MARIE-JULIE FAVE
REINTRODUCTION DU CISCO DE FUMAGE (COREGONUS HOYI) DANS LE LAC ONTARIO : DIVERSITÉ GÉNÉTIQUE ET CONSANGUINITÉ
Mémoire présenté à la Faculté des études supérieures de l'Université Laval dans le cadre du programme de maîtrise en biologie pour l'obtention du grade de maître es sciences (M.Se.)
Département de biologie FACULTÉ DES SCIENCES ET GÉNIE UNIVERSITÉ LAVAL QUÉBEC
Octobre 2006
© Marie-Julie Favé, 2006 RESUME
La gestion active des populations est souvent désirée et doit être réalisée de façon à minimiser les risques génétiques. Afin de déterminer la meilleure source de cisco de fumage (Coregonus hoyi) pour une réintroduction dans le lac Ontario, le polymorphisme de 10 locus microsatellites a été analysé pour des échantillons de C. hoyi des lacs Huron, Michigan, Supérieur et Nipigon ainsi que pour des échantillons de C. artedi et de ciscos des zones profondes du lac Ontario. Les populations de C. hoyi sont génétiquement diversifiées malgré des baisses d'abondance connues et sont différentiées entre les lacs. Aussi, nos résultats suggèrent que les individus du lac Ontario sont plus étroitement liés aux individus des lacs Huron et Michigan qu'à ceux des lacs Supérieur et Nipigon. Par la suite, des simulations montrent qu'un grand nombre de géniteurs, un rapport des sexes équilibré, une haute proportion de croisements efficaces, un pool de géniteurs diversifié ainsi qu'un grand nombre d'individus introduit minimiserait la consanguinité dans la nouvelle population.
ABSTRACT
Active population management is often desired and should be designed so as to minimize genetic risks. In order to détermine the best source of bloater {Coregonus hoyi) for a reintroduction in Lake Ontario, we analyzed genetic polymorphism at 10 microsatellite loci in samples of C. hoyi from Lakes Huron, Michigan, Superior and Nipigon as well as samples of C. artedi and deepwater ciscoes from Lake Ontario. C. hoyi populations are genetically diversified despite known démographie déclines and they are significantly differentiated among lakes. Also, our results suggest that Lake Ontario ciscoes are more closely related to ciscoes from Lake Huron and Michigan than to ciscoes from Lake Superior or Nipigon. Computer simulations demonstrate that a high number of breeders, a balanced sex ratio, a high proportion of effective crosses, a genetically diverse pool of breeders and a high number of individuals introduced each year would minimize inbreeding in the reintroduced population.
H AVANT-PROPOS
Une multitude de gens et d'organismes ont contribué, chacun à leur façon, à l'accomplissement de ce projet :
Je tiens avant tout à remercier ma directrice Julie Turgeon pour m'avoir accueillie dans son laboratoire après une simple petite rencontre, pour m'avoir accordé sa confiance et guidée dans mes nombreux apprentissages, de la pipette à la rédaction finale. Merci pour toutes les discussions que nous avons eues, scientifiques comme humanistes, sérieuses comme ludiques, critiques comme naïves. Je dois également un gros merci à Pierre Duchesne pour avoir programmé les équations qui ont servi lors des simulations exécutées dans le Chapitre 2, sans qui la réalisation de cette deuxième partie aurait été bien plus laborieuse, voire impossible. Je le remercie pour ses recommandations toujours éclairées, précises et efficaces. Merci aux nombreuses personnes qui ont collaboré à l'échantillonnage des ciscos dans les Grands Lacs, soit Rick Salmon (OMNR), Lloyd Mohr (OMNR), Bruce Morrison (OMNR), Owen Gorman (USGS) et Kim Scribner (Michigan State University). Je remercie la Great Lakes Fishery Commission pour l'initiation du projet ainsi que pour son support financier nécessaire à la réalisation de celui-ci. Je remercie également pour leur soutien financier, le FGÉS-CRSNG et le département de biologie.
Merci aux membres du labo JT pour les discussions quotidiennes qui font avancer les choses et qui refont notre petit (ou grand) monde, pour les conseils judicieux ainsi que pour les bons moments passés ensemble: Julien April, Olivier Rey, Cécilia Hernandez, Marie- Claude Gagnon, Ariane Dubé-Linteau, Mélissa Tremblay, Julie Jeukens, Emilie Bilodeau, Milouda Achaboune, Sébastien Bélanger et Kevin Foulché. Pour m'avoir changé les idées aux moments propices, merci à mes amis cépassiens, avec qui les plus fous délires libérateurs sont permis et aux fidèles grimpeurs matinaux, toujours prêts à se dépasser afin d'assouvir notre dépendance aux endorphines, à l'adrénaline et aux grands espaces...
Et bien sûr, un merci tout particulier à François et Martine, pour leur continuel support, pour m'avoir appris à poursuivre mes idées et pour toujours m'encourager à accomplir ce qui me rend heureuse.
ni Ce mémoire comporte une introduction et une conclusion générales rédigées en français. Les chapitres I et II sont rédigés en anglais car ils sont destinés à être publiés dans des revues spécialisées.
IV TABLE DES MATIERES
RÉSUMÉ II ABSTRACT II AVANT-PROPOS III TABLE DES MATIERES V LISTE DES TABLEAUX ET FIGURES VII INTRODUCTION GÉNÉRALE 9
RÉINTRODUCTIONS EN NATURE 9 LA FAUNE ICHTHYENNE DES GRANDS LACS LAURENTIENS 12 RÉINTRODUCTION DE C. HOYI DANS LE LAC ONTARIO 13 OBJECTIFS DE RECHERCHE 14 CHAPITRE 1: 16 PATTERNS OF GENETIC DIVERSITY IN GREAT LAKES BLOATERS (COREGONUS HOYI) FOR FUTURE REINTRODUCTION IN LAKE ONTARIO ... 16 RÉSUMÉ 17 ABSTRACT 17 INTRODUCTION 19 Laurentian Great Lakes Fish Fauna 20 Reintroduction ofC. hoyi inLake Ontario 21 MATERIALS AND METHODS 22 Biological material 22 DNA extraction and microsatellite analysis 23 Genetic diversity and historical demography 23 Genetic relationships among Great Lakes bloaters andLake Ontario ciscoes 24 RESULTS 25 Genetic diversity and historical demography 25 Relationships among populations and individual assignment tests 26 DISCUSSION 28 Genetic diversity and bottlenecks 28 Differentiation within and among lakes 31 CONCLUSIONS 32 ACKNOWLEDGMENTS 33 FIGURE LEGEND 34 ANNEXE A : NOMBRE D'INDIVIDUS (N), NOMBRE D'ALLÈLES (A), HÉTÉROZYGOTIE
OBSERVÉE (HO) ET ATTENDUE (HE) PAR LOCUS ET MULTILOCUS POUR CHAQUE ÉCHANTILLON 44 CHAPITRE 2: 48 INBREEDING DYNAMICS IN REINTRODUCED, AGE-STRUCTURED POPULATIONS OF HIGHLY FECUND SPECIES 48 RÉSUMÉ 49 ABSTRACT 49 INTRODUCTION 51 MATERIALS AND METHODS 52 Breeding design and inbreeding coefficient. 52 Dynamics of inbreeding coefficient in a reintroduced population 53 RESULTS 55 Effects of breeding design on FH 55 Dynamics ofFwin the reintroduced population 56 DISCUSSION 57 ACKNOWLEDGMENTS 60 FIGURE LEGEND 61 ANNEXE B: DERIVATION OF RECURRENCE EQUATIONS TO COMPUTE F VALUES 64 Discrète générations 64 Three âge classes 65 Nine âge classes 66 CONCLUSION GÉNÉRALE 69 Résumé de la recherche 69 Limites et perspectives .71 BIBLIOGRAPHIE 73
VI LISTE DES TABLEAUX ET FIGURES
Liste des tableaux
Chapitre 1
Table 1: Information on ciscoe samples, along with détails on genetic diversity (NA: number of alleles; NAR: allelic richness; Ho and He: observed and expected heterozygosity, respectively), and results of démographie tests, (P-values for heterozygosity déficit and excess (Cornuet and Luikart 1996); M ratio of Garza and Williamson (2001)) 35
Table 2: Microsatellite loci used for the characterization of Great Lakes ciscoes: primer séquences with fluorescent label, PCR conditions (primer concentration, TA, and number of cycles), PCR multiplex and migration set, allele size range (bp), and total number of alleles observed 37
Table 3: Pairwise FST values between C. hoyi samples from Potential Donor Lakes (HUR: Lake Huron; MCH: Lake Michigan; SUP: Lake Superior; NIP: Lake Nipigon) and ciscoes from Lake Ontario (ONT-ARTEDI, ONT-DEEP). FST values significantly différent from 0 (Bonferroni corrected) are in bold 38
Table 4: Average probability (LnP(D)) and standard déviation across 20 runs of STRUCTURE for K genetic partitions among bloaters from ail Potential Donor Lakes 39
Table 5: Assignment of individual ciscoes from Lake Ontario a) to three genetic partitions defined by STRUCTURE - see Results-, with mean coefficient of membership (q) in parenthèses and b) to Potential Donor Lakes with GENECLASS, with average first score in parenthèses (HUR: Lake Huron; MCH: Lake Michigan; SUP: Lake Superior; NIP: Lake Nipigon)) 40
Liste des figures
Chapitre 1
Figure 1: Sampling locations for C. hoyi in Potential Donor Lakes (•) and for ciscoes in Lake Ontario (•: contemporary and historical scale samples of C. artedi; "*": deepwater ciscoes caught in 2002) 41
Figure 2: Temporal variation in allelic richness (NAR) and demographical index (M ratio) in historical ciscoes from Lake Ontario (Hay Bay, Canada) in relation with relative commercial ciscoes catch in the Canadian waters of Lake Ontario (grey bars; Baldwin et al. 2005) 42
Figure 3: Unrooted DcE-based neighbor-joining phenogram relating ail samples from Potential Donor Lakes and Lake Ontario ciscoes (ONT-ARTEDI: récent sample of C. artedi; ONT-DEEP: deepwater ciscoes caught in 2002). Bootstrap values above 60 are shown 43
vu Chapitre 2
Figure 1: Inbreeding coefficient (FH) generated by crosses with variable numbers of breeders (NB), M:F sex ratios (• ~ 1:1; • ~ 1:4; * ~ 1:20), and pools of breeders characterized by high (a, b, and c; F = 0.24) or weak genetic diversity (d; F = 0.41) as a function of the proportion of effective crosses (PEC). Dashed Unes represent the F value for an infinité Wright-Fisher population of similar diversity. Bars represent variance among 200 simulations. Missing data points represents situations with unrealizable calculations because of excessively stringent conditions (a, PEC = 10; d, PEC = 10 and 20) 62
Figure 2: Simulated demography and inbreeding in a new population of bloaters created by supplementation during 15 years. a) Age-3 bloater population size b) breeders inbreeding coefficient (Fw) compared to that of introduced individuals (FH = 0.24; horizontal dashed Une). In both graphs, the number of individuals supplemented each year is indicated by différent Une types 63
Figure B.l: Supplementation System with three âge classes J, Wl, and W2 68
vin INTRODUCTION GÉNÉRALE
L'homme a un impact global incontestable sur plusieurs fonctions des écosystèmes qui va jusqu'à exercer un contrôle sur ceux-ci. Un grand nombre d'espèces ont été surexploitées tout autour du monde, principalement au cours des derniers siècles où l'accroissement de la population humaine fut substantiel. En réponse à cet impact grandissant, de nombreuses espèces ont vu leurs populations décliner, diminuant parfois jusqu'à l'extinction locale ou globale. Les priorités de conservation devraient donc porter sur des politiques de gestion qui renforceraient la protection contre la surexploitation et toutes les conséquences qui y sont associées. Malheureusement, dans certaines circonstances, la protection complète n'est plus possible à cause de la disparition de l'espèce dans une partie de son aire de répartition historique. Une telle situation ne laisse souvent que la création volontaire d'une nouvelle population pour restaurer la diversité écologique originale et son potentiel évolutif futur.
Une population ayant une diversité génétique élevée augmente ses chances d'adaptation et de survie à long terme dans un environnement où les conditions biologiques sont changeantes (Keller et al. 1994, Hughes et Stachowicz 2004). La diversité génétique nécessiterait d'être maximisée dans les programmes de réintroduction d'espèces en milieu naturel afin d'en augmenter le succès par la persistance de l'espèce sur une longue période de temps (Hedrick et Kalinowski 2000, Hughes et Stachowicz 2004, Tallmon et al. 2004). Une telle pratique doit donc être réalisée avec attention et considération de la biologie de l'espèce, de sa démographie et de ses structure et diversité génétiques (Bowen 1999, Moritz 2002).
Réintroductions en nature Les pratiques de conservation consistent souvent à prélever des adultes sauvages matures, à effectuer une ou plusieurs reproductions en captivité pour ensuite relâcher les juvéniles le plus rapidement possible dans la région où l'espèce est disparue localement (réintroduction) ou est en sévère déclin (supplémentation). Bien qu'une espèce puisse atteindre de hautes abondances en captivité, entraînées par une meilleure fertilité et une survie plus élevée, l'objectif ultime pour toutes les espèces devrait consister en la réintroduction en milieu naturel (Frankham et al. 2002). Plusieurs espèces dont les populations ont été historiquement réduites au seuil de l'extinction ont été récemment l'objet de programmes de réintroduction et sont en voie de rétablissement en milieu naturel, telles que l'oryx d'Arabie (Marshall and Spalton 2000), le sabre d'Argent du Mauna Kea (Friar et al. 2000), la crécerelle de l'île Maurice (Groombridge et al. 2000) et la panthère de Floride (Pimm et al. 2006). Par contre, d'autres espèces tel le touladi (Salvelinus namaycush) dans les Grands Lacs, peinent encore à retrouver des niveaux d'abondance appréciables et leur persistance dépend encore de l'introduction volontaire de spécimens élevés en captivité (Page et al. 2004).
L'établissement d'une population à partir d'un nombre limité d'individus entraîne le risque de causer un effet fondateur et de générer des niveaux élevés de consanguinité dès la création de la nouvelle population. Un tel effet est observé lorsque la diversité génétique des descendants se trouve diminuée à cause du nombre réduit de parents fondateurs et qu'une ascendance commune s'installe entre les rejetons. Il en sera amplifié si ces individus fondateurs sont apparentés, ou s'ils ont été échantillonnés dans une population ayant une faible diversité génétique. Aussi longtemps que la population reste petite et isolée, la diversité génétique est érodée par la fixation aléatoire des allèles et des mutations délétères tendent à s'exprimer à cause de l'efficacité réduite de la sélection naturelle. (Lande 1995, Lynch et al. 1995, Frankham et al. 2002). De ce fait, les individus consanguins démontrent souvent une survie réduite par rapport aux individus diversifiés lorsqu'ils sont soumis à une menace environnementale (Keller et al. 1994, Saccheri et al. 1998, Sherwin et al. 2000). Même lors de circonstances où la mortalité peut être attribuée à des causes environnementales, ce sont les facteurs génétiques qui, en définitive, déterminent quels individus survivront ou non. La consanguinité peut être un phénomène difficilement identifiable sans équivoque mais doit prendre une place centrale dans un programme de gestion de population et de réintroduction (Frankham et al. 2002).
Afin d'augmenter la probabilité de succès d'un programme de réintroduction, on cherche à maximiser la diversité génétique tandis qu'on cherche à minimiser la consanguinité. Ce but sera atteint en considérant deux aspects simultanément. Premièrement, les géniteurs fondateurs doivent être prélevés autant que possible dans une population génétiquement diversifiée, où aucune baisse d'abondance drastique récente
10 n'est suspectée. Ce point de départ assure déjà un pool génétique relativement diversifié et représentatif de l'espèce. Par exemple, des programmes de réintroduction du touladi {Salvelinus namaycush) dans les Grands Lacs ont été conçus de manière à respecter la diversité et divergence génétique observée entre les différents morphotypes présents (Page et al. 2004, 2005) afin de rétablir la diversité originale. Deuxièmement, la stratégie de reproduction en captivité doit être adroitement planifiée afin de minimiser l'ascendance commune des rejetons en fonction des contraintes biologiques et logistiques induites par le programme d'élevage en captivité. Un plan de croisement factoriel, où la production d'œufs totale de chaque femelle est divisée de manière à ce que chaque partie soit fécondée par un mâle différent, maximise à la fois la production de rejetons et la taille efficace de la population générée (Fiumera et al. 2004).
Le deuxième risque rencontré est celui d'introduire des individus trop différenciés génétiquement dans une population résiduelle et alors y causer une supplantation génique si ces nouveaux migrants ont une valeur adaptative plus élevée ou que leur nombre excède de beaucoup celui des résidents (Frankham et al. 1995, Rhymer et Simberloff 1996). Un cas classique de mélange génétique est celui des canards malards (Anas platyrhynchos) qui se sont hybrides intensément avec des populations d'espèces indigènes en Australie, en Nouvelle-Zélande et à Hawaii. Ces hybridations ont contribué au déclin de espèces endémiques locales et à la perte du phénotype original (Mank et al. 2004). Le croisement d'individus localement adaptés avec des individus génétiquement divergents amène également la possibilité de créer un phénomène de dépression d'outbreeding, où les descendants hybrides se trouvent défavorisés à cause du démantèlement de complexes de gènes co-adaptés et de l'expression d'allèles néfastes récessifs. Un tel phénomène fut identifié par des taux de survie réduits chez les hybrides FI et F2 issus de croisements entre deux populations génétiquement différentiées de saumon rosé (Onchorhynchus gorbuscha) (Gilk et al. 2004). Par ailleurs, la vigueur d'individus hybrides est identifiée dans divers systèmes biologiques (Sehaussen 2004), où l'introduction de migrants dans des populations consanguines peut apporter du nouveau matériel génétique bénéfique (Ebert et al. 2002, Whitlock et al. 2000). Aussi, une dépression é'outbreeding peut être considérée comme un phénomène à court terme puisque les combinaisons génétiques défavorables seront éventuellement éliminées par sélection naturelle (Frankham et al. 2002).
Il La faune ichthyenne des Grands Lacs Laurentiens La communauté de poissons indigènes des Grands Lacs Laurentiens a souffert de modifications majeures depuis le début du vingtième siècle. L'exploitation commerciale intense et les introductions involontaires d'espèces exotiques agissant en synergie avec la pollution aquatique ont aggravé le déclin de plusieurs espèces indigènes ou ont empêché leur rétablissement (Christie 1973, Mills et al. 1991). Un exemple frappant d'un tel déclin ayant pour cause des influences anthropogéniques est celui des ciscos {Coregonus spp., excepté C. clupeaformis). Les ciscos ont eu une importance économique et culturelle incontestable, plus particulièrement aux États-Unis. La pêche aux 'chubs', florissante jusque dans les années 50, exploitait alors toutes les formes de cisco de profondeur des Grands Lacs. Avec C. artedi et C. clupeaformis, ils formaient alors la plus grande biomasse de poissons des Grands Lacs, les ciscos comptant pour la plus grande biomasse des eaux profondes (Smith 1995, Baldwin 1999). Aujourd'hui, les ciscos sont un groupe décimé et plusieurs formes qui occupaient les Grands Lacs historiquement, dont certaines endémiques, font l'objet de préoccupation. Le statut de C. kiyi est considéré comme préoccupant alors que d'autres espèces sont menacées (C. zenithicus, C. reighardi, C. nigripinnis) ou disparues (C. alpenae, C. johannae) (Cosepac 2006). Globalement, la plupart des formes de cisco de profondeur sont maintenant menacées ou disparues, à l'exception du cisco de fumage (C. hoyî) qui est toujours abondant dans certains lacs.
La taxonomie des ciscos des Grands Lacs a posé de nombreux problèmes à cause de leur extrême variation phénotypique et écologique, et ce, malgré les efforts continus depuis Koelz (1928) pour la résoudre. Des travaux récents sur la phylogéographie de C. artedi (Turgeon et al. 1999, Turgeon et Bernatchez 2001a,b, 2003) ont révélé une partie de leur histoire évolutive complexe. L'analyse des polymorphismes mitochondriaux de C. artedi a démontré l'existence de deux races glaciaires qui, après avoir été isolées durant la dernière glaciation du Pléistocène, sont entrées en contact secondaire lors de la recolonisation postglaciaire. Ainsi, les différentes formes endémiques de ciscos des Grands Lacs ont fort probablement évolué suite à une période d'hybridation intensive des deux lignées glaciaires de C. artedi. Des analyses de microsatellites ont indiqué que l'évolution des phénotypes caractérisant les multiples formes endémiques est survenue très récemment, soit après le contact secondaire des deux races glaciaires de C. artedi. De plus, ces formes endémiques ont probablement évolué indépendamment dans chaque lac où elles sont, ou
12 étaient, rencontrées (Turgeon et Bernatchez 2003). Chaque écomorphotype rencontré est donc l'expression unique d'une partie du pool génique constitué du mélange, à divers degrés de deux lignées distinctes, ce pool étant apparemment toujours en diversification.
Historiquement, quatre formes indigènes de ciscos de profondeur occupaient le lac Ontario, soit C. hoyi, C. reighardi, C. nigripinnis et C. kiyi, ainsi qu'une forme pélagique, C. artedi. Ces quatre premières espèces ont un jour formé la plus grande biomasse dans les régions profondes de chacun des Grands Lacs et étaient des sources de nourriture importantes pour le touladi (S. namaycush) et la lotte (Lota Iota) (Smith 1995). Le lac Ontario fut peut-être le lac qui eut le plus haut rendement de pêche de tous les Grands Lacs (Smith 1995). Mais au tournant du 20e siècle, C. nigripinnis, le plus gros des ciscos, était déjà disparu du lac Ontario et entre 1927 et 1942, C. kiyi et C. reighardi disparurent également (Christie 1973). Le cisco de fumage était à l'époque plutôt dédaigné, étant de plus petite taille que les autres espèces mais il acquit de l'importance une fois les plus grosses espèces décimées (Scott et Crossman 1973). À ce jour, toutes les formes de cisco de profondeur du lac Ontario sont considérées éteintes, laissant uniquement C. artedi et C. clupeaformis comme coregonidés encore présents (Christie 1973). Cependant, quelques prises occasionnelles de ciscos de profondeur laissent croire qu'il pourrait toujours exister une population réduite dans les strates profondes du lac. Depuis 2002, moins d'une dizaine de C. hoyi ont été attrapés dans le lac Ontario (Bruce Morrisson, OMNR, comm. pers.). Il reste toujours abondant dans les lacs Huron, Michigan, Supérieur et Nipigon; il n'a par ailleurs jamais été présent dans le lac Érié.
Réintroduction de C. hoyi dans le lac Ontario En dehors des aspects sociaux et culturels rattachés à la réhabilitation de C. hoyi dans le lac Ontario, le principal gain d'une réintroduction identifié par une majorité de scientifiques (voir Baldwin 1999) serait celui d'occuper l'habitat pratiquement vide des profondeurs du lac. La chaîne alimentaire est actuellement dépourvue de benthivore majeur, les populations de chabots de profondeur (Myoxocephalus thompsoni), le seul autre planctivore benthique saisonnier, étant assez basses (Smith 1995) et constituées presque seulement d'adultes (Robert O'Gorman, USGS, comm. pers.). Elle se termine donc dans un cul-de-sac, laissant de larges populations de mysidés non consommées et occasionnant une perte d'énergie en raison du flux énergétique interrompu. La
13 réintroduction d'un planctivore indigène se nourrissant de ces invertébrés pourrait recréer le lien trophique manquant et ainsi rétablir les transferts d'énergie originaux (Baldwin 1999, Stewart et al. 2002). Le rétablissement d'un planctivore aurait aussi l'autre avantage direct de fournir une source d'alimentation à S. namaycush et L. Iota. Contrairement aux autres coregonidés du lac Ontario, la distribution verticale dans la colonne d'eau de C. hoyi chevauche presque entièrement celle de S. namaycush, et il en fut une des proies favorisées historiquement (Scott et Crossman 1973).
Certains aspects intrinsèques à C. hoyi devront bénéficier d'une attention plus particulière lors de l'élaboration du plan de réintroduction. Certaines populations de coregonidés sont connues pour être possiblement porteuses de la maladie bactérienne des reins (Bacterial Kidney Disease, BKD) qui est un pathogène obligatoire des salmonidés. Il est connu pour infecter des populations captives et naturelles de poissons et sa transmission s'effectue soit entre adultes, soit d'un parent à la progéniture par les œufs (Warren 1983). Des individus exempts de pathogènes devront être choisis comme géniteurs afin de prévenir l'introduction de la maladie dans les populations du lac Ontario. D'un autre côté, la dynamique de la population suite à la réintroduction relève des caractéristiques inhérentes à l'espèce et à l'environnement. La fertilité, l'âge à maturité, la longévité, la probabilité de survie, les interactions de dépendance à la densité et la capacité de support du milieu devront être considérés comme autant de facteurs clés pouvant influencer l'évolution de la population au cours des années. C. hoyi est connu pour sa démographie peu commune, influencée principalement par le rapport des sexes. Une dynamique cyclique du rapport des sexes s'échelonnant sur 30 ans a été découverte dans le lac Michigan (Bunnell et al. 2006) et est suspectée dans le lac Huron (Schaeffer 2004). Les années de rapport des sexes équilibré sont associées à un fort recrutement tandis que les années de rapport des sexes biaisé vers les femelles à un faible recrutement (Bunnell et al. 2006).
Objectifs de recherche Dans le cadre de cette recherche, deux objectifs principaux se démarquent et chacun fera l'objet d'un chapitre de ce mémoire. Le premier objectif porte sur l'identification d'une source de cisco de fumage à des fins de réintroduction dans le lac Ontario. Cette source devra être une population (1) génétiquement diversifiée et ne montrant pas de
14 signes de déclin récent afin de minimiser la consanguinité dans la population réintroduite ; et (2) ayant une ascendance partagée maximale avec les ciscos du lac Ontario pour minimiser les risques de dépression d'outbreeding chez la potentielle population réduite de cisco de profondeur du lac Ontario et augmenter la probabilité d'expression de l'écomorphotype désiré. Pour ce faire, le polymorphisme de locus microsatellites de l'ADN seront analysés pour des échantillons de C. hoyi provenant des lacs Huron, Michigan, Supérieur et Nipigon. Des échantillons de C. artedi et de quelques ciscos de profondeur du lac Ontario seront également analysés afin d'obtenir un portrait de la diversité génétique des ciscos du lac Ontario.
Le deuxième objectif cherche à optimiser le plan de croisement artificiel nécessaire à la création d'individus à réintroduire en milieu naturel de façon à minimiser la consanguinité des individus fondateurs, ainsi qu'à évaluer la dynamique de cette consanguinité à plus long terme pour différents scénarios de réintroduction. Afin d'optimiser le plan de croisement, j'ai exécuté des simulations de croisements artificiels d'individus en utilisant les fréquences alléliques d'un échantillon diversifié et exempt de signe de goulot d'étranglement provenant du Chapitre 1. Par la suite, la dynamique de la population réintroduite a été étudiée en réalisant des simulations de réintroduction en milieu naturel à l'aide d'un modèle mathématique évaluant la dynamique de la consanguinité, couplé à un modèle démographique adapté à l'espèce étudiée. Les résultats obtenus pourront guider la réintroduction du cisco de fumage dans le lac Ontario de façon à maximiser les chances de réussite quant à la conservation d'une diversité génétique élevée à long terme et à l'expression du phénotype désiré.
15 CHAPITRE 1:
PATTERNS OF GENETIC DIVERSITY IN GREAT LAKES BLOATERS (COREGONUS HOYI) FOR FUTURE REINTRODUCTION IN LAKE ONTARIO
Running title: Reintroduction of bloaters in Lake Ontario
Names of author(s): Marie-Julie FAVÉ and Julie TURGEON* (correspondence author)
Institution address: Département de biologie Université Laval, Québec, Canada, G1K 7P4
Full mailing address : Département de biologie Université Laval, Québec, Québec, Canada G1K 7P4 Phone : 418-656-3135 Fax : 418-656-2043 Email : [email protected]
Keywords : reintroduction, genetic diversity, bottleneck, Coregonus hoyi, ciscoe, Laurentian Great Lakes
16 Résumé
La faune originale des ciscos des Grands Lacs Laurentiens a souffert de plusieurs extinctions à l'échelle locale et globale. Les ciscos de fumage (Coregonus hoyi) sont présumés disparus du lac Ontario et la réintroduction de cette espèce benthique est considérée. En tenant compte des fluctuations documentées de cette espèce dans les autres Grands Lacs et de son origine intralacustre récente, nous cherchons à identifier une source génétiquement diversifiée et similaire de cisco de fumage par l'analyse de la diversité et de la structure génétique de C. hoyi en utilisant 10 locus microsatellites. Malgré des déclins démographiques bien documentés, nous n'avons pas trouvé d'évidence de goulot d'étranglement dans 12 échantillons de ciscos de fumage provenant de quatre lacs donneurs potentiels (Huron, Michigan, Supérieur et Nipigon). Par contre, des goulots d'étranglement ont été détectés dans des échantillons historiques de C. artedi du lac Ontario, suggérant que les méthodes génétiques standard de détection des goulots d'étranglement ne peuvent détecter que des étranglements très sévères et de longue durée pour des populations ayant naturellement de hautes abondances. Les patrons de différentiation génétique suggèrent également que les quelques ciscos de profondeur récemment capturés dans le lac Ontario font partie d'une petite population de C. hoyi restée non détectée durant plusieurs années, les individus la composant étant plus similaires aux ciscos de fumage des lacs Huron et Michigan, lesquels ne sont pas génétiquement différenciés. En se basant sur les critères génétiques de grande diversité génétique, d'absence de goulot d'étranglement et de la similarité avec les ciscos du lac Ontario, nous suggérons que les ciscos de fumage provenant des lacs Huron ou Michigan, peu importe le site particulier, seraient un choix de géniteurs judicieux pour effectuer une réintroduction de C. hoyi dans le lac Ontario.
Abstract
The originally diverse ciscoe fish fauna of the Laurentian Great Lakes has suffered many extinctions and local extirpations. Bloaters {Coregonus hoyi) are presumed extirpated from Lake Ontario and the reintroduction of this deepwater fish is under considération. Given the démographie fluctuations of this species in the other Great Lakes
17 and its reœnt intralacustrine origin, we sought to identify a genetically diverse and similar source of bloaters via an analysis of genetic diversity and population structure of C. hoyi using 10 microsatellite loci. Despite well-documented démographie déclines, we found no genetic évidence of bottlenecks in 12 bloater samples from the four potential donor lakes (Huron, Michigan, Superior and Nipigon). By contrast, évidence of bottlenecks among historical samples of C. artedi from Lake Ontario suggested that standard genetic methods frequently used to identify population bottlenecks can only detect very severe and long- lasting démographie déclines in naturally large populations. Patterns of genetic differentiation also suggested that the few deepwater ciscoes recently caught in Lake Ontario are part of a small undetected C. hoyi population most similar to bloaters of Lake Huron and Lake Michigan, which are not genetically differentiated. On the basis of genetic criteria, and given the high genetic diversity, the absence of significant bottlenecks and the similarity to Lake Ontario ciscos, we conclude that bloaters from any location within Lake Huron or Lake Michigan would be judicious sources of breeders for reintroducing C. hoyi in Lake Ontario.
18 Introduction
Human impact on the environment is undeniable. For many species, it has repeatedly led to huge démographie déclines driving local populations and species to extinction. Conservation priority should obviously be given to improved management and protection policies. Unfortunately, the near or total extirpation of a species from a portion of its native range too often prevents this idéal approach. In such cases, reintroduction and/or supplementation of vacant site(s) is the only option to restore ecological diversity and its associated biological functions. Reintroduction and supplementation imply the release of artificially-reared individuals, or alternatively the translocation of wild-caught individuals, in an area where a native species is extirpated or severely declined, respectively. Thèse conservation practices, however, should be conducted with care and considération for the species biology, health status, genetic structure and diversity (Bowen 1999, Moritz 2002).
The first genetic risk associated with establishing a new population is a founder effect whereby genetic diversity is reduced because of the small number of individuals used for the reintroduction. This effect can be further amplified if founding individuals are taken from a genetically depauperate population, e.g. because of a past bottleneck (Ramstad et al. 2004), and it sets the stage for increased inbreeding in the new population. Care should therefore be taken to sélect founders so as to reflect the original genetic diversity (Frankham et al. 2002, Moritz 1999). Indeed, reduced diversity in small populations can negatively affect fitness-related traits such as reproductive capacity or survival; it can also increase the probability of expression of deleterious mutations, and eventually bring a population to extinction (Ryman and Laikre 1991, Frankham 1996, Hartl and Clark 1997, Frankham et al. 2002, Keller and Weller 2002,). For example, reduced genetic variability has been associated with weak résistance to bacterial pathogens and viruses in animal species (O'Brian and Evermann 1988, Sherwin et al. 2000). Similarly, high genetic diversity was associated with enhanced survivorship under changing biological conditions in eelgrass Zostera marina (Hughes and Stachowicz 2004), indicating that it should be maximized to improve the likelihood of long-term persistence of reintroduced species.
19 Another genetic risk, often considered in supplementation programs is the possibility of creating an outbreeding dépression as a resuit of crosses between indigenous and introduced individuals (Hedrick and Kalinowski 2000). Indeed, reproductively isolated populations may evolve unique co-adapted gène complexes spécifie to local environment (Lynch 1991), and disruption of those complexes by hybridization with individuals bearing différent genomic architecture may negatively affect fitness-related traits in offspring. For example, Gilk et al. (2004) hâve shown that FI and F2 hybrids between genetically differentiated pink salmon (Onchorhynchus gorbuschà) populations had significantly reduced survival rates. On the other hand, évidence for hybrid vigor is reported in many Systems (e.g. Seehausen 2004), indicating that the introduction of foreign individuals in inbred or bottlenecked populations can often be valuable by providing new genetic material and bénéficiai alleles (Ebert et al. 2002, Whitlock et al. 2000, Tallmon et al. 2004). Furthermore, an outbreeding dépression could be a short-term phenomenon since the unfit genetic combinations would eventually be nearly eliminated by natural sélection (Frankham et al. 2002).
Laurentian Great Lakes Fish Fauna The native fish community of the Laurentian Great Lakes has suffered major modifications since the beginning of the 20th century (Crossman 1991). Intense commercial fisheries, inadvertent exotic species introductions and water pollution hâve exacerbated the downfall of many indigenous species and/or prevented their recovery (Christie 1973, Mills et al. 1991, Ricciardi 2001). Thèse modifications hâve had effects on the original fish biodiversity of the area as well as on lake ecosystem functions (Horns et al. 2003, Ebener 2005, Holey and Trudeau 2005, Mills et al. 2005)
One striking example of a human-induced native species décline in the Great Lakes is that of the ciscoes (Coregonus spp. except C. clupeaformis). Each lake once possessed several ecomorphotypes displaying différences in trophic traits, diet, spawning season, and habitat depth (McPhail and Lindsey 1970, Scott and Crossman 1973, Smith and Todd 1984, Turgeon et al. 1999). Many of thèse were endémie to the Great Lakes and enhanced global biodiversity of the zone (Scott and Crossman 1973). Moreover, thèse unique ecotypes likely evolved recently within each lake following extensive secondary contact between two glacial races (Turgeon and Bernatchez 2001, 2003). Nowadays, the pelagic
20 lake herring {C. artedi) is présent in ail of the Great Lakes, while most deepwater ciscoes are either extinct or restricted to Lake Superior and/or Lake Nipigon (Philips and Ehlinger 1995). The only exception is the bloater (C. hoyi), which remains abundant in Lakes Huron, Michigan, Superior and Nipigon (it never occurred in shallower Lake Erie).
In Lake Ontario, the décline in forage fishes such as C. hoyi followed the gênerai pattern observed in the other Great Lakes, with three out of four deepwater forms gone extinct by the mid-20th century, namely C. nigripinnis, C. reighardi and C. kiyi (Christie 1973). Bloaters persisted longer but became scarce by the 60s, and the last documented catch dates back to 1983 (Baldwin 1999). Bloater's décline coincided with the expansion of invading smelt populations {Osmerus mordax), but the présence of other non-indigenous species, notably the alewife {Alosa pseudoharengus) and the sea lamprey {Petromyzon marinus), may also hâve accelerated the final démise of C. hoyi (Christie 1973, Christie 1974, Stedman and Argyle 1985). However, rare occasional catches of deepwater ciscoes since 2002 (less than ten individuals) suggest that there could still exist an undetected remnant population of C. hoyi in the deepest layers of the lake, or alternatively, that lake herrings are gradually invading the underexploited deep layers of the lake.
Reintroduction ofC. hoyi in Lake Ontario The reintroduction of C. hoyi in Lake Ontario has been identified as a priority for achieving a healthy lake ecosystem through the restoration of the native fish fauna (Baldwin 1999, Stewart et al. 1999, Eshenroder and Krueger 2002). Beyond a gain in native biodiversity, the principal benefit of C. hoyi rehabilitation in Lake Ontario would be to re-establish the original energetic pathways linking benthic production, which is now left largely unconsumed, with top predators in the upper layers of the lake. Indeed, bloaters feed on epibenthic amphipods {Mysis relicta, Diporeia spp.) and perform diel vertical migrations (TeWinkel and Fleischer 1999), thus providing a trophic link with their indigenous predators, namely Salvelinus namaycush and Lota Iota.
In this paper, we use genetic criteria to identify a source of bloaters that would minimize genetic risks associated with its planned reintroduction in Lake Ontario. We first aimed at identifying genetically diverse, non-bottlenecked population(s) that would help minimize inbreeding in the introduced population. Concerns for this potential effect are
21 justified because of the known unstable demography of bloaters in ail of the Great Lakes (Schaeffer 2004, Baldwin et al. 2005, Bunnell et al. 2006). Secondly, we wished to identify a source population of C. hoyi with maximum shared ancestry with Lake Ontario ciscoes in order to minimize the risk of outbreeding dépression in the putatively remnant deepwater- like ciscoes of Lake Ontario. To this effect, we analyzed variation at 10 microsatellite loci in C. hoyi samples from ail potential donor lakes. In addition to that, we also tested for the origin of the putatively remnant deepwater-like ciscoes of Lake Ontario, composed of either individuals derived from pelagic C. artedi progressively invading the deeper layers of the lake, or from a truly remnant C. hoyi population that was undetected for décades. We hypothesize that if the genetic differentiation between those deepwater ciscoes and Lake Ontario C. artedi individuals is non-significant, the deepwater population originates from recently derived C. artedi individuals invading the vacant benthic habitat. The alternate hypothesis of a significant genetic differentiation would indicate an historical isolation between the two populations that allowed for a détectable divergence, and thus confirming the remnant origin of a deepwater C. hoyi population.
Materials and Methods
Biological material Tissue samples (adipose fin or muscle) were obtained from 551 bloaters collected from three différent sectors within each potential donor lake (PDL), namely Lake Huron, Lake Michigan, Lake Superior, and Lake Nipigon (Table 1). Given the virtual absence of C. hoyi in Lake Ontario, we used samples of C. artedi individuals as a surrogate sample of the genetic diversity of Lake Ontario ciscoes. Indeed, because of the intralacustrine diversification of Great Lakes ciscoes (Turgeon and Bernatchez 2003), extant lake herrings are the closest living relatives and share maximal ancestry with bloaters that once inhabited Lake Ontario. Thus, we analyzed tissue from 51 C. artedi (ONT-artedi) and five deepwater ciscoes (ONT-deep) caught in Lake Ontario in 2004 to portray Lake Ontario modem genetic diversity. Ail samples were preserved in 95% ethanol.
In addition, 116 Lake Ontario C. artedi scale samples from an historical temporal séries (1968 to 1994) were characterized along with 28 EtOH-preserved samples from 1997 used by Turgeon and Bernatchez (2003) (Table 1). Ail of thèse samples were caught
22 in Hay Bay (Canada) and were obtained from the Glenora Fisheries Research Station (Ontario Ministry of Natural Resources). This temporal séries was used to assess the ability of our markers to detect known population décline (see below).
DNA extraction and microsatellite analysis DNA from récent samples was extracted either by using Miniprep extraction plates and a vaccum pump (Millipore Montage BAC96 Miniprep Kit) following the manufacturer's instructions or with a classic phenol-chloroform protocol (Sambrook and Russell 2001). Ail individuals were amplified at 10 microsatellite loci, among which five were developed for this study (Table 2). Ail multiplex amplifications were performed in a 15 juL total reaction volume including -20 ng of genomic DNA, 20 //M of each dNTP, IX buffer (500 mM KC1, 200 mM Tris-HCl pH 8.8, Triton X 1%), 1.5 mM MgCl2, 1U Taq polymerase and 0.15-0.6 |J.M primer. Amplification profiles were as described in Turgeon et al. (1999). Diluted PCR products were pooled into one of two genotyping sets and separated by capillary electrophoresis on an ABI 3100 automated sequencer (Table 2). Alleles were sized using GeneMapper V. 3.7.
DNA from scale samples was extracted following the protocol of Nielsen et al. (1997) with the following modifications: scales were digested 2 hours at 55 °C, 10 juL of RNAse and 10 juh of proteinase K were added 45 minutes before the end of the digestion, and the second elution step was performed with 10 //L of distilled water. Two microsatellite loci (cisco-106 and cisco-183) were discarded because amplification was very difficult on low DNA concentration samples. Each locus was individually amplified by two successive PCR reactions; 1.5//L IX BSA was added to the reaction and amplification was performed with 5 additional cycles.
Genetic diversity and historical demography
Estimated number of alleles (NA), expected (HE) and observed (Ho) heterozygosity were calculated using GENETIX V.4.05.2 (Belkhir et al. 2004). We examined the data for the présence of allelic "drop-out" events associated with low quality or degraded DNA
using MICROCHECKER V.2.2.3 (Van Oostershout et al. 2004). Allelic richness (NAR, Petit et al. 1998) was estimated independently for récent and historical samples given the important variation in sample sizes and the différent number of loci considered (Table 1).
23 Tests for déviation from Hardy-Weinberg and linkage equilibria were conducted using exact tests in GENEPOP V.3.4 (Raymond and Rousset 1995). Bonferroni corrections (Rice 1989) for multiple comparisons were carried out using FSTAT V.2.9.3.2 (Goudet 2001).
We employed two methods to identify past bottlenecks in each PDL sample. First, data were analyzed with BOTTLENECK V.1.2.02 (Cornuet and Luikart 1996), assuming a two-phase mutation model (TPM) (variance=30%, SMM=70%). Under mutation-drift equilibrium, important réduction in effective population size is expected to yield a significant heterozygosity excess given the observed number of alleles (hereafter HE test). Second, we computed the M statistic of Garza and Williamson (2001), i.e. the ratio between the mean number of alleles and the allele size range. One locus (cisco-126) was not included in this calculation because its compound nature (Turgeon et al. 1999) biased calculations towards very low M values. According to Garza and Williamson (2001), an M value equal to or smaller than 0.68 is a conservative indication of an important réduction in the effective size of a population, while an M value greater than 0.82 is indicative of populations that hâve not experienced substantial size fluctuations. Moreover, given that bloater populations are known to undergo large démographie fluctuations (e.g. Bunnell et al. 2006), we used the temporal sample séries available for C. artedi from Lake Ontario to assess the potential of our genetic markers at detecting a réduction in genetic diversity and/or a past bottleneck. This was achieved by comparing M and NAR values to the relative abundance of ciscoes (as estimated by commercial catches) during the period covered by our temporal samples séries (i.e. 1968 to 1997).
Genetic relationships among Great Lakes bloater s and Lake Ontario ciscoes Global patterns of relationships among Great Lakes bloaters and Lake Ontario ciscoes were first assessed with conventional analyses. Inter-population DCE distances (Cavalli-Sforza and Edwards 1967) were estimated and a Neighbor-Joining tree was constructed using PHYLIP V.3.65 (Felsenstein 2005). Then, pairwise FST values among ail récent samples were estimated with FST (Weir and Cockerham 1984) using ARLEQUIN V.2.0 (Schneider et al. 2000). Given the small number of deepwater ciscoes caught in Lake Ontario (ONT-deep; N = 5), we investigated the effect of this small sample size on patterns of genetic differentiation identified with the latter analysis. To this effect, virtual samples were created by randomly choosing 5 individuals from a PDL C. hoyi sample and
24 assessing differentiation. This was repeated ten times and FST significance levels obtained were compared to the actual FST significance level between ONT-deep and ONT-artedi.
Given the very récent ancestry of ail ciscoe types in the Great Lakes, we also used an analytical approach seeking to group PDL bloaters with similar ancestral genetic characteristics. Similarity of Lake Ontario ciscoes to thèse référence groups was then assessed. The référence groups were defined with the Bayesian method implemented in STRUCTURE V.2.1 (Pritchard et al. 2000). Ail PDL individuals were included to détermine the number of cohesive genetic clusters (K) présent in the System. Twenty runs were performed for each K value (K =1 to 9), with 10 000 burn-in and 10 000 MCMC itérations. Evanno et al. (2005) recently used second-order rate of change in Ln P(D) between successive K values in order to enhance the ability of STRUCTURE to detect the real number of clusters when K >2 but, this procédure did not always yield better results (Waples and Gaggiotti 2006). Therefore, we chose to use the original advice of Pritchard et al. (2000) and selected the value of K associated with the highest posterior probability of the data (LnP(D)), while also seeking to minimize variability among runs.
Having identified three cohesive groups as référence clusters (see Results), a second analysis was performed with STRUCTURE to estimate the genomic proportion of each Lake Ontario individual's shared ancestry (q) in each cluster. Then, each individual was assigned to the référence clusters of highest q value, but only if over 50%.
Thèse assignments were then compared to those obtained with assignment tests performed with GENECLASS V.2.0 (Piry et al. 2004) using the Bayesian statistical approach of Rannala and Mountain (1997), which is more efficient than distance-based methods (Cornuet et al. 1999). Given that the true population of origin of Lake Ontario ciscoes is obviously not any of the PDLs, we forced the assignment of each spécimen to one of the PDL by not allowing the exclusion of any population of origin.
Results
Genetic diversity and historical demography
25 Considérable microsatellite gène tic variation was detected in ail samples (Table 1, detailed information available in Appendix A). In PDL samples, the observed heterozygosity (Ho) was rather high (0.57 to 0.78), and allelic richness (NAR) varied from 10.9 to 13.1 alleles per sample. In Lake Ontario historical samples, for which only 8 loci were considered, NAR ranged from 5.1 to 7.9 alleles per sample. Exact probability tests across populations did not show any significant departure from Hardy-Weinberg equilibrium. Similarly, no significant déviation from linkage equilibrium was found for any pair of loci (P > 0.05).
Very little évidence of past bottlenecks was found among PDL samples with either of the two tests employed. First, Wilcoxon sign rank HE test (Cornuet and Luikart 1996) did not yield significant results for any samples (Table 1). Likewise, M ratios seldom exceeded the range of values indicative of drastic démographie changes (0.68-0.82) in PDL bloater samples. Only one sample from Lake Nipigon (NIP-A) yielded M ratios close or below the cut-off value indicative of a population bottleneck, while other samples showed values between 0.69 and 0.79 (Table 1).
For the Lake Ontario temporal séries, M ratios ranged from 0.57 to 0.83, with samples from 1968, 1991 and 1994 having M ratios clearly below the 0.68 threshold value, and the 1997 sample exhibiting an M ratio slightly over 0.82 (Table 1). Variation of M ratio and NAR over the period covered by the temporal séries followed, albeit with some lag for M, the historical records of relative ciscoes catches in the Canadian waters of Lake Ontario (Figure 2). The important réduction in population size, which reached its lowest level around 1987, was reflected by a drop in allelic richness in the 1986 sample as well as a décline in M ratio from 1986 to 1994. A sensible démographie rebound between years
1989 and 1995 was also followed by a small increase in NAR and M ratio. Finally, a second decrease in NAR and M ratio was observed in 2004, once again following an important drop in ciscoes abundance between 1995 and 2000.
Relationships among populations and individual assignment tests The neighbor-joining DCE phenogram relating PDL bloaters samples and Lake Ontario ciscoes readily suggests greater genetic similarity within lakes as well as between nearby lakes (Fig. 3). Samples from Lake Nipigon clearly defined a distinct cluster and
26 formed a very well-supported group with bloaters from the geographically proximate Lake Superior. Although of uncertain taxonomic status, ciscoes from the deepest section of Lake Ontario (ONT-deep) strongly grouped with the lake herring sample (ONT-artedi; bootstrap support of 99%, Fig. 3). It is worth noting that the branch linking the latter sample is not excessively long even if leading to a différent taxon. Finally, samples from Lakes Huron and Michigan did not show any clear grouping pattern but stood between the Ontario and Superior-Nipigon clusters, reflecting their geographical setting.
The pattern of population differentiation, which was globally significant (FST = 0.027; P < 0.00001), reflected similar trends. First, there was no significant differentiation within any of the PDLs. Likewise, there was no évidence for significant differentiation between samples from Lake Huron and Lake Michigan. By contrast, nearly ail other comparisons between PDLs yielded significant FST ranging from 0.010 (NIP-B/SUP-B) to 0.058 (MCH-A/NIP-A), with only one FST value not significantly différent from zéro (SUP-B/NIP-C). The lake herring sample from Lake Ontario was significantly différent from ail PDL (FST: 0.033 - 0.098) as well as from ONT-deep (0.062, Table 3). The small sample from ONT-deep was also differentiated from most PDL samples, but less so from bloater samples from Lake Huron or Lake Michigan. Differentiation was likely not due to the small sample size of ONT-deep, as none of the pseudo-samples of size N = 5 were differentiated from their source sample (P > 0.05).
Bayesian clustering analysis identified three genetic clusters among PDL individuals. Evidence for three clusters was provided on the one hand by the highest probability value for K = 3, and on the other hand by the much larger standard déviations among runs as K further increased (Table 4). Ail but 20 individuals unambiguously belonged to one of thèse référence clusters, with highest coefficients of membership (q) averaging 0.83 ± 0.17 (Table 5a). Each partition included a majority of individuals from Lake Superior, Lake Nipigon or from both Lakes Huron and Michigan, and thèse partitions are referred to accordingly hereafter (i.e. Sup, Nip, and HurMch).
Lake Ontario ciscoes were easily assigned to one of thèse partitions, with a high average coefficient of membership of 0.79 ± 0.14 (Table 5a). Nineteen C. artedi were classified in HurMch and 32 were classified in Sup. No individuals were classified in Nip.
27 Four out of the five deepwater ciscoes were classified with strong support in HurMch (q > 0.67, Table 5a) while the other deepwater cisco could not confidently be assigned. The assignaient procédure with GENECLASS was also successful, with 84% of individuals confidently reassigned to their lake of origin. As for Lake Ontario individuals, 27 C. artedi were assigned to Lake Huron, 11 to Lake Michigan, eight to Lake Superior and five to Lake Nipigon (Table 5b). Individuals assigned to Lakes Huron and Michigan were assigned with higher average q scores (0.72 and 0.73 respectively) than individuals assigned to Lakes Superior or Nipigon (0.66 and 0.64 respectively). Deepwater ciscoes were assigned either to Lake Huron (N = 2) or Lake Michigan (N = 3). Even if more individuals were assigned to Lake Michigan than Huron, those assigned to Lake Huron were assigned more confidently (0.90).
Discussion
Genetic diversity and bottlenecks Our analyses revealed that ail contemporary Great Lakes bloater populations hâve retained substantial genetic diversity, and that there are neither strong nor widely distributed genetic évidence of démographie bottlenecks. Thèse results are ail the more surprising given the high commercial exploitation during the last century and the well- documented drastic fluctuations in ciscoe population abundances (Christie 1974, Baldwin et al. 2005, Bunnell et al. 2006). Indeed, analysis of commercial catch data and long- lasting monitoring programs (Christie 1973, Baldwin et al. 2005) hâve clearly demonstrated that bloater and lake herring populations hâve been reduced below exploitation levels on several occasions during the 20th century. For example, in Lake Superior, abundance of deepwater ciscoes inferred from catch data reached a maximum abundance slightly above one million fish at the end of the 70s, but dropped below 50 thousands fish during three periods in the 20th century (Baldwin et al. 2005). Similar trends are also documented from the other Great Lakes (Horns et al. 2003, Baldwin et al. 2005, Fleischer et al. 2005). In Lake Michigan, Bunnell et al. (2006) recently analyzed a 30 year-long data séries and they hâve convincingly suggested that population fluctuations may be under natural endogenous cycles whereby high and low recruitment phases alternate at about 15 years intervais. Thus, despite repeated periods of population réductions, widely distributed genetic évidence for bottleneck is lacking.
28 It appears that the M ratio and HE tests, albeit widely used as standard tests in conservation genetics, often fail to detect the genetic signatures expected from major démographie déclines in real endangered populations. Indeed, failures to detect significant loss of diversity in cases of known démographie déclines hâve been reported for a variety of organisms such as marsupials (Jones et al. 2004), reptiles (Kuo and Janzen 2004), mammals (Whitehouse and Harley 2001, Harley et al. 2005), insects (Watts et al. 2006), and fish (Lippe et al. 2006). Among the potential causes invoked by thèse authors, the pattern of démographie décline, the species lifespan, a past founder effect and the type of genetic markers may hâve played a rôle in the case of the Great Lakes bloaters. First, a graduai rather than sudden pattern of décline has been put forward as a possible cause for the absence of impact on the number of alleles and heterozygosity (Kuo and Janzen 2004, Lippe et al. 2006). For the majority of Great Lakes ciscoes, population décline was long and graduai and span over more than 80 years, with huge historical abundances apparently preventing steep population drops in abundances. Moreover, although population réductions reported in ciscoes appear enormous, thousands of individuals were still counted at lowest historical abundances. Thèse minimum, yet large population sizes probably prevented any significant allelic loss, and may in fact not qualify as real bottlenecks. Similarly, a graduai décline is reported to reduce the power of the HE test at detecting bottlenecks (Harley et al. 2005). Second, Kuo and Janzen (2004) and Lippe et al. (2006) suggested that the long lifespan of species could also prevent the détection of bottlenecks with both tests. Although not explicitly invoked, this factor may also hâve played a rôle in the failure to detect known bottlenecks in long-lived species such as éléphant (Whitehouse and Harley 2001), Tasmanian devils (Jones et al. 2004), and rhinocéros (Harley et al. 2005). With regards to bloaters, they are known to live between 6 and 11 years (Scott and Crossman 1973) and this moderate lifespan could be an aggravating factor contributing to the failure of the tests. Third, Jones et al. (2004) proposed that past founder effect could so severely reduce the original genetic diversity that no clear pattern of allelic loss could be detected after subséquent bottlenecks. In the case of the Great Lakes ciscoes, however, the opposite argument seems to apply. Indeed, extensive genetic diversity is présent in Great Lake ciscoes (Table 1, Appendix A, Todd 1981, Turgeon et al. 1999, Turgeon and Bernatchez 2003) and very large long-term effective population sizes in thèse forage fishes has apparently prevented major losses of
29 genetic variation. In fact, only very severe and prolonged bottlenecks such as in Lake Ontario hâve caused a réduction in population abundance sufficient to generate a strong genetic érosion signal (Table 1, Figure 2).
Finally, failure to detect bottlenecks may be due to the lack of resolution caused by mutation modalities or allele frequency distribution of genetic markers, as has been suggested by Whitehouse and Harley (2001). This potential explanation, however, is rejected for Great Lakes bloaters. Indeed, the ability of our markers to identify bottlenecks is validated by the results for the Lake Ontario historical séries, with M ratios below or on the threshold value of 0.68 in 1968, 1978, 1991 and 1994. Knowing that Lake Ontario harbors the most severely impacted populations of ciscoes, thèse results support the hypothesis that extrême and long-lasting low abundances are needed to provoke détectable réductions of genetic diversity in naturally large populations such as bloaters or lake herrings. Hère, the décline of Lake Ontario C. artedi population started around 1941 and abundances of less than 50 000 fish per year are consistently reported since 1955. The fact that the two most récent samples do not exhibit M ratios characteristic of bottlenecked populations may indicate some recovery or at least, stability of the population size in the last décade or so. In fact, M ratios tend to increase gradually even in reduced but demographically stable populations (Garza and Williamson 2001).
Incidentally, the inadequacy of available tests to detect bottleneck is further suggested by the few significant results obtained in this study. For example, the détection of a bottleneck in Lake Nipigon (NIP-A, Table 1) is unexpected given that Lake Nipigon is the only Great Lake that never supported a commercial ciscoe fishery. While no long-term population data is available for this lake, it seems legitimate to suppose that it has not been disturbed to the same extent as the other lakes. Ail thèse results suggest that the M and HE tests used to detect genetic bottlenecks are far from infallible, and that they should be applied and interpreted very carefully. Indeed, when démographie data are not available, such as is often the case in non-commercial species, severely declining populations may well be identified as not having suffered severe abundance réductions, and the lack of évidence for bottlenecks could mislead conservation décisions.
30 In relation to the reintroduction of bloaters in Lake Ontario, the lack of évidence for bottlenecks so severe as to significantly affect genetic diversity is a positive finding. Indeed, it appears that the récent evolutionary history of the Great Lakes ciscoes, which involved massive hybridization fostering genetic diversity, as well as the large size and resilience of populations of this long-lived species hâve prevented undesirable diversity érosion. In terms of diversity, then, any one of the Great Lakes is equally fit to provide the reproductive adults needed for reintroducing deepwater ciscoes into Lake Ontario.
Differentiation within and among lakes Ail analyses coincide with the lack of genetic differentiation within any of the Potential Donor Lakes, the great similarity between bloaters from Lake Huron and Michigan, and the significant genetic differentiation of bloaters from Lake Superior and Lake Nipigon (Figure 3, Table 3). When considered in conjunction with the analyses including ciscoes from Lake Ontario, thèse results provide background information pertinent to the sélection of a source of bloaters for artificial rearing and subséquent reintroduction in Lake Ontario. First of ail, the absence of significant spatial genetic substructure within each PDL indicates that the précise location for collecting C. hoyi in the selected donor lake is not relevant. More importantly, this gênerai resuit suggests that ciscoes recently caught in the deepest section of Lake Ontario likely consist of a remnant, undetected population of bloaters. Indeed, given that sympatric bloaters are not genetically differentiated within any of the PDL, the significant differentiation between deepwater ciscoes and sympatric lake herring suggest that they belong to a différent ciscoe lineage. As hypothesized in the introduction, if the deepwater ciscoes were merely derived from lake herrings having recently recolonized the vacant deeper strata of Lake Ontario, it is unlikely that they would be so significantly differentiated from the récent ancestral stock (FST = 0.062, P < 0.0001). Certainly, the fact that only five deepwater ciscoes were available for genetic analyses could be seen as an important weakness, and we do not deny that more samples would hâve been of great value. However, ail tests with small virtual populations revealed non significant genetic differentiation from the original samples, indicating that a sample of only five individuals was sufficient to detect genetic differentiation. While this interprétation confirms that the original genetic diversity of Lake Ontario bloaters can no longer be recovered, it also indicates a tangible risk of outbreeding
31 dépression and underlines the importance of choosing a PDL most genetically similar to the Lake Ontario deepwater ciscoes.
Patterns of genetic similarity and assignment analyses indicate that Lake Huron and Lake Michigan represent the best sources for the reintroduction of bloaters in Lake Ontario. Ail analyses indicated that both lakes are equally suitable as sources of bloaters. Indeed, there was no support for genetic differentiation among samples from thèse lakes
(Table 3), thèse were most similar and intermingled on the DCE phenogram (Figure 3), and a Bayesian analysis clearly indicated that a majority of individuals from Lake Huron and Michigan define a cohesive genetic partition (HurMch, Table 5a). In turn, Lake Ontario deepwater ciscoes were ail assigned to this partition or to thèse lakes on the basis of extensive shared ancestry or high assignment scores, respectively (Table 5a, b). With regards to Lake Superior and Lake Nipigon, they defined distinct genetic partitions, and ail samples from thèse lakes were significantly differentiated from the best suited donor lakes. Bloaters from Lake Nipigon were most strongly differentiated from ail the other Great Lakes (Table 3), suggesting that this lake would be the least suited lake to obtain bloaters from. Finally, it is worth noting that ail of thèse results are consistent with previous results indicating that geography is the best predictor of genetic similarity, and that Lake Nipigon belongs to the western group of cisco populations whereas alleles typical of the Atlantic glacial race predominate in ail other Great Lakes (Turgeon and Bernatchez 2003).
Conclusions
This study surveyed patterns of neutral genetic variation in Great Lakes ciscoe populations that are relevant to the future re-introduction of C. hoyi in Lake Ontario. Our results first indicate that bloaters from ail potential donor lakes are genetically very diverse. The absence of genetic signature revealing démographie déclines was rather unexpected given the known fluctuations of commercially-exploited bloater populations during the past century. While suggesting that the available tests may often be unable to detect bottlenecks, thèse results also indicate that the original diversity of huge forage fish populations such as bloaters can prevent the érosion of genetic diversity, even with récurrent periods of abundance below exploitation level. Second, our analyses confirmed that the few available deepwater ciscoes of Lake Ontario likely represent a remnant
32 population of bloaters, and that thèse are most similar to Lake Huron and Lake Michigan populations. Given that there was no apparent spatial structure within Lake Huron or Lake Michigan, we conclude that bloaters from any location within either lake would be suitable as a source for reintroduction. Thèse recommendations, based solely on genetic criteria, will obviously need to be weighted in light of fish health and logistical considérations.
Acknowledgments
This work was supported by a Great Lakes Fishery Commission grant to JT. We would like to thank those who provided samples: Rick Salmon (OMNR), Lloyd Mohr (OMNR), Bruce Morrison (OMNR), Owen Gorman (USGS), and Kim Scribner (Michigan State University).
33 Figure legend
Figure 1: Sampling locations for Coregonus hoyi in Potential Donor Lakes (•) and for ciscoes in Lake Ontario (•: contemporary and historical scale samples of Coregonus artedi; ^ : deepwater ciscoes caught in 2002).
Figure 2: Unrooted DcE-based Neighbor-joining phenogram relating ail samples frorn Potential Donor Lakes and Lake Ontario ciscoes (ONT-ARTEDI: récent sample of C. artedi; ONT-DEEP: deepwater ciscoes caught in 2002). Bootstrap values above 60 are shown.
Figure 3: Temporal variation in alielic richness (NAR) and demographical index (M ratio) in historical ciscoes from Lake Ontario (Hay Bay, Canada) in relation with relative commercial ciscoes catch in the Canadian waters of Lake Ontario (grey bars; Baldwin et al. 2005).
34 Table 1: Information on ciscoe samples, along with détails on genetic diversity (NA: number of alleles; NAR: allelic richness; Ho and He: observed and expected heterozygosity, respectively), and results of démographie tests, (P-values for heterozygosity déficit and excess (Cornuet and Luikart 1996); M ratio of Garza and Williamson (2001)).
Sample Information NAR récent NAR temporal Ho He M ratio H déficit H excess Code N type Area samples (10 loci) séries (8 loci) P-value P-value
Hur-A 47 C.hoyi Goderich, CAN 14.3 11.8 0.69 0.75 0.76 0.07 0.95
Hur-B 48 C. hoyi Lion's Head, CAN 14.3 12 0.72 0.76 0.73 0.25 0.78 Hammond Bay, Hur-C 51 C.hoyi USA 16.2 12.9 0.71 0.77 0.72 0.25 0.78
Mch-A 50 C. hoyi Waukegan, USA 16.4 13.1 0.7 0.75 0.76 0.04 0.98
Mch-B 48 C.hoyi Beaver Island, USA 13.7 11.8 0.71 0.76 0.77 0.50 0.54 Frankfurt/Ludington, Mch-C 49 C.hoyi USA ' 14.8 12.3 0.72 0.75 0.79 0.31 0.72 Stockton/Basswood Sup-A 50 C.hoyi Islands, USA 13.2 11 0.7 0.74 0.69 0.35 0.69
Sup-B 34 C.hoyi Thunder Bay, CAN 12 11.6 0.74 0.76 0.73 0.46 0.58
Sup-C 48 C.hoyi Marquette, USA 13.4 11.3 0.69 0.74 0.72 0.19 0.84
Nip-A 53 C.hoyi Humboldt, CAN 15.3 12.3 0.75 0.78 0.63 0.28 0.75
Nip-B 29 C.hoyi The Willows, CAN 11.1 10.9 0.78 0.76 0.78 0.62 0.42
Nip-C 44 C.hoyi Gros Cap, CAN 14.9 12.4 0.77 0.78 0.79 0.72 0.31
Ui Sample Information NA NAR récent NAR temporal Ho He M ratio H déficit H excess Code type Area samples (10 loci) séries (8 loci) P-value P-value deepwater Ont-deep 5 ciscoes Rocky point, CAN 4.5 - 0.64 0.59 - - -
Ont-artedi 51 C. artedi Hay Bay, CAN 12.8 6.1 0.7 0.71 0.72 0.08 0.94
1997 28 C. artedi Hay Bay, CAN 10.8 7.3 0.65 0.67 0.83 0.01 0.99 ciscoe 1994 22 scales Hay Bay, CAN 9.8 7.9 0.57 0.69 0.57 0.13 0.90 ciscoe 1991 24 scales Hay Bay, CAN 8.3 6.8 0.66 0.68 0.64 0.16 0.88 ciscoe 1986 17 scales Hay Bay, CAN 6.4 6.1 0.61 0.65 0.70 0.58 0.47 ciscoe 1978 28 scales Hay Bay, CAN 6.9 6.6 0.63 0.68 0.68 0.68 0.37 ciscoe 1968 25 scales Hay Bay, CAN 5.9 5.1 0.65 0.62 0.66 0.58 0.47
U) Table 2: Microsatellite loci used for the characterization of Great Lakes ciscoes: primer séquences with fluorescent label, PCR conditions (primer concentration, TA, and number of cycles), PCR multiplex and migration set, allele size range (bp), and total number of alleles observed.
Locus Primer séquence (5'-3') Label Primer T°A Multiplex- Allele Total allele (uM) (N cycles) Migration set size range number cisco-90 * F: CAGACATGCTCAGGAACTAGf 6-FAM 0.3 57 (30) 1-1 104-127 9 R: CTCAAGTATTGTAATTGGGTACt bwf2** F: CGGATACATCGGCAACCTCTG PET 0.3 57 (30) 1-1 146-201 11 R: AGACAGTCCCCAATGAGAAAA cisco-59 *** F: AGTTGTGTTAGAGGCACAGC NED 0.6 57 (30) 1-1 163-248 35 R: GATAGCTCCCAGGGTTAGTT cisco-181 * F: GGTCTGAATACTTTCCAAATGCAC 6-FAM 0.2 60 (25) 2-1 145-362 52 R: CCATCCCTTTGCTCTGCC cisco-200 * F: GGTTAGGAGTTAGGGAAAATATG VIC 0.6 60 (25) 2-1 154-264 38 R: GTTGTGAGGTAGGCCTGG cisco-106 *** $ F: TCGTCGTCAGGTGAACAG VIC 0.4 55 (30) 3-2 200-286 44 R: GGATTATTTAAAGGCCCAGT cisco-126 * F: GCCAGAGGGGTACTAGGAGTATG NED 0.15 55 (30) 3-2 150-202 7 R: GCAGAGAAAGAGCCTGATTGAAC cisco-157 * F: CTTAGATGATGGCTTGGCTCC VIC 0.2 55 (30) 4-2 129-164 20 R: GGTGCAATCACTCTTACAACACC cisco-179 *** F: GTCTGTAAGGGCCTTGTCCAC PET 0.15 55 (30) 4-2 176-210 19 R: GTATCATCTATAAGGAGGCAGAGGC cisco-183 *** t F: TGGCTATATTCGACTACCTTG 6-FAM 0.6 55 (30) 4-2 161-292 46 R: CCCCATATATCAGAATGAGC * Turgeon et al. (1999) ** Patton et al. 1997 *** This paper t Poly-A tailed primer $ Locus not used for scale samples Table 3: Pairwise FST values between C. hoyi samples from Potential Donor Lakes (HUR: Lake Huron; MCH: Lake Michigan; SUP: Lake Superior; NIP: Lake Nipigon) and ciscoes from Lake Ontario (ONT-ARTEDI, ONT-DEEP). FST values significantly différent from 0 (Bonferroni corrected) are in bold.
ONT- ONT- ARTEDI DEEP HUR-A HUR-B HUR-C MCH-A MCH-B MCH-C SUP-A SUP-B SUP-C NIP-A NIP-B ONT-DEEP 0.062 HUR-A 0.033 0.064 HUR-B 0.035 0.063 -0.008 HUR-C 0.033 0.068 0.001 -0.009 MCH-A 0.038 0.055 0.001 0.000 -0.002 MCH-B 0.052 0.075 0.005 -0.003 0.002 -0.004 MCH-C 0.032 0.051 -0.007 -0.006 -0.016 -0.011 -0.005 SUP-A 0.044 0.123 0.021 0.018 0.030 0.038 0.038 0.034 SUP-B 0.049 0.124 0.017 0.006 0.020 0.031 0.027 0.023 0.001 SUP-C 0.052 0.129 0.027 0.021 0.033 0.041 0.044 0.035 0.004 0.003 NIP-A 0.098 0.138 0.053 0.042 0.048 0.058 0.049 0.046 0.042 0.016 0.047 NIP-B 0.089 0.140 0.038 0.035 0.034 0.047 0.039 0.039 0.035 0.010 0.040 0.000 NIP-C 0.085 0.119 0.036 0.034 0.028 0.039 0.029 0.031 0.031 0.002 0.032 0.001 -0.004
oc Table 4: Average probability (LnP(D)) and standard déviation across 20 runs of STRUCTURE for K genetic partitions among bloaters from ail Potential Donor Lakes.
K Average LnP(D) Standard LnP(D) déviation
1 -21393 10.13
2 -21010 11.80
3 -20827 47.68
4 -20929 233.97
5 -20919 177.32
6 -21243 618.91
7 -21738 1109.34
H -21798 950.64
9 -23012 3428.36
39 Table 5: Assignment of individual ciscoes from Lake Ontario a) to three genetic partitions defined by STRUCTURE - see Results-, with mean coefficient of membership (q) in parenthèses and b) to Potential Donor Lakes with GENECLASS, with average first score in parenthèses (HUR: Lake Huron; MCH: Lake Michigan; SUP: Lake Superior; NIP: Lake Nipigon)) a) Source PARTITION sample NIP SUP HURMCH
NIP 120 (0.90) 2 (0.96) 3 (0.65)
SUP 17 (0.78) 93 (0.89) 22 (0.69)
MCH 15 (0.68) 22 (0.77) 110(0.80)
HUR 8 (0.72) 28 (0.87) 110(0.82)
ONT-ARTEDI 0 32 (0.80) 19 (0.78)
ONT-DEEP 0 0 4 (0.83) b) Source sample NIP SUP HUR MCH
NIP 118(0.96) 3 (0.85) 3 (0.65) 1 (0.37)
SUP 3 (0.72) 115 (0.90) 12 (0.76) 13 (0.58)
HUR 3 (0.68) 5 (0.68) 114 (0.83) 17 (0.68)
MCH 0 9 (0.69) 17 (0.69) 116 (0.84)
ONT-ARTEDI 8 (0.66) 5 (0.64) 27 (0.72) 11 (0.73)
ONT-DEEP 0 0 2 (0.90) 3 (0.63)
40 Figure 1: Sampling locations for C. hoyi in Potential Donor Lakes (•) and for ciscoes in Lake Ontario (•: contemporary and historical scale samples of C. artedi; ~*" : deepwater ciscoes caught in 2002).
Lake Nipigon 250km
NIP-C
Lake Superior
OtïT-ARTEDI ONT-SCALES
Lake Michigan
OKY-DEEP
Lake Ontario Lake Erie Figure 2: Temporal variation in allelic richness (NAR) and demographical index (M ratio) in historical ciscoes from Lake Ontario (Hay Bay, Canada) in relation with relative commercial ciscoes catch in the Canadian waters of Lake Ontario (grey bars; Baldwin et al. 2005).
11.0 y 0.90
10.0 -- 0.80 - 0.70 9.0 -I - 0.60 i 8.0 - 0.50 • -s o 7.0 3 - 0.40 I
6.0 0.30
- 0.20 5.0- - 0.10 4.0 0.00 1965 1970 1975 1980 1985 1990 1995 2000 2005 2010 Year
4 Kl Figure 3: Unrooted DcE-based neighbor-joining phenogram relating ail samples from Potential Donor Lakes and Lake Ontario ciscoes (ONT-ARTEDI: récent sample of C. artedi; ONT-DEEP: deepwater ciscoes caught in 2002). Bootstrap values above 60 are shown.
SUP-C SUP.A
NIP-C
NIP-A
ONT-ARTEDI NIP-B
ONT-DEEP HUR-C HUR-A
MCH-A
43 ANNEXE A : Nombre d'individus (N), nombre d'alleles (A), hetérozygotie observée (Ho) et attendue (He) par locus et multilocus pour chaque échantillon
Echantillon Locus Multilocus
Code c90 bwO c59 cl81 c200 clO6 cl26 cl57 cl79 cl83 N A H» He ONT-ARTEDI N 49 48 48 44 48 42 50 49 46 46 51 12.8 0.7 0.71 A 3 5 14 23 21 21 3 10 7 21 Ho 0.61 0.54 0.82 0.79 0.89 0.92 0.06 0.75 0.79 0.90
He 0.63 0.58 0.90 0.64 0.90 0.93 0.06 0.76 0.87 0.78 ONT-DEEP N 5 5 5 5 5 5 5 5 5 4 5 4.5 0.64 0.59 A 2 1 4 6 5 7 2 5 6 7
Ho 0.32 0.00 0.66 0.70 0.76 0.84 0.32 0.68 0.78 0.84
He 0.40 0.00 0.60 0.80 0.80 0.80 0.40 0.80 0.80 1.00 1997 N 27 28 25 28 26 - 28 28 25 - 28 10.8 0.65 0.67 A 4 4 13 30 14 - 2 11 9 -
HQ 0.62 0.48 0.84 0.95 0.71 - 0.22 0.82 0.74 -
He 0.56 0.46 0.96 0.79 0.54 - 0.25 0.82 0.80 - 1994 N 21 22 19 19 22 - 20 19 19 - 22 9.8 0.57 0.69 A 3 6 11 19 18 - 3 8 10 -
Ho 0.37 0.69 0.80 0.92 0.91 - 0.14 0.80 0.84 -
He 0.19 0.73 0.74 0.58 0.86 - 0.05 0.74 0.68 - 1991 N 24 23 11 21 22 - 24 23 19 - 24 8.3 0.66 0.68 A 3 7 6 17 12 - 2 8 11 -
Ho 0.57 0.63 0.73 0.79 0.86 - 0.28 0.75 0.85 -
He 0.54 0.52 0.73 0.62 0.82 0.33 0.78 0.95 _ Echantillon Locus Multilocus
Code c90 bw£2 c59 cl81 c200 clO6 cl26 cl57 cl 79 cl83 N A H» He 1986 N 16 16 9 12 12 - 15 15 15 - 17 6.4 0.61 0.65 A 3 5 6 n 10 - 2 6 8 - Ho 0.57 0.46 0.77 0.87 0.86 - 0.23 0.77 0.69 -
He 0.75 0.50 0.33 0.67 0.75 - 0.27 0.87 0.73 - 1978 N 27 24 4 15 9 - 28 27 18 - 28 6.9 0.63 0.68 A 3 5 4 14 9 - 2 11 7 - Ho 0.63 0.63 0.66 0.91 0.85 - 0.24 0.85 0.69 -
He 0.81 0.63 0.50 0.73 0.56 - 0.29 0.85 0.67 - 1968 N 25 23 7 20 12 - 24 25 13 - 25 5.9 0.65 0.62 A 3 5 3 9 10 - 2 9 5 - Ho 0.60 0.66 0.62 0.68 0.88 - 0.15 0.83 0.54 -
He 0.92 0.65 0.57 0.55 1.00 - 0.17 0.88 0.46 - HUR-A N 47 45 39 46 47 39 45 43 43 38 47 14.3 0.69 0.75 A 4 6 17 28 21 24 3 10 8 22
Ho 0.46 0.64 0.83 0.93 0.89 0.94 0.32 0.81 0.74 0.92
He 0.43 0.58 0.82 0.65 0.85 0.85 0.31 0.88 0.72 0.84 HUR-B N 47 47 46 35 4^ 45 44 48 48 47 48 14.3 0.72 0.76 A 4 7 22 24 18 24 5 9 10 20
Ho 0.53 0.62 0.88 0.93 0.91 0.94 0.35 0.79 0.75 0.90
He 0.60 0.57 0.80 0.86 0.91 0.84 0.32 0.75 0.73 0.85 Echantillon Locus Multilocus
Code c90 bwf2 c59 cl81 c200 clO6 cl26 cl57 cl79 cl83 N A Ho He HUR-C N 50 50 49 49 50 41 48 44 42 37 51 16.2 0.71 0.77 A 6 7 27 30 19 24 4 10 11 24
Ho 0.52 0.68 0.90 0.95 0.91 0.94 0.29 0.80 0.80 0.93
He 0.52 0.58 0.82 0.80 0.90 0.83 0.25 0.75 0.74 0.89 MCH-A N 50 49 46 40 41 47 49 50 47 40 50 16.4 0.7 0.75 A 5 5 20 29 22 30 5 13 13 22
Hc 0.40 0.57 0.91 0.95 0.90 0.95 0.33 0.80 0.79 0.90
He 0.38 0.55 0.80 0.70 0.90 0.96 0.29 0.80 0.77 0.90 MCH-B N 48 47 38 37 40 37 45 43 37 38 48 13.7 0.71 0.76 A 5 4 20 22 18 24 4 9 9 22 Ho 0.40 0.61 0.89 0.93 0.89 0.93 0.48 0.78 0.76 0.93
He 0.35 0.55 0.74 0.78 0.93 0.89 0.58 0.67 0.70 0.87 MCH-C N 49 48 31 33 44 40 48 49 48 46 49 14.8 0.72 0.75 A 4 5 20 25 21 25 3 11 S 26 Ho 0.38 0.60 0.89 0.94 0.89 0.95 0.37 0.81 0.78 0.92
He 0.33 0.63 0.68 0.88 0.86 0.88 0.31 0.88 0.81 0.96 SUP-A N 50 46 37 49 38 39 à6 46 37 50 13.2 0.7 0.74 A 4 6 19 19 19 24 3 8 6 24
Ho 0.57 0.65 0.87 0.64 0.88 0.93 0.46 0.80 0.62 0.92
He 0.72 0.70 0.84 0.51 0.79 0.74 0.61 0.65 0.61 0.81 Echantillon Locus Multilocus
Code c90 bwG c59 cl81 c200 clO6 cl26 cl57 cl79 cl83 N A Ho He SUP-B N 34 33 23 34 27 28 27 32 31 19 34 12 0.74 0.76 A 4 4 15 19 19 20 4 10 7 18
Ho 0.57 0.66 0.84 0.78 0.92 0.90 0.51 0.84 0.66 0.92
He 0.85 0.61 0.61 0.56 0.93 0.64 0.78 0.94 0.84 0.68 SUP-C N 47 45 39 45 38 36 45 43 39 32 48 13.4 0.69 0.74 A 4 7 21 16 20 23 3 9 9 22 Ho 0.61 0.59 0.87 0.63 0.91 0.93 0.46 0.79 0.72 0.92
He 0.79 0.49 0.72 0.42 0.76 0.72 0.56 0.79 0.72 0.91 NIP-A N 52 52 47 50 47 47 52 53 49 41 53 15.3 0.75 0.78 A 4 7 21 24 22 25 3 12 9 26
Ho 0.61 0.56 0.91 0.93 0.92 0.90 0.53 0.83 0.69 0.91 He 0.52 0.63 0.96 0.82 0.87 0.70 0.56 0.83 0.76 0.80 NIP-B N 28 26 26 23 26 27 29 28 27 22 29 11.1 0.78 0.76 A 4 3 13 15 16 23 3 9 8 17 Ho 0.54 0.60 0.81 0.91 0.90 0.92 0.50 0.81 0.72 0.90
He 0.64 0.54 0.81 0.83 0.96 0.70 0.76 0.82 0.74 0.95 NIP-C N 41 40 39 3S 39 41 42 42 41 42 44 14.9 0.77 0.78 A 4 4 21 20 19 24 4 11 13 29
Ho 0.56 0.59 0.85 0.92 0.91 0.94 0.56 0.81 0.75 0.94
He 0.46 0.55 0.95 0.89 0.95 0.78 0.50 0.88 0.73 0.95 CHAPITRE 2 : INBREEDING DYNAMICS IN REINTRODUCED, AGE-STRUCTURED POPULATIONS OF HIGHLY FECUND SPECIES
Running title: Inbreeding dynamics in reintroduced populations
Names of author(s): Marie-Julie FAVÉ Pierre DUCHESNE Julie TURGEON* (correspondence author)
Institution address: Département de biologie Université Laval, Québec, Canada, G1K 7P4
Full mailing address : Département de biologie Université Laval, Québec, Québec, Canada G1K 7P4 Phone : 418-656-3135 Fax : 418-656-2043 Email : [email protected]
Keywords : reintroduction, inbreeding, breeding design, demography, Coregonus hoyi
48 Résumé
La réintroduction est une stratégie de gestion de populations utilisée dans le but de repeupler une région avec une espèce indigène maintenant localement disparue, en passant par une période de supplémentation. Lors d'une telle procédure, le type de croisements peut rapidement générer de la consanguinité au sein de la nouvelle population, celle-ci pouvant être subséquemment augmentée par le scénario de réintroduction utilisé. Nous avons d'abord simulé le degré de consanguinité (F) résultant de croisements alors que le nombre de géniteurs, le rapport des sexes, la proportion de croisements efficaces et le niveau de diversité génétique varient. Un nombre élevé de géniteurs, un rapport des sexes équilibré, une grande proportion de croisements efficaces et une population source génétiquement diversifiée contribuent généralement à des valeurs de F plus basses. Par contre, un nombre moyennement élevé (>= 20) de géniteurs combiné à un rapport des sexes équilibré ou modérément biaisé produisent des valeurs moyennes de F se rapprochant de la valeur minimale même avec des proportions de croisements efficaces aussi basses que 40%. La variabilité de F est négligeable pour toutes les combinaisons de paramètres, excepté avec un nombre très réduit de géniteurs (5) et un rapport des sexes très biaisé (< 1:19). Nous avons aussi exécuté des simulations pour évaluer la dynamique de la consanguinité à long terme dans une population réintroduite selon différents scénarios démographiques. Notre principale conclusion est que le nombre de juvéniles introduit chaque année est un facteur décisif dans l'établissement à long terme de la valeur de F dans la population réintroduite. Des niveaux réduits de supplémentation (102) génèrent rapidement une population complètement consanguine alors que des hauts niveaux de supplémentation (>104) produisent des valeurs de F stables et proches de celle des individus introduits. En suivant un régime de supplémentation en alternance, la valeur de F se rapproche de celle correspondant à la plus grande supplémentation. Les diverses simulations ont été basées sur l'histoire de vie et la démographie spécifique du cisco de fumage (Coregonus hoyi), dont la réintroduction dans le lac Ontario est envisagée.
Abstract
Reintroduction programs aim at reinstalling a self-sustained population into the wild via a period of supplémentation with captive-bred individuals. This procédure can rapidly generate inbreeding among offspring because of the mating scheme and this inbreeding
49 might be further enhanced by the reintroduction scénario. First, we used simulations to assess the conséquences of breeding designs on mean inbreeding index F among offspring when the genetic diversity of breeders, the number and sex ratios of breeders, and the proportion of successful crosses vary. A high number of breeders, a balanced sex ratio, a high proportion of effective crosses and a genetically diverse source population generally contribute to lower F values. However, moderately high (>= 20) numbers of breeders combined with ail but the most biased sex ratios produced mean F values near minimal even with proportions of effective crosses as low as 40%. The variability in F was negligible in ail parameter combinations except with a very small number of breeders (5) and very biased sex ratios (< 1:19). We also simulated the long-term inbreeding dynamics in the introduced population under various démographie scénarios. Our main finding was that the annual number of introduced offspring is a décisive factor in establishing long- term F values in the supplemented population. Low supplementation levels (102) quickly generated an almost completely inbred population whereas high levels (>104) produced stable F values close to that of the introduced offspring. With alternating supplementation régimes, F eventually reached levels close to that corresponding to the largest supplementation. Simulations were run based on the life history and spécifie demographics of the bloater (Coregonus hoyi), whose reintroduction in Lake Ontario is being considered.
50 Introduction
Anthropogenic influences such as overexploitation, physical habitat dégradation, pollution and exotic species introductions hâve triggered population déclines and range contractions in many species (Hills et al. 1999, Channell and Lomolino 2000, Palumbi 2001, Balmford et al. 2003). In such cases, captive breeding of animais is often considered as a managing strategy to increase the reproductive success and survival probabilities of endangered species. Even if populations can reach relatively high abundances in captivity, reintroduction in natural environment must remain the ultimate long term objective for the majority of species (Frankham et al. 2002). Reintroduction programs are labor-intensive management stratégies whereby adult individuals are taken in the wild, artificially bred, and the resulting offspring are released in an area formerly occupied by the species. Several species that had been driven to the edge of extinction are now on their way to wild résurgence because of reintroduction programs, such as the Arabian oryx (Marshall and Spalton 2000), the Mauna Kea silversword (Friar et al. 2000), the Mauritius kestrel (Groombridge et al. 2000), and the Florida panther (Pimm et al. 2006).
Reintroduction generally implies the foundation of a new population with the offspring of a limited number of breeders and this will rapidly induce mating among relatives, thus increasing inbreeding levels in the introduced population. As long as such a population remains small and isolated, it will continue facing genetic threats leading to further inbreeding. In addition, genetic drift will likely cause a loss of genetic diversity allowing deleterious mutations to be expressed while the small population size will tamper the effectiveness of sélection (Lande 1995, Lynch et al. 1995, Frankham et al. 2002). Inbred individuals also often suffer reduced survival when facing an environmental challenge (Keller et al. 1994, Saccheri et al. 1998, Sherwin et al. 2000), thus jeopardizing the success of the reintroduction program.
Inbreeding can be difficult to detect and its influence on fitness varies among species (Hedrick and Kalinowski 2000), but maximum avoidance of this phenomenon should be central in any captive breeding and/or reintroduction programs (Frankham et al. 2002). This goal can be attained by simultaneously giving due considération to the choice of breeders, the breeding strategy, and to key factors affecting the species demography. Ideally, unrelated breeders should be obtained from a genetically diverse population for
51 which no serious bottlenecks are suspected (Chapter 1). This guarantees a diversified gène pool to choose from for artificial breeding. Then, the breeding strategy must be designed to minimize the loss of available diversity (and the concomitant increase in identity-by- descent). This objective implies paying attention to both the number of captive breeders and the mating scheme employed (Waples and Do 1994, Duchesne and Bernatchez 2002, Fiumera et al. 2004). For example, Fiumera et al. (2004) hâve shown that complète factorial designs, where the total egg production of each female is divided in such a way that each part is fertilized by a différent maie, maximize the offspring production as well as the population effective size. Finally, the duration and extent of the reintroduction efforts, in interaction with characteristics intrinsic to the species population dynamics and the environment, will ultimately détermine the resulting level of genetic diversity in the new population. Certainly, fecundity, âge at maturity, longevity, survival probability, density- dependence interactions and carrying capacity of the habitat are key factors influencing the évolution of the population over the years.
Hère, we use the proposed reintroduction of the bloater (Coregonus hoyi) in Lake Ontario as a case study to investigate the genetic conséquences of a reintroduction program. Once abundant, this fish species was driven to extinction in Lake Ontario by 1982, and its reintroduction has been identified as a restoration priority for a healthy Great Lakes ecosystem (Baldwin 1999, Stewart et al. 1999, Eshenroder and Krueger 2002). Using empirical and/or realistic parameter values pertaining to this particular case, we appraise the associated genetic risks by performing computer simulations investigating the évolution of inbreeding over time. We chose to simulate the évolution of the inbreeding coefficient F, i.e. the probability of randomly choosing two identical alleles from a gène pool (Crow and Kimura 1970) since F is the most basic parameter describing genetic diversity (Waples and Do 1994, Duchesne and Bernatchez 2002). While inspired from a spécifie conservation project, this model should be applicable to any age-structured species with overlapping générations for which both maies and females gamètes can be collected and artificially fertilized. Many fish, amphibians and invertebrate species fall in this category.
Materials and Methods
Breeding design and inbreeding coefficient
52 We used the symbolic calculator Maple V. 9.5 (Maplesoft) to calculate the inbreeding coefficient (F) resulting from various breeding schemes. Empirical estimâtes of allele frequencies at 10 microsatellite loci in C. hoyi were used as a gène pool to form virtual breeder génotypes. A gender was assigned to each breeder and a proportion of randomly chosen crosses among ail possible crosses were simulated. The inbreeding coefficient was computed independently for each locus over 200 distinct virtual reproductions to obtain mean F values by locus as well as over ail loci, along with variance. The resulting mean F value was used as the inbreeding coefficient of reintroduced hatchery-born juvéniles (FH) in the démographie model detailed below.
Four parameters having a potential impact on the offspring mean F value were investigated, namely the genetic diversity of breeders, the number and sex ratios of breeders, and the proportion of effective crosses. High and low diversity gène pools were represented by allelic frequencies documented in samples from the large Lake Michigan bloater stock and from the putative small relie population of deepwater ciscoes from Lake Ontario, respectively (Chapter 1). The number of breeders was set to 5, 20, and 50, while sex ratios were unbiased (i.e. M:F « 1:1), slightly (M:F ~ 1:4) or highly (M:F » 1:20) biased towards females. Thèse values encompass the highly skewed sex ratios observed in natural populations of bloaters (Brown 1970, Bunnell et al. 2006) and also correspond to common breeding practices wherein fewer maies than females are used in artificial fertilization trials. Finally, knowing that pooling gamètes prior to fertilization can lead to huge variance in reproductive success among crosses (Simon et al. 1986, Withler 1988, Brown et al. 2000), we included variation in the proportion of effective crosses (PEC) for ail conditions tested, including a complète factorial design, i.e. 100% successful crosses. Ail results were compared to the mean F values expected in an infinité Wright-Fisher population, where F is calculated, for each locus, as:
n 2 th F = ^Tifi where n = number of alleles and fi is the frequency of the i allele
Dynamics oj'inbreeding coefficient in a reintroduced population We used a modified version of an existing supplementation model (Duchesne and Bernatchez 2002) to investigate the évolution of Fw, the inbreeding coefficient within the reintroduced population. As per the original model, the System comprises two main
53 components i.e. a hatchery (H) and a wild (W) populations, the latter being the new, introduced population. However, given that this is a reintroduction scénario, wild breeders from the reintroduced population are not used to refresh the captive breeding stock (parameter R in Duchesne and Bernatchez 2002). Moreover, hatchery-born juvéniles introduced in the wild are given a constant inbreeding coefficient, FH. Note that, given a spécifie hatchery breeding design, FH is the expected F value of the offspring. The model also differs in that there are overlapping générations, such that ail breeders are pooled during reproduction, irrespective of âge class. Détails on récurrence équations employed to détermine the value of Fw over any number of years are provided in Appendix B.
The évolution of inbreeding over time was integrated in the démographie context of a population originating from reintroduced, hatchery-born juvéniles. The dynamics of F within the reintroduced population was simulated with Excel (Microsoft Office). The user- defined démographie parameters included longevity, number of age-classes, âge at maturity, survival probabilities over the lifespan, fecundity per female, proportion of fertilized females and sex ratio. The management parameters were the number of individuals introduced at any given year, the duration of the reintroduction program, and FH. Using values pertinent to bloaters, the wild component was therefore divided into nine age-classes: juvéniles aged 0 to 3 are non-reproductive while âges 4 to 8 years are breeders (Brown 1970, Scott and Crossman 1973). The survival probability of age-0 class was estimated at 1% from gênerai estimâtes commonly used in fishery management (Hilborn and Walters 1992). Over the typical bloater's lifespan, survival probabilities of age-class 1 to 8 were calculated from bloater trawl surveys reporting abundance for each age-class in Lake Michigan (Brown 1970). Once mature, the number of eggs produced per female was set to 10 000, a figure that is consistent with values reported in Scott and Crossman (1973) (1241 eggs/oz * average adult bloater weight of 7 oz = 8 700 eggs/female) and the mean number of eggs per female in Lake Michigan (11 000 eggs/female; Emery and Brown 1978). In order to account for density-dependence effects known to affect C. hoyi (Hilborn and Walters 1992, Bunnell et al. 2006), the total population fecundity varied as a function of the total population abundance. On the one hand, female fertility was gradually decreased to 25 eggs/female with increasing population size, reflecting a density effect often observed in massive spawning species (Hilborn and Walters 1992). On the other hand, the proportion of fertilized females varied between 0.1 at low abundance and 0.8 at high abundance to introduce reduced population growth at low abundances and so simulate
54 an Allée effect (Stephens and Sutherland 1999). Thèse values were selected to obtain a maximum abundance of age-3 individuals between 3xlO6 and 4xlO6. This maximum abundance was inferred from the maximum age-3 fish abundance estimated for Lake Michigan over the last 38 years (7xlO6 individuals) and the availability of suitable deep habitat in Lake Ontario relative to that in Lake Michigan. To this effect, we used a Great Lakes bathymétrie map and measured the surface area that spans the depth interval suitable for adult bloaters (40-120 m; Crowder and Crawford 1984) in both Lakes Michigan and Ontario. We estimated that this depth habitat in Lake Ontario was half that available in Lake Michigan, and we therefore inferred that the maximum abundance of age-3 fish in Lake Ontario was also half that of Lake Michigan (i.e. 3.5 x 106). Finally, as was recently proposed by Bunnell et al. (2006), the sex ratio cycled over a 30-years period between observed values for Lake Michigan, i.e. M : F oscillated between 0.2 and 0.5.
Using this démographie context, we examined the dynamics of F over a period of 50 years with 104 individuals introduced each year during 15 years as default parameters. We also investigate how F is affected by the number of individuals released each year (constant over the entire reintroduction program), the effect of variable number of individuals released among years, and the effect of the duration of the reintroduction program.
Results
Effects ofbreeding design on F H The influence of the proportion of effective crosses (PEC), the sex ratio and the total number of breeders originating from a genetically diverse or depauperate source on F is depicted in Figure 1. First of ail, a gênerai increase in FH is observed for any breeding plan when breeders are from a weakly diversified population (Figure 1 a, b, c vs. ld). For a given number of breeders and sex ratio, inbreeding decreases as the proportion of effective crosses increases. This effect is particularly strong with a small number of breeders (Figure la) and is less marked when more breeders are used (Figure lb, c). With 50 breeders involved in the reproduction, the effect of the variation in PEC is almost nonexistent
(Figure le). Also, there is a slight decrease in the variance of FH among reproductions as
PEC increases. For ail tested parameter combinations, the FH value stabilizes for PEC equal to or higher than 50%.
55 The sex ratio influences the value towards which FH converges as PEC increases (Figure 1). Unbalanced sex ratios generate higher FH values as well as higher variance. The addition of only a small number of maies induces a réduction in both the mean and variance of FH for a given PEC and number of breeders (e.g. Figure lb). For moderate to high number of breeders (20 and 50), an unbiased sex ratio leads to a relatively constant FH that approaches that expected in a Wright-Fisher population, and the variance among reproductions becomes very small as PEC increases.
Moderately high (>= 20) numbers of breeders combined with ail but the most biased sex ratios produced mean FH values near minimal even with proportions of effective crosses as low as 40%. The variability in FH was negligible in ail parameter combinations except with very small numbers of breeders (5) and very biased sex ratios (1:19).
Dynamics ofFwin the reintroduced population Population sizes obtained with baseline values start with a period of low abundance, followed by an exponential growth phase and then fluctuate around the proposed carrying capacity while experiencing cyclical dynamics (Figure 2a). When the first group of introduced individuals reaches maturity, Fw obviously takes on the value generated by the mating design utilized (FH). The inbreeding coefficient then stays about constant over the years, reflecting that of the introduced individuals (Figure 2b).
The number of individuals introduced each year has a profound impact on the early demography of the population as well as on the resulting levels of inbreeding (Figure 2). A small number of released individuals lengthen the low abundance phase that précèdes exponential growth. For example, when introducing 106 individuals, an abundance of 500 000 is reached after only 10 years while a similar abundance is reached after 24 years when only 102 individuals are introduced. The prolonged low abundance phase observed with such a low introduction effort is associated with a dramatic increase in the inbreeding level, which then approaches unity after only 10 years (Figure 2b). In comparison, 3 4 releasing 10 individuals générâtes maximal Fw values of about 0.28, and releasing 10 individuals or more produces Fw values (0.23) even below the value generated by the
artificial mating design (FH = 0.24). However, above a certain level of supplementation, the number of reintroduced individuals only affects demography and has little influence on
56 the inbreeding level. Indeed, the introduction of 104 or 106 individuals per year greatly increases breeders population size (Figure la), while inbreeding F values remain very similar to FH in both cases (Figure lb). Finally, it is worth mentioning that the effect of higher mortality rates of age-0 hatchery-born individuals on demography and inbreeding is similar to that of reducing the number of individuals introduced per year, i.e., a longer low abundance phase and an increased in the ultimate Fw value.
Interestingly, when alternating between years of high (104) and very low (102) numbers of introduced individuals, the ultimate Fw value still converges towards 0.26 as in the control case when 104 individuals are introduced. Similarly, when alternating between years of low (103) and very low (102) numbers of individuals introduced, the ultimate F value converges towards 0.38, approaching the Fw value generated when 103 individuals are introduced each year. Introducing a variation in the number of supplementation years does not influence the démographie pattern nor the Fw dynamics significantly, as long as the supplementation lasts long enough to allow a first génération of breeders to reproduce in the wild before stopping introducing individuals (results not shown).
Discussion
We assessed the impact of breeding design and reintroduction scheme on the levels and évolution of inbreeding in a reintroduced population of highly-fecund and age- structured species with overlapping générations. Simulations indicate that every step of this conservation practice influences the inbreeding level that will characterize a new, introduced population. First, the level of genetic diversity among breeders used to establish the population will détermine the baseline inbreeding level of the supplemented offspring (FH) as well as that of the new population (Fw). Thus, it is worth surveying available donor populations to sélect individuals with a diverse allelic array (Appendix A). Next to securing a high degree of genetic diversity among potential breeders, the breeding scheme is of utmost importance for the production of genetically diversified offspring. Balanced or moderately biased sex ratios among a reasonable number of breeders (e.g. > 20) allows for some proportion of ineffective crosses (up to 50%) while still yielding the lowest possible inbreeding level among offspring. In contrast, the combinations of some parameter values had négative synergetic effects that were at times drastic. For example, highly female- biased sex ratios inevitably promoted high inbreeding that could not be significantly
57 attenuated by the addition of female breeders to the system (higher NB) or by increasing the proportion of effective crosses.
The variety of factors resulting in biased sex ratios of breeders highlights the relevance of considering this parameter in the context of artificial breeding practices. Indeed, many artificial reproduction programs tend to increase progeny production by fertilizing many females with few maies in order to reduce efforts put into capturing wild adult maies. Sex ratios are thus often female-biased in artificial reproduction plans since managers are constrained to minimize effort to production costs (Hilborn and Walters 1992, Wedekind 2002). By applying this strategy, however, the réduction in logistical costs will often occur at the expense of genetic costs. Sex ratio of a breeding program can also be involuntary biased because of sampling difficulties and/or naturally skewed sex ratios (see Wedekind 2002 for a review). Hère again, it is probably advisable to deploy additional sampling efforts in the wild population to reduce the disparity in gender représentation among breeders. Finally, the effective sex ratio in a pool of breeders containing equal numbers of maies and females may nevertheless be skewed if some maies prove stérile or with low-quality sperm. Failed reproduction by unfit breeders raises variance in reproductive success among breeders as it reduces the proportion of effective crosses (PEC). Variance in reproductive success among artificially-bred individuals is reported for several species (e.g. Simon et al. 1986, Withler 1988, Brown et al. 2000). It is a classic theoretical cause of reduced genetic variation in any population (Hartl and Clark 1997) that is experimentally supported by low effective population sizes and genetic variation in harem lines of Drosophila melanogaster (Briton et al. 1994). Likewise, Fiumera et al. (2004) hâve suggested that complète factorial designs, where 100% of crosses are assumed to be effective (i.e. PEC = 100%) is the best breeding strategy to maximize effective population size. However, our simulations showed that a very high PEC may not be necessary to reduce inbreeding if the sex ratio is not excessively skewed. We conclude that paying attention to the apparent sex ratios in a breeding stock, which is easily monitored, is probably a simple strategy to avoid major inbreeding increases among offspring. Given that the addition of only a small number of maies to the reproductive individuals pool may significantly reduce FH, a rule-of-thumb is to use ail available mature maies.
In addition to the breeding scheme, the reintroduction scénario can also be designed to minimize inbreeding among individuals. Hère again, économie and logistical constraints
58 may easily shape the extent of the program and affect démographie and genetic outeomes. Care should primarily be given to the number of juvéniles released in the wild each year as new individuals hâve a large impact on Fw dynamics within the introduced population. For the studied case, the addition of only 100 individuals per year can generate an almost completely inbred population (F=0.95) in 10 years. As more individuals are released annually, the risk of generating high inbreeding is lowered and inbreeding tends towards that of an idéal population. Also, adding more individuals each year shrinks the low abundance phase of the exponential growth curve and therefore minimizes the duration of risks associated will small populations, such as Allée effects and sensitivity to catastrophes (Stephens and Sutherland 1999, Frankham et al. 2002). As for inbreeding per se, introducing a limited number of individuals does produce a dramatic inbreeding incrément. However, there is a threshold level above which releasing more individuals can no longer ameliorate the Fw value. Obviously, the levels of supplementation favorable to achieving the lowest possible inbreeding level (FH) are contingent upon the demography of the species of interest. While modeling and predicting population abundance is a major challenge, our model nevertheless broadly defines the magnitude of the supplementation effort necessary to avoid a largely inbred and probably unfit population.
Simulation results indicated that cases with variable supplementation levels among years, which is bound to happen in real life situations, are not necessarily disastrous. Indeed, if a year of relatively unsuccessful reintroduction occurs, either because of low effective supplementation level or low survival of juvéniles after reintroduction, the resulting increase in inbreeding level can be compensated by the release of higher numbers of juvéniles in a subséquent year. Years of low supplementation levels are almost inévitable as a resuit of the many opérations needed to release juvéniles in the wild. The impacts of sporadic failures accumulate if rapidly countered by increased efforts.
Our model offers guidance to décide on breeding designs and reintroduction plans for highly fecund species with age-structured populations. However, it does not account for the possible decrease in fertility and survival that higher inbreeding levels could cause as a resuit of low supplementation. This effect is demonstrated in several biological Systems (e.g. Keller et al. 1994, Saccheri et al. 1998, Sherwin et al. 2000) and should be kept in mind when predicting the dynamics of the population. Likewise, adaptation to captivity is a serious issue in captive breeding programs since the survival of individuals when
59 released in the natural environment can be severely affected (Frankham et al. 2002, Frankham 2005). A distinctive feature of this reintroduction System is that new captive breeders are obtained from a very large external population. This feature contributes to maintain minimal inbreeding as compared to supplementation régimes based on domestic strains or hatchery breeders obtained from the supplementation population itself (Waples and Do 1994, Duchesne and Bernatchez 2002).
Although the use of appropriate mating designs and supplementation levels can help prevent excessive inbreeding, the long term dynamics of any wild population remains under control of many random factors. Therefore, reintroduced populations should be closely monitored on a long-term basis in order to collect empirical data bearing on démographie fluctuations and genetic diversity.
Acknowledgments
This work was supported by a Great Lakes Fishery Commission grant to JT.
60 Figure legend
Figure 1: Inbreeding coefficient (F) generated by crosses with variable numbers of breeders (NB), M:F sex ratios (• ~ 1:1; • ~ 1:4; * ~ 1:20), and pools of breeders characterized by high (a, b, and c; F = 0.24) or weak genetic diversity (d; F = 0.41) as a function of the proportion of effective crosses (PEC). Dashed lines represent the F value for an infinité Wright-Fisher population of similar diversity. Bars represent variance among 200 simulations. Missing data points represents situations with unrealizable calculations because of excessively stringent conditions (a, PEC = 10; d, PEC = 10 and 20).
Figure 2: Simulated demography and inbreeding in a new population of bloaters created by supplementation during 15 years. A) Age-3 bloater population size b) breeders inbreeding coefficient (Fw) compared to that of introduced individuals (FH = 0.24; horizontal dashed line). In both graphs, the number of individuals supplemented each year is indicated by différent line types.
61 Figure 1: Inbreeding coefficient (FH) generated by crosses with variable numbers of breeders (NB), M:Fsex ratios (• ~ 1:1; • ~ 1:4; * « 1:20), and pools of breeders characterized by high (a, b, and c; F = 0.24) or weak gène tic diversity (d; F = 0.41) as a function of the proportion of effective crosses (PEC). Dashed lines represent the F value for an infinité Wright-Fisher population of similar diversity. Bars represent variance among 200 simulations. Missing data points represents situations with unrealizable calculations because of excessively stringent conditions (a, PEC = 10; d, PEC = 10 and 20).
a) NB = 0.45 " 0.40 0.35 0.30 0.25 -
b) 0.45 NR = 20 0.40
0.35 ï (1:19) 0.30 -• (5:15) 0.25 ^ (10:10)
o 0.45 c) o NB= 50 0.40 tu 0.35 (1:49) 1 0.30 ^ (15:35) 0.25 ™ (25:25) 0.20
0.60 0.55 NB = 0.50 (1:19) 0.45 .j. ••.N = 50 (25:25) 0.40 .1. • _•_ B 0.35 20 40 60 80 100 Percentage of effective crosses Figure 2: Simulated demography and inbreeding in a new population of bloaters created by supplementation during 15 years. a) Age-3 bloater population size b) breeders inbreeding coefficient (Fw) compared to that of introduced individuals (FH = 0.24; horizontal dashed line). In both graphs, the number of individuals supplemented each year is indicated by différent line types.
4.5 i a) Number of iudividuals introduced per year ct/5 4.01 o — 106 3.5 10 4 ...... 103 N 3.01 --102 "03 g 2.5 J
3 On 2.0 O 1.5 J
0.0 b) •| 1.00
0.80 o
0-60
0.40
.§ 0.20
CQ 0.00 0 10 20 30 40 50 60 Years following the first introduction year
63 ANNEXE B: Dérivation of récurrence équations to compute F values
The system comprises two main components i.e. a hatchery and a wild populations. At each génération, the hatchery produces juvénile spécimens which are introduced in the wild. The juvéniles from the hatchery hâve constant inbreeding coefficient FH over générations since it is provided with fresh breeders at each génération randomly sampled from a very large pool of breeders. The wild component is further divided into nine âge classes of spécimens, three juvénile (non-reproductive) and six reproductive. Reproduction is monoecious diploid. Ail breeders are pooled during reproduction, irrespective of âge class. First we dérive the récurrence équations for simpler Systems, namely one with discrète générations and one with three âge classes, one juvénile and two reproductive.
Discrète générations This model is a spécial case of the basic model described in Duchesne & Bernatchez (2002). The System of récurrence associated with the gênerai model (p.49) is:
2 2 FH(k) = (1-R) WF(H) + 2(1-R)R FHw(k-l) + R WF(W)
FHw(k)= (l-R)C WF(H) + (1-R)(1-C) FHW(k-l) + RC FHW(k-l) + R(l-C) WF(W) 2 2 Fw(k) = C WF(H) + 2C(1-C) FHW(k-l) + (1-C) WF(W) where H = hatchery population W = population of breeders C = proportion of spécimens contributed by H to W at time k R = proportion of spécimens contributed by W to H at time k For the sake of simplicity C and R are used in place of C(k) and R(k), respectively, and should not be mistaken for constants.
WF(X) = 1/(2 Nx(k-l)) + (1-1/(2 Nx(k-1))) Fx(k-1) for X = H, W
Nx = efficient size of breeder population X (Hartl and Clark 1997) Clearly within the spécifie supplementation process that we are considering R = 0 and since Fn(k) remains constant over générations, it will be denoted FH. Therefore the récurrence équations become:
FH(k) = FH 2 2 Fw(k) =C FH +2C(l-C)FHW(k-l) + (l-C) WF(W)
(A CFH + (l-C)FHw(k-l)
However FHw(l) = FH and so
FHW(k)= C FH + (1-C) FH
Consequently, the System for discrète générations reduces to:
1) FH(k) = FH 2 2 2) Fw(k) = [C + 2C(1-C)] FH + (1-C) WF(W)
Note that the équation FHw(k) = FH is superfluous since we are monitoring Fw(k) over générations.
Three âge classes This system comprises one non reproductive âge class, J, and two reproductive âge classes, Wl and W2. Together Wl and W2 make up W, the population of breeders. The proportions of Wl and W2 breeders within W are denoted pl(k-l) and p2(k-l), respectively. The three âge classes system is illustrated at the end of the appendix.
Again we assume that
1) FH(k) = FH If one picks two spécimens from J(k), then thèse may corne:
1) both from H with F = FH ii) one from H and the other from W(k) i.e. either from Wl(k-l) with probability pl(k-l) or from W2(k-1) with probability p2(k-l) iii) both may corne from W(k-l) as a resuit of the reproduction process resulting in
F - WF(W) = 1/(2 Nw(k-l)) + (1-1/(2 Nw(k-1))) Fw(k-1) Weighting each of the above three events by its probability of occurrence one gets: 2 2 2) Fj(k) = C FH + 2C(l-C)[pl(k-l)FHwl(k-l) + p2(k-l)FHw2(k-l)] + (1-C) WF(W) Ail âge classes, except J, inherit the same F value as in the previous génération:
3) FWi(k) = Fj(k-l)
4) FW2(k) = Fwi(k-l) If one takes a spécimen from H and a spécimen from J, then they either corne i) both from H with F = FH ii) one from H and the other from W(k-l) i.e. either from Wl(k-l) with probability pl(k-l) or from W2(k-1) with probability p2(k-l)
65 Given that C = proportion of J spécimens contributed by H we get after weighting:
5) FHJ(k)= C FH + (1-C) [pl(k-l)FHWi(k-l) + p2(k-l)FHw2(k-l)]
Fnwi(k) and Fnw2(k) are inherited from the previous génération:
6) FHwi(k) = FHj(k-l)
7) FHW2(k) = FHWi(k-l)
Since, as in the discrète générations System, FRj(k) = FHwi(k) = FHw2(k) = FH , we get:
1) FH(k) = FH 2 2 2) Fj(k) = [C + 2C(1-C)] FH + (1-C) WF(W) 3) Fwi(k) = Fj(k-l)
4) Fw2(k) = Fwi(k-l) where
WF(W) = 1/(2 Nw(k-1)) + (1-1/(2 Nw(k-1))) Fw(k-l) 2 2 Fw(k-l) = [pl(k-l)] Fwi(k-l) + 2 pl(k-l) p2(k-l) FwlW2(k-l) + [P2(k-1)] FW2(k-l) which may be approximated as
Fw(k-1) ~ pl(k-l) FWi(k-l) + P2(k-1) FW2(k-l)
Nine âge classes This System comprises three non reproductive âge classes, Jl, J2, J3, and six reproductive âge classes, Wl, W2, W3, W4, W5, W6. The totality of the reproductive âge classes, i.e. ail the breeders, are referred to as W. The proportions of Wl, W2, W3, W4, W5, W6 breeders within W are denoted pl(k-l), p2(k-l), p3(k-l), p4(k-l), p5(k-l), p6(k- 1), respectively.
Reasoning along the same lines as with the three âge classes System, we obtain the following set of récurrence équations:
FH(k) =FH 2 Fn(k) = C FH + 2C(l-C)[pl(k-l)FHWi(k-l) + p2(k-l)FHW2(k-l) + p3(k-l)FHw3(k-l) + 2 p4(k-l)FHW4(k-l) + p5(k-l)FHW5(k-l) + p6(k-l)FHW6(k-l)] + (1-C) WF(W)
FJ2(k) = Fj!(k-1), FJ3(k) = FJ2(k-l)
Fwl(k)= FJ3(k-l), FW2(k) = FWi(k-l), FW3(k) = FW2(k-l),
FW4(k)= FW3(k-l), FW5(k) - FW4(k-l), FW6(k) = Fw5(k-1), FHJi(k)= C FH + (1-C) [pl(k-l)FHwi(k-l) + p2(k-l)FHw2(k-l) + p3(k-l)FHW3(k-l) + p4(k- l)FHW4(k-l) + p5(k-l)FHW5(k-l) + p6(k-l)FHW6(k-l)]
FHJ2(k)=: FHJ1(k-l), FHJ3(k)= FHJ2(k-l), FHwi(k)= FHj3(k-l), FHW2(k)= FHWi(k-l), FHW3(k)=
FHw2(k-l), FHW4(k)= FHw3(k-l), FHW5(k)= FHW4(k-l), FHW6(k)= FHW5(k-l)
However assuming Fe(k) = FH leads to:
Therefore the nine âge classes (Jl, J2, J3,W1, W2, W3, W4, W5, W6) System reduces to:
1) FH(k) = FH 2 2 2) Fn(k) = [C + 2C(1-C)] FH + (1-C) WF(W)
3) FJ2(k) = F,i(k-1)
4) FJ3(k) = FJ2(k-l)
5) Fwi(k) = FJ3(k-l)
6) FW2(k) = Fwi(k-l)
7) FW3(k) = FW2(k-l)
8) FW4(k) = FW3k-l)
9) FW5(k) = FW4(k-l)
10) FW6(k) = Fw5(k-1) where
WF(W) = 1/(2 Nw(k-1)) + (1-1/(2 Nw(k-l))) Fw(k-1)
Fw(k-l) ~ pl(k-l) Fwi(k-l) + p2(k-l) FW2(k-l) + P3(k-1) Fw3(k-1) + p4(k-l) FW4(k-l)
+ p5(k-l) Fws(k-l) + p6(k-l) Fw6(k-1) C = proportion of Jl from H Nw(k-1) = effective size of W at génération k-1 Figure B.l: Supplementation system with three âge classes J, Wl, and W2
p2(k-1)
Ij
H W1(k) W2(k) i
Reproduction • Survival * Supplementation --*•
68 CONCLUSION GENERALE
Résumé de la recherche Dans l'optique d'une réintroduction du cisco de fumage dans le lac Ontario, les objectifs principaux de la présente recherche étaient de (1) identifier une source de cisco de fumage génétiquement diversifiée, exempte de signe de déclin important récent et montrant une ascendance partagée maximale avec les ciscos du lac Ontario et (2) examiner les conséquences du plan de croisement et de réintroduction sur la production de consanguinité chez les individus réintroduits et leur descendance. Les résultats de l'analyse des polymorphismes génétiques microsatellites démontrent qu'une grande diversité génétique est toujours présente dans les populations étudiées malgré des décennies d'exploitation commerciale intense. Aucun signe évident de déclin significatif affectant la diversité et la structure génétique des populations de cisco de fumage n'a été identifié bien que des fluctuations démographiques considérables aient été enregistrées tout au long du siècle dernier (Christie 1973, Baldwin et al. 2005). Ceci constitue une conclusion positive démontrant la capacité qu'ont certaines espèces à soutenir des niveaux d'exploitation élevés tout en conservant une grande diversité génétique. Bien entendu, cette capacité s'effrite lorsque l'intensité de l'exploitation dépasse les limites viables pour la population et une extinction locale ou globale devient alors inévitable, à l'exemple de plusieurs populations d'écomorphotypes de ciscos des Grands Lacs maintenant disparus (Christie 1973, COSEPAC 2006)).
La recherche de signature génétique témoignant d'une baisse importante de l'abondance des populations de ciscos a permis de tester l'efficacité de deux méthodes couramment utilisées en génétique de conservation, soit le test sur l'hétérozygotie (test HE) (Cornuet et Luikart 1996) et l'indice « M » de Garza et Williamson (2001). Notre faculté à détecter de tels phénomènes avec nos données a été validée suite à la détection d'un goulot d'étranglement dans une série temporelle comprenant des échantillons de ciscos du lac Ontario provenant d'une population où une baisse démographique sévère et prolongée a été reportée (Baldwin et al. 2005). Étant donné l'échec des deux méthodes à identifier les baisses d'abondance dans les autres populations de cisco de fumage, l'efficacité de ces dernières est relativisée et semble être fonction des caractéristiques biologiques de l'espèce, du niveau de diversité génétique précédent le déclin et du patron
69 général de celui-ci. Ces résultats renforcent l'idée de l'importance des données historiques dans les études de génétique de conservation. L'optimisation du protocole ayant servi à l'extraction de l'ADN et l'amplification de motifs microsatellites provenant d'écaillés de spécimens de musée âgés de près de 40 ans constitue une contribution positive de cette étude.
Chez les populations sources potentielles, le degré d'ascendance partagée entre les populations reflète la distribution géographique des populations, les individus provenant de lacs limitrophes montrent une plus grande similarité génétique entre eux (ex. Huron et Michigan) que des individus provenant de lacs éloignés (ex. Huron et Nipigon). La plus grande similarité entre les ciscos de deux taxons différents du lac Ontario qu'entre les ciscos du même taxon provenant de lacs différents apporte un certain support aux conclusions de Turgeon et Bernatchez (2003) proposant une diversification intralacustre des ciscos des Grands Lacs. Par ailleurs, les ciscos du lac Nipigon semblent former un groupe quelque peu différent des autres ciscos. Leur divergence génétique avec les ciscos des autres lacs est plus importante, plus particulièrement avec les lacs Ontario, Huron et Michigan. À cause de sa plus grande dissimilitude génétique avec les ciscos du lac Ontario et de l'identification d'une possible baisse d'abondance sévère dans le secteur Humboldt, le lac Nipigon devrait donc être écarté comme source de géniteurs en vue d'une réintroduction dans le lac Ontario. De leur côté, les lacs Huron et Michigan seraient des choix judicieux pour le rétablissement du cisco de fumage dans le lac Ontario à cause de leur grande diversité génétique, de l'absence de baisse récente importante dans la population ayant affecté la diversité génétique et de leur similarité avec les ciscos du lac Ontario. Le choix ultime de la source devra donc être établi en fonction de l'état de santé général des poissons et des capacités logistiques pour le transfert d'individus vers les installations d'élevage en captivité.
Les simulations de croisement ont établi que le nombre total d'individus fertilisés, le rapport des sexes ainsi que le pourcentage de croisements efficaces influencent la consanguinité des rejetons produits par la reproduction artificielle. Des situations extrêmes, telles qu'un plan de croisement avec un rapport des sexes fortement biaisé, un nombre total d'individus réduit ou des individus provenant d'une population montrant une faible diversité génétique ont été identifiées comme génératrices de hauts niveaux de
70 consanguinité. Par contre, un rapport des sexes modérément biaisé ainsi qu'un faible pourcentage de croisements efficaces peuvent rester sans effet significatif sur la consanguinité des rejetons en autant qu'un grand nombre de géniteurs soient inclus dans les croisements. Une fois la réintroduction initiale effectuée, le niveau d'ascendance partagée atteint après plusieurs générations dépend du nombre d'individus réintroduits à chaque année et de leur consanguinité. Un grand nombre d'individus relâché chaque année amènera une stabilité dans les niveaux de consanguinité qui tendront alors vers les valeurs théoriques attendues dans une population idéale infinie. D'autre part, un nombre très réduit aura pour effet de générer rapidement, après quelques générations seulement, une consanguinité quasi complète dans la population. Ces simulations ont permis d'assigner des valeurs concrètes d'individus relâchés à des niveaux de consanguinité à l'intérieur d'une population.
Limites et perspectives Comme toutes les recherches, cette étude comporte des limites sur quelques uns de ses aspects. Un nombre plus élevé de ciscos provenant des couches profondes du lac Ontario aurait apporté un bénéfice considérable aux analyses de similarité entre les populations. D'un autre côté, récolter un grand nombre de spécimens provenant d'une population probablement très réduite et vulnérable n'entre vraisemblablement pas dans les objectifs d'un plan de conservation et d'échantillonnage non invasif. Puisqu'en définitive, c'est le rétablissement du phénotype original du cisco de fumage qui est recherché, il aurait été intéressant de coupler les données génétiques à des mesures de variation des caractères morphologiques des ciscos provenant des différents lacs. Cela aurait permis d'identifier les sources de ciscos ayant la plus grande ressemblance morphologique avec les ciscos récoltés dans les couches profondes du lac Ontario. Le modèle de réintroduction développé se base sur les meilleurs paramètres biologiques que nous ayons pu trouver ou évaluer mais il importe de spécifier que plusieurs sont des approximations ou des données acceptées dans la littérature pour d'autres espèces voisines. Une meilleure connaissance de l'histoire de vie du cisco de fumage aurait fourni une plus grande exactitude au modèle et par conséquent, une capacité supérieure à modéliser la dynamique de la population.
Malgré ses limites, cette étude est un cas tangible de génétique de conservation qui guidera la réintroduction du cisco de fumage dans le lac Ontario. Cette étape initialisera le
71 projet de rétablissement de la faune ichthyenne native des Grands Lacs Laurentiens, dans lequel plusieurs autres écomorphotypes de ciscos sont ciblés, notamment C. zenithicus et C. kiyi. Cet objectif binational est d'une part original dans l'idée de rétablir des espèces ichthyennes puisque les plans de réintroduction à des fins non lucratives sont en grande partie dirigés vers les gros mammifères et les oiseaux, les poissons étant fortement sous- représentés (Seddon et al. 2005). D'autre part, le succès de ce projet apportera des bénéfices directs aux communautés locales, tels qu'une pêche sportive au touladi plus vigoureuse et, à plus long terme, l'éventuel rétablissement d'une exploitation commerciale du cisco de fumage (Baldwin 1999). Les facteurs génétiques étant reconnus comme devant faire partie de tout plan de gestion de population (Frankham et al. 2002), le modèle mathématique développé pour exécuter les simulations pourra servir d'outil aux gestionnaires des populations non seulement de poissons, mais également pour d'autres espèces hautement fécondes tels que les amphibiens, les insectes ou divers invertébrés.
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