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La cause de la mort des Ongulés sauvages consommés par les vautours (bouquetins et surtout chamois) dans la zone d’étude n’a donc pu être déterminée jusqu’à maintenant, sauf un cerf à Chamaloc (Diois) : par balle44. Car on doit s’abstenir de perturber les curées de vautours, que ce soit sur animaux sauvages ou domestiques, sauf nécessité absolue. Pour le bétail, s’il y a suspicion d’attaque par Loup, les nécessités de constat pour éventuelle indemnisation font examiner la charogne. Cette différence explique que, dans la zone d’étude, on dispose pour le bétail de preuves de curée de vautours sur proies de loups, alors qu’elles sont encore attendues pour les Ongulés sauvages.

Photos A. BULTINGAIRE

Photos 27 . – Curée sur brebis en estive tuées par loups : vautours fauves Gyps fulvus et vautours moines Aegypius monachus à proximité immédiat des sites de lâcher préconisés pour le Gypaète Gypaetus barbatus, dans le sud de la Réserve Naturelle des Hauts Plateaux du Vercors, juin 2005 (vers la Croix du Lautaret).

Le Loup vivant essentiellement d’Ongulés sauvages c’est néanmoins une certitude rationnelle que son retour augmente les disponibilités alimentaires de cette origine pour les charognards. 44

Des chasseurs ont aussi observé des vautours fauves arrivant sur chamois ou sangliers abattus par eux.

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6. Charniers à vautours A Rémuzat (nord des Baronnies) et à Chamaloc (nord du Diois en bordure du Vercors) des charniers sont approvisionnés toute l’année pour Vautour fauve, Vautour moine et Percnoptère. Si besoin était, ils offriraient pour le Gypaète une « assurance trophique » à la mauvaise saison. On connaît le rôle joué par la fourniture de telles ressources complémentaires dans la restauration de la population de Gypaète des Pyrénées. C. – CLIMAT « L’altitude importe peu, le gypaète étant présent du niveau de la mer (en Corse, en Crète et même dans la dépression de la Mer Morte) jusqu’au massif de l’Everest. Il peut supporter un climat continental froid (en Asie centrale)…la pluviométrie importe peu, bien qu’elle puisse être un facteur défavorable pendant la reproduction : les montagnes soumises à la mousson dans l’Himalaya, les Pyrénées occidentales ou le Caucase avec une forte pluviométrie, les hauts plateaux éthiopiens au taux d’humidité important peuvent être de bons habitats, comme les montagnes subdésertiques de l’Atlas, du Yémen ou de l’Altaï » (Terrasse). L’importance des précipitations et de la nébulosité expliquerait au moins en partie la relativement faible productivité des couples des Pyrénées occidentales, sans suffire néanmoins à empêcher l’espèce de prospérer dès lors que nourriture et protection lui sont assurées. Dans les Alpes orientale, un climat particulièrement plus rude en hiver provoque un mouvements saisonniers des individus non nicheurs vers le versant sud de la chaîne, singularité par rapport aux Alpes occidentales. Bioclimatiquement, la zone d’étude appartient surtout aux Alpes occidentales méridionales. Seuls le Dévoluy, le haut Vercors et le haut Diois, soit bien moins de la moitié de la zone d’étude, subissent un enneigement relativement important et durable (qui tend à diminuer au fil des décennies. Caractéristique particulièrement favorable à la prospection alimentaire par toutes les espèces de vautours, Gypaète inclus, le nombre de jours de brume est, dans la zone d’étude, beaucoup plus réduit que plus au nord dans les Alpes. Ceci à la seule exception du Vercors qui, sauf sa frange sud et son coin sud-ouest appartient déjà aux Préalpes du nord. Etages de végétation et essences dominantes. Étages représentés dans la zone d’étude, distribution par massifs : - alpin : en Dévoluy et sur les crêtes les plus hautes de l’est du Vercors ; - subalpin, à Pin-à-crochet Pinus uncinata. Non représenté dans les Baronnies et, sauf ses crêtes orientales, le Diois ; - montagnard à boisements mésophiles de Hêtre Fagus silvatica. Le Sapin Abies pectinata, répandu dans le Vercors, est très localisé ailleurs ; - collinéen de la série supra-méditerranéenne du Chêne blanc Quercus pubescens45, avec nombreuses espèces méditerranéennes46 ; - méditerranéen : dans les Baronnies, où passe la limite septentrionale de l’Olivier. L’ouest de ce massif (Nyonsais) est la région la plus ensoleillée de France. Seul, le Vercors, à l’exception de son sud-ouest et de sa bordure méridionale, a un climat 45

Des pins Pinus dominent souvent localement dans les étages collinéens et montagnards : P. sylvestris (essence pionnière, durablement présente sur les substrats les plus défavorables ou/et dans les zones les plus sèches) P. nigra austriaca (largement introduit par les forestiers). Inversement, la série du Rouvre Quercus petraea (= sessiliflora) est limité à la bordure collinéenne nord et nord-ouest du Vercors. 46 Juniperus phoenicea, Spartium junceum, Genista scorpius, Dorycnium suffruticosum, Catananche caerulea, Aphyllanthes monspessuliensis, etc. et, plus, localement Pinus halepensis, Quercus ilex, Juniperus oxycedrus, J. thurifera, etc. Projet de réintroduction du Gypaète Barbu dans les Préalpes occidentales: pertinence stratégique, faisabilité, biotope et site de lâcher au PNRV. Jean-Pierre Choisy 2010

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montagnard arrosé analogue à celui de la Chartreuse et autres Préalpes du nord. Bien que n’étant pas déterminant pour la présence du Gypaète, LE CLIMAT DE LA ZONE D’ETUDE EST DONC GLOBALEMENT TRES FAVORABLE : au niveau tant de la survie hivernale que du taux de réussite de la reproduction. Ceci ne peut qu’être un avantage pour la dynamique de population locale.

CONCLUSIONS à ce niveau de l’analyse : Les habitats favorables abondent, notamment falaises et éboulis calcaires. Les importantes populations d’Ongulés sauvages, présentes toute l’année, resteront essentielles pour le Gypaète, notamment à la mauvaise saison. Le retour du Lynx et surtout du Loup augmente certainement la disponibilité en charognes de Sanglier, Chamois, Mouflon, Cerf et surtout de Chevreuil. Chamois et surtout Bouquetin, partout particulièrement favorables pour le Gypaète du fait de leurs habitats rupestres, le seront encore plus dans la zone d’étude du fait de son fort taux de boisement. Cette prépondérance minimise les conséquences pour le présent travail d’une estimation moins précise des effectifs de Chevreuil et de Sanglier. La mise en alpage du bétail local, diversifié et l’arrivée des transhumant, à forte prépondérance d’ovins, offre à tous les charognards un maximum estival des ressources alimentaires dans la nature. Dans le cycle annuel de reproduction du Gypaète, il se situera de la fin de l’élevage au nid à la dispersion des juvéniles. Ce maximum trophique sera donc extrêmement favorable au succès de l’élevage comme à celui de l’émancipation des jeunes oiseaux inexpérimentés ainsi qu’à leur taux de survie au cours de cette phase vulnérable. Ceci qu’ils s’envolent d’une aire ou d’un taquet. Les charniers fréquentés par Vautour fauve, Vautour moine et Percnoptère offrent une sécurité supplémentaire en cas de disette temporaire. Ils sont d’ailleurs exploités occasionnellement par Milan noir Milvus migrans, Milan royal M. milvus et, parfois, Aigle royal Aquila chrysaetos. On y a déjà vu deux Pygargues Haliaetus pygargus et un…Aigle des steppes Aquila rapax (Tessier, assoc.Vautours en Baronnies) ! Toutefois, il convient de souligner que, malgré une pression d’observation très élevée, aucune des trente données de Gypaète dans la zone d’étude, une seule (mai 2008) provient d’un site de charnier, sans qu’on l’ai vu s’y poser. Ceci tend à confirmer l’importance dans le milieu naturel de la zone d’étude des disponibilités alimentaires pour le Gypaète.

LA ZONE D’ETUDE OFFRE A LA REINTRODUCTION DU GYPAETE : - VASTE HABITAT OPTIMAL ; - SECURITE ALIMENTAIRE : DEJA ASSUREE ET CROISSANTE.

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II. IMPACT DES ACTIVITES HUMAINES Une densité humaine remarquablement faible47 pour l’Europe occidentale (même lors de l’augmentation relative lors de la saison touristique estivale concentrée sur les mois de juillet et d’août) est indiscutablement un contexte favorable au retour de la grande faune. Ce facteur favorable ne saurait suffire : dans le passé comme de nos jours, les exemples abondent de grande faune exterminée en dépit d’une densité humaine très faible, qu’il s’agisse d’excès de prélèvements cynégétiques ou bien de destruction de “nuisibles” ou prétendus comme tels. A. – DESTRUCTIONS DIRECTES Dans la zone d’étude, le maximum de densité humaine se situe au milieu du XIX° siècle. Or, c’est au cours des cent années suivantes que l’extermination de la grande faune (dont l’Ours et le Loup) y connaît son paroxysme et ce, en dépit d’un effondrement de la densité humaine. Jusqu’aux années 1960 incluses, observer un seul Ongulé restait une expérience rare et mémorable, à l’exception des stations les plus difficiles d’accès ou encore de très rares territoires hors chasse, parfois au bénéfice d’une seule espèce, généralement le Chamois. Quelle que soit la densité humaine, le retour général de la grande faune dont a bénéficié la zone d’étude48 aurait été impossible sans une prise de conscience en faveur de la protection de la nature, de la conservation et de la restauration de la biodiversité et, pour les espèces chassées, de la nécessité de gérer les prélèvements. Ceci avec une très réelle amélioration de la perception de la faune en général, des grands Rapaces en particulier et des comportements son égard, du respect effectif de la protection légale de ces oiseaux par la grande majorité, chasseurs ou non. On le doit à l’action persévérante, pendant des décennies, de minorités, chasseurs ou non-chasseurs, initialement très faibles mais très motivées, donc très actives. Cette action qui commence à porter ses premiers fruits dans les années 1970, continue à se développer jusqu’à nos jours. Des photos telles que celles de bouquetins et chamois, pages précédentes ou page suivante, traduisent une réalité actuelle qui ne pouvait être qu’un rêve pour les générations précédentes. L’état des populations d’Aigle royal et plus encore de Vautour fauve, le début de renouveau du Percnoptère et du Vautour moine, montre que le poison n’est plus un problème dans la zone d’étude. Bien entendu, on ne peut totalement exclure le moindre acte délictueux ponctuel. Mais la situation actuelle de l’Aigle royal, du Vautour moine, du Percnoptère et du Vautour fauve montre que L’IMPACT DEMOGRAPHIQUE ACTUEL DES DESTRUCTIONS DIRECTES PAR L’HOMME EST NUL OU NEGLIGEABLE. Il n’y a aucune raison de penser qu’il en soit autrement pour le Gypaète. B. - PERTURBATIONS L’évolution positive de la perception de la Faune, la grande notamment, et des comportements de l’Homme à son égard induit en retour une tolérance bien plus grande de celle-ci à la présence et aux activités humaines, sous réserve qu’elles ne détruisent pas, ou n’altèrent pas excessivement, leurs biotopes. 47 48

A l’exception de l’extrême nord de la zone d’éude proche de l’agglomération grenobloise. Cf. infra § sur ce thème in quatrième partie.

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Photo R. MATHIEU, Observatoire de la faune drômoise CORA-FRAPNA

Photo 28. – Chamois R. rupicapra installé à proximité de l’Homme. Impensable jusqu’aux années 1970 : un des nombreux indices du changement radical de la perception de la grande faune sauvage par Homo sapiens et de son comportement à l’égard de celle-ci dans la zone d’étude. Ici, un noyau de population de vingt-cinq chamois à 150 m. d’altitude à St Marcel-lès-Sauzet sur la bordure de la vallée du Rhône avec Chêne vert Quercus ilex et, en bas de versant Peuplier noir Populus nigra, à quelques centaines de mètres d’une des autoroutes les plus fréquentées d’Europe hors de la zone d’étude stricto sensu mais en bordure.

Mais ceci n’exclut nullement qu’on puisse nuire involontairement à la grande Faune en général, aux Rapaces rupestres et notamment au Gypaète en particulier, notamment en perturbant sa reproduction et en la faisant échouer, d’où impact négatif sur la dynamique des populations, voire sur leur pérennité. 1. Nature et mode d’action des perturbations Fondé sur une approche statistique de données de terrain précises dans les Pyrénées, le travail récent d’Arroyo & Razin (2006) “Effect of human activities on bearded vulture behaviour and breeding success in the French Pyrenees”, est LA référence actuelle pour identifier les causes de perturbation pouvant faire échouer la nidification du Gypaète ainsi que comprendre leur mode d’action. Sont particulièrement décisifs pour l’impact des activités humaines sur la nidification du

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Gypaète le NIVEAU DE BRUIT des activités, leurs DATES - dans les périodes les plus sensibles du cycle annuel de reproduction ou hors d’elles - et la DISTANCE au site de nidification à laquelle elles sont exercées. Sans prétendre rendre compte de la richesse de cette publication, qu’on trouvera in extenso dans les ANNEXES I quelques citations doivent en souligner des points majeurs : “Very noisy activities (such as infrastructure works, motorbikes, forestry or military activities and helicopters) seemed to be those most strongly affecting bearded vulture behaviour and breeding success. Noise is particularly transmitted in alpine habitats, as relief provokes echoes and increases resonance. These activities thus provoked a reaction in bearded vultures even if far away from the nest (2 km)….These activities should, therefore, be largely avoided around the nests in order to maximise bearded vulture productivity. where disturbance was most important in terms of leading to failure. In the Spanish Pyrenees breeding success was overall related to the failure during the hatching and incubation periods (Margalida et al., 2003).” “Breeding success was significantly negatively associated with the frequency of very noisy activities in a territory during a breeding event, as the probability that nests were left unattended was higher in those territories… Hunting also strongly affected bearded vulture behaviour during the pre-laying period, as it was associated with the absence of the pair from the territory in a significantly high proportion of cases…which may be associated with the fact that hunting as exercised in the French Pyrenees (with groups of people) is also potentially noisy as well as having a visual impact during the pre-laying period (October– December). « Very noisy activities and hunting most frequently provoked nest unattendance even when occurring far (>1.5 km) from the nest. People on foot or cars/planes only affected bearded vulture behaviour if close ( 2,23 qui est le seuil pour p = 0,05, 10 ddl.

TABLEAU VI . – Temps de vol cumulé des juvéniles de Gypaète Gypaetus barbatus : moyenne supérieure de 74 % depuis le développement de l’estivage du Vautour fauve Gyps fulvus au cours du premier mois suivant l’envol du taquet au Mercantour. D’après L. Zimmerman in V. Coirié. Seul Monte-Carlo n’a volé qu’approximativement autant d’heures (8% en moins) que la moyenne d’avant 2002, les cinq autres l’ont fait beaucoup plus : Rocca et Fontvieille plus de deux fois plus, Guillaumes près de trois fois.

À notre connaissance, c’est la première fois que le fait est mis en évidence. Il est probable que les parents des gypaètes juvéniles nés dans la nature jouent le même rôle pour ceux-ci. L’aptitude des gypaètes envolés du taquet à utiliser de même les vautours fauves mérite d’être soulignée. Projet de réintroduction du Gypaète Barbu dans les Préalpes occidentales: pertinence stratégique, faisabilité, biotope et site de lâcher au PNRV. Jean-Pierre Choisy 2010

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B. – APRES ACQUISITION DE LA MAITRISE DU VOL Citations de Coirié 2007 Lors du suivi des vautours fauves l’été 2007, la mise au taquet de « deux gypaètes le 22 mai 2007 (Rocca et Fontvieille), au sein même de la zone fréquentée par les vautours fauves, a permis d’observer quelques relations entre les deux espèces.Les vautours fauves ont fréquenté de façon importante le site de réintroduction sur un couloir de passage entre le massif du Mounier et le plateau de Longon (et) régulièrement utilisé comme reposoir diurne mais aussi comme lieu propice aux ascendances thermiques. » 1. Territorialisme à distance du taquet « De mi-juillet à mi-août, les gypaètes ont commencé l’apprentissage du vol… A plusieurs reprises, nous avons pu observer ces deux jeunes oiseaux à la poursuite de vautours fauves de passage. Ce comportement peut être assimilé à de la territorialité : (en effet) ce phénomène n’a été constaté que sur le site même de réintroduction. (Au contraire) en d’autres lieux, les gypaètes sont (restés) pacifiques vis-à-vis des vautours. » 2. Grégarisme à distance du taquet « nous avons pu, à plusieurs reprises, observer les gypaètes en compagnie des vautours …aussi bien sur le reposoir nocturne de la Barre Sud du Mounier que sur…des curées » 3. Découverte de carcasses « …le Vautour fauve, prompt à trouver les carcasses disponibles et à les consommer, a (ainsi) permis aux jeunes gypaètes l’accès à une ressource trophique très abondante. Des cassages d’os ont été constatés en ces lieux. » 4. Moindre dispersion, conséquences probables « ...la présence des vautours et l’abondance en nourriture qui en découle entraîne vraisemblablement une fixation plus longue des gypaètes sur le massif du Mounier avant leur départ automnal en phase d’erratisme, Plus cette imprégnation au site sera longue, plus les chances de retour pourraient être importantes. »

CONCLUSIONS à ce niveau de l’analyse : LA PRESENCE PERENNE DE VAUTOURS FAUVES DANS LA ZONE D’ETUDE : - d’abord FACILITERA L’ACQUISITION .DE LAMAITRISE DU VOL PAR LES GYPAETES QUITTANT LE TAQUET ; .

- ensuite FAVORISERA : - LA DECOUVERTE DE NOURRITURE ET DE REPOSOIRS ; - UN SEJOUR PROLONGE DES GYPAETES MAITRISANT LE VOL.

CES FACTEURS ECOLOGIQUES ET ETHOLOGIQUES MAXIMALISERONT LA PROBABILITE DE SURVIE DES JUVENILES, PUIS D’ATTACHEMENT AU BIOTOPE. Il est probable que la réintroduction préalable du Vautour fauve ait déjà contribué de même au succès ultérieur de celle du Vautour moine dans les Causses et contribue à celui des opérations en cours dans les Préalpes, bien qu’on ne dispose pas d’observations aussi circonstanciées que pour le Gypaète au Mercantour. Le versant occidental du Glandasse offrent aussi des sites potentiels avec falaises, éboulis, bouquetins, chamois et, sur l’alpage entre lui-même et le cirque d’Archiane, des troupeaux de brebis en estive. Projet de réintroduction du Gypaète Barbu dans les Préalpes occidentales: pertinence stratégique, faisabilité, biotope et site de lâcher au PNRV. Jean-Pierre Choisy 2010

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Photo P. LARDANCHET, CORA

Photo 47. – Versant occidental du Glandasse.

Photo P. BEAUDOIN, FRAPNA

Photo 48. – Le dessus du Glandasse vu de la pointe sud : Entre vallée de la Drôme à l’ouest (à gauche sur la photo) et cirque d’Archiane à l’est (à droite

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. Photo R. MATHIEU, Observatoire de la faune drômoise CORA-FRAPNA

Photo 49 – Nord du versant ouest du Glandasse à sa jonction de la bordure méridionale du Vercors à l’ouest de la photo. Fréquenté du fait d’un sentier d’accès aux Hauts Plateaux, mais surveillance continuelle habituelle lors de lâchers au taquet.

Photo R. MATHIEU, Observatoire de la faune drômoise CORA-FRAPNA

Photo 50. – L’est de la bordure méridionale du Vercors un peu à l’ouest de sa jonction avec le Glandasse. L’accès d’un éventuel taquet se ferait par le plateau.

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Le sud Hauts-Plateaux du Vercors (Glandasse inclus) bordant les falaises biotopes de la principale population de bouquetins offrent des reliefs rocheux escarpés, modestes mais suffisant pour un taquet, environnés d’alpages, prés-bois, bosquets clairs et épars, essentiellement de Pin-à-crochets Pinus uncinata, sans boisement étendus ni denses.

Photos R. MATHIEU, Observatoire de la faune drômoise CORA-FRAPNA….

Photos 51. - Le sud des Hauts des Plateaux du Vercors :

photo du haut paysage typique, photo du bas : détail de petits sites escarpés.

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Photos R. MATHIEU, Observatoire de la faune drômoise CORA-FRAPNA Photos 52. - Le sud des Hauts des Plateaux du Vercors : sites escarpés moins modestes que photos 46. Sur celui du bas des bouquetins commencent à être vus.

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COMBE MALE Ce site pourrait être considéré comme une variante de celui de Chamaloc (cf. supra début du chapitre) : les avantages logistiques sont presque équivalents, sans le risque d’une ligne électrique, avec des pentes à couvert ligneux notablement plus faible. QUINT Si une distance de 14 ou 15 km à l’ouest du plus important noyau de population de bouquetins était considérée comme négligeable, des sites de la bordure méridionale du Vercors, tel que celui-ci-dessous, seraient excellents.

Photo R. MATHIEU, Observatoire de la faune drômoise CORA-FRAPNA

Photo 53. – Bordure méridionale du Vercors : falaises du Quint.

INCONVENIENTS : - encore sans population de bouquetins ; - ligne à moyenne tension non balisée ; - hors de la Réserve Naturelle des Hauts Plateaux du Vercors. AVANTAGES : - excellente position géographique : à l’ouest de la zone d’étude ; - chamois, mouflon, cerf ; - brebis et autres transhumants ; -

- domaniaux étendus et propriété départementale en réserve de chasse

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IV. LE PARC NATUREL REGIONAL DU VERCORS : CADRE ET AGENT DE REALISATION La réalisation de l’opération impliquera une concertation du Parc Naturel Régional avec les organismes directement concernés, tant publics (Ministère de l’Environnement, DIREN, etc.) qu’associatifs (FCBV, LPO Misson Rapaces, LPO Drôme, LPO Isère, CORA Faune sauvage, etc.). Seront des partenaires techniques privilégiés : -

les associations, Vautours-en-Baronnies et International Bearded-Vulture Monitoring et Agir pour la Sauvegarde des Territoires et des Espèces Remarquables ou Sensibles (ASTERS). Cette dernière ayant réalisé l’opération de réintroduction du Gypaète achevée en Haute-Savoie, participe à la reproduction de l’espèce en captivité et coordonne sa réintroduction dans les Alpes françaises ;

-

le Parc National des Ecrins, qui centralise les données de Gypaète en Dauphiné. Son territoire, au centre d’un zone d’estivage massif de vautours fauves non nicheurs et quelques vautours moines, provenant surtout de la population de la zone d’étude, le sera nécessairement également par des gypaètes qui y seront lâchés.

Le protocole commun de réintroduction et de suivi entre ASTERS et le Parc National du Mercantour fournit un modèle (Geng, Heuret & Rouillon) qui pourrait être adapté au contexte local et aux connaissances actuelles. La compétence de l’association étant reconnue en la matière, ce qui suit analyse les aptitudes à réaliser l’opération de l’autre partenaire : le Parc Naturel Régional du Vercors.

A. - UN CONTEXTE GENERAL DE RETOUR LA GRANDE FAUNE En France comme dans l’ensemble de la chaîne des Alpes, il n’y a guère de territoires qui puissent se comparer à celui du Parc Naturel Régional du Vercors par le retour de sa grande faune54, avec ou sans réintroduction : 1. Retours spontanés Sanglier Sus scrofa : dès le premier quart du XX° siècle ; Héron cendré Ardea cinerea, tous cours d’eau, localement avec nidification ; Grand cormoran Phalacrocorax carbo : en hivernage sur les grands cours d’eau encadrant Parc Naturel Régional du Vercors au nord et à l’ouest avec remontée à l’intérieur des Préalpes en petit nombre le long des basses vallées ; Loup Canis lupus : la source des pionniers a été le renouveau de l’espèce en Italie péninsulaire d’où elle a gagné les Alpes françaises par le Mercantour, dont la zone d’étude. Les difficultés de la gestion de sa cohabitation avec l’élevage font trop souvent oublier que ce qui permet au Loup de vivre toute l’année c’est d’abord le renouveau spectaculaire des Ongulés sauvages, base de sa subsistance : en Europe en général, dans le Parc du Vercors et massifs contigus en particulier.

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On peut toujours discuter la limite inférieure de la faune dite « grande ». On y a inclus ici la totalité des Ongulés, ainsi que les deux Rongeurs de taille très supérieure à celle des autres espèces de l’ordre en Europe : Marmotte et Castor. Du point de vue démographique ce sont toutes des espèces qui, autrefois, n’avaient pu compenser prélèvements ou/et destructions, en dépit de la persistance d’habitats favorables. Projet de réintroduction du Gypaète Barbu dans les Préalpes occidentales: pertinence stratégique, faisabilité, biotope et site de lâcher au PNRV. Jean-Pierre Choisy 2010

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À l’échelle locale, certaines espèces, naguère réduites à une minorité de stations du territoire du Parc Naturel Régional du Vercors, ont fait un très large retour spontané dans les biotopes adéquats du massif : Chamois R. rupicapra55, Rapaces en général, dont les plus remarquables sont Aigle royal Aquila chrysaetos, Jean-le-Blanc Circaetus gallicus, Autour Accipiter gentilis, Pèlerin Falco peregrinus, Grand-duc Bubo bubo. 2.

Réintroductions56

Cerf Cervus elaphus, Chevreuil C. capreolus, d’abord, puis Marmotte Marmotta marmotta, en 1989-90 Bouquetin Capra ibex et, de 1996 à 2008, Vautour fauve Gyps fulvus. C’est dans le Vercors qu’ont vécu, au moins jusqu’à la fin des années 1930, les derniers individus d’Ours Ursus arctos des Alpes françaises. Une étude de faisabilité (Erome & Michelot 1990) couvrant leur ensemble a conclu que le Vercors et le Haut-Diois étaient, de nos jours, leur zone la plus favorable au retour de l’espèce. Le Parc Naturel Régional du Vercors a organisé un colloque avec des spécialistes européens de l’espèce qui ont visité les biotopes et validé l’étude et ses conclusions. Non réalisé le projet n’est nullement officiellement abandonné. D’autres espèces pourraient encore être réintroduites. 3. Processus mixtes Lynx Lynx lynx : retour spontané à partir de la Suisse, où l’espèce avait été antérieurement réintroduite. Retour permis par le renouveau des Ongulés, surtout celui du Chevreuil ; Castor Castor fiber : retour spontané à partir du bas bassin du Rhône, où l’espèce avait survécu. Concerne le sud du PNRV pour le bassin de la Drôme. Réintroduction dans la région grenobloise, d’où il a gagné la Gresse, jusque dans l’est du PNR du Vercors ; Vautour moine Aegypius monachus. L’estivage d’individus de la population réintroduite avec succès dans les Causses, régulier depuis 2002 aux confins du Diois et du Vercors, nulle part ailleurs dans les Alpes avant les premiers lâchers dans les Baronnies et, depuis ces derniers, la présence de l’espèce en toutes saisons, traduisent la qualité des biotopes, la diversité et l’abondance des Ongulés sauvages et domestiques, donc de leurs cadavres. Un couple semble se cantonner à Chamaloc ; Vautour percnoptère Neophron percnopterus. Le retour de ce picoreur de bribes, nettoyeur de carcasses, est le résultat d’une synergie : - démographique = source d’oiseaux : la population jadis relictuelle de la région méditerranéenne française (Gallardo 2006) désormais en renouveau ; - écologique = nourriture : charniers à Vautour fauve et renouveau de la grande faune ; - éthologique = stimulus : l’espèce a évolué en finisseur de restes et suiveur des grands vautours. 55

En dépit de certains sous-peuplements locaux du fait d’attributions de prélèvements excessives.

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L’introduction d’espèces exotiques peut user de moyens analogues : les lâchers. Mais, processus majeur d’altération de la biodiversité, elle est diamétralement opposée aux politiques visant à la conserver ou à la restaurer. La zone d’étude n’est pas totalement indemne des séquelles de cette « maladie infantile » de la protection de la nature d’antan qu’exprimait l’expression désuète : « Société d’Acclimatation ». On doit y déplorer l’introduction du Mouflon Ovis gmelini. La gestion cynégétique actuelle des populations subsistantes de l’est et du sud-ouest Vercors ainsi que des confins méridionaux de la zone d’étude et du Ventoux ne vise plus à développer ces populations d’exotiques, bien au contraire. Le précédent du Mercantour permet d’espérer que, particulièrement dans les zones à fort enneigement, le retour du Loup contribuera à supprimer ou réduire cette aberration biogéographique. Les tentatives d’introduire le Sika Cervus (Sika) nippon et le Faisan vénéré Syrmaticus reevesi ont, elles, échoué… heureusement !

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La réintroduction du Vautour fauve, comme dans les trois autres massifs qui en ont bénéficié en France, a ici joué dans le Parc Naturel Régional du Vercors ce rôle éco-éthologique. Une quarantaine d’années après la disparition de l’espèce au nord de la rivière Drôme, son retour y a recommencé en 2001, d’abord, centré sur le site de réintroduction de Chamaloc. Le charnier de cette commune du Diois est toujours fréquenté par le couple qui, depuis 2008, niche avec succès dans le sud-ouest du Vercors (cf. supra) comme par des non nicheurs. B. - A LA CONVERGENCE DE GRANDS AXES POLITIQUES DU .PARC NATUREL REGIONAL DU VERCORS 1. Une politique ferme, ancienne, persévérante de conservation et restauration de la biodiversité LE RETOUR DE TREIZE GRANDES ESPECES, LE RENOUVEAU SPECTACULAIRE DE PLUSIEURS AUTRES dont deux grands Carnivores, quatre Ongulés, trois vautours C’EST D’ABORD ET FONDAMENTALEMENT UNE TRES REMARQUABLE RECONSTITUTION DE LA BIODIVERSITE FAUNISTIQUE. Ceci que le Parc Naturel Régional du Vercors ait pris l’initiative de ces retours,

qu’il en ait simplement bénéficié ou que, comme dans le cas du Loup, sans l’avoir voulu, il ait assumé la gestion de la cohabitation avec l’espèce. Or, par leur originalité taxonomique, leurs habitats ou/et leur place dans le fonctionnement des écosystèmes, ces grands animaux ont un poids dans la biodiversité qui dépasse de beaucoup le seul nombre de leurs espèces, relativement modeste. LA POLITIQUE DE CONSERVATION ET RESTAURATION DE LA BIODIVERSITE DU PARC NATUREL REGIONAL DU VERCORS, DE LONGUE DATE ET PERSEVERANTE, est pour beaucoup dans ces retours et renouveaux, notamment en ce qui concerne les espèces réintroduites par lui directement (Bouquetin, Vautour fauve et, en coopération, Marmotte) ou indirectement (Percnoptère, Vautour moine). Dès son entrée en fonction, et à diverse reprises ensuite, le Président Pillet, au sommet de la direction politique du Parc Naturel Régional du Vercors, avait affirmé, réaffirmé ensuite, que la conservation et la restauration de la biodiversité en général, les réintroductions en particulier, étaient au nombre des missions majeures de cet organisme. La présidente Pic, qui lui succédé depuis la première rédaction de ce travail, n’a nullement remis en cause cette orientation. Celle-ci irait de soi dans un Parc National. Dans un Parc Naturel Régional, syndicat mixte de communes et collectivités territoriales, cette prise de position publique ferme montre une volonté politique du Bureau qui mérite d’être soulignée. 2. Retombées pour d’autres politiques a. - POLITIQUE AGRICOLE La réintroduction des vautours en général est une contribution à la gestion des charognes en zone d’élevage extensif élégante, efficace et économique (cf. Chassagne et in ANNEXES II Choisy 2004 b). Le retour du Gypaète, consommateur des squelettes, complètera la guilde des grands charognards. b. - POLITIQUE DE DEVELOPPEMENT DU TOURISME RURAL Le développement du tourisme rural est l’un des grands axes de l’action du Parc Naturel Régional du Vercors. LES PROFESSIONNELS DU TOURISME SONT DE PLUS EN PLUS CONSCIENTS DE L’ATOUT MAJEUR QUE CONSTITUE UN RETOUR AUSSI SPECTACULAIRE DE LA GRANDE FAUNE, à trois niveaux, d’intensité et d’extension inverses : - SUPPORT MEME DU TOURISME NATURALISTE « FAUNE », dont la clientèle, minoritaire mais en expansion, est fortement motivée par l’extrême diversité. La région Rhône-Alpes est actuellement la seule d’Europe où, depuis 2009, pondent des couples des quatre Projet de réintroduction du Gypaète Barbu dans les Préalpes occidentales: pertinence stratégique, faisabilité, biotope et site de lâcher au PNRV. Jean-Pierre Choisy 2010

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vautours du continent, bien que le Vautour moine n’ait pas encore élevé de jeunes, les trois autres que le Gypaète le faisant dans la zone d’étude. Le retour du Gypaète donnera aux Préalpes du Dauphiné méridional un caractère ornithologique exceptionnel par la cohabitation des quatre Vautours d’Europe. - SUPPLEMENT TRES APPRECIE DU TOURISME DE PLEIN AIR : une probabilité élevée de rencontrer la grande faune est extrêmement appréciée par un nombre considérable de randonneurs. Les plus « rentables » sont actuellement, sans conteste, Bouquetin et Vautour fauve : grandes espèces observables sans contraintes d’horaire, à distance modérée, voire courte, et souvent en nombre. Ces deux espèces sont désormais associées aux Hauts Plateaux du Vercors dans l’esprit des randonneurs, la seconde également plus largement : du Diois au sud du Vercors. - MARQUAGE FORT DU TERRITOIRE DANS LA REPRESENTATION QUE S’EN FAIT LE GRAND PUBLIC SUSCEPTIBLE D’Y VENIR SEJOURNER : « si la grande faune prospère, c’est donc une région à la nature préservée, donc agréable pour y passer ses vacances ». Écologiquement un peu simpliste mais il faut bien constater que, au niveau des représentations, ça fonctionne ! c . – CULTURE : UN FAIT SOCIAL CONCERNANT LE NIVEAU POLITIQUE Le retour de la grande faune en général, du Gypaète en particulier, ajoute à l’émotion57 du plus grand nombre face aux grands animaux une dimension culturelle certaine ; objet de recherches scientifiques sur divers aspects de la biologie des espèces, support à la créativité artistique du photographe, du peintre, du sculpteur, de diffusion dans le public des connaissances acquises et des œuvres créées. d. – PERCEPTION

DU TERRITOIRE PAR SES HABITANTS COMPTE AU NIVEAU POLITIQUE

:

UN AUTRE FAIT SOCIAL PRIS EN

Bien entendu, le monde associatif concerné (naturalistes, protecteurs de la nature), souvent partenaire, notamment pour des études et des suivis, perçoit très positivement ce qui précède. Peut-être moins attendue : l’exploitation de quatre cents questionnaires d’enquête58 retournés a montré à la rubrique « Ce qui plaît le plus aux habitants sur le territoire du Parc » que le premier rang revenait à « l’environnement naturel ». Si on considère que « les services de proximité » et « la proximité d’un pôle urbain », légitimes préoccupations matérielles, n’arrivent qu’aux quinzième et seizième places, soit en fin de classement, « la faune » au septième rang, après « les relations humaines » et avant « la flore », occupe un rang très honorable. En dépit de problèmes très réels de cohabitation avec quelques espèces (essentiellement Loup et Sanglier), le retour de la grande faune est globalement perçu par la grande majorité des habitants comme valorisant le territoire, sans qu’on puisse en réduire les motivations aux retombées touristiques escomptées ou à la possibilité de chasser certains Ongulés. Cette composante socio-culturelle, trop souvent négligée, mériterait d’être davantage prise en considération, car tout ce qui contribue à donner aux ruraux une perception positive du territoire qu’ils habitent est un facteur sociologiquement dynamisant.

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Emotion complexe : esthétique, certes, mais non pas exclusivement. Enquête de mai-juin 2006 dans le cadre du renouvellement décennal de la charte du Parc Naturel Régional du Vercors, bureau d’étude EDATER. 58

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La réussite des réintroductions antérieures, leur perception positive par l’opinion publique dominante, y compris au fil des ans de la part d’une fraction initialement réticente, constitue un précédent créant a priori positif en faveur de nouvelles opérations : facteur favorable à leur réussite. 3. Réintroductions : des compétences et des moyens Les projets de réintroduction de toutes espèces qui peuvent être portés ou soutenus par une structure officielle telle que Parc National, Réserve Naturelle ou Parc Naturel Régional disposent généralement de moyens de réalisation généralement bien supérieurs à ceux d’autres projets, par ailleurs tout aussi justifiés pour la restauration de la biodiversité. Le biotope de la zone d’étude le plus favorable au lâcher de Gypaète se trouve dans le PARC NATUREL REGIONAL DU VERCORS, aux confins de la RESERVE NATURELLE DES HAUTS PLATEAUX DU VERCORS, la plus vaste de France, gérée par le Parc, et d’une RESERVE BIOLOGIQUE FORESTIERE gérée par l’Office National des Forêts.

Photo T. PUJOL

Photo 54. – La Réserve Naturelle des Hauts Plateaux du Vercors, vue aérienne partielle .

vers le N-NE, d’un planeur à 1500 m. au dessus de la Grande Cabane. Pour une vue analogue de la partie sud cf. supra photo 33.

Ces statuts représentent de réels avantages, tant du point de vue du personnel disponible et de la réglementation en vigueur que de l’obtention de financements. Le statut de Réserve Naturelle de la partie du Parc Naturel Régional du Vercors de la zone plus favorable à la réalisation des lâchers donne des moyens réglementaires et de garderie pour prévenir ou supprimer d’éventuels perturbations très au-dessus de la moyenne, ainsi que pour la surveillance pendant le séjour des juvéniles au taquet et le suivi ultérieur, notamment pendant la phase d’acquisition de l’indépendance. Ces personnels sont déjà engagés dans le suivi de la reproduction du Vautour fauve. Projet de réintroduction du Gypaète Barbu dans les Préalpes occidentales: pertinence stratégique, faisabilité, biotope et site de lâcher au PNRV. Jean-Pierre Choisy 2010

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Le biologiste de terrain du Parc Naturel Régional du Vercors concerné par les réintroductions de Bouquetin et de Vautours fauve, par leur suivi et par celui du Percnoptère et du Vautour moine, a acquis une compétence reconnue, tant sur le plan technique que scientifique : par sa contribution au progrès des connaissances de l’éco-éthologie de ces espèces. Preuve en est la fréquence des consultations ou demandes de coopération qu’il reçoit en tant qu’expert pour : - critique constructive d’études de faisabilité de réintroduction du Bouquetin dans d’autres Parcs Naturels Régionaux français (et même un projet relatif au Chamois dans le Massif Central) ; - critique constructive d’études de projets de réintroduction de Vautour fauve en Roumanie, en Bulgarie, en Italie ; - co-rédaction d’un guide méthodologique pour la réintroduction du Vautour fauve, à destination de chargés d’études de faisabilité à l’étranger ; - échange et diffusion d’informations sur les mouvements de vautours en Europe à distance des population en Europe à la belle saison (isolés jusqu’en Scandinavie, groupes jusqu’en Allemagne et aux Pays-Bas). Il participe régulièrement aux réunions de travail, séminaires, colloques français et internationaux consacrés aux quatre espèce de Vautours d’Europe et prépare plusieurs articles sur la biologie de ces espèces et/ou leur réintroduction. Tant le biologiste que les gardes évoqués ci-dessus, outre qu’ils ont une grande pratique des trois autres espèces de Vautours d’Europe ont suivi le stage d’une semaine sur le Gypaète organisé dans les Pyrénées par l’Atelier Technique des Espaces Naturels avec la coopération du FIR – LPO. Dans le cadre du suivi local et à grande distance des trois autres espèces de vautours, un réseau très étoffé d’observateurs est animé, déjà collectant et rediffusion des informations sur les vautours de toutes espèces. Il est tout autant prêt à fonctionner pour le Gypaète et a d’ailleurs déjà fourni certaines des données de cette espèce dans la zone d’étude. La diffusion des actualités et documents se fait depuis la fin de 2007 via un site : http://www.parc-du-vercors.fr/blog-nature On y trouvera également documents et articles de diverses provenances. Ce réseau coopère déjà avec celui de l’IBM consacré au suivi du Gypaète dans l’ensemble de la chaîne alpine.

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CONCLUSIONS à ce niveau de l’analyse

LE TERRITOIRE DU PARC NATUREL REGIONAL DU VERCORS EST EMINEMMENT FAVORABLE A LA REALISATION D’UNE OPERATION REINTRODUCTION DU GYPAETE DANS LA ZONE D’ETUDE, PAR LES COMPOSANTES LES PLUS DIVERSES DE SON CONTEXTE LOCAL : ..- RESTAURATION SPECTACULAIRE DE LA GRANDE FAUNE en cours depuis des décennies ; ..- POLITIQUE DE CONSERVATION ET RESTAURATION DE LA BIODIVERSITE. Pour la grande faune, réintroductions de certaines espèces, gestion de difficiles problèmes de cohabitation avec d’autres, revenues seules ; ..- OPINION PUBLIQUE : perception largement positive tant du retour de la grande faune que de la politique ci-dessus ; ..- IMPORTANTES STRUCTURES OFFICIELLES DE GESTION - ET DE PROTECTION DE L’ESPACE NATUREL engagées dans ces politiques, dont le Parc NATUREL REGIONAL DU VERCORS qui dispose en matière de REINTRODUCTION de : a) COMPETENCES TECHNIQUES ET SCIENTIFIQUES ; b) MOYENS DE REALISATION ; c) BILAN : REUSSITES ANTERIEURES.

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Ph o t o J . T R A V E R S I E R

, O b s e r v a t o i r e d e l a f a u n e d r ô m o i s e C O R A F R A P N A

Photo 55. – Toutes les conditions sont réunies pour retour du Gypaète dans la zone d’étude : falaises, éboulis, bouquetins, contexte humain, moyens de réalisation et de suivi, Parc Naturel Régional du Vercors et Réserve Naturelle des Hauts Plateaux. Projet de réintroduction du Gypaète Barbu dans les Préalpes occidentales: pertinence stratégique, faisabilité, biotope et site de lâcher au PNRV. Jean-Pierre Choisy 2010

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CONCLUSION La restauration du Casseur d’os, dynamique dans les Alpes et les Pyrénées, est globalement à peine amorcée à l’échelle de l’Europe. Même dans les Pyrénées, elle est loin d’être achevée : cent quarante couples sur dix-neuf mille kilomètres carrés des Pyrénées ne font qu’un seul couple pour cent trente six kilomètres carrés. Dans les Alpes les effectifs actuels n’atteignent qu’un huitième de ceux des Pyrénées sur une aire dix fois plus étendue : le sous-peuplement relatif est donc énorme. Avec dix-sept couples, dont onze nicheurs, la pérennité du Gyapète dans les Alpes reste très fragile. Le lent accroissement de la reproduction en liberté dans les Alpes, l’achèvement opérationnel, atteint en Haute-Savoie, approché au Stelvio-Engadin, sont très encourageants. Mais un relâchement prématuré des efforts risquerait fort de les gaspiller. Ils doivent, au contraire, être exploités stratégiquement par la poursuite, et même l’intensification des lâchers. La très récente augmentation de la production de gypaètes en liberté en donne les moyens. Mais les effectifs disponibles de ces précieux oiseaux restent modestes par rapport aux besoins et leur coût individuel élevé. Deux raisons pour améliorer la stratégie de réintroduction sur la base des connaissances actuellement disponibles : le présent travail a pu s’appuyer, aux échelles spatiales les plus diverses, sur une somme de connaissances, sur des publications récentes et solides, dont on trouvera certaines en ANNEXES. Certes, la réintroduction du Gypaète reste une démarche expérimentale, n’excluant pas totalement l’imprévu et dont le suivi continue à enrichir la connaissance de l’espèce. Mais on est désormais très loin de l’époque « héroïque » des pionniers59. Deux nouvelles opérations sont à la fois possibles et, par leur position géographique, stratégiquement prioritaires à l’échelle des Alpes comme de l’Europe. L’une d’elles est l’objet du présent travail. Elle seule bénéficie actuellement d’un état d’avancement du projet et d’une structure, le territoire du Parc Naturel Régional du Vercors, permettant un très prochain passage à l’acte. De ce fait, elle disposera donc chaque année d’un contingent étoffé d’oiseaux : un facteur essentiel du rendement démographique des lâchers.

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Le succès actuel de la réintroduction du Vautour fauve dans le sud de la France, du Gypaète dans les Alpes ne doit pas faire oublier les premières tentatives. Outre la hardiesse des pionniers, leur détermination leur valeur d’exemple, il faut souligner que, malgré leurs échecs, ces premières expériences ont été riches d’enseignements. « Imaginée dès le début de ce siècle (le XXe), notamment par A. RICHARD, premier président de « Nos Oiseaux », la réintroduction du Gypaète dans les Alpes ne s’est concrétisée qu’à parttr des années septante. Un premier projet franco-suisse vit le jour sous l’égide, entre autres, de P. GEROUDET » (Arlettaz) et de G. AMIGUES. Invités d’honneur du colloque au Grand Bornand marquant les 20 ans de la réintroduction l’automne 2006, ces deux pionniers ont présenté un récit vivant et remarqué de cette première tentative mondiale, quelques semaines avant que le premier soit enlevé à l’admiration, à la reconnaissance et à l’affection des naturalistes, francophones et autres. Projet de réintroduction du Gypaète Barbu dans les Préalpes occidentales: pertinence stratégique, faisabilité, biotope et site de lâcher au PNRV. Jean-Pierre Choisy 2010

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Éventuellement complétée dans les Causses, une opération à l’extrême ouest des Alpes sera une étape essentielle pour la constitution en une même métapopulation des gypaètes de cette chaîne et de ceux des Pyrénées. Ceci, tant du point de vue de la dynamique de population que de la diversité génétique et de la probabilité de pérennité, constituera une amélioration considérable pour chacune de ses composantes comme pour ce vaste ensemble, central dans la poursuite de la restauration de l’espèce en Europe.

Photo R. MATHIEU, Observatoire de la faune drômoise CORA-FRAPNA

Photo 56. – Couple de Gypaète Gypaetus barbatus dans les Pyrénées espagnoles. PARTICIPER A LA REINTRODUCTION DU GYPAETE EST DANS LE DROIT FIL DE LA POLITIQUE PERSEVERANTE DE RESTAURATION DE LA BIODIVERSITE DU PARC NATUREL REGIONAL DU VERCORS SUR SON TERRITOIRE, AVEC D’AUTRES RETOMBEES LOCALES TRES POSITIVES. AUX

ECHELLES ALPINE ET CONTINENTALE CETTE OPERATION LUI UN ROLE DE PREMIER PLAN A JOUER, A LA HAUTEUR DE CETTE RESPONSABILITE.

OFFRE

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N’ Y FAUT QU’INTELLIGENCE STRATEGIQUE ET VOLONTE POLITIQUE, DONT ON NE SAURAIT DOUTER.

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Mingozzi T. & Estève R. (1997) Analysis of a historical extirpation of the Bearded Vulture (Gypaetus barbatus) in the western Alps (France-Itay) : former distribution and causes of extirpation. Biological Conservation 79, 155-171. Mathieu R. & Choisy J.-P. (1982) L’Aigle royal (Aquila chrysaetos) dans les Alpes méridionales françaises de 1964 à 1980. Essai sur la distribution, les effectifs, le régime alimentaire et la reproduction. Le Bièvre IV-1, pp 1-32. Otto S., Greßmann G. & coll. (2006). Le Gypaète barbu dans les Alpes. Nationalparkrat Hohe Tauern & Foundation for the Conservation of the Bearded Vulture. Schaub M., Zink R., Beissmann, Sarrazin F. & Arlettaz R. (2009). When to end releases in reintroduction programmes : demographic rates and population viability analysis of bearded vultures in the Alps. Journal of Applied Ecology, 46 92-100. Terrasse J.-F. (2001) avec la coopération de Coton C. et Géroudet P. Le Gypaète barbu. Description, moeurs, observation, mythologie Ed. Delachaux & Niestlé. 208 pp. Terrasse M. (2006). Evolution des déplacements du Vautour fauve Gyps fulvus en France et en Europe. Ornithos 13-5 : 273-299 Tessier C., Henriquet S. & Eliotout B. (2003) Le Vautour moine Aegypius monachus dans les Préalpes provençales. Etude de faisabilité de la réintroduction du Vautour moine dans le massif des Baronnies (Drôme) et les Gorges du Verdon (Alpes de Haute Provence). Association Vautours en Baronnies & Ligue pour la Protection des Oiseaux Provence-Côte d’Azur. 64 pp. Villaret J.C. (1987) Projet de réintroduction du Bouquetin en Isère ; sélection des sites. Direction Départementale de l’Agriculture de l’Isère & Centre Ornithologique Rhône-Alpes. 119 pp. Wilmers C., D. Stahler, R. Crabtree, D. Smith & W. Getz (2003a) Resource dispersion and consumer dominance : scavenging at wolf- and hunter-killed carcasses in Greater Yellowstone, USA. Ecology Letters 6 : 996-1003 Wilmers C., D. Stahler, R. Crabtree, D. Smith, K Murphy & W. Getz (2003b) Trophic facilitation by introduced predators : grey wolf subsidies to scavengers in Yellowstone National Park. Journal of Animal Ecology 72 : 909- 916 Wilmers C. & W. Getz (2004) Simulating the effect of wolf-elk population dynamics on resource flow to scavengers. Ecological Modelling 177 : 193-208 Wilmers C. & W. Getz (2005) Gray wolves as Climate Change Buffers in Yellowstone. PLoS Biol Vol. 3(4), e92. 571-576 +

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Projet de réintroduction du Gypaète Barbu dans les Préalpes occidentales: pertinence stratégique, faisabilité, biotope et site de lâcher au PNRV. Jean-Pierre Choisy 2010

ANNEXES I

ARTICLES ET DOCUMENTS HORS PARC NATUREL RÉGIONAL DU VERCORS

Schaub M., Zink R., Beissmann, Sarrazin F. & Arlettaz R. (2009) When to end releases in reintroduction programmes : demographic rates and population viability analysis of bearded vultures in the Alps. Journal of Applied Ecology, 46 92-100.

Hirzel A., Posse B., Oggier P.-A., Crettenand Y., Glenz C. & Arlettaz R. (2004) Ecological requirements of reintroduced species and the implications for release policy: the case of the bearded vulture.Journal of Applied Ecology 41, 1103–1116 Blackwell Publishing, Ltd.

Arroyo B. & Razin M. (2006) Effect of human activities on bearded vulture behaviour and breeding success in the French Pyrenees. Biological Conservation 128, 276-284.

Heuret J. (1999) Choix du nouveau site de réintroduction des gypaètes barbus en HauteSavoie. Programme LIFE Nature, Conservation du Gypaète barbu dans les Alpes françaises. 4 pp.

Geng M., Heuret J. & Rouillon A. (2000). – Protocole commun de réintroduction et de suivi des Gypaètes barbus dans les Alpes françaises. ASTERS & P.N. du Mercantour.

Journal of Applied Ecology 2009, 46, 92– 100

doi: 10.1111/j.1365-2664.2008.01585.x

When to end releases in reintroduction programmes: demographic rates and population viability analysis of bearded vultures in the Alps

Blackwell Publishing Ltd

Michael Schaub1,2*, Richard Zink3,4, Helmut Beissmann5, François Sarrazin6 and Raphaël Arlettaz1,2,7 1

Institute of Ecology and Evolution, Division of Conservation Biology, University of Bern, Baltzerstrasse 6, CH-3012 Bern, Switzerland; 2Swiss Ornithological Institute, CH-6204 Sempach, Switzerland; 3Research Institute of Wildlife Biology, University of Veterinary Medicine, Savoyenstrasse 1, A-1160 Vienna, Austria; 4International Bearded Vulture Monitoring, Hohe Tauern National Park, A-9844 Heiligenblut, Austria; 5Konrad Lorenz Institute for Comparative Ethology, Austrian Academy of Science, Savoyenstrasse 1, A-1160 Vienna, Austria; 6UMR 5173 MNHN-CNRS-UPMC Conservation des espèces, restauration et suivi des populations, Muséum National d’Histoire Naturelle, 61 rue Buffon, 75005 Paris, France; and 7The Ecology Centre, University of Queensland, St Lucia, Qld 4072, Australia

Summary 1. Reintroductions are commonly used for re-establishing self-sustainable populations in formerly inhabited areas. Reintroductions are expensive, and thus, it is worth performing a thorough demographic analysis of current and likely future population trajectories to guide strategic decisions on release policy. 2. Bearded vultures Gypaetus barbatus were exterminated from the Alps in the late 19th century, mainly due to human persecution. To re-establish them, captive-bred young have been released annually since 1986. Since the first successful breeding in the wild in 1997, the population has increased to 9 pairs in 2006. It is not known, however, for how long releases should be continued to obtain a self-sustaining, viable population. 3. We estimated age-specific survival probabilities with a mark–resighting model and quantified fecundity rates of released individuals. Using the resulting demographic estimates, we built a stochastic population model to estimate population growth rates, and explored the value of continuing to release birds for varying periods into the future. 4. Annual survival probabilities were high (first year of life, 0·88; later years, 0·96); average annual fecundity was 0·6 fledglings per breeding pair. Using the estimated survival probabilities, projected population growth rates would increase with additional years of releases. Yet, the population would grow, even if releases had stopped after 2006. Only if mortality increased by ≥ 50% would the population start to decline. 5. Synthesis and applications. Our population dynamics model provides essential information to optimize decision-making within a major reintroduction programme. From a demographic viewpoint, releases of captive-raised bearded vultures can be ceased in the Alps. The resources freed could be redirected towards a close demographic surveillance of the free-ranging population, with periodic evaluation of its viability and the option to release birds if deemed necessary. Birds available from the captive stock could be used for reintroductions in other areas where the bearded vulture is extinct. Key-words: Alps, conservation, fecundity, Gypaetus barbatus, population growth rate, parameter uncertainty, survival probability

*Correspondence author. E-mail: [email protected] © 2008 The Authors. Journal compilation © 2008 British Ecological Society

Population viability of bearded vultures 93

Introduction Animal population reintroductions and translocations are likely to become a key tool in conservation biology in the 21st century (Sarrazin & Barbault 1996; Seddon, Armstrong & Maloney 2007). Reintroductions usually involve the intentional release of individuals from captive-reared stock into a species’ historical range, or translocation of individuals from thriving populations into relict populations. Reintroductions make sense only when the principal cause of extinction has been eliminated (Griffith et al. 1989). Although reintroductions and translocations are currently widely used to reinstall or restock populations, strategic decisions about release policy within such programmes are still too often based on empirical rules of thumb rather than on appropriate, quantitative scientific assessment. This is often associated with a lack of clearly defined quantitative goals and/or insufficient monitoring of the success or failure of the chosen management (Sarrazin & Barbault 1996; Seddon 1999; Armstrong & Seddon 2008). Several techniques developed by population biologists exist, which can assist in taking appropriate strategic decisions (Norris 2004). Often, released and translocated animals are individually marked, and therefore, vital rates can be estimated using capture–recapture models. Knowledge of these rates allows us to conduct population viability analyses which can provide decisive insights into management (Beissinger & Westphal 1998). Quantitative demographic analyses of reintroduced species are scarce (Sarrazin & Barbault 1996) and biased towards successful projects (Seddon, Armstrong & Maloney 2007). In order to orient future strategic decisions, we applied the demographic approach to the bearded vultures Gypaetus barbatus (Linnaeus) which have been reintroduced into the European Alps. This reintroduction programme is one of the largest and most publicized European reintroduction projects ever conducted. The bearded vulture is a large (4·5–7·1 kg) scavenging raptor that mainly feeds on bones of medium-sized wild and domestic ungulates, and inhabits mountain ranges in Eurasia and Africa. It went extinct in the Alps between the late 19th and the early 20th century (Mingozzi & Estève 1997) mainly due to shooting and poisoning. In 1986, an international reintroduction programme, based on the release of birds born and reared in captivity was launched (Frey 1992). By 2005, 137 individuals had been released. The first successful reproduction of released birds in the wild took place in 1997, and by 2006, 9 breeding pairs were established across the range, with habitat preferences for limestone areas with abundant populations of ibex Capra capra L. and chamois Rupicapra rupicapra L. (Hirzel et al. 2004). It has been proposed that releases in the Alps should be ceased as soon as the mean yearly number of wild-born fledglings equalled the average number of yearly released young (n = 6·5; Zink 2005a). A linear model of the number of wild-born fledglings against year predicted that this number would be greater than 6·5 by 2007. This strategy may be erroneous, as it focuses on productivity alone, a parameter whose relevance for population dynamics in a long-lived species is likely to be low

(Lebreton & Clobert 1991), and as it does not consider other relevant demographic parameters. Moreover, as the annual number of released individuals is used as the target, the management decision is not objective: the ultimate goal is the establishment of a naturally, self-sustaining population in the wild. This study aims to estimate for how long further releases of young will be necessary for ensuring the long-term viability of the Alpine bearded vulture population. However, we focused on the establishment of a self-growing population (Armstrong & Seddon 2008) as a first step towards viability, without considering density dependence since we had no reliable estimate of the carrying capacity. We used a demographic model that incorporates all key demographic parameters estimated directly from data on the released individuals. We explicitly considered uncertainty in the parameter estimates for the population modelling to ensure careful management recommendations (Ellner & Fieberg 2003). Such a demographic assessment is also central for an optimal allocation of financial resources as every young has accumulated costs of up to A70 000 by the moment of its release (Frey 1998). Finally, we estimated the sensitivity of the population growth to changes in survival probabilities to explore the possible impact of an increased use of illegal poisoned baits. These may be used against the naturally expanding wolf Canis lupus L. population in the Alps (Valière et al. 2003), and might represent a serious threat to bearded vultures.

Methods RELEASE OF YOUNG

Young bearded vultures reared in different zoos were released at an age of about 3 months (~3 weeks before fledging) in artificial eyeries at four sites well scattered across the entire Alpine range (Frey 1992). Starting in 1986, up to three birds were released per site annually, amounting to 137 birds released by 2005. The released birds were fed artificially until they were independent. All birds were marked individually prior to release with colour rings and with an individual pattern of bleached wing or tail feathers. The latter marks enable individual recognition until the termination of the first moult (until 2–3 years of age; Arlettaz 1996).

DATA COLLECTION

Throughout the Alps, professional ornithologists and hundreds of volunteer birdwatchers have monitored movements of the released birds since the beginning of the release programme. The birds were monitored before the start of wing and tail feather moult at 1–2 years of age, using the patterns of bleached feathers, and later by recording the individual colour ring codes. Moulting patterns, if discernable on pictures of birds in flight, were used when a good time series of photographic documentation was available for a given bird (Arlettaz 1996). In addition, recoveries of dead birds were recorded. Observations were transferred into a central data bank (International Bearded Vulture Monitoring, Vienna; Zink 2005b), where a reliability check was performed; specifically, double entries for a same bird in distant areas on the same day were eliminated, as they were indications of misidentifications. Such errors were scarce (< 5% of the observations). We restricted our analysis to observations relating to birds of certain identity.

© 2008 The Authors. Journal compilation © 2008 British Ecological Society, Journal of Applied Ecology, 46, 92–100

94 M. Schaub et al. The general survey demonstrated the dispersal potential of the species: many individuals moved several hundred kilometres from the release sites, and some individuals eventually settled far from these sites (Arlettaz 1996; Hirzel et al. 2004; Zink 2004). As a consequence, we consider the whole Alpine population as a single functional demographic unit in our analysis. Monitoring of pair formation and reproductive success was conducted by trained biologists. The good spatial and temporal survey coverage as well as a high number of sightings of identified birds by independent observers within a given area suggests that the chance of failing to locate a territorial pair is low, and the chance of failing to detect a breeding pair is close to zero.

ESTIMATION OF SURVIVAL PROBABILITIES

From 1986 to 2005, 137 individuals were released, five died before they fledged and 132 individuals were included in our analyses. Naturally born individuals (33 up until 2006) were not considered, because they were not marked individually. To estimate survival probabilities, we considered mark–resight data of the 132 individuals from 1986–2005, and dead recoveries up until May 2006 from the whole Alps. Only resightings (n = 250) from the months June–October in each year were included in order to meet the assumption of capture–recapture models that resightings shall be obtained within a short period of time. Additionally, 17 dead recoveries from throughout the year were considered. We used a probabilistic multistate capture–recapture model (Nichols et al. 1992) to estimate annual survival probabilities jointly from the mark–resighting data and the dead recoveries. The model was constructed in such a way that an immediate resighting effect could be modelled (i.e. individuals that were seen in the preceding year had a higher probability to be seen in the current year than individuals that were not seen in the preceding year), which was detected by a goodness-of-fit test. We used U-CARE (Choquet et al. 2001) and E-SURGE (Choquet, Rouan & Pradel 2009) to analyse these data (see Supporting Information, Appendix S1).

(denoted a4), the second assumes equal survival probabilities in the establishment and the territorial phases (a3), and the third assumes equal survival probabilities in the prospecting, establishment and territorial phases (a2). We also considered models in which survival probabilities varied across years. Only additive models (i.e. annual variations were the same in all age classes) were included due to their low number of parameters. Interactive models would have had very limited power to detect differential temporal variation for each age class given the small sample size. Resighting probabilities were likely to depend on the age of the individuals because bleached feathers are lost during the first moult and because of the different behavioural patterns during the four phases described above. For example, during dispersal, birds may move to sites where fewer observers are active, which would decrease detection probability. We considered the same three age-class models as above for the resighting probabilities, as well as an additional model where the resighting probabilities did not depend on age. We always considered an immediate resighting effect, and also included models where the resighting probabilities had an additive time effect. If temporal variation in resighting had been present, it would have been induced by varying resighting effort, which would affect all individuals in a similar way. We considered two age classes for the dead recovery probability. The probability of recovering dead birds might be higher in the first year than in later years because released birds usually remain close to the release sites, where the observation effort is much higher. We also included models without an age effect and models with additive time effects. We conducted model selection in two steps because of the potentially large number of models. For the first modelling step, we considered 33 models (combination of 8 models for resighting with 4 models for dead recovery, plus a model close to that used to assess the goodness-of-fit). We identified the smallest set of models whose Akaike weights (wi) sum to 0·95 (95% confidence set). In the second modelling step, we combined the structures of resighting and dead recovery included in the 95% confidence set with the six a priori defined models for survival. Finally, we calculated the model averaged mean for the parameters of interest based on the wi.

MODEL SELECTION

Our aim was to obtain reliable survival estimates of released bearded vultures in order to perform a population viability analysis. We formulated different models and performed model selection based on the Akaike’s Information Criterion adjusted for small sample size and overdispersion (QAICc, Burnham & Anderson 2002). The life history of bearded vultures can be decomposed into four phases: juvenile, prospecting, establishment and territorial phases, respectively. The first year of life corresponds to the juvenile phase, when young become progressively independent from the adults. The next 2 years (2–3 years of age) can be considered as a dispersal phase, when immatures prospect and evaluate the landscape on a broad scale. Subadults at 4–5 years of age enter an establishment phase when they get more and more sedentary. Finally, in their sixth year of life, bearded vultures become fully territorial, adopting a definitive adult plumage; this is also when first breeding attempts may take place. Survival probabilities might differ between these phases. For instance, we expected lower survival in the juvenile and prospecting phases, compared to the establishment and territorial phases, because there are risks when birds enter an unfamiliar environment without the assistance of parents. Based on these life phases, we considered three different models for the age-specific changes in survival. The first model considers different survival probabilities for each phase

ESTIMATION OF FECUNDITY

Fecundity (Ft) was estimated as the production of fledglings in year t divided by the number of territories occupied by adult pairs in year t. A territory was considered as occupied from the year of first successful breeding onwards as long as an adult pair was present.

POPULATION MODELLING

Based on our life-history trait estimates, we constructed a post-breeding census, stage-classified projection model (Caswell 2001) with 12 stages. Six stages refer to the six age classes before maturity (J and 1–5; Fig. 1), four stages refer to mature individuals that have not yet reproduced (6–9), one stage refers to breeders (B) and one stage refers to mature non-breeders (NB). Survival probabilities were agedependent as identified in the survival analysis. We assumed that reproduction starts at 6 years of age (Bustamante 1996; Brown 1997; Antor et al. 2007). Each year, half of the still inexperienced breeders (classes 6–9, Fig. 1) start to reproduce (α = 0·5), and at 10 years of age all are assumed to have reproduced at least once. Always present in bearded vulture populations (Carrete et al. 2006), non-breeders were incorporated by assuming that a fraction of the potential

© 2008 The Authors. Journal compilation © 2008 British Ecological Society, Journal of Applied Ecology, 46, 92–100

Population viability of bearded vultures 95 Fig. 1. Sketch of the life cycle of a bearded vulture population with post-breeding census as used for the present modelling. The nodes refer to the different stages with J, juveniles; 1–9, 1 to 9-year-old, inexperienced breeders; B, breeders; NB, non-breeders. The recruitment transitions are shown with broken lines, the survival transitions with solid lines. S1, first year survival; S2, annual survival from age 1 to 3; S3, annual survival from age 3 to 5; S4, annual adult survival; F, reproductive success; sr, sex ratio; α, probability that an as yet inexperienced mature individual starts to reproduce; δ, breeding probability. Note that only the female segment of the population is shown; the complete model includes males as well. breeders skips reproduction each year. Furthermore, we assumed that fecundity of the reproducing individuals does not change with age. We explicitly modelled both sexes, since random deviations from an even sex ratio can affect population growth negatively in a small population (Legendre et al. 1999). The number of breeding pairs in a given year was assumed to be equal to the smallest number of reproducing individuals of either sex. In order to incorporate demographic stochasticity, we modelled survival, fecundity, probability of starting reproduction, and breeding probability as binomial processes for each sex independently. The binomial process for fecundity was chosen because bearded vultures have at most one fledgling per year (Margalida et al. 2003). Mathematical details for the population model are provided in Supporting Information, Appendix S2. Based on the estimated survival probabilities and on the number of released and wild born individuals, we calculated the number of individuals theoretically alive in each age class by 2006. The estimated number of experienced breeders amounted to 50 individuals (25 breeding pairs). However, we used the actually observed number of breeding pairs in year 2006 (9) and assumed that the remaining individuals will never reproduce. This leads to a very conservative scenario. Based on this initial stage-specific population size vector (7, 7, 6, 4, 5, 3, 2, 1, 0, 0, 9, 0) for each sex, we used simulation to model the population development. We modelled population growth over the next 25 years, since this is a time horizon relevant to management recommendations regarding future release policy. We estimated the population growth rate with a linear regression model of the logarithm of the annual number of breeding pairs against time (Caswell 2001). Ten thousand populations with these features were simulated to generate mean and 95% confidence intervals of the population growth rate. Simulations were performed in  ( Development Core Team 2004), and code is available in Supporting Information, Appendix S3. To account for the uncertainty regarding the estimated survival probabilities in the population modelling, we generated for each iteration specific values from a beta distribution using the model averaged estimates of the mean and the variance of all age-specific survival probabilities. Uncertainty of fecundity was accounted for by creating for each iteration a binomial random variable using the total number of fledglings and the total number of breeding events as parameters. These generated parameter values were held constant across time for the given iteration. Our main interest was to decide whether further releases of young are essential to ensure an optimal population development, and if so, for how long releases should continue. We therefore considered scenarios reflecting various durations of releases in the future. First,

we assumed that no further release took place after the releases in 2006. In the next cases, we assumed that 3 females and 3 males would be released each year for the next 5, 10 or 25 years. The uncertainty about some demographic parameters (fecundity, breeding probability) was accounted for in different scenarios. Our sample size was too small to test whether fecundity was constant or stochastic; consequently, we considered two options. In the constant case, we used the observed mean reproductive success (F ) in the simulations. To model environmental stochasticity, we randomly chose a yearspecific, observed fecundity ( Ft). Because the fluctuations were very wide when the population size was low, we only considered annual fecundities from 1999 onwards. Further uncertainty surrounds the breeding probabilities (δ). We considered four scenarios (δ = 1, 0·8, 0·6 or 0·4) to include a range from optimistic to pessimistic. The value of δ = 1 is very optimistic and unlikely to be true, since floating non-breeders occur in many bearded vulture populations (Carrete et al. 2006). By contrast, the value of δ = 0·4 is pessimistic in the long term, although it is similar to what is observed currently in the Alps. Based on observations in the Pyrenees, we regard δ = 0·8 to be the most realistic value (Carrete et al. 2006). Thus, in total we considered 32 different scenarios (4 different duration of releases × 4 different breeding probabilities × 2 different fecundities).

Results ESTIMATION OF SURVIVAL PROBABILITIES

Modelling of the resighting and dead recovery probabilities revealed that five top-ranked models yield a summed QAIC weight of > 0·95 (Supporting Information, Table S1). In these models, neither resighting nor dead recovery probabilities were time-dependent, but there was considerable uncertainty regarding the age structure in both parameters. We combined the resighting and dead recovery structure of these five models with the six different models of survival, obtaining 30 models in the final modelling step (Supporting Information, Table S2). Modelling uncertainty was considerable again, in particular regarding the age structure of the resighting and dead recovery probabilities. The four top-ranked models had constant survival probabilities across time and incorporated two age classes (first year vs. older individuals, QAIC weights summing up to 0·56), whilst lower-ranked models differentiated between three to four age classes.

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Fig. 2. Model-averaged probabilities of survival (a), resighting (b) and dead recovery (c) for bearded vultures in the Alps in relation to their age and previous resighting history. The over-dispersion coefficient was 9 = 1·224. The vertical lines show the unconditional 95% confidence intervals.

The model averaged survival probabilities were lowest in the first year, as expected. Thereafter, however, they did not change much with increasing age (Fig. 2a). The 95% unconditional confidence intervals were relatively wide, reflecting uncertainty in the parameter estimates. The model averaged resighting probabilities were highest in the year immediately following the release (Fig. 2b). If an individual was not seen in a year, the probability of recording it in the subsequent year, given that it survived, was very low. The dead recovery probabilities for young birds were higher than for older birds (Fig. 2c), but their confidence intervals were very large.

FECUNDITY AND POPULATION DEVELOPMENT

Bearded vultures reintroduced into the Alps from 1986 started to reproduce in the wild in 1996, totalling 55 breeding events with 33 fledglings by 2006. Average fecundity was thus 0·6, but there were considerable annual fluctuations (Fig. 3).

Fig. 3. Number of breeding pairs, total number of fledglings, and fecundity (number of fledglings per territorial breeding pair and year) of bearded vultures in the Alps from 1996 to 2006.

ASSESSMENT OF DIFFERENT RELEASE STRATEGIES

The projected average population growth rates over the next 25 years were > 1, regardless of the duration of releases, the different options for fecundity and the different breeding probabilities (Fig. 4). They increased with increasing duration of releases. Mean population growth rates were higher when fecundity was constant than when affected by environmental stochasticity, but the difference was marginal. Increasing breeding probability affected population growth positively, but this effect declines the longer the releases continued. The confidence intervals of the population growth rate covered 1 only in the situation where releases stop immediately after 2006 and when the breeding probability is very low (0·4). Thus, a population decline cannot be ruled out completely under this pessimistic scenario, although it remains improbable. The population growth rate strongly declined with increasing mortality (Fig. 5). Mortality needs to increase more than about 50% to render the mean population growth rate less than 1, indicating that such an increase can be supported even if no further individuals were released after 2006 and even if the breeding probability was low.

Discussion This study represents the first attempt to estimate life-history traits of a free-ranging population of bearded vultures in natural conditions with reliable methods. This was possible because individuals were systematically marked from the beginning of the reintroduction and because we applied modern demographic estimation and analytical methods. Based on empirical estimates of vital rates, we could evaluate the growth rate of the Alpine bearded vulture population over the next 25 years under different durations of releases of captive-reared young while accounting for uncertainty in the estimates of the demographic parameters. Overall, our model suggests that the population will further increase even if releases cease after 2006, if breeding probability was low and if mortality increased slightly. We are confident that releasing young can cease without endangering the established population, as long as new factors of mortality such as poisoned baits does not increase mortality by more than 50%, fecundity remains the same on average and no catastrophic events occur.

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Population viability of bearded vultures 97

Fig. 4. Estimated population growth rates of Alpine bearded vultures averaged over 25 years in relation to different release strategies (no further releases after 2006, or releases of 3 males and 3 females each year for another 5, 10 and 25 years) and to different breeding probabilities, when fecundity is either constant or affected by environmental stochasticity. The vertical lines show the limits of the 95% confidence intervals.

Fig. 5. Sensitivity of the population growth rate (averaged over 25 years) of Alpine bearded vultures to a linear increase of mortality rates across all age classes with variable breeding probabilities, when fecundity is constant and when no individuals were released. The starting values without increase of mortality (0 on the x-axis) refer to the average survival rates (Fig. 2). The vertical lines show the limits of the 95% confidence intervals.

Our population model is based on some simplifying assumptions, which may all impact on the modelling results. First, we did not consider environmental stochasticity for survival and age at first reproduction. If stochasticity occurs, the population growth rates would be lower (Tuljapurkar 1989)

and our conclusions too optimistic. However, given the survival probabilities that are close to 1 and the high sensitivity of population growth rate to survival, temporal variation of adult survival is expected to be small, and therefore, it is likely that any overestimation of the population growth rate is only slight. The sensitivity of the population growth rate to changes in the probability of starting to breed (α) is low, and therefore, variability needs to be strong in order to have any significant impact on population dynamics. Secondly, we assumed that adults that had not yet reproduced in the year 2006 would not reproduce in the future. If this assumption is wrong, which is very likely, then the population growth rate would in fact be higher than our conservative estimates. Taken together, we believe that the population growth rate estimates presented here are realistic. Population growth is generally highly sensitive to changes in adult survival in long-lived species (Lebreton & Clobert 1991). Several studies have confirmed this to be the case for bearded vulture populations (Bustamante 1996; Bustamante 1998; Bretagnolle et al. 2004; this study) but previously, little was known about actual survival probabilities of free-ranging bearded vultures. Based on reliable methods, our estimates of survival recognize two age classes, where first year survival is slightly lower (0·88) than thereafter (0·96). A larger sample size and longer time series would be necessary to get more precise estimates, to detect finer age-structures and to assess the magnitude of temporal variation of survival probabilities. Moreover, data from naturally born individuals would be required to test if the release has costs in terms of survival, as observed in griffon vultures Gyps fulvus Hablizl (Sarrazin & Legendre 2000; Le Gouar et al. 2008). Brown (1997) estimated survival of bearded vultures from South Africa using age ratio methods, which produces accurate estimates only under restrictive assumptions (Conn, Doherty & Nichols 2005). He obtained similar survival estimates for adults as the current study, but much lower estimates for young individuals during their first 4 years of life (~0·6). The only other information about species-specific survival probabilities comes from zoos, where bearded vultures appear to survive better than in the wild (year 1, 0·92; years 2–6, 0·99; year 6+, 0·97; Bustamante 1996). Although we have estimated survival probabilities using a method which accounts for imperfect detection of both live and dead individuals, the precision of our estimates was relatively low. This mainly reflects the relatively small sample size (low number of individuals) and the heterogeneity of resighting probabilities among individuals. Although the monitoring was very intense, it was difficult to re-sight individuals once they had remained undetected for a year. Once they lost their conspicuous bleached wing feathers due to moulting, the resighting probability dropped dramatically because identification was then dependent on recording their colour rings or slight details in the moulting pattern, which is more difficult. Once birds became territorial, the probability of identification increased again because they mostly stayed within the same territory and observers invested a great deal of effort to identify territorial birds. Thus, resighting declined after the initial wing feather moult (second and third year of life,

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i.e. prospecting phase) and increased during the establishment and territorial phases, which is well modelled with the trapdependent resighting probabilities applied in this study. As population dynamics are very sensitive to variation in survival and because little is yet known about the important demographic aspects (e.g. magnitude of temporal variation, sex-specific variation, survival of wild-born individuals), survival is a key parameter for future monitoring. The number of natural births will probably increase in the Alps in the near future, but so far naturally fledged birds have not been marked in order to avoid disturbances. This impedes any chance of monitoring survival of the reconstituting population. Genetic sampling (e.g. of feathers collected in the eyrie after fledging) seems to be a promising non-invasive method (Gautschi et al. 2000). If a mark–resighting-based monitoring programme, as the one developed here, is to be applied to genetic data, it must be ensured that repeated samples across years are collected from young and adult birds. In addition, sufficiently large sample sizes will be required, otherwise it will be difficult to detect small but relevant variations in survival probabilities. Furthermore, genetic tracking will presumably be expensive and its coverage will be less comprehensive than resightings of individually marked birds. In this respect, integrated population models, which combine demographic information from different sources may actually be useful to make the most efficient use of this diverse information (Schaub et al. 2007). The average fecundity of 0·6 fledglings per breeding pair per year is a rough estimate due to the small size of the Alpine population. Yet, this value is within the range observed in other areas. In the Pyrenees, fecundity declined with increasing population density from about 0·8 to 0·4 within 25 years (Carrete, Donazar & Margalida 2006). Bearded vultures were more productive in South Africa (0·89, Brown 1997), but much less so in Corsica (France, 0·18, Bretagnolle et al. 2004) than in the Alps. As evidenced in other raptors (e.g. Krüger & Lindström 2001), fecundity in bearded vultures may decline with increasing density due to habitat heterogeneity (suboptimal habitats colonized secondarily) and/or interference (Carrete, Donazar & Margalida 2006; Carrete et al. 2006). To date, there is no indication that fecundity is regulated by density in the Alps. Based on our estimated survival probabilities and on the number of naturally born and released individuals, there should be 50 mature (at least 6 years old) individuals alive in 2006. Only 18 (9 breeding pairs, 36%) of these 50 adults were actual breeders that year, which in theory gives 32 additional mature individuals. Although non-breeding floaters are common in bearded vulture populations (Carrete et al. 2006), the comparative figure for the Alps seems to be very high. An important reason why so many mature individuals do not reproduce could be due to inverse density-dependent phenomena such as Allee effects (Derdedec & Courchamp 2007). First, the slightly biased sex ratio in the released individuals (43% males, 57% females, n = 118 sexed individuals), could lead to mating problems typical of small populations scattered over a huge area. Secondly, the local density of available

partners may still be too low in the Alps to allow mating choice and pairing to operate properly. For example, two of the released individuals only started to reproduce in their 13th and 17th year of life, respectively, which is an unusually old age for first reproduction in bearded vultures (Brown 1997; Antor et al. 2007). The observed annual population growth rate calculated from the number of breeding pairs between 1999 and 2006 was much higher (1·245) than the highest estimate drawn from our model (1·113). This difference cannot be explained by demographic mechanisms (e.g. immigration). The most likely reason is a sudden acceleration in the formation of new pairs and reproduction thereof, as the likelihood of new pair formations increases as a non-linear function of the number of mature birds within the population if an Allee effect occurs. We thus predict a decline in the proportion of non-breeding adults in the future, when this initial boosting mechanism will be over, with a progressive decrease of the observed population growth rate to values similar to those of our population modelling. Although the Alpine bearded vulture population is presently increasing, with further releases judged superfluous, caution must be exercised with regard to any additional alteration of survival. Our model shows that the population would currently be capable of sustaining a 50% increase in mortality, even at very low breeding probabilities, which provides a buffer against potentially new emerging threats. Yet, in a population which consists of 50 individuals older than 6 years, an increase from 2 to 3 yearly fatalities would already lead to critical mortality levels. There is thus a real risk that the illegal practice of depositing poisonous baits against wolves currently recolonizing the Alps may obliterate the reintroduction effort. Tight monitoring of the poisoning situation is therefore essential to protect the bearded vulture population.

RECOMMENDATIONS FOR FUTURE MANAGEMENT

The model presented here shows that continued release of young bearded vultures into the Alps would enhance population growth rate, corroborating previous predictions (Bustamante 1998; Bretagnolle et al. 2004). However, the present analysis also demonstrates that the population has been self-sustainable since 2006. From a purely demographic viewpoint, we therefore recommend ending releases in the Alps and redirecting reintroduction efforts towards other areas where the species is now extinct (e.g. Sardinia, Balkans). Demographic management in the Alps should now concentrate on close surveillance of the breeding pairs, with systematic collection of data on fecundity and survival probabilities of wild-born birds, which may differ from captive-reared birds. This requires systematic marking of young plus sampling of genetic material at the eyrie, practices avoided so far to minimize disturbance. In the future, analyses combining demographic and genetic information should be performed periodically. Further releases from captive populations should remain an option if the wild population declines in the future.

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Population viability of bearded vultures 99 This study illustrates the relevance of detailed population modelling studies for orienting strategic decisions in largescale reintroduction programmes. Even when the population size is still small, the acquired information may prove invaluable for directing conservation effort. Finally, reintroduction projects provide unique opportunities to gather data on the vital rates of free-ranging species which usually remain inaccessible for demographic investigations. Ironically, species that have become extinct in the wild, but have subsequently been rehabilitated in nature, may well be better understood than thousands of surviving species for which knowledge of their population dynamics would greatly assist conservation management.

Acknowledgements We express our sincere thanks to the monitoring centres (Mercantour-, Ecrins, Vanoise-, Gran Paradiso-, Stelvio- and Hohe Tauern National Parks, Alpi Marittime Natural Park, ASTERS, Stiftung Pro Bartgeier, Réseau Gypaète Suisse Occidentale, Provincia Autonoma di Trento, Foundation for the Conservation of the Bearded Vulture) for allowing access to their data. We also thank hundreds of volunteers who regularly check bearded vultures across the Alps and report their observations. Marc Kéry, Lukas Jenni, Jean-Dominique Lebreton and two anonymous reviewers provided important comments on earlier drafts of the paper. Additional financial support was provided by the Stiftung Pro Bartgeier.

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Zink, R. (2005b) Monitoring an evaluation tool to determine reintroduction success. Bearded Vulture Reintroduction into the Alps Annual Report, 2005, 85–87. Received 23 May 2008; accepted 14 October 2008 Handling Editor: Des Thompson

Supporting Information Additional Supporting Information may be found in the online version of this article: Appendix S1 Details about the multistate-capture–recapture model

Appendix S2 Mathematical details about the bearded vulture population model Appendix S3  code for running the stochastic population model Table S1. Modelling result for resighting and recovery probabilities Table S2. Modelling result for survival probabilities Please note: Wiley-Blackwell are not responsible for the content or functionality of any supporting materials supplied by the authors. Any queries (other than missing material) should be directed to the corresponding author for the article.

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Journal of Applied Ecology 2004 41, 1103–1116

Ecological requirements of reintroduced species and the implications for release policy: the case of the bearded vulture Blackwell Publishing, Ltd.

ALEXANDRE H. HIRZEL*†, BERTRAND POSSE‡, PIERRE-ALAIN OGGIER‡, YVON CRETTENAND§, CHRISTIAN GLENZ¶ and RAPHAËL ARLETTAZ*‡** *Zoological Institute–Conservation Biology, University of Bern, Baltzerstrasse 6, CH-3012 Bern, Switzerland; †Laboratory of Conservation Biology, Department of Ecology and Evolution, University of Lausanne, CH-1015 Lausanne, Switzerland; ‡Bearded Vulture Network Western Switzerland, Nature Centre, CH-3970 Salgesch, Switzerland; §Game, Fishery and Wildlife Service, Canton of Valais, Rue de l’Industrie 14, CH-1950 Sion, Switzerland; ¶Laboratory of Ecosystem Management, Institute of Environmental Science and Technology, Swiss Federal Institute of Technology, CH-1015 Lausanne, Switzerland; and **Swiss Ornithological Institute, Valais Field Station, Nature Centre, CH-3970 Salgesch, Switzerland

Summary 1. Species undergoing reintroduction offer a unique opportunity for clarifying their specific niche requirements because they are likely, if sufficiently mobile, to colonize the most suitable habitats first. Information drawn from the individuals released first might thus be essential for optimizing species’ policy as reintroductions proceed. 2. Bearded vultures were extirpated from the European Alps about a century ago. An international reintroduction programme using birds reared in captivity was launched in 1986; up to 2003, 121 individuals had been released at four different locations. Subsequent dispersion throughout the range has been far from homogeneous, resulting in a clumped occurrence of the first breeding pairs within three main zones that do not necessarily coincide with release areas. 3. In order to discern ecological requirements we performed a geographical information system (GIS) analysis of bearded vulture sightings collected in Valais (Swiss Alps) from 1987 to 2001. This area harbours no release site, is situated in the core of the Alpine range and has been visited by birds from all four release points. 4. During the prospecting phase (1987–94, mostly immature birds), the most important variable explaining bearded vulture distribution was ibex biomass. During the settling phase (1995–2001), the presence of birds (mostly maturing subadults) correlated essentially with limestone substrates, while food abundance became secondary. 5. The selection of craggy limestone zones by maturing bearded vultures might reflect nesting sites that are well protected against adverse weather, as egg laying takes place in the winter. Limestone landscapes, in contrast to silicate substrates, also provide essential finely structured screes that are used for bone breaking and temporary food storage, particularly during chick rearing. Finally, limestone substrates provide the best thermal conditions for soaring. 6. Synthesis and applications. Extrapolated to the whole Alpine range, these findings might explain both the current distribution of the subadult/adult population and the absence of breeding records for bearded vultures around release sites in landscapes dominated by silicate substrates. As reintroduced bearded vultures tend to be philopatric, we suggest that population restoration would be more efficient if releases were concentrated within large limestone massifs. This case study of the bearded vulture illustrates the need for continual adaptive management in captive release programmes.

© 2004 British Ecological Society

Correspondence: Raphaël Arlettaz, Zoological Institute, Division of Conservation Biology, University of Bern, Baltzerstrasse 6, CH-3012 Bern, Switzerland (fax +41 31 631 45 35; e-mail [email protected]).

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Key-words: Alps, ecological niche factor analysis (ENFA), geographic information system (GIS), Gypaetus barbatus, habitat selection, niche modelling, reintroduction, species conservation Journal of Applied Ecology (2004) 41, 1103–1116

Introduction

© 2004 British Ecological Society, Journal of Applied Ecology, 41, 1103–1116

As part of modern strategies to counter the risks of biodiversity loss, reintroduction schemes are becoming more and more common world-wide (Sarrazin & Barbault 1996; Carroll et al. 2003). Usually relying upon individuals stemming from captive stocks (Wedekind 2002), they aim at reinstalling wild populations of extinct species, especially those that have been directly (overkilling) or indirectly (habitat loss, predator or competitor introduction) extirpated by humans (O’Toole, Fielding & Haworth 2002; Richards & Short 2003; Ripple & Beschta 2003; Schaub, Pradel & Lebreton 2004). Alternatively, they aim to reinforce relict populations of critically endangered species (Hodder & Bullock 1997; Wanless et al . 2002). However, reintroduction programmes will only be effective if the ecological requirements of a species or local population are well understood. Species extirpated before the onset of modern ecology are of particular concern because knowledge of their functional position within a local ecosystem (i.e. their ecological niche) is often poorly documented, with information often scarce or anecdotal (Breitenmoser 1998). On the other hand, expanding populations of released species provide an opportunity to unravel species’ ecological needs. This is particularly true if the species shows a high degree of dispersal, when it might be assumed that individuals will first settle in those areas that offer the most suitable conditions. The information drawn from first released individuals can thus serve to estimate species’ preferences, thereby guiding subsequent releases within an ongoing reintroduction programme. Reintroduced individuals therefore offer an opportunity to identify species’ requirements more easily than long-established populations. We illustrate this with a study of resource selection in a newly reintroduced population of bearded vultures Gypaetus barbatus Linnaeus 1758 in the Alps. The bearded vulture, or lammergeier, is a large scavenging raptor that feeds primarily on bones. It was extirpated from the European Alps between the end of the 19th century and the beginning of the 20th century, with the date of extinction varying among populations (Arlettaz 1996; Mingozzi & Estève 1997). A large-scale reintroduction programme, based on release to the wild of birds just prior to fledging that were born in captivity, was launched in 1986 in Austria (Frey 1992). This was followed by regular releases in three further Alpine countries: France from 1987 onwards, and Switzerland and Italy since 1991. Up to July 2003, a total of 121 individuals has been reintroduced into the

Alps (Frey 2002; see list in Robin et al. 2004). About 60–65% of the released birds are believed to have survived (Frey 2002; Zink 2002), although this figure may underestimate mortality because released birds are monitored only passively, principally through patterns of bleached feathers, a marking technique that enables individual recognition only until the first moult (i.e. until 2–3 years of age). Also, the limited number of marking combinations available, as well as the difficulty of reading markings accurately, particularly by inexperienced observers, has generated some confusion about sighting records (Zink 2002). Feather bleaching has nevertheless illustrated the dispersal potential of the species: individuals released as far away as central Austria (375 km), eastern Switzerland (250 km) and southern France (230 km) have subsequently been located in Valais in the Swiss Alps. The first successful breeding event of the released population occurred in 1997 in Haute-Savoie, France (Heuret & Rouillon 1998). It was followed by 13 further chicks that have fledged between 1998 to 2003, and there were six breeding attempts in 2002 (three in France and three in Italy). However, no successful breeding has occurred in Austria and Switzerland, in spite of numerous observations of mature individuals in these countries. Several birds released in the Swiss National Park (the only Swiss release site) settled in nearby Italy. The absence of established breeding pairs is particularly striking for Austria, as birds were reintroduced there from the very beginning of the programme. At least one pair bond was established close to the release site in Rauris, Austria, some years after the beginning of the reintroduction effort, and a total of seven successive pair bonds was formed over the years (Zink 2002). The reasons for this lack of success remain largely poorly understood. Although no releases occurred in the Alps of Valais (south-west Switzerland), they are close to the French release site in Haute-Savoie (30 km from the Swiss border). Bearded vultures were observed in the area soon after the first French releases in 1987 (Arlettaz 1996). A network of observers was formed in Valais in 1988, and the sightings checked meticulously for reliability before being recorded into a database. Although several individuals, including some adults, have attempted to settle in Valais, no pair has attempted to breed and the few pairs consist of subadult individuals. In the mid-1990s, we noticed that the distribution of bearded vultures in Valais was heterogeneous, with the species showing a strong preference for particular areas but avoiding other apparently suitable mountainous zones.

1105 Reintroducing the bearded vulture

During the last decade, with the growing availability of digitalized spatial data, geographic information systems (GIS) have received growing interest from conservation biologists. These tools allow better understanding of the species–habitat links and underpin many spatial predictions in wildlife management (Augustin, Mugglestone & Buckland 1996; Guisan & Zimmermann 2000; Jaberg & Guisan 2001; Cabeza et al. 2004; Johnson, Seip & Boyce 2004; Rushton, Ormerod & Kerby 2004). By performing a GIS-based habitat analysis using the Valais sighting data set, we sought to identify which key environmental factors explained the distribution of bearded vultures. An extrapolation of these findings to other zones within the Alps could enable an understanding of why maturing and adult bearded vultures occupy and breed in some areas, while others remain uninhabited. Recommendations can then be made for improving the ongoing international reintroduction programme.

Data and methods We analysed the relationships between two GIS data sets: the distribution of vulture sightings and a set of environmental descriptors. This allowed us to quantify and model bearded vulture’s ecological requirements.

 

© 2004 British Ecological Society, Journal of Applied Ecology, 41, 1103–1116

The study area was the canton of Valais in the southwestern Swiss Alps. It covers 5191 km2 (about 2·7% of the overall area of the Alpine massif, which is 191 000 km2 wide), modelled by a 100 × 100-m resolution raster map, i.e. 519 124 1-ha grid cells, overlaid on the hectometric Swiss coordinate system (plane projection). Four categories of environmental descriptors were included in the analysis: (i) topographical (continuous variables), comprising altitude, slope and exposition (or aspect); (ii) geological (presence–absence), comprising compact limestone, marl, gneiss, granite, rocky area, scree and water (rivers and lakes); (iii) anthropogenic (presence– absence), comprising buildings; and (iv) biological, comprising forest and meadows (presence–absence), ibex Capra ibex and chamois Rupicapra rupicapra (biomass), sheep (density) and distance to release site. The actual environmental variables were derived from information maps to provide quantitative (a requirement of ecological niche factor analysis, ENFA; Hirzel et al. 2002) and integrative information. In fact, vulture behaviour and resource selection are probably not influenced by the quality of a single hectare but rather by characteristics of a wider area (which we assumed was circular). We envisaged three possible scales: (i) a sight-field scale of 500-m radius; (ii) a flight-search scale of 2000-m radius; and (iii) a long-range exploration scale (unlimited). The two first scales provided occurrence–frequency maps computed by means of a circular moving window, while the third provided a map of the distance to the closest occurrence. Accord-

ingly, each presence–absence descriptor generated three variables. We used the module CircAn of Biomapper (Hirzel, Hausser & Perrin 2002) and the module Distance of Idrisi32 (Eastman 1999) to perform these GIS operations. The topographical descriptors were averaged by means of a 2000-m radius circular moving window and provided four variables: altitude, slope, northness (cosine of aspect) and eastness (sine of aspect); all were averaged on the moving window. We also computed the standard deviation of the altitude. Wild ungulate biomass maps were built as follows. For the chamois, data were from a detailed census conducted in 1997 –98 (Glenz et al. 2001) of the 24 cantonal gamekeeper districts and the federal and cantonal protected areas. For ibex, we had similar census data for every colony. In order to increase the spatial accuracy of the density estimations, we combined the census information to potential distribution maps of each species (Hausser 1995) to produce density maps (individuals ha−1). These densities were finally multiplied by average sex- and age-dependent body mass (male, female and young: chamois, 30, 24 and 16 kg, respectively; ibex, 84, 33 and 22 kg, respectively; Game, Fishery & Wildlife Service, Valais, 1997, unpublished data) to give the biomass per grid cell (= kg ha−1). Additionally, we computed a map of ibex observation density (kilometric resolution, database from the Centre Suisse de Cartographie de la Faune, 1999, unpublished data). We used our own data on sheep density (individuals ha−1) for each summering pasture, and their spatial coordinates (Cantonal Veterinary Service, Valais, 1997, unpublished data). As information about the spatial extent of these pastures was lacking, we assumed a circular 2000-m radius shape for all of them. Where two such pastures were overlapping, we retained only the largest density. The distributions of all environmental descriptors were, as far as possible, rendered more symmetrical by the Box–Cox standardizing algorithm (Sokal & Rohlf 1981).

   Throughout the Alps, ornithologists have been monitoring birds’ movements since the beginning of the release programme. In Valais, a network of observers (Bearded Vulture Network Western Switzerland), in collaboration with the Cantonal Game, Fishery and Wildlife Service has collected and checked 1398 sightings, which stem from at least 29 different individual birds, from 1987 to 31 December 2001. Most identified birds (19 out of 29, c. 65%) originated from the release site in Haute-Savoie. Data recorded included date of observation, geographical location and, if known, the identity of marked birds. Although the observation effort varied with the occurrence and effort of observers, we assumed that our data were representative of the actual geographical occupancy of the area by bearded vultures. Additionally, we controlled for any possible bias in observation

1106 A. H. Hirzel et al.

Fig. 1. Top: location of Valais in the European Alps and Switzerland. Bottom: hill-shade map of the study area (Valais, Switzerland) showing 1-km2 squares with bearded vulture observations from 1987 to 1994 (black squares) and from 1995 to 2001 (white circles). The geographical subdivision of the study area is depicted by numbers (see Fig. 4).

© 2004 British Ecological Society, Journal of Applied Ecology, 41, 1103–1116

clustering by subdividing the study area into major valley systems for which observation effort within a golden eagle Aquila chrysaetos monitoring programme was quantified (P. A. Oggier). Bearded vulture sightings were also recorded systematically in the same area by the same observer, and an index of frequency of observations per observation time unit and year (1990– 2001) could be estimated for each zone separately. Both species have converging soaring habits and frequently use similar routes. This enabled us to assess whether clusters of bearded vulture sightings were observation effort-dependent or reflected actual habitat preferences by the species. The colonization of Valais showed two distinct chronological phases: (i) a prospecting phase (1987– 94), when immature (i.e. mostly 1–3 years old) individuals mainly visited the south-western parts of Valais, i.e. the valleys south of the main Rhône valley axis, at the periphery of the release site in France; (ii) a settling phase (1995–2001), when mostly maturing birds (sub-

adults; ≥ 3 years old) attempted to settle down in the north-west of Valais. In order to investigate this change of behaviour and its possible link to new patterns of habitat selection, the observation data were divided into two subsets (Fig. 1). The second data set included a marginal number of observations of immature birds, which rendered our spatial analysis conservative because an even greater contrast would have been found if those immatures could have been removed from this secondphase analysis. In the survey, vulture sightings were recorded at a 1km resolution. We built two presence 100-m resolution maps, hereafter observation maps, by assigning each record to the central hectare of the kilometric square where the bird had been seen.

  A variety of methods have been developed to model species’ habitat and potential distribution (Guisan &

1107 Reintroducing the bearded vulture

Zimmermann 2000; Rushton, Ormerod & Kerby 2004). The majority are based on presence–absence species’ data sets. They make the intuitive assumption that the presence of a species is an indicator of suitable habitat and its absence an indicator of unsuitable habitat. However, there are many cases where these assumptions are not correct. In many cases, absence data are either unavailable (e.g. museum data, herbarium, atlas data) or unreliable (e.g. cryptic species, metapopulation following extinction–recolonization dynamics, invading species) (Hirzel, Helfer & Métral 2001; Peterson 2001; Hirzel et al. 2002; Peterson et al. 2002; Engler, Guisan & Rechsteiner 2004). In the case of the bearded vulture, absences were unreliable for two main reasons. (i) Being philopatric, this raptor is slowly spreading from its release site, therefore lack of sighting in some locations might be caused either by unsuitable conditions (true absence) or by the site being too far and yet unreached (false absence). (ii) This bird explores a wide area every day, making any systematic sampling of absences difficult. The first reason is particularly problematic as the case of a spreading species has been shown to fool a presence–absence-based method (generalized linear model; Hirzel, Helfer & Métral 2001). Therefore, we had to use a presence-only approach and we selected the ENFA (Hirzel et al. 2002) as it has been shown to be robust to spreading-species bias (Hirzel Helfer & Métral 2001); it has been applied to several studies based on presence-only data (Zaniewski, Lehmann & Overton 2002; Dettki, Lofstrand & Edenius 2003; Reutter et al. 2003; Thomas 2003; Brotons et al. 2004; Engler, Guisan & Rechsteiner 2004). A further advantage of the ENFA is that there is no descriptor selection, a sensitive process when stepwise procedures are involved. Instead, the ENFA computes a weight for all descriptors indicating their importance for the species’ niche and their correlations.

    

© 2004 British Ecological Society, Journal of Applied Ecology, 41, 1103–1116

The ecological niche of a species is potentially shaped by a large number of variables, with various levels of importance. Moreover, most of these variables exhibit some degree of correlation. ENFA extracts all relevant information from these variables while discarding their correlations and the background noise. It does so in a similar way to principal components analysis (PCA) by computing new, uncorrelated factors, a few of them summarizing most of the information. The main difference between PCA and ENFA is the nature of the data sets (here a data set is a population of vectors, the components of which are the values of the descriptors recorded at a grid cell). The PCA is computed on a single data set and its factors (or components) seek to find the directions that maximize the descriptor variances in the multidimensional environmental space. In contrast, the ENFA is based on two data sets: (i) the global set stores the descriptor values for all cells in the study area, and (ii) the species set stores these values for only

Fig. 2. Geometrical interpretation of marginality and specialization factors. The two-dimensional distribution of the global and species sets are symbolized by white and dotted ellipses, with a crossed-circle marking their centroids. The marginality factor (M) is the axis passing through both centroids. Once marginality has been extracted, both distributions have a common centroid and the specialization factor (S) is the axis maximizing the ratio of global variance σG to species variance σS; it is intermediary between the axes of maximal global variance (dotted line) and minimum species variance (dashed line). See text for further details.

those cells where the species is present; it is therefore a subset of the global set. The ENFA factors result from the comparison between these two sets, and they are therefore directly interpretable. The first ENFA factor maximizes the absolute value of the marginality, defined as the standardized difference between the species mean and the global mean on all descriptors. It is geometrically figured by the line passing through the centroids of both the species and global sets (Fig. 2a). The marginality coefficients range from −1 to +1. Positive or negative values indicate a species’ optimum higher (respectively lower) than the average conditions in the study area. Once the marginality factor has been extracted, the global and species sets centroids are coinciding. All the subsequent factors maximize the specialization, defined as the ratio of the global variance to the species variance. A specialization factor is geometrically figured by a line intermediary between the direction of maximum global breadth and the direction of minimum species breadth (Fig. 2b). Specialization coefficients range from −1 to +1, but only their absolute value is meaningful. A high value indicates a narrow niche breadth in comparison with the available conditions. Finally, there are as many factors as descriptors. The first one explains all the marginality and some part of the specialization. The subsequent factors explain the remaining specialization in decreasing amounts.

1108 A. H. Hirzel et al.

Table 1. Environmental descriptors retained for the habitat analysis. Except when stated otherwise, they were derived from the GEOSTAT database (Swiss Federal Office of Statistics, Neuchâtel, Switzerland). EGV = ecogeographical variables Variable category

Environmental descriptor

EGV code

Topographical

Average altitude in a 2000-m radius Average slope in a 2000-m radius SD of altitude in a 2000-m radius Average northness in a 2000-m radius* Average eastness in a 2000-m radius†

ELEV SLOPE SDELEV NORTH EAST

Geological

Frequency of limestone area in a 2000-m radius Distance to limestone area Rock frequency in a 2000-m radius Rock frequency in a 500-m radius Distance to granite area Scree frequency in a 2000-m radius Scree frequency in a 500-m radius Distance to screes Water frequency in a 2000-m radius Water frequency in a 500-m radius

CALC-2K CALC-D ROCK-2K ROCK-500 GRANIT-D SCREE-2K SCREE-500 SCREE-D WATER-2K WATER-500

Anthropogenic Biological

Building frequency in a 2000-m radius Forest frequency in a 2000-m radius Forest frequency in a 500-m radius Ibex biomass index‡ Ibex frequency in a 2000-m radius‡ Chamois biomass index‡ Sheep density in a 2000-m radius‡ Distance to release site§

BUILD-2K FOREST-2K FOREST-500 IBEX-BM IBEX-2K CHAM-BM SHEEP-2K RELEASE-D

*Cosine of aspect. †Sine of aspect. ‡Derived from Swiss Federal Game statistics, Bern, Switzerland. §Computed in the GIS.

© 2004 British Ecological Society, Journal of Applied Ecology, 41, 1103–1116

Usually, the most significant part of the information is gathered in a few of the first factors, thus reducing the problem complexity. The factors are given by their coefficients along the environmental descriptors and provide information about the species’ marginality and specialization on each of them. Moreover, the global marginality and specialization coefficients integrate all these descriptor-specific scores, providing general clues about the species’ niche. The global marginality ranges from 0 to 1 and indicates how far, all descriptors being accounted for, the species optimum is from the average conditions in the study area. The global specialization ranges from 1 to infinity; for ease of interpretation, the global tolerance coefficient, defined as the inverse of the specialization, is usually preferred as it ranges from 0 to 1. It is an indicator of the species’ niche breadth. It must be noted though that these coefficients are relative to the study area and can only be used to compare species modelled with the same set of predictors. A detailed mathematical demonstration of the ENFA is beyond the scope of this paper and we refer the interested reader to our basic description (Hirzel et al. 2002). The ENFA analysis was first applied to all the available environmental descriptors of the full set of observations (immatures and subadults pooled together); this was done in order to select the variables relevant for the bearded vulture distribution. Then, the ENFA was applied with the reduced descriptor set (listed in

Table 1) to both observation subsets. This provided two ecological niche models, one for the prospecting phase (1987–94) and one for the settling phase (1995– 2001). All these analyses were made within the eco-GIS package Biomapper 2·1 (Hirzel et al. 2002).

   The two models were then used to compute a habitat suitability map by means of the geometric mean algorithm (Hirzel & Arlettaz 2003). This algorithm works in the multidimensional environmental space defined by the most significant ecological niche factors computed by the ENFA. The species set defines a cloud of points in this environmental space, the density of which varies greatly and is assumed to be positively correlated with the suitability of any particular combination of descriptor values. This density is modelled for every hypervolume element (voxel) of this space by the geometric mean of its distances to all observation points; the higher the density of observations around a given voxel, the lower the mean distance. These distances are transformed into habitat suitability indices by delineating hypersurfaces (or envelopes) linking all voxels that have the same value (like the altitude isolines of a topographic map). An envelope suitability index is computed as the proportion of observation points outside it; for instance, the envelope 0·9 encloses 10% of the observations and leaves out 90% of them. These envelopes

1109 Reintroducing the bearded vulture

are then transported to the classical geographical space to produce a habitat-suitability map. We chose to keep only the envelopes 0·5 and 0·9. The inner envelope (enclosing 50% of the observations) geographically defines a region we called core habitat. The geographical area located between the two envelopes (40% of observations) was termed marginal habitat. The area outside the external envelope (10% of observations) was considered unsuitable. See Hirzel & Arlettaz (2003) for further information about this algorithm.

Results The first ENFA including all environmental variables and all the observations showed that some variables were not relevant to the bearded vulture distribution: all gneiss- and marl-related variables, 500- and 2000-m radius granite frequency, distance to water, distance to rock, 500-m radius frequency of human buildings and distance to them, all meadow-related variables, and distance to forest. The retained variables are listed in Table 1 and were used for all subsequent analyses.

  Observation points not used to calibrate the model were held on a validation set. Two indices could then be computed: (i) the absolute validation index (AVI), which is the proportion of validation points occurring in the predicted core habitat; and (ii) the contrast validation index (CVI), which is the AVI minus the AVI that would have been obtained with a hypothetical model that would predict core habitat for all cells of the study area. The latter index gives an indication of how well the model discriminates poor from good habitat. Both AVI and CVI were submitted to a cross-validation process (Sokal & Rohlf 1981; Manly 1991; Fielding & Bell 1997), allowing the computation of confidence intervals: the observation data set was partitioned into 100 subsets of which, alternately, 99 were used to calibrate the model (calibration set) and 1 to validate it (validation set).

 

(1987–94)

For this period, 310 observations were analysed. The ENFA computed a global marginality coefficient of 0·72 and a global tolerance coefficient of 0·66, indicating that the vulture was living in conditions rather uncommon in the study area but that its niche breadth was rather wide. By comparing the ENFA eigenvalues wit the MacArthur’s broken-stick distribution (MacArthur 1960; Hirzel et al. 2002), the first five factors were kept as significant for the subsequent analyses, explaining 70% of the information (100% of the marginality and 41% of the specialization). The marginality factor explained little of the specialization (6%), meaning that the vulture niche breadth was not particularly narrow for the variables for which its optimum was the furthest from the average conditions. A slightly negative marginality coefficient (Table 2, factor 1) for altitude indicated that, on average, the

Table 2. Correlation between ENFA factors and the environmental descriptors for the prospecting phase (1987–94). The percentages indicate the amount of specialization accounted for by the factor (moreover, factor 1 explains 100% of the marginality)

ELEV SLOPE SDELEV NORTH EAST CALC-2K CALC-D ROCK-2K ROCK-500 GRANIT-D SCREE-2K SCREE-500 SCREE-D WATER-2K WATER-500 BUILD-2K FOREST-2K FOREST-500 IBEX-BM IBEX-2K CHAM-BM SHEEP-2K RELEASE-D © 2004 British Ecological Society, Journal of Applied Ecology, 41, 1103–1116

Factor 1† (6%)

Factor 2‡ (18%)

Factor 3‡ (13%)

Factor 4‡ (8%)

Factor 5‡ (7%)

– +++ +++ –– – ++ – + + 0 ++ + –– ++ + – + 0 ++ +++++ + ++ –––––

********* * * * * * 0 *** * 0 * 0 0 * 0 *** 0 * 0 0 0 0 *

0 ** * * 0 0 0 ** ** 0 * ** 0 * 0 ********* * ** 0 0 0 0 *

* * 0 ** 0 **** **** **** ** **** ** ** 0 ** 0 ** 0 * 0 0 * * *

*** 0 * * * 0 0 **** ** 0 ** **** ***** * * *** **** 0 0 * 0 0 0

†Marginality factor. The symbol + means that the vulture was found in locations with higher values than average. The symbol – means the reverse. The greater the number of symbols, the higher the correlation. 0 indicates a very weak correlation. ‡Specialization factor. The symbol * means the vulture was found occupying a narrower range of values than available. The greater the number of asterix, the narrower the range. 0 indicates a very low specialization.

1110 A. H. Hirzel et al.

Fig. 3. Habitat suitability map computed for the (a) prospecting (1987–94) and (b) settling (1995–2001) phases showing the spatial distribution of the core (black), marginal (dark grey) and unsuitable habitats (white).

© 2004 British Ecological Society, Journal of Applied Ecology, 41, 1103–1116

bearded vulture was found at lower altitude (2067 m) than the Valais average (2157 m). Furthermore, the high value of the first specialization factor (Table 2, factor 1) for this predictor indicated a narrow niche breadth, meaning that birds were rarely seen flying far from this altitude (SD = 503 m). Similar reasoning on the other coefficients showed that the favoured areas had steeper slopes than average (31° vs. 28°, respectively); their average northness (−0·11) and eastness (−0·02) and their relatively high marginality indicated a preference for slopes orientated towards the south or south-west. Nevertheless, the bearded vulture showed a very low level of specialization on these three variables. Other outstanding landscaperelated features were specialization for rocky areas (mainly limestone and screes), relatively high frequency of water, avoidance of human settlements and some

specialization for areas with a slightly higher forest frequency than average (at the 2000-m radius scale). The highest marginality was related to ibex and sheep presence and proximity to the release site; however, the vultures were again very tolerant regarding these variables (all five specialization coefficients were null or very low). Considering the sensitivity to different scale patterns, the vulture was almost always more marginal at the 2000-m than 500-m radius scale, or at distances greater than 2000 m; the specialization showed the same tendency. The cross-validation gave a mean AVI of 0·49 (SD = 0·13) and a mean CVI of 0·34 (SD = 0·13) (both values cannot be greater than 0·5). This means that, while the presence prediction power was very good, it could be due to a general overestimation of the habitat suitability (Fig. 3).

1111 Reintroducing the bearded vulture

Table 3. Correlation between ENFA factors and the environmental descriptors for the settling phase (1995–2001). The percentages indicate the amount of specialization accounted for by the factor (moreover, factor 1 explains 100% of the marginality)

ELEV SLOPE SDELEV NORTH EAST CALC-2K CALC-D ROCK-2K ROCK-500 GRANIT-D SCREE-2K SCREE-500 SCREE-D WATER-2K WATER-500 BUILD-2K FOREST-2K FOREST-500 IBEX-BM IBEX-2K CHAM-BM SHEEP-2K RELEASE-D

Factor 1† (10%)

Factor 2‡ (16%)

Factor 3‡ (9%)

Factor 4‡ (9%)

Factor 5‡ (7%)

Factor 6‡ (6%)

–– + ++ ––– ++ ++++++ ––––– + + + + 0 0 0 0 + + 0 0 ++ + ++ –––

****** ** *** *** * * 0 ***** * * * ** ** * 0 0 ** * ** 0 0 * 0

******** 0 * * ** * ** **** 0 0 * 0 0 0 0 ** *** 0 * * 0 0 **

** *** * * 0 * *** ***** 0 * * * * ** 0 0 ***** * * * 0 0 **

****** **** *** * ** * * *** * * * * 0 * 0 *** * * * 0 0 * **

*** * * * 0 ** **** * * 0 * ** ****** * * 0 ** * ***** * * 0 **

†Marginality factor. The symbol + means that the vulture was found in locations with higher value than average. The symbol – means the reverse. The greater the number of symbols, the higher the correlation. 0 indicates a very weak correlation. ‡Specialization factor. The symbol * means the vulture was found occupying a narrower range of values than available. The greater the number of asterix, the narrower the range. 0 indicates a very low specialization.

 

© 2004 British Ecological Society, Journal of Applied Ecology, 41, 1103–1116

(1995–2001)

More observations (1088) were available for this phase, but the global tolerance coefficient remained almost identical (0·65 vs. 0·66), whereas the global marginality coefficient was larger (0·84 vs. 0·72). By comparison of the eigenvalues with the MacArthur’s broken-stick distribution (MacArthur 1960; Hirzel et al. 2002), the first six factors were significant and were used in the subsequent analyses, explaining 79% of the information (100% of the marginality and 58% of the specialization). The marginality factor explained slightly more of the specialization (10% vs. 6%). In this phase, the situation was far more contrasted, with a few variables accounting for most of the marginality and specialization (Table 3). The most striking feature was the high marginality related to limestone areas: the bearded vulture tended to be seen in limestone environments (an average of 28% of limestone area in a 2000-m radius circle around observation points) or close to them (mean distance 468 m). There was some evidence of specialization on this variable, indicating a narrow niche breadth. The distance to release site was less marginal than for immatures, indicating that the mature birds had spread further. The average altitude was slightly lower than for immatures (mean = 1864 m, SD = 638 m), whereas the marginality for forest frequency was similar. The preference for southwards slopes was stronger among settling adults but with a tendency

towards south-eastern slopes. Ibex- and sheep-related variables lost their outstanding marginality, but vulture distribution was still biased towards them. The cross-validation gave a mean AVI of 0·5 (SD = 0·23) and a mean CVI of 0·45 (SD = 0·23). The contrast value was greater than in immatures, confirming the fact that this map is obviously more accurate (Fig. 3).

   .   The frequency of bearded vulture sightings was not dependent on local observation effort (Fig. 4). In geographical subunits 1 and 2 (compare Fig. 4 with Fig. 1), the observation ‘reward’ was definitely biased towards bearded vultures, actually confirming a more dense presence of the raptor in the north-western Valais Alps.

Discussion  The ecological requirements of reintroduced bearded vultures colonizing Valais differed markedly between the prospecting (1987–94) and settling (1995–2001) phases. Bearded vultures were globally more selective during the settling phase than during the prospecting phase. Habitat suitability maps also had a better predictive