vigo, 36310 Lagoas–Marcosende, Vigo, España (Spain).– Juan Moreira, Depto. de Biología (Zoología), Fac. de Ciencias, Univ. Autónoma de Madrid, Cantoblanco, 28049 Madrid, España (Spain). Corresponding author: Eva Cacabelos. E–mail: evacacabelos@yahoo.es

ISSN: 1578–665X

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Introduction Polychaetes play a major role in the functioning of benthic communities in terms of recycling and reworking of marine sediments (Hutchings, 1998). They are often the numerically dominant macrobenthic taxon in marine sediments (Jumars & Fauchald, 1977). Due to their high diversity of trophic behaviours and their great ability to adapt to different habitats (Fauchald & Jumars 1979), they are considered good indicators of community structure in benthic invertebrate assemblages (Olsgard et al., 2003). Moreover, some polychaetes are either sensitive or tolerant to a variety of perturbations and therefore have been regarded as indicators of marine environmental conditions (Grall & Glémarec, 1997; Tomassetti & Porrello, 2005). In the search for strategies and management plans to achieve sustainable use of protected areas, quick measures of their biodiversity are needed. In this sense, among the polychaetes, Terebellida species frequently represent an important component of the benthic habitat assemblages in terms of abundance and biomass, and they have been found to be a good indicator of polychaete species richness and main faunal patterns in biodiversity studies of marine benthic communities (Olsgard et al., 2003). The order Terebellida consists of large, long–lived and taxonomically well–defined polychaetes that occur at all depths and in all sediment types (Olsgard et al., 2003). They are usually tubicolous and surface deposit–feeders (Fauchald & Jumars, 1979), irrigating the substrate and actively transporting oxygen into their burrows, and therefore considerably affecting the sediments and their associated faunal communities (Fauchald & Jumars, 1979). Numerous faunistic and ecological works on the macrobenthic communities of the Galician coasts (NW Iberian peninsula), and specially their highly productive rias, have been carried out in recent years (Garmendia et al., 1998; Olabarria et al., 1998; Troncoso et al., 2005; Moreira et al., 2006). The soft–bottom polychaete faunas of the rias have been exhaustively studied (Parapar et al., 2000; Moreira et al., 2006; Cacabelos et al., 2008a; Lourido et al., 2008). Nevertheless, the role of terebellids is less understood. We here describe the diversity and assemblage structure of the order Terebellida inhabiting intertidal and subtidal soft substrata at Ensenada de San Simón (NW Iberian Peninsula), a Special Conservation Zone of the Nature 2000 Network. Furthermore, we investigated the main abiotic factors structuring the Terebellida populations at Ensenada de San Simón and tested whether the Terebellida is a useful indicator group to predict the species’ richness in soft bottoms of the Galician rias, as suggested by Olsgard et al. (2003) for the North Atlantic. Material and methods Ensenada de San Simón is located in the inner part of the Ría de Vigo, between 42º 17' and 42º 21' N and between 8º 37' and 8º 39' W (fig. 1).

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Intertidal and shallow subtidal areas of this inlet are colonised by Zostera noltii Hornem. 1832 and Z. marina L. meadows, and their soft bottoms are mainly muddy with high organic matter contents (Vilas et al., 1995). The inlet is subjected to large freshwater inputs, resulting in salinity fluctuations on both a tidal and seasonal basis (Nombela & Vilas, 1991). In addition, culture of mussels on rafts is a common practice in the inlet. Terebellida specimens were collected in Ensenada de San Simón during XI and XII 99. Twenty–nine sites were sampled (fig. 1) with a van Veen grab (0.056 m2; five replicates per site) and samples were sieved through a 0.5 mm mesh. The retained material was fixed in 10% buffered formalin, and fauna were sorted from the sediment and preserved in 70% ethanol for later identification. Temperature and pH were measured in situ both from the water and the sediment. Additional samples were taken at each site for later sediment analyses (calcium carbonate and total organic matter contents and grain–size analysis; see Cacabelos et al., 2008b for further details). We determined several cological indices (total abundance, number of species, Shannon–Wiener’s diversity index, Pielou’s evenness index and Soyer’s frequency index) depending on the presence and abundance of the species for each site. Assemblages were determined using non–parametric multivariate techniques (Plymouth Routines of the Multivariate Ecological Research software package, PRIMER; Clarke & Warwick, 1994), and SIMPER analysis was used to identify which species contributed most to dissimilarity among the groups of sites determined by classification and ordination analyses. Relationships between abundance of Terebellida and environmental variables were studied using Spearman’s non–parametric correlation coefficients and the BIOENV procedure (PRIMER package). Environmental variables expressed in percentages were previously transformed by log (x + 1) and all of them were normalised. Results Physical characteristics of the water and sediments of Ensenada de San Simón are shown in table 1. Subtidal soft bottoms were characterised by a predominance of muddy sediments (silt/clay fraction: 67.1% ± 5.4, mean ± SE) with a high total organic matter (17.7% ± 1.8) and low calcium carbonate content (6.8% ± 0.8). Sandy sediments were present in intertidal areas (silt/clay fraction: 37.7% ± 11.0), where the lower organic matter contents were found (12.9% ± 3.6). Polychaetes were the numerically dominant macrobenthic taxon in Ensenada de San Simón (Cacabelos et al., 2008b). A total of 4,814 specimens of Terebellida belonging to five families and ten species were identified (table 2). The inner intertidal area of the inlet, colonised by the seagrasses Zostera marina and Z. noltii, showed very low abundance or total absence (sites 1, 2, 4, 5, 6, 10, 15 and 29) of Terebellida. Most of the Terebellida were found in muddy subtidal

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143

Río Xunqueira

N Ría de Vigo

Pontesampaio Río Oitabén–Verdugo

Ensenada de San Simón

Iberian peninsula

Ría de Vigo

12 13 14

16 17 18 19 22

Soutoxusto

10 15

42º 10' N

Río

1 km

Redondela

Alve

dos

23 24 5 10 2115 28 26 27 29

8º 36' O

Fig. 1. Location of Ensenada de San Simón (Ría de Vigo) and position of the 29 sampling sites. Fig. 1. Localización de la Ensenada de San Simón (Ría de Vigo) y posición de las 29 estaciones de muestreo.

bottoms (99.6% of total abundance). The ampharetids Ampharete finmarchica and Melinna palmata were the numerically dominant species, representing 60.2% and 34.4% ,respectively, of total abundance, and reaching densities up to 2,114 and 990 ind./m2, respectively. The aforementioned species appeared distributed all over the subtidal part of the inlet, increasing their densities along the central channel: in sites 14, 17, 19, 21, 22, 23, 26 and 27, densities of both species considered together ranged from 1,161 to 2,711 ind./m2. According to Soyer’s frequency (F) index, only two out of ten species were characterized as Constant (F ≥ 50, namely M. palmata and A. finmarchica), three species as Common (25 < F < 49; Lagis koreni, Lanice conchilega and Terebellides stroemi) and the rest as Rare (F < 25). The number of Terebellida species in the inlet showed a direct relationship with the polychaete and the overall macrobenthic species richness (Cacabelos et al., 2008a; fig. 2). Ecological indices of sites at which Terebellida were found are shown in table 3. The highest densities were recorded at sites 22, 27, 26 and 14 (2,239.3 to 2,767.9 ind./m2), due to the high abundance of A. finmarchica and M. palmata. These sites, together with sites 19 and 21, showed the largest number of species (6–10). The Shannon–Wiener’s diversity index reached maximum values up to 1.5 in sites 9 and 25.

Dendrogram obtained through cluster analysis based on abundance data showed three main groups (fig. 3): Group A, composed of sites 11, 3 and 7, Group B (sites 28, 13, 24, 8, 9 and 25) and Group C (sites 12, 18, 22, 26, 27, 21, 23, 16, 17, 14 and 19). Site 20 appears clearly separated from the others due to the presence of only one specimen of Polycirrus sp.; Ordination of sites through MDS analysis confirmed the results of the dendrogram (stress: 0.04). The physical features of these assemblages are shown in table 4. Group A is poorly represented in terms of number of species; sites composing this group are located in the marginal part of the inlet, in shallow sediments subjected to strong variations of salinity close to the mouth of the river Oitabén–Verdugo and the small freshwater discharge near the western harbour (fig. 1). Sites from group B are muddy bottoms with low densities of species but high diversity indices due to the low dominances of L. koreni, M. palmata, A. finmarchica and L. conchilega. Finally, group C is characterized by deeper subtidal sites showing coarser sediments (table 4). The species that contributed most to the similarity and dissimilarity among groups of sites are listed in table 5. A. finmarchica strongly contributed to the similarity within the groups A, B and C, whereas L. koreni showed a high ratio coefficient in group B. A. finmarchica and M. palmata showed a

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Table 1. Parameters measured in water and sediment and bathymetry in sampling sites at Ensenada de San Simón (based in Cacabelos et al., 2008a): Tªw. Temperature of water (ºC); Q50. Median grain size (mm); Gravel, sand and silt/clay fractions in %; D. Depth (m); Tªs. Temperature of sediment (ºC); OM. Organic matter content (%); CO3. Calcium carbonate content (%). Tabla 1. Parámetros medidos en el agua y en el sedimento y batimetría en los lugares de muestreo en la Ensenada de San Simón (basados en Cacabelos et al., 2008a): Taw. Temperatura del agua (ºC); Q50. Mediana del tamaño de grano (mm); Fracciones de grava, arena y limo/arcilla en porcentajes; D. Profundidad (m); Tas. Temperatura del sedimento (ºC); OM. Contenido de materia orgánica (%); CO3. Contenido en carbonato cálcico (%). Site Tªw

Q50

Gravel

Sand

Silt/clay

Bottom type

Tªs

CO3

13.6

0.009

0.084

16.833

83.083

Mud

1.6

14.5

26.52

5.52

13.8

0.014

2.211

32.929

64.86

Mud

1.6

13.9

23.30

5.60

13.9

0.075

56.736

43.264

Sandy mud

1.6

13.8

19.05

6.12

14.5

0.320

17.793

73.985

8.222

Muddy sand

1.6

14.3

2.16

6.00

14.8

1250

29.959

64.363

5.678

Muddy sand

1.8

14.7

4.90

7.33

13.7

1150

21129

76826

2045

Very coarse sand 1.6

13.9

0.95

11.98

13.1

0.145

0.32

74.302

25.378

Sandy mud

3.4

12.1

3.95

6.31

13.1

0.040

0.631

35.894

63.475

Mud

3.2

10.1

10.88

5.80

12.9

0.010

0.871

27.674

71.455

Mud

2.9

14.7

18.12

4.28

17.3

0.008

2.333

97.667

Mud

2.9

36.93

4.28

12.4

0.010

8.858

91.142

Mud

3.6

12.7

26.50

4.81

12.4

0.009

1.103

19.183

79.714

Mud

3.8

12.3

19.93

2.12

12.4

0.012

3.014

22.995

73.991

Mud

3.5

12.3

23.00

2.36

15.8

0.013

7.074

24.365

68.561

Mud

4.6

12.1

19.78

2.28

15.6

0.740

3.493

94.385

2.122

Coarse sand

1.8

14.7

1.00

8.35

17.3

0.008

0.953

15.479

83.568

Mud

4.2

20.9

21.47

4.53

17.5

0.015

4.723

31.111

64.166

Mud

3.7

18.4

18.93

5.90

16.1

0.010

1.982

19.971

78.047

Mud

4.5

15.20

4.52

16.1

0.010

13.941

86.059

Mud

4.7

17.2

21.05

4.53

14.1

0.210

11.831

77.685

10.484

Muddy sand

2.6

14.1

1.80

4.85

17.7

0.010

0.587

26.354

73.059

Mud

14.6

19.50

4.61

16.6

0.013

1.028

37.383

61.589

Mud

10.4

16.7

12.98

5.51

18.2

0.011

1.254

25.148

73.598

Mud

5.9

16.3

22.17

5.40

18.1

0.005

12.589

87.411

Mud

4.1

19.2

21.42

4.07

0.012

6.847

31.79

61.363

Mud

1.6

16.9

23.72

5.47

16.5

1500

40.166

48.34

11.494

Muddy sand

28.2

16.1

7.22

40.46

16.1

0.013

9.297

28.211

62.492

Mud

11.5

16.8

10.60

8.61

17.2

0.011

2.367

19.031

78.602

Mud

4.7

16.3

22.32

4.61

0.013

0.093

31.707

68.2

Mud

14.33

4.45

similar pattern of distribution along the inlet according to their average abundance. The trichobranchid, T. stroemi, was only present in group C and also greatly contributed to average similarity within this group.

Cluster analysis based on abundance data of species showed three main groups (fig. 4). The first group was composed of Amphitritides gracilis, Sabellaria spinulosa and Paramphitrite tetrabranchia, species

Animal Biodiversity and Conservation 34.1 (2011)

145

180

Number of taxa

160 140 120 100 80 60 40 20

2 4 6 8 10 12 Number of Terebellida species

Fig. 2. Relationship between total number of benthic polychaete species (●) and total benthic taxa (○) (data from Cacabelos et al., 2008b and 2008a respectively) and number of Terebellida species (y = 12.09x + 32.22, R2 = 0.73 and y = 6.15x – 8.43, R2 = 0.82). Fig. 2. Relación entre el número total de especies de poliquetos bentónicos (●) y taxones bentónicos totales (○) (datos de Cacabelos et al., 2008b y 2008a, respectivamente) y el número de especies de Terebellida (y = 12,09x + 32,22; R2 =0,73 y = 6,15 x – 8,43; R2 = 0,82).

present in site 27 and scarcely distributed along the inlet. The second group was defined by A. finmarchica and M. palmata at an 87.6% similarity level; these taxa were distributed all along the subtidal bottoms of the inlet appearing in very high densities. The third group was formed at a lower similarity level (45%) and was composed of L. koreni, L. conchilega, T. stroemi, Pista cristata and Polycirrus sp., which appeared also in the subtidal bottoms of the inlet but in low densities, only surpassing the value of 42 ind./m2 in the case of P. cristata in site 26, where it reached a beak density of 282 ind./m2. The combination of temperature of bottom water, coarse sand, fine sand, very fine sand and carbonate content showed the highest correlation with distribution and abundance of Terebellida (BIOENV, pw = 0.50). The sabellariid, S. spinulosa, showed high positive correlations with the coarse fractions of sediment (gravel, very coarse and coarse sand) and mean grain size (Spearman’s correlation coefficient (rs = 0.73–0.91), depth (rs = 0.82) and carbonate content (rs = 0.94). On the other hand, A. finmarchica, P. tetrabranchia, P. cristata and Polycirrus sp. were also positively correlated with depth (rs > 0.5), while the last two species showed positive correlations with coarse fractions of sediment and carbonate content (rs > 0.7), and T. stroemi was positively correlated with temperature of sediment (rs= 0.54). Discussion This study showed that Ensenada de San Simón has a diverse Terebellida fauna; some species presented very high densities in the deep and muddy habitats of

the mouth of the inlet. In general, diversity values of Terebellida were high in comparison to other Galician rías. For example, in the muddy sands of Ría de Aldán (Lourido et al., 2008), with high organic matter contents, A. finmarchica, M. palmata and T. stroemi were also abundant, but they reached smaller densities than in San Simón, only up to 564 ind./m2 for A. finmarchica and up to 21 ind./m2 for the other two species. Three major Terebellida assemblages were determined in Ensenada de San Simón through multivariate analyses. Discrimination between these groups was mainly correlated to the Ampharetidae density, as shown by the SIMPER analysis. A. finmarchica and M. palmata showed the highest similarity (87.6%, see also fig. 4) and were much more abundant within the deeper muddy bottoms of the mouth of the inlet. These species showed high similarity since both were distributed in wide ranges of depth and sediment temperatures, appearing from intertidal bottoms to 28.2 m and from 10.1ºC to 20.9ºC, within the higher silt/clay and organic matter contents (up to 91% and 26% respectively). Other surface deposit–feeder polychaetes, such as paranoids (Cacabelos et al., 2008a), have also shown high abundances in these bottoms,. This is in agreement with the reported dominance of this trophic group in the assemblages from intertidal and shallow areas of other Galician rías (Anadón, 1980; Mazé et al., 1993; Moreira et al., 2006) and estuaries in Portugal (e.g. Gaudêncio & Cabral, 2007). Some of the Terebellida from San Simón have been reported as target species (Greham, 1986; Hiscock et al., 2004). M. palmata was a key species in determining the detected groups of sites in Ensenada de San Simón, its densities being much

146

Cacabelos et al.

Table 2. List of identified Terebellida species found in Ensenada de San Simón, indicating the number of individuals found at each site in the inlet (ind./m2). Order Terebellida

Sites 8

Family Sabellariidae Johnston, 1865

Sabellaria spinulosa Leuckart, 1849

Family Pectinariidae Quatrefages, 1865 Lagis koreni Malmgren, 1866

7.1

10.7

25.0

10.7

Family Ampharetidae Malmgren, 1866 Ampharete finmarchica (Sars, 1866) 14.3

3.6

25.0

Melinna palmata Grube, 1870

3.6

42.9

60.7

3.6

42.9

3.6 228.6

17.9

32.1 1,310.7 846.4

614.3

Family Terebellidae Malmgren, 1867

Amphitritides gracilis (Grube, 1860) Lanice conchilega (Pallas, 1766)

3.6

10.7

42.9

Paramphitrite tetrabranchia Holthe, 1976

Pista cristata (Müller, 1776) Polycirrus sp.

3.6

Family Trichobranchidae Malmgren, 1866 Terebellides stroemi Sars, 1835

7.1

10.7

Similarity

100

20 11 3

28 13 24 9

25 12 18 22 26 27 21 23 16 17 14 19

Fig. 3. Dendrogram using Bray–Curtis similarity coefficient showing the classification of sites. Fig. 3. Dendrograma confeccionado utilizando el coeficiente de similitud de Bray–Curtis, que muestra la clasificación de las estaciones.

Animal Biodiversity and Conservation 34.1 (2011)

147

Tabla 2. Lista de especies de Terebellida identificadas, encontradas en la Ensenada de San Simón, indicándose el número de individuos encontrados en cada estación de la ensenada (ind./m2).

Sites 20

32.1

14.3

32.1

14.3

7.1

7.1 14.3

25.0

7.1

10.7

435.7

582.1

92.9 450.0

621.4

267.9 710.7

0.0 792.9 1717.9

0.0 321.4 992.9 582.1

825.0 50.0 35.7

10.7 1,642.9 2,114.3 150.0

7.1 260.7

325.0

28.6

7.1

3.6

21.4

10.7

28.6

3.6

7.1

10.7 282.1

3.6

7.1 3.6

7.1

10.7

28.6

3.6

25.0

39.3

3.6

17.9

3.6

25.0

28.6

Amp fin

Mel pal

Lag kor

10.7

Similarity

40 60 80

100 Amp gra

Sab spi

Par tet

Lam con

Ter str

Pis cri

Pol sp

Fig. 4. Dendrogram using Bray–Curtis similarity coefficient showing the classification of species. Species codes: Amp gra. Amphitritides gracilis; Sab spi. Sabellaria spinulosa; Par tet. Paramphitrite tetrabranchia; Amp fin. Ampharete finmarchica; Mel pal. Melinna palmata; Lag kor. Lagis koreni; Lan con. Lanice conchilega; Ter str. Terebellides stroemi; Pis cri. Pista cristata; Pol sp. Polycirrus sp. Fig. 4. Dendrograma confeccionado utilizando el coeficiente de similitud de Bray–Curtis, que muestra la clasificación de las especies. Código de las especies: Amp gra. Amphitritides gracilis; Sab spi. Sabellaria spinulosa; Par tet. Paramphitrite tetrabranchia; Amp fin. Ampharete finmarchica; Mel pal. Melinna palmata; Lag kor. Lagis koreni; Lan con. Lanice conchilega; Ter str. Terebellides stroemi; Pis cri. Pista cristata; Pol sp. Polycirrus sp.

148

Cacabelos et al.

Table 3. Faunistic parameters at each site where species of Terebellida were found: Ns. Number of species/0.28 m2; N. Total abundance/m2; J'. Pielou's evenness; H'. Shannon Wiener's diversity index (log2). Tabla 3. Parámetros faunísticos en cada estación donde se hallaron especies de Terebellida: S. Número de especies por 0,28 m2; N. Abundancia total/m2; J'. Índice de equitatividad de Pielou; H'. Índice de diversidad de Shannon Wiener (log2). Site

14.29

–

3.57

–

39.29

0.75

1.49

121.43

0.79

1.58

7.14

1.00

278.57

0.50

0.79

60.71

0.91

1.45

2,239.29

0.47

1.22

1,085.71

0.59

1.19

1,278.57

0.57

1.33

364.29

0.56

0.89

1,207.14

0.48

1.23

3.57

–

1,132.14

0.39

1.00

2,767.86

0.40

1.12

1,435.71

0.69

1.10

92.86

0.82

1.30

53.57

0.94

1.89

2,296.43

0.44

1.40

2,521.43

0.25

0.84

189.29

0.58

0.91

higher than those observed for example by Oyenekan (1988) in British waters (up to 990 ind./m2). A great impact of the high density of this tube–builder over the associated community may be inferred: M. palmata is a large and gregarious polychaete and it has a significant impact on the community structure in soft–bottom subtidal habitats, such as those of copepods (Olafsson et al., 1990), by forming faecal casts on the sediment surface, which could affect the associated meiofauna community. Moreover, the dynamics of its populations can be useful to determine the environmental status of the littoral and sublittoral muds and sublittoral mixed sediments (Hiscock et al., 2004). For example, the abundance of this species increases in habitats suffering organic enrichments, such as those impacted by salmon farming (Kempf et al., 2002). The high densities of M. palmata in San Simón could be related with the high organic matter content of the sediments, derived from both the hydrodynamic conditions in the area and the pellet production of the mussels cultured there. Guillou & Hily (1983) detected a high organic matter flux over their studied area, but the redox potential conditions avoided the settlement of M. palmata and therefore the establishment of large populations. Only after favourable redox conditions were reached could organic enrichment positively contribute to the growth of the M. palmata. This species also served as indicator species of the environmental conditions, increasing its density after an increase of fine sediments derived from some anthropogenic source (Dauvin et al., 2007). However, in San Simón the effect of the high organic matter content on M. palmata density can be restricted depending on the state of the populations and on the interaction with other abiotic factors, since no significant correlation was detected through statistical analyses. Another terebellid, T. stroemi, can be used to detect metals in sublittoral muds, and its density decreases in the case of organic enrichment or after trawling pressures. Therefore, the results shown here can be useful for further comparative analysis to determine the ecological status along a protected area, i.e. the Ensenada de San Simón, and to take the appropriate control measures.

Table 4. Physical features of the groups of sites determined according to Terebellida fauna in Ensenada de San Simón (values: mean ± SD): G. group; Q50. Median grain size (mm); OM. Total organic matter content (%); CO3. Calcium carbonate content (%). Table 4. Características físicas de los grupos de lugares determinados según la fauna de Terebellida en la Ensenada de San Simón (valores: media ± DS); G. Grupo; Q50, Tamaño medio del grano (mm); OM. Contenido total de materia orgánica (%); CO3. Contenido de carbonato cálcico (%), G

Depth

Q50

Gravel

2.9 ± 0.6

0.18 ± 0.04

0.1 ± 0.1

3.3 ± 0.4

0.02 ± 0.01

9.5 ± 2.5

0.16 ± 0.15

Sand

Silt/clay

CO3

46.6 ± 19.6 53.3 ± 19.6

16.5 ± 6.6

5.8 ± 0.5

2.3 ± 1.0

25.0 ± 3.5

72.7 ± 4.0

19.9 ± 2.0

4.4 ± 0.5

6.6 ± 3.9

27.0 ± 3.3

66.4 ± 6.7

17.4 ± 1.6

8.4 ± 3.6

Animal Biodiversity and Conservation 34.1 (2011)

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Table 5. Results of SIMPER analysis. Species were ranked according to their average contribution to similarity/dissimilarity within/between site groups determined by cluster analysis: Av.Sim. Average similarity; Av.Diss. Average dissimilarity; Av.Abund. Average abundance (ind./m2); Av.Sim/Diss. Average similarity/dissimilarity; Ratio. Value of similarity (or dissimilarity)/standar deviation, Sim(Diss)/SD; Contrib.%.Percentage of contribution; Cum.%. Percentage of cumulative similarity/dissimilarity. Tabla 5. Resultados del análisis SIMPER. Se clasificaron las especies de acuerdo con su contribución media a la similaridad/disimiliaridad dentro/entre los grupos de estaciones determinados por los análisis de clasificación: Av.Sim. Similaridad media; Av.Diss. Disimilaridad media; Av. Abund. Abundancia media (Ind./m2); Av.Sim/Diss. Similaridad/disimilaridad media; Ratio. Valor de similaridad (o disimilaridad)/ desviación estándar, Sim(Diss)/SD; Contrib.%. Porcentaje de la contribución; Cum.%. Porcentaje de similaridad/disimilaridad acumulada.

Av.Sim.

Group A

Av.Abund

Ratio

Contrib.% Cum.%

69.36 7.14

Ampharete finmarchica Group B

Av.Sim/Diss

82.22

69.36

5.61

100

A. finmarchica

54.76

30.65

6.26

37.28

Melinna palmata

22.62

24.48

4.82

29.77

67.05

Lagis koreni

9.52

Group C

72.80

23.13

11.13

28.13

95.18

524.68

27.37

4.2

37.6

A. finmarchica

909.74

26.03

5.31

35.76

73.36

Terebellides stroemi

16.56

10.68

4.02

14.66

88.02

Av.Diss

Av.Abund

Av.Sim/Diss

Ratio

M. palmata

Groups A & B

56.81

Group A

Group B

Contrib.% Cum.%.

L. koreni

0.00

9.52

19.02

7.75

33.47

M. palmata

1.19

22.62

17.71

1.93

31.17

64.63

A. finmarchica

7.14

54.76

10.87

1.82

19.14

83.77

Groups A & C

76.15

Group A

Group C

M. palmata

1.19

524.68

24.8

2.9

32.57

A. finmarchica

7.14

909.74

18.57

4.14

24.38

56.95

T. stroemi

0.00

16.56

11.21

3.18

14.73

71.68

11.04

6.04

L. koreni Groups B & C

48.30

M. palmata

0.00 Group B 22.62

Polychaetes were the numerically dominant macrobenthic taxon in the Ensenada de San Simón (Cacabelos et al., 2008b). The proportion of species of Terebellida in this inlet showed a direct relationship with the overall polychaete species richness found in the inlet (Cacabelos et al., 2008a); this is in agreement with previous results reported by Olsgard et al. (2003) for marine sediments. Therefore, the number of species of the order Terebellida may be used to predict the overall species richness of the polychaetes in any given area, at least in muddy sediments, and some Terebellida species can also

Group C 524.68

1.2

11.45

7.94

3.28

79.61

23.71

be used as indicator species. Moreover, due to the large size of these polychaetes and their relatively rapid identification, this order can be very useful for conservation needs such as mapping of biodiversity (Olsgard et al., 2003). Acknowledgements The authors wish to thank the team at the Adaptations of Marine Animals laboratory for their invaluable help with sample collection.

150

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Strandings of cetaceans and sea turtles in the Alboran Sea and Strait of Gibraltar: a long–term glimpse at the north coast (Spain) and the south coast (Morocco) E. Rojo–Nieto, P. D. Álvarez–Díaz, E. Morote, M. Burgos–Martín, T. Montoto–Martínez, J. Sáez–Jiménez & F. Toledano Rojo–Nieto, E., Álvarez–Díaz, P. D., Morote, E., Burgos–Martín, M., Montoto–Martínez, T., Sáez–Jiménez, J. & Toledano, F., 2011. Strandings of cetaceans and sea turtles in the Alboran Sea and Strait of Gibraltar: a long–time glimpse of the north coast (Spain) and the south coast (Morocco). Animal Biodiversity and Conservation, 34.1: 151–163. Abstract Strandings of cetaceans and sea turtles in the Alboran Sea and Strait of Gibraltar: a long–term glimpse at the north coast (Spain) and the south coast (Morocco).— A total of 13 species of cetaceans and three species of marine turtles were found in this study. Data were collected by eight independent and self–regulated stranding networks, providing information about 1,198 marine mammal (10 odontocetii, three mysticetii and one phocidae) and 574 sea turtle stranding events between 1991 and 2008. Trends in the strandings were analysed in relation to species composition and abundance, and their geographic and seasonal distribution. The most abundant species recorded were the striped dolphin and the loggerhead turtle. Some of the strandings, such as the humpback whale, harbour porpoise, hooded seal and olive ridley turtle, were considered 'rare' because their distribution did not match the pattern of the study. When the north and south coasts in the study area were compared, pilot whales stranded more frequently in the north, while delphinid species stranded more in the south coast, and loggerhead turtles stranded more frequently in the north while leatherback turtles stranded more in south coast. Key words: Strandings, South–western Mediterranean, Distribution, Marine turtle, Cetacean, Conservation. Resumen Varamientos de cetáceos y tortugas marinas en el mar de Alborán y el Estrecho de Gibraltar: un vistazo a largo plazo de la costa norte (España) y la costa sur (Marruecos).— En este estudio se registraron un total de 13 especies de cetáceos y tres especies de tortugas marinas, proviniendo los datos de redes de voluntarios que prestan asistencia en los varamientos. Se recogió información de 1.198 mamiferos marinos (10 odontocetos, tres misticetos y un fócido) y 574 tortugas marinas entre los años 1991 y 2008. Se analizaron las tendencias de los varamientos en relación a la composición de especies, su abundancia y su distribución geográfica y estacional. Las especies más comunes fueron el delfín común y la tortuga boba. Algunos de los varamientos, como la ballena jorobada, la marsopa común, la foca de casco o la tortuga olivácea, pueden considerarse "anómalos" puesto que su distribución se escapa a los patrones del estudio. Comparando la costa norte del área de estudio con la sur, los calderones y tortugas bobas vararon con mayor frecuencia en la costa norte, mientras que las especies de delfines y las tortugas laúd vararon con mayor frecuencia en la costa sur. Palabras clave: Varamientos, Sudoeste mediterráneo, Distribución, Tortuga marina, Cetáceo, Conservación. E. Rojo–Nieto, P. D. Álvarez–Díaz, E. Morote, M. Burgos–Martín, T. Montoto–Martínez & J. Sáez–Jiménez, DELPHIS–Ecologistas en Acción Cádiz, c/ San Alejandro s/n., 11510 Puerto Real, Cádiz, España (Spain).– F. Toledano, PROMAR–Ecologistas en Acción Almería, Apdo. 15, 04770 Adra, Almería, España (Spain). Corresponding author: Elisa Rojo–Nieto. E–mail: elisa.rojonieto@gmail.com

ISSN: 1578–665X

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Introduction Andalucía is a region in the south of Spain where two geographic systems meet: the Gulf of Cádiz (Atlantic Ocean) and the Alboran Sea (Mediterranean Sea). Andalucian coastal waters (south of Spain) are characterized by high biodiversity and productive habitats for commercial species; the area is characterized by nutrient inputs due to upwelling processes in the Gulf of Cadiz (García Lafuente & Ruiz, 2007), the Alboran Sea (Reul et al., 2005), and the water bodies mixing processes at the Strait of Gibraltar (Echevarria et al., 2002). In the Mediterranean Sea, 20 species of cetaceans (Notarbartolo di Sciara, 2002), and three species of marine turtles (Jerez et al., 2010) have been recorded. Small and large species of cetaceans are distributed in this area not only as migrators (fin whale, Balaenoptera physalus, L. 1758; Jonsgård, 1966), but also as resident populations (short–beaked common dolphin, Delphinus delphis, L. 1758; Cañadas et al., 2002). Cetaceans are particularly vulnerable to threats deriving from human activities. The Mediterranean Sea supports a high human density in the coastal zone. Chemical pollution, marine debris, climate change, land–based changes (agriculture, industry, tourism, etc.), depletion of marine resources and acoustic contamination may all contribute to the degradation and loss of cetacean habitat. As a result, the natural factors causing cetacean mortality have intensified, such as the morbillivirus epizootic that affected the striped dolphin (Stenella coeruleoalba, M. 1833) in 1990 (Aguilar & Borrell, 1994). In addition, collisions with ships and incidental captures by fisheries are sources of direct mortality (Notarbartolo di Sciara, 2002). Cetaceans may also be affected by oil spills by causes such as contamination of their prey items (Moore & Clarke, 2002). Three of the seven species of marine turtles can still be found in the Mediterranean Sea: the loggerhead sea turtle (Caretta caretta, L. 1758), the green turtle (Chelonia mydas, L. 1758) and the leatherback turtle (Dermochelys coriacea, V. 1761). All of them are currently classified as 'endangered' species (IUCN, 2007). Bycatch in drifting longlines has often been considered to be the main threat for immature loggerhead sea turtles, C. caretta, throughout the Mediterranean (Aguilar et al., 1995; Margaritoulis et al., 2003; Lewison et al., 2004; Deflorio et al., 2005; Camiñas et al., 2006). To be effective, conservation actions require basic biological information about the species. Obtaining abundant estimates is a priority to assess the status of the cetacean species in the Mediterranean Sea and to evaluate the impact that human threats may have on the populations (Gómez de Segura et al., 2006). The information collected could contribute to the development of conservation plans by identifying coastal areas of high importance to cetaceans (Pierce et al., 2010). To promote the maintenance of biodiversity, most species of cetaceans and marine turtles are protected under the European Union Habitats and Species

Rojo–Nieto et al.

Directive (92/43/EEC). The Spanish Law on Conservation of Wild areas and Species (4/1989) recognizes the need to protect cetaceans and marine turtles, and establishes the legal basis for their conservation, providing a national catalogue of threatened species. The Law on Conservation of Nature (9/2001) sets out rules for protection, and the Law on Natural Heritage and Biodiversity (42/2007) includes the incorporation of the Habitats Directive into Spanish Law. Royal decree 1727/2007 completes this legislation, establishing measures for the protection of cetacean species in Spanish waters (Pierce et al., 2010). The monitoring of cetacean strandings in the south of Spain and north of Morocco has been undertaken since 1990 by several non–governmental volunteer organizations through the Marine Animal Stranding Network. Stranding networks are made up of volunteers based at local environmental NGOs, academic associations, and veterinary clinics and also by independent people who respond to or provide professional advice on handling stranding events. Data gathered from stranding events can facilitate management in several ways. It provides an overview of distribution and stranding trends usually observed in the region, it helps monitor stranding patterns (spatial and temporal) by identifying unusual mortality events, and it adds to existing knowledge on distribution of cetaceans already obtained from terrestrial sightings and aerial and shipboard surveys. Furthermore, stranded animals provide information on population movement patterns or residency of a given species (Norman et al., 2004). It may also be possible to draw correlations between beached species and their parent populations in the region (Woodhouse, 1991). This work links data from eight independent and self–regulated stranding networks in Spain and Morocco, providing information about 1198 cetacean and 574 sea turtle stranding events from both sides of the Gibraltar Strait and the Alboran Sea. Trends in stranding reports are analysed in relation to species composition and abundance, geographic and seasonal distribution and size of stranded animals and gender. No attempts have been made to explain the cause of strandings except in general terms. Material and methods Records of marine mammals and turtles stranded on beaches in the study area were collected from several sources; mainly comprised by volunteers from universities and from the NGO Ecologistas en Acción. Data of this study were recorded from 1991 to 2008. Only some punctual data were recorded before that period. The volunteer network is distributed along the coast forming a marine mammal and turtle stranding network, and allowing performance of an effective stranding time–response in the whole area (fig. 1). The stranding network receives alerts from diverse sources such as state agencies like the police corps and coast guards, an emergency phone number, and also from local residents and tourists who may encounter a dead or injured marine mammal (or turtle).

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Spain

Morocco

Gulf of Cadiz

Cadiz Malaga Motril

Almeria

Ceuta

Alboran Sea Melilla

Morocco

Nador, Al Hocemas and Driouch

Fig. 1. Study area indicating the location of volunteer groups. Fig. 1. Área de studio donde aparecen indicadas las localizaciones de los diferentes grupos de voluntarios.

In case of a stranding event, the network immediately sends the closest volunteer (team) out to confirm the report, investigate the animal, collect data about location, beach morphology, weather conditions, sea conditions and physical condition of the animal (alive or dead), to decide the suitable response. If the animal is still alive, qualified personnel such as veterinarians and staff members from the rescue centres go to the site to assist the animal with medical care and, if necessary, to transport it to the nearest rehabilitation centre. Some of the work teams do not have qualified personnel at all times. However, all the volunteers are trained in marine mammal health assessment and supportive care, so they are able to proceed with keeping the animal in situ, checking vital signs, informing general public about the situation and waiting for the authorised personnel to arrive (DELPHIS, 2005). When the stranded animal is dead on the beach (the most frequent case), data are collected according to established protocols (Geraci & Lounsbury, 1993): species identification, general measurements (standard length, head height, body width), body state and body condition. If the body is still in a fresh stage, the authorised team personnel proceed to do a necropsy and collect the samples (tissues, teeth, etc.). Samples

are stored and/or delivered to university research groups who are carrying out studies on cetacean and marine turtles (i.e. teeth for age studies, muscle for isotope analyses for trophic research, etc.). Species This paper contains data from thirteen cetacean species, one phocid species, and four marine turtle species: D.d. Delphinus delphis, common dolphin; S.c. Stenella coeruleoalba, striped dolphin; T.t. Tursiops truncatus, bottlenose dolphin; G.m. Globicephala melas, pilot whale; G.g. Grampus griseus, risso’s dolphin; Z.c. Ziphius cavirostris Cuvier’s beaked whale; B.a. Balaenoptera acutorostrata, minke whale; B.p. Balaenoptera physalus, fin whale; P.p. Phocaena phocaena, harbour porpoise; M.b. Mesoplodon bidens Sowerby’s beaked whale; P.m. Physeter macrocephalus sperm whale; M.n. Megaptera novaeangliae, humpback whale; P.c. Pseudorca crassidens, false killer whale; C.c. Cystophora cristata, hooded seal; D.s.i. Unidentified dolphin (usually D.d. or S.c.); Ca.ca. Caretta caretta, loggerhead sea turtle; D.c. Dermochelys coriacea, leatherback turtle; C.m. Chelonia mydas, green turtle; L.o. Lepidochelys olivacea, olive ridley sea turtle.

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Table 1. Strandings of cetaceans recorded in the study area: C. Cadiz; A. Almeria; G. Granada; Ce. Ceuta; Me. Melilla; M. Morocco; other cases (a M. novaeangliae; b P. crassidens; c P. macrocephalus; d C. cristata; e P. phocoena; f M. bidens). (* The year is shown between brackets.)

T. truncatus

Other year*

S. coeruleoalba Me

7(1990) 1(1967)

1991

1992

1993

1994 1995

1996 1997

1998

2000

1999

2001

2002

115

2003 2004

1 13

2005

2006

2007

2008

G. melas

G. griseus M

Other year* 1(1962) 1(1974)

1991

1992

1993

1994 1995

1996

1997

1998

1999

2000

2001

2002

2003

2004

2005

2006

2007

2008

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Tabla 1. Varamientos de cetáceos registrados en el área de estudio: C. Cádiz; A. Almería; G. Granada; Ce. Ceuta; Me. Melilla; M. Marruecos; otros casos (a M. novaeangliae; b P. crassidens; c P. macrocephalus; d C. cristata; e P. phocoena; f M. bidens). (* El año se especifica entre paréntesis.)

D. delphis A

Unidentified dolphin

Me M

Z. cavirostris

G Ce Me M

12(1990) 1(1967) 1(1974)

2 6

14 4

115

22 2

21 6

14 13

27 3

B. physalus C

Other cases

B. acustorata Me M

G Ce Me M

1(1960) 1(1986)

1 2

1e 1c

1 1 ,1 ,1 1

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Table 2. Strandings of turtles recorded in the study area: C. Cadiz; A. Almeria; G. Granada; Ce. Ceuta; Me. Melilla; M. Marruecos; a L. olivacea; b C. mydas; c Undefined. Tabla 2. Varamientos de tortugas registrados en el area de studio. (Para las abreviaturas, ver arriba.) Caretta caretta

Ce Me

Other cases

Dermochelis coriacea M

G Ce

Me M

C bC cC

1990

1997

1998

1999

2000

2001

103

2002

2003

2004

2005

1 2

1 1

1 2

1 1 2

2006

2007

1 5 2

2008

1 6

Study area The Alboran Sea is the most western area in the Mediterranean Sea. The western boundary is the Strait of Gibraltar, and the Eastern end is a line drawn from the Gata Cape (Almería) to the African coastline. In the Alboran Sea, upwellings take place, providing nutrient enrichment of the surface waters (Reul et al., 2005), and thus making this area one of the most productive regions in the Mediterranean Sea (Cañadas et al., 2002). Furthermore, it is an important route for migratory species (such as cetaceans and marine turtles) between the Atlantic Ocean and the Mediterranean Sea (DELPHIS, 2009). This makes the Alboran Sea region rich in cetacean and turtle diversity, with stable (Gómez de Segura et al., 2006) and migratory populations (Castellote et al., 2010). The study area is shown in figure 1. Statistical analysis Principal component analysis and cluster multivariate analysis were performed using the software Statistica 6.0 (StatSoft®). Principal component analysis (PCA) enabled ascertainment of associations between variables, thus reducing the dimensionality of the data table. This reduction is accomplished by diagonalisation of the correlation matrix data, which transforms the 'n' standardised original variables into 'n' uncorrelated (orthogonal) ones (weighed linear combinations of the original variables) called principal components (PCs).

The eigenvalues of the PCs are the measure of their associated variance; the participation of the original variables in the PCs is given by loadings, and the individual transformed observations are called scores. A varimax rotation (VF) allows to 'clean up' the PCs by increasing the participation of the variables with higher contribution, while simultaneously reducing the variables with less contribution. Therefore, the number of original variables contributing to each VF is reduced at the cost of a loss of orthogonality. Cluster analysis is an unsupervised pattern recognition technique that uncovers intrinsic structure or underlying behaviour of a data set without making prior assumptions about the data. The objective is to classify the variables, or cases (sampling stations) of the system into categories or clusters based on their proximity or similarity. Hierarchical agglomerative cluster analysis was carried out on the standardised data by means of the Wards method of linkage, using Euclidean distances as a measure of similarity (Massart & Kaufman, 1983). Results and discussion Cetacea Table 1 and 2 show the data of strandings obtained from 1991 to 2008 (with some punctual data of strandings occurring before this period) in six areas on the coast of the Alboran Sea and Strait of Gibraltar. A total

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40 35 30 25 20 15 10 5

T.t. S.c. D.d. D.s.i. G.g. G.m. P.p. Z.c. P.m. B.p. B.a. M.b. M.n. M.c. C.c.

100 90 80 70 60 50 40 30 20 10 0

Ca.ca.

D.c.

L.o.

C.m.

Unidentified

Fig. 2. Percent of strandings for each species: A. Cetaceans; B. Turtles. (Species codes are indicated in Material and methods section.) Fig. 2. Porcentaje de varamientos por especies: A. Cetáceos; B. Tortugas. (Los códigos de las especies se encuentran indicados en la sección Material and methods.)

of 1,198 cases were recorded for 14 species in the study area (10 odontocetii, three mysticetii and one phocidae). This is the widest compilation of stranding data to date in the study area. These data may also be used as an approximation to which species are present in the study area, but never as an approximation to a size of population, because the number of beached animals depends on several factors, such as habitat (more or less close to the platform), feeding behaviour, impact factors, etc., and they are not only linked to number of members. Spatial distribution Figure 2A shows the percentages of stranding of each species. The data have been standardised with respect to the total number of strandings to avoid

errors in comparisons. The species that stranded most frequently were dolphins (striped and common), followed by the pilot whales (Risso’s dolphin and long–finned pilot whale). The striped dolphin was the most abundantly stranded (34.3%) cetacean species in the western Mediterranean. Common dolphins were common two decades ago, but the continuous incidental killings of common dolphins in some areas such as the Alboran Sea and the Strait of Gibraltar may have caused a significant decline even when the population was still abundant (Forcada & Hammond, 1998). In the Andalucian coast, the striped dolphin is catalogued as a 'vulnerable' species, and the common dolphin is catalogued to be at 'critical risk' (Libro Rojo de los Vertebrados Amenazados de Andalucía, 2001); this

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North coast

South coast

45 40 35 30 25 20 15 10 5 B

T.t. S.c. D.d. D.s.i. G.g. G.m. P.p. Z.c. P.m. B.p. B.a. M.b. M.n. M.c. C.c.

North coast

100

South coast

90 80 70 60 50 40 30 20 10 0

Ca.ca.

D.c.

L.o.

C.m.

Unidentified

Fig. 3. Comparison between strandings recorded on the north and south coasts. (Species codes are indicated in Material and methods section.) Fig. 3. Comparación entre varamientos registrados en la costa norte y en la costa sur de la zona de estudio. (Los códigos de las especies se encuentran indicados en la sección Material and methods.)

indicates that the high number of deaths may be important for the conservation of the species. Other species are uncommonly stranded; however, some strandings were maintained over the years (e.g. Balaenoptera physalus). A significant point to highlight in this study is the first stranding record of the species Sowerby’s beaked whale in the Gulf of Cadiz, according to the authors’ knowledge and bibliography. There was a remarkably high occurrence of stranding of pilot whales in Almeria in 2007 (seven cases). According to Fernández et al. (2008) this was due to

a morbillivirus epizootic. The increase in strandings of Cuvier’s beaked whales in Almeria since 2005 is also interesting. In 2003, in Morocco, a program was carried out to monitor the dolphins captured in fishing nets. These data are accounted together with the normal data of strandings for that year. Figures 3a shows a comparison between the strandings that occurred in the north coast, Alboran Sea and Strait of Gibraltar (Andalucian coast), and the south coast (Ceuta and Melilla). The data have been standardized with respect to the total number

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Table 3. Factorial analysis with cetacean data. Factor loadings (Varimax norm.); * Marked loadings > 0.700000.

Table 4. Factorial analysis with cetacean data. Factor loadings (Varimax norm.); * Marked loadings > 0.700000.

Tabla 3. Análisis factorial con los datos de cetáceos según las localizaciones. Cargas factoriales (Varimax normalizada); * Cargas marcadas > 0,700000.

Tabla 4. Análisis factorial con los datos de cetáceos según las especies. Cargas factoriales (Varimax normalizada); * Cargas marcadas > 0,700000.

Principal Components

Variable

Factor 1

Variable

Factor 2

Principal Components Factor 1

Factor 2

Almeria

0.841953* 0.450509

Tursiops truncatus

0.827994* 0.152275

Granada

0.966429* 0.080902

Stenella coeruleoalba

0.186432 0.960231*

Cadiz

0.907535* 0.055762

Delphinus delphis

0.012270 0.935003*

Ceuta

0.439687 0.818370*

Unidentified dolphin

0.891106* 0.410057

Melilla

–0.106394 0.822135*

Ziphius cavirostris

0.891778* 0.108618

Morocco

0.697791

0.593599

Globicephala melas

0.743288* 0.463590

Eigenvalue

3.882011

1.186645

Grampus griseus

0.275200 0.859294*

% Total variance

64.70018

19.77741

Balaenoptera physalus 0.719895* 0.022054

Cumulative eigenvalue 3.882011

5.068656

Balaenoptera acustorata 0.685323 –0.198398

of strandings on each coast (where total stranding cases on each coast were equal to 100). It was found that for the species with a high number of strandings (dolphins and pilot whales) there was a difference in frequency on the beached animals; dolphins stranded more on the south coast (although the difference was not significant, 89% vs. 86%), and pilot whales stranded more on the north coast (8.2% vs. 4.3%). This could be due to the distribution patterns in the study area and/or the oceanographic water circulation. Table 3 shows the results of the factorial analysis (PCA) to establish the determinant factors for location and species variables of the strandings. In the case of locations, there were two principal factors that were related with the north and south coasts. In the case of species, there were two principal factors: one of them related with the two species that stranded more frequently (common and striped dolphin), and the other related to the remaining species. Figure 4A shows cluster analysis results for the associations between the strandings of the different species. For this analysis, the results of factorial analysis by species (table 4) were taken into consideration. The most important factor in this analysis was the delphinid species, possibly due to their representative number of cases. Two well–differentiated groups were found in the cluster analysis: one of them with the common dolphin, striped dolphin and unidentified dolphin (because it is one of the precedent species), and the other group included the rest of the delphinid species. These two groups were representative of the stranding frequency.

Eigenvalue

4.783991 2.134516

% Total variance

53.15546 23.71684

Cumulative eigenvalue

4.783931 6.918507

Figure 4B shows the zoning of the study area according to the strandings that occurred. Cluster analysis was carried out using the components of the two factors by species. Figure 4 shows that south zones were highly related and constituted a group. Northern areas did not make a differentiated group, but certain trend in strandings may be noticed. Further studies would be necessary to corroborate that trend. Temporal trends Figure 5 shows the temporal trends of strandings of different species (the data have been standardized with respect to the total strandings of each species). These temporal trends were only made for those species with more than 10 recorded animals. During the study period, most of the strandings occurred in spring; however, data were not analysed by season. Future analysis would be necessary to show whether different seasonal trend patterns exist. Turtles Table 2 shows the data obtained for the period from 1991 to 2008 (with some punctual data recorded before this period) in six locations on the coasts of the Alboran Sea and the Strait of Gibraltar. Four species

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A T.t G.g. Z.c. G.m. S.c. D.s.i. D.d. 10

50 60 Linkage distance

0.4

0.6

0.8

1.0

1.2 1.4 Linkage distance

1.6

1.8

2.0

0.00

0.05

0.10

0.30

0.35

0.40

100

Almeria Granada Ceuta Marruecos Melilla Cadiz C Cadiz Granada Melilla Marruecos Ceuta Almeria

0.15 0.20 0.25 Linkage distance

0.45

Fig. 4. Cluster analysis results by cetacean species (A); by location, taking cetacean species into consideration (B); and by location, taking turtle species into consideration (C). Fig. 4. Resultados del análisis de grupos por especies de cetáceos (A); por localización geográfica considerando las especies de cetáceos (B) y por localización geográfica considerando las especies de tortugas (C).

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0.2 0.1

0.3 0.2

0.2 0.1

1991 1992 1993 1994 1995 1996 1997 1998 1999 2000 2001 2002 2003 2004 2005 2006 2007 2008

0.1

0.2

0.1

0.35 0.30 0.25 0.20 0.15 0.10 0.5 0

0.25 0.20 0.15 0.10 0.5 0

1990 1997 1998 1999 2000 2001 2002 2003 2004 2005 2006 2007 2008

1991 1992 1993 1994 1995 1996 1997 1998 1999 2000 2001 2002 2003 2004 2005 2006 2007 2008

0.1

1991 1992 1993 1994 1995 1996 1997 1998 1999 2000 2001 2002 2003 2004 2005 2006 2007 2008

0.3

1991 1992 1993 1994 1995 1996 1997 1998 1999 2000 2001 2002 2003 2004 2005 2006 2007 2008

0.3

I 0.2

1991 1992 1993 1994 1995 1996 1997 1998 1999 2000 2001 2002 2003 2004 2005 2006 2007 2008

0.3

1991 1992 1993 1994 1995 1996 1997 1998 1999 2000 2001 2002 2003 2004 2005 2006 2007 2008

0.2

1990 1997 1998 1999 2000 2001 2002 2003 2004 2005 2006 2007 2008

0.3

1991 1992 1993 1994 1995 1996 1997 1998 1999 2000 2001 2002 2003 2004 2005 2006 2007 2008

0.1

1991 1992 1993 1994 1995 1996 1997 1998 1999 2000 2001 2002 2003 2004 2005 2006 2007 2008

C 0.5 0.4 0.3 0.2 0.1 0

0.4 0.3

1991 1992 1993 1994 1995 1996 1997 1998 1999 2000 2001 2002 2003 2004 2005 2006 2007 2008

0.3 0.2 0.1 0

Fig. 5. Temporal trends of strandings in the study area from 1991 to 2008. Data are standardised: A. T. truncatus; B. S. coeruleoalba; C. D. delphis; D. Unidentified dolphin; E. Z. cavirostris; F. G. melas; G. G. griseus; H. B. physalus; I. B. acustorata; J. C. caretta; K. D. coriacea. Fig. 5. Tendencias temporales de los varamientos en la zona de estudio, de 1991 a 2008. Los datos han sido estandarizados: A. T. truncatus; B. S. coeruleoalba; C. D. delphis; D. DelfĂn sin identificar; E. Z. cavirostris; F. G. melas; G. G. griseus; H. B. physalus; I. B. acustorata; J. C. caretta; K. D. coriacea.

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were found stranded in the study area. Data were grouped by location (table 2). A total of 574 animals were recorded and this is also the widest compilation recorded in the area to date. These data may also be used as an approximation of knowledge of the species present in the study area, but not as an approximation of population abundance. Loggerhead sea turtle and leatherback turtle were the most frequent and abundant species in the Western Mediterranean (92.9% and 7.8% respectively). Two other species were found on the beaches, C. mydas and L. olivacea. These strandings were considered 'rare' in this area; the first one because the population was located in the Eastern Mediterranean, and the second was an extremely rare stranding because in the Mediterranean Sea it is considered neither a migration nor resident species in the area. Both species only occurred once in the study period. Spatial distribution Figure 2B shows the percentages of strandings for each species (data are standardized for comparisons). The species that beached most frequently (more than 90%) was the loggerhead sea turtle, followed by the leatherback turtle. Figure 3B shows a comparison between the strandings occurring on the north coast and the south coast of the Alboran Sea and the Strait of Gibraltar. Data have been standardized with respect to the total strandings on each coast (total strandings on each coast is equal 100). The results for the most abundant and representative species (loggerhead sea turtle and leatherback turtle) revealed slight spatial difference in frequency of the beached animals: the loggerhead sea turtle stranded more on the north coast (although the difference was not significant, 93.3% vs. 82.5) and the leatherback turtle stranded more on the south coast (5.4 vs. 17.4%). The reasons for this could be due to the distribution patterns of each species in the study area and/or by the oceanographic characteristics. In this case, the factorial analysis was not used because there were only two factors (the two main species). Cluster analysis was used to zone the study area according to stranding occurrence (fig. 4C). Cluster analysis was carried out using the two main species. Figure 4C shows how southern areas were highly related and formed a group (including one of the northern areas). The north area did not make a differentiated group, but as found in the cetacean analysis, the cluster analysis showed a certain trend. Further studies would be necessary to contrast that trend. Temporal trends Figure 5B shows the temporal trends of the two main species. Data were standardised with respect to the total stranding of each species. These temporal trends were only obtained for those species with more than 10 stranding events. Over the years, most of the strandings occurred in summer; however, data for this study were not analysed by seasons. Future analysis will show whether there are patterns in these trends. Although little information is available to determine the causes of these deaths, the reason appears to

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be mainly human induced for the marine turtles, but cetacean deaths were less clear. This paper does not attempt, however, to explain the causes of mortalities. Conclusions This study is the widest compilation made to date of cetacean and sea turtles stranded in the Strait of Gibraltar–Alboran Sea: 1,198 strandings of cetaceans and 574 strandings of sea turtles were recorded. Thirteen species of cetaceans, one species of pinnipeds and four species of sea turtle were identified. Two of these species were recorded for the first time in the area of Cadiz (Mesoplodon bidens and Lepidochelys olivacea). A trend can be observed in the strandings for cetaceans and sea turtles. There are two differentiated areas: the north coast and south coast of the study area. This zoning could be due to the distribution patterns of the species in the study area and/or to oceanographic characteristics. Further studies with future data will provide information to ascertain whether the observed variations were isolated events or part of a defined zoning. For cetaceans, specifically in the case of delphinids, two groups can be defined according to stranding occurrence; one of them is formed of common and striped dolphins, and the other group consists of the remaining delphinids. The first group was by far the most important group of cetaceans that stranded in the study area. Acknowledgements The authors wish to thank Ecologistas en Acción, PROMAR–EA, DELPHIS–EA, GUELAYA–EA, SEPTEM NOSTRA–EA, ALBORÁN–EA, AZIR Association, Moustakbalm Association, and Moubadara Association for providing the data, and for all their collaboration. The authors are grateful to all volunteers for their altruistic work over the years. The authors acknowledge the patience and efforts of the reviewers while critically reading the manuscript. References Aguilar, A. & Borrell, A., 1994. Abnormally high polychlorinated biphenyl levels in striped dolphin (Stenella coeruleoalba) affected by the 1990–1992 Mediterranean epizootic. Science of The Total Environment, 154: 237–247. Aguilar, R., Mas, J. & Pastor, X., 1995. Impact of the Spanish swordfish longline fisheries on the loggerhead sea turtle Caretta caretta population in the Western Mediterranean. In: Proceedings of the 12th Annual Workshop on Sea Turtle biology and Conservation NOAA Tech Memo NMFS–SEFSC, 361: 1–6 (J. I. Richardson & T. H. Richardson, Eds.). Gland, Switzerland. Camiñas, J. A., Báez, J. C., Valeiras, Z. & Real, R., 2006. Differential loggerhead by–catch and direct mortality due to surface longlines according to boat

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strata and gear type. Scientia Marina, 70: 661–665. Cañadas, A., Sagarminaga, R. & García–Tiscar, S., 2002. Cetacean distribution related with depth and slope in the Mediterranean waters off southern Spain. Deep–Sea Research I, 49: 2053–2073. Castellote, M., Clark, C. W. & Lammers, M. O., 2010. Population identity and migration movements of fin whales (Balanoptera physalus) in the Mediterranean Sea and Strait of Gibraltar. Journal of Cetacean Research and Management, SC/62/SD2. Deflorio, M., Aprea, A., Corriero, A., Santamaría, N. & De Metrio, G., 2005. Incidental captures of sea turtles by swordfish and albacore longlines in the Ionian sea. Fisheries Science, 71: 1010–1018. DELPHIS, 2005. Informe de varamientos. Cetáceos y Tortugas marinas en la provincia de Cádiz 2004–2005. – 2009. Informe de varamientos. Cetáceos y Tortugas marinas en la provincia de Cádiz. 2003–2009. http://www.ecologistasenaccion.org/IMG/pdf_Informe_de_varamientos_de_Cadiz_2003–2009.pdf Echevarría, F., García Lafuente, J., Bruno, M., Gorsky, G., Goutx, M., González, N., García, C. M., Gómez, F., Vargas, J. M., Picheral, M., Striby, L., Varela, M., Alonso, J. J., Reul, A., Cózar, A., Prieto, L., Sarhan, T., Plaza, F. & Jiménez–Gómez, F., 2002. Physical–biological coupling in the Strait of Gibraltar. Deep–Sea Research Part II: Topical Studies in Oceanography, 49(19): 4115–4130. Fernández, A., Esperon, F., Herraez, P., Espinosa, A., Clavel, C., Bernabe, A., Sanchez–Vizcaino, J., Verborgh, P., De Stephanis, R., Toledano, F. & Bayon, A., 2008. Morbillivirus and Pilot Whale Deaths, Mediterranean Sea. Emerging Infectious Diseases, 14: 792–794. Forcada, J. & Hammond, P., 1998. Geographical variation in abundance of striped and common dolphins of the western Mediterranean. Journal of Sea Research, 39: 313–325. García–Lafuente, J. & Ruiz, J., 2007. The Gulf of Cádiz pelagic ecosystem: A review. Progress in Oceanography, 74(2–3): 228–251. Geraci, J. R. & Lounsbury, V. J., 1993. Marine mammals ashore: a field guide for strandings. Texas A & M Univ. Sea Grant College Program. Galveston, TX. 305. Gómez de Segura, A., Crespo, E. A., Pedraza, S. N., Hammond, P. S. & Raga, J. A., 2006. Abundance of small cetaceans in waters of the central Spanish Mediterranean. Marine Biology, 150(1): 149–160. IUCN, 2007. International Union for the Conservation of Nature and Natural Resource Red List of Threatened Species. Jerez, S., Motas, M., Cánovas, R. A., Talavera, J., Almela, R. M. & Del Río, A. B., 2010. Accumulation and tissue distribution of heavy metals and essential elements in loggerhead turtles (Caretta caretta) from Spanish Mediterranean coastline of Murcia. Chemosphere, 78(3): 256–264. Jonsgård, A., 1966. The distribution of Balaenopteridae in the North Atlantic Ocean. In: Whales,

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dolphins and porpoises: 114–124 (K. S. Norris, Ed.). Univ. California Press, Berkeley. Lewison, R. K., Freeman, S. A. & Crowder, L. R., 2004. Quantifying the effects of fisheries on threatened species: the impact of pelagic longlines on loggerhead and leatherback sea turtles. Ecological Letters, 7: 221–231. Libro Rojo de los Vertebrados Amenazados de Andalucía, 2001. Consejeria de Medio Ambiente, Junta de Andalucía, Spain, 2001. Margaritoulis, D., Argano, R., Baran, I., Bentivegna, F., Bradai, M. N., Camiñas, J. A., Casale, P., De Metrio, G., Demetropoulos, A., Gerosa, G., Godley, B. J., Haddoud, D. A., Houghton, J., Laurent, L. & Lazar, B., 2003 Loggerhead turtles in the Mediterranean sea: present knowledge and conservation perspectives. In: Loggerhead sea turtles: 175–198 (A. B. Bolten & B. E. Witherington, Eds.). Smithsonian Books, Washington, DC. Massart, D. L. & Kaufman, L., 1983. Interpretation of analytical chemical data by the use of cluster analysis. Wiley, New York. Moore, S. E. & Clarke, J. T., 2002. Potential impact of offshore human activities on grey whales. Journal of Cetacean Research and Management, 4(1): 19–25. Norman, S. A., Bowlby, C. E., Brancato, M. S., Calambokidis, J., Duffield, D., Gearin, P., Gornall, T. A., Gosho, M. E., Hanson, B., Hodder, J., Jeffries, S. J., Lagerquist, B., Lambourn, D. M., Mate, B., Norberg, B., Osborne, R. W., Rash, J. A., Riemer, S. & Scordino, J., 2004. Cetacean strandings in Oregon and Washington between 1930 and 2002. Journal of Cetacean Research and Management, 6(1): 87–99. Notarbartolo di Sciara, G., 2002. Cetacean species occurring in the Mediterranean and Black Seas. In: Cetaceans of the Mediterranean and Black Seas: state of knowledge and conservation strategies. (G. Notarbartolo di Sciara, Ed.) A report to the ACCOBAMS Secretariat, Monaco, February 2002. Section, 3: 17. Pierce, G. J., Caldas, M., Cedeira, J., Santos, M. B., Llavona, A., Covelo, P., Martinez, G., Torres, J., Sacau, M. & López, A., 2010. Trends in cetacean sightings along the Galician coast, north–west Spain, 2003–2007, and inferences about cetacean habitat preferences. Journal of the Marine Biological Association of the United Kingdom, 90(8): 1–14. Reul, A., Rodríguez, V., Jiménez–Gómez, F., Blanco, J. M., Bautista, B., Sarhan, T., Guerrero, F., Ruíz, J. & García–Lafuente, J., 2005. Variability in the spatio–temporal distribution and size–structure of phytoplankton across an upwelling area in the NW–Alboran Sea, (W–Mediterranean). Continental Shelf Research, 25(5–6): 589–608. Woodhouse, C. D., 1991. Marine mammal beachings as indicators of population events. Marine mammal strandings in the United States. Proceedings of the second marine mammal stranding workshop. US Dep. Commer., NOAA Tech. Rep. NMFS, 98: 111–115.

Editor executiu / Editor ejecutivo / Executive Editor Joan Carles Senar

Secretaria de Redacció / Secretaría de Redacción / Editorial Office

Secretària de Redacció / Secretaria de Redacción / Managing Editor Montserrat Ferrer

Museu de Zoologia Passeig Picasso s/n 08003 Barcelona, Spain Tel. +34–93–3196912 Fax +34–93–3104999 E–mail mzbpubli@intercom.es

Consell Assessor / Consejo asesor / Advisory Board Oleguer Escolà Eulàlia Garcia Anna Omedes Josep Piqué Francesc Uribe

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Status of marine protected areas in Egypt M. Samy, J. L. Sánchez Lizaso & A. Forcada

Samy, M., Sánchez Lizaso, J. L. & Forcada, A., 2011. Status of marine protected areas in Egypt. Animal Biodiversity and Conservation, 34.1: 165–177. Abstract Status of marine protected areas in Egypt.— Egypt has sought to protect its natural resources and marine biodiversity by establishing a network of six MPAs that are generally located in the Gulf of Aqaba and the Red Sea; most of them include interconnected marine and terrestrial sectors based on conserving coral reefs and accompanying systems. We assessed the present status of MPA networks that showed a set of important results manifested in some strengths (i.e. proper selection according to specific criteria, management plans, etc.), and also some weaknesses (i.e. a relatively small protected proportion of the Egyptian marine territorial waters, significant pressures mainly by tourism activities, etc.). Finally, some recommendations are proposed from this work (i.e. incorporate more habitats that are not well represented in the network, especially on the Mediterranean Sea; establishing a touristic carrying capacity of each area; etc.) to improve the current situation. Key words: Marine reserves, Fishing, Tourism, Conservation, Sustainable development, Egypt. Resumen Estado de las áreas marinas protegidas en Egipto.— Egipto ha establecido una red de seis áreas marinas protegidas (AMPs), situadas principalmente en el Golfo de Aqaba y el mar Rojo para proteger sus recursos naturales y su biodiversidad marina. La mayoría incluyen sectores terrestres y marinos interconectados con el fin de conservar los arrecifes de coral y otros sistemas acompañantes. El estado actual de la red de AMPs se manifiesta mediante algunos puntos fuertes (selección basada en criterios apropiados, existencia de planes de gestión, etc.) y también algunos puntos débiles (protección de una proporción relativamente pequeña de las aguas territoriales egipcias; presiones significativas de algunas actividades, principalmente el turismo, etc.). Finalmente, se proponen algunas recomendaciones (incorporación de más hábitats que no están bien representados en la red actual, particularmente en el Mediterráneo; establecimiento de una capacidad de carga turística para cada área; etc.) para mejorar la situación actual. Palabras clave: Reservas marinas, Pesca, Turismo, Conservación, Desarrollo sostenible, Egipto. M. Samy, J. L. Sánchez Lizaso & A. Forcada, Unidad de Biología Marina, Dept. de Ciencias del Mar y Biología Aplicada, Univ. de Alicante. P. O. Box 99, E–03080 Alicante, España (Spain). Corresponding author: M. Samy. E–mail: mahamed_samy_1984@hotmail.com

ISSN: 1578–665X

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Introduction Marine biodiversity is important to conserve for many reasons. It allows the environment to adapt to changing conditions, it is a source of food and raw materials, and marine ecosystems are the most important elements controlling global climate (Norse, 1993). Preserving marine biodiversity for the sake of knowledge itself is also important. Generally, the greatest levels of marine biodiversity are found in tropical countries which are developing (Gray, 1997). However, being poorer than their developed country counterparts in general, they have fewer facilities, equipment, trained staff and resources available to devote to marine biodiversity conservation (Gray, 1997). The Egyptian waters have great biodiversity. There are more than 5,000 species, including 800 species of seaweeds and seagrasses, 209 species of coral reefs, more than 800 species of molluscs, 600 species of crustacean and 350 species of echinodermata (NCS, 2007, 2009). The Egyptian marine environment is distinguished by specific habitats and threatened species, especially coral reefs, mangrove trees, seagrasses, marine mammals (17 species), marine turtles (four species), sharks (more than 20 species), sea cucumber, bivalves, and many birds (white–eyed gulls, sooty falcons, ospreys) (NCS, 2007, 2009). Over the past few years, Egypt has paid special attention to issues of natural resources protection and signed many international conventions related to natural protection. It has also established a system and legislation for conservation of natural heritage, environment and natural resources for the benefit of the present and next generations (NCS, 2005). Nature conservation in Egypt is the responsibility of the Ministry of State for Environmental Affairs (MSEA). Specifically, the Egyptian Environmental Affairs Agency (EEAA) and the Nature Conservation Sector (NCS) are the governmental bodies that are responsible for establishing and managing the National Protected Area Network in Egypt (EcoConServ, 2004; NCS, 2006a, 2006b). Since Egypt sought to fulfil its own natural conservation goals and the international convention signed by the country, it was necessary to establish legislation and a legal framework to begin the process of protecting its habitats. Law 102/1983 provides the legislative framework for establishing and managing protected areas in Egypt, which are defined as 'any area of land or coastal or inland water characterized by special flora, fauna and natural features having cultural, scientific, tourism or aesthetic value' (NCS, 2006b). Furthermore, Law No. 4/1994 for Environment is supportive to Law No. 102/1983, especially in the areas outside the declared protected areas. Although Law No. 4/1994 is focused mainly on pollution issues, many provisions have implications for nature conservation and hunting management in Egypt (Hanafy, 2005). To date, 28 protected areas (PAs) have been declared since the passage of Law No. 102/1983 and the declaration of Ras Mohammed National Park (the first PA in Egypt in 1983). The present network covers

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almost 14.5% of the country’s land and marine areas and physiographic regions, along with other sites of importance such as biodiversity hot spots, cultural heritage sites, geological formations and landscapes of outstanding natural beauty (Fouda et al., 2006; NCS, 2006b; Ghazali & GEPA MU Staff, 2008). The 28 PAs were selected according to specific criteria including biological value, habitat representation, structural/ geologic value, cultural heritage value, importance to traditional cultures, research opportunity, educational opportunity, recreational value, economic value, urgency for protection (condition of area), degree of threat, management concerns (relationship with other programmes or parties) and enforcement potential (Baha El Din, 1998; NCS, 2006b). These criteria assess the degree to which each area contributes to the fulfilment of the objectives of the Protected Area network (Baha El Din, 1998; NCS, 2006b). Other considerations include location, size and shape determination, as well as the spatial relationship between individual Protected Areas (Baha El Din, 1998; NCS, 2006b). Based on the major sensitive habitats, the strategy categorized the Egyptian protected areas into four categories: Marine Protected Areas (six areas), Wetland Protected Areas (eight areas), Desert Protected Areas (10 areas) and Geological Protected Areas (four areas) (Hanafy, 2005; NCS, 2005). With regard to the economic benefits of marine biodiversity, the revenues of marine activities and tourism, including diving, snorkelling and other activities, are more than € 3 billion per year (NCS, 2007, 2009). The economic benefits of coastal–marine tourism in Egypt go well beyond the direct revenue generated by the dive clubs and the snorkelling operators. Hotels and resorts prosper from diver–related tourism, as do other service industries like bars, cafes, launderettes and Internet cafes. Therefore, calculating the total economic benefits of coral reefs from the tourism industry involves much more than simply adding up the number of reef–related tourists and the value–add of the dive and snorkel industry (Herman, 2003). Additionally, the indirect benefits provided as ecosystem services should be considered (Costanza et al., 1997). The importance of conserving marine biodiversity of Egypt prompted this study that was undertaken to analyse the status of Marine Protected Areas in Egypt, including both coasts of the Mediterranean and the Red Sea. For each MPA, taken into consideration are the general parameters (size, year of establishment, etc.), regulation and zoning, management resources, monitoring, education programs, problems and threats and socio–economic activities. Finally, general and specific recommendations are made for the management of the Egyptian MPAs that would help to improve the situation. Furthermore, an objective of this work was to analyse the total Egyptian marine surface protected. Additionally, as all the information that exists about the Egyptians MPAs is scattered, conflated and inconsistent and there is not a detailed database or document, another objective of this work was to collect, in one document, all the essential and current information about the Egyptians MPAs.

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Lebanon Mediterranean Sea

Syria

Iraq

West Bank Gaza Strip

Sallum Marine Protected Area

Jordan

Israel

Abu Galum Managed Resources Protected Area Nabq Managed Resources Protected Area Ras Mohammed National Park Saudi Arabia Egypt

Libya

Red Sea

Wadi El Gemal Hamata Protected Area N

Elba Protected Area

Sudan

100

200

400 km

Fig. 1. Marine areas currently protected in Egypt. Fig. 1. Áreas marinas actualmente protegidas en Egipto.

Material and methods The study area The study focused on the marine portion of the current protected areas (PAs) that are located within the Egyptian territorial waters, including the Mediterranean and the Red Sea. Other coastal protected areas that are located on the Egyptian Mediterranean coast (lagoons at the end of the Nile delta) were excluded; they are classified according to the EEAA as wetland protected areas as long as they are interior brackish waters and are only connected to the Mediterranean by narrow inlets. To date, there are six MPAs in Egypt (fig. 1): Ras Mohammed National Park, Nabq Managed Resources Protected Area (including Dahab), Abu Galum Protected Area in South Sinai Governorate (in the Gulf of Aqaba), Gebel Elba Protected Area (including the Red Sea islands), Wadi El Gemal–Hamata Protected Area (in the Red Sea Governorate) and the recently declared Sallum Marine Protected Area, which is the first Egyptian MPA on the Mediterranean Sea coast (Herman, 2003;

EcoConServ, 2004; Hanafy, 2005; NCS, 2005). They include interconnected marine and terrestrial sectors based on conserving coral reefs and accompanying systems, marine biome, mangrove bushes, marine islands and adjacent mountain and desert areas. Data collection The study was mainly based on revision of bibliography (books, documents, articles, reports and other grey literature) to collect, for each MPA considered, the information about general description of the area (location, size, year of establishment, etc.), zoning and uses regulation, management resources, monitoring, problems and threats in the PA, socio–economic activities and benefits from the PA. In spite of all the bibliographic data collected, some specific information of some MPAs was missing (i.e. current number of staff, current budget, etc.). This information was not published in any previous work, so it was necessary to conduct interviews with the directors and/or managers of some PAs. The interviews were different from one area to another depending on

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the missing information. However, the main questions concentrated on the actual management resources. These interviews were carried out by phone or by email. Spatial analysis The surface of the marine portion of the PAs and of the closed areas/no–take zones (NTZs) of most of the protected areas studied were not specified in any of the consulted bibliography nor in the interviews, therefore they were estimated. The maps with the limits of each PAs were georeferenced in a geographic information system. Afterwards, the surfaces of the PAs and of the closed areas/NTZs were estimated. Results Ras Mohammed National Park (RMNP, Red Sea) Ras Mohammed National Park is the oldest and the best–known protected area of Egypt (NCS, 2006b, 2009). It was established as a National Park (IUCN Protected Area Category II) in 1983 by the Law No. 102 of 1983, decree 1068/1983, and adjusted by prime ministerial decree 2035 for 1996 with a total area of 836 km² (land portion: 239 km²; marine portion: 597 km², in which three closed areas cover 2.99 km²) (Fouda, 1984; Baha El Din, 1998; Pearson & Shehata, 1998; Shehata, 1998; Herman, 2003; PERSGA, 2004; NCS, 2006b, 2009). Coral reefs fringe Ras Mohammed from all directions and descend to 100 m into the sea (Baha El Din, 1998; Smith & McMellor, 2005; NCS, 2006b). Littoral habitats include a mangrove Avicennia marina community, salt marshes, intertidal flats and seagrass beds, as well as a diversity of shoreline configurations (Baha El Din, 1998; Hegazy et al., 2002; NCS, 2006b). Fishing is prohibited in inshore areas, around Sharm El Sheikh and within the Ras Mohammed National Park. However, other recreational activities (such as diving, snorkelling and water sports) are allowed in the PA except in the three closed areas where only the scientific research is allowed (Wood, 2007; Sayed Abu Bakr, pers. comm.). Many monitoring programmes are conducted by the protected area staff including: coral reef, fishes, invertebrates and birds. There are a total staff of 22 people in the PA, varying from a manager to researchers and ticket collectors. Moreover, the PA have the necessary infrastructures and equipment to conduct essential surveillance, monitoring programmes, scientific research and basic services and guidance for visitors such as: a visitor centre, a diving centre, a workshop, four laboratories, an experimenting hatchery, toilets, dry toilets, guiding and informational signs in all parts of the PA, maps on the main gate and visitor centre, also two patrolling boats, a research boat and a (4 × 4) car. However, the infrastructures are somehow old and need supporting and updating. On the other hand, the area receives about € 87,500/year from the government. All the income generated by the PA (approximately € 1,952,000/year) comes from entrance fees (about €

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4 for foreign visitors and € 0.8 for Egyptians), collection of penalties and sanctions of any violation (Sayed Abu Bakr, per comm.). However, this finance generation goes to the central fund of the NCS. Although all efforts are made, some problems and threats are still ongoing mainly from the tourism pressure on the area and direct physical impacts on the reefs caused by the visiting divers and snorkellers. Tourism activity in and around the Ras Mohammed National Park is intense, and several studies have estimated the carrying capacity of Red Sea reefs in this area (Hawkins & Roberts, 1994; Smith & McMellor, 2005), with the most prominent suggesting a carrying capacity of around 6,000 to 8,000 dives per year (Hawkins & Roberts, 1994). Even the sites receiving the lowest numbers of visitors exceed this by almost 100%, while the heavily dived sites exceed the recommended levels by over 10 times (Smith & McMellor, 2005). Kotb et al. (2004) reported major indirect threats from tourism in the form of landfills, dredging and sedimentation, sewage discharge and effluent from desalination plants, all associated with continued coastal development. Pollution caused by tourism boats including waste, garbage, plastic bags and water bottles, also forms a source of the problem. Furthermore, anthropogenic impacts on coral reefs can be assumed to be cumulative, with natural impacts causing coral deterioration in the area (Smith & McMellor, 2005). Some benefits from the protection were detected within the PA, such as the higher abundances of several commercial species, particularly among the groupers (Roberts & Polunin, 1992). On other hand, Tawfik (2004) estimated that the recreational value of the coral reefs in only Ras Mohammed National Park amounted to about € 113 to € 152 million per year; this excludes the value of the many ecological services provided by coral reefs, and nor does it take into account the employment opportunities arising from tourism, such as recreational activities which the Ministry of Tourism estimated create about 200,000 jobs for every million visitors (Borhan et al., 2003; Tawfik, 2004). Abu Galum Managed Resources Protected Area (AGMRPA, Red Sea) The PA has been established as a Managed Resource Protected Area (IUCN PA management category VI) in 1992 by Law No. 102 of 1983, prime ministerial decree 1511 for 1992 declaration of the area, and adjusted by decree 33 for 1996 declaration of the whole Gulf of Aqaba (Egyptian Side) Natural Protectorate (Baha El Din, 1998; Herman, 2003; NCS, 2006b, 2009). The area has a total surface of 458 km² (land portion: 337 km²; marine portion: 121 km², in which four NTZs cover 52.72 km²) (Baha El Din, 1998; Herman, 2003; NCS, 2006b, 2009). The area includes a remote and pristine stretch of beach along the Gulf of Aqaba coast fringed by rich coral reefs and many adjoining marine and coastal habitats: seagrass, lagoon, spawning areas, rock and sand shores (Goodman & Meininger 1989; Ibrahim, 1993; Baha El Din, 1998; Herman, 2003; NCS, 2006b, 2009).

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All recreational activities are allowed in the marine portion, while artisanal fishing is allowed only for local people (Bedouins). However, fishing is prohibited in 4 NTZs (Wood, 2007; Khaled El Haddad, per. comm.). Many monitoring programmes are conducted by the protected area staff and include coral reefs, bivalves, flora, fauna, and Bedouin settlement. There is a total staff of 15 people in the PA, varying from a manager to researchers and ticket collectors. Moreover, the infrastructures and equipment in the area are considered poor to conduct the essential surveillance, monitoring programmes, scientific research and basic services and guidance for visitors. For instance, the area does not have its own patrol boat. The boat used for patrolling is owned by a NGO. They are expecting a new research boat from the government in the next few months. However, the area contains some infrastructures and equipment that could help in the meantime until sooner updates, such as a visitor centre and a house for staff with solar energy system and water tanks, toilets, shelters, garbage boxes, dumping site, cafeterias, a cabin car, computers, GPS. On the other hand, the area receives about € 7,000/year from the government (exceptionally about € 636,320 from the government for 2010/2011 to support the old infrastructures). All the income generated by the PA that comes from entrance fees (about € 2.4 for foreign, € 0.8 for Egyptian and camping € 8) and collection of penalties and sanctions of any violation (Khaled El Haddad, per. comm.) goes to the central fund of the NCS. Fishing and tourism are the main socio–economic activities in the PA. With regard to fishing, landing sites are mainly in El–Rasasah, Hasat El–Hagar, El–Reheibat, El–Hebeisha, Om Faey, Makser Ayed and El–Sokhna area, where each place is about 1 km along the coast. The mean number of fishermen is about 10 fishermen/day; they use hand lines, nets and sometimes shell collecting. Catches are composed mainly of parrot fishes, some species from surgeon groupers, snapper, jacks, sweet lips, spangled emperor, sky emperor and big–eye emperor. Furthermore, the area has good tourism potentiality (400 visitors/ day) in forms of diving, snorkelling and camping. The mean number of divers is 75–100 per day, while the mean number of campers is about 10 per day. Nabq Managed Resources Protected Area (NMRPA, Red Sea) The Nabq Managed Resources Protected Area (NMRPA) was established as a Managed Resource Protected Area (IUCN PA management category VI) in 1992 by the Law N. 102 of 1983, prime ministerial decree 1511 for 1992 declaration of the area, and adjusted by decree 33 for 1996 establishing Dahab Marine Protected Area as part of NMRPA and declaration of the whole Gulf of Aqaba (Egyptian Side) a natural protectorate (Herman, 2003; Mabrouk, 2007; NCS, 2009). The area has a total surface of 586.5 km² (land portion: 464.6 km²; marine portion: 121.9 km², in which five closed areas cover 97.27 km²) (Mabrouk, 2007; NCS, 2009). The PA includes a variety of marine

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habitats, having one of the northern–most mangrove Avicennia marina communities in the world, coral reefs and seagrass (Ibrahim, 1993; Mabrouk, 2007; NCS, 2009; Riegel & Luke, 1997a). The marine portion of NMRPA (excluding Dahab MPA) falls under four management zones of varying protection levels: strict natural zone, no–take zone, recreational zone and multiple use zone (Mabrouk, 2007). The strict natural zone (marine 91.27 km²) includes a scientific reserve for about 15 km of the coastline where all activities are prohibited except scientific research (Mabrouk, 2007). While four NTZs (marine 6 km²) cover about 5 km of the coastline where all fishing is prohibited, recreational activities (boating, scuba diving, snorkelling, reef walking), and scientific research are permitted (Benzoni et al., 2006; Mabrouk, 2007). Activities in Dahab MPA are not regulated. Many monitoring programmes are conducted by staff at the protected area. These programmes include: coral reefs, mangrove rehabilitation, flora, fauna, and Bedouin settlement (Mabrouk, 2007). There are a total staff of 19 people in the PA, varying from a manager to researchers and ticket collectors. However, infrastructures and equipment to conduct the essential surveillance, monitoring programmes, scientific research are poor. For instance, the area has only one boat and one car for patrolling. Nevertheless, it has infrastructure and equipment for staff accommodation, basic services and guidance for visitors such as a visitor centre and 3 houses for staff, solar energy system and water tanks, toilets, shelters, two hand craft workshops for Bedouin products. The area receives around € 7,000/year from the government, while finance generation goes to the central fund of the NCS. The income generated by the NMRPA comes from entrance fees (about € 4 for foreign, € 0.8 for Egyptian), collection of penalties and sanctions of any violation (Khaled El Haddad, pers. comm.). Although all efforts are made, some problems and threats still remain, generally by tourism, pollution and fishing (Mabrouk, 2007). Pollution in the form of solid wastes (mainly plastic bags and bottles) comes from tourists and the Bedouin community (Mabrouk, 2007). Also ships that pass the Gulf discharge the ballast water and generate pollution that drafts to the shore of the PA (PERSGA, 2001; Mabrouk, 2007). Physical contact by anchors and anchor chains of tourism boats and yachts are potential sources of coral reef damage (Mabrouk, 2007). Fishing in Dahab MPA is not regulated and causes a conflict with other recreational activities like diving and snorkelling (Ashworth & Ormond, 2005; Mabrouk, 2007). Moreover, local Bedouin women harvest invertebrates daily on shallow reef flats using a traditional metal spear (Ashworth & Ormond, 2005; Mabrouk, 2007). Finally, dugong and turtles are being caught as by–catch and hit by fast moving boats (Mabrouk, 2007). Fishing and tourism are the main socio–economic activities in the PA. The fishermen village, El Ghargana, is the only settlement of Bedouin on Nabq coast and is the main landing place of all fishermen. The

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number of permanent fishermen is around 20, but increases up to 40 in summer time, with a total yield from each fishing site that ranges between 1.9 and 6.2 t km–2 yr–1 (Mabrouk, 2007). Catch is composed of Scaridae, Siganidae, Acanthuridae, Lethrinidae, Mugilidae, Kyphosidae, Haemulidae, Labridae and Serranidae (Mabrouk, 2007). Some benefits from the protection were detected within the PA. Since 1995, fishery has been regulated, and after five years of protection the abundance of the main target fish families was found to be significantly greater (Galal, 1999; Galal et al., 2002). However, Galal et al. (2002) reported that fishing by Bedouins in Nabq PA had led to a significant decrease in the abundance and mean length of some serranids and lethrinids. Moreover, the high diversity of the area gives it tourism potential with an average of 18,000 visitors/year (Mabrouk, 2007). The area depends on three main recreational activities: diving, snorkelling and wind surf. Wadi El Gemal–Hamata Protected Area (WEGHPA, Red Sea) Wadi El Gemal–Hamata Protected Area (WGHPA) has been established as National Park (IUCN Protected Area Category II) in 2003 by the Law No. 102 of 1983, and prime ministerial decree 134/2003 (Baha El Din, 2003). It has a total area of 7,450 km² (land portion: 5,850 km²; marine portion: 1,600 km², in which three NTZs cover 305.57 km²) (Baha El Din, 1998, 2003; Herman, 2003; Mansour, 2003; NCS, 2009). The shores of the region are heterogeneous in nature, encompassing rocky, sandy and muddy beaches (Baha El Din, 1998, 2003; Mansour, 2003). The marine part of the protected area encompasses a strip of marine waters of an average width of 15 km along 110 km of the coast. This marine portion includes all the important coral reefs in the region, as well as marine islands, seagrass meadows, mangrove stands, intertidal pavement with algae, intertidal sand (Baha El Din, 1998, 2003; Mansour, 2003). The marine area falls under four management zones of varying protection levels: Strict natural zone, No–take zone, Recreational zone and Multiple use zone (Baha El Din, 2003). All fishing forms are prohibited in the three NTZs, while the recreational activities (boating, scuba diving, snorkelling, reef walking) and scientific research are permitted (Baha El Din, 2003). Many monitoring programmes are conducted by the protected area staff including coral reefs, mangrove rehabilitation and mooring maintenance, which are performed by a NGO called HEPCA (Hurghada Environmental Protection and Conservation Association) (USAID/Egypt, 2008). There is a total staff of 50 people in the PA, varying from a manager to researchers and technicians. Moreover, the area contains good infrastructures and equipments to conduct the essential surveillance, monitoring programmes, scientific research, accommodation for staff, basic services and guidance for visitors such as: a central administration office, a conference hall, a visitor centre, houses for staff, toilets, kitchen,

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also seven (4 x 4) vehicles, four patrol boats, heavy machinery (two soil–moving equipment, one truck), communications (three satellite phones, three mobile phones made available until radio network established), computers… (Baha El Din, 2003; USAID/ Egypt, 2008). Furthermore, the area is funded from a combination of sources: the government (EEAA) (that provides 28% of the budget that is dedicated to salaries of permanent staff), the Samadai Fund of the Red Sea Governorate (36%) and USAID (The United States Agency for International Development) funding from the LIFE Red Sea Project (28%). In addition, the park receives in–kind support for mooring maintenance at dive sites from a NGO (HEPCA) (8%) (Baha El Din, 2003; USAID/Egypt, 2008). The total budget provided by the previous funding sources in 2008 was about € 131,750, while the income generated by the PA in 2007 was about € 2,959,595. This revenue comes from fees (for tour boats and diving operators about € 1.5 per person), collection of penalties and sanctions of any violation (USAID/Egypt, 2008), and goes to the central fund of the NCS. Ongoing problems and threats are generally due to tourism, pollution and fishing (Baha El Din, 2003). In areas without buoy moorings, many boat crews temporarily moor their vessels to reefs using steel that is a potential source of coral reef damage (Baha El Din, 2003). Many forms of pollution are impacting the PA, including solid waste (mainly plastic bags and bottles) coming from tourists and the Bedouin community. In addition, there are no approved sewage outfalls in the PA (Baha El Din, 2003). On the other hand, Barrania & Ibrahim (2003) reported that non–indigenous fishermen in the PA have introduced gill nets named 'sabeeb' that have a smaller mesh size than that legally permitted. They also use ring nets on corals that can lead to physical destruction. Riegel & Luke (1997b) also reported that a very small number of fishermen may still use explosives. Furthermore, dugong and turtles are being caught and hit by fast moving boats. Fishing and tourism are the main socio–economic activities in the PA. Generally, fishing is not a traditional activity of the local people. However, two groups target fishery resources in the PA: local fishermen (the Ababda tribe is the only tribe that has fishing traditions among the local people) and migratory fishermen from other governorates (Baha El Din, 2003). The principal fishing methods used by the traditional fishermen are hand lines, gill nets and trammel nets, while the new settler fishermen are replacing the traditional fishermen, and have less knowledge about the local ecology and sustainable fishing practices, and use some illegal fishing tactics (Baha El Din, 2003; Barrania & Ibrahim, 2003). The main target species by both fishermen groups are: groupers, snappers and grunts (Baha El Din, 2003). There are three main landing sites in the PA: Sharm El Luli (11 boats/55 fishermen), Qulan Village (four boats/20 fishermen) and Hamata Harbour (10 boats/50 fishermen) (Baha El Din, 2003; Barrania & Ibrahim 2003). On the other hand, the area has potential tourism activities concentrated mainly in the marine environment, with only 10% in the terrestrial environment. In 2007, the total

Animal Biodiversity and Conservation 34.1 (2011)

number of visitors was 69,860 with the most important activities being diving (27,631 divers) and snorkelling (23,357 snorkellers) (Mohammed Besar & Mohammed Abbas, pers. comm.). Dive boats regularly visit the reefs in the area and the numbers of boats and visitors are increasing in correlation with the establishment of new hotels (USAID/Egypt, 2008; Mohammed Besar & Mohammed Abbas, pers. comm.). Elba PA (including Red Sea Islands PA) (EPA, Red Sea) The Elba Protected Area was established as a managed resources protected area (IUCN PA management category VI) in 1986 by the Law No. 102 of 1983, prime ministerial decree 450 for 1986 declaration of the area, and adjusted by prime ministerial decree 1186 for 1986 and prime ministerial decree 642 for 1995 (Ghazali & GEPA MU Staff, 2008; NCS, 2009). The area has a total surface of 26,500 km² (land portion: 24,500 km²; marine portion: 2,000 km², closed area surface could not be estimated) (Marchetti & Genena, 2002; USAID/ Egypt, 2007; Ghazali & GEPA MU Staff, 2008; NCS, 2009). The coast and the 22 islands included within the PA support a diverse terrestrial flora and fauna, as well as a rich marine ecosystem including: rocky shoreline, sandy shores, tidal flats, lagoons, salt marsh, mangroves, extensive fringing reefs and seagrass beds (Goodman, 1985; Arnold, 1997; Baha El Din, 1997a; 1997b, 1998; Mekki, 1997; Marchetti & Genena, 2002; USAID/Egypt, 2007; Ghazali & GEPA MU Staff, 2008). The marine area falls under four management zones of varying protection levels: Strict natural zone, No–take zone, Recreational zone and Multiple use zone (Ghazali & GEPA MU Staff, 2008). Many monitoring programmes are conducted by protected area staff, and include: coral reefs, mangrove rehabilitation, flora, fauna, Bedouin settlement (Usama El Ghazali, pers. comm.). There is a total staff of 59 people in the PA, varying from a manager to researchers and technicians. Although the area contains good infrastructures and equipment such as: a visitor centre, a field station, five outputs/control unit, six (4 x 4) vehicles, three patrol boats, they are considered somewhat poor because of the vast surface area which needs more infrastructure and equipment to conduct the essential surveillance (Ghazali & GEPA MU Staff, 2008). The area receives about € 28,000/year from the government, in addition to USAID/LIFE Project that currently involves in developing some activities in GEPA, including handicrafts, solid waste management and public awareness (Marchetti & Genena, 2002; Ghazali & GEPA MU Staff, 2008). Finance generation comes from entrance fees, collection of penalties and sanctions of any violation and the sale of PA products such as handicrafts (Marchetti & Genena, 2002; Ghazali & GEPA MU Staff, 2008). As in all PAs, all the income generated goes to the central fund of the NCS. Some problems and threats continue in the PA, caused generally by fishing, tourism and pollution. Fishing activities in the region are unsustainable because of the illegal fishing techniques and equipment (the case is the same as WGHPA) (Marchetti & Genena,

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2002; Ghazali & GEPA MU Staff, 2008). In addition, as the PA currently does not have potential tourism, it is perceived not to have economic value, thus there is little incentive for decision makers to embrace and support protection of the area (Jameson et al., 1999; Marchetti & Genena, 2002). Finally, pollution in the form of haphazard disposal of solid waste of urban settlements remains a problem because there are no solid waste management systems and waste is thus dumped haphazardly (Marchetti & Genena, 2002; Ghazali & GEPA MU Staff, 2008). Fishing is the main socio–economic activity in the PA. However, it is not a traditional practice of the local tribes as most do not consume fish as part of their diet. Most fishing is done by the Ababda tribe, which is the only tribe that has fishing traditions, and by commercial fishing boats that come from outside the region (Marchetti & Genena, 2002; Ghazali & GEPA MU Staff, 2008). Sallum MPA (SMPA, Mediterranean Sea) Sallum is the first Egyptian MPA on the Egyptian Mediterranean coast, established in March 2010 by Law No. 102 of 1983 and prime ministerial decree 533 for 2010 declaration of the area. It has total area of 1064.2 km² (land portion: 80 km²; marine portion: 984.2 km²) (Environics, 2009). The area encompasses marine and coastal habitats including tidal flats, coastal plains, seagrass meadows, and shallow and intermediate depth marine habitats (Environics, 2009). Three zones have been proposed: core, buffer, and transition as a management zoning scheme for the recently declared area (Environics, 2009). Many criteria were used for Sallum to be declared as a marine protected area: i) the uniqueness and rarity of the protected area that entails unique habitats and geographical features, as mentioned previously; ii) its biological diversity (over 160 species of resident and migratory avifauna, 30 species of reptiles and amphibians, 57 species of macrobenthic organisms and at least 55 commercial marine species including molluscs, crustaceans and fish) and its importance for threatened, endangered and declining species (over 30 species of mammals, some of which are endangered, five marine species of special and global concern, in addition to 11 terrestrial species listed in the Red list of the IUCN 2008); iii) its representativeness of marine and coastal environments, habitats and species of the Mediterranean Sea; iv) the connectivity that it provides, as it is associated and geographically closely linked with similar environments in the countries of the Mediterranean region, which qualifies the area to be a part of the Mediterranean Sea network of marine protected areas; and v) it will replicate similar zones that will be declared at the local level: El Shuwaila and Ras El Hekma, which are proposed within the plan of the Egyptian protected areas network (Environics, 2009). Fishing and tourism are the main socio–economic activities in the PA. Fishermen are the most important category of users of the marine resources of the Gulf of Sallum, and they depend on these resources as a

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Table. 1. Marine habitats represented in the marine protected areas network, and habitats recommended to be incorporated in the future: RMNP. Ras Mohammed National Park; AGPA. Abu Galum Protected Area; NMRPA. Nabq Marine Resources Protected Area; WEGHPA. Wadi El Gemal Hamata Protected Area; GEPA. Gabal Elba Protected Area; SMPA. Sallum Marine Protected Area. Tabla 1. Hรกbitats marinos representados en la red de รกreas marinas protegidas y hรกbitats que se recomienda incorporar en el futuro. (Para las abreviaturas, ver arriba.) Habitats

Red Sea

Med. Sea

RMNP

AGPA

NMRPA

Coral reefs

Reef Fringing

Seagrass meadows

Mangrove stands

Sandy shores

GEPA

Tidal flats

Intertidal pavement with algae

Subtidal sand

Lagoons

Coastal plains Islands

Intertidal sand

Salt marshes

Intertidal flats

SMPA

Muddy shores Rocky shores

WEGHPA

Pelagic habitats

Shallow marine water

Intermediate marine water

Detritic bottoms Coralligenous Bathyal sands Bathyal muds

major source of income (Environics, 2009). According to the statistics of 2007, most of the 49 licensed fishing boats (few are powered and most are sailed) are currently registered in the Department of Matrouh. However, there are another 30 boats registered in the East Port of Alexandria that were fishing in the region of Matrouh to Sallum (Environics, 2009). The main fishing gears used are fishing nets, long lines and some bottom trawls. There is no service to the industry such as refrigerators; even the process of selling and marketing the catch depends on the fish market in the east port of Alexandria (Environics, 2009). The total number of commercial species in the Gulf was 55 (five molluscs, three crustaceans, five cartilaginous fish and 42 bony fish) (Environics, 2009). With regard to tourism, the town is the western

entrance to Egypt and receives about 7,000 tourists yearly in the winter and 10,000 tourists in the summer, mostly from North African neighbours such as Libya, Tunisia and Algeria (Environics, 2009). Discussion Assessment of the present status of MPAs showed a set of important findings manifested in some strengths and weaknesses in the network of the Egyptian MPAs. Generally, MPAs in Egypt are meeting some of their conservation objectives despite the many difficulties that they have. They also have been properly selected according to many criteria that are listed by several authors (Kelleher & Kenchington, 1992; Day

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Table 2. Allocation of staff and budget in the marine protected areas. (For abbreviations, see tble 1.) Tabla 2. Asignación de personal y presupuesto en las áreas marinas protegidas. (Para las abreviaturas, ver tabla 1.)

RMNP

AGPA

NMRPA

WEGHPA

GEPA

Annual budget

€ 87,500

€ 7,000

€ 37,000

€ 28,000

Number of staff

22 people

15 people

19 people

50 people

59 people

Area km²

836 km²

458 km²

586.5 km²

7,450 km²

26,500 km²

Annual budget/100 km²

€ 10,466

€ 1,528

€ 1,193

€ 497

€ 105

2.6 people

3.3 people

3.2 people

0.7 people

0.2 people

Person/100 km²

& Roff, 2000), such as their biological value, habitat representation, geologic value and recreational value. These criteria allowed the objectives of the Protected Area network to be met. The good selection is clearly evident in the MPAs networks of the Red Sea, which are well located and have a good connection since the whole Gulf of Aqaba is protected with south Sinai MPAs (Ras Mohammed, Nabq and Abu Galoum) and the south of the Red Sea MPAs (Wadi el Gemal and Elba including Red Sea islands). Even the newly created Sallum MPA is well located, taking into account the adjacent Ras El Hekma and El Shuwaila proposed areas, which are planned for the near future (around 2014) (Environics, 2009). Monitoring programmes in the MPA network are adequate for assessing the status of protected habitats and existing resources, and the PA Management Unit staff’s technical skills are generally good. Furthermore, as it is a serious concern to prepare a management plan for all MPAs to track effective management or develop a business plan (Kelleher & Kenchington, 1992; Thomas & Middleton, 2003), management plans are already available for four MPAs (Ras Mohammed NP, GEPA, NMRPA and WGHPA) out of the six MPAs. Additionally, an advanced business plan is already available for WGHPA. The MPAs in Egypt had many international initiatives to promote marine protection including international and regional programs (e.g. UNDP, GEF, etc.), projects (e.g. a project to assess the coastal area of Sallum to be declared as MPAs carried in 2009 under association with IUCN, also currently ongoing since 2008 the Life Red Sea project in both WGHPA and GEPA), organization (e.g. IUCN, PERSGA, etc.) and cooperation with other countries (Egyptian–Italian’s cooperation in the form of BioMap project in the end of nineties) (Marchetti & Genena, 2002; PERSGA, 2004; NCS, 2006b; USAID/Egypt, 2007, 2008). The managers, rangers and staff members at MPAs have taken advantage of the involvement in such initiatives (courses, workshops, conferences, campaigns for evaluation of resources and biodiversity, etc.) and many of them have received training in Egypt

or abroad, others have obtained PhDs or master’s degrees (NCS, 2006b). In the last 20 years, the network of MPAs of Egypt has achieved a good reputation and has attracted tourism as one of the best and most important spots for diving and recreation all over the world (Borhan et al., 2003; Tawfik, 2004; Mabrouk, 2007; USAID/Egypt, 2008). During this time, Egypt has increasingly been seen as a regional model for other Arab states and Middle Eastern countries in terms of protected–area management and biodiversity protection. From another perspective, the Egyptian MPAs network has some negative and weak points. Egypt has protected about 5,424.1 km² of the 56,981 km² of Egyptian marine territorial waters, a relatively small proportion (only 9.5%), which is under the 10% to 20% recommended by IUCN and others (Ballantine 1991; Kelleher et al., 1995; Roberts & Hawkins, 2000; Sánchez Lizaso et al., 2000). In addition, the total area of NTZs (which reflect the real protection) is about: 1,052.55 (without considering GEPA and Sallum PA) of the 56,981 km² of the Egyptian marine territorial waters, a very small proportion (only 1.85%). Moreover, the MPA network protects a disparity representation of Egyptian marine habitats and ecosystems, concentrated on the Red Sea coast, and still has only one MPA (Sallum MPA) on its Mediterranean coast protecting about 984.2 km² of the 26,125.8 km² of Egyptian Mediterranean territorial waters (only 3.77%) (table 1). Many other habitats still need to be more represented inside the MPAs of Egypt (e.g. rocky habitats, bathyal habitats and pelagic habitats). Therefore, it is recommended to incorporate more habitats into the network, especially the Mediterranean Sea habitats that are not yet protected (detritic bottoms, coralligenous, bathyal sands and bathyal muds) (table 1). On the other hand, adding these habitats will raise the proportion of the protected territorial water taking into account to close more area as NTZs inside the MPAs. Most of MPAs in Egypt are under–resourced, far below the norm for developing countries or even for Africa. The main limitations to effective management

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are considered to be the very low levels of government funding, the fewer staff than needed and a disparity allocation of both funding and staff (table 2). Inadequate management resources and poor infrastructure facilities are also important constraints. Moreover, the income of each MPA goes to a central fund in the NCS that subsidises other PAs in the whole Egyptian PAs network that do not generate funds. The five MPAs (of the Red Sea) can sustain themselves through tourism income, while the money reinvested in each MPA is less than 10% of what it generates. For instance, Ras Mohammed Marine Park generated about € 1,538,752 in finance year 2004–2005, of which only about € 156,060 was reinvested in the park (Harper, 2006). Governments should provide core support to their MPAs for essential requirements (Kelleher, 1999; Roberts & Hawkins, 2000). Hence, there is an urgent need to increase funding of Egypt’s MPA network, and to ensure that it is addressed by the government of Egypt in a sustained long term manner. Also, the involvement of NGOs and the private sector, 'Friends of Parks', corporate sponsorships and private donations are viable approaches in declining budgets and worsening economic situations and have a good experience of success in many countries (Kelleher, 1999; Riedmiller, 1998, 2003; Mulongoy & Chape, 2004). There is also the need to move away from external funds, such as funding by donors, which is primarily on a project–by–project basis and for a relatively short period of time. Finally, entrance fees to these MPAs are significantly lower than fees for comparable natural attractions in other developing countries. It is therefore recommended to increase entrance fees for each MPA to be equivalent to their comparable developing countries. The MPAs should have a good sea–going capacity, with offshore research and monitoring facilities. Also a comprehensive staff audit should be undertaken in the near future to review the disposition of staff in the PA system. At the beginning of this decade, there were some initiatives to evaluate different services provided by marine ecosystems. These included both management and exploitation costs (e.g. user investment, stakeholders, such as diving centres, other recreational activities, etc.), especially evaluation of coral reefs and mangrove areas in the Red Sea as they are considered the most important ecosystems in the Egyptian waters (Herman, 2003). These studies showed that by far the most important use for reefs is as tourist attractions, although the reefs do have value for fishing, shoreline protection, research and other uses. Because the reefs are such an important component of nature–based tourism, and because such tourism is a crucial component of Egypt’s strategy for sustainable tourism development, it is vital that the reefs be protected from overuse and abuse that would undermine a key asset for Egypt and its economy. Results of these studies indicate that investing in reef protection will prove profitable, as the reefs are a key part of Egypt’s tourism development strategy. Protecting Egypt’s world–class reefs would mean that, year after year, the Red Sea would continue to attract the diving community, who

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spend significantly more money for their vacations than do average tourists (Herman, 2003). These studies provided decision–makers with a potent piece of information that supports the fundamental principle that, in the long run, investing in protecting and managing the environmental and natural resource base that supports tourism in the Red Sea will be good for Egypt’s economy. Recently, the NCS has realized the importance of the economic issue (cost–profit management) and an advanced business plan is already available for WGHPA awaiting the preparation of such plans for the rest of MPAs (Herman, 2003; USAID/Egypt, 2008). Although all efforts are made by the NCS to protect the marine biodiversity and marine resources of Egypt, there are four major threats that are still impacting MPAs in Egypt: recreational use, coral reef deterioration, pollution and illegal fishing. These activities will continue to threaten the resources within PAs until some actions are taken to go ahead to solve these issues. It is obvious that the continued development of the tourism industry is the major and most threatening pressure on MPAs in Egypt (Baha El Din, 2003; Smith & McMellor, 2005; Mabrouk, 2007), since it is a common cause of all other threats such as pollution caused directly by tourists (littering) or indirectly from landfills, dredging and sedimentation, sewage discharge and effluent from desalination plants (Baha El Din, 2003; Kotb et al., 2004; Mabrouk, 2007; Ghazali & GEPA MU Staff, 2008). Also coral reef deterioration by physical impacts on the reefs caused by divers, snorkellers and anchor chains in coral reef areas (Baha El Din, 2003; Smith & McMellor. 2005; Mabrouk, 2007). Therefore, to mitigate the problem it is necessary to establish a tourism capacity for each area, and limit the number of tourists, although this will affect the income of these PAs since it comes mainly from tourism. The conservation and management of coral reefs is a priority issue in Egypt, since it is the most important source of income to MPAs through tourism and diving activities. A certain number of diving activities per day according to the carrying capacities of each area and coral reef cover should be implemented. Direct anchoring should be prohibited on coral reefs. Pollution within PAs is made up of two sources: tourism (in the form of littering, solid wastes and sewage), and oil spills (made by vessels passing through the Red Sea) (PERSGA, 2001; Bashat, 2003). Hence, some actions should be taken such as: the implementation of on the spot fines for littering and solid wastes; no discharge of sewage into the sea or on land and no discharge of liquid or solid waste should be allowed from vessels in or adjacent to the PA; in addition, all sewage should be treated and sludge can be used as fertilizer; finally, the protected area management unit should conduct regular patrols to ensure that vessels operating in the PA do not dispose of liquid or solid waste and do not produce oil discharge. Illegal fishing mainly consists of artisanal fishing either by local Bedouins or the non–indigenous fis-

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hermen coming from outside the area (Baha El Din, 2003; Mabrouk, 2007; Ghazali & GEPA MU Staff, 2008). The relationship and communication between MPA rangers, responsible authorities and Bedouin should be fostered. Moreover, increasing patrols, especially at night, would help to alleviate this problem, as boats of non–indigenous fishermen were often fishing in the early hours of the morning while there is no surveillance. All fishermen active in each PA should receive a license from the EEAA and be registered by the Protected Area Management Unit (PAMU). If any violation occurs, deterrent fines and sanctions should be implemented. Egyptian MPAs are individually vulnerable as a result of poor law enforcement. The PAMU can detect violations, but then the law is not applied because they have to rely on the police and the judiciary to carry it through. In addition, the low support from the local communities elevates the problem because they do not respect the MPA regulation. The NCS should take an active role in discussions and agreements with the police and judiciary, at local and national levels, to ensure the detection of violations and the application of law. Also, the community outreach programmes should be improved in all PAs to ensure that local stakeholders benefit from support and participate in the PA’s management. Finally, regular consultations should be maintained with indigenous community representatives, such as tribal leaders (sheikhs). Conclusion The network of MPAs of Egypt is generally good, having the principal bases such as a sufficient legal framework, good selection of PAs, protection of essential habitats and resources, well trained staff, management plans, and a very high attraction of tourism providing a high income. All these aspects are conducive to aiding a good environment for effective management and protection of the natural resources and marine habitats. However, some constraints need to be addressed for the correct management of these PAs: mainly the lack of management resources and funding, and some impact problems. Finally, once the current problems are solved and by the declaration of the proposed PAs in the near future (the Egyptian Red Sea coast will be totally protected, and two new MPAs on the Mediterranean Sea coast, will increase the proportion protected over the 10% recommended by the IUCN), the Egyptian MPA network might become a good example of coastal management. Acknowledgements Special thanks to CIHEAM (International Centre for Advanced Mediterranean Agronomic Studies) for giving Samy a scholarship to carry out this work. We would also like to thank each one of protected areas management staff who cooperated with us by providing information through interviews.

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Aqaba, Egypt: A mechanism of integrated coastal management. ITMEMS Proceedings. Smith, D. J. & McMellor, S., 2005. Monitoring Program report Ras Mohammed National Park. Baseline Conservation Value Index assessment of Ras Mohammed National Park July–September 2005. Tawfik, R. T., 2004. Recreational value of coral reefs—an application to coral reefs in Ras Mohammed National Park. Master’s thesis, Univ. of York, UK. Thomas, L. & Middleton, J., 2003. Guidelines for Management Planning of Protected Areas. IUCN Gland, Switzerland and Cambridge, UK. USAID/Egypt, 2007. Elba Protected Area Marine Biological Field Survey and Coastal Sensitivity Mapping. LIFE Red Sea Project. – 2008. Wadi El Gemal–Hamata Protected Area Business Plan. LIFE Red Sea Project. Wood, L. J., 2007. MPA Global: A database of the world’s marine protected areas. Sea Around Us Project, UNEP–WCMC & WWF. www.mpaglobal.org

Editor executiu / Editor ejecutivo / Executive Editor Joan Carles Senar

Secretaria de Redacció / Secretaría de Redacción / Editorial Office

Secretària de Redacció / Secretaria de Redacción / Managing Editor Montserrat Ferrer

Museu de Zoologia Passeig Picasso s/n 08003 Barcelona, Spain Tel. +34–93–3196912 Fax +34–93–3104999 E–mail mzbpubli@intercom.es

Consell Assessor / Consejo asesor / Advisory Board Oleguer Escolà Eulàlia Garcia Anna Omedes Josep Piqué Francesc Uribe

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Effects of sea bass and sea bream farming (Western Mediterranean Sea) on peracarid crustacean assemblages V. Fernandez–Gonzalez & P. Sanchez–Jerez

Fernandez–Gonzalez, V. & Sanchez–Jerez, P., 2011. Effects of sea bass and sea bream farming (Western Mediterranean Sea) on peracarid crustacean assemblages. Animal Biodiversity and Conservation, 34.1: 179–190. Abstract Effects of sea bass and sea bream farming (Western Mediterranean Sea) on peracarid crustacean assemblages.— Benthic soft–bottom assemblages are good indicators of environmental disturbance, such as coastal aquaculture, considering their rapid response in terms of diversity and abundance. The aim of this study was to evaluate the response of peracarid assemblages to the release of waste from coastal farming as these organisms play an important ecological role. Abundance and species richness did not show significant differences between farm and control localities but did show a high spatial variability at the two studied scales. Non–metric multi– dimensional scaling (MDS) analysis showed a separation between farms and controls, indicating that peracarid assemblages are modified as a result of aquaculture activities, and some species such as Ampelisca spp. showed statistical differences. Peracarids, at both species and community level, may therefore be applied as helpful indicators to assess benthic effects of coastal farming. Key words: Benthos, Aquaculture, Impact, Indicators, Management, Sustainability. Resumen Efectos del cultivo de la lubina y la dorada (Mediterráneo occidental) sobre las comunidades de crustáceos peracáridos.— Las comunidades bentónicas de fondos blandos son buenas indicadoras de perturbaciones ambientales, tales como la acuicultura costera, teniendo en cuenta sus cambios relativamente rápidos en términos de diversidad y abundancia. El objetivo del presente estudio es evaluar la respuesta de las comunidades de peracáridos a la liberación de desechos de las instalaciones de acuicultura costeras, dado el importante papel ecológico de estos organismos. La abundancia y la riqueza de especies no mostraron diferencias significativas entre áreas con impacto y de control, pero si una importante variabilidad espacial a las dos escalas estudiadas. El análisis no métrico de escalas multidimensionales (EMD) mostró una separación entre las piscifactorías y los controles, lo que indica que las comunnidades de peracáridos se ven modificadas como resultado de las actividades relacionadas con la acuicultura, donde algunas especies, como Ampelisca spp. mostraron diferencias significativas. Por lo tanto, los peracáridos, tanto a nivel de especie como de comunidad, pueden ser utilizados como buenos indicadores para evaluar el efecto de la acuicultura sobre el fondo marino en ambientes costeros. Palabras clave: Bentos, Acuicultura, Impacto, Indicadores, Gestión, Sostenibilidad. V. Fernandez–Gonzalez & P. Sanchez–Jerez, Dept. of Marine Sciences and Applied Biology, Univ. of Alicante, P. O. Box 99, 03080 Alicante, España (Spain). Corresponding author: Victoria Fernandez–Gonzalez. E–mail: victoria.fernandez@ua.es

ISSN: 1578–665X

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Introduction Aquaculture activities have increased greatly in coastal marine areas during the last decades (Mazzola et al., 2000; Mirto et al., 2000; Borja, 2002; Klaoudatos et al., 2006; Sánz–Lázaro & Marín, 2006; Sutherland et al., 2007; Grego et al., 2009). This situation has been induced by progressive advances in cage building, which facilitated mooring of cage farms and their establishment on relatively deep bottoms and exposed sites (Maldonado et al., 2005). Since floating cages for intensive aquaculture started to appear, general concern has increased for the potential impact of this activity on marine ecosystems (Mazzola et al., 2000; Mirto et al., 2000; Klaoudatos, 2002; Sánz–Lázaro & Marín, 2006; Sutherland et al., 2007). These effects include: organic enrichment, derived from excess of uneaten food and fish excretions, chemical pollution, related with medicines and antifouling products, genetic effects and introduction of non–native species, resulting from both the escapes and alterations of adjacent benthic and pelagic fauna (Borja, 2002; Dempster et al., 2002; Macías et al., 2005; Holmer et al., 2007; Borja et al., 2009). From among these possible impacts, the most evident effect of fish cages on seabeds is the accumulation of organic matter, which generates significant changes in the chemical, physical and biological characteristic of the sediment (Karakassis et al., 2000; Mirto et al., 2002; Klaoudatos, 2002; Maldonado et al., 2005; Martí et al., 2005; Marbà et al., 2006; Sánz–Lázaro & Marín, 2006; Lampadariou et al., 2008; Grego et al., 2009; Mirto et al., 2010). These effects can be noted within a range of tens to hundreds of meters (Mazzola et al., 1999; Mirto et al., 2002; Aguado–Giménez & García–García, 2004; Tomassetti et al., 2009). Additionally, the increase of organic matter and sediment structure is affected by silting, increased oxygen demand, anoxic sediment generation and toxic gases (Borja, 2002; Martí et al., 2005). All of these effects could modify the structure and characteristics of the benthic assemblages (Mazzola et al., 1999, 2000; Mirto et al., 2000; Maldonado et al., 2005; Martí et al., 2005; Marbà et al., 2006; Klaoudatos et al., 2006; Lampadariou et al., 2008). Due to their small size, high abundance, direct relation with the sediment, high turnover and fast response time to perturbations, Benthic fauna are presently utilized as a useful indicator to detect environmental changes due to pollution (Boyra et al., 2004; Sutherland et al., 2007; Grego et al., 2009; Fabi et al., 2009). Crustaceans are one of the most important taxa in the benthic fauna, in terms of diversity and abundance. Several groups belonging to this taxon are very ecologically sensitive organisms. As a consequence, a high number of species appear as good indicators of different environmental conditions. Several studies have effectively applied copepods harpacticoids (e.g. copepods–nematods index; Raffaeli & Mason, 1981), ostracods (Ruiz et al., 2005), cumaceans (Corberá & Cardell, 1995) and amphipods (Conradi et al., 1997; Gómez–Gesteira & Dauvin, 2000; Sánchez–Jerez et al., 2000; Guerra–García & García–Gomez, 2001) for assessing different types of environmental impacts. However, there are no studies

Fernandez–Gonzalez et al.

directly assessing the effects of coastal farming on peracarid assemblages in the Western Mediterranean. Consequently, to evaluate the environmental impact of fish farming using peracarid assemblages we applied a multi–control impact design with a spatial replication at different scales to understanding the natural spatial variability with regards to the influence of the fish farming activity. Material and methods Study area and sampling method Three Mediterranean fish farms located east off the coast of Guardamar del Segura (Alicante, SE Spain: 38° 5' 45.88'' N; 0° 36' 15.84'' W) were selected for the study. All farms cultured sea bream (Sparus aurata) and sea bass (Dicentrarchus labrax). In addition, three control zones in the same area were also selected. They were located at least 1.5 km away from the farms to minimize the potential interactions with dispersed farm wastes. Samples were collected in March 2009. Regarding fish farm–impact monitoring, punctual sampling can be relevant, because if important environmental and biotic parameters are affected, the differences between controls and farms should be detectable at any time (Maldonado et al., 2005). To study benthic community, three random replicates were collected at each site using a Van Veen grab (0.04 m2), sieved in seawater through a 500 µm mesh and preserved in 4% formalin. In the laboratory, the peracarids were separated, identified at the lower possible taxonomic level and counted. An additional sample was collected at each location for sediment analysis. Sediment particle size was determined by the wet sieve method, and organic matter content by incinerating a known dried sample in a muffle furnace at 450°C for 4 h (Buchanan, 1984). Data analysis We tested the differences of peracarid assemblages between control areas and farms using both univariate and multivariate statistical analyses. Univariate analysis We analysed the number of species and total abundance of the peracarids, and abundance of the most important species using analysis of variance (ANOVA). The experimental design incorporated three factors: control/farm (fixed and orthogonal with two levels), locality (random and nested in treatment, with three levels), and site (random and nested in Locality, with two levels). Prior to ANOVA, heterogeneity of variance was tested with Cochran’s C–test and data were √ x + 1 transformed in cases where the variances were significantly different, with P < 0.05, and log (x + 1) transformed where the variance was still heterogeneous (Underwood, 1997). Post hoc Student–Neuman Kuels (SNK) tests were used if significant differences were found.

Animal Biodiversity and Conservation 34.1 (2011)

Control

Farm

80 70

40 30

Mud

% Fine sand

% Medium sand

F3–S2

F3–S1

F2–S2

F2–S1

F1–S2

F1–S1

C3–S2

C3–S1

C2–S2

C2–S1

C1–S2

C1–S1

Organic matter (%)

Particle size distribution (%)

100

181

% Coarse sand

% OM

Fig. 1. Granulometric structure of sediment at the studied zones, expressed as the relative abundance (dry weight percentage) of the different grain–size fractions and organic matter content in samples: C. Control; F. Farm. Fig. 1. Estructura granulométrica del sedimento en las zonas estudiadas, expresada como abundancia relativa (porcentaje de peso seco) de las distintas fracciones según el diámetro de partícula y contenido de materia orgánica de las muestras: C. Control; F. Granja.

Multivariate analysis of assemblage structure Non–parametric multidimensional scaling (MDS) was used as the ordination method to explore differences in the peracarid assemblage composition (Clarke & Warwick, 1994). For this test, data were transformed with fourth root and the similarity matrix was calculated using the Bray–Curtis index. The percentage similarities (SIMPER) procedure was then used to calculate the contribution of each species to the dissimilarity between control time and impact (PRIMER software; Clarke, 1993). A permutation test (PERMANOVA software; Anderson, 2004) was used to analyse differences of the overall species composition following the same experimental design as the univariate analysis. Results Characterization of the sediment The seabed was dominated by soft non–vegetated substrates in which the predominant sediment type was mud, but significant differences were found between farm and control treatments. Remarkable similarity was found in grain–size structure among farm sediments since all of them contained a silt/clay (< 0.063 mm) proportion higher to 90%. However,

sediment structure for control sites was different, with large variations in the sand proportion across sites. The highest fine sand content was measured in sampling site C2–S2 (fig. 1). Results of two–way ANOVA test (Control/Farm and Locality factors) showed significant differences for coarse sand and fine sand proportion, which were higher in control areas, and for mud proportion higher in farm areas (table 1). In general, levels of organic matter in sediment samples were relatively high (fig. 1). Organic content of the sediment was lower in control areas (mean value 9.12%) than in farm areas (mean value 10.63%), but without significant differences. The minimum value recorded in the sediment was in C2–S2, which was related with the fine sand proportion. Peracarids assemblages A total of 708 individuals were found: amphipods (64.97%), of which 55.40% were gammarids and 9.75% caprellids, tanaids (20.20%), cumaceans (14.41%) and isopods (0.42%). Therefore, amphipod gammarids were the most abundant group, especially due to the high abundance values recorded for Ampelisca spp. (30.65% of total abundance). Tanaids were the second most representative taxonomic group, with Apseudes latreillei contributing with 18.50% of the total abundance.

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Table 1. Results of analysis of variance (ANOVA) for sediment variables: S. Source; C/F. Control/farm; L. Locality; R. Residual; CT. Cochran test; T. Transformation; MS. Mean square; P. Level of significance; df. Degrees of freedom; ns. Non–significant. Tabla 1. Resultados del análisis de varianza (ANOVA) para las variables del sedimento: S. Fuente; C/F. Control/granja; L. Localidad; R. Residual; CT. Test de Cochran; T. Transformación; MS. Media de los cuadrados; P. Nivel de significación; df. Grados de libertad; ns. No significativo. Coarse sand MS

Medium sand MS

Fine sand

C/F

1.2706 0.047

0.0956 0.299

3.8579 0.0083

598.71 0.031

L(C/F)

0.1573 0.713 0.0672 0.413

0.1641 0.8962

56.571 0.806 2.039 0.769

24 0.2913

Mud

F vs.

6.836 0.141 L(C/F) R

0.0579

0.6422

143.19

4.517

0.5009

0.7743

0.5401

0.7803

0.4723

None

√ (x + 1)

Ln (x + 1)

ArcSin (%)

Regarding species richness and total abundance, a similar pattern was observed for both variables. For species richness, there was a slight decrease from control (mean value 8.08 species) to farm areas (mean value 4.72 species). In relation to the mean total abundance, it was also lower in farm

Farm

Control 10

Total abundance

Species richness

Number of species

Abundance (ind./m2)

None

areas (29.30 ± 4.16 ind./m2) than in control areas (89.37 ± 8.16 ind./m2) (fig. 2). A high spatial variability was found between localities and sites, which probably contributed to the absence of significant differences in both variables between farm and control treatments (table 2).

100

Fig. 2. Values (± SE) of total abundance (ind./m2) and number of species of peracarid. Fig. 2. Valores (± EE) de abundancia total (ind./m2) y de número de especies de peracáridos.

Animal Biodiversity and Conservation 34.1 (2011)

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Table 2. Results of analysis of variance (ANOVA) for species richness, total abundance and abundance of the most important peracarid species. C/F. Control/farm; L. Locality; S. Site; R. Residual; CT. Cochran test; T. Transformation; MS. Mean square; P. Level of significance; df. Degrees of freedom; ns. Non–significant. Tabla 2. Resultados del análisis de varianza (ANOVA) para la riqueza específica, la abundancia total y la abundancia de las especies de peracáridos más importantes: C/F. Control/granja; L. Localidad; S. Sitio; R. Residual; CT. Test de Cochran; T. Transformación; MS. Media de los cuadrados; P. Nivel de significación; df. Grados de libertad; ns. No significativo.

Species richness

C/F

64.000 0.263

78.624 0.197

17.228 0.050

L(C/F)

37.805 0.083

33.061 0.063

2.2568 0.485

S (C/F x L)

12.777 0.026

8.1704 0.023

2.3123 0.000

4.3333

0.6820

0.2701

0.2500

0.2739

0.3871

Ampelisca spp.

Source

Total abundance P

None

√ (x + 1)

Apseudes latreillei MS

Ppooling

F vs.

0.0207

L(C/F)

ns √ (x + 1)

Liropus elongatus

Caprella dilatata

Source

F vs.

C/F

1.0783 0.5554

1.0354 0.2563

1.1062 0.0234

L(C/F)

2.6101 0.5299

0.5911 0.1262

0.0869 0.3304

S (C/F x L)

2.9798 0.0001

0.2119 0.2796

0.0607 0.7143

0.3664

0.1585

0.0984

0.2808

0.3782

0.3354

Ln (x + 1)

√ (x + 1)

Medicorophium runcicorne

Iphinoe tenella

Jassa marmorata Source

F vs.

C/F

10.028 0.1748

1.9124 0.3420

2.7778 0.4997

L(C/F)

3.6944 0.4495

1.6477 0.1506

5.0556 0.0041

S (C/F x L)

3.4722 0.0370

0.6564 0.0144

0.3889 0.6251

1.2778

0.1934

0.5278

0.6087

0.2986

(P < 0.01)

None

0.2105

√(x+1)

ns None

The MDS analysis, based on species abundance (fig. 3), showed a separation between farm and control assemblages. Peracarid assemblages of the different farms were more similar between them than the control assemblages, indicating a homogenisation

of species structure in these areas. At control areas, species compositions showed higher variability, highlighting the separation of C2–S2 from the others which was evident and probably due to the presence of a higher proportion of fine sand.

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C2–S2 (2)

C2–S2 (3)

F3–S1 (3)

2D stress: 0.2

F3–S1 (1) F2–S2 (2) F3–S2 (3) F1–S2 (2) F2–S2 (3) F1–S1 (2)F2–S1 (3) C1–S2 (3) F1–S1 (3) F2–S1 (2) F1–S2 (3)F1–S1 (1) F1–S2 (1) F2–S1 (1) C2–S2 (1) F3–S2 (2) C3–S1 (1) F3–S1 (2) C3–S1 (2) F2–S2 (1) C1–S2 (2) C3–S2 (2) C3–S2 (3) C3–S2 (1) C2–S1 (1) C2–S1 (3) C1–S1 (3) C3–S1 (3) C2–S1 (2)

C1–S1 (2) C1–S1 (1)

C1–S2 (1)

Control Farm

Transform: fourth root Resemblance: S17 Bray Curtis similarity Fig. 3. Non–metric multi–dimensional scaling (MDS) plot in two dimensions for benthic peracarid species abundance: C. Control; F. Farm; S. Site. (The number indicates the replicate samples.) Fig. 3. Análisis no métrico de escalas multidimensionales (MDS) en dos dimensiones a partir de los valores de abundancia de las especies de peracáridos bentónicos. C. Control; F. Granja; S. Sitio. (El número indica las distintas réplicas.)

The PERMANOVA test indicated no significant differences for species composition between farms and control areas, but it revealed a high variability among localities (p = 0.001) as well as among sites (p = 0.01) (table 3). The SIMPER analysis showed that the gammarid amphipods Ampelisca spp., Jassa marmorata and Medicorophium runcicorne, the tanaid Apseudes latreillei, the caprellid amphipods Liropus elongatus and Caprella dilatata and the cumaceans Iphinoe tenella and Bodotria scorpioides were the species that contributed most to the dissimilarity between farm and control areas (table 4). These species were also most responsible for the similarity within the farm and control samples. The abundance values of Ampelisca spp., A. latreillei, L. elongates, I. tenella and B. scorpioides were higher in control areas, while C. dilatata, J. marmorata and M. runcicorne were more abundant in farm sediments (fig. 4), but only two of these species presented significant differences for the abundance between farm and control treatments: Ampelisca spp. and Caprella dilatata (P < 0.05, table 2). Significant differences at localities and sites for many of the variables reflected a high variability

of the peracarid abundance at a scale of hundreds of meters to tens of kilometres, thus reducing the power of the ANOVA. Discussion The study of the peracarid crustacean assemblages under aquaculture influence showed that the species Ampelisca spp. and Caprella dilatata are affected by fish farming activities. A drastic decrease in the total abundance and species richness was detected, even though significant differences were not found. In addition, changes in the sediment structure due to an increase of finer material were also detected. This silting has been previously described associated with organic enrichment from fish aquaculture waste (Sutherland et al., 2001; Borja, 2002; Porrello et al., 2005; Sánz–Lázaro & Marín, 2006; Aguado–Giménez et al., 2007). However, the use of grain size distribution as an impact indicator is not appropriate but it is a very useful parameter for describing the environment and interpreting some phenomena (Aguado–Giménez et al., 2007).

Animal Biodiversity and Conservation 34.1 (2011)

Table 3. Results of PERMANOVA analysis for peracarid assemblages: C/F. Control/farm; L. Locality; S. Site; R. Residual; df. Degrees of freedom; MS. Mean square; P. Level of significance. Tabla 3. Resultados del análisis de PERMANOVA de las poblaciones de peracáridos: C/F. Control/ granja; L. Localidad; S. Sitio; R. Residual; df. Grados de libertad; MS. Media de los cuadrados; P. Nivel de significación. Source

P(perm)

F vs.

C/F

10090

0.198

L(C/F)

S (L (C/F)) 6

5,287.6 0.001 S (C/F x L) 1,957.3 0.010

24 1,177.9

Total

The effects of aquaculture activities on bottom sediments are well known and have been reported worldwide (e.g. Hall et al., 1990; Wu, 1995; Karakassis et al., 1998, Borja, 2002: Borja et al., 2009). In the Mediterranean Sea, these effects are well documented for sea bream (Sparus aurata) and sea bass (Dicentrarchus labrax) farming (e.g. Karakassis et al., 1998; Mazzola et al., 1999; Mirto et al., 2002; Vita et al., 2002; Aguado–Giménez & Garcia, 2004, Maldonado et al., 2005; Tomassetti & Porrello, 2005; Marbá et al., 2006; Aguado–Giménez et al., 2007, Tomassetti et al., 2009) and shellfish farming, particularly mussel farming (Mirto et al., 2000; Fabi et al., 2009).

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Fish farm sediments are assumed to represent organic enriched conditions (Hall et al., 1990; Hargrave et al., 1993; Delgado et al., 1997; Karakassis et al., 1998) if they are compared to reference areas. However, in this study, significant differences in organic matter loads between farm and control areas could not be detected. Other studies (e.g. Maldonado et al., 2005; Aguado–Giménez et al., 2007) showed a similar lack of differences because the magnitude of this increase in organic matter is different between farms and mainly depends on local variables such as hydrographic regime, sediment type, water depth, as well as management variables such as fish production, efficiency of feeding method and feed quality (Tomassetti et al., 2009). Previous studies have demonstrated that the changes originated on the bottom can cause a strong impact on the structure and characteristics of the benthic communities, including effects on bacterial assemblages (Mirto et al., 2000; La Rosa et al., 2004); meiofauna (Mazzola, 1999, 2000; Sutherland, 2007; Grego et al., 2009; Mirto et al., 2010), macrofauna (Edgar et al., 2005; Tomassetti et al., 2009; Fabi et al., 2009) or seagrass species (Delgado et al., 1999; Ruiz et al., 2001; Marbá et al., 2006). Even though several taxonomic groups, such as polychaeta (Tomassetti & Porrello, 2005; Sutherland et al., 2007) or nematode (Mirto et al., 2002), have been proposed as tools for monitoring the impact of organic enrichment following intensive aquaculture activities, few studies have focused on the effects of fish farming on benthic macrocrustacean assemblages. Hall–Spencer & Bamber (2007) described how epifaunal and infaunal benthic crustacean communities are affected for salmon farming on maerl bottoms. In our study, the peracarid assemblages also seem to be an adequate indicator of sea bass and sea bream farming activity. Specific richness and total abundance

Table 4. Result of SIMPER analysis and mean abundances (± SE) of most important peracarid species. Tabla 4. Resultado del análisis SIMPER y abundancias medias (± EE) de las especies más importantes de peracáridos.

Average dissimilarity = 64.51 farm and control

Species Ampelisca spp.

Contrib.(%)

Cum.(%)

Mean abundance Control

Farm

9.17

24.44 ± 3.13

5.69 ± 0.98

6.75

15.92

15.42 ± 5.99

2.78 ± 0.82

Liropus elongatus

6.74

22.66

4.86 ± 1.14

0.97 ± 0.39

Caprella dilatata

6.33

28.98

0.56 ± 0.17

3.06 ± 0.47

Apseudes latreillei

6.32

35.31

0.83 ± 0.31

3.47 ± 0.73

Medicorophium runcicorne

6.06

41.36

1.25 ± 0.37

6.11 ± 1.59

Iphinoe tenella

5.41

46.78

2.36 ± 0.49

0.97 ± 0.31

Bodotria scorpioides

5.17

51.94

2.36 ± 0.65

0.28 ± 0.13

Jassa marmorata

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Fernandez–Gonzalez et al.

Control

Farm

Abundance (ind./m2)

30 25 20 15 10 5 0

Asp

Cdi

Mru

Jma

Lel

Bsc

Ite

Ala

Fig. 4. Mean abundance of most important peracarid species (ind./m2 ± SE): C. Control; F. Farm. Asp. Ampelisca spp.; Cdi. Caprella dilatata; Mru. Medicorophium runcicorne; Jma. Jassa marmorata; Lel. Liropus elongatus; Bsc. Bodotria scorpioides; Ite. Iphinoe tenella; Ala. Apseudes latreillei. Fig. 4. Abundancia media de las especies más importantes de peracáridos (ind./m2 ± EE): C. Control; F. Granja. Asp. Ampelisca spp.; Cdi. Caprella dilatata; Mru. Medicorophium runcicorne; Jma. Jassa marmorata; Lel. Liropus elongatus; Bsc. Bodotria scorpioides; Ite. Iphinoe tenella; Ala. Apseudes latreillei.

in sediments was lower beneath the cages compared to control areas. Similarly, a drastic reduction (50–70%) in crustacean fauna abundances has been reported from other fish farm areas in the Mediterranean (Mazzola et al., 1999, 2000; Mirto et al., 2000; La Rosa et al., 2001; Klaoudatos et al., 2006). At the lowest taxonomic level, this study revealed that some species and genera are sensitive to fish farming. The most important genus in this regard was Ampelisca spp., which was highly sensitive to farm effects. The genus Ampelisca showed a high sensitivity to significant increases in organic matter but also to toxins in the sediment (especially PCBs, pesticides, metals and PAHs) (Gómez–Gesteira & Dauvin, 2000), compounds that may be found in sediments beneath the cages (Tsapakis et al., 2010). The abundances of cumaceans I. tenella and B. scorpioides were also drastically decreased in farm sediments. There is a limited number of cumacean species with adaptative strategies in response to eutrophication (Corberá & Cardell, 1995), so this taxon is considered sensitive to polluted areas. Another species, the tanaid A. latreillei, was apparently affected to organic enrichment because its presence was reduced in farm sediments. Other authors have reported that this species may be vulnerable to hypoxic sediments (Gray et al., 2002; Guerra–García & García–Gómez, 2006; Sánchez–Moyano et al., 2002; Sánchez–Moyano & García–Gómez, 1998). However, the present work showed that this species was mainly associated with one control, which was characterized by fine sand sediments. These results

are in agreement with other works (Bakalem et al., 2009; Bouchet & Sauriau, 2008; Marín–Guirao et al., 2005; Moreira et al., 2008; Lourido et al., 2008, De–la–Ossa–Carretero et al., 2010). On the other hand, other species such as Jassa marmorata, Caprella dilatata and Medicorophium runcicorne increased their presence in farm sediments. In the case of M. runcicorne, species belonging to the family Corophiidae are generally linked to muddy and disturbed areas continuously exposed to toxics in the sediment (Diviacco & Bianchi, 1987; Guerra–García et al., 2003; Carvalho et al., 2006; Vázquez–Luis et al., 2008). Medicorophium runcicorne have been reported in yachting harbours where their presence seems to be influenced by factors other than grain size and organic matter, such as low hydrodynamics, higher sedimentation rate and availability of larvae (Guerra–García & Garcia–Gomez, 2009). Similar results have been found for the gammarid amphipod J. marmorata; their higher abundance at farm sediments must be due mainly to their trophic requirements and living habits since these species are tube–builders and deposit–feeders (Conradi et al., 1997; Guerra–García et al., 2003). But this species and C. dilatata, which is associated with buoys (Guerra–García et al., 2006), are present in fouling communities on aquaculture installations, so their presence in sediments could be due to a direct influence from water column structures to the seabed. We found different structures of peracarid assemblages at control and farm sites. However, the high variability found in these soft–bottoms prevented finding statistical differences. The use of a more robust

Animal Biodiversity and Conservation 34.1 (2011)

hierarchical design (high replication at several spatial scales) is needed to detect significant changes on assemblages, and these results were widespread as an indicator of environment involvement by aquaculture (Underwood, 1997). Conclusions This study revealed that peracarid assemblages are modified at sediments affected by fish farming, as has been described for other faunal groups such as polychaetes and nematods assessing responses to different environmental conditions. Peracarids may also therefore be used for this purpose. Moreover, peracarids play an important role as trophic resources for other crustaceans and macrofauna such as fish populations (Bell & Harmelin–Vivien, 1983; Edgar & Shaw, 1995; Sanchez–Jerez et al., 1999; Stergiou & Karpouzi, 2002; Stål et al., 2007). For example, they are a key component in the diet of key soft–bottom species like Mullus barbatus, where the proportion of crustaceans consumed can reach up to 70% of total prey (Aguirre Villaseñor, 2000). Consequently, peracarids assemblages are useful tools for describing the environmental impact of fish farming activities, but additionally they are important for assessing the potential effects on the trophic webs. Acknowledgements We are grateful to the staff at the fish farms that gave us access and help for the study. We thank Pablo Arechavala López, Maite Vázquez–Luis and Damián Fernandez–Jover for their cooperation throughout this work. This study forms part of the project 'Selección de indicadores, determinación de valores de referencia, diseño de programas y protocolos de métodos y medidas para estudios ambientales en acuicultura marina' and was funded by the Planes Nacionales de Acuicultura (JACUMAR). References Aguado–Giménez, F. & García–García, B., 2004. Assessment of some chemical parameters in marine sediments exposed to offshore cage fish farming influence: a pilot study. Aquaculture, 242: 283–296. Aguado–Giménez, F., Marín, A., Montoya, S., Marín–Guirao, L., Piedecausa, A. & García–García, B., 2007. Comparison between some procedures for monitoring offshore cage culture in western Mediterranean Sea: Sampling methods and impact indicators in soft substrata. Aquaculture, 271: 357–370. Aguirre Villaseñor, H., 2000. Aspectos biológicos y ecológicos del salmonete de fango Mullus barbatus L. 1758 y del salmonete de roca Mullus surmuletus L. 1758 del Mediterráneo Noroccidental. Ph. D. Thesis, Univ. Politécnica de Cataluña. Anderson, M. J., 2004. PERMANOVA 2–factor: a

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Review of the effects of protection in marine protected areas: current knowledge and gaps C. Ojeda–Martínez, J. T. Bayle–Sempere, P. Sánchez–Jerez, F. Salas, B. Stobart, R. Goñi, J. M. Falcón, M. Graziano, I. Guala, R. Higgins, F. Vandeperre, L. Le Direach, P. Martín–Sosa & S. Vaselli Ojeda–Martínez, C., Bayle–Sempere, J. T., Sánchez–Jerez, P., Salas, F., Stobart, B., Goñi, R., Falcón, J. M., Graziano, M., Guala, I., Higgins, R., Vandeperre, F., Le Direach, L., Martín–Sosa, P. & Vaselli, S., 2011. Review of the effects of protection in marine protected areas: current knowledge and gaps. Animal Biodiversity and Conservation, 34.1: 191–203. Abstract Review of the effects of protection in marine protected areas: current knowledge and gaps.— The effectiveness of marine protected areas (MPAs) and the conservation of marine environments must be based on reliable information on the quality of the marine environment that can be obtained in a reasonable timeframe. We reviewed studies that evaluated all aspects related to the effectiveness of MPAs in order to describe how the studies were conducted and to detect fields in which research is lacking. Existing parameters used to evaluate the effectiveness of MPAs are summarised. Two–hundred and twenty–two publications were reviewed. We identified the most commonly used study subjects and methodological approaches. Most of the studies concentrated on biological parameters. Peer reviewed studies were based on control vs. impact design. BACI and mBACI designs were used in very few studies. Through this review, we have identified gaps in the objectives assigned to MPAs and the way in which they have been evaluated. We suggest some guidelines aimed at improving the assessment of the effects of protection in MPAs. Key words: Marine conservation, Management, Assessment, Descriptors, Subject of study, Marine protected areas. Resumen Revisión de los efectos de la protección en las áreas marinas protegidas: conocimiento y deficiencias actuales.— La efectividad de las áreas marinas protegidas (AMPs) y la conservación del medio ambiente marino debe basarse en información fiable sobre la calidad del medio marino que pueda obtenerse en un plazo de tiempo razonable. Se revisaron estudios que evalúan aspectos relacionados con la efectividad de las AMPs con el fin de describir cómo se realizaron los estudios y detectar donde existen vacíos en la investigación. En este estudio se enumeran los parámetros existentes para evaluar la efectividad de las AMPs. Se revisaron 224 publicaciones. Identificamos los objetos de estudio más utilizados y los enfoques metodológicos. La mayoría de los estudios se centran en el estudio de parámetros biológicos. Los estudios publicados se basaron en el diseño control frente a impacto. En muy pocos estudios se utilizaron diseños de muestreo BACI y mBACI. A través de esta revisión, se han identificado deficiencias en los objetivos de las AMPs y en la manera como han sido evaluados. Como conclusión sugerimos algunas pautas para mejorar la evaluación de los efectos de la protección en estas zonas. Palabras clave: Conservación marina, Gestión, Evaluación, Descriptores, Objetos de estudio, Áreas marinas protegidas. Celia Ojeda–Martínez, Just T. Bayle–Sempere & Pablo Sánchez–Jerez, Dept. de Ciencias del Mar y Biología Aplicada, Univ. de Alicante, CP 99, 03080 Alicante, España (Spain).– Fuensanta Salas, Dept. de Ecología e Hidrología, Fac. de Biología, Univ. de Murcia, 30100 Murcia, España (Spain).– Ben Stobart & Raquel Goñi, Centro Oceanográfico de Baleares, Inst. Español de Oceanografía, Muelle Poniente s/n., Palma de Mallorca, España (Spain).– Jesús M. Falcón, Grupo de Investigación BIOECOMAC, Depto. de Biología Animal (Ciencias ISSN: 1578–665X

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Marinas), Univ. de La Laguna, c/ Astrofísico Francisco Sánchez s/n., 38206 La Laguna, Tenerife, Canary Islands, España (Spain).– Mariagrazia Graziano, Dipto. di Biologia Animale, Univ. di Palermo Via archirafi 18, 90123 Palermo, Italia (Italy).– Ivan Guala, Fondazione IMC–International Marine Centre, Onlus, Loc. Sa Mardini 09072 Torregrande, Oristano, Italia.– Ruth Higgins & Fréderic Vandeperre, Dept. of Oceanography and Fisheries, Univ. of the Azores, PT–9901–862 Horta, Portugal (Portugal).– Laurence Le Direach, GIS Posidonie, Centre d’Océanologie de Marseille, Parc Scientifique et Technologique de Luminy, Case 901, 13288 Marseille Cedex 09, Francia (France).– Pablo Martín–Sosa, Canarian Oceanographic Centre, Spanish Inst. of Oceanography, Avda 3 de Mayo, 73, Edificio Sanahuja, 38005 S/C Tenerife, Canary Islands, España (Spain).– Stefano Vaselli, Dept. of Biology, Univ. of Pisa, Italy. *Corresponding author: Just T. Bayle–Sempere. E–mail: bayle@ua.es

ISSN: 1578–665X

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Introduction

Material and methods

Coastal marine environments host key habitats for many endangered marine populations, yet their accessibility and proximity to heavily inhabited areas makes them vulnerable to over exploitation through fishing, and to direct anthropogenic impacts. Traditionally, the demand of food in coastal areas makes fishing one of the most important activities impacting these areas. Fishing exerts direct pressure on the environment as well as on fish stocks, and there is unequivocal evidence that fishing has reduced the abundance and size of the most targeted and valuable species (Chapman & Kramer, 1999; Edgar & Barrett, 1999; McClanahan et al., 1999; Chiappone et al., 2000; Willis et al., 2003; Williamson et al., 2004). Poor planning and overpopulation of coastal areas has added to the problem due to resulting pollution and excessive recreational use (Bellan–Santini et al., 1994). In recent years, Marine Protected Areas (MPAs) have been increasingly seen as a way of reducing the intensity of these impacts (Ward et al., 1999). Since the creation of the first MPA in 1935 (Doumenge, 1993), MPAs have been established throughout the world as a management tool for compensating the effects of human impacts on the coastal marine environment (Agardy, 1994). Specifically, MPAs are implemented to reduce the effects of overfishing of coastal marine stocks, preserve marine biodiversity and protect key habitats (Francour et al., 2001; Halpern, 2003). They also provide a sustainable socioeconomic development for human communities in coastal areas (Sainsbury & Sumaila, 2003). MPAs have been strongly advocated as a tool for the management of fisheries as they conserve fish stocks, increase the number and fecundity of the breeding population, increase the abundance of juveniles and act as nurseries and areas of biodiversity conservation (Bell, 1983; Russ & Alcala, 1998; Garcia–Charton et al., 2004). Recent empirical evidence suggests that establishing well–designed and managed marine reserves results in a rapid increase in the size and abundance of exploited species (Gell & Roberts, 2003; Lubchenco et al., 2003), thus reversing the detrimental effects of fishing (Dugan & Davis, 1993; Roberts & Hawkins, 2000). However, there is an increasing need to understand the long–term overall effectiveness of MPAs operating around the oceans (Pomeroy et al., 2005), as most studies assess different effects of MPAs over short time periods and at a local scale. Clearly, there is a need to assess the ability of MPAs to achieve their management objectives, taking into account the expectations of managers, monitoring needs and constraints (Pelletier et al., 2005). This study therefore aims to: (a) provide a synthesis of studies that have been carried out to evaluate the effects of MPAs in terms of their objectives; (b) identify areas concerning the use of study subjects, descriptors and the most commonly used methods of investigation; (c) analyse the different kinds of results on the effects of protection; and (d) reveal areas where our understanding is poor and future research is necessary.

This study is based on a comprehensive search of papers available through published literature, together with a classical bibliographical search, from which a database of specific research on MPA evaluation parameters was constructed. We made keyword searches using 'MPA(s)', 'indicators', 'ecological indicators' and 'social indicators'. References not published in journals were obtained through a classical search in several governmental institutions, research centres and universities. Great effort was put into obtaining technical reports, though there was considerable difficulty in obtaining these due to their restrictive circulation; for this reason, most came from the Southern European countries. A database was made including these fields: year of publication, reference type, evaluation type (peer reviewed or technical reports), location of the study, geographical area, sampling design, study subject, considered taxa, variables selected, sampling method used, if confirmation applied, main results obtained, significant differences found and whether reserve effects were detected. Where studies covered several topics, we designated the primary topics as those given most attention by the author. We took into account every descriptive parameter that was used to measure the effects of protection. Not all studies gave information for each field. Where more than one reference by the same author clearly presented the same information, only one was included in our database, with a peer reviewed paper given preference over a technical report. Papers selected by the search that were revisions, or did not provide quantitative data, were not included in the database. Each study was ranked by an expert panel of scientists that scored manuscripts by summing values assigned to sampling design quality, statistical analysis performed and type of editorial evaluation (table 1). Results Our search of the literature led to the selection of 224 studies conducted between 1983 and 2006. A higher proportion (peer reviewed: 70.98%, n = 159; technical reports: 29.02%, n = 65) of these studies were peer–reviewed papers that increased during the nineties (fig. 1). Most peer studies and technical reports reported significant differences between protected and unprotected areas (58.49% and 64.61%, respectively). Study purposes and subjects While the level of understanding in a particular subject area cannot be quantified in terms of absolute numbers of papers written, they do provide an indication of the extent of attention paid to the different subjects and highlight less studied areas that require further study. Both peer reviewed studies and technical reports mainly concentrated on the 'Effects on populations' and the 'Effects on assemblages', the latter being more frequent

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Table 1. Study rank according to their sampling design quality, statistical analysis performed and type of editorial evaluation. Tabla 1. Clasificación de estudios en categorías de acuerdo con su calidad de diseño de muestreo, con su análisis estadístico y su tipo de evaluación editorial. Samping design ranking mBACI

Rank Several protected vs. several unprotected

before/after replicated in space and time Beyond BACI

1 protected vs. 2 or more unprotected

before/after replicated in space and time

mACI

Several protected vs. several unprotected in

space and time (only after establishment)

ACI

1 protected vs. 2 or more unprotected in

space and time (only after establishment)

C vs. I

Protected vs. unprotected replicated only in

space

4 4 3 2

Protection levels (fully vs. buffer vs. general) replicated in space and time

Protection levels (fully vs. buffer vs. general) replicated in space

Protection levels (fully vs. buffer vs. general) replicated in time

Fixed transects (C vs. I) replicated in time

Others: no spatio–temporal replication

Statistical renking Statistically analysed

Statistically non–analysed

Publishing ranking Published (peer–reviewed)

Published (non–peer–reviewed)

Not–published

in peer reviewed studies (table 2). To a lesser extent, 'Effects on fishing yields' and 'Socioeconomic indirect effects' were also considered, while other topics such as spillover, direct socioeconomic effects and ecological indirect effects have rarely been addressed. In the review conducted, we could not find any 'larval exportation' and 'direct socioeconomic effects studies'. The study purposes used does not remain constant over time; peer reviewed publications assessed more different types of study purposes in the period from 1994 to 2006 (fig. 2). Both peer reviewed studies and technical reports used most 'all fishes' to assess the effects of protection (table 3). 'Commercial fishes' were most frequently used in peer–reviewed papers. There

is a lack of studies for 'charismatic species' and 'Exploitative uses', within peer reviewed journals. Most studies considered the biological effects of protection (89.28%, n = 200), though there were a few studies that analysed socioeconomic effects using biological subjects (e.g. number of contacts with key species). Fishes were by far the most widely used taxa, though in some cases molluscs, crustaceans, echinoderms, gorgonian, seagrasses and algae, amongst others, were also considered. In socioeconomic studies, divers were the most frequently used subjects. In some cases, socioeconomic studies also considered travel costs, contingent behaviour and prices that tourists would pay.

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80 Peer–reviewed publications

Number of publications

Technical reports

60 50 40 30 20 10 0

1983–85

1986–89

1990–93

1994–97 Years

1998–01

2002–04

2005–08

Fig. 1. Number of peer–reviewed publications and technical reports from 1983 to 2006. Fig. 1. Número de publicaciones revisadas por pares y de informes técnicos desde 1983 hasta 2006.

Variables selected

Study approach and confirmation

Parameters on 'population structure' were the most studied variables, followed by variables on 'Assemblage structure', while behavioural studies must be highlighted for being scarce in technical reports (table 4). In general, most of the reviewed studies (84.37%) used parameters to assess changes to key species, populations and/or habitats potentially affected by the protection. Almost all the reviewed studies used parameters on the evaluation of protection on some very restricted biological subjects, specifically those that assess the condition of some species and/or habitats.

Correlative studies were the most frequently applied to the assessment of the effects of protection, with only few peer–reviewed studies using an experimental approach. Most of the studies, both peer reviewed and technical reports, did not use any techniques to confirm whether the parameters assessed could be used in the future as indicators (Oreskes et al., 1994). Within those peer–reviewed studies and the technical reports that stated significant differences among the parameters evaluated, 70.96% and 78.57% respectively presented higher values of these parameters within the protected areas. Only 28.93% and 15.38% respectively exhibited non–significant differences. A considerable number of studies did not report back if any differences where found between protected and unprotected areas with the parameters studied. This can be due to an ineffective protection for several reasons (e.g. low effective protection, few protection years) but these studies where not very explicit on these reasons.

Sampling designs Sampling designs were grouped into fourteen different types (table 5). Most peer–reviewed studies used a 'control vs. impact' sampling design, while technical reports mostly used 'only in protected replicated in time and space' sampling. The percentage of studies having what we would consider a 'good quality' design was higher in peer reviewed studies, though even here these more complex and logically suitable sampling designs were scarce. Sampling designs incorporating spatial and/or temporal hierarchical replication were more frequent from 1994 onwards. Interestingly, our review highlighted a great frequency of peer–reviewed studies that had 'no replication' and/or were carried out only within the protected area. Although they appear in the first studies reviewed, they are also frequent in recent years (fig. 3).

Study ranking Peer–reviewed studies tended to have the highest ranking due to their higher quality sampling designs (table 6). Although most of the studies had some type of statistical analysis, peer–reviewed papers tended to use more rigorous methods and produce more quality results, a quality undoubtedly arising due to the review method applied.

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Table 2. Study purposes used by technical reports and peer–reviewed publications, in number (n) and frequency (%). Tabla 2. Número (n) y frecuencia (%) de los diseños de muestreo utilizados en los informes técnicos y en las publicaciones revisadas por pares. Technical reports

Peer–reviewed publications

Study purposes

Effects on populations

47.69

42.77

Effects on assemblages

20.00

30.19

Effects on habitats

1.54

6.92

Effects on fishing yield

13.85

5.66

Larval spillover

0.00

Adult spillover

0.00

5.66

Direct socioeconomic effects

0.00

0.63

Ecological indirect effects

0.00

3.77

Socioeconomic indirect effects

16.92

4.40

The results of the comparisons within peer–reviewed studies and the authors’ origin revealed that the studies best ranked were carried out by Australian–New Zealand authors, followed by North American authors (fig. 4). Papers from southern European researchers presented very heterogeneous values in the ranking. Discussion In this review, we highlight the differences among peer–reviewed studies and technical reports, primarily in terms of sampling designs and statistical analyses. Other fields considered only exhibited minimum quantitative differences between these two kinds of publications. Our results emphasize the very narrow range of methods and parameters used to assess the effects of protection in MPAs, as well as the low number of specific objectives proposed for MPAs tackled in the literature. Although our search for technical reports (the 'grey literature') was very thorough and particularly concentrated on MPAs reports, we consider that due to the difficulty of obtaining such information, there is a bias towards Southern European countries conducted in recent years (older reports are less likely to be listed online or available). For this reason, our results on grey literature should be restricted to this area since other areas are under–represented. Regarding peer–review papers, we consider our sampling to be a good representation of global research trends as they are readily available online and inter–country availability is not an issue. If a review lacks a comprehensive search strategy, it is likely to suffer from a degree of publication bias. On the other hand, when the results of research are negative (Hull, 1999; Underwood, 1999), they are usually not

published. As a consequence, reviews which fail to include these negative studies may overestimate the true effect of an intervention, resulting in false positive conclusions being drawn. It could be resolved if they had initially commissioned a comprehensive, systematic review of all the evidence (Scargle, 2000; Glasziou et al., 2001; Higgins & Green, 2005). Our review was undertaken with such a comprehensive search strategy to obtain every type of study, although if the studies were not published, they can’t be obtained and some bias may exist. MPAs have been considered a suitable management tool since the 1960s. However, in the early years, few studies were conducted and those that were did not suitably assess the benefits of MPAs or reserve effects or did not include all species that benefit from protection. Peer–reviewed publications increased mainly from the 1990s when MPAs became more popular, reaching 1306 declared MPAs around the world (Kelleher et al., 1995). The decline in papers after 2000 might simply reflect the delay between undertaking more complex research and getting it published or because this issue lost newness between the scientific community. Driving forces of past trends in MPA study Very few of the proposed objectives of MPAs have been examined (Jones, 1994; Rowley, 1994). Research effort has tended to concentrate on the conservation of biodiversity and fisheries resources, probably due to higher socioeconomic demand for such lines of study, but also possibly due to relative ease of study. This was already evidenced by IUCN, 2006 and has remained thus during subsequent years. In these areas uses are limited and within them the fishery. Therefore, one of the best descriptors is

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

1983–85

1986–89

Efects on populations Effects on fishing yield Direct social economic effects

1990–93

197

1994–97

1998–01

Effects on assemblages Larval spillover Ecological indrect effects

2002–04

2005–08

Effects on habitats Adult spillover Socioeconomic indirect effects

Fig. 2. Variation of the different types of study purposes for technical reports and peer–reviewed publications from 1983 to 2006: TR. Technical reports; P. Peer–reviewed publications. Fig. 2. Variación de los diferentes tipos de efectos estudiados en los informes técnicos y publicaciones revisadas por pares desde 1983 hasta 2006: TR. Informes técnicos; P. Publicaciones revisadas por pares.

related with the changes in fisheries as this would be the first response expected and logical to prove, and yet another reason for it to be the most studied. Other effects in MPAs are always going to be weaker and more difficult to demonstrate and study because they include more complex interactions. During this period of time, few other new study purposes were considered but without covering all possibilities or necessities. The reasons for this absence cannot be inferred from our review. We believe that, to a certain extent, research has gone this way due to inadequate funding of research and MPAs management. For example, Natura 2000 will receive only 3–5% of subsides for natural resources allocated by the European Union. At best, the figure may rise to US$18/ha/year and is quite far from the US$1,000/ha/years needed to finance marine parks (IUCN, 2006). At the same time, there is also a tendency for certain study lines to be favoured by the background and personal preferences of researchers and decision makers and to the lack of considerations of certain subjects. It is likely that both of these factors have led to research not covering the prime objectives of MPAs (e.g. the effects on fishing yield). Fishes and decapods have been well studied, as they are relatively easy to sample and because they are the first organisms that show changes due to protection, but other more complicated and costly subjects to study, such as 'the cascade effect as an ecological indirect effect due to protection' have been less well covered. The ease with which fish and crustaceans can be studied is obviously the reason

most population and/or assemblage descriptors, such as abundance, biomass, number of species and size, are the most commonly used, and explains the narrow range of generic methodologies used to record the data. Considering MPAs tend to be created using an ecosystem approach that takes into account the global links in the marine environment (Bohnsack, 1999), it seems that many studies fall into the error of excluding many potential study subjects and thus do not meet the study purposes. The complexity of the studies must increase searching for more complex and more difficult to prove interactions. This would explain the absence of studies evaluating the effects of protection at different stages and on components related with the design and functioning of MPAs: e.g. functional effects of enforcement on the management of MPAs, mitigate effects on the impacts, regulatory effects on main socioeconomic sectors affecting many coastal areas such as fishing and tourism. The opportunistic approach to the creation of many MPAs (McArdle, 1997; Roberts, 2005), in which neither size or the adequate scale needed for the MPA to accommodate the development of most of the species being protected are considered, has likely contributed to the inadequate state of research which is too general and vague. The politics of protection, which has centred mostly on species, is also deficient (Roberts, 2005). We therefore consider that MPAs research should contemplate a compromise in which study purposes are more evenly dispersed amongst the physical, bio–ecological and socioeconomic study lines. One

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Table 3. Study subjects used by technicl reports and peer–reviewed studies, in number (n) and frequency (%). Tabla 3. Número (n) y frecuencia (%) de los objetivos utilizados por los informes técnicos y publicaciones revisadas por pares. Technical reports

Peer–reviewed publications

Study subject

Algae

0.00

3.77

All invertebrates

4.62

10.06

Non–commercial Invertebrates

6.15

4.40

Commercial invertebrates

16.92

7.55

All fishes

40.00

32.08

Non–commercial fishes

3.08

3.77

Commercial fishes

12.31

27.04

Exploitative uses

0.00

Charismatic species

0.00

0.63

Non–exploitative uses

16.92

6.92

Others

0.00

3.77

of the main stumbling blocks for the study of MPAs has been the lack of suitable methodologies to cover complicated topics such as spillover. This deficiency is, in our opinion, partly due to the lack of adequate investment needed to develop new techniques that go beyond purely observational methods. The necessary added investment has only really been available in some regions such as North America, Australia, New Zealand and Japan (OECD, 2007) and is reflected in the higher ranking of studies coming from these areas. Of course, these areas were also pioneers of MPAs and therefore have a longer track record of such research. The differences in study ranking are also, in part, due to differences in experimental design training of researchers and the adoption of different study approaches, for example, the controversy between BACIPS (Stewart–Oaten et al., 1986) and BACI methodology (Green, 1979) or a more recent one, beyond–BACI (Underwood, 1991) and MBACI (Underwood, 1993). These methodological approaches have been embraced elsewhere over recent years. There is a gradual increase of a wide range of methodological approaches and the enlargement of those proved as more sturdy methodologies from a logical and statistical point of view. However, the results presented by technical studies are biased, as it was much easier to obtain studies of Southern Europe due to geographical proximity and Southeast Asia by being more available, obtaining results influenced by these events. However, the transfer of knowledge has not been as effective as the technical reports. Once again, the causes of this fact cannot be concluded directly from our review. Taking into account our experience, we can infer deficiencies in the training of the consult-

ants that made the technical reports, making studies without a minimum replication. Consultants involved in technical reports lack any previous educational training in research, and they have to face the job market only with the knowledge acquired at university; these gaps in the study curriculum for professionals and/ or the small investment of institutions limit transfer of skills between consultants, managers and scientists. MPAs research: the way forward As the number of MPAs and associated capital and social investment increases, it becomes more and more important for managers to base decisions on sound scientific and social knowledge. There is therefore a growing need for reliable information on the patterns, processes and ecological consequences that protection has on communities. While descriptive studies have contributed greatly to our understanding of the structure of biological communities harboured by MPAs, we do not yet know the effects of protection on parameters such as adult biomass export, larval spillover, etc. It is clear that such lines of study need to be encouraged. Monitoring is also essential to all MPA management programmes in order to be able to effectively track changes over time with a review of management decisions accordingly. This is particularly relevant if MPA managers are to meet their objectives (Kelleher et al., 1995). Not only is better funding necessary to maintain long term monitoring programs, but these should strive to homogenize methodologies in order to allow posteriori comparisons at different temporal and spatial scales, overall when national and transnational programs promoting MPAs exist.

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Table 4. Variables used by technical reports and peer–eviewed publications, in number (n) and frequency (%). Tabla 4. Número (n) y frecuencia (%) de las variables utilizadas en los informes técnicos y publicaciones revisadas por pares. Technical reports

Peer–reviewed publications

Variables

Population structure variables

56.92

52.20

Assemblage structure variables

20.00

35.22

Behaviour

0.00

3.14

Exploitative variables

7.69

2.52

Non–exploitative variables

13.85

1.89

Others

1.54

5.03

Enough time has passed to now be able to assess the state of MPA related science and to recommend selected data collection methods that robustly capture data on the effects of protection due to MPAs in the context of all their proposed objectives. While the geographic scope of data collection methodologies

is often initially designed for use at the local level, the use of consistent methodologies across larger regions is desirable. New protocols should outline steps necessary to obtaining environmental (including physical, biological conditions and ecological), fisheries and socioeconomic field, laboratory and

Table 5. Sampling designs used by technical reports and peer–reviewed publications, in number (n) and frequency (%). Tabla 5. Número (n) y frecuencia (%) de las variables utilizadas por los informes técnicos y publicaciones revisadas por pares. Technical reports

Peer–reviewed publications

Sampling design category

No spatial and/or temporal replication

1.54

12.58

Fixed transects (C vs. I) replicated in time

7.69

4.40

Only in protected levels replicated in time and space

38.46

13.84

Protected levels (fully vs. buffer vs. general) replicated in time

1.54

2.52

Protected levels (fully vs. buffer vs. general) replicated in space

1.54

8.81

Protected levels (fully vs. buffer vs. general) replicated in space and time

21.54

10.69

C vs. I

26.15

23.27

ACI

1.54

10.06

mACI

0.00

4.40

mACI time

0.00

6.29

BACI

0.00

BACI time

0.00

2.52

mBACI

0.00

mBACI time

0.00

0.63

200

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

Ojeda–Martínez et al.

1983–85

1986–89

1990–93

No spatial and/or temporal replication Only in protected levels replicated in time and space Protected levels (fully vs. buffer vs. general) replicated in space C vs. I mACI BACI mBACI

1994–97

1998–01

2002–04

2005–08

Fixed transects (C vs. I) replicated in time Protected levels (fully vs. buffer vs. general) replicated in time Protected levels (fully vs. buffer vs. general) replicated in space and time ACI mACI time BACI time mBACI time

Fig. 3. Variation of the different types of sampling design category for technical reports (TR) and peer– reviewed publications (P) from 1983 to 2006. Fig. 3. Variación de las categorías de diseño de los diferentes tipos de muestreo de informes técnicos (TR) y publicaciones revisadas por pares (P) desde 1983 hasta 2006.

office–based data relevant to management objectives and the health of the considered marine systems (Oakley, 2003). The use and selection of standardised protocols is not new and has previously been proposed at different forums (Goñi et al., 2000). It is essential that chosen indicators of the effects of protection are easily interpreted by managers and stakeholders so that they can contribute to efficient and transparent management (Mangi et al., 2007). At the same time, in view of the overlapping of different processes in space and time, it is essential that suitable scales for the evaluation of the effects of protection are chosen (Garcia–Charton et al., 2004; McClanahan et al., 2007). While it is true that selection of appropriate spatial and temporal scales used for detecting the effects of protection tends to be intuitive, it is essential that such sampling decisions are made with great care. An optimal strategy consists of studying the patterns of interest at multiple, simultaneous scales, identifying relevant scales of variability and then listing a series of hypotheses and testing them to account for the observed patterns (Underwood, 1997). This being the case, we are obliged to use increasingly complex sample designs requiring adequate spatial and temporal replication, with several control and impacted sites (Underwood & Chapman, 2003). Choice of these sites must be

Table 6. Rank obtained for technical reports and peer–reviewed publications, in number (n) and frequency (%). Tabla 6. Número (n) y frecuencia (%) obtenido en la clasificación de los informes técnicos y publicaciones revisadas por pares.

Technical reports

Peer–reviewed publications

Rank

0.00

1.26

18.46

6.92

15.38

12.58

44.62

5.66

16.92

27.67

4.62

23.27

0.00

18.87

0.00

3.77

Animal Biodiversity and Conservation 34.1 (2011)

ences. This scarcity is likely due both to researchers’ lack of awareness of its importance and the lack of reference information needed to confirm the assessed effects are due to the effects of protection. This highlights the need for long–term data series, preferably initiated before the time of protection (before data) and/or the need for good independent control sites. We must also change the tendency to only publish positive results in peer–reviewed journals, as this is clearly not beneficial for the correct interpretation of the effects of protection. In particular, the current trend to conduct meta–analyses on published MPA work is clearly weighted in favour of studies that show positive results for protection.

8 7 6 Rank

201

5 4 3 2 1

Conclusions A–NZ CA NA NE SAf SAm SE SAs Fig. 4. Rank of peer–reviewed publications related to the authors’ origin (A–NZ. Australian– New Zealand; CA. Central America; NA. North American; NE. North European; SAf. South African; SAm. South American; SE. South European; SAs. Southeast Asian. Fig. 4. Clasificación de las publicaciones revisadas por pares según el origen de los autores: A–NZ. Australia–Nueva Zelanda; CA. Centro América; NA. Norte América; NE. Norte Europa; SAf. Sur África; SAm. Sur América; SE. Sur de Europa; SAs. Sudeste Asiático.

made taking into account the dangers of pseudo replication (Hurlbert, 1984). The benefits lie in the power of resulting analyses and increased certainty of the results of protection (Underwood & Chapman, 2003). Wide heterogeneity has been detected in the papers studied, therefore the comparison and proper assessment of the politics used in MPAs as tools of management should be convenient to standardise methods of sampling when gathering field data. Using the same patterns would allow to compare and analyse the long term series and will ease the comparison in a wide range the local studies using meta–analysis, using either spatial (Mosquera et al., 2000; Coté & Reynolds, 2000) or temporal comparison (Ojeda–Martínez et al., 2007). Meta–analysis data originated from several independent studies can be analysed quantitatively, providing major advantages over traditional synthesis and reviews (Hedges & Olkin, 1985; Gurevitch & Hedges, 1993). It is noteworthy that confirmation, the process of verifying that any parameter used really responds to protection and not some other driving force, Oreskes et al., 1994, has been established in very few studies and is generally based on the comparison of results with historical long–term data or bibliographic refer-

Our review highlights the high heterogeneity among studies assessing the benefits of MPAs. A lot of emphasis is placed on the planning of MPAs and the evaluation of certain study purposes, study subjects and variables. Many parameters are studied but the study of the success of the protection itself is given less consideration. Insufficient attention is given to monitoring the extent to which MPAs achieve their objectives as a basis for taking action to improve management programs. Considering many of the studies are funded by the same institutions, there should be a concerted effort to require researchers to adopt standard methodological techniques that would allow widespread comparison and more cohesive management practices. Changes to the way we collect data and the questions asked by researchers are clearly required for effective, economically sound development of MPA policies. There is a pressing need for an integrated approach that treats MPAs as a whole instead of as a collection of separate biotic, social and economical entities. Acknowledgements This work was carried out with financial support from the Commission of the European Community, specific RTD program 'Specific Support to Policies', SSP–2003–006539 'European Marine Protected Areas as Tools for Fisheries Management and Conservation (EMPAFISH)'. It does not necessarily reflect its views and in no way anticipates the Commission’s future policy in this area. Thanks are due to Dr. Tim Dempster, SINTEF Norway and Dr. Fernando Tuya from Centre for Ecosystem Management, Edith Cowan University, for reviewing the English version. References Agardy, M. T., 1994. Advances in marine conservation: the role of marine protected areas. Trends in Ecology and Evolution, 9: 267–270. Bell, J. D., 1983. Effects of depth and marine reserve fishing restrictions on the structure of a

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Editor executiu / Editor ejecutivo / Executive Editor Joan Carles Senar

Secretaria de Redacció / Secretaría de Redacción / Editorial Office

Secretària de Redacció / Secretaria de Redacción / Managing Editor Montserrat Ferrer

Museu de Zoologia Passeig Picasso s/n 08003 Barcelona, Spain Tel. +34–93–3196912 Fax +34–93–3104999 E–mail mzbpubli@intercom.es

Consell Assessor / Consejo asesor / Advisory Board Oleguer Escolà Eulàlia Garcia Anna Omedes Josep Piqué Francesc Uribe

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Biodiversity and characterization of marine mycota from Portuguese waters E. Azevedo, M. F. Caeiro, R. Rebelo & M. Barata

Azevedo, E., Caeiro, M. F., Rebelo, R. & Barata, M., 2011. Biodiversity and characterization of marine mycota from Portuguese waters. Animal Biodiversity and Conservation, 34.1: 205–215. Abstract Biodiversity and characterization of marine mycota from Portuguese waters.— The occurrence, diversity and similarity of marine fungi detected by the sum of direct and indirect observations in Fagus sylvatica and Pinus pinaster baits submerged at two Portuguese marinas are analyzed and discussed. In comparison with the data already published in 2010, the higher number of specimens considered in this study led to the higher number of very frequent taxa for these environments and substrata; the significant difference in substrata and also in fungal diversity detected at the two environments is also highlighted, in addition to the decrease in fungal similarity. Because the identification of Lulworthia spp., Fusarium sp., Graphium sp., Phoma sp. and Stachybotrys sp. down to species level was not possible, based only on the morphological characterization, a molecular approach based on the amplification of the LSU rDNA region was performed with isolates of these fungi. This was achieved for three isolates, identified as Fusarium solani, Graphium eumorphum and Stachybotrys chartarum. To achieve this with the other isolates which are more complex taxa, the sequencing of more regions will be considered. Key words: Marine fungi, Wood baits, Fungal diversity, Ascomycota, Anamorphic fungi, Sequence alignment. Resumen Biodiversidad y caracterización de los hongos marinos de las aguas portuguesas.— Se analiza y discute la presencia, la diversidad y la similitud de los hongos marinos detectados mediante la suma de observaciones directas e indirectas utilizando cebos de Fagus sylvatica y Pinus pinaster sumergidos en dos puertos deportivos portugueses. En comparación con los datos ya publicados en 2010, el mayor número de especímenes aquí considerados condujo a un mayor número de taxones muy frecuentes en estos sustratos y medios ambientales; también debe destacarse la diferencia significativa en los sustratos y también en la diversidad fúngica en los dos medios ambientales, además de la disminución de la similitud fúngica. Dado que no fue posible la identificación de Lulworthia spp., Fusarium sp., Graphium sp., Phoma sp., y Stachybotrys sp. hasta el nivel de especie, basándose únicamente en la caracterización morfológica, se llevó a cabo un estudio molecular basado en la amplificación de la región LSU ADNr con extractos de dichos hongos. Ello se consiguió en tres extractos puros, identificados como de Fusarium solani, Graphium eumorphum y Stachybotrys chartarum. Para llevar a cabo este proceso con otros extractos puros pertenecientes a taxones más complejos, se considerará la secuenciación de más regiones. Palabras clave: Hongos marinos, Cebos de madera, Diversidad fúngica, Ascomycota, Hongos anamórficos, Alineación de secuencias. E. Azevedo, R. Rebelo, M. F. Caeiro & M. Barata, Fac. de Ciências, Univ. de Lisboa.– E. Azevedo, R. Rebelo & M. Barata, Centro de Biologia Ambiental (CBA), Fac. de Ciências, Univ. de Lisboa.– E. Azevedo & M. F. Caeiro, Centro de Estudos do Ambiente e do Mar (CESAM), Univ. de Aveiro. Corresponding author: Egidia Azevedo, Depto. de Biologia Vegetal, Fac. de Ciências, Univ. de Lisboa, Campo Grande, Edifício C2, 4º Piso, 1749-016 Lisboa, Portugal. E–mail: egazd@hotmail.com ISSN: 1578–665X

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Introduction Fungi have been known to exist in marine environments since early times. Hyde et al. (2000) highlighted the first reports of marine fungi up until 1846; however, interest in marine mycology only increased worldwide with Barghoorn & Linder (1944). In natural marine environments many substrata are good sources for marine fungi detection. The most studied have been wood substrata (Barghoorn & Linder, 1944; Koch, 1974; Koch & Petersen, 1996; Gonzálvez et al., 2001; Lintott & Lintott, 2002; Jones et al., 2006; Ravikumar et al., 2009), halophytes as Spartina spp. (Gessner & Kohlmeyer, 1977; Barata, 1997, 2002; Torzilli et al., 2006), Phragmites australis (Poon & Hyde, 1998; Wong & Hyde, 2002) and Juncus roemarianus (Kohlmeyer & Volkmann Kohlmeyer, 2001, 2002), as well as algae (Kohlmeyer & Volkmann–Kohlmeyer, 2003; Zucaro et al., 2008) and sea foam (Kohlmeyer & Kohlmeyer, 1979; Steinke & Lubke, 2005). Marine mycota associated to sand dunes plants (e.g. Arundo donax, Agropyron junceiforme and Ammophila arenaria) are poorly explored (Kohlmeyer & Kohlmeyer, 1979; Jones et al., 2009). Other substrata like corals, tropical sea grasses, crustacean and mollusk shells and soft rocks, are yet to be intensively investigated (Hyde et al., 2000; Jones et al., 2009). Considerable progress has been made in inventorying endophytes from marine hosts including seagrass (Alva et al., 2002; Sakayaroj et al., 2010). The diversity found comprises mostly anamorphic fungi and sterile mycelia and some isolates revealed to be producers of cellulases and xylanases (Alva et al., 2002). The baiting method, often used by mycologists for ecological studies, also yields pure cultures of marine fungi, representative of particular and/or selected environments (Vrijmoed et al., 1982, 1986; Alias & Jones, 2000; Azevedo et al., 2010). Pinus spp. and Fagus sylvatica are woods often used to inventory marine fungi in submerged conditions (Byrne & Jones, 1974; Grasso et al., 1990; Vrijmoed et al., 1982, 1986; Azevedo et al., 2010), because lignocellulosic substrata are colonized throughout submersion by a great variety of lignicolous species. Among the marine fungi isolated by Azevedo et al. (2010), five taxa could not be identified to species level based only on morphology: Lulworthia spp., Fusarium sp., Graphium sp., Phoma sp. and Stachybtrys sp. In temperate waters Lulworthia species are among the most frequently detected fungi in submerged woods (Byrne & Jones, 1974; Mouzouras et al., 1985; Grasso et al., 1985, 1990; Azevedo et al., 2010) and halophytes (e.g. S. maritima) (Barata, 1997). Several investigations reported Fusarium species in sediments, sand dunes and recovered submerged twigs of Tamarix aphylla (Jones et al., 2009), on coral reefs (Morrison–Gardiner, 2002) and in wood baits submerged in marine environment (Azevedo et al., 2010). Graphium species are also found in wood in marine environments (Vrijmoed et al., 1982, 1986; Gonzálvez et al., 1998; Maria

Azevedo et al.

& Sridhar, 2003; Azevedo et al., 2010). Phoma species are widespread, occurring in a variety of environments and ecological niches; they are less explored in marine environment in which Phoma species completely new to science are regularly found (Aveskamp et al., 2010). Finally, Stachybotrys species are also detected in marine environment (Landy & Jones, 2006; Jones et al., 2009; Azevedo et al., 2010), having been considered important by Jones et al. (2009) to document the occurrence of these taxa in the sea and to discover their ecological role. One goal of this work was to present and analyze, in a comprehensive manner, the data from the survey of Azevedo et al. (2010) concerning the occurrence, diversity and similarity of the marine mycota detected in wood baits before and after incubation in moist chambers (direct and indirect observations respectively). We further proposed to compare the results from this and other surveys carried out in temperate waters. A second goal was to present and discuss the results of a molecular approach performed to characterize the isolates that had not been possible to identify down to species level based only on morphological characters. Material and methods Sampling strategies Two marinas located on the western coast of Portugal, Cascais (38º 40' N 09º 25' E) and Sesimbra (38º 26' N 09º 06' W), were selected for the submersion of wood baits from Pinus pinaster Aiton and Fagus sylvatica L., as described by Azevedo et al. (2010) and shown in figure 1. The experimental design of baits is presented in table 1 and figure 1. The wood baiting technique involved a previous overnight soaking of the baits in distilled sterilized water followed by 20–minute autoclave sterilization at 121ºC. After submersion, collections were performed periodically each eight to 10 weeks, on a total of six collections, at each marina, The baits were examined as soon as possible after collection under the dissecting microscope to detect spores and fruit bodies. Microscopic characterizations were performed under the light microscope (Leitz Laborlux S with Normarski) in slides prepared with seawater as mounting media and microphotographs were taken (fig. 2). Thereafter, identifications were made following the dichotomous keys of Kohmeyer & Kolhmeyer (1979), Kohlmeyer & Volkmann–Kolhmeyer (1991) and Hyde & Sarma (2000). The baits were analyzed by direct observation and then incubated in moist chambers for 12 months. They were re–examined on a monthly basis, following the procedures described by Vrijmoed (2000). The isolates of marine fungi subjected to molecular analysis were obtained by the single spore method (Azevedo et al., 2010).

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1.5 mm

Fig. 1. A. Cascais marina; B. Sesimbra marina; C. Set of wood baits before submersion; D. Pinus pinaster bait colonized with marine organisms after six months of submersion; E. Box of wood baits at the moment of submersion; F. Fagus sylvatica bait colonized with basidiocarps of Nia vibrissa. Fig. 1. A. Puerto deportivo de Cascais; B. Puerto deportivo de Sesimbra; C. Serie de cebos de madera, antes de sumergirlos; D. Cebo de Pinus pinaster colonizado por organismos marinos después de seis meses de inmersión; E. Caja de cebos de madera en el momento de sumergirlos; F. Cebo de Fagus sylvatica colonizado con basidiocarpos de Nia vibrissa.

Analysis of fungal occurrence, diversity and similarity Frequencies of occurrence, expressed as percentages, were calculated taking the results from direct and indirect observations together. Marine fungi were classified as 'very frequent', 'frequent' or 'infrequent' based on Tan et al. (1989).

The average numbers of fungi per bait, species richness (S), Shannon (H’) and evenness (E) diversity indices, as well as the Sorenson similarity index (Cs), were calculated as described by Figueira & Barata (2007). The values of Shannon Index were compared applying a t–test as proposed by Hutcheson (Zar, 1999).

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Azevedo et al.

Table 1. Experimental design of the baits. Tabla 1. Diseño experimental de los cebos.

Total number of baits

288

Total number of baits in each marina

144

Total number of each type of wood bait 144 Total number of each type of wood baits in each marina Dimension of the baits Depth of submersion

72 20 x 20 x 60 mm 2m

Dates of submersions 20 XII 06 (Cascais) Dates of final collections

06 II 07 (Sesimbra) 07 I 08 (Cascais) 21 II 08 (Sesimbra)

DNA extraction, PCR amplification and sequencing The cultures selected for molecular analysis were grown in Malt extract broth prepared with sea water on a rotary shaker at 200 rpm for 6–15 days at 20ºC. Fungal biomass was harvested, washed three times with sterile distilled sea water and frozen in liquid nitrogen to be ground into a fine powder with a mortar and pestle. DNA was extracted following the instructions of Nucleospin Plant DNA extraction Kit (Machery–Nagel, Germany). A partial LSU rDNA sequence was amplified with LROR and LR5 primers (Viglays & Sun, 1994) and

PCR reactions were carried out in a total volume of 25 µl with Phire Hot Start DNA polymerase (Finnzymes Oy., now Thermo Scientific) and 1 µl DNA sample, following the manufacturer’ instructions. The amplification program consisted of an initial 3–minute denaturation step at 98ºC followed by 35 cycles of (i) denaturation (98ºC for 10''), (ii) annealing (58.5ºC for 10'') and (iii) elongation (72ºC for 30 '') and a final extension of 1' at 72ºC. After a sample being resolved on 0.7% agarose gel, PCR products were purified by Jet quick DNA Clean Up Kit (Genomed GmbH), according to the manufacturer’s instructions, and sent to be sequenced by a commercial lab. Direct sequencing was performed by STAB VIDA (Portugal), using the same set of primers and the the big dye terminator kit on ABI automated DNA sequencer. BioEdit Sequence Alignment Editor v7.0.9.0 (Hall, 1999) and ClustalW (Thompson et al., 1997) with default parameter settings were used for alignment and to obtain the consensus sequences. The obtained consensus sequences were compared to data in GenBank (National Center for Biotechnology Information, Bethesda, USA) online (www.ncbi.nih. gov), with GenBank BLASTn search engine. Results Marine fungi occurrence, diversity and similarity Table 2 presents the marine fungi detected by direct and indirect observations. The taxa are listed by decreasing values of frequency of occurrence in the ensemble of the two marinas; only infrequent fungi for both marinas were not listed. Diversity and similarity indices per environment and per substratum are presented respectively in tables 3 and 4.

Fig. 2. Lulworthia sp.: A. 15 day–old colony on corn meal agar made with 50% seawater; B. Ascocarp; C. Asci; D. Ascospores with conic apical chambers (arrow). Fusarium solani (JF746155): E. Eight day– old colony on potato dextrose agar (PDA) made with distilled water; F. Macroconidia with five septa; G. Monophialide with a slimy head of microconidia; H. Macro and microconidia. Graphium eumorphum (JF746156): I. 15 day–old colony on PDA; J. Synemmata; K. Annellidic cells with conidia (arrow); L. Conidia. Phoma sp. (JF746158): M. 15 day–old colony on PDA; N. Pycnidium; O, P. Conidia. Stachybotrys chartarum (JF746157): Q. Eight day–old colony on PDA; R. Rough dark conidiophore (arrow); S. Conidia in wet mass (arrow); T. Rough dark spores. Fig. 2. Lulworthia sp.: A. Colonia de 15 días de edad sobre harina de maíz agar hecho con 50% de agua de mar; B. Ascocarpo; C. Ascos; D. Ascosporas con cámaras apicales cónicas (flecha). Fusarium solani (JF746155): E. Colonia de ocho días de edad sobre agar papa dextrosa (PDA) hecho con agua destilada; F. Macroconidios con cinco septos; G. Monophialide con conidios agregados en una masa mucilaginosa; H. Macro y microconidios. Graphium eumorphum (JF746156): I. Colonia de 15 días de edad sobre PDA; J. Synemmata; K. Células anelídicas con conidios (flecha); L. Conidios. Phoma sp. (JF746158): M. Colonia de 15 días de edad sobre PDA; N. Picnidio; O, P. Conidios. Stachybotrys chartarum (JF746157): Q. Colonia de ocho días de edad sobre PDA; R. Conidióforo oscuro y rugoso (flecha); S. Conidios agregados en una masa mucilaginosa (flecha); T. Esporas oscuras y rugosas.

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252.86 µm E

51.67 µm G

31.25 µm P

43.75 µm S

25.0 µm

26.67 µm

43.75 µm

262.2 µm Q

98.3 µm

17.5 µm

36.67 µm

39.06 µm I

43.75 µm T

25.0 µm

29.02 µm

210

Azevedo et al.

Table 2. Frequency of occurrence of marine fungi (in %): C. Cascais marina (144 baits); S. Sesimbra marina (144 baits); C + S. Cascais + Sesimbra marinas (288 baits); Fs. Fagus sylvatica (144 baits); Pp. Pinus pinaster (144 baits). Tabla 2. Frecuencia de presencia de hongos marinos (en %): C. Puerto deportivo de Cascais (144 cebos); S. Puerto deportivo de Sesimbra (144 cebos); C + S. Puertos deportivos de Cascais + Sesimbra (288 cebos); Fs. Fagus sylvatica (144 cebos); Pp. Pinus pinaster (144 cebos). Environment

Substrata

C + S

Lulworthia sp.

71.88

74.31

69.44

97.92

45.14

Cirrenalia macrocephala (Kohlmer.) Meyers & Moore

46.18

43.06

49.31

13.19

79.17

Corollospora maritima Werdermann

36.81

41.67

31.94

27.78

45.83

Zalerion maritima Anastasiou

36.81

31.94

41.67

14.58

59.03

Cerisosporopsis halima Linder

33.33

44.44

22.22

28.47

38.19

Halosphaeria appendiculata Linder

29.51

37.50

21.53

46.53

12.50

Trichocladium achrasporum (Meyers & Moore) Dixon

17.01

15.28

18.75

3.47

28.86

Periconia prolifica Anastasiou

11.81

18.75

4.86

20.83

1.39

Remispora quadriremis (Hohnk) Kohlm.

10.07

9.72

10.42

Richness (S)

– 19.44 19

Total number of specimens

949

477

472

415

530

Average number of fungi per bait

3.30

3.31

3.28

2.88

3.68

Marine fungal diversity was higher at Sesimbra than at Cascais (table 3), the difference being highly significant for both types of baits: F. sylvatica (t322.2 = –3.73; P < 0.001) and P. pinaster (t521.2 = –4.49; P < 0.001). Considering the total number of baits, the fungal diversity was higher for P. pinaster than for F. sylvatica (t813.1 = – 2.46; P < 0.01) (table 3). This is a significant difference that was also observed separately in each marina: for Cascais (t453.8 = –3.66; P < 0.001) and for Sesimbra (t335 = –1.92; P < 0.05). Comparing the two marinas for mycota similarity, the Sorenson index presented a mean value for all analyzed situations (tables 3, 4), except for the comparison between the two types of baits submerged at Sesimbra marina (table 4). Lignicolous marine mycota occurrence in temperate locations Table 5 lists the very frequent and frequent marine fungi recorded in this and in other surveys carried out with submerged woods in temperate waters. Sequence analysis of the selected fungi Comparisons were made between partial sequences of the LSU rDNA region from our isolates and sequences from Genbank. Our sequences ranged between 886 and 915 base pairs. Concerning Lulworthia spp., the results from alignments and comparisons of sequences from selected

isolates are until now inconclusive to achieve species level (data not shown). Our isolate of Fusarium sp. (JF746155) shared 99% maximum identity with Fusarium solani (Mart.) Sacc. (EU719659, AY097317, AY097316 and FJ34532), as well as with Fusarium lichenicola C. Massal (AY097325) with query coverage of 99%. Our isolate of Graphium sp. (JF746156) evidenced maximum identities of 98% with Scedosporium apiospermum Sacc. ex Castell. & Chalm. (FJ345358) and 99% with Pseudallescheria boydii (Shear) McGinnis, A. A. Padhye & Ajello (AY882372) with query coverage of 100% and 95%, respectively. The consensus sequence of our isolate of Phoma sp. (JF746158) showed 98% of maximum identity with Loratospora aestuarii Kohlm. & Volkm.– Kohlm. (GU301838) and Coniothyrium obiones Jaap (DQ678054) with query coverage of 99%. When BLASTn was directed to 'Phoma', the results pointed out values of 98% maximum identity and 96% query coverage, with Phoma septicidalis Boerema (GQ387600, GQ387599, GQ387601), Phoma glaucispora (Delacr.) Noordel. & Boerema (GU238078), Phoma violicola P. Syd. (GU238156), Phoma fallens Sacc. (GU238074), Phoma vasinfecta Boerema, Gruyter & Kesteren (GU238151), Phoma dimorphospora (Speg.) Aa & Kesteren (GU238069), Phoma carteri Gruyter & Boerema (GQ387594, GQ387593), Phoma flavigena Constant. & Aa (GU238076), Phoma betae A. B. Frank (EU754179, EU754178), Phoma heteromorphospora Aa & Kesteren (EU754188, EU754187), and Phoma

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Table 3. Comparison of diversity indices per environment. Tabla 3. Comparación de los índices de diversidad por ambiente. Cascais

Sesimbra

Total (144 baits) 15

Cascais

Sesimbra

F. sylvatica (72 baits)

Cascais

Sesimbra

P. pinaster (72 baits) 12

H’

2.21

2.48

1.87

2.27

2.10

2.41

0.82

0.79

0.81

0.77

0.84

0.81

0.58

(J = 11, a = 15, b = 23)

0.64 (J = 9, a = 10, b = 18)

apiicola Kleb. (GQ387601) all of them with a query coverage of 96%. The isolate of Stachybotrys sp. (JF746157) shared 100% maximum identity with Stachybotrys chartarum (Ehrenh, ex Link) Hughes (AY489712) with query coverage of 99%. Discussion Marine mycota collected from wood baits submerged in temperate regions This analysis includes the total mycota detected on the survey of Azevedo et al. (2010). The data of frequency of occurrence highlight the increase of the very frequent fungi (four taxa) in relation to the results reported by Azevedo et al. (2010) because C. maritima and Z. maritima (very frequent fungi) and R. quadriremis (frequent fungus) were not detected by direct observation. The average number of fungi per Fagus sylvatica and Pinus pinaster baits increased respectively from

0.63 (J = 10, a = 12, b = 20)

1.70 to 2.88 and from 1.92 to 3.68. This shows how the incubation on moist chambers significantly contributed for the differentiation of reproductive structures from the marine fungi mycelia already present when direct observations were carried out (Azevedo et al., 2010; table 2). The diversity was significantly higher at Sesimbra than at Cascais and in P. pinaster than in F. sylvatica baits; a highly significant value was obtained when comparisons were done only with baits from Cascais. It is to be stressed that no significant differences were found between the two types of baits from Sesimbra marina when comparisons were made only with results of direct observations (Azevedo et al., 2010). The values of fungal similarity (Cs) decreased for all analyzed situations when compared with the results presented by Azevedo et al. (2010). This evidences the advantages of using different types of substrata and subjecting them to long incubation periods in order to achieve better inventories of marine fungal communities. Evenness values indicate that individuals recorded for each species were more evenly abundant in Cascais marina and for P. pinaster baits.

Table 4. Comparison of the diversity indices per substratum: Fs. Fagus sylvatica; Pp. Pinus pinaster. Tabla 4. Comparación de los índices de diversidad por sustrato: Fs. Fagus sylvatica; Pp. Pinus pinaster.

Cascais + Sesimbra

(144 baits) S

(72 baits)

Sesimbra

(72 baits)

H’

2.16

2.32

1.87

2.10

2.27

2.41

0.73

0.76

0.81

0.84

0.77

0.81

(144 baits)

Cascais

0.73 (J = 15, a = 19, b = 22)

0.64 (J = 7, a = 10, b = 12)

0.79 (J = 15, a = 18, b = 20)

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Azevedo et al.

Table 5. Very frequent and frequent marine fungi recorded in submerged wood at temperate locations: Fs. Fagus sylvatica; Pp. Pinus pinaster; Ps. Pinus sylvestris; Q. Quercus sp.; P. Populus sp.; L. Larix sp.; + Present; – Absent; In bold, exclusive taxa to baiting method; * This paper. Tabla 5. Hongos marinos frecuentes y muy frecuentes registrados en maderas sumergidas en áreas templadas: Fs. Fagus sylvatica; Pp. Pinus pinaster; Q. Quercus sp.; P. Populus sp.; L. Larix sp.; + Presente; – Ausente; en negritas, taxones exclusivos del método de los cebos; * Este estudio.

Taxa

Portugal

Ceriosporopsis halima Linder

England England +

–

Italy

–

Denmark –

Cirrenalia macrocephala (Kohlm.) Meyers & Moore

–

Corollospora maritima Werdermann +

–

Halosphaeria appendiculata Linder

–

Lulworthia fucicola Suth.

–

Lulworthia sp.

–

Marinospora calyptrata (Kohlm.) Cavaliere

–

Marinospora longissima (Kohlm.) Cavaliere

–

Monodictys pelagica (T. W. Johnson) Jones

–

Periconia prolifica Anastasiou *V

–

Remispora maritima Linder

–

Remispora quadriremis (Hohnk) Kohlm *

–

Trichocladium achrasporum (Meyers & Moore) Dixon *

–

Zalerion maritima (Linder) Anastasiou

–

Dictyosporium pelagicum (Linder) G. C. Hughes ex E. B. G. jones

–

Fs, Pp

Fs, Ps

Q, L

Harbour or marinas installations

Type of wood

Open sea waters Number of samples examined Richness (S) References

288 26 Azevedo et al. (2010)

Studies in temperate open coastal waters relative to wood inhabiting fungi are based both in submerged and in drift or intertidal wood. When comparing the results of the survey of Azevedo et al. (2010) with

–

134

–

145

1,440

Byrne Mouzouras Grasso Grasso Petersen & Jones

et al.

& Koch

(1974)

(1985)

(1990)

(1997)

other surveys carried out in temperate waters, differences found in fungal richness (table 5) could be due to the different nature of the woods used, to duration and depth of submersions in sea water and also to

Animal Biodiversity and Conservation 34.1 (2011)

different abiotic conditions (oxygen, temperature, salinity) to which the woods were subjected as well as to the number of analyzed samples. Lulworthia species were the most common fungi (present in five surveys), followed by Remispora maritima (observed in four surveys) C. maritima, H. appendiculata and M. pelagica (observed in three surveys) (table 5). The most common species can be considered species that play an important role in wood degradation (Alias & Jones, 2000). Additionally, considering the results expressed in table 2, it is to be emphasized that, for some of these taxa, there are references to production of enzymes and biocompounds. Bucher et al. (2004) reported production of cellulase, xylanase and peroxidase for one isolate of Lulworthia sp., and laccase for T. achrasporum. In relation to C. maritima, Jensen & Fenical (2002) found that an isolate of this fungus was able to produce a new secondary metabolite (Corollosporine) and Bucher et al. (2004) referred the production of cellulase and xylanase. Sequence analysis of the selected fungi The sequence data obtained suggest that our isolate of Fusarium sp. is closely related to Fusarium solani and F. lichenicola. However, the morphological characters are only compatible with the descriptions of Domsch & Gams (1980) and Samson et al. (2002) for F. solani as well as with the dichotomous key presented by Samson et al. (2002) for Fusarium species. Taking together morphological and molecular data, our isolate was considered to be Fusarium solani. Concerning Graphium sp., the result indicating identity with Scedosporium apiospermum was evaluated, although the morphological features of our isolate (figs. 2I, 2J, 2K, 2L) did not correspond with the description of this fungus (www.mycobank.org). The molecular results also revealed a close relation to the teleomorph Pseudollescheria boydii. It is worth nothing that Graphium eumorphum (Sacc.) is described as anamorph of this fungus (www.mycobank.org). The morphological features of our isolate are in accordance with the original description of Saccardo (www.indexfungorum.org), with slight differences on the length of conidia. For this reason, our isolate was considered Graphium eumorphum. For the isolate of Phoma sp., our molecular results pointed out members of two other genera (L. aestuarii and C. obiones) as well as 11 species of Phoma that have never been described for marine habitats (Jones et al., 2009) Eight of these species of Phoma are included in clade 7 (Leptosphaeriaceae and Pleosporaceae) in the study performed with 159 species of Phoma and its associated teleomorphs by Aveskamp et al. (2010). These authors recognize the complexity of this group, which is considered to be one of the largest fungal genera. This explains why a better identification of our isolate was not achieved, also because only one DNA region was accessed by sequence data. Concerning our isolate of Stachybotrys sp., comparisons of sequence data support the coincidence

213

found between our morphological characterization and the one made by Samson et al. (2002) for Stachybotrys chartarum (= S. atra corda). S. atra was referred by Jones et al. (2009) for marine environments, however indicating conidia dimensions slightly smaller. For this reason our molecular data were determinant in considering our isolate as Stachybotrys chartarum (= S. atra). Finally, regarding Lulworthia spp., our results pointed out the necessity of further analysis to accomplish the objective of characterizing the Portuguese isolates. The phylogenetic trees recently proposed for Lulworthiales comprise many isolates to be identified down to species level (Campbell et al., 2005; Jones et al., 2009) as well. We intend to contribute for the establishment of phylogenetic relationships within this taxon with the molecular characterization (still currently underway) of our isolates. In conclusion, this molecular approach pursuing a contribution for the identification of these Portuguese isolates down to species level showed to be valuable as this goal could be achieved for three of them (Fusarium solani, Graphium eumorphum and Stachybotrys chartarum). However, the sequencing of more regions always allows more accurate results. This procedure will be mandatory for more complex taxa such as Lulworthia spp., and Phoma sp., those that, in this study, remain to be better characterized. Acknowledgements We thank Francisco Caeiro for reviewing this manuscript. This work was partially funded by CBA and CESAM. References Alias, S. A. & Jones, E. B. G., 2000. Colonization of mangrove wood by marine fungi at Kuala Selangor stand, Malaysia. Fungal Diversity, 5: 9–21. Alva, P., Mckenzie, E. H. C., Poiting, S. B., Pena– Murrala, R. & Hyde, K. D., 2002. Do sea grasses harbour endophytes? In: Fungi marine environments: 167–178 (K. D. Hyde, Ed.). Fungal Diversity Press, Hong Kong. Aveskamp, M. M., De Gruyter, J., Woudenberg, J. H. C., Verkley, G. J. M. & Crous, P. W., 2010. Highlights of the Didymellaceae: A polyphasic approach to characterise Phoma and related pleosporalean genera. Studies in Mycology, 65: 1–60. Azevedo, E., Rebelo, R., Caeiro, M. F. & Barata, M., 2010. Diversity and richness of marine fungi on two Portuguese marinas. Nova Hedwigia, 90(3–4): 521–531. Barata, M., 1997. Fungos marinhos superiores associados com Spartina maritima em estuários da Costa Portuguesa. Ph. D. Thesis, Lisbon Univ. – 2002. Fungi in the Halophyte Spartina maritima in salt marsh. In: Fungi marine environments: 179–193 (K. D. Hyde, Ed.). Fungal Diversity Press,

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Succession of fungi on wood Avicennia alba and A. lanata in Singapore. Canadian Journal Botany, 67: 2686–2691. Thompson, J. D., Gibson, T. J., Plewniak, F., Jeanmougin, F. & Higgins, D. G., 1997. The CLUSTALX windows interface: flexible strategies for multiple sequence alignment aided by quality analysis tools, Nucleic Acids Research, 25: 4876–4882. Torzilli, A. P., Sikaroodi, M., Chalkley, D. & Gillevet, P. M., 2006. A comparison of fungal communities four salt marsh plant using automated ribosomal intergenic spacer analysis (ARISA). Mycologia, 98(5): 690–698. Vrijmoed, L. L. P., 2000. Isolation and culture of higher fungi In: Marine Mycology – a practical approach: 1–20 (K. D. Hyde & S. B. Poiting, Eds.). Fungal Diversity Press, Hong Kong. Vrijmoed, L. L. P., Hodgkiss, I. J. & Thrower, L. B.,

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Editor executiu / Editor ejecutivo / Executive Editor Joan Carles Senar

Secretaria de Redacció / Secretaría de Redacción / Editorial Office

Secretària de Redacció / Secretaria de Redacción / Managing Editor Montserrat Ferrer

Museu de Zoologia Passeig Picasso s/n 08003 Barcelona, Spain Tel. +34–93–3196912 Fax +34–93–3104999 E–mail mzbpubli@intercom.es

Consell Assessor / Consejo asesor / Advisory Board Oleguer Escolà Eulàlia Garcia Anna Omedes Josep Piqué Francesc Uribe

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Elysia timida (Risso, 1818) three decades of research F. Giménez–Casalduero, C. Muniain, M. González–Wangüemert & A. Garrote–Moreno

Giménez–Casalduero, F., Muniain, C., González–Wangüemert, M., Garrote–Moreno, A., 2011. Elysia timida (Risso, 1818) three decades of research. Animal Biodiversity and Conservation, 34.1: 217–227. Abstract Elysia timida (Risso, 1818) three decades of research.— During the last 30 years, studies on Elysia timida (Risso, 1818) have addressed various aspects related to food sources, photosynthetic efficiency of kleptoplasts, population genetics, chemical ecology and reproductive biology, both in the Mediterranean Sea and in the Mar Menor coastal lagoon. E. timida shows a strong specific interaction with Acetabularia acetabulum, retaining functional chloroplasts for at least 45 days and obtaining extra energy in periods when food resources are scarce. It shows control of parapodia, avoiding pigment photodestruction under oversaturated light conditions. The chemical ecological relationships established between E. timida and its potential predator fish, Thalassoma pavo, have also been evaluated, and it has been found that that the extracts of the mollusc contain repellent and unpalatable polypropionate compounds. Population genetics has demonstrated the genetic divergence between populations showing high and significant values of FST and genetic distances, and at least six privative alleles that are not shared with Mediterranean populations have been detected in lagoon populations. This sacoglossan is a poecilogonic species, and its lagoon populations show a greater reproductive output than Mediterranean populations; they produce a greater number of egg masses and embyros per individual, and the capsules have a wider diameter. Key words: Elysia timida, Kleptoplasts, Environmental stress, Chemicals ecology, Genetic divergence, Poecilogonic specie. Resumen Elysia timida (Risso, 1818) tres décadas de investigación.— Durante los últimos 30 años los estudios sobre Elysia timida (Risso 1818) han abordado diversos aspectos relacionados con sus fuentes de alimentación, la eficacia fotosintética de los cleptoplastos, la genética de poblaciones, la ecología química y la biología reproductiva, tanto en Mar Mediterráno como en la laguna costera del Mar Menor. E. timida presenta una fuerte interacción específica con Acetabularia acetabulum, reteniendo los cloroplastos funcionales durante al menos 45 días y obteniendo energía extra durante los periodos en que los recursos alimentarios escasean. Mediante el control de los parapodios evita la fotodestrucción de los pigmentos en condiciones de sobresaturación lumínica. También se han evaluado las relaciones ecológicas y químicas entre E. timida y su depredador potencial, el pez Thalassoma pavo, detectándose que los extractos del molusco contienen componentes polipropionados que son repelentes y de gusto desagradable. La genética de poblaciones ha demostrado la existencia de divergencia genética entre las poblaciones, presentando valores altos y significativos de FST y de distancias genéticas, detectándose al menos seis alelos privativos en las poblaciones lagunares los cuales no son compartidos por las poblaciones del Mediterráneo. Este sacogloso es una especie poecilogónica, y sus poblaciones de la laguna muestran un mayor esfuerzo reproductivo que las poblaciones mediterráneas; producen un número mayor de masas de huevos y de embriones por individuo, y las cápsulas tienen un mayor diámetro. Palabras clave: Elysia timida, Cleptoplastos, Estrés ambiental, Ecología química, Divergencia genética, Especie poecilogónica. F. Giménez–Casalduero & A. Garrote–Moreno, Dpto. de Ciencias del Mar y Biología Aplicada, Univ. de Alicante, Campus de Sant Vicent del Raspeig, Apartado 99, E–03080 Alicante; C. Muniain, Inst. de Investigación e Ingeniería Ambiental, Univ. de San Martín, Buenos Aires, Argentina; M. González–Wangüemert, Centro de Ciências do Mar (CCMAR), Univ. do Algarve, Campus de Gambelas, 8005–139 Faro, Portugal. Corresponding author: F. Giménez–Casalduero. E–mail: francisca.gimenez@ua.es ISSN: 1578–665X

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35 30 25 20 15 10

2006–2010

2001–2005

1996–2000

5 1991–1995

E. timida (Risso, 1818) is a gregarious opisthobranch mollusc that lives in shallow waters characterised by low–energy hydrodynamic conditions. These molluscs seek light in areas covered by stones or a thick layer of sand (Bouchet, 1984; Marín & Ros, 1988, 1992; Thompson & Jaklin, 1988; Giménez–Casalduero et al., 2002; Giménez–Casalduero & Muniain, 2006). They share these environments with the chlorophyceae algae, Acetabularia acetabulum (Linneaus, 1758), along the Mediterranean coast, including coastal lagoons (Ortea et al., 1997). E. timida occurs naturally in populations feeding on the fleshy, siphoned green alga, A. acetabulum, although it has been linked with other algae that are commonly found in photophilic environments such as Padina pavonia (Linneaus) (Bouchet, 1984; Ballesteros, 1985), or Ulva sp., Bryopsis sp., Enteromorpha sp., and also phaeophyceae algae, such as Halopteris filicina (Grateloup) and Colpomenia sinuosa (Mertens ex Roth) (Rahat, 1976). However, A. acetabulum is its optimal diet (Ros & Rodriguez, 1985; Marín, 1988; Marín & Ros, 1987, 1989, 1991, 1993). Some authors have suggested a co–evolution between the two species (Marín & Ros, 1992, 2004). E. timida populations associated with A. acetabulum meadows exceed densities of about 6 ind./m2 (Marín & Ros, 1992). This gregarious nature is linked to a limited dispersion ability and a strong habitat preference. Lagoon environments are sometimes under conditions of extreme environmental stress due to significant changes in salinity and temperature, which may cause strong selective pressures on organisms (Gamito et al., 2005; González–Wanguëmert et al., 2004, 2006, 2009). E. timida has been recorded from the Mediterranean Sea, Gibraltar Strait, Canary Islands, Cape Verde Island and Sao Tomé Island, including coastal lagoons (Bouchet, 1984; Ballesteros, 1979; García– Gómez, 2002; Giménez–Casalduero, 1997a, 1997b, 1999; Marín & Ros, 1987, 1988, 1991, 1992; Ortea et al., 1997; Rahat, 1976; Ros, 1976, 1977; Swennen, 1961; Templado, 1982; Thompson & Jaklin, 1988; Türkmen & Demirsoy, 2009; Wirtz & Anker, 2009). However, if we take into account the low dispersal

1986–1990

The sacoglossan: habitat, interaction and geographical distribution

The sacoglossan show a degree of feeding specialization that is similar to that of terrestrial insects. E. timida feeds with priority on the green alga, A. acetabulum. In the field, these animals are white and are conspicuous in winter or spring, but they are very cryptic when the algal food is calcified (Marín & Ros, 1992). The white colour of E. timida could serve two main purposes: chloroplasts exploitation and predation avoidance. The animal uses the white of the parapodia to control the amount of light used for the photosynthesis of kleptoplasts. The design of the mollusc coloration may appear relatively cryptic during periods of algal calcification (summer: white animal on white background) and relatively conspicuous during periods characterized by low algal calcification (autumn: white animal on green background). Experiments conducted on feeding preference (Giménez–Casalduero, 1997b), have shown that the white models were more cryptic in the summer background than in the autumn background (fig. 2). This suggests that the optimal colour for a possible prey is white rather than green or red, in particular for E. timida, bearing in mind its defensive characteristics

1981–1985

Results

Aposematic versus cryptic coloration

1976–1980

In recent decades, the sacoglossan Elysia timida (Risso, 1818) (Gastropoda, Opisthobranchia, Sacoglossa) has attracted the attention of numerous researchers (fig 1). The reasons for this growing interest are the easy accessibility of its habitat, the large number of individuals in a small areas (due to its gregarious behaviour associated with a low dispersal ability), the extraordinary physiological and ecological adaptations (great versatility and adaptability to stress conditions) and particular reproductive and survival strategies.

ability of this species and the genetic differentiation results found among populations at a small spatial scale (González–Wanguëmert et al., 2006), the amphi–Atlantic status of E. timida, cited from the Western Atlantic coasts by Ortea et al. (1997) and Valdés et al. (2006), this does not seem consistent. It is more likely a new variety in the Western Atlantic, although this suggestion needs to be corroborated genetically.

Number of records

Introduction

Giménez–Casalduero et al.

Fig. 1. Accumulative numbers of Elysia timida manuscripts pooled in five–year periods. Fig. 1. Número acumulativo de manuscritos de Elysia timida, agrupados en períodos de cinco años.

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Green

Number of attacks

80 70

Red

60 50

White

40 30 20 10 0 Summer

Spring

Autumn

Fig. 2. Crypsis degree measured as number of attacks of the green, red and white artificial models over the three backgrounds (autumn, spring and summer), which represents the seasonal change of the algae A. acetabulum (modified from Giménez–Casalduero, 1997b). Fig. 2. Grado de cripsis medido como el número de ataques a los modelos artificiales verdes, rojos y blancos, sobre los tres fondos (otoño, primavera y verano), los cuales representan el cambio estacional del alga A. acetabulum (modificado de Giménez–Casalduero, 1997b).

Number of models eaten

and population dynamics. The colour design of a prey is considered to be cryptic when the potential predator cannot distinguish it from the sea floor on which it lies.

This strategy is very common in many elysiidae, one example being Bosellia mimetica (Trinchese, 1891), which feeds on Halimeda tuna (Ellis et Solander) and

Green

Red

White

10 5 0

Summer

Spring

Autumn

Fig. 3. The ability of algae to refuge measured as number of attacks of the green, red and white artificial models over the three backgrounds (autumn, spring and summer), which represent the seasonal change of the algae A. acetabulum. In this experiment the background was three–dimensional, simulating the algae structure with white plastic stick; the density of the stick was proportional to the alga density during the three seasons simulated (modified from Giménez–Casalduero, 1997b). Fig. 3. Capacidad de refugio de las algas, medido como número de ataques a los modelos artificiales verdes, rojos y blancos, en las tres estaciones (otoño, primavera y verano), los cuales representan el cambio estacional del alga A. acetabulum. En este experimento el fondo fue tridimensional, simulando la estructura de las algas mediante palitos blancos de plástico; la densidad de los palitos fue proporcional a la densidad del alga durante las tres estaciones simuladas (modificado de Giménez–Casalduero, 1997b).

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is virtually indistinguishable. Aposematic coloration is used as a warning to potential predators of their toxic or unpleasantness. The close relationship between E. timida and A. acetabulum, apparently favours a design that can be either cryptic or aposematic. Gendron & Staddon (1983) showed that if the probability of detecting a prey is inversely related to both the search rate and the degree of crypsis, specialist herbivores could be subject to greater selective pressure by natural enemies than generalist animals. Laboratory experiments show that E. timida is less consumed when the calcified algal biomass increases (fig. 3) (Giménez–Casalduero, 1997b). The nutritional quality decreases, but it provides an effective refuge to avoid predation (Hacker & Steneck, 1990; Duffy & Hay, 1991). Some sacoglossans however have adapted their diet to algae with chemical defence, because they subsequently avoid predation and accidental ingestion by macro–herbivores, or because the sequestered secondary metabolites from the algae protect them from their own predators (Hay, 1991; Hay et al., 1994) A. acetabulum contains no deterrent metabolites and E. timida must synthesize its own defence (Gavagnin et al., 1994). Photosynthetic efficiency of kleptoplasts The sacoglossans are specialized herbivores that can retain chloroplasts intact within the cells of the gastrointestinal tract (Marín & Ros, 1988). The manipulation of unchanged foreign structures is a complex mechanism involving both cell recognition and acceptance, and physiological adaptation within the digestive glands. Also, its operation as a photosynthetic source is an extraordinary evolutionary feature (Marín & Ros, 1991; Rumpho et al., 2000, 2001). From the point of view of chloroplasts’ autonomy it is interesting to note the time period that chloroplasts retain their normal functions within their host cells. Several authors have demonstrated the photosynthetic capacity of E. timida that consists of their using symbiont chloroplasts (kleptoplasts) retained in their intestinal diverticula (Ros & Rodriguez, 1985; Marín & Ros, 1989; Wägele & Johnsen, 2001; Evertsen et al., 2007; Händerler et al., 2009, 2010). In addition, Greene (1970) and Greene & Muscatine (1972) not only described the release of organic compounds in chloroplast symbionts but noted that the metabolism of symbiotic chloroplasts can be modified by the association. They found abundant evidence showing the transfer of photosynthates from the chloroplast to the host tissue sacoglossans. In the case of isolated Acetabularia, chloroplasts have been reported that C14 labelled only chlorophyll joins the first 70', after which it is extracted from the algal tissue (Trench & Smith, 1970). The chloroplast–animal association is unable to synthesize chlorophyll, glycolipids, ribulose–bisphosphate carboxylase or membrane proteins. Therefore chloroplasts are unable to grow or have a real division when they are inside the host animal (Trench & Ohlhorst, 1976). This inability to synthesize DNA and RNA or regenerate plastidial proteins means that

Giménez–Casalduero et al.

elysoidea must obtain chloroplasts in each generation (Marín & Ros, 1993; Wägele et al., 2010). Effectiveness in this symbiotic system depends on the ability of the host to incorporate active chloroplasts and retain them functionally active in its glands for as long as possible (a high turnover rate would not provide benefits). Holding capacity of active chloroplasts is revealed by analysing the changes in the proportion of chlorophyll a over time compared to other pigments within the body of sacoglossa. Chlorophyll a has been described as one of least stable pigments over time as compared to carotenes or degradation products from chlorophyll, which have an extraordinary resistance to destruction. The experiment carried out by Giménez–Casalduero (1997b) on E. timida showed a lower degradation rate for chlorophyll than for the remaining pigments during the experimental time (starved and dark conditions) (fig. 4). However, the absence of light seems to be a determining factor for the incorporation of new chloroplasts. This result was interpreted as the ability of sacoglossa to assimilate or eliminate the remaining pigments, because E. timida tends to retain and maintain high levels of chlorophyll a, declining its net concentration over time more slowly than the other pigments. In an effective system, the retained chlorophyll must endure being photosynthetically active for long periods of time. Gross oxygen production per milligram of chlorophyll reaches values between 2.910 and 3.397 mg O2/mg of chlorophyll a in E. timida tissues and these levels are maintained over time, even in starvation periods (table 1) (Giménez–Casalduero, 1997b) Metabolic benefits and consequences As mentioned above, the chloroplasts retained in the intestinal diverticula of E. timida are able to perform photosynthesis (Ros & Rodriguez, 1985; Marín & Ros, 1989). The use of photosynthetic energy has also been demonstrated by Giménez–Casalduero & Muniain research (2006, 2008). These latter works studied the photosynthetic parameters of E. timida in the Mar Menor lagoon environment, and the researchers found that the results of P–I curves were explained by the model based on the kinetics of Michaelis–Menten (1913) and described by Pérez (1989). This model showed a rapid saturation of the photosynthetic apparatus at relatively low irradiance values, seeming to contradict the animal habitat (shallow and lit areas). However, this fast saturation might be influenced by a control of parapodia, which would avoid pigment photodestruction under oversaturated light conditions. During the experimental period (Giménez–Casalduero & Muniain, 2006), E. timida individuals were observed closing and opening their parapodia, illustrating the ability of sacoglossans to regulate the photosynthetic production described by Rahat & Monseline (1979), Monseline & Rahat (1980) and Jesus et al. (2010): parapodia were completely opened when individuals were exposed to low irradiance and gradually closed with increasing the irradiance values. Furthermore, in the field, the sacoglossan is commonly found

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Chlorophyll a / other pigments

26 24 22 20 LF

18 Days

Fig. 4. Rate of chlorophyll a versus other photosynthetic pigments (± SE) in the tissue of E. timida specimens kept in a laboratory for 28 days at a constant temperature (25ºC) with four treatments: LF. Light photoperiod 12:12 light and darkness feeding A. acetabulum; LS. Light photoperiod 12:12 light and darkness without food; DF. Continuous darkness feeding A. acetabulum; DS. Continuous darkness without food (modified from Giménez–Casalduero, 1997b). Fig. 4. Proporción de clorofila a frente a otros pigmentos fotosintéticos (± ES) en el tejido de especímenes de E. timida mantenidos en el laboratorio durante 28 días a una temperatura constante (25ºC) con cuatro tratamientos distintos: LF. Fotoperiodo lumínico 12:12 de luz y oscuridad alimentándose de A. acetabulum; LS. Fotoperiodo lumínico 12:12 de luz y oscuridad sin alimento; DF. Oscuridad continuada aimentándose a base de A. acetabulum; DS. Oscuridad continuada sin alimento (modificado de Giménez–Casalduero, 1997b).

with completely closed parapodia on a well–lit day (Giménez–Casalduero & Muniain, 2006). The production versus respiration relationship (P/R) was initially used to estimate compensation irradiance for symbiotic associations between corals and zooxanthellae: (i) P/R = 1 indicates that most of the fixed carbon during photosynthesis is consumed on respiration by the host and the 'guest'; (ii) P/R > 1 means that the amount of fixed carbon by photosynthesis exceeds the basal metabolic requirements of the host; (iii) P/R < 1 indicates that the host has to feed on other sources to fulfil the amount of carbon required for its basal metabolic needs (McCloskey et al., 1978). Theoretically, the estimated mean P/R values of E. timida populations from the Mar Menor coastal lagoon are above 0.91, these values are considered high in the coral–zooxanthellae symbioses. If we consider that sacoglossans are more active than corals and possess higher respiration values, then production values of chloroplasts should also be higher to obtain similar indices. Therefore, photosynthetic efficiency of symbiont chloroplasts on E. timida is fairly high. The parameters of the P–I curve obtained by Giménez–Casalduero & Muniain (2006) provide a useful baseline information to perform further studies on photosynthetic energy generated from the relationship between the mollusc and the

acquired chloroplasts, as well as on the importance of such energy for primary metabolism of lagoon sacoglossans (Giménez–Casalduero & Muniain, 2006; Jesus et al., 2010). Poecilogony is an intraspecific variation in the mode of larval development and can be found in single or different populations (Bouchet, 1989; Krug, 1998; Krug et al., 2007). E. timida is considered a poecilogonic species showing direct and lecitotrophic development in the Mediterranean area (Rahat, 1976; Marín & Ros, 1993; Garrote–Moreno, 2007). Direct development increases the likelihood of offspring survival in the absence of food limitation. However when the food is scarce, lecitotrophic development allows exploration of new areas with A. acetabulum. So, is there coevolution between E. timida and A. acetabulum? Or, conversely, is it an adaptive response of E. timida to environmental factors, such as the greater or lesser availability of food? Energy invested in reproduction is also affected by the presence of 'extra energy' from symbiotic chloroplasts. In general, the number of spawn is greater if animals have extra energy input; a disturbance of normal conditions initiates a massive reproductive activity and the number of eggs per spawn is greater in starved conditions (Giménez–Casalduero & Muniain, 2008).

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Table 1. Gross oxygen production (GOP, mg O2/ mg chlorophyll a) in Elysia timida for specimens kept in a laboratory under starved conditions for 28 days at a constant temperature (25ºC) with two treatments: GOP–L. With light photoperiod (12:12 light and darkness); GOP–D. Continuous darkness (modified from Giménez–Casalduero, 1997b). Tabla 1. Producción total de oxígeno (GOP, mg O2/mg de clorofila a) en Elysia timida, para especímenes mantenidos en el laboratorio sin recibir alimento durante 28 días a una temperatura constante (25ºC) bajo dos tratamientos: GOP–L. Con fotoperiodo lumínico (12:12 luz y oscuridad); GOP–D. Oscuridad continuada (modificado de Giménez– Casalduero, 1997b). Days 0

GOP–L

GOP–D

2.91

2.25

1.31

2.47

0.68

3.397

1.58

The sacoglossan reaction in a starved situation is to invest energy in reproduction, using a mating behaviour and a particularly highly efficient co–occurrent sperm transfer (Schmitt et al., 2007). Reproductive effort is very high the first few days, using both photosynthetic and metabolic energy (which explains a decrease of up to 20% in size the first 18 days) (Giménez–Casalduero, 2008). Adaptations to environmental stress conditions Estuaries and coastal lagoons constitute transitional environments whose main characteristic is the instability of their physical–chemical parameters, mainly the concentration of salt (Cognetti & Maltagliati, 2000; Gamito et al., 2005). Physiological adaptations are very important for survival in these extreme environments. These adaptations imply a high energetic cost that is reflected in growth and reproduction rates; thus genetic selection aimed at minimizing such costs can be assumed (Wright, 1977; Remmert, 1988). When environmental conditions change unpredictably in space and time, variation in life–history traits can be an adaptive response to selection (Meyers & Bull, 2002). For instance, stable dispersal dimorphisms can evolve when the quality of habitat patches varies over time and there is spatial heterogeneity in environmental fluctuations (Mathias et al., 2001). Two morphs of E. timida have been described in the Mar Menor coastal lagoon and shallow environments of the Southwest Mediterranean coast, showing differences in body

size, colour and chlorophyll concentration (Giménez– Casalduero, 1997a). It is important to stress, hewever, that differences in these parameters have also been found between lagoon (regardless of morphotype) and Mediterranean populations (Giménez–Casalduero, 1999). For example, simultaneously comparing the mean dry weight and mean chlorophyll a concentration from two lagoon morphotypes and the Mediterranean population, we recognized differences between all of them (fig. 5). These differences have recently been confirmed by Jesus et al. (2010). Moreover, changes have been observed in the spawn of populations from Mar Menor lagoon and the Mediterranean Sea, with the reproductive effort in lagoon sacoglossa being higher than in Mediterranean populations (Giménez–Casalduero, 1997b, 1999). The variability in the number of eggs per spawn is smaller in the Mediterranean population than the lagoon one which varies from 88 to 725 eggs per spawn (fig. 6). In some species, such as Elysia viridis (Montagu, 1804) the increase in reproductive success may result from a change in salinity of environment (Hagerman, 1970). High salinity of the Mar Menor lagoon could influence the increase in E. timida reproductive output, but it does not explain the differences among environments or among morphs. According to Giménez–Casalduero (1997a, 1997b), the differences between brown and green lagoon animals, could have diverse explanations: i) We could be facing a new variety of E. timida, considering the Mar Menor lagoon as a shelter for this species (high salinity, extreme temperature changes, isolation from the Mediterranean Sea, few predators) and the capacity of this elysoidea to have a direct larval development and being a strong candidate to develop genetic adaptations in lagoon conditions; ii) Another possible scenario would be the higher versatility converning food in this species, with a great capacity to adapt to new food resources. González–Wangüemert et al. (2006) estimated the degree of genetic divergence between populations living inside and outside the Mar Menor coastal lagoon. They showed genetic identity values ranging from 0.9 to 0.87 in the Mediterranean and coastal lagoon populations. These values indicated a subspecies–rank separation according to several authors (Avise, 1974; Thorpe, 1983; Mariani et al., 2002). However, it is important to stress that there are different methods of delimiting species and subspecies (Sites & Marshal, 2003). However, significant genetic differences (FST values and Nei’s genetic distances) were found between lagoon and marine populations, confirmed by the principal component analysis. This genetic differentiation (Mar Menor/ Mediterranean Sea populations) has been confirmed using other molecular markers and species (González–Wangüemert & Pérez–Ruzafa, in review; Vergara–Chen et al., 2010a, 2010b). Nevertheless, no genetic differences were found between the two E. timida morphs previously described (fig. 7). In fact, according to data from Giménez–Casalduero (1999), both morphs can occur either in the Mediterranean or Mar Menor populations, which implies

Mg chlorophyll a / dry weight (g)

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Dry weight (g)

3.5 3 2.5 2 1.5 1 0.5 0

0.0035 0.003 0.0025 0.002 0.0015 0.001 0.0005 0

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Mediterranean Lagoon

Dun

Green

Green Population

Mazarrón

Fig. 5. Mean chlorophyll a (A) and mean dry weight (B) concentration values in the tissues of Mediterranean and lagoon populations of Elysia timida (modified from Giménez–Casalduero, 1997b, 1999). Fig. 5. Concentración media de clorofila a (A) y peso seco medio en los tejidos de las poblaciones mediterránea y de laguna de Elysia timida (modificado de Giménez–Casalduero, 1997b, 1999).

Number of eggs/spawn

800

600

400

200

Mar Menor

Mazarrón

Fig. 6. Box plot of number of eggs per spawn in the Mediterranean and lagoon (Mar Menor) E. timida. The box itself represents 50% of all cases, and extends from the 25th to the 75th quartiles. The line inside the box shows the media. Points beyond the whiskers (outliers) are drawn individually. (Source of data from: Marín & Ros, 1993; Giménez–Casalduero, 1997a, 1997b, 1999; unpublished data, 2007). Fig. 6. Diagrama de cajas del número de huevos por puesta de E. timida en el Mediterráneo y en la laguna (Mar Menor). La caja por sí misma representa el 50% de todos los casos, y se extiende desde el cuartil 25 al 75. La línea del interior de las cajas representa la media. Los puntos situados más allá de las líneas (datos atípicos o outliers) se han dibujado individualmente. (Procedencia de los datos: Marín & Ros, 1993; Giménez–Casalduero, 1997a, 1997b, 1999; datos no publicados, 2007).

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Giménez–Casalduero et al.

Axe II

–1 Mar Menor –1

Mediterranean Sea 0

Axe I

Fig. 7. Factorial correspondence analysis on E. timida genotypes: ■ Dun individuals; □ Green individuals. Mediterranean Sea: B. Tabarca Beach; S. South Tabarca; M. Mazarrón Harbour; G. Gachero. Mar Menor coastal lagoon: P. Perdiguera Island; U. Urrutias. The dashed line separates Mar Menor from Mediterranean samples (modified from Gonzalez–Wangüemert et al., 2004).

Mg chlorophyll a / dry weight (g)

Fig. 7. Análisis factorial de correspondencia de los genotipos de E. timida: ■ Individuos pardos; □ Individuos verdes. Mar Mediterráneo: B. Playa de Tabarca; S. Sur de Tabarca; M. Puerto de Mazarrón; G. Gachero. Laguna costera del Mar Menor: P. Isla Perdiguera; U. Los Urrutias. La línea discontinua separa las muestras del Mar Menor de las muestras mediterráneas (modificado de González–Wangüemert et al., 2004).

4.5 4

3.5 3

2.5 2 1.5 1 0.5 0

Days

Fig. 8. Mean chlorophyll a concentration values (± SE) in the tissue of E. timida specimens kept in a laboratory for 28 days at a constant temperature (25º) and fed with A. acetabulum, for treatment light photoperiod 12:12 light and darkness (LS) and continuous darkness (DS) (modified from Giménez– Casalduero, 1997b). Fig. 8. Concentración media de clorofila a (± EE) en los tejidos de especímenes de E. timida mantenidos en el laboratorio durante 28 días a una temperatura constante (25ºC) y alimentados con A. acetabulum con los tratamientos fotoperiodo lumínico 12:12 de luz y oscuridad (LS) y oscuridad continua (DS) (modificado de Giménez–Casalduero, 1997b).

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Editor executiu / Editor ejecutivo / Executive Editor Joan Carles Senar

Secretaria de Redacció / Secretaría de Redacción / Editorial Office

Secretària de Redacció / Secretaria de Redacción / Managing Editor Montserrat Ferrer

Museu de Zoologia Passeig Picasso s/n 08003 Barcelona, Spain Tel. +34–93–3196912 Fax +34–93–3104999 E–mail mzbpubli@intercom.es

Consell Assessor / Consejo asesor / Advisory Board Oleguer Escolà Eulàlia Garcia Anna Omedes Josep Piqué Francesc Uribe

Animal Biodiversity and Conservation 34.1 (2011)

Animal Biodiversity and Conservation Animal Biodiversity and Conservation (abans Miscel·lània Zoològica) és una revista inter disciplinària 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 amb especial énfasis als estudis que d'una manera o altre tinguin relevància en la biología de la conservació. La revista no publica compilacions bibliogràfiques, catàlegs, llistes d'espècies o cites puntuals. Els estudis realit zats 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 http://www.bcn.cat/ABC, de manera que permet una difusió mundial dels seus articles. Tots els manuscrits són revisats per l'editor execu tiu, 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.

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. Si s'opta per la versió impresa, s'han d'enviar quatre còpies del treball juntament amb una còpia en disquet a la Secretaria de Redacció. Cal incloure, juntament amb l'article, una carta on es faci constar que el treball està basat en investigacions originals no publicades anteriorment 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. Quan l'article sigui acceptat, els autors hauran d'enviar a la Redacció una còpia impresa de la versió final acompanyada d'un disquet indicant el progra ma utilitzat (preferiblement en Word). Les proves d'impremta enviades a l'autor per a la correcció, seran retornades al Consell Editor en el termini de 10 dies. Aniran a càrrec dels autors les despeses degudes a modificacions substancials introduïdes per ells en el text original acceptat. ISSN: 1578–665X

El primer autor rebrà 50 separates del treball sense càrrec a més d'una separata electrònica 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. La redacció del text serà impersonal, i s'evitarà sempre la primera persona. Els caràcters cursius s’empraran per als noms científics de gèneres i d’espècies i per als neologis mes 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. 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 designacions de sèries numèriques (I, II, III, etc.) seran acceptats previ acord amb l'editor. Nom de l’autor o els autors. Abstract en anglès que no ultrapassi les 12 línies mecanografiades (860 espais) i que mostri l’essència del manuscrit (introducció, material, mètodes, resultats i discussió). S'evitaran les especulacions i les cites bibliogràfiques. Estarà encapçalat pel títol del treball en cursiva. Key words en anglès (sis com a màxim), que orientin sobre el contingut del treball en ordre d’importància. Resumen en castellà, traducció de l'Abstract. De la traducció se'n farà càrrec la revista per a aquells autors que no siguin castellanoparlants. Palabras clave en castellà. © 2011 Museu de Ciències Naturals

Adreça postal de l’autor o autors. (Títol, Nom, Abstract, Key words, Resumen, Pala bras clave i Adreça postal, 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, mèto des d’estudi i d’anàlisi de les dades i zona d’estudi. Resultados. En aquesta secció es presentaran úni cament les dades obtingudes que no hagin estat publicades prèviament. Discusión. Es discutiran els resultats i es compa raran amb treballs relacionats. Els suggeriments de recerques futures es podran incloure al final d’aquest apartat. Agradecimientos (optatiu). Referencias. Cada treball haurà d’anar acom panyat 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 spe cies 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 conserva tion 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 vari ation 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. Anim. Biodivers. Conserv. 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 & 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 ressen yades 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 au tors. La mida màxima de les figures és de 15,5 cm d'amplada per 24 cm d'alçada. S'evitaran les figures tridimensionals. Tant els mapes com els dibuixos han d'incloure l'escala. Els ombreigs preferibles són blanc, negre o trama. S'evitaran els punteigs ja que no es reprodueixen bé. Peus de figura i capçaleres de taula. Seran clars, concisos i bilingües en la llengua de l’article i en anglès. Els títols dels apartats generals de l’article (Intro ducción, 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.

Animal Biodiversity and Conservation 34.1 (2011)

Animal Biodiversity and Conservation Animal Biodiversity and Conservation (antes Miscel·lània Zoològica) es una revista inter disciplinar, publicada desde 1958 por el Museo Ciencias Naturales de Barcelona. Incluye artículos de investigación empírica y teórica en todas las áreas de la zoología (sistemática, taxonomía, mor fología, biogeografía, ecología, etología, fisiología y genética) procedentes de todas las regiones del mundo, con especial énfasis en los estudios que de una manera u otra tengan relevancia en la biología de la conservación. La revista no publica compila ciones bibliográficas, catálogos, listas de especies sin más o citas puntuales. 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 re gistrada en todas las bases de datos importantes y además está disponible gratuitamente en internet en http://www.bcn.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 garan tizar su calidad. El proceso de revisión es rápido y constructivo, y se realiza vía correo electrónico siem pre 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 propie dad de la revista. Ésta se reserva los derechos de autor, y ninguna parte del trabajo podrá ser reprodu cida sin citar su 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. Si se opta por la versión impresa, deberán remitirse cuatro copias juntamente con una copia en disquete a la Secretaría de Redacción. Debe incluirse, con el artículo, una carta donde conste que el trabajo versa sobre investigaciones originales no publicadas anteriormente y que se somete en ex clusiva 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 permisos necesa rios y que han cumplido la normativa de protección animal vigente. Los autores pueden enviar también sugerencias para asesores. Cuando el trabajo sea aceptado los autores de berán enviar a la Redacción una copia impresa de la versión final junto con un disquete del manuscrito ISSN: 1578–665X

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preparado con un procesador de textos e indican do el programa utilizado (preferiblemente Word). Las pruebas de imprenta enviadas a los autores deberán remitirse corregidas al Consejo Editor en el plazo máximo de 10 días. Los gastos debidos a modificaciones sustanciales en las pruebas de im prenta, introducidas por los autores, irán a cargo de los mismos. El primer autor recibirá 50 separatas del trabajo sin cargo alguno y una copia electrónica en for mato 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 de ben 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 ningu no, 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. La redacción del texto deberá ser impersonal, evitándose siempre la primera persona. 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. 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 designacio nes 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 esen cia del manuscrito (introducción, material, métodos, resultados y discusión). Se evitarán las especula © 2011 Museu de Ciències Naturals

ciones y las citas bibliográficas. Irá encabezado por el título del trabajo en cursiva. Key words en inglés (un máximo de seis) que especifiquen el contenido del trabajo por orden de importancia. Resumen en castellano, traducción del abstract. Su traducción puede ser solicitada a la revista en el caso de autores que no sean castellano hablantes. Palabras clave en castellano. Dirección postal del autor o autores. (Título, Nombre, Abstract, Key words, Resumen, Palabras clave y Dirección postal conformarán la primera página.) Introducción. En ella se dará una idea de los ante cedentes del tema tratado, así como de los objetivos del trabajo. Material y métodos. Incluirá la información referente a las especies estudiadas, aparatos utilizados, me todología de estudio y análisis de los datos y zona de estudio. Resultados. En esta sección se presentarán úni camente los datos obtenidos que no hayan sido publicados previamente. Discusión. Se discutirán los resultados y se compara rán 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 spe cies 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 conserva tion 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 vari ation 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. Anim. Biodivers. Conserv. 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 & Redford (1991)...", "...las prospecciones realizadas (Begon et al., 1999)...". Tablas. Se numerarán 1, 2, 3, etc. y se reseñarán todas en el texto. Las tablas grandes deben ser más estrechas y largas que anchas y cortas ya que deben ajustarse a la caja de la revista. Figuras. Toda clase de ilustraciones (gráficas, figuras o fotografías) se considerarán figuras, se numerarán 1, 2, 3, etc. y se citarán todas en el texto. Pueden incluirse fotografías si son imprescindibles. Si las fotografías son en color, el coste de su publicación irá a cargo de los autores. El tamaño máximo de las figuras es de 15,5 cm de ancho y 24 cm de alto. Deben evitarse las figuras tridimensionales. Tanto los mapas como los dibujos deben incluir la escala. Los sombreados preferibles son blanco, negro o trama. Deben evitarse los punteados ya que no se reproducen bien. Pies de figura y cabeceras de tabla. Serán claros, concisos y bilingües en castellano e inglés. Los títulos de los apartados generales del artículo (Introducción, Material y métodos, Resultados, Dis cusión, Agradecimientos y Referencias) no se nume rarán. 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.

Animal Biodiversity and Conservation 34.1 (2011)

Animal Biodiversity and Conservation

Manuscripts

Animal Biodiversity and Conservation (formerly Miscel·lània Zoològica) is an interdisciplinary journal published by the Natural Science Museum of Barce lona since 1958. It includes empirical and theoretical research from around the world that examines any aspect of Zoology (Systematics, Taxonomy, Morphol ogy, Biogeography, Ecology, Ethology, Physiology and Genetics). It gives special emphasis to studies related to Conservation Biology. The journal does not publish bibliographic compilations, listings, catalogues or collections of species, or isolated descriptions of a single specimen. 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 regis tered in all principal data bases and is freely available online at http://www.bcn.cat/ABC, assuring world–wide access to articles published therein. All manuscripts are screened by the Executive Edi tor, 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 prop erty of the journal, which reserves copyright, and no published material may be reproduced or cited without acknowledging the source of information.

Manuscripts must be presented in DIN A–4 format, 30 lines, 70 keystrokes per page. Maintain double spacing throughout. Number all pages. Manuscripts should be complete with figures and tables. Do not send original figures until the paper has been accepted. The text may be written in English, Spanish or Cata lan, though English is preferred. The journal provides linguistic revision by an author’s editor. Care must be taken to use correct wording and the text should be written concisely and clearly. Scientific names of gen era and species as well as untranslatable neologisms must be in italics. Quotations in whatever language used must be typed in ordinary print between quota tion marks. The name of the author following a taxon should also be written in lower case letters. When referring to a species for the first time in the text, both common and scientific names should be given when possible. Do not capitalize common names of species unless they are proper nouns (e.g. Iberian rock lizard). Place names may appear either in their original form or in the language of the manuscript, but care should be taken to use the same criteria throughout the text. Numbers one to nine should be written in full within the text except when preceding a measure. Higher numbers should be written in numerals except at the beginning of a sentence. Specify dates as follows: 28 VI 99 (for a single day); 28, 30 VI 99 (referring to two days, e.g. 28th and 30th), 28–30 VI 99 (for more than two consecu tive days, e.g. 28th to 30th). Footnotes should not be used.

Information for authors Electronic submission of papers is encouraged (abc@bcn.cat). The preferred format is DOC or RTF. All figures must be readable by Word, embedded at the end of the manuscript and submitted together in a separate attachment in a TIFF, EPS or JPEG file. Tables should be placed at the end of the document. If a printed version is sent, four copies should be forwarded to the Editorial Office, together with a copy on computer disc. A cover letter stating that the article reports original research that has not been published elsewhere and has been submitted exclusively for consideration in Animal Biodiversity and Conservation is also necessary. When animal manipulation has been necessary, the cover letter should also specify that the authors follow current norms on the protec tion of animal species and that they have obtained all relevant permits and authorisations. Authors may suggest referees for their papers. Once an article has been accepted, authors should send a paper copy and an electronic copy of the final version. Please identify software (preferably Word). Proofs sent to the authors for correction should be returned to the Editorial Board within 10 days. Expenses due to any substantial alterations of the proofs will be charged to the authors. The first author will receive 50 reprints free of charge and an electronic version of the article in PDF format. ISSN: 1578–665X

Formatting of articles Title. Must be concise but as informative as possible. Numbering of parts (I, II, III, etc.) should be avoided and will be subject to the Editor’s consent. Name of author or authors. Abstract in English, no longer than 12 typewritten lines (840 spaces), covering the contents of the article (introduction, material, methods, results and discussion). Speculation and literature citation should be avoided. The abstract should begin with the title in italics. Key words in English (no more than six) should express the precise contents of the manuscript in order of relevance. Resumen in Spanish, translation of the Abstract. Summaries of articles by non–Spanish speaking au thors will be translated by the journal on request. Palabras clave in Spanish. Address of the author or authors. (Title, Name, Abstract, Key words, Resumen, Palabras clave and Address should constitute the first page.) Introduction. Should include the historical back ground of the subject as well as the aims of the paper. © 2011 Museu de Ciències Naturals

Material and methods. This section should provide relevant information on the species studied, materi als, methods for collecting and analysing data, and the study area. Results. Report only previously unpublished results from the present study. Discussion. The results and their comparison with re lated studies should be discussed. Suggestions for future research may be given at the end of this section. Acknowledgements (optional). References. All manuscripts must include a bibliog raphy 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 spe cies 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 conserva tion biology. In: Dispersal: 358–372 (T. J. Clober, E. Danchin, A. A. Dhondt & J. D. Nichols, Eds.). Oxford University Press, Oxford. * Ph. D. Thesis: Merilä, J., 1996. Genetic and quantitative trait variation in natural bird populations. Ph. D. 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. Anim. Biodivers. Conserv. References must be set out in alphabetical and chrono

logical order for each author, adding the letters a, b, c,... to papers of the same year. Bibliographic citations in the text must appear in the usual way: "...according to Wemmer (1998)...", "...has been defined by Robinson & Redford (1991)...", "...the prospections that have been carried out (Begon et al., 1999)..." Tables. Must be numbered in Arabic numerals with reference in the text. Large tables should be narrow (across the page) and long (down the page) rather than wide and short, so that they can be fitted into the column width of the journal. Figures. All illustrations (graphs, drawings, photo graphs) should be termed as figures, and numbered consecutively in Arabic numerals (1, 2, 3, etc.) with reference in the text. Glossy print photographs, if essential, may be included. The Journal will publish colour photographs but the author will be charged for the cost. Figures have a maximum size of 15.5 cm wide by 24 cm long. Figures should not be tridimen sional. Any maps or drawings should include a scale. Shadings should be kept to a minimum and preferably with black, white or bold hatching. Stippling should be avoided as it may be lost in reproduction. Legends of tables and figures. Legends of tables and figures should be clear, concise, and written both in English and Spanish. Main headings (Introduction, Material and methods, Results, Discussion, Acknowledgements and Refer ences) should not be numbered. Do not use more than three levels of headings. Manuscripts should not exceed 20 pages including figures and tables. If the article describes new taxa, type material must be deposited in a public institution. Authors are advised to consult recent issues of the journal and follow its conventions.

123–132 M. G. Pennino, J. M. Bellido, D. Conesa & A. López– Quílez Trophic indicators to measure the impact of fishing on an exploited ecosystem 133–140 A. Garcia, S. Cecchetti, M. N. Santos, S. Mattiucci, G. Nascetti & R. Cimmaruta Population structure of Atlantic swordfish (Xiphias gladius L. 1758) (Teleostea, Xiphiidae) using mitochondrial DNA analysis: implications for fisheries management 141–150 E. Cacabelos, J. Moreira, A. Lourido & J. S. Troncoso Ecological features of Terebellida fauna (Annelida, Polychaeta) from Ensenada de San Simón (NW Spain) 151–163 E. Rojo–Nieto, P. D. Álvarez–Díaz, E. Morote, M. Burgos–Martín, T. Montoto–Martínez, J. Sáez– Jiménez & F. Toledano Strandings of cetaceans and sea turtles in the Alboran Sea and Strait of Gibraltar: a long–term glimpse at the north coast (Spain) and the south coast (Morocco)

165–177 M. Samy, J. L. Sánchez Lizaso & A. Forcada Status of marine protected areas in Egypt 179–190 V. Fernández–González & P. Sánchez–Jerez Effects of sea bass and sea bream farming (Western Mediterranean Sea) on peracarid crustacean assemblages 191–203 C. Ojeda–Martínez, J. T. Bayle–Sempere, P. Sánchez– Jerez, F. Salas, B. Stobart, R. Goñi, J. M. Falcón, M. Graziano, I. Guala, R. Higgins, F. Vandeperre, L. Le Direach, P. Martín–Sosa & S. Vaselli Review of the effects of protection in marine protected areas: current knowledge and gaps 205–215 E. Azevedo, M. F. Caeiro, R. Rebelo & M. Barata Biodiversity and characterization of marine mycota from Portuguese waters 217–227 F. Giménez–Casalduero, C. Muniain, M. González– Wangüemert & A. Garrote–Moreno Elysia timida (Risso, 1818) three decades of research

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

Índex / Índice / Contents Animal Biodiversity and Conservation 34.1 (2011) ISSN 1578–665X

1–10 A. Martínez–Ortí A new hygromiid for the Iberian malacofauna: Candidula corbellai n. sp. (Gastropoda, Pulmonata)

71 J. L. Sánchez Lizaso, J. Bayle Sempere & P. Sánchez Jerez Marine biology, biodiversity and conservation: A foreword to the SIEBM 2010 Conference

11–21 S. C. Nikolov, D. A. Demerdzhiev, G. S. Popgeorgiev & D. G. Plachiyski Bird community patterns in sub–Mediterranean pastures: the effects of shrub cover and grazing intensity

23–29 K. A. Otter, B. W. Murray, C. I. Holschuh & K. T. Fort Rare insights into intraspecific brood parasitism and apparent quasi–parasitism in black–capped chickadees 31–34 M. Lozano, J. Baro, T. García, A. Frías, J. Rey & J. C. Báez Loggerhead sea turtle bycatch data in artisanal fisheries within a marine protected area: fishermen surveys versus scientific observations 35–45 D. Denis & U. Olavarrieta ¿Existe isomorfía en los huevos de las especies de la familia Ardeidae (Aves, Ciconiiformes)? 47–66 R. I. Ruiz–C., C. Román–Valencia, B. E. Herrera–M., O. E. Peláez & A. Ermakova–A. Variación morfológica de las especies de Astyanax, subgénero Zygogaster (Teleostei, Characidae)

83–99 G. A. Rivera–Ingraham, F. Espinosa & J. C. García– Gómez Conservation status and updated census of Patella ferruginea (Gastropoda, Patellidae) in Ceuta: distribution patterns and new evidence of the effects of environmental parameters on population structure 101–111 L. M. Ferrero–Vicente, Á. Loya–Fernández, C. Marco– Méndez, E. Martínez–García & J. L. Sánchez–Lizaso Soft–bottom sipunculans from San Pedro del Pinatar (Western Mediterranean): influence of anthropogenic impacts and sediment characteristics on their distribution 113–122 M. García–Rodríguez, A. Fernández &. A. Esteban Biomass response to environmental factors in two congeneric Mullus, M. barbatus and M. surmuletus, off Catalano–Mediterranean coast of Spain: a preliminary approach