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Genetic variation, population structure and mating system in bigleaf maple (acer macrophyllum pursh) Iddrisu, Mohammed Nurudeen 2005

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G E N E T I C V A R I A T I O N , P O P U L A T I O N S T R U C T U R E A N D M A T I N G S Y S T E M IN B I G L E A F M A P L E {ACER MACROPHYLLUM P U R S H ) by M O H A M M E D N U R U D E E N IDDRISU Ing. For. University of P inar del R io , 1993. A T H E S I S S U B M I T T E D IN P A R T I A L F U L F I L M E N T O F T H E R E Q U I R E M E N T S F O R T H E D E G R E E O F D O C T O R O F P H I L O S O P H Y in T H E F A C U L T Y O F G R A D U A T E S T U D I E S (Forestry) T H E U N I V E R S I T Y O F BRIT ISH C O L U M B I A May, 2005 © Mohammed Nurudeen Iddrisu, 2005 ABSTRACT Ecolog ica l characterist ics and life history traits of long lived woody plants influence their levels of genet ic variation. To embark upon sound management , utilization and conservat ion of plant spec ies , a thorough understanding of genet ic p rocesses affecting their pers istence is essent ia l . In this thesis, I studied genet ic diversity, population structure, and mating system as well as compared genetic diversity and inferred dif ferences in genetic p rocesses in cont inuous versus fragmented populat ions of bigleaf maple (Acer macrophyllum Pursh). Bigleaf maple is one of the most abundant hardwood spec ies in the Paci f ic Northwest and its native range extends from latitude 3 3 ° N to 5 1 ° N along the Paci f ic coast of North Amer i ca . Genet ic diversity, est imated using isozyme markers, revealed a mean expected heterozygosity (H E ) of 0 . 1 5 2 similar to other North Amer ican angiosperm trees. The level of population differentiation was moderately low ( F S T = 0 . 0 5 4 ) , indicating extensive gene flow among populat ions. Est imated outcrossing rates in two populat ions were high ( 9 5 % ) but significantly less than one, with no biparental inbreeding evident. A relatively high level of correlated matings was found, consistent with 2 - 5 effective pollen donors per tree, indicating low adult density and limited pollinator d ispersal . Seedl ing and adult populat ions p o s s e s s similar levels of genet ic variation regardless of whether populat ions are fragmented or cont inuous. However, seedl ing cohorts have higher levels of inbreeding than adult cohorts, on average, in both cont inuous and fragmented populat ions. Ana lys is of spatial genetic structure indicates non-random distribution of genotypes in all three fragmented populat ions and one of the three cont inuous populat ions. I found a significant positive autocorrelation (p/,= 0 . 2 0 ) among individuals located up to 1 0 0 m apart in all three f ragmented populat ions and among individuals located at approximately 1 0 0 - 2 0 0 m apart (p,y = 0 . 1 4 ) in one of three cont inuous populations. Finally, for quantitative traits, p rovenances and famil ies within provenances showed significant genet ic variation for height growth and bud flush traits, but not for diameter growth. Individual heritabilities for all traits were general ly low to moderate ( 0 . 1 5 - 0 . 2 1 ) , and family heritability was higher only for bud f lush. Compar i son of Q S T and F S T in this study (mean Q S T = 0 . 1 7 > mean F S T = 0 . 0 9 ) suggests the involvement of select ion for different phenotypes in different populat ions of bigleaf maple. TABLE OF CONTENTS Abstract ii Table of Contents iv List of Tables viii List of Figures xi List of Appendices xii Acknowledgements xiii Dedication xv Published papers xvi Chapter One Genera l Introduction 1 Thes is overview 2 Chapter Two Literature Rev iew 4 Biology and silvics of Acer macrophyllum Pu rsh 4 Genet ic variation and structure in natural populat ions 5 Effects of population s ize on genet ic variation 6 Effects of population s ize on mating sys tems 8 Effects of fragmentation on genet ic variation in plant populations 10 Effects of fragmentation on spatial genet ic structure 13 Molecular and quantitative variation 14 Chapter Three Genet ic variation, population structure and mating sys tem in bigleaf maple [Acer macrophyllum) 19 Introduction 19 Materials and methods 20 Isozyme assay 21 Data analysis 22 Resul ts 24 Al le le f requency distribution 24 Genet ic diversity 24 Genet ic structure 25 Mating system 26 Discuss ion 27 Genet ic variation 27 Populat ion genetic structure and gene flow 27 Mating system ... .29 Implications for management and conservat ion 31 Chapter four Effects of forest fragmentation on genet ic variation and spatial genetic structure in natural populations of bigleaf maple (Acer macrophyllum) 41 Introduction 41 Materials and methods 44 Populat ions and sampl ing 44 Electrophoresis 45 Data analysis 45 Genet ic structure 46 Spatial autocorrelation analys is 46 Simulat ions 49 Resul ts 50 Al le le f requencies 50 Genet ic diversity 50 Levels of inbreeding 51 Bott leneck test 51 Genet ic structure 51 Spatial genetic structure 52 Simulat ions 53 D iscuss ion 54 Effects of fragmentation on genet ic variation and inbreeding 54 Inbreeding in adults versus seedl ings 56 Populat ions structure 57 Spatial genetic structure 58 Computer simulat ions of fragmentation effects 60 Chapter five Genet ic variation and population structure in bigleaf maple: a compar ison of a l lozyme markers and quantitative traits 74 Introduction 74 Materials and methods 76 Quantitative traits 76 Data col lect ion.. 77 Ana lys is 77 Isozyme variation 79 Resul ts 80 Quantitative traits 80 Molecular genetic variability 81 Discuss ion 8 2 Quantitative traits 8 2 Bud flush 8 3 Genet ic correlations 8 4 Correlat ions with climatic var iables 8 4 F S T V S Q S T 8 5 Chapter 6 Conc lus ions 9 5 Major f indings 9 6 Recommendat ions 9 8 References 100 LIST OF TABLES 3.1. Distribution of allele f requencies at 10 loci in eight natural mature populations of bigleaf maple (Acer macrophyllum) 34 3.2. Summary of genetic diversity within eight mature natural populat ions of bigleaf maple (Acer macrophyllum) based on 10 a l lozyme loci 35 3.3. Total gene diversity (HT), genetic diversity within populat ions (Hs), expected heterozygosity (H0), al leles per locus (NA), fixation index over the total populations (FIT), fixation index within population (F/s), and genetic differentiation among populations (FST) for eight mature natural populat ions of bigleaf maple (Acer macrophyllum) at nine polymorphic loci 36 3.4. Est imates of multi- locus outcrossing rates (tm), s ingle- locus outcrossing rates (ts), biparental inbreeding (tm-ts), parental inbreeding coefficients (F) and correlation of paternity among sibl ings (rp) 37 3.5. Compar i son of within-population genetic diversity for Acer macrophyllum with average va lues for all plants, woody spec ies , woody ang iosperms, and for maple spec ies 38 4.1. Summary of population information for adult trees and seedl ings of bigleaf maple Acer macrophyllum 62 4.2 a. Al le le f requencies for nine loci for adults in cont inuous and fragmented populations of Acer macrophyllum 63 4.2 b. Al le le f requencies for nine loci studied for seedl ings in cont inuous and fragmented populations of Acer macrophyllum 64 4.3. Genet ic diversity est imates for adults and seedl ings in cont inuous and fragmented populations of Acer macrophyllum 65 4.4. Wi lcoxon s igned ranked test for recent bottleneck (Cornuet and Luikart 1996) in Acer macrophyllum populations under the Infinite Al le les Mode l 66 4.5 a & b. Genet ic diversity statistics for the eight polymorphic isozyme loci for (a) cont inuous populations and (b) f ragmented populat ions 67 4.6. Pairwise FST between adult f ragmented and cont inuous populat ions of Acer macrophyllum 68 4.7. Expected percentage of a l lozyme diversity retained over 250-year period based on computer simulat ions B O T T L E S I M (Kuo and J a n z e n 2003) for adult populations of Acer macrophyllum in f ragmented and continuous forests assuming 125-year generat ion length 69 5.1. Locat ions of bigleaf maple sampled populat ions for provenance trials and least square means for growth and bud f lush traits 88 5.2. A N O V A results for F approximations for the hypothesis of no family or provenance effect 89 5.3. Componen ts of var iance, individual heritabilities (h 2i), family heritabilities (h2f) and population differentiation (Q S r ) among growth and bud f lush traits 90 5.4. Genet ic correlations (above diagonal) and family phenotypic correlations (below diagonal) between seedl ing traits for bigleaf maple provenances in British Co lumb ia 91 5.5. Correlat ion coefficients between quantitative traits and cl imatic var iables based on 14 provenance means 91 5.6. Genet ic diversity est imates for 14 juvenile populat ions of Acer macrophyllum 92 5.7. Est imates of Wright 's F-statistics for eight polymorphic loci in British Co lumb ia bigleaf maple populations 93 LIST OF FIGURES 2.1. Native range of Acer macrophyllum (bigleaf maple) 18 3.1. Geograph ica l locations of eight Acer macrophyllum mature populat ions natural populat ions 39 3.2. U P G M A cluster analys is of Nei 's genetic d is tances between eight mature populat ions of Acer macrophyllum 40 4.1. Geograph ica l locations of sampled bigleaf maple populat ions 70 4.2. Distribution of allele f requencies for adults (a) and seedl ing (b). Fil led bars are cont inuous populations and open bars f ragmented populat ions 71 4.3 (a-c). Spat ia l correlograms of coancestry coefficients (p,y) for cont inuous populat ions of Acer macrophyllum. Dashed l ines represent upper and lower 9 5 % conf idence limits for p,y under the null hypothesis that genotypes are randomly distributed 72 4.3 (d-f). Spat ia l correlograms of coancestry coefficients (p,y) for f ragmented populat ions of Acer macrophyllum. Dashed l ines represent upper and lower 9 5 % conf idence limits for p,y under the null hypothesis that genotypes are randomly distributed 73 5.1. Locat ions of sampled populations of bigleaf maple provenance trials 94 LIST OF APPENDICES I. Enzyme , buffer sys tems and recipes for h istochemical staining solut ions 129 II. Al le le f requency distribution of ten loci of bigleaf maple provenance trials 130 ACKNOWLEDGEMENTS I would first like to acknowledge with deep appreciat ion the Department of Foreign Affairs and International Trade for the Award of Canad ian Commonwea l th Scholarsh ip through the International Counc i l for Canad ian Studies. Funding for research was made avai lable initially from the B C Ministry of Forests, Resea rch Branch through Dr. C h e n g Y ing and compl imented by a Natural Sc i ences and Engineer ing Resea rch Counc i l ( N S E R C ) grant to Dr. Kermit Rit land. This study would not have been completed without the much needed fel lowship and additional funding for research provided to me by my co-superv isor Dr. Sa l ly Ai tken through the Centre for Forest G e n e Conservat ion via the Forest Genet ics Counc i l of B C from the Forest investment Accoun t of B C and the N S E R C Industry Junior Chai r in Populat ion Genet ics . I would like to thank my co-supervisors Drs Sal ly Ai tken and Kermit Rit land and committee member Dr. Jeannette Whitton for their gu idance, support and constructive comments . Spec ia l thanks go to Dr. Sal ly Ai tken who spent an extra time on my draft, chal lenging me to write concisely and encouraging me to think critically and realistically. My s incere thanks also go to Dr. Caro l Rit land for her initial involvement in my committee, for providing fresh perspect ive on my research during the initial s tages, planning my field trips and supervis ing my lab work. To Dr. C h e n g Y ing , thanks very much for serving on my committee, for the guidance and numerous fruitful comments and suggest ions on my research, for your fr iendship and fatherly advice and for providing the quantitative data and encouraging me to work on bigleaf maple. To Mr. Don Pigott for sharing your expert ise with me and the great help in collection of samp les in the field. I wish to a lso thank the following: - My co l leagues, Cherdsak L iewlaksaneeyanawin , Char les C h e n , Yan ik Berube and Hugh We l lman . - Drs. Tongl i W a n g , Andreas Hamann for their help and adv ice and P ia Smets for being so opened to d iscuss ing issues beyond academics . - My wife Yanela and children Leandro and Neina for their enormous sacr i f ices and support throughout the duration of my program. - To my mother, brothers and sister and my extended family members for their encouragement and advice. DEDICATION To the memory of my late father- M b a Iddi (Gushe i -Naa) "After great pain a formal feeling comes" Emily Dickinson Chapter One G E N E R A L INTRODUCTION Forests are decl ining in most regions world-wide, and this has caused grave concern among scientists and policy makers throughout our world. Much of the world's biodiversity is harbored in forests, particularly tropical forests which are est imated to contain up to 70 percent of the world's spec ies (Groom 1994). However, over the last two centuries, the exponential growth of human populat ions coupled with growth of cities, industrialization and agriculture has led to widespread destruction and degradat ion of many forested and other natural sys tems ( F A O 1997). Approximately half of the world's forest area has been c leared or degraded s ince the beginning of the Holocene (Groombridge and Jenk ins 2002). Currently, about 30 percent of the world's land area is covered with forest ( F A O 2001). Many of these forests are partially converted to agricultural or urban use, resulting in the loss of s o m e unique characterist ics that were previously present. The main goal of population genet ics is to understand the origin, distribution and maintenance of genetic diversity, which is the raw material for evolutionary change (Ledig 1992; Hartl and Clark 1997). In smal l populat ions in particular, genes undergo genet ic drift and as a consequence genet ic diversity is randomly and continuously lost (Vitalis and Couvet 2001). Moreover, genet ic drift in smal l populat ions can be more important than select ion in determining the fate of new al leles (Whitlock 2000). In subdiv ided populat ions, the maintenance of neutral al leles depends on the relative strength of local genetic drift and the extent of gene flow as a homogeniz ing force (Slatkin 1995). C h a n g e s in genet ic structure and levels of diversity in subdiv ided or smal l populat ions of forest tree spec ies will depend on severa l factors, such as the magnitude and frequency of forest destruction and degree of isolation among fragmented forests (Bawa 1994). Forest management pract ices also affect genetic diversity (Savola inen and Karrkainen 1992; E l -Kassaby and Namkoong 1994). Harvest ing can lead to a reduction in stand density, which may result in increased levels of inbreeding and a decl ine in genetic diversity (Murawski and Hamrick 1992; Buchert et al. 1997). Knowledge of mating sys tems, the levels and distribution of genet ic variation, and factors influencing its maintenance is necessary for effective forest management and conservat ion programs (Er iksson et a l . 1995). In recent years, there has been an accumulat ion of data concerning the patterns of genet ic variation in many North Amer ican coniferous tree spec ies . However, relatively little information is avai lable about genetic variation in temperate woody ang iosperms. This situation is particularly true for intolerant, early success iona l and shrubby trees, the majority of which are non-commercia l and suffer s o m e degree of habitat fragmentation. To improve our understanding of the genet ic structure of forest trees, it is necessary to broaden the scope of study to include ang iosperms with different mating sys tems, pollination vectors, patterns of d ispersal , and evolutionary histories. Life history and ecological factors that would promote genetic diversity of woody early success iona l spec ies , like Acer macrophyllum (bigleaf maple), are likely to be similar to those of relatively long-lived spec ies . Thesis overview In order to embark on any useful conservat ion program for any spec ies , knowledge of how genet ic variation is partitioned among and within populat ions is the first necessary step. The main goals of chapter two are: 1) to review the literature on the biology of Acer macrophyllum; 2) to review the bas ic ideas that shape our thinking about genetic diversity and mating patterns in smal l and subdiv ided populat ions; 3) to review the role of evolutionary forces in explaining genetic differentiation for neutral genet ic markers and quantitative traits; and 4) to address genetic consequences of forest fragmentation and spatial genet ic structure in natural populations. In chapter three I use isozyme markers to investigate the genet ic variation, population structure and mating sys tem in bigleaf maple. This is an important bas ic step to understanding the population genet ics of this spec ies . I hypothesized that fragmentation may lead to erosion of genet ic variation. In chapter four this is tested by compar ing seedl ing and mature cohorts in fragmented and non-fragmented populat ions. I a lso examine the extent of spatial structuring within stands of bigleaf maple trees and examine how structuring is affected by fragmentation. In chapter five I study the quantitative variation in height, diameter and bud flush traits and compare quantitative genet ic differentiation among populat ions and genet ic differentiation at neutral loci. I then conc lude by summariz ing the major f indings in chapter six, providing speci f ic recommendat ions for management and suggest ing areas for future research that will enhance our knowledge for management and conservat ion of bigleaf maple genetic resources for present and future generat ions. Chapter two LITERATURE REVIEW BIOLOGY AND SILVICS OF BIGLEAF M A P L E (Acer macrophyllum Pursh) The maple family, A c e r a c e a e , includes two genera, Dipteronia and Acer. Dipteronia contains only two spec ies of smal l trees both native to central Ch ina . Acer contains about 148 spec ies of smal l t rees and shrubs that are widely scattered throughout the Northern Hemisphere but are most abundant in the eastern Himalayan Mounta ins and in central Ch ina (Peterson et. al 1999). Thirteen spec ies of maples are indigenous to the North Amer i ca , ten of which are native to C a n a d a (Farrar 1995). Three of C a n a d a ' s ten maple spec ies are native to British Co lumbia : Acer macrophyllum Pursh (bigleaf maple); Acer glabrum Torr.var. dauglasii (Hook.) Dipple (Douglas or Rocky Mountain maple); and Acer circinatum Pursh (vine maple). Whi le substantial information is avai lable on the si lvics, management and genet ics of maple spec ies in eastern North Amer i ca , the extent to which this information is appl icable to bigleaf maple is unknown (Peterson et al. 1999). S o m e plant b iogeographers initially suggested that because of the isolating effect of P le is tocene continental ice sheets on plant distributions (Ritchie 1987), bigleaf male is actually more closely related to some of the As ian and European maples than to the maples in eastern North Amer i ca based on taxonomic features (El ias 1980). This conclus ion was further supported by molecular phylogenetic studies conducted on Acer by Acker ley and Donoghue (1998). The native range of bigleaf maple extends from latitude 33° N to 51° N., mostly within 300 km of the Paci f ic Coas t (Fig 1.1) and it is the Paci f ic Northwest's second most abundant spec ies of hardwood after red alder (Niemiec et a l . 1995). Bigleaf maple grows over a wide range of temperature and moisture condit ions, from the cool , moist marit ime cl imate of Coas ta l British Co lumb ia to the warm, dry cl imate of southern Cal i fornia. It often occurs on coarse gravel soil in mixed stands with Alnus rubra (red alder), Populus trichocarpa (black cottonwood), Thuja plicata (western red cedar), Pseudotsuga menziesii (Douglas-fir) and Tsuga heterophylla (western hemlock) (Farrar 1995). Bigleaf maple is able to produce f lowers and s e e d s as early as 10 years after germinating from seed on open and high productivity sites. S e e d crops are produced every year, and can be prodigious, especia l ly in open-grown trees. F lowers of bigleaf maple are relatively smal l but insect-poll inated. It is po lygamous and both staminate and perfect f lowers are mixed in the same dense cyl indrical, racemes (Minore and Z a s a d a 1990). The fruit is a double samara with slightly divergent wings and a hairy seed case . The flowering period is usually from early Apri l to May. Fruit r ipens by Sep tember or October and seed dispersal occurs from October through January (Ruth and Muerle 1958). S e e d d ispersal is primarily by wind and gravity but d ispersal by some smal l mammals (mice, wood rats, and squirrels) and birds has been reported (Fowels 1965). S e e d s are not dormant and germinate soon after d ispersal . Bigleaf maple is moderately shade tolerant and an excel lent shade tree. Its wood is known to have good properties for use as furniture but it is neither as strong nor as hard as sugar maple (Kerbes 1968). The wood of bigleaf maple is very popular in the piano industry, where it is the most preferred spec ies for piano f rames. It a lso has severa l industrial and domest ic uses such as decorat ive face veneer, container materials, moulding, hardwood flooring, kitchen utensils, pallets, turnery and hardwood plywood, as well as for f irewood (Kerbes 1968; N iemiec et a l . 1995). GENETIC VARIATION AND S T R U C T U R E IN NATURAL POPULATIONS Ecolog ica l characterist ics and life history of plant spec ies influence levels of genet ic variation. Important spec ies characterist ics assoc ia ted with levels of variation include taxonomic status, regional distribution, geographic range, life form, mode of reproduction, seed d ispersal mechan ism, and success iona l status (Hamrick and Godt 1989). Plant breeding sys tems are a primary determinant of genetic structure in plant populat ions, because they alter the probability of random mating among individuals within a population. Outcrossing mechan isms reduce the rate of inbreeding and therefore maintain genetic variability within populat ions (Richards 1986). Self ing tends to dec rease genet ic variation within populat ions and promote genetic differentiation among populat ions (Barrett and Kohn 1991). Inbreeding can occur through either self ing or consangu ineous matings (Ritland 1985). Most trees are predominantly outcrossing but their rel iance on either wind pollination or a wide variety of biotic pollination agents generates considerable variation in outcrossing rates and mating patterns among individuals (Aizen and Feis inger 1994; Hamrick et al. 1991). G e n e flow occurs through seed and pollen d ispersal , and dec reases the level of genet ic differentiation among populat ions (Hamrick and Godt 1989). S e e d d ispersal has a greater homogeniz ing effect than pollen d ispersal because seed transmits both maternal and paternal genes whereas pollen transmits only paternally inherited genes (Nason and Hamrick 1997). Within a given geographical region, seemingly large, cont iguous populat ions often consist of subpopulat ions that are linked along temporal and spatial sca les . This network of populations is defined as meta-populat ion (Hanski 1997). Non-random associat ion of genotypes within meta-populat ions creates further structuring at relatively smal l sca les (Muona et al . 1991). Effects of population size on genetic variation Smal l populations undergo p rocesses predicted by genet ic drift theory and population structure models (Tempelton et al . 1991). Kimura and Crow (1964) demonstrated that for a diploid population the expected heterozygosity (He) at equil ibrium is a direct function of mutation and effective population s ize: 1 + ANeM where Ne is the effective population s ize and ju is the mutation rate. Thus the larger the population, the higher the heterozygosity that can be maintained, all e lse being equal . For example , Drosophila populat ions and mammal spec ies exhibit heterozygosit ies of 12% and 5-6%, respectively for a l lozyme loci with corresponding value of Ne/u of 0.035 and 0.015 (Ohta 1992). A meta-analysis of genetic diversity in common and rare plants in the s a m e genus conc ludes, however, that historically large populat ions that have recently become fragmented may still harbor significant genet ic diversity despite current smal l population s ize. However, in most c a s e s the predicted correlation between genetic diversity and population s ize holds (Gi tzendanner and Solt is 2000). The complexity of random effects of genetic drift on allele f requencies in finite populat ions is summar ized by Kimura (1955) and Wright (1951) in which after one generat ion of random mating, a population with initial heterozygosity (H0) would be expected to dec rease by a proportion of on average such that in generat ion t the expected value of expected heterozygosity (Hartl and Clark 1989) is: H, 1 -2JV. H,, Decl ine in population s ize due to deforestation or fragmentation in already subdiv ided populat ions further increases the probability of loss of al leles and enhances the decl ine in heterozygosity due to genet ic drift. However, such effects do not derive only from direct reduction in effective population s ize, because the magnitude of genetic drift can be predicted as a s imple function of census population s ize only when the population characterist ics meet the assumpt ions of the Fisher-Wright drift model (Cabal lero 1994). Usual ly this is not the case (Hartl and Clark 1989) and Ne tends to be considerably smal ler than the census population s ize N (Frankham 1995). Severa l factors are responsible, including unbalanced sex ratio, unequal fecundity among individuals and population s ize fluctuations (Falconer 1989; Futuyma 1986; Hartl and Clark 1989; Y e h 2000). Aldr ich and Hamrick (1998) found that reproduction of the tree Symphonia globulifera in a 38.5 ha circular plot was dominated by numerous smal l groups of remnant pasture land trees which exper ienced a post-fragmentation increase in fecundity leading to a secondary constriction of the fragmentation bottleneck. Similarly in New Zea land mistletoe {Perexia tetrapetala) pollination and seed set was more than four-fold higher in isolated than cont inuous forest (Kelly et al . 2000). Effects of population size on mating systems The mating system of plant spec ies is an important biological characterist ic because it is a key determinant of genet ic variation, genetic structure and evolutionary potential of plant populat ions (Clegg 1980; Brown 1990). Modes of poll ination, population s ize and density, and plant and floral architecture are all likely to influence mating sys tems (Clegg 1980). Plant mating sys tems are character ized by 1) proportion of outcrossing versus self ing; 2) consangu ineous matings, and 3) the level of correlated paternity, def ined as the proportion of full-sib pairs among outcrossed maternal progenies (Ritland 1989a). Reduct ions in population s ize and increases in the degree of isolation and fragmentation of populat ions can lead to increases in inbreeding (e.g. Farris and Mitton 1984; Murawski et al . 1994; Rai jmann et al . 1994). T h e s e effects may be of particular s igni f icance in woody ang iosperms because population s izes and densit ies can be significantly reduced as a consequence of forest harvesting pract ices and other land use pract ices. In smal l populat ions, e levated levels of inbreeding are expected (Barrett and Kohn 1991). Under these condit ions, select ion can purge early-acting lethal and semi- lethal recess ive al leles from populat ions as they become exposed as homozygotes (Lande and S c h e m s k e 1985; Char lesworth and Char lesworth 1987). However, mutations of mild deleterious effects may become fixed in a process cal led mutational meltdown (Lynch 1985). The level of correlated paternity def ines the probability that a seed tree draws two male gametes from the same pollen donor. This can be regarded as the inverse of effective pollination neighbourhood s ize, ana logous to Wright 's neighbourhood s ize , when consider ing only the d ispersal var iance of male gametes (Austerlitz and S m o u s e 2001). The level of correlated paternity, together with setting rate, will determine the degree of departure from random mating and the signi f icance of genetic drift under the isolat ion-by-distance model (Ritland 1989a). Correlated mating may also influence patterns of select ion and competit ion among sibl ings (Karron and Marshal l 1990). A m o n g the major factors that may enhance correlated paternity in wind-poll inated plants are pollen limitation (Surles et al . 1990), spatially restricted pollen d ispersal (Smouse and Sork 2004), asynchronous floral phenology, unequal male fecundity (Er icksson and A d a m s 1989; Burczyk and Prat 1997), and low conspeci f ic density (Smouse and Sork 2004). Reduct ions in population s ize may directly affect the mating system for three reasons. First, in smal l populat ions the number of local compatible mates is reduced, which even under random mating will increase the likelihood of correlated mating and self-fertilization (Surles et al . 1990). S e c o n d , total pollen availability will dec rease in smal l plant populations, which may result in reduced seed set or increased self ing (Larson and Barrett 2000). The impact of pollen limitation in wind poll inated spec ies , however, remains unclear (Koenig and Ash ley 2003). Third, as a consequence of the typically leptokurtic shape of the pollen d ispersal curve in plants (Levin and Kerster 1974), pollen pool diversity around individual trees in smal l populations may be reduced by the absence of a broad spectrum of long-distance pollen donors (Adams 1992; Ellstrand 1992). Ev idence from exper iments on conifer spec ies suggests that both the quantity and diversity of avai lable pollen in smal l s tands may be significantly lower than in large populat ions (Sarvas 1962; Kosk i 1970, 1973). Little is known, however, about the precise consequences of this potential pollen pool impover ishment for the mating system of particular spec ies . Most studies deal ing with the consequences of smal l population s ize on plant mating sys tems have focused on the outcrossing rate and reproductive output of insect pollinated spec ies , in which the interaction between the spatial structure of populat ions and the pollen foraging behaviour of pollinators poses an addit ional chal lenge (Levin and Kerster 1974; van Treuren et al . 1993; Hauser and Loeschcke 1994; Kennington and J a m e s 1997; Rout ley et a l . 1999). Al though no significant effects of smal l population s ize were detected in some of these studies, a general trend towards increased selfing and reduced seed set has been observed as population s izes decrease . E F F E C T S OF FRAGMENTATION ON GENETIC VARIATION IN PLANT POPULATIONS Habitat d isturbances causing forest fragmentation can impact the genet ic structure of spec ies . Fragmentat ion can disrupt the pre-existing genet ic structure of populat ions, alter genetic p rocesses and result in a net loss and redistribution of genet ic diversity. Differentiation among populat ions will tend to increase as once cont iguous populat ions become subdiv ided into smal l , isolated fragments (Barrett and Kohn 1991). The first genet ic consequence of a reduction in population s ize is the loss of rare al leles, and over time a concomitant dec rease in heterozygosity will occur through genetic drift (Barrett and Kohn 1991). Addit ionally, fragmentation can impact plant-pollinator and plant-seed d isperser interactions (Aizen and Feis inger 1994). A shift in landscape pattern is expected to result in altered animal foraging behaviour (Dirzo and Miranda 1991). General ly , alterations of pollinator behaviour will tend to limit pollen d ispersal , increase the level of inbreeding within populat ions, reduce the rate of inter-population pollen d ispersal and therefore increase among-populat ion differentiation (Bawa 1990). Similarly, a change in foraging behaviour of seed d ispersers could reduce seed flow among existing populat ions, and also dec rease colonizing events, thus reducing the establ ishment of new populat ions (Aizen and Feis inger 1994). Recent ly, two broad approaches for detecting effects of forest fragmentation on genet ic variation have been used: (i) compar ison of f ragmented and unfragmented (continuous) populations, and (ii) analys is of relat ionships between measures of genet ic diversity and indices of fragmentation (e.g. population s ize, isolation, or different age cohorts). Studies of these sorts have produced diverse results. In severa l cases , important genetic effects have been detected. Fragmentat ion has been assoc ia ted with a decl ine in allelic r ichness in a number of cases . For example, in 17 fragmented populations of the perennial Swainsona recta, B u z a et al . (2000) reported a significant reduction in the presence of rare al leles in smal l populations. Similar relat ionships were reported by van Treuren et al . (1991) for Salvia pratensis and Scabiosa columbaria and Prober and Brown (1994) for Eucalyptus albens. Y o u n g et al . (1999) found reduced allelic diversity in smal l populations of Rutidosis leptorrhynchoides, a perennial and self- incompatible spec ies . Further, they argued that the associat ion of dec reased genet ic diversity with low seed production was a consequence of parallel reductions in the number of al leles present at loci controlling self-incompatibil ity (SI). Eros ion of allelic r ichness at SI loci in smal l populat ions has also been found in the rare lakeside daisy Hymenoxys acaulis (DeMauro 1993). G e n e flow between fragments might restore lost al leles very quickly but only when lost al leles are still present in the post-fragmentation metapopulat ion as a whole. For, instance, in a study of Acer macrophyllum (sugar maple) in C a n a d a , Young et a l . (1993) compared genet ic variation in eight patchy populat ions with variation in another eight cont inuous control populat ions. They assumed that genet ic variation in the large control populations represented variation in the pre-fragmentation population, and that the present patchy populat ions were derived from once cont inuous populat ions. A compar ison of genetic diversity parameters between fragmented and control populat ions found genetic diversity (as measured by percentage polymorphic loci, allelic diversity and heterozygosity) was not significantly lower in the fragmented populat ions, nor was there any increase in inbreeding. However, the total number of al leles was six fewer in the fragmented populat ions, which was attributed to possib le founder effects. A s outlined above, not all measures of genetic diversity are expected to be sensit ive to founder effects. However, decl ines in va lues of expected heterozygosity and allelic r ichness have been reported. For example, Prober and Brown (1994) demonstrated that smal l populat ions (< 500 reproductive individuals) of Eucalyptus albens that were less than 250 m from a larger population had a higher allelic r ichness than more isolated smal l populat ions. T h e s e results are crucial as they point out a threshold up to which gene flow from a large population can maintain genetic diversity, but beyond which genetic diversity can decl ine. Similarly, a significant reduction in genet ic diversity and increased genetic differentiation was documented in fragmented relative to cont inuous populat ions of the tropical tree Pithecellobium elegans (Hall et a l . 1996). A relatively high correlation between population s ize and genet ic diversity was also reported by Rai jmann et al . (1994) for Gentiana pneumomanthe. However in a number of other studies on relatively recently f ragmented populations, there were no clear relationships between genet ic diversity and population s ize (van Treuren et a l . 1991; Fore et al . 1992; Young et al . 1993, 2000; B u z a et al . 2000). Ev idence for more rapid genet ic erosion in smal l isolated populations than in less isolated populations was reported by Dayanandan et a l . (1999). They found that genet ic d istance between adult and seedl ing cohorts in fragmented populat ions of Carapa guianensis in Cos ta R i ca was greatest in the most isolated population, which was a lso the only one in which allelic diversity was lower in the adult cohorts. The results from these studies s e e m to suggest that spec ies with similar life history characterist ics such as those mentioned above may be particularly vulnerable to the consequences of fragmentation s ince they typically exist at low densit ies and are predominantly outcrossed (Hamrick and Godt 1989; O'Mal ley and Bawa 1987). Effects of fragmentation on spatial genetic structure Spat ial genetic structure is the non-random distribution of genetic variation among sexual ly reproducing individuals (McCau ley 1997). The spatial distribution of genotypes within plant populat ions is inf luenced by many ecological and evolutionary p rocesses such as limited seed and pollen d ispersal (Wright, 1943; S c h o e n and Latta, 1989; Bacil ieri e t a l . 1994), adult density (Knowles et al . 1992; Hamrick e t a l . 1993; Hamrick and Nason , 1996; V e k e m a n s and Hardy 2004), colonizat ion and disturbance history (Epperson and Chung , 2001 ; Parker et a l . 2001), spatial and temporal patterns of seedl ing establ ishment (Ellstrand, 1992; Schnabe l and Hamrick, 1995; Parker e t a l . 2001), differential select ion and micro-environmental select ion (Linhart et a l . 1981; Slatkin and Arter, 1991) and forest fragmentation (Dol igez and Joly 1997). Of these factors, probably the most widely studied influence on spatial genetic structure is pattern of gene dispersal (Hamrick et al . 1993; Ennos 1994; Hamrick and Nason 1996). Ka l isz et al . (2001) descr ibed general scenar ios of seed and pollen d ispersal under which genet ic structure could develop: (i) If at the sca le of investigation, seed d ispersal is local ized while pollen d isperses long d is tances or randomly, spatial clustering of full and half-sibs will result in the development of significant spatial structure in the absence of inbreeding (e.g., Peaka l l and Beatt ie, 1996; Ka l isz et al . 2001). (ii) If pollen d ispersal is a lso restricted, this will result in inbreeding thereby reinforcing the buildup of more intense genetic structure (Wright 1943; Barbujani 1987). (iii) In contrast, if seeds are widely and independently d ispersed then regardless of whether pollen d isperses long or short d is tances, seed dispersal will effectively randomize the spatial distribution of genetic variation within populat ions (e.g. Dewey and Heywood 1988; Loisel le et al . 1995). In most studies of spatial genet ic structure of tree spec ies with wind d ispersed seeds , whether animal or wind poll inated, many authors have reported either weak or no spatial genetic structure (e.g., Acer saccharum (Perry and Knowles 1991; Young et a l . 1993), Quercus spp.(Streiff et a l . 1998), Psychotria officinalis (Loisel le et al . 1995); Carapa procera (Dol igez and Joly 1996); Pinus spp (Parker et al . 2001; Epperson et al . 2003); Vitelarria paradoxa (Kelly et al . 2004)). Their results expla ined spatial genetic structure by overlapping seed shadows and extensive gene flow via pol len. W o o d y insect-poll inated spec ies with seeds widely d ispersed by birds a lso show weak genetic structure (Dewey and Heywood 1988, Chung et al . 2000). A lack of spatial structure was found for other spec ies by Soka l and Oden (1978), Dol igez and Joly (1997), and Chung et al . (2000). They attributed their results to extensive gene flow, wide seed d ispersal , self incompatibility and dispersal agents. In conclus ion, fragmentation effects on population genet ics of forest tree populat ions are complex and difficult to predict. Theoret ical considerat ions in particular are perhaps more useful in understanding empir ical results rather than predicting them (Young et al.1996). However it is worth noting that both theoretical and empir ical studies suggest that fragmentation can exert some effects on genet ics of f ragmented populat ions. M O L E C U L A R AND QUANTITATIVE VARIATION Forest trees are long-l ived, sess i le organisms that are exposed to large temporal f luctuations in their environmental condit ions. Consequent ly , the demands placed on the adaptabil ity of trees are extremely high compared to other organisms. To fulfill these demands , forest tree spec ies need to maintain large amounts of genetic variation for the preservation of adaptabil ity and survival to subsequent generat ions (Muller-Starck and Gregor ius 1985). On the other hand, the existence of populat ions of healthy plants with little or no detectable genetic variation shows that long-term survival is possib le. Without variability, however, such spec ies will be unable to adapt to new environmental condit ions. Acquis i t ion of sufficient information on the extent and pattern of genet ic diversity, population differentiation across spec ies ranges, and the ecological and genetic relationship among individuals and among populat ions, are essent ia l for establ ishing guidel ines on conservat ion and utilization of genetic resources (Er iksson et al . 1995). The need to understand genet ic structure stems from the necessi ty to answer the essent ia l quest ion of whether one population or many different populat ions will be an effective collection of all the important al leles for a particular spec ies (Bradshaw and Stettler 1995). The answer to this quest ion is critical for the efficient management of natural forests or for any effort to restore deforested habitats by reintroduction (Namkoong 1988). Hereditary bas is of differentiation in morphology and development have been demonstrated in studies of intraspecific variation in quantitative traits beginning over two centur ies ago (see review by Langlet 1971). However, because patterns of geographic variation within spec ies are influenced by different select ive pressures, barriers to gene flow and genet ic drift, the maintenance of genetic variation in natural populat ions thus become very complex (Grant and Linhart 1996). The importance of genet ic drift and gene flow as evolutionary forces in natural populat ions has been thoroughly addressed in studies using molecular markers. However, using such markers to determine genet ic structure has several limitations in providing information that could be useful to define conservat ion strategies (Lynch 1996). Despi te these limitations, molecular markers have been proposed as an indirect indicator of quantitative genetic variation avai lable for adaptation (Petit et a l . 1998). However Wa ldmann and Anderson (1998) indicated three major shortcomings of this approach: (i) The higher mutation rates of quantitative trait characters suggests that the recovery t imes after a bottleneck will be shorter for polygenic variation than for single locus polymorphism. (ii) W h e n non-addit ive var iance is high, the expected loss of additive var iance caused by genetic drift fol lows a different pattern than the reduction in s ingle- locus heterozygosity. (iii) The effect of smal l population s ize on genetic variation is expected to differ for monogenic and polygenic characters owing to select ion having different effects on these types of characters. In view of these dif ferences, planning conservat ion efforts based exclusively on marker gene loci may be mis leading. For this reason quantitative genet ic analys is is an important compl iment in studies of spec ies (Lynch 1996; Storfer 1996). Compar i sons of d ivergence in neutral genet ic markers (as measured by FST) and polygenical ly-control led quantitative traits (as measured by QSf, Wright 1951) al low for an assessmen t of the relative importance of natural select ion and genetic drift as a cause of population differentiation in quantitative traits (Spitze 1993; Prout and Barker 1993; Long and Singh 1995; Podo lsky and Hartsford 1995; Bonin et al . 1996; Y a n g et a l . 1996; Wa ldmann and Anderson 1998; Lynch et al . 1999; G o n z a l e z -Mart inez et al . 2002; Merila' and Crnokrak 2001; M c K a y and Latta 2002; Howe et al . 2003). Higher d ivergence in quantitative traits than in neutral markers (Qsr > FST) is indicative of directional select ion favouring different genotypes in different populat ions, whereas the opposite (Qsr < FST) suggests that the s a m e genotypes are favoured in different populations, i.e. stabil izing select ion. However, if the two measures do not differ significantly (Qsr = FST), then patterns of variation for both neutral markers and quantitative traits are both a s s u m e d to reflect only the act ions of genetic drift and gene flow (Merila and Crnokrak 2001). Comparat ive studies of quantitative trait and neutral marker d ivergence are relevant from a conservat ion genet ics perspect ive as management dec is ions often rest on population genetic ana lyses conducted with neutral molecular markers (e.g. Moritz 1994; Haig 1998; R e e d and Frankham 2001). For instance, operat ional definitions of evolutionarily significant units ( E S U s ) are based on d ivergence in neutral or nearly neutral markers (Moritz 1994; Moritz et al . 1995). However, quantitative characters are more likely to be related to f i tness and therefore to population pers istence (Lynch 1996; Storfer 1996). Never theless, the question of whether the levels of variation and the degree of population differentiation are correlated between neutral genet ic markers and genet ic variation in quantitative traits remains controversial (Pfender et a l . 2000; Meri la and Crnokrak 2001; R e e d and Frankham 2001 ; Latta and M c K a y 2002). In a meta-analysis of 18 studies reporting QST and FST values for 20 spec ies , Meri la and Crnokrak (2001) did indeed find a positive correlation between the two d ivergence indices across different studies (see reviews in Crnokrak and Meri la 2002; Latta and M c K a y 2002; M c K a y and Latta 2002; Howe et al . 2003). However, studies compar ing the predictive power of neutral markers as indicators of d ivergence in quantitative traits among populat ions within spec ies are still lacking (but see Steinger et al . 2002), as are studies examining the sensitivity of Q s r est imates to genotype-environment interactions. Figure 2.1. Native range of Acer macrophyllum (Bigleaf maple). Chapter three GENETIC VARIATION, POPULATION S T R U C T U R E AND MATING S Y S T E M IN BIGLEAF MAPLE (Acer macrophyllum Pursh) 1 INTRODUCTION Studies of the genetic variation and population structure of woody ang iosperms with molecular genetic markers have shown that they p o s s e s s relatively high levels of genet ic variation within populat ions, but little among population differentiation (Hamrick et al . 1992; Love less 1992). Genet ic variation enab les spec ies to survive and adapt to changing environments, therefore, information on genetic variation of forest tree spec ies is fundamental to management and conservat ion (Er iksson et a l . 1995). The most important determinants of genet ic variation are natural select ion, mutation, genet ic drift, migration and the mating system (Hartl and Clark 1997). However, human activities such as deforestation, air pollution and forest fragmentation can modify the direction and ampli tude of these evolutionary forces and alter genetic variation of natural forest resources (Lande 1988). Measurement and characterizat ion of this variation, particularly in relation to human activities, are important first steps towards developing strategies to preserve the genet ic variation of native forest tree spec ies (Hamrick et al . 1991). Bigleaf maple (Acer macrophyllum Pursh) occurs along the Paci f ic coast of North Amer i ca , in populat ions of scattered individuals or as smal l groves, in associat ion with both conifers and broad- leaved trees. Its f lowers are po lygamous and both staminate and perfect f lowers are mixed in the s a m e dense cylindrical, racemes (Minore and Z a s a d a 1990). Poll ination is primarily by insects (Minore and Z a s a d a ' A version of this chapter has been pubished. Mohammed N . Iddrisu and Kermit Ritland. Genetic variation, population structure and mating system in bigleaf maple (Acer macrophyllum Pursh). Can. J. Bot. 82: 1817-1825 (2004). 1990). It is an early success iona l spec ies with consistent seed production. The seeds are double samaras with slightly divergent wings and can be d isseminated by wind for long d is tances. Its scattered distribution over the coastal Paci f ic Northwest and southwestern British Co lumb ia in rural a reas makes it a prime woodlot spec ies (Minore and Z a s a d a 1990). In the past, bigleaf maple populations have suffered much habitat disturbance due to agricultural pract ices, and the marketing of bigleaf maple wood products may in the long run lead to an accelerated loss of its genetic resources. This could dec rease the opportunities for the genet ic improvement and conservat ion of bigleaf maple. However, there has been little study of the genetic structure of bigleaf maple, and this is needed to develop a management strategy des igned to maintain stable, productive and sustainable forest populat ions of this spec ies . The objectives of this study were to: (1) determine the amount and distribution of genetic variation among bigleaf maple (Acer macrophyllum) populat ions, (2) est imate the mating system in two populat ions from the Lower Main land of British Co lumb ia , and (3) recommend a strategy for the management and conservat ion of bigleaf maple genetic resources. I hypothesize that bigleaf maple is predominantly an outcrossing spec ies with extensive gene flow, and therefore, has much genetic variation within populat ions but little population differentiation. MATERIALS AND METHODS S e e d s from eight populat ions representing the range-wide distribution of Acer macrophyllum were col lected. Populat ion locations are given in Fig. 3.1. S e e d s were col lected from 36 adult t rees in Jer icho and 40 trees in Fraser populat ions respectively. In each of these populat ions, seed progenies (progeny arrays) with about 30 s e e d s per mother tree were col lected and used for estimating outcrossing rates. For the rest of the populat ions sampled in the southern portion of the spec ies range, seeds were col lected from 20 adult trees in each population with the except ion of Artie where only 14 adult trees were sampled . The Artie population was at the western edge of the spec ies distribution in Oregon and no additional trees could be found after travelling 3 km westward towards the coast. Individual sampled trees were spaced approximately 30 meters apart as minimum. A s much as possib le, seeds were col lected before the first rainfall. If they were damp from fog and dew or col lected after the first rains, they were dried indoors at room temperature until seeds and paper sacks felt dry, then stored in large plastic garbage bags with holes bored for ventilation at 2-4 °C until germination. Isozyme assay S e e d s were de-winged and germinated in Petri d ishes on filter papers moistened with distilled water, then kept in a 4°C refrigerator for 5-8 d before seed dissect ion and enzyme extraction. Individual coty ledons with emerging radicles were exc ised and placed in separate wel ls in microtiter plates for enzyme extraction. The freshly exc ised material was ground in 2-3 drops of extraction buffer: 0.283g germanium dioxide, 25ml_ water, 0.0917g diethyldithiocarbonic ac id, 0.1g sodium bisulfate, (0.16M) 2.67ml_ phosphate buffer at pH 7, 2 .67mL D M S O , 17mL 2-phenoxythenol and 0.66ml_ p-mercapthoethanol. The extracts were absorbed onto filter paper wicks ( 3 x 1 3 mm), loaded onto 12% starch gels. Ge l s were cooled overnight to 4°C before loading samples . Samp le wicks were removed after half an hour of electrophoresis. The voltage was then set and run from 5 to 7 hours. Rec ipes for h istochemical staining solut ions fol lowed Murphy et al . (1996). Buffer sys tems used were: lithium borate pH 8.3, 80 m A (Ridgeway et al . 1970); and morphol ine citrate pH 8.0, 50 m A (Clayton and Tretiak 1972). In all, 33 enzyme sys tems were initially sc reened for polymorphism and 10 putative a l lozyme loci for 6 enzymes were resolved clearly and consistently, thus selected for analysis. T h e s e were glutamic dehydrogenase (GDH; 1 locus), phosphog lucose isomerase (PGI; 2 loci), leucine aminopet idase (LAP; 2 locus), isocitic dehydrogenase (IDH; 1 locus), asparate aminotransferase (AAT; 2 loci), and 6-phosphogluconate dehydrogenase (6-P G D ; 2 loci). For enzyme sys tems with multiple loci, the most anodal migrating locus (fastest locus) was ass igned as 1 and other loci were ass igned increasing numbers with decreas ing migrating distance. At each locus, the most common allele was arbitrarily designated as 1 and the others 2, and so on. Data analysis Est imates of the following quantit ies were obtained: allele f requencies, mean number of al leles per locus (A), percentage polymorphic loci (%P) at 9 9 % criterion, observed heterozygosity (H0) and expected heterozygosity (HE=1-£pi2, where p, the f requency of the ith allele). This analysis was performed with B I O S Y S - 2 (Will iam C . Black IV, Department of Microbiology, Co lo rado State University), a modified version of the B I O S Y S - 1 program by Swofford and Se lander (1981). Wright 's FST (Wright 1965) was computed for individual loci of the eight populat ions. Nei 's (1978) genetic distance (D) was computed between all pairs of populat ions. A dendrogram of genet ic relationships among populations was constructed from these d is tances using the unweighted pair group method ( U P G M A , Sneath and Soka l 1973), and standard error bars calculated with Rit land's (1989b) genet ic d istance and clustering program (GDD) . Dendrogams plotted using these procedures help to v isual ize the genetic relationship among populat ions. In each population, departures of genotypic f requencies from Hardy-Weinberg expectat ions were character ized by estimating Wright 's inbreeding coefficient as F = 1 - Hc/HE. G e n e p o p (Raymond and Rousse t 1995) was used to est imate P- va lues from exact test of departure from Hardy-Weinberg equil ibrium using the Markov chain method with 1000 iterations (Guo and Thompson , 1992). Linear regression was performed to study the relationship between the expected heterozygosity and latitude, heterozygosity and elevation, and between genet ic d istance and geographic distance using S - P L U S statistical software ( S - P L U S 6, Insightful, Corp . 2001). In addit ion, isolation by distance was tested using Rousse t ' s method (1997), which involves a l inear regression of pairwise FST/(1-FST) on the natural logarithm of geographic d is tances between populations. Signi f icance was tested statistically using the Mantel test (Mantel 1967) in the program IBD version 1.4 (Bohonak 2002). This test a s s e s s e s whether the pairwise geographic d istance matrix correlates with the pairwise genet ic d istance matrix. The degree of genet ic isolation (gene flow) was est imated by Nm, the number of migrants per generat ion. Nm was est imated by two methods, by the relationship between FSr and A / m a n d by the method of private al leles. From Wright (1951): Nm = (1-FST)/4FST, where FST is the proportion of the total genetic diversity among populat ions. I used Genepop (Raymond and Rousset , 1995) to est imate Nm based on the private al leles (unique al leles found in only one population) method developed by Slatkin (1985), using the f requency and distribution of rare al leles among populat ions. Mating system parameters were est imated using the mixed mating model of Rit land and Ja in (1981), as implemented in M L T R (Ritland 1990). S ingle- locus (fs) and multi- locus (tm) est imates of population outcrossing rates were est imated for the two populat ions (Jericho and Fraser) in which progeny arrays were sampled . Maternal genotypes were inferred from progeny arrays for these two populat ions following Brown and Al lard (1970) and used to est imate genet ic diversity parameters for the northern range of the spec ies distribution. Mult i locus outcrossing rates {tm) were compared with mean single- locus (fs) rates to detect any selfing due to biparental inbreeding (tm-ts). The correlation of outcrossed paternity (rp) was est imated following Rit land's (1989a) sibling pair model . T h e s e parameters and the average s ingle- locus inbreeding coefficient of maternal parents (F) were a lso est imated via the M L T R program (Ritland 1990). RESULTS Allele frequency distribution Seedl ing genotypes were scored for a total of 10 loci in two enzyme systems. S o m e loci were apparently monomorphic but were inconsistent in resolution and were exc luded from the analysis (SKD-1, SKD-2, MDH-2). Others were variable but could not be used for the analys is because of overstaining of one locus on top of another, e.g. PGM-1, or because of very faint banding patterns that could not be properly interpreted, e.g. G-6P-1, G-6PDH, ME-1, ACO-1, ACO-2 and UGUT-1. Al le le f requencies for each population are given in Tab le 3.1. A total of 24 al leles were detected in this study. One locus (LAP-2) was monomorphic across all populat ions sampled . Of the nine polymorphic loci, 1 locus (AAT-2) was polymorphic in only one population, 2 loci (6PG-2, GDH) were polymorphic in two populat ions, and one other locus (LAP-1) was polymorphic in six populat ions. Two loci (6PG-1, PGI-1) were polymorphic in seven populat ions and the remaining 3 (AAT-1, IDH, PGI-2) were polymorphic across all populat ions. Two of the eight populat ions had a total of 3 private al leles (i.e. al le les found in only one population). T h e s e private al leles were unique to the two northern populat ions (Jericho, AAT-2-2, AAT-2-3; Fraser, IDH-1-3). Var iab le loci exhibited just 2 al leles per locus with the except ion of Jer icho and Fraser populat ions that exhibited 3 al leles at severa l loci (Table 3.1). Genetic diversity Genet ic variation statistics are summar ized for all populat ions in Tab le 3.2. A l le les per locus averaged 1.71, and ranged from 1.5 (Artie) to 2.2 (Jericho). On average, 61 .2% of the loci were polymorphic, ranging from 5 0 % (Artie) to 8 0 % (Jericho). The expected heterozygosity within populat ions ranged from 0.102 (Jericho) to 0.189 (Artie), and averaged 0.152 (Table 3.2). A significant negative relationship was found between expected heterozygosity and latitude (R2= 0.71; p=<0.05). A n unexpected relationship was also found between expected heterozygosity and al leles per locus (R2= 0.89; p=<0.05), and expected heterozygosity and percent polymorphic loci (R2= 0.53; p=<0.05). Genetic structure Observed heterozygosit ies varied from 0.108 to 0.160, with an average of 0.118. The mean observed heterozygosity was 2 2 % lower than the expected heterozygosity (0.152). Hardy-Weinberg equil ibrium was rejected for three of the eight populat ions (P < 0.05), which showed a def ic iency of heterozygotes. Observed heterozygosit ies were found to be slightly lower than the expected va lues within most populat ions. This heterozygote def ic iency is reflected in the mean inbreeding coefficient (F = 0.166). The sample from Siletz had the highest inbreeding coefficient of 0.334, while Jer icho had a slight e x c e s s of heterozygosity (F =-0.050) (Table 3.2). The mean fixation index ( F / S ; Wright 1951) across loci and populations was 0.193, suggest ing A c e r macrophyllum populat ions have some degree of inbreeding. The proportion of genetic variation due to dif ferences among populations was F S T = 0.054 (Table 3.3), indicating that 94 .6% of the genet ic variation resides within populat ions. The relatively low value indicates little population differentiation which may be due to high levels of gene flow, as est imated by the indirect method (/Vm = 4.39) and for the private al leles method (A/m = 4.10). Nei 's genet ic d istance between populat ions was low averaging 0.011 (SD=0.005) and ranging from 0.001 to 0.042. The dendrogram of genet ic relationship among the eight populat ions is shown in Fig. 3.2. Signif icant clusters of populations occur when the length of the shaded portion (thicker bar) is less than half that of (thinner bar); the method for determining the error is based upon the among- locus var iance of genet ic distance between c lades (Ritland 1989b). The most genetical ly distinct population is Artie, which is separated from the rest of the populations (Fig. 3.2). Severa l statistically significant groups are evident, but the cluster between Oakvi l le and Helmick was not statistically significant, as a large geographic d istance a lso exists between the 2 populat ions (228 km). The similarity between the Jer icho and Fraser populat ions is obvious in that they both share the highest value of allelic diversity, lowest level of heterozygosity and with little inbreeding as compared to the other populat ions. The reason for their similarity could be that they share the most common recent ancestor. No significant correlation was found between genetic d is tances and geographic d is tances (Mantel test, r= 0.36, one tailed p=0.059). Mating system Est imated multi- locus outcrossing rates for the two populations were high (Table 3.4). S ingle- locus outcrossing rates est imate (ts) ranged from 0.939 to 0.942, with an average of 0.94 for the two populat ions. Mult i- locus est imates of outcrossing rates (tm) ranged from 0.941 to 0.950 and averaged 0.945 for the two populat ions. S ing le- locus and multi- locus outcrossing est imates for all populat ions differed significantly from unity ( M ) . Dif ferences in multi- locus and s ingle- locus {tm-ts) can indicate biparental inbreeding, however the difference in this case was essent ial ly zero, indicating no biparental inbreeding (Table 3.4). Individual tree outcrossing rate est imates were heterogeneous in the two populat ions. Both populations exhibited predominant outcrossing, with a large portion of trees having outcrossing rates equal to or greater than 0.90. Maternal inbreeding coefficients (F) were low and did not differ significantly from zero in either population. The correlation of outcrossed paternity rp (probability that s ibs shared the s a m e father) for both populations ranged from 0.234 to 0.544 and averaged 0.389. This value is high, indicative of few effective pollen donors (A/ e p= 1/rp= 2.57, see Ritland 1989a). DISCUSSION Genetic variation Compared with the genet ic diversity found in other woody angiosperms, Acer macrophyllum has a higher percentage of polymorphic loci (P= 61.2%) and expected heterozygosity (/-/E=0.152) than average. However, the number of al leles per locus in bigleaf maple (A=1.71) is slightly less than that in woody spec ies , on average, and slightly higher than the mean in all woody angiosperm (Table 3.5). Dif ferences in the amount of genet ic variation among populations, particularly dif ferences in expected hetrozygosity relative to polymorphism, may reflect the action of different genetic p rocesses . Jer icho and Fraser were the two populat ions with the lowest levels of heterozygosity, and higher polymorphism, with no ev idence of deviation from random mating. In this c a s e it s e e m s more likely that the low heterozygosity may be a reflection of low overall population genet ic variation. This is possibly the result of genet ic drift or select ion in some parts of the northern range of this spec ies . The most genetical ly distinct population is Artie, which is separated from the rest of the populat ions (Fig.3.2). This separat ion in my view is anomalous in that there is no clear environmental explanation as to why it should be different. Isozymically, it differs from other populat ions in having the highest expected heterozygosity (0.189), yet it is the only population which is invariant at the PGI-1 locus, thereby having the lowest proportion of polymorphism. In addit ion, the sample s ize of 14 individuals is smal l compared to the rest of the populat ions which could partially be responsible for its genet ic dist inct iveness. Population genetic structure and gene flow Differentiation among populations, as measured by Nei 's genet ic d istance (D), averaged 0.011 for the eight populat ions of bigleaf maple in this study. This value is similar to that observed in 22 populat ions of Alnus crispa (D = 0.012, Bousquet et al . 1987) but higher than those observed for Acer saccharum (D = 0.003, Perry and Knowles 1989; D = 0.007, Young et al . 1993). The value of genet ic distance (0.011) in this study is probably due to large geographic d is tances. The mean geographic d istance between pairs of population was 250 km and the largest d istance was 562 km suggest ing absence of isolation by distance across the spec ies range. Most of the sites sampled were similar in terms of cl imate and edaph ic factors. T h e s e factors may promote low differentiation among populations. This result is supported by a positive but non-signif icant correlation between genetic and geographic d istances. Simi lar results have been found for Acer saccharum (Young et a l . 1993) and Populus tremuloides (Hyun et al . 1987). The level of population differentiation in bigleaf maple, F S r = 0.054, is similar to that reported for Acer saccharum (FST= 0.049, Young et al . 1993; F S T = 0.033, Perry and Knowles 1989), Alnus crispa (Fsr = 0.051, Bousquet et al . 1987). T h e s e low va lues reflect extensive gene flow via pollen or seed , or recent colonizat ion (Huh 1999). Spec ies with more pollen or seed movement should have less genet ic differentiation than spec ies with restricted gene flow. In support of these predictions, Fore et al . (1992) observed an average seed d ispersal d istance of up to 100 metres for sugar maple (Acer saccharum). C o m m o n life history traits such as al logamy, wind d ispersal of seed , high reproductive capacity, longevity and success iona l behaviour could account for the low differentiation observed. The apparent lack of associat ion between geographic and genetic d is tances somewhat indicated a tendency towards isolation by distance but not statistically significant. Simi lar f indings have been found for Alnus rubra (Xie et al . 2002) that has similar patchy distribution to Acer macrophyllum and also occup ies a similar geographic range. Relat ionships between genet ic and geographic d is tances has been observed for several tree spec ies (e.g. Camellia japonica, Wende l and Parks 1985; Tsuga mertensiana, Al ly et a l . 2000; Pseudotsuga menziesii, Y e h and O'Mal ley 1980), indicating that for these spec ies isolation by distance may be an important factor in population differentiation. However, most studies that have demonstrated such significant assoc ia t ions sample a larger number of populat ions or involve spec ies that cover larger longitudinal or latitudinal distribution than the current study. The observed geographic separat ion between the two British Co lumb ia populat ions (Jericho and Fraser) and the rest of the populat ions may be of recent origin as hypothesized by Pie lou (1991); it is therefore reasonable to suggest that strong gene flow through pollen and seed have overcome the effects of genetic drift so that physical ly separated smal l patchy bigleaf maple populat ions within each geographic region share a more or less cont inuous common gene pool as those cont inuously distributed. For instance, Si letz is widely separated geographical ly (> 500 km) from Jer icho and Fraser but is genetical ly similar. Mating system High outcrossing rates were found in the two populat ions sampled using progeny arrays (Table 3.4). T h e s e outcrossing est imates are similar to other temperate ang iosperm tree spec ies , e.g. Fagus sylvatica (mean t = 0.96, Ross i et al . 1996), Alnus crispa (t = 0.95, Bousquet et al . 1987), Quercus lobata (f = 0.96, Sork et al . 2001), and Eucalyptus urophylla (t = 0.91, House and Bell 1996). Al though Acer macrophyllum is poll inated by insects, wind pollination can not be ruled out. W ind -pollination has been reported for other North Amer ican Acer spec ies that were originally thought to be only insect poll inated (e.g. Acer grandidentatum, Barker et a l . 1982; and Acer saccharum, Gabr ie l and Garrett 1984). Interestingly, six of eight populat ions of bigleaf maple had an e x c e s s of homozygotes, and the est imated F was 0.166. T h e s e results again suggest that some inbreeding and selfing occurs in most populat ions (Table 3.1). Assum ing inbreeding equil ibrium and assuming all inbreeding is due to self ing, this level of inbreeding can be explained by a selfing rate of 2F/(1 + F) = 0.28. S o m e inbreeding in bigleaf maple may result from gei tonogamous poll inations by bumblebees (Bombus spp.), through positive assortat ive mating (i.e., preferential mating among similar genotypes, Sul l ivan 1983), or through mating among relatives. Acer macrophyllum has perfect as well as staminate f lowers and the pollination mechan ism is mainly entomophi lous (Minore and Z a s a d a 1990). The movement of pollinators among adjacent f lowers within the crown or between adjacent crowns of related neighbours could cause inbreeding and selfing (Gonza lez-As torga and Nunez Far fan2001) . I found no significant difference between multi- locus (tm) and s ingle- locus (rs) est imates of outcrossing rates in either bigleaf maple populations, indicating an absence of consangu ineous mating. This result is in agreement with the study by Sork et al . (2001) for Quercus lobata, which occurs in open landscape and is patchily distributed, and for Stemmadenia donnell-smithii by J a m e s et al . (1998). Biparental inbreeding has been reported for some woody spec ies , including western larch (El-K a s s a b y and Jaqu ish 1998), Fagus sylvatica (Ross i et a l . 1996) and Eucalyptus marginata (Millar et al . 2000). Two significant results I obtained were that "correlated matings" (the fraction of outcrossed sibling pairs that share the s a m e father) were high, especial ly for the Fraser population (rp= 0.544, Tab le 3.5). This correlation is inf luenced by two factors: (i) multiple deposi ts of pollen from a single male parent, or (ii) repeated mating among a relatively smal l number of neighbours nearer to one another. A value near one-half implies that only a few pollen donors (1-2) must have sired the majority of seeds within each tree, while the value of 0.234 for the Jer icho population indicates 4-5 effective pollen donors per tree. The lower number of effective pollen donors at Fraser could be due to features of this habitat, compared to Jer icho. Fraser is a roadside population which may be expected to show reduced outcrossing rates due to disturbances. However, the outcrossing rates were similar between the two populat ions, suggest ing that high outcrossing was being maintained in the relatively disturbed Fraser population due to insect or pollen movement along a corridor of trees (Chase et a l . 1996). The other significant result was that, despite the ev idence of moderate inbreeding (F / s >0) in this spec ies , I found high levels of outcrossing (t = 0.95). Al lozyme-based est imates of outcrossing rates based upon seed progeny could give upwardly b iased est imates of outcrossing if selfing reduces the germination capacity, or if filled s e e d s are used for estimating outcrossing rates (e.g., the effects of embryonic lethals due to selfing on seed development and the formation of empty seeds are not accounted for; Rajora et a l . 2000, 2002). In this study, I used entirely germinated and filled seeds s ince these yielded interpretable enzyme bands. However, this may have upwardly b iased est imates of outcrossing. Self-fertil ization has been found to adversely affect both embryo development and seed germination in conifers (Sorensen 1969). This a lso holds for some spec ies in the maple family; for example, Gabr ie l (1962) examined the interior of carpels of sugar maple seeds {Acer saccharum), and suggested that the low seed set after self ing may be related primarily to post-zygotic abortion. In addition, Gabr ie l (1962) noted a reduction in viability of sel fed seeds compared to outcrossed seeds , e.g., inbreeding depress ion. IMPLICATIONS FOR MANAGEMENT AND CONSERVATION The observed genetic differentiation among bigleaf maple populations probably reflects the combined effects of ecologica l , evolutionary and biogeographic factors such as pollen and seed dispersal mechan isms. Consider ing the relatively large geographic sca le in this study, my results indicate lower than expected levels of differentiation among populations of bigleaf maple, in light of the fact that most populat ions in this study have trees patchily distributed, and that trees are poll inated by insects. This study also found that bigleaf maple is predominantly outcrossing (t = 0.95) and lacks biparental inbreeding. High outcrossing is related to its floral biology and characterist ics, e.g., protogyny. Perhaps the most interesting finding was the smal l numbers of effective pollen donors, which probably reflects the relatively low density of populat ions coupled with limited pollinator movement. Correlated paternity results in an increased genet ic re latedness of progeny, and a dec reased genetic diversity of individual tree progenies, which may limit local adaptive responses (James e t a l . 1998). Understanding genetic diversity and population genetic structure is not only crucial in the conservat ion of spec ies under threat of extinction, but it is a lso essent ia l for the maintenance of healthy populat ions and the breeding of w idespread indigenous tree spec ies (Millar and Westfal l 1992). The genet ic variation and population genet ic structure revealed in this study are instructive for making conservat ion plans and developing breeding strategies. The est imate of FST= 0.054 in this study indicated up to about 9 5 % of the total genetic diversity resides within populat ions. Therefore for such a predominantly outcrossing spec ies (tm = 0.95) with insect pollination and seed d ispersal by wind, it may be advisable to sample fewer populat ions but more individuals per population for breeding purposes. Measu res of genetic diversity based on number of al leles (allelic r ichness) are important, especia l ly in the field of conservat ion genet ics. S ince one goal of a conservat ion program is to maintain as many al leles as possib le, cho ices of populat ions to conserve in situ should be based on allelic r ichness of the population (Marshal l and Brown 1975). In view of this, I suggest that although a few in situ populat ions would contain most of the existing genet ic variation in the spec ies , it would be essent ia l to a lso consider the populat ions from the northern range; thus, Jer icho and Fraser would be favored. Al though inbreeding perse does not lead to loss of al leles nor alter their f requencies in a population, it leads to increased in homozygosi ty, and thus may dec rease the mean fi tness of the population. To this end, populat ions displaying extensive inbreeding would not be desirable for future in situ gene conservat ion. The lack of correlation between genet ic and geographical d is tances in the spec ies distribution suggests that when sampl ing, we may not necessar i ly need to sample sites evenly across the spec ies range. Al though bigleaf maple is widely distributed without any current threat of extinction, effective in situ conservat ion and reasonable management of its populat ions in the wild will promote and enhance its adaptability to changing environments, and also sustain its gene pool for future genet ic improvement. Table 3.1. Distribution of allele f requencies at 10 loci in eight natural mature populat ions of bigleaf maple (Acer macrophyllum). Populations Locus Allele J E R C F R A S A R T C C A S C E L B E HELM O A K V SILE AAT-1 1 0.942 0.958 0.857 0.900 0.895 0.789 0.833 0.917 2 0.058 0.042 0.143 0.100 0.105 0.211 0.167 0.083 AAT-2 1 0.904 1.000 1.000 1.000 1.000 1.000 1.000 1.000 2 0.077 0.000 0.000 0.000 0.000 0.000 0.000 0.000 3 0.019 0.000 0.000 0.000 0.000 0.000 0.000 0.000 IDH 1 0.923 0.854 0.786 0.850 0.850 0.850 0.850 0.850 2 0.077 0.125 0.214 0.150 0.150 0.150 0.150 0.150 3 0.00 0.021 0.000 0.000 0.000 0.000 0.000 0.000 6PG-1 1 0.981 1.000 0.786 0.825 0.825 0.800 0.875 0.850 2 0.019 0.000 0.214 0.175 0.175 0.200 0.125 0.125 6PG-2 1 0.885 0.875 1.000 1.000 1.000 1.000 1.000 1.000 2 0.038 0.063 0.000 0.000 0.000 0.000 0.000 0.000 3 0.077 0.062 0.000 0.000 0.000 0.000 0.000 0.000 PGI-1 1 0.962 0.896 1.000 0.816 0.853 0.789 0.875 0.825 2 0.019 0.042 0.000 0.184 0.147 0.211 0.125 0.175 3 0.019 0.063 0.000 0.000 0.000 0.000 0.000 0.000 PGI-2 1 0.885 0.938 0.714 0.825 0.775 0.875 0.875 0.800 2 0.077 0.021 0.286 0.175 0.225 0.125 0.125 0.200 3 0.038 0.042 0.000 0.000 0.000 0.000 0.000 0.000 GDH 1 0.981 0.917 1.000 1.000 1.000 1.000 1.000 1.000 2 0.019 0.083 0.000 0.000 0.000 0.000 0.000 0.000 LAP-1 1 1.000 1.000 0.538 0.825 0.781 0.850 0.850 0.800 2 0.000 0.000 0.462 0.175 0.219 0.150 0.150 0.200 LAP-2 1 1.000 1.000 1.000 1.000 1.000 1.000 1.000 1.000 Note: JERC = Jericho; FRAS = Fraser; ARTC = Artie; CASC = Cascadia ;ELBE = Elbe; HELM= Helmick; OAKV = Oakville; SILE = Siletz. Table 3.2. Summary of genet ic diversity within eight mature natural populat ions of bigleaf maple (Acer macrophyllum) based on 10 al lozyme loci. P O P U L A T I O N N A %P Ho HE F 1. Jer icho 40 2.2 80.0 0.108 0.102 - 0 . 0 5 0 N y (0.2) (0.029) (0.026) 2. Fraser 36 2.0 60.0 0.112 0.105 - 0 . 0 8 6 N S (0.3) (0.036) (0.033) 3. Artie 14 1.5 50.0 0.160 0.189 0 . 1 0 5 N S (0.2) (0.060) (0.066) 4. C a s c a d i a 20 1.6 60.0 0.121 0.164 0 2 2 2 N S (0.2) (0.037) (0.046) 5. E lbe 20 1.6 60.0 0.109 0.172 0.285* (0.2) (0.035) (0.049) 6. Helmick 20 1.6 60.0 0.118 0.176 0.332* (0.2) (0.038) (0.049) 7. Oakvi l le 20 1.6 60.0 0.117 0.148 0 . 1 8 6 N S (0.2) (0.033) (0.041) 8. Si letz 20 1.6 60.0 0.102 0.163 0.334* (0.2) (0.031) (0.047) Mean 23.75 1.71 61.2 0.118 0.152 0.166* Note: N, sample size; A, average number of alleles per locus; %P, percent polymorphic loci; H0, observed and heterozygosity ; HE, expected heterozygosity; F, inbreeding coefficient; and numbers in parenthesis, standard errors. Exact test of departure from Hardy-Weinberg equilibrium * P < 0.05, N S not significant after sequential Bonferroni correction (Rice 1989) Table 3.3. Total gene diversity (HT), genetic diversity within populat ions (Hs), expected heterozygosity (H0), al le les per locus (NA), fixation index over the total populat ions (FIT), fixation index within population (F / s ) , and genetic differentiation among populat ions (FST) for eight mature natural populat ions of bigleaf maple (Acer macrophyllum) at nine polymorphic loci. Locus HT Hs Ho NA Fis FIT FST AAT-1 0.202 0.202 0.169 2.00 0.116 0.143 0.030 AAT-2 0.024 0.023 0.024 3.00 -0.088 -0.013 0.067 IDH 0.254 0.258 0.222 3.00 0.125 0.134 0.009 6PG-1 0.231 0.226 0.134 2.00 0.396 0.431 0.057 6PG-2 0.059 0.056 0.060 3.00 -0.101 -0.029 0.065 PGI-1 0.219 0.213 0.194 3.00 0.060 0.105 0.048 PGI-2 0.279 0.276 0.180 3.00 0.308 0.335 0.039 GDH 0.025 0.024 0.026 2.00 -0.076 -0.015 0.056 LAP-1 0.277 0.252 0.171 2.00 0.300 0.390 0.128 Mean 0.174 0.170 0.131 2.55 0.193 0.236 0.054 Table 3.4. Est imates of multi- locus outcrossing rates (tm), single locus outcrossing rates (fs), biparental inbreeding (tm-ts), parental inbreeding coefficients (F) and correlation of paternity among sibl ings (rp). Populat ion tm ts tm.ts F rp Jer icho 0.941 0.939 0.002 0.052 0.234 (0.057) (0.008) (0.052) (0.010) (0.053) Fraser 0.950 0.942 0.008 0.053 0.544 (0.067) (0.054) (0.034) (0.009) (0.067) Mean 0.945 0.940 0.005 0.052 0.389 Note: Standard errors in parentheses. Table 3.5. Compar ison of within-population genet ic diversity for Acer macrophyllum with average va lues for all plants, woody spec ies , woody ang iosperms, and for Acer spec ies . Categories A %P HE GST/FST Reference All plant species 1.52 34.6 0.113 0.228 Hamrick et al. 1992 Woody species 1.76 49.3 0.148 0.084 Hamrick et al. 1992 Woody angiosperms 1.68 45.1 0.143 0.102 Hamrick et al. 1992 Acer saccharum 2.80 87.5 0.150 0.012 Fore et al. 1992 Acer saccharum 1.95 38.2 0.110 0.033 Perry and Knowles 1989 Acer saccharum 2.03 53.7 0.109 0.049 Young et al. 1993 Acer platanoides 1.92 53.9 0.128 0.120 Rusanen et al. 2000 Acer platanoides 2.0 54.5 0.132 0.009 Rusanen et al. 2003 Acer cam pest re 3.15 100 0.287 - Bendixen 2001 Acer macrophyllum 1.71 61.2 0.152 0.054 This study Note: See tables 3.2 and 3.3 for de finition of variables. -130 -125 -120 -115 5 0 1 V v M . v 1 , 5 0 ' ^ e n c h o # F r a s e r - a 48 n n 48 'Artie if O* Oakvi l le ° Elbe 46 y _|<m_ ;> y 46 • F m PI " 0 50 100 | j ( Siletz f O • H e i m i c k I O Cascadia 44 n n 44 "130 .125 -120 -115 Figure 3.1. Geograph ic locations of eight Acer macrophyllum mature natural populat ions. tiMi^ MiMmWiM ARTIC FRASER Hi JERICHO OAKVILLE HELMICK ELBE SILETZ CASCADIA Figure 3.2. U P G M A cluster analysis of Nei's genetic distances between eight mature populations of Acer macrophyllum. Chapter four E F F E C T S OF FOREST FRAGMENTATION ON GENETIC VARIATION AND SPATIAL GENETIC S T R U C T U R E IN NATURAL POPULATIONS OF BIGLEAF MAPLE {Acer macrophyllum Pursh) INTRODUCTION Worldwide, human development is rapidly encroaching upon and subdividing many remaining natural areas. Fragmentat ion of the landscape produces remnant vegetation patches, surrounded by a matrix of different vegetat ion types or unvegetated land uses . Habitat fragmentation, the breaking up of cont inuous forest into smal ler patches, can reduce population s ize and increase population isolation (Young et a l . 1993; Andren 1994). It a lso reduces the availability of suitable colonizat ion sites for the establ ishment of new populations (Wilcox and Murphy 1985). Studies of natural plant populations have shown that population s ize is an important factor in determining the amount of genetic variation maintained within sexual ly mature populations and the distribution of this variation among individuals (Sampson e t a l . 1988). Habitat fragmentation may erode genetic diversity and increase population differentiation, affecting population viability in the short or long term (Young et a l . 1996). T h e s e effects are due mainly to increased genet ic drift and inbreeding in habitat f ragments with smal l census s izes and reduced gene flow between fragments (Young et a l . 1996). The subdivision caused by fragmentation will promote local population differentiation if gene flow barriers are establ ished and subpopulat ions diverge due to genetic drift (Bac les et a l . 2004). In addition to isolation, the genetic structure of natural populations prior to fragmentation may determine how large the impact of fragmentation or habitat loss would be. For example , if f ragments are large enough to maintain the genetic structure of the original populat ion, differentiation among fragments may be less marked. On the other hand, if f ragments are smal l and scattered, they will be more likely to contain a b iased sample of the original genet ic variation, and differentiation will be further promoted if isolation persists (Nason and Hamrick 1997). The impact of fragmentation varies among organisms, depending on the effects of fragmentation on reproduction, d ispersal and gene flow, and the original distribution of genet ic diversity (Young 1996). Another potential consequence of fragmentation is a change in mating sys tem. In plant populat ions in particular, inbreeding can increase due to either increased self-pollination or through an increased probability of mating between individuals sharing recent common ancestry (Rai jmann et a l . 1994; Young et a l . 1996). Resul ts avai lable from empirical studies of the effects of habitat fragmentation on several ang iosperm tree spec ies from the tropics have indicated that fragmentation can have significant genetic consequences . Reduced population s ize and increased isolation assoc ia ted with habitat fragmentation may cause a reduction in genetic variation, increased population differentiation between habitat f ragments (Wilcove 1987; Templeton et a l . 1990; C h a s e et al .1996; White et a l . 1999), and increased inbreeding (Lee et a l . 2000; Fuchs et a l . 2003) as predicted by theory. However, there is a lso growing empir ical ev idence for enhanced gene flow between isolated trees in forest f ragments (Young et a l . 1993). It appears that the effects of habitat fragmentation on the genet ic behavior of tree spec ies are more varied and complex than first thought (Young 1996; Aldr ich and Hamrick 1998). Populat ions of long-lived woody perennials s e e m to be resistant to changes in genetic diversity due to long generation t imes, overlapping generat ions and high levels of gene flow (White et a l . 1999; Merwe at a l . 2000). However, genet ic losses are more observable in seedl ing cohorts than in adult cohorts because seedl ings reflect the genet ic effects of reduced present-day levels of gene flow and population s ize (Lee et a l . 2000). Bigleaf maple is a consp icuous spec ies in the temperate coastal rainforests of the Paci f ic Northwest. It grows in a variety of soi ls throughout its range and it is usually a smal l to medium-s ized tree. The trees are usually scattered or in smal l groves in associat ion with conifers and other broad- leaved spec ies . A s this spec ies is commonly found in remnant forests surrounded by pasture lands and agricultural f ields, it is likely that forest fragmentation has divided formerly larger Acer macrophyllum populat ions into smal ler and isolated patches in some parts of its range. Accordingly , Acer macrophyllum populat ions have exper ienced reductions in effective population s ize and spatial extent, which may have significantly reduced genet ic variation and altered population genet ic structure. It is therefore hypothesized that forest patch populations (fragments) of Acer macrophyllum will have less genetic variation and increased levels of inbreeding than more cont inuous populations. In this study, I compared genetic diversity, genet ic structure and inbreeding level in seedl ing and adult cohorts from both fragmented and cont inuous populations. In addition I used computer simulat ions to forecast the decl ine in genet ic variation due to forest fragmentat ion. The speci f ic quest ions addressed are: 1. Does genet ic diversity differ between adult bigleaf maple populat ions of trees in f ragmented and continuous forests? 2. Is inbreeding different in seedl ing than that in adult cohorts? 3. Is there spatial genetic structure in bigleaf maple populat ions? 4. If spatial genet ic structure exists, does it differ between cont inuous and fragmented populat ions? If Acer macrophyllum genetic variation is being affected by fragmentat ion, we would expect: a) less overall genet ic diversity in f ragmented forests than in cont inuous forests; b) lower inbreeding in cont inuous than in f ragmented landscapes ; c) stronger spatial genetic structure in fragmented than in cont inuous populat ions. MATERIALS AND METHODS Populations and sampling Six study populat ions were located on Vancouver Island (Fig 4.1). These compr ised three populat ions that occurred in areas in which forests have been fragmented over the past 150 years due to agriculture and urban development, and three in relatively cont inuous forests. The intent of the sample design was to have a control (continuous populations) against which the genetic effects of fragmentation could be tested. Populat ions sampled occurred between 20 and 150 m in elevation. For each population and at each sampl ing site, severa l Eas t -Wes t transects were establ ished, each approximately 100 m wide. A long these transects, 50 seedl ings were col lected from the forest floor from each of the six populat ions, and terminal buds of lateral shoots were col lected from each of 50 adult trees in each populat ion. A s much as possib le, seedl ing and mature trees were samp led at least 30 m apart to avoid sampl ing c losely related individuals. The total area from which tree samp les were col lected varied widely due to the patchy nature and variable density of populat ions, ranging from 79 ha at Rosewal l C reek to 312 ha for Nitinat (Table 4.1). After col lect ion, seedl ing and bud samples were wrapped in a luminum foil, labeled and frozen in liquid nitrogen. Upon return to University of British Co lumb ia , samp les were immediately transferred into a -20°C freezer for the seedl ings and to a -80°C freezer for bud samp les until electrophoresis. Electrophoresis Horizontal starch gel electrophoresis was used to obtain a l lozyme data for seedl ings and adults for the same 10 loci descr ibed in chapter three, with the except ion of L A P - 1 , which was monomorphic across all adult populat ions and was not included in this analys is . The extraction buffer used to grind emerging leaf t issues from seedl ings and bud t issues from adult trees was the s a m e as descr ibed in chapter three. Data Analysis Standard genetic diversity parameters (allele f requencies, average number of al leles per locus (A), observed heterozygosity (Ho), and expected heterozygosity (HE) were est imated for both seedl ing and adult cohorts in all populat ions. Th is analysis was performed with B I O S Y S - 2 (W.C. Black IV, Department of Microbiology, Co lorado State University), a modified version of the B I O S Y S - 1 program, by Swofford and Se lander (1981). Departures in genotype f requencies from Hardy-Weinberg expectat ions were tested using a Markov chain method following the algorithm of G u o and Thompson (1992). Exact tests for these departures were conducted at each of the variable loci, and l inkage disequil ibrium was tested between pairs of variable loci. The inbreeding coefficient F / s (Wright 1951) was est imated following We i r and Cocke rham (1984). Al l calculat ions and tests above were performed using G e n e p o p (version 3.1 d) (Raymond and Rousset , 1995). I a lso used B O T T L E N E C K (version 1.2.02) descr ibed by Cornuet and Luikart (1996) to test for historical reductions in population s ize . If a population has been through a bottleneck it should show a signature of reduced allelic r ichness compared to expected heterozygosity, as rare al leles are lost faster than heterozygosity dec reases . After a bottleneck, the expected heterozygosity (HE) computed from allele f requencies for a sample of genes should be larger than the heterozygosity expected (Heq) based on the number of al leles in the same sample , assuming the population is at mutation-drift equil ibrium (Cornuet and Luikart, 1996). Isozymes are expected to conform to the infinite allele model (IAM), where each new mutation gives rise to a new allele different from all existing ones (Kimura and Crow, 1964), thus data were analyzed under this model . To test for a def ic iency or e x c e s s in HE, the Wi lcoxon s igned-ranks test was used as it has more power than the sign-test and can be used effectively with fewer loci (Cornuet and Luikart, 1996; Piry et a l . 1999). Genetic structure The genetic structure of populations was a s s e s s e d according to Wright 's (1965) F-statistics following Wei r and Cocke rham (1984). T h e s e fixation indices were used to measure deviat ions from Hardy-Weinberg equil ibrium attributable to individuals within local populat ions (F / s ) , variation among populations (FST,) and variation among individuals relative to all populations pooled (FIT). The signi f icance of these parameters was tested based on 1800 permutations of al leles among individuals within samp les , genotypes among samples , and al leles among samp les , respectively. M e a n s and standard errors were obtained by jackknif ing over loci. A boostrap conf idence interval (CI) of 9 5 % was considered significant when conf idence intervals did not overlap zero. These calculat ions were made using the program F S T A T (Goudet, 2000). Spatial autocorrelation analysis of genetic variation Spat ial genetic structure within populations was a s s e s s e d using Cocke rham 's (1969) est imates between all possible pairs of individuals at different inter-tree d is tances in each population for the adult cohort. This method provides a powerful test of spatial genetic structure (Hardy and V e k e m a n s 2002). The coancest ry coefficient (p,y) has been used in a number of studies recently (e.g., Loisel le et a l . 1995; Peaka l l and Beatt ie, 1996; Burke et a l . 2000; Ka l i sz et a l . 2001 ; Parker et a l . 2001). The parameter p,j was est imated for each distance c lass using the software program Spat ial Pattern Ana lys is of Genet ic Diversity ( S P A G e D i ) 1.1 (Hardy and V e k e m a n s 2002). The software uses the estimator descr ibed by Loisel le et a l . (1995) as fol lows: Pij = M&-P)(Pi-P) + 2 /cp(1-p) (8/c + 1 ) ° 5 -1 where p, and p 7 are the f requencies of homologous al leles at a locus for individuals / and j; p is the mean frequency for that allele; and k = n(n -1) / 2, the number of possible pairs between n individuals located in each distance c lass . The second term in the equation adjusts for bias assoc ia ted with a finite sample s ize and results in p,y having an expected value of zero for a population in Hardy-Weinberg equil ibrium. The results were combined across loci to estimate coancestry by weighting the va lues for each locus by its polymorphic index, 2 p,- (1 - pi). For a population in Hardy-Weinberg equil ibrium, the coancestry between individuals is a measure of the inbreeding coefficient of their hypothetical offspring with expected va lues of 0.25 for pairs of full-s ibs, 0.125 for half-sibs, and 0.0625 for first cous ins. Individual tree locations were identified by a coordinate grid sys tem using a hand-held G loba l Posit ioning Sys tem instrument ( G P S Garmin Model 12XL) . E a c h tree was mapped on a North-South and Eas t -West (x,y respectively) grid using G P S data to construct inter-tree distance matrices for spatial autocorrelat ion analys is. With this procedure, each scored genotype was ass igned to its corresponding spatial location within each population. Eleven to fourteen distance c lasses were used for the spatial autocorrelation analys is. Al l populations had the following intervals: 0 - 50 m, 50 - 100 m and nine 100 m interval up to 1000 m for all populations. Rosewal l had two addit ional distance c lasses (1000 - 1200 m and 1200 - 1500 m), and Maple Bay , Ye l low Point, and Ni t inatan addit ional four c lasses ; ( 1 0 0 0 - 1200 m, 1 2 0 0 - 1500 m, 1 5 0 0 - 2 0 0 0 m, 2000 - 2500 m). Distance c lasses were chosen so that each contained at least 30 pairwise compar isons. Th is analysis tests whether pairs of trees within a specif ied distance interval exhibit the s a m e al leles more often than expected by chance under a random spatial distribution (Hardy and V e k e m a n s 2002). M e a n est imates of coancestry were obtained over all pairs of individuals for the distance c l asses descr ibed above. W h e n p,y = 0, there is no genetic correlation between the f requencies of al leles in individuals at the spatial sca le of interest; when p,y > 0, individuals in a given distance c lass are more closely related than expected by chance; and conversely, when p,y < 0, individuals within a given d is tance c lass are less related than expected by chance. Est imates of coancest ry were tested for signi f icance with a randomizat ion procedure that generated populat ions with a random spatial distribution of genotypes (i.e. no spatial structure). In each plot, intact multi locus genotypes were randomly drawn, with replacement, from the sampled data and ass igned to points occupied by plants; new p,y va lues were then calculated. This randomizat ion procedure was repeated 499 t imes for each plot, giving (together with the originally sampled data) 500 p,y va lues, from which 9 5 % conf idence intervals were constructed. A p,y estimate falling outside this conf idence limit is cons idered significant. If genet ic structure exists, then we expect a pattern of significant va lues at shorter distance c l asses becoming non-significant or negative with increasing distance. Finally, to test whether the s lope (b) of the correlograms obtained for py was statistically significant, py est imates were permuted (999 t imes) with respect to the upper bound (m) of each distance c lass under the null hypothesis b=0. Simulations I used the simulation program B O T T L E S I M (Kuo and J a n z e n 2 0 0 3 ) to estimate the current levels of genet ic variation in fragmented and cont inuous populat ions, forecast their future genetic diversity levels and make recommendat ions with respect to sustainable population s ize . The program al lows specif icat ion of an arbitrary population s ize and projects the decl ine genetic diversity due to genet ic drift based on the actual allele f requencies estimated from the genotypic data input. In order to project the most realistic projections of decl ine in genet ic variation for Bigleaf maple, I used the over- lapping generation model of the program. Other parameters during the simulation process were set as fol lows: degree of generation overlap = 1 0 0 (i.e. all individuals start with random age value that is within the longevity limit), monoecy with random mating and self ing reproductive sys tem, expected longevity = 1 2 5 years, age of reproductive maturation = 1 0 years (Minore and Z a s a d a 1 9 9 0 ) , number of years simulated = 2 5 0 years , effective population s izes N E = 5 0 and N E = 1 0 0 respectively for both cont inuous and fragmented populations, and number of iterations = 1 0 0 0 . I compared the empir ical data to the simulation results in order to determine whether the levels of genet ic variation found in fragmented populat ions will be lower than those in cont inuous populat ions. If the fragmented populat ions show levels of genetic variation that are lower relative to cont inuous populat ions, then the empirical data are consistent with the hypothesis that fragmentation affects Acer macrophyllum populat ions. In contrast, if there is no decrease in va lues of A o a n d HE respectively in fragmented populations relative to continuous populat ions, then the empir ical data are consistent with the hypothesis that fragmentation has not affected populat ions of A c e r macrophyllum. R E S U L T S Allele frequencies Eight of the nine loci were polymorphic in at least one of the populat ions examined for both seedl ings and adults. In all populations samp led , G D H was monomorphic for both adults and seedl ings. Two to three al le les were detected for each polymorphic locus, with a total of 18 al leles observed for adults in cont inuous populat ions and 20 al leles in fragmented populations (Table 4.2 a). In seedl ings, a total of 19 al leles were observed in continuous populat ions and 18 in fragmented populat ions (Table 4. 2 b). The majority of the al leles were common and distributed widely ac ross most populat ions, but a few rare al leles were private, unique to one population, or found only in couple of populations. The distribution of al lele f requencies was the typical U-shaped, with most al leles either rare or nearing fixation (Fig 4.2a-b). Exact test for l inkage disequil ibrium did not yield any significant va lues for seedl ings or adults in any populat ions, indicating independence of loci used in this study. Genetic diversity In all six populat ions, both seedl ing and adult cohorts p o s s e s s e d similar levels of genetic variation regardless of whether populations were fragmented or cont inuous (Table 4.3). The mean number of al leles per locus for adults averaged 1.66 in cont inuous populations and 1.60 in fragmented populat ions, these est imates were 1.73 and 1.66, respectively, in seedl ings. Expected heterozygosity w a s slightly higher in adults in fragmented (0.134) than continuous populat ions (0.120), but slightly lower for seedl ings (0.130 versus 0.140). Seed l ings in cont inuous populat ions had a slightly higher proportion of polymorphic loci than seedl ings in f ragmented populat ions or in the adult cohorts (Table 4.3). Levels of Inbreeding Within all adult populations, genotypic f requencies showed signif icant departures from Hardy-Weinberg expectat ions (P<0.05) with an e x c e s s of homozygotes. In the seedl ing cohort, one cont inuous population (Elk Fal ls) and one fragmented population (Maple Bay) did not deviate significantly from Hardy-Weinberg expectat ions. F/s varied considerably among both loci and populations, with overal l va lues of 0.20 for adults and 0.28 for seedl ings in cont inuous populations, and 0.25 in adults and 0.37 for seedl ings in fragmented populations, suggest ing substantial levels of inbreeding (Table 4.3). Bottleneck test There was no significant bottleneck signature in any of the populat ions. Recent ly bott lenecked populations should show a mode shift of distribution in allele f requencies so that al leles in low frequency c lasses (<0.1) become less abundant than intermediate and high f requency c lasses . The bottleneck program did not show any significant mode shift, thus, all allele f requency distributions were U-shaped (Fig 4 .2a -b), moreover, the Wi lcoxon test detected more heterozygosity than expected under mutation-drift equil ibrium (Table 4.4). Genetic structure Eight polymorphic loci were consistently scored , of which PGI -2 , L A P - 2 , and IDH had high gene diversit ies (HT) in excess of 20%, suggest ing adequate variation for appreciable genetic structure (Table 4.5 a-b). F / s and F , r est imates were positive and significantly greater than zero, suggest ing a deficit of heterozygotes. However, individual loci showed a great deal of variation in their fixation indices. For instance, A A T - 1 , 6 P G - 2 , P G I - 1 , and PGI -2 , show significant excess of heterozygotes, while A A T - 2 , IDH, and L A P - 2 show a significant deficiency, in a manner consistent with inbreeding (Table 4.5 a-b). There was a low but significant amount of genetic differentiation among adult populations both in the cont inuous and the fragmented forests, with a mean FST across loci of 0.015 (95% CI = 0.005 - 0.035) for cont inuous populat ions (Table 4.5 a) and 0.031 (95% CI = 0.010 - 0.056) for fragmented populat ions (Table 4.5 b). Pairwise FST est imates among populations (Table 4.6) were a lso low in all c a s e s suggest ing extensive gene flow between these populat ions or a recent common ancestral population (Table 4.6). Spatial genetic structure Analys is of spatial genet ic structure revealed significant, positive spatial genetic structure in four of the six populat ions. Al l f ragmented populat ions (Fig 4.3 d-f) had significant, positive multi locus coancestry (p,y) coefficients at inter-tree d is tances of up to 100 m distance for Rosewal l and Maple Bay, and up to 200 m for Ye l low Point. For instance, in the 50-100 m distance c lass for these fragmented populat ions, p,y ranged from 0.13 to 0.30, averaging 0.22. This estimate suggests that in f ragmented populat ions of Acer macrophyllum, trees sampled up to approximately 100 m apart are likely to be nearly as similar as full-sibs. Beyond the 50 - 100 m distance c lass , genet ic structuring remained significant (a = 0.05) but less pronounced up to approximately 600 m , then became non-significant or negative up to 2500 m. Of the cont inuous populations, only Elk Fal ls revealed significant spatial genet ic structure, with a positive coancestry coefficient (p,y = 0.14) for only the 100 - 200 m distance c lass . The overall s lopes of the correlograms for all three fragmented populat ions and for Elk Fal ls were negative and significant (P<0.05), indicating spatial genet ic structure for these populat ions. Simulations Est imates for simulat ions of the observed number of al le les (Ao) and expected heterozygosity (HE) likely to be retained over a 250-year period compared to the current levels are summar ized in Tab le 4.7. The observed number of al leles decl ine slightly faster than expected heterozygosity consistent with theoretical predictions (Nei et a l . 1975; Chakraboty 1980). However the decl ine is much faster when NE = 50 than when NE - 100 (Table 4.7). Based on actual allele f requencies, over 9 0 % of expected heterozygosity would be retained for both fragmented and cont inuous populations over two generat ions (after 250 years) irrespective of whether NE - 50 or NE =100 (Table 4.7). D I S C U S S I O N Effects of fragmentation on genetic variation and inbreeding Habitat fragmentation can cause a loss of population genet ic variation in two ways. First, a transient reduction in population s ize could result in a substantial loss of al leles (Frankel and Sou le 1981; Young et a l . 1996). The extent to which this occurs is dependent on the extent and pattern of forest loss, and its co inc idence with any fine-sca le genet ic structure. A n immediate loss of heterozygosity would , however, only be evident if the population s ize was greatly reduced. S e c o n d , subsequent to this initial allelic loss, f ragmented populations that remain smal l and isolated for severa l generat ions will continue to lose al leles due to genet ic drift, further reducing levels of genetic variation within the stands (Barrett and Kohn 1991; El lstrand and E lam 1993). Heterozygosity is mostly affected by intermediate-frequency al leles (Taggart et a l . 1990), whereas rare al leles are the most likely to be lost in smal l or f ragmented populat ions and high f requency al leles are likely to become f ixed. Near ly all of the dif ferences observed in number of al leles in this study were caused by low frequency al leles (<0.10). ( LAP-2 , 2 being the one exception). There are three likely explanation for the maintenance of genet ic variation in fragmented populat ions: (i) There have been insufficient generat ions s ince fragmentation for detectable loss of diversity through genet ic drift and inbreeding or for mutation and genet ic drift to generate differences among populat ions; (ii) Despite fragmentation, effective population s ize (A/e) remains large so that initial loss of heterozygosity is very smal l , s ince the proportionate reduction in expected heterozygosity AHe = —!—. Even though populations are f ragmented, they could still have hundreds or even thousands of individuals (depending on the neighborhood s ize) ; (iii) G e n e flow is sufficient and there was no isolation by distance in fragmented population. I hypothesized that one effect of fragmentation would be dec reased heterozygosity and polymorphism in seedl ings compared to adults in fragmented populat ions, but did not detect ev idenced of this. O n the contrary, polymorphism was higher in seedl ings than adults in both cont inuous and fragmented populat ions, and most al leles found in seedl ings were common and widespread across all populations, just as in adults, suggest ing high gene flow. Gonza lez -As to rga and Nunez-Far fan (2001) found low f requency al leles in seedl ings of Brongniartia vazquezii a monoec ious, animal poll inated shrub in Central Mex ico , which were not found in adult populations and attributed this to gene flow. There is no reduction of genet ic variation in fragmented compared to continuous populations in either adults or seedl ings, indicating there may be substantial gene flow among fragmented populat ions or that fragmented populations have large effective population s izes (Levin and Kerster 1974; Young e t a l . 1996). Simi lar studies conducted by Young et al . (1993) on Acer saccharum a lso found no reduced genet ic variation in fragmented populations compared to cont inuous populat ions. Comb ined with higher mean levels of polymorphism in f ragmented populat ions, this indicated increased gene flow may be a consequence of fragmentation. Similarly, Fore et a l . (1992) compared inter-population genet ic d ivergence between pre-fragmentation and post-fragmentation seedl ing and adult cohorts in 15/4. saccharum populations in Ohio, U S A . In their study, genet ic d ivergence in post-fragmentation cohorts was less than half that of cont inuous populat ions, indicating a reduction in genetic differentiation s ince fragmentation, and suggest ing increased inter-population gene flow. It appears therefore that maple spec ies may be resilient to fragmentation. The results obtained from this study are somewhat in contrast to similar studies conducted on some angiosperm trees spec ies in the tropics that showed some effects of fragmentation on the overall genet ic structure, (e.g. Swietenia humilis; Whi te et a l . 1999; and Spondias mombin, (Nason and Hamrick 1997)).The authors attributed their results mainly to the demographic and reproductive character ist ics of these spec ies in the tropics. For instance, many tropical trees occur at low densi t ies, are pollinated by animals, have high outcrossing rates, and have breeding sys tems that involve complex mechan isms of self-incompatibil ity (Bawa 1990; Hamrick and Murawski 1990). Inbreeding in adults versus seedlings In this study a high proportion of the a l lozyme loci were not in Hardy-Weinberg equil ibrium, with significant inbreeding (F/s) (Table 4.3). Deviat ions from Hardy-Weinberg expectat ions due to nonrandom mating within f ragmented populat ions of either adults or seedl ings would be expected across all loci. However , the high values of F/s were not consistent for individual loci or within a particular cohort or population type. For instance, adults in all populations had significant inbreeding, but seedl ings in the Elk Fal ls and Maple Bay populations did not differ from Hardy-Weinberg equil ibrium. Two genetical ly unlinked loci (6PG-1 and L A P - 2 ) contributed to the overall high est imates of F / s in both adults and seedl ings. For, instance if these two loci are removed from seedl ings in Rosewa l l , F/s est imated drops from 0.57 to 0.39, reducing the inbreeding estimate by 3 3 % . There are two possib le explanat ions for the high F/s observed in this study. First, if population sub-structuring has been ignored in sampl ing, the inbreeding coefficient would be overest imated, i.e., the patchy distribution of related individuals may generate a Wah lund effect (Barbujani 1987). S e c o n d , there may be a significant amount of inbreeding occurr ing in these populat ions (see chapter three). Fragmentat ion, coupled with local ized pollinator movement and seed d ispersal , may have resulted in higher correlated paternity or selfing for these sampled populations compared to cont inuous populat ions, caus ing a def ic iency of heterozygotes. A s suggested by S h e a (1990), inbreeding could a lso result from differential select ion pressures resulting from micro-environmental variations favoring related individuals. Whi le the ev idence of inbreeding was shown in both cont inuous and fragmented populations for both adults and seedl ings, seedl ings had higher levels of inbreeding compared to their adult cohorts in four of the six populat ions (all but Elk Fal ls and Maple Bay). Posit ive va lues of F/s at the seedl ing stage may be due to partial self ing. In Shorea leprosula, an insected poll inated dipterocarp with the highest isozyme heterozygosity (HE = 0.40) ever recorded in long lived plant spec ies (Lee et a l . 2000), a higher inbreeding level found for seedl ings in natural populat ions compared to adults was attributed to select ion against homozygotes between the seedl ing and adults s tages. However in Eucalyptus regnans, e levated inbreeding found in natural populations compared to seedl ings in a seed orchard was expla ined by spatial genetic structure (Muona et a l . 1990). This a lso could be the case with A. macrophyllum, in which trees in natural populat ions somet imes exist in c lumps. Inbreeding may also be due to mating between relatives in these c lumps. Population structure M e a n FST est imates indicate weak but significant population differentiation in this spec ies , among both fragmented (FST - 0.031) and cont inuous (FST = 0.015) populations (Table 5.5 a , b). This is consistent with earl ier f indings of low inter-population differentiation in a range-wide genetic study of this spec ies (chapter three), and suggests that stands sampled on Vancouver Island, B C , may essent ial ly form single large population with weak within-population structure. Spatial genetic structure The degree of spatial genetic clustering within a population is determined by a variety of genet ic and demographic factors including population s ize , micro-environmental select ion, seed and pollen d ispersal , plant density, temporal variation in population reproductive rates, patterns of competi t ion-induced mortality and other sources of mortality, spatial sca le of gap formation and possib ly other details of the regeneration process (Frankel et a l . 1995). In particular, the magnitude and spatial sca le of genet ic structure is strongly inf luenced by seed d ispersal mechan isms and adult density that character ize individual spec ies (Hamrick et a l . 1993; Hamrick and Nason 1996; Dol igez and Joly 1997; Ka l isz e t a l . 2001). Spat ia l genetic structure in this study revealed apparent dif ferences between cont inuous and fragmented populations. The distribution of genotypes in all f ragmented populat ions was non-random with significant positive va lues for coancestry (p,y) up to 100 m or 200 m whereas in two of the three cont inuous populat ions, genotypes were distributed randomly (non-signif icant coancestry est imates) with weak or no spatial genetic structure. The observed (p,y) va lues for two of the cont inuous populat ions are much less than that expected for full or half-sibs at all d is tances, suggest ing overlapping of seed shadows from maternal parents. The strong spatial genetic structure observed in fragmented populat ions could be due to the high degree of seed production favoring regeneration of seedl ings in the neighborhood of the mother trees as seeds of bigleaf maple are d ispersed by wind and gravity. In contrast, trees of bigleaf maple that grow in cont inuous forests develop narrow a crown that is supported by a stem free of branches for more than half of its total height due to strong competit ion for light (Minore and Z a s a d a 1990). This habit may lead to low seed production or a smal ler seed shadow thereby reducing the number of sibl ings that are likely to develop around the neighborhood of the mother trees (Kelly et a l . 2004). The higher spec ies r ichness (number of spec ies per hectare) of the forest ecosys tem in continuous forests and the habitat of d ispersal agents may also lead to a reduction in spatial genetic structure in cont inuous populat ions. For example, some smal l mammals such as mice, wood rats, squirrels and birds as reported by Fowels (1965), can collect fruits from different bigleaf maple trees and d isperse them in the forest. Hamrick et a l . (1993) and Hamrick and Nason (1996) suggested that plant spec ies with high adult densit ies have weaker f ine-scale genet ic structure than spec ies with lower densit ies. During sampl ing, cont inuous populat ions appeared to have higher spec ies densit ies than fragmented populat ions. Th is study is consistent with the f indings of Gapare and Aitken (2005) for Picea sitchensis in which core populat ions with higher densit ies had no spatial structure, while peripheral populations with lower density had strong spatial genetic structure up to 500 m. In addit ion, V e k e m a n s and Hardy (2004) re-analyzed data for six spec ies and compared spatial genet ic structure for differing population densit ies with spec ies classi f ied as low or high density. In each of the six pairwise compar isons, populat ions with low densit ies consistently revealed spatial genetic structure. T h e s e findings suggest that relatively high population density in cont inuous forests compared to f ragmented populations could have a strong influence on spatial genetic structure. The lack of spatial genet ic structure observed in the two cont inuous populations (Nitinat and Port Alberni) is similar to that observed for a number of other tree spec ies (e.g. Pinus contorta (Epperson and Al lard 1989); Picea mariana (Knowles 1991); Pinus banksiana (Xie and Knowles 1991); and Neolitsea sericea (Chung et a l . 2000)). In contrast, in some tree spec ies with restricted seed d ispersa l , significant spatial genet ic structure was detected. For, example Quercus rubra (Sork et a l . 1993); and Quercus petraeae (Streiff et a l . 1998) exhibit spatial genet ic structure at short spatial sca les which was attributed to large, gravity d ispersed s e e d , as pollen movement by wind is known to be extensive in those spec ies . Strong spatial genet ic structure was also found in tree spec ies featuring restricted dispersal of both seed and pollen due to pollination by smal l insects and seed dispersal by gravity, for example , in Eurya emarginata (Chung et a l . 2000). The high va lues of F/s (Table 4.3) indicate that some populat ions of bigleaf maple exper ience appreciable inbreeding. In the four populat ions where genet ic structure was evident, aggregat ion of genotypes most likely resulted from limited seed d ispersal . Simi lar studies conducted recently by Kevin et a l . (2004) in two Shorea spp. attributed spatial genetic structure to limited seed d ispersal . Computer simulations of fragmentation effects The results obtained from the computer simulat ions suggest that fragmented populat ions would maintained well over 9 0 % of their genet ic variation over a 250-year period even with an effective population s ize of NE = 50. W h e n the effective population s ize is increased to NE = 100 in the simulation both cont inuous and fragmented populations on average retained over 9 7 % of expected heterozygosity. These results suggest that an effective population s ize of at least 100 can effectively minimize the decl ine of genetic diversity for bigleaf maple populat ions. Overal l the simulation projections corroborates my earlier f indings about maintenance of genet ic variation in fragmented populations and might further help explain why fragmented populat ions maintain the same levels of genetic variation as the cont inuous populat ions. It has been suggested that fragmentation represents a significant threat to the long-term survival of many plant spec ies (Templeton et a l . 1990; Young et a l . 1996). In addit ion, it has been argued that erosion in genet ic variation is one of the important consequences that fragmentation may have on plants spec ies that remain in the smal ler patches due to genetic drift, reduction in gene flow, and elevated inbreeding (Templeton et a l . 1990; Young and Merr iam 1994). In this thesis, using both empir ical data and simulat ion projections, I have shown that fragmentation has not led to overall reduction in genet ic variation nor elevated levels in inbreeding at this t ime in Acer macrophyllum and is not likely to in the near future un less populat ions are very smal l and no gene flow occurs . Lastly, I argued that I did not detect any significant impacts of fragmentation on the overall genet ic variation on bigleaf maple populat ions. Th is argument must be taken with caution in view of the fact that with only three fragmented and three cont inuous populations, statistical power was weak. In a power test conducted using the S A S Analyst module ( S A S Inc. 1994) with both a one- and two- tailed test with a = 0.05, statistical power ranged from 6% to 12.5%. However, no trends were seen in overall genetic variation between continuous and fragmented populat ions, so the lack of significant dif ferences did not s e e m to be a function of the power of the statistical tests. Table 4 .1 . Summary of population information for adult trees and seedl ings of Acer macrophyllum. Forest type Populat ion S ize of plot or forest area [ha) No. of trees ana lysed No. of seedl ings ana lysed Cont inuous Nitinat 312 40 50 Elk Fal ls 102 50 50 P. Alberni 112 50 50 Fragmented Maple Bay 170 46 50 Rosewal l 79 45 50 Yel low Pt. 278 50 50 Total 281 300 Table 4.2 a . Al le le f requencies for nine loci for adults in cont inuous and fragmented populat ions of Acer macrophyllum. Cont inuous Fragmented Locus Al le les NIT E L F P A L M B Y R W C Y P T AAT-1 1 0.990 1.000 1.000 0.957 0.978 0.969 2 0.010 0.000 0.000 0.043 0.022 0.031 A A T - 2 1 0.857 0.733 0.900 0.910 0.845 0.944 2 0.143 0.244 0.100 0.090 0.131 0.056 4 0.000 0.022 0.000 0.000 0.024 0.000 IDH 1 0.900 0.809 0.840 0.716 0.784 0.864 2 0.100 0.191 0.160 0.261 0.216 0.136 4 0.000 0.000 0.000 0.023 0.000 0.000 6 P G - 1 1 0.944 1.000 0.920 1.000 0.901 1.000 2 0.056 0.000 0.008 0.000 0.099 0.000 6 P G - 2 1 0.860 0.990 0.920 1.000 0.922 1.000 2 0.140 0.010 0.080 0.000 0.078 0.000 P G M 1 1.000 0.970 1.000 1.000 1.000 0.938 2 0.000 0.030 0.000 0.000 0.000 0.063 PGI -2 1 0.850 0.818 0.859 0.744 0.833 0.878 2 0.000 0.000 0.000 0.023 0.000 0.000 3 0.150 0.182 0.141 0.233 0.167 0.122 G D H 1 1.000 1.000 1.000 1.000 1.000 1.000 L A P - 2 1 0.949 1.000 0.910 0.837 0.956 0.739 2 0.051 0.000 0.090 0.163 0.044 0.261 Note: Populations abbreviations. NIT= Nitinat; ELF = Elk Falls; PAL = Port Alberni; MBY = Maple Bay; RWC = Rosewall Creek; YPT = Yellow Point. Table 4.2 b. Al le le f requencies for nine loci studied for seedl ings in cont inuous and fragmented populations of Acer macrophyllum. Cont inuous Fragmented Locus Al le les NIT E L F P A L M B Y R W C Y P T AAT-1 1 0.900 1.000 1.000 0.925 1.000 1.000 2 0.100 0.000 0.000 0.075 0.000 0.000 A A T - 2 1 0.906 0.768 0.904 0.780 0.917 0.936 2 0.094 0.232 0.096 0.220 0.083 0.064 IDH-1 1 0.833 0.727 0.800 0.739 0.833 0.811 2 0.167 0.273 0.200 0.261 0.167 0.189 6 P G - 1 1 0.851 1.000 0.978 1.000 0.851 0.978 2 0.149 0.000 0.022 0.000 0.149 0.022 6 P G - 2 1 0.956 0.929 0.917 0.952 0.978 0.929 2 0.044 0.071 0.083 0.048 0.022 0.071 PGI-1 1 1.000 0.928 1.000 1.000 1.000 0.901 2 0.000 0.072 0.000 0.000 0.000 0.099 PGI -2 1 0.775 0.761 0.807 0.784 0.788 0.818 2 0.050 0.102 0.023 0.102 0.038 0.023 3 0.175 0.136 0.170 0.114 0.175 0.159 G D H 1 1.000 1.000 1.000 1.000 1.000 1.000 L A P - 2 1 0.939 1.000 0.837 1.000 0.959 0.867 2 0.061 0.000 0.153 0.000 0.041 0.133 3 0.000 0.000 0.010 0.000 0.000 0.000 Table 4 .3 . Genet ic diversity est imates for adults and seedl ings in cont inuous and fragmented populations of Acer macrophyllum. Populations N % P Ho He Continuous Nitinat Adults Seedlings Elk Falls Adults Seedlings P. Alberni Adults Seedlings Mean Adults seedlings 50 47 49 47 50 47 1.6 1.8 1.6 1.6 1.5 1.9 1.6 1.8 60 70 60 50 50 70 56.6 63.3 0.101±0.037 0.091 ±0.042 0.103±0.048 0.124±0.048 0.083±0.030 0.090±0.035 0.095 0.102 0.116±0.037 0.142±.0436 0.122±0.053 0.144±0.056 0.113±0.036 0.134±0.043 0.120 0.140 0.17* 0.39* 0.25* 0.100 ,NS 0.17* 0.36* 0.20 0.28 Fragmented Maple Bay Adults 44 1.7 50 0.104±0.046 0.146±0.055 Seedlings 47 1.6 50 0.109±0.044 0.133±0.054 Rosewall Adults 44 1.7 60 0.095±0.036 0.130±0.043 Seedlings 47 1.8 70 0.076±0.037 0.130±0.042 Yellow Pt Adults 48 1.6 60 0.095±0.029 0.126±0.042 Seedlings 47 1.8 70 0.074±0.032 0.128±0.039 0.34" 0.13 NS 0.20* 0.57* 0.22* 0.41* Mean Adults 1.7 56.6 0.098 0.134 0.25 seedlings 1.7 63.3 0.086 0.130 0^ 37 Table 4.4. Wi lcoxon s igned ranked test for recent bottleneck (Cornuet and Luikart 1996) in Acer macrophyllum populations under the Infinite Al le les Mode l . Number of loci Wi lcoxon test with HE excess Exp H E > Heq H E < Heq Populat ion P P Nitinat 3.11 0.9609 0.0546 Elk Fal ls 2.86 0.5781 0.5000 P. Alberni 2.47 0.7187 0.3437 Maple Bay 3.62 0.3711 0.6796 Rosewal l C reek 2.79 0.9453 0.0781 Yel low Point 2.99 0.6562 0.4218 Note: Exp = expected number of loci with a heterozygosity excess; H E = expected Heterozygosity; Heq = heterozygosity expected at mutation drift-equilibrium. Table 4.5. Genetic diversity statistics for the eight polymorphic isozyme loci for continuous populations (a) and fragmented populations (b). (a) Locus Hf Hs F/s F[T FST AAT-1 0.008 0.008 -0.002 0.001 0.003 AAT-2 0.279 0.272 0.229 0.258 0.038 IDH 0.269 0.271 0.626 0. 622 -0.010 6PG-1 0.111 0.108 0. 075 0.098 0.015 6PG-2 0.133 0.130 -0.100 -0.054 0.042 PGI-1 0.020 0.020 -0.020 -0.001 0.019 PGI-2 0.268 0.268 -0.179 -0.186 -0.006 LAP-2 0.215 0.212 0.392 0.402 0.016 Mean 0.170 0.168 0.140 0.149 0.015 S E 0.175 0.171 0.012 95%CI (-0.077, 0.480) (-0.079, 0.486) (-0.005, 0.035) (b) Locus Hj Hs FJS FIT FST AAT-1 0.063 0.063 -0.025 -0.032 -0.007 AAT-2 0.183 0.182 0.517 0.519 0.006 IDH 0.337 0.339 0.595 0. 600 0.010 6PG-1 0.165 0.154 0. 267 0.281 0.045 6PG-2 0.051 0.048 -0.073 0.000 0.068 PGI-1 0.032 0.032 -0.056 -0.002 0.051 PGI-2 0.283 0.283 -0.167 -0.147 0.017 LAP-2 0.270 0.257 0.377 0.421 0.071 Mean 0.174 0.169 0.178 0.204 0.031 S E 0.185 0.179 0.014 95%CI (-0.070, 0.507) (0.041, 0.518) (0.010, 0.056) Table 4.6. Pai rwise FST between adult f ragmented ( M B Y , R W C Y P T ) and cont inuous (NIT, E L F , P A L ) populations of Acer macrophyllum. NIT E L F P A L M B Y R W C Y P T NIT — E L F 0.024 — P A L 0.009 0.020 — M B Y 0.031 0.016 0.018 --R W C 0.004 0.017 0.005 0.013 — Y P T 0.034 0.029 0.023 0.023 0.047 --Note in parenthesis: Populations abbreviations as in Table 4.2a. Table. 4.7. Percentage of a l lozyme diversity retained over 250-year period based on computer simulations B O T T L E S I M (Kuo and J a n z e n 2003) for adult populations of Acer macrophyllum in f ragmented and cont inuous forests assuming 125-year generat ion length. Populat ions N E = 50 N E = 100 A 0 HE AO HE Continuous Nitinat 92.78 94.01 96.15 97.53 Elk Falls 85.88 95.14 88.57 97.23 P. Alberni 92.67 93.71 96.46 98.00 Fragmented Maple Bay 91.61 95.32 94.95 97.36 Rosewall 91.41 93.77 95.46 97.81 Yellow Pt. 89.60 93.77 96.69 97.08 -130' 52" -125" -120" ®2-": 3R7 I' 43' -130" -125" -120" km 0 50 100 Figure 4 .1 . Geograph ic locations of bigleaf maple populat ions sampled on Vancouver Island. c O '€ o Q. O 0.45 0.40 0.35 0.30 0.25 0.20 0.15 0.10 0.05 0.00 J l o o o o O o o o o o o ^ k> CO 4^  cn 05 bo CD o o o o O o 1 o 1 o 1 o 1 —X o k) o co o O cn o o o CD o CO o o o o o o o o o o o o o cn -v ro CO 4^  cn CD bo CD p o o 6 6 o o o o —». —k o k) o co o 4^  O 09 CD o o bo o CD o o Allele frequency class Figure 4.2. Distribution of allele f requencies for adults (a) and seedl ing (b). Fil led bars are cont inuous populations and open bars f ragmented populat ions. w o c (0 o 0.30 0.20 0.1 0 0.00 -0.1 0 -0.20 Port A lb e mi 4 5 6 L n d i s t a n c e (m) Figure 4.3 (a-c). Spatial correlograms of coancestry coefficients (p,y) for continuous populations of Acer macrophyllum. Dashed lines represent upper and lower 95% confidence limits for pj, under the null hypothesis that genotypes are randomly distributed. e M a p l e B a y in w o c (0 o o 0.30 1 0.20 -0.10 -0.00 -0.10 -0.20 Yellow Point 5 6 7 Ln distance (m) Figure 4.3 (d-f). Spatial correlograms of coancestry coefficients (p/j) for fragmented populations of Acer macrophyllum. Dashed lines represent upper and lower 95% confidence limits for p# under the null hypothesis that genotypes are randomly distributed. Chapter five GENETIC VARIATION AND POPULATION STRUCTURE IN BIGLEAF MAPLE: A COMPARISON OF ALLOZYME MARKERS AND QUANTITATIVE TRAITS I N T R O D U C T I O N Knowledge of genetic variation and population structure is necessary for understanding and conserving the evolutionary potential of populations (Wright 1951). Patterns of genetic variation can be detected at both among and within-population levels (e.g. Hamrick et al. 1992; Xie and Ying 1996). Levels of genetic variation and degree of genetic control also vary among traits, ages and environments (Mullin et al. 1995; Aitken et al. 1995; W u et al. 1995; Xie and Ying 1996). Langlet (1971) noted that the maintenance of intra-specific variation in natural populations of plants is complex. Patterns of geographic variation result from the joint actions of underlying mechanisms that affect the associations between environmental and genetic heterogeneity, such as different selection pressures, levels of gene flow, and genetic drift (Lindhart and Grant 1996). The use of molecular markers has several limitations in providing information that could be used to define conservation and management strategies (Lynch 1996). This is because the primary aim of conservation genetics is to quantify and maintain the evolutionary potential of a species. For this reason, studies should include an assessment of genetic variation for traits affecting fitness, many of which are polygenic (Petit et al. 2000). Most molecular genetic markers are considered selectively neutral, while the pattern of quantitative trait variation is likely to be driven by environmental factors resulting in different selection pressures in different locations (Petit et al. 2000; Lynch et al. 1999). A comparison of molecular and quantitative measures of genetic variation allows insights into the different modes of evolution in sub-divided populations. Studies of patterns of genetic differentiation of quantitative traits are not uncommon in forest trees. However, only a fraction of these allow for the estimation of Qsr, the parameter estimating the portion of total quantitative genetic variation due to among-population differences (Spitze 1993). It is expensive to conduct sufficiently large common garden experiments with trees to include both populations and families within populations to obtain among and within-population genetic diversity estimates. As a result, the availability of such estimates in the literature is biased towards small, short-lived organisms (Ritland 2000). Studies that have reported joint estimates of quantitative (Qsr) and molecular (Fsr) estimates of genetic variation among populations in plants include, Pseudotsuga menziesii (Rehfeldt 1978; Campbell 1986), Populus balsamifera (Riemennschneider et al. 1992), Picea glauca (Li et al. 1992; Jaramillo-Correa et al. 2001), Daphnia obtusa (Spitze 1993; Lynch et al.1999), Clarkia dudleyana (Podolsky and Holtsford 1995), Pinus contorta ssp. latifolia (Yang et al. 1996), Populus tremuloides (Thomas et al. 1997), Quercus petraeae (Kremer et al. 1997), Cerastium arvense (Quiroga et al. 2002), and Pinus pinaster (Gonzalez-Martinez et al. 2002). Merila and Crnokrak (2002), in their meta-analysis comparing such studies of genetic differentiation at marker loci and quantitative traits, found that the degree of genetic differentiation coding quantitative traits (Q Sr) typically exceeds that of presumably neutral genetic markers (FSr)- These results have been attributed to the role of differential natural selection among populations in determining the population genetic structure of quantitative traits. In forests of the Pacific Northwest of North America, bigleaf maple {Acer macrophyllum) is an important component of biodiversity, and a species of growing economic importance. However, breeding programs have not yet been initiated. Common garden experiments including provenance trials are being conducted to screen genetic variation in natural populations and to allow selection of the best available genotypes for reforestation or for breeding (Wright 1976). In addition, provenance research also aims to define the genetic and environmental components of phenotypic variation between trees from different geographic regions (Morgenstern 1997). In this chapter, I use a provenance/progeny common garden experiment to estimate quantitative genetic parameters, and compare genetic differentiation among populations at allozyme loci with quantitative variation. M A T E R I A L S A N D M E T H O D S In 1995 and 1996, seeds from 14 populations were collected from across the portion of the species range of distribution in British Columbia by the Ministry of Forests (Table 5.1 and Figure 5.1). Nine out of the 14 populations were located on Vancouver Island. Populations selected ranged from 48°22' to 50°21' N latitude, 121°23' to 126°35' W longitude, and 14 to 600 m elevation (Table 5.1). QUANTITATIVE TRAITS Seeds from 148 open-pollinated families from 14 provenances in total were sown in 614 Styroblocks® in mixture of peat and vermiculite in early December 1995, and maintained at 5.2°C minimum and 10°C maximum temperature in a greenhouse. Regular misting four times per day continued throughout the germination period. Germination started in mid-January, 1996. Germination of all provenances was nearly complete by the end of February. Seedlings were fertilized in February and March and moved to a cooler greenhouse for acclimatization on March 18. Overall, only about 45.3% of the seedlings germinated. Germination rates ranged from 8-86% among population and 0-100% among families. However there was no clear geographic pattern in germination rate. The common garden test was planted from April 29 to May 1, 1997 at Surrey, British Columbia. The experiment was laid out in a split-plot design in four randomized blocks with provenances as main plots, and five-tree family rows as subplots. A total of 2925 seedlings were planted. Maintenance of the experiment to ensure high survival and good seedling growth included weeding, watering and fencing against deer. Data collection Height growth was measured at the end of the second year in the field (1998), third (1999) and fourth year (2000). Diameter was measured for all provenances at the end of the growing season in year three (1999). Phenological data (bud flush) was monitored and recorded two to three times a week from March 2002 to mid-May 2002. Julian bud flush data was defined by when the first unfolded leaf was observed. With the exception of bud flush, data measurements of height and diameter were made available to me from the B C Ministry of Forests. Analysis Analysis of variance (ANOVA) was conducted using P R O C G L M (SAS Institute Inc. 1990) for height, diameter and bud flush traits using the following general linear models; Y i j k |= n+ Bi + Pj + PBj, + F(P) k ( l , + F ( P ) B k ( i j ) + 8 I P ) (1) Where: Y = measurement of seedling / from family k in provenance j in block / u, = overall mean Bj = effect of block i Pj = effect of provenance j PBjj = interaction effect of block with provenance F(P)k(j) = effect of family within provenance F(P)Bk(jj) = interaction effect of block with family within provenance e = experimental error All effects in the model were assumed to be random. Variance components for all traits were estimated using the P R O C V A R C O M P (METHOD=REML) procedure (SAS Institute Inc. 1990). In addition, the G L M procedure type III sums of squares (SAS, 1990) was used to estimate the proper F-test for family and provenance effects with the null hypotheses (Ho): No family or provenance effects. I used provenance-by-block interaction (PB) as the error term to test provenance effect and family within provenance-by-block interaction (F(P)B) as the error term to test for family effect. The amount of genetic variation in growth traits and bud flush was quantified by estimating the family variances and testing their significance (P < 0.05). Individual tree and family heritabilities were estimated as follows: Individual heritability: h 2 = 3 o 2 F ( P ) / o 2 F ( P ) + o 2 B X F ( P ) + O 2 E (2) Family heritability: h2f = O 2 F ( P / 0"2F<P) + O2BXF<P) lb + a2Eln (3) All variables are defined above with the exception of b and n which are number of blocks and number of trees in plot respectively. For estimating individual heritability, the additive genetic variance was estimated as three times the family variance (instead of four times the family variance) as suitable for half-sib progenies. It is assumed that Acer macrophyllum open-pollinated progenies are more closely related than half-sibs in view of the high inbreeding in this species and the relatively low number of effective pollen parents (Iddrisu and Ritland 2004). The standard errors of heritability estimates were calculated following Dickerson (1969). To assess the associations among traits for both growth and bud phenology (bud flush), genetic correlations (rg) between pairs of traits were calculated following Falconer (1989) as follows: rg = C o v F ( x , y ) / ( o 2 F x o 2 F y ) 1 / 2 (4) where Cov F(x,y) is the family covariance between traits x and y, and o 2 F x and o 2 F y are their corresponding family variances. Cov F ( x , y ) was calculated using the following relationship: C o v F ( x , y ) = ( a 2 F ( x + y ) - a 2 F x - a 2 F y ) / 2 (5) Phenotypic correlations for each pair of traits, as well as correlations between traits and geographic and climatic variables were estimated as Pearson's product moment correlations using the P R O C C O R R procedure (SAS Institute Inc. 1990). Climatic data were obtained using a method developed by Hamann and Wang (2005). Wright's (1951) F-statistics provide a useful measure of the level of population genetic structure at neutral marker loci by quantifying the proportion of total allelic variation found within versus among populations. Similarly, population differentiation for quantitative traits can be estimated using QST (Spitze 1993) which is analogous to the F S T estimate for marker loci. It is estimated as: QST= o 2 G B / ( 2 a 2 G w + a 2 G B ) (6) where O 2 G B is the among population component of variance and o 2 G w the within population component of variance. The neutral expectation for QST is equivalent to Fsrfor selectively neutral genetic markers (Lande 1992). ISOZYME VARIATION Vegetative buds were collected in February 2001 from two of the four blocks (1460 trees) from all 14 populations in the common garden experiment. Buds were stored at -80°C until analyzed by isozyme electrophoresis. The trees sampled from each population for isozyme analysis were the same as those used for quantitative genetic analysis. Electrophoresis buffer systems and loci assayed are those described in chapter three. Genetic data analyses were performed using B IOSYS-2 , a modified version of the BIOSYS-1 program (Swofford and Selander 1981). The following parameters were estimated: allele frequencies, mean number of alleles per locus (A), percent of loci that were polymorphic (%P) (with the most common allele having a frequency of 99% or less), observed heterozygosity (Ho) and expected heterozygosity (HE = 1- Ip, 2, where p, is the frequency of the ith allele). To investigate the extent of population structuring and differentiation, Fsr (Wright 1965) was estimated for individual loci across the 14 populations. R E S U L T S Quantitative traits Provenances differed significantly in growth traits (p<0.001). All growth traits had similar patterns of variation at all ages. Provenance means for height, diameter and bud flush are presented in Table 5.1. The highest growth rates were observed in trees from Hope, Squamish, Port Alberni and Qualicum. The difference between the most productive provenance (Hope) and the least productive (Woss) in terms of height was about 36%. The first bud flush was recorded on Julian day 105 day. All buds in the trial completely flushed by the 129 t h day. In general, flushing was variable among trees within provenances. Woss, Sayward and Owl, the three northernmost provenances (Table 5.1) flushed first and Metchosin, the southernmost provenance, flushed last, this latitudinal trend was weak. For height growth and bud flush, block, population and family within population effects were all highly significant (p<0.001) (Table 5.2). The family variance for height increase slightly from 1% of the total variance for height-2 to 2.3% for height-4 (Table 5.3). Estimates of individual and family heritability for height were relatively low, ranging from 0.15 to 0.18. Family heritability ranged from 0.37 for height-2 to 0.40 for height-4 (Table 5.3). Timing of bud flush had the highest heritability estimate both for individuals (0.21) and families (0.91) (Table 5.3). Estimates of QST values varied from 0.12 for bud flush to 0.26 for height-2 and averaged 0.17. For height, QST values seemed to decrease with age (Table 5.3). There were strong genetic correlations among heights at all ages (Table 5.4). Diameter was also strongly correlated with height growth. Timing of bud flush was weakly and negatively correlated with all growth traits. Phenotypic correlations were strongly correlated among growth traits at all ages and significant at P = 0.01 (Table 5.4). Height growth in all years was correlated with degree days above 5°C and bud flush was mainly correlated with continentality (Table 5.5). Molecular genetic variability Eight of the 10 loci analysed were polymorphic in at least one population. In all populations, G D H and LAP-1 were monomorphic. The percent of loci that were polymorphic (%P) varied among populations between 30% and 60%, averaging 43.5%. The mean number of alleles per locus (A) ranged from 1.3 to 1.5, averaging 1.37 (Table 5.6). The expected heterozygosity within populations ranged from 0.071 to 0.134 and averaged 0.127 across the 14 populations studied. The proportion of inter-population genetic differentiation among populations ( F S T ) indicated that the vast majority of total variation resided within populations, with approximately 9% of the total variation occurring among populations (Table 5.7). Locus specific estimates ranged from 0.0653 for 6PGD-1 to 0.2243 for LAP-2 . DISCUSSION QUANTITATIVE TRAITS The provenances of Acer macrophyllum sampled did not exhibit high germination rate under nursery conditions, with an overall mean of 45.3% and a range from 0% to 86% among provenances. The low germination rate, assuming seeds were healthy and well handled, may suggest that growth conditions in the nursery were not optimal and under such conditions seedlings may not fully express genetic variation at an early age (Bongarten and Hanover 1985). In addition, seedling growth can suffer following transplanting from the nursery, which could also impact expression of genetic differences in early stages (Namkoong and Conkle 1976; Camussi et al. 1995). Provenances like Hope and Chilliwack that showed higher germination rates in the nursery (greenhouse) also showed higher height and diameter growth (Table 5.1). Genetic variation in growth traits for bigleaf maple seedlings both among and within provenances was detected at an early age. Both provenance and family variance components were significant for bud flush and height at all ages but not for diameter. This pattern is similar to that reported for young lodgepole pine (Wu et al. 1995). Although provenance, block and family within population effects were significant for height, the largest variance component was block by family within provenance interaction (Table 5.3). The narrow-sense heritabilties for individual (h 2) and family (h 2 f) were moderate and remained stable with age for height (Table 5.3). The individual heritability estimate for height growth in bigleaf maple in this study is on the low side compared with those reported for forest trees by Cornelius (1994). Franklin (1979) found diminishing heritability estimates for height growth with age, as competition increased with canopy closure. Other studies, however, have reported different age trends of individual heritability for height growth. For instance, Cotterill and Dean (1988) observed an increase in individual heritability for radiata pine (Pinus radiate) following thinning, followed by a decrease. On the contrary, Xie and Ying (1996) reported a decrease followed by an increase after thinning a lodgepole pine (Pinus contorta) early selection test. It is therefore difficult to find a consistent pattern for heritabilities for growth traits with age or silvicultural treatment. Genetic parameters for quantitative traits need to be interpreted with caution, as they are applicable only to the defined base population, reference unit of selection and specific environments where studies are performed (Zobel 1984). Bud flush Several studies of bud flush phenology have reported that it is under moderate to strong genetic control (reviewed in Howe et al. 2003). For example, Howe et al. (2000) and Bradshaw and Stettler (1995) reported that heritability for bud flush was moderate for F 2 hybrid poplar. Other studies indicate that bud flush is under strong genetic control in Douglas-fir (Aitken and Adams 1997), in Populus trichocarpa (Thomas et al. 1997) as well as in other angiosperm and coniferous tree species (Bongarten and Hanover 1985; Chuine et al. 2000) than in bigleaf maple. In this study, bud burst phenology for bigleaf maple varied significantly among families with moderate estimates for heritability. Notwithstanding, geographically based patterns of genetic variation have been observed for bud flush (e.g., Howe et al. 2000). For some species, trees from northern locations and high elevations will tend to flush earlier than those from southern locations, especially in common garden tests, because they have been exposed to shorter frost free seasons in their native environment, leading to selection of genotypes that have either a lower chilling requirements to break bud dormancy, or a lower heat sum or threshold temperature to initiate growth therefore begin growing earlier in a common garden than populations from milder climates in the spring (Farmer 1993). For example, northern provenances (Owl, Sayward and Woss) flushed slightly earlier than the southern provenances on average. If chilling requirements are met, bud flush is mainly in response to heat accumulation in the spring (Lavender 1981). In this study, it is presumed that, chilling requirements were met and thus bud flush timing differences among families may reflect different heat sums or threshold temperature required for bud flush (Li and Adams 1993). This result corroborates the findings of Perry and Wu (1960) from another maple {Acer rubrum), in which buds from northern provenances flushed earlier than southern provenances or at the same time, depending on the temperature. The test site (Surrey, BC) experienced mild winter temperatures, and according to Hunter and Lechowitz (1992), under such natural conditions, the lack of chilling temperature will be less important than the lack of forcing temperature as an agent to speed up bud flush. Genetic correlations There were strong genetic correlations observed among growth traits (Table 5.4) . These high genetic and phenotypic correlations could be due to either pleiotropy or maternal effects (contribution of the maternal parent to the offspring phenotype via some mechanism other than the transmission of genes, e.g. seed size). The presence of maternal effects can bias estimates of seedling genetic variance, heritability and genetic correlations, especially for height (Lambeth 1980). Hence, it would be useful to study growth patterns of bigleaf maple seedlings over more growing seasons, to investigate the extent of maternal effects and age trend in genetic control of growth traits (Lambeth 1980). High age-to-age genetic correlations between growth traits detected in this study suggest that selection for fast growing trees can be done at the early ages. However the interval between ages two, three and four is too short a time to realize significant changes in family ranks with tree age for the tested families. Therefore, caution should be taken when interpreting such genetic correlations, since they might be lower over long intervals (Rweyengeza et al. 2003). Correlations with climatic variables Correlations between geographic and climatic variable were moderate (Table 5.5) which may reflect the capacity of bigleaf maple to adapt to varying environmental conditions (Jaramillo-Correa et al. 2001). Height growth was significantly correlated with mean annual temperature, and degree days above 5°C (DD5) and bud flush correlated with temperature differential (TD). Since differentiation in quantitative traits (Qsf- see below) is observed for these traits, according to Jaramillo-Correa et al. (2001) these quantitative traits may be under differential selection in response to regional differences in climatic factors. For instance, the mean annual temperature (MAT) and degree days above 5°C (DD5) for Owl average 4.79°C and 891 respectively whiles that in Hope was 10.41°C and 2022. As one would expect, trees from the milder climate (Hope) exhibited higher growth rates than Owl (Table 5.1). F S T vs QST Over the past few years, joint estimates of differentiation for quantitative traits and for molecular marker loci have shown two main patterns. Some species, such as Daphnia obtusa (Lynch et al. 1999) and Arabidopsis thaliana (Kuittinen et al. 1997), have a quantitative population structure essentially identical to that for molecular markers suggesting genetic structure for both quantitative and genetic markers is determined by drift and gene flow, whereas other plant species such as Quercus petraeae (Kremer et al. 1997) or Clarkia dudleyana (Podolsky and Holtsford 1995) have highly divergent populations to quantitative traits. The latter pattern is found in coniferous species. Yang et al. (1996) found differences between allozyme (FST= 0.019) and quantitative genetic differentiation for specific gravity (Qsr = 0.133), stem diameter (Qsr = 0.166), stem height (Qsr = 0.195) and branch length (Qsr= 0.161) in Pinus contorta. In this study, estimates of Qsr for five quantitative traits varied from (Qsr = 0.12) in bud flush, diameter, and height-4 to (Qsr = 0.26) for height-2 (Table 5.3). By comparing estimates of differentiation from quantitative traits (QST) and isozymes (Fsr) we can examine whether evolutionary processes involved in quantitative and isozyme variation in Acer macrophyllum are similar or not. In meta-analyses of published results that compared population structure in markers with that of quantitative traits, Mckay and Latta (2002) and Merila and Crnorkak (2001) found that mean QST is typically larger than but poorly correlated with mean F S T across 29 species of plants, vertebrates and invertebrates. Spitze (1993) suggested three possible outcomes from the comparison of FST and QSf. 1) If QST > FST, the implication is that natural selection rather than genetic drift alone must have been involved in shaping or favouring different phenotypes in different populations; 2) if FST = QST, then genetic drift alone could be responsible in the population divergence and this could be evident in smaller populations; and 3) if QST < FST, then it is most likely that natural selection is convergent in that the same phenotypes are favoured in different populations. Comparison of average estimates of Q s r and FST in this study (QST = 0.17 > FST = 0.09) according to Merila and Crnokrak (2001) provides evidence of involvement of differential selection in shaping phenotypic variation in different populations. Growth traits such as height have been reported to be under differential selection in Pinus contorta (Yang et al. 1999) and Picea glauca (Jaramillo-Correa et al. 2001) because individual trees must grow rapidly to escape suppression from competition from neighbouring trees yet have a sufficiently conservative growth pattern to avoid frost injury, the risk of which varies locally. Acer macrophyllum is an early successional species, relatively shade tolerant and growing across a wide range of sites and climatic conditions. Differential adaptation to regional and local patterns of precipitation, temperature and other climatic variables seems to be the explanation for the divergence in these traits. Notwithstanding, it is worth noting that, geographic and environmental scale of sampling will affect the magnitude of QST-Some studies using isozyme markers have reported greater differences between QST and Fsr than the current study (e.g., Prout and Barker 1993; Spitze 1993; Long and Singh 1995; Yang et al. 1996; and Waldmann and Anderson 1998). For several other tree species, Q S T values are relatively low for timing of bud flush but high for growth cessation or timing of bud set (Howe et al. 2003). Merila and Crnorkak (2001) and Latta and Mckay (2002) have reviewed the basic assumption underlying comparative studies of population genetic structure which included assumption of neutrality of allozymes in these comparative studies. It is worth noting that, in some instances genetic variances within populations have been overestimated because of non-genetic (maternal) effects which could lead to a downward bias of Q S r (Waldmann and Anderson 1998). One way to resolve this, as proposed by Merila and Crnorkak (2001), would be to compare the consistency of QST estimates and direct measures of selection in different populations for different traits. Table 5.1. Bigleaf maple populations sampled for provenance trials, and least square means for growth and bud flush traits (with standard errors in parenthesis). Population LAT LONG ELEV HT-2 HT-3 HT-4 DIA BF Metchosin 48.36 123.55 40 135.5 (0.04) 184.9 (0.05) 231.4 (0.09) 32.5 (0.01) 122.5 (0.05) Maple Bay 48.83 123.63 14 132.2 (0.03) 179.5 (0.04) 240.6 (0.06 30.7 (0.01) 120.8 (0.120) Chilliwack 49.15 122.00 142 135.9 (0.04) 181.6 (0.04) 236.6 (0.05) 28.6 (0.01) 121.7 (0.10) P.AIberni 49.26 124.85 15 139.8 (0.03) 192.2 (0.05) 263.2 (0.05) 34.2 (0.01) 121.9 (0.07) Qualicum 49.33 124.36 80 138.2 (0.03) 194.5 (0.04) 243.9 (0.05) 32.7 (0.01) 120.2 (0.06) Hope 49.36 123.38 90 160.2 (0.03) 216.8 (0.04) 247.5 (0.04) 33.5 (0.01) 120.6 (0.20) Courtenay 49.66 125.03 70 126.8 (0.03) 174.9 (0.05) 225.8 (0.05) 30.1 (0.01) 120.9 (0.04) Gold River 49.75 124.73 200 112.2 (0.03) 162.1 (0.05) 201.0 (0.05) 31.4 (0.01) 121.5 (0.8) Squamish 49.78 123.13 50 143.4 (0.03) 193.2 (0.06) 246.3 (0.05) 32.3 (0.01) 120.7 (0.05) Lang Bay 49.78 124.36 25 136.3 (0.03) 186.6 (0.04) 236.7 (0.05) 30.2 (0.01) 120.3 (0.07) Cowichan 49.81 124.21 200 116.0 (0.03) 166.8 (0.05) 233.8 (0.04) 31.9 (0.01) 121.0 (0.20) Woss 50.21 126.58 160 95.8 (0.03) 147.6 (0.05) 235.5 (0.06) 25.1 (0.01) 120.2 (0.03) Sayward 50.31 125.93 50 131.5 (0.04) 180.0 (0.04) 237.9 (0.06) 33.2 (0.01) 119.6 (0.05) Owl 50.35 124.73 600 118.9 (0.04) 171.8 (0.04 213.0 (0.05) 28.6 (0.01) 118.5 (0.05) Note: LAT = latitude (°N), LONG = longitude fW) , ELEV = elevation (m), HT-2 = second year height (cm), HT-3 = third year height (cm), HT-4 = forth year height (cm), DIA = third year diameter (cm), BF = bud flush (Julian days). Table 5.2. A N O V A results for F approximations for the hypothesis of no family or provenance effect. Trait Source DF SS MS F Pr> F B 3 58.24 19.41 154.85 0.0001 P 13 62.78 4.83 38.52 0.0011 HT-2 BxP 39 53.11 1.36 10.86 0.0001 F(P) 134 87.42 0.65 5.2 0.0004 BxF(P) 386 160.12 0.41 3.31 0.0001 B 3 112.85 37.62 210.32 0.0001 P 13 71.96 5.54 30.95 0.0183 HT-3 BxP 39 90.60 2.32 12.99 0.0001 F(P) 134 159.79 1.19 6.67 0.0004 BxF(P) 386 291.72 0.76 4.23 0.0001 B 3 250.22 83.41 283.01 0.0001 P 13 77.85 5.99 20.32 0.0455 HT-4 BxP 39 120.45 3.09 10.48 0.0001 F(P) 134 169.69 1.27 4.30 0.0383 BxF(P) 383 466.50 1.22 4.13 0.0001 B 3 2.30 0.77 126.01 0.0001 P 13 1.59 0.12 20.05 0.0841 DIA BxP 39 2.69 0.07 11.33 0.0001 F(P) 134 4.36 0.03 5.35 0.0001 BxF(P) 386 7.04 0.02 3.00 0.0001 B 3 822.71 274.24 7.61 0.0001 P 13 12625.82 971.22 26.96 0.0150 BF BxP 39 15395.30 394.75 10.96 0.0001 F(P) 134 17591.27 131.28 3.64 0.0285 BxF(P) 386 46934.67 121.59 3.38 0.0001 Note: DF = degree of freedom, SS = sum of squares, MS = mean sum of squares, F = F-value approximation, Pr > F= probability of greater F-values occurring by chance. Table 5.3. Components of variance, individual heritabilities (h2j), family heritabilities (h2f) and population differentiation (QST) among growth and bud flush traits. Trait B P BxP F(P) BxF(P) E h 2 h 2 f QST HT-2 0.028 0.022 0.019 0.010 0.075 0.127 0.15(0.06) 0.37 (0.04) 0.26 HT-3 0.054 0.021 0.031 0.019 0.144 0.184 0.17(0.02) 0.38 (0.01) 0.16 HT-4 0.093 0.019 0.036 0.023 0.165 0.190 0.18(0.05) 0.40 (0.01) 0.12 DIA + 0.001 0.004 0.001 0.005 0.004 0.001 - - 0.12 BF 0.032 0.086 0.058 0.110 0.210 0.780 0.29 (0.01) 0.91 (0.02) 0.12 + : Not significant, all other variables are significant for all effects at the P<0.001. Table 5.4. Genetic correlations (above diagonal) and family phenotypic correlations (below diagonal) between seedling traits for bigleaf maple provenances in British Columbia. Trait HT-2 HT-3 HT-4 DIA BF HT-2 0.99 0.94 0.78 -0.19 HT-3 0.95 0.97 0.79 -0.22 HT-4 0.89 0.93 0.75 -0.15 DIA 0.69 0.70 0.64 -0.11 BF -0.32 -0.41 -0.39 -0.20 Table 5.5. Correlation coefficients between quantitative traits and climatic variables based on 14 provenance means. HT-2 HT-3 HT-4 DIA BF L A T + + -0.47 -0.42 -0.22 -0.35 0.45 E L E V + + -0.44 -0.38 -0.46 -0.42 -0.05 MAT 0.58* 0.51 0.32 0.45 -0.15 TD -0.24 -0.25 -0.10 -0.27 0.55* M A P -0.15 -0.15 0.20 -0.23 0.25 A H M 0.23 0.21 -0.10 0.28 -0.36 DD5 0.62* 0.54* 0.33 0.42 -0.13 Note: * Significant at P<0.05 after sequential Bonferroni adjustment (Rice 1989). ^Abbreviations as in table 5.1. MAT= mean annual temperature, TD = temperature differential, MAP = mean annual precipitation, AHM = annual heat: moisture index, DD5 = degree days above 5°C. Table 5.6. Genetic diversity estimates for 14 juvenile populations of Acer macrophyllum. Population A % P Ho H E Courtenay 1.43 50 0.103 0.126 Hope 1.43 50 0.101 0.134 Sayward 1.31 40 0.090 0.127 Squamish 1.37 40 0.090 0.128 Cowichan 1.37 50 0.090 0.119 Metchosin 1.37 50 0.080 0.133 Maple Mt 1.43 60 0.110 0.111 Owl 1.37 40 0.103 0.123 Woss 1.50 50 0.096 0.138 Pt. Alberni 1.31 30 0.112 0.127 Qualicum 1.37 40 0.092 0.119 Lang Bay 1.31 30 0.079 0.119 Chilliwack 1.31 30 0.109 0.148 Gold River 1.37 50 0.121 0.137 Mean 1.37 43.5 0.098 0.127 Table 5.7. Estimates of Wright's F-statistics for eight polymorphic loci in British Columbia bigleaf maple populations. Locus FIS FIT FST AAT-1 0.721 0.747 0.091 AAT-2 0.050 0.124 0.106 IDH 0.103 0.101 0.106 6PG-1 0.314 0.370 0.085 6PG-2 0.314 0.391 0.126 PGI-1 0.110 0.203 0.108 PGI-2 -0.123 0.114 0.146 LAP-2 0.291 0.322 0.068 Mean 0.222 0.301 0.090 -130' 52' 50' -125' -1 20" Owl 1- G o l Ccu PJO.ua " . Hope -130'- -125' -120' 0 50 100 52' 50" 43' Figure 5.1. Locations of sampled populations of bigleaf maple provenance trials. Chapter Six C O N C L U S I O N S In British Co lumbia there is a trend towards greater understanding and utilization of angiosperm trees, because of their important contributions to the diversity and sustainabil i ty of British Co lumbia 's forest ecosys tems as well as the value of their wood. Respons ib le management and utilization of this hardwood resource could provide employment opportunities in forestry and va lue-added sectors. In addit ion, these trees are a desirable ecosys tem component , adding to the structural and spec ies diversity of British Co lumbia 's forests. Forest fragmentation is a growing problem because of human population growth and land use conversion of forests. Therefore, what we encounter today in some areas are smal l patches of original habitat for spec ies restoration and conservat ion of genetic diversity. In this scenar io, a thorough understanding of genetic p rocesses affecting genes , individuals and populat ions, and thus affecting the persistence of this spec ies in modified landscapes , is essent ia l for designing sound conservat ion pract ices. M y study has contributed towards our understanding of some important components of genet ics of bigleaf maple. By document ing genet ic variation and population structure, both at the quantitative and molecular levels, investigating the mating sys tem, and compar ing genet ic diversity and genetic p rocesses in continuous versus fragmented populat ions of bigleaf maple, I have provided information needed to manage and conserve this spec ies . Major findings In chapter three I showed that natural populations of bigleaf maple harbour moderate levels of genetic variation. However, at the northern range of the spec ies distribution (Jericho and Fraser populations), polymorphism was high yet expected hetrozygosit ies were low compared to more southern populat ions. In addition there was no ev idence of deviation from random mating in northern populat ions, in contrast to populations from the southern portion of the range, which had substantial inbreeding in three out of the six populat ions. Inbreeding in bigleaf maple may result from gei tonogamous poll inations by bumble bees (Bombus spp), from assortative mating, or from mating among relatives. In addit ion, as pollination is mainly by insects, the movement of pollinators among adjacent f lowers within a crown or between adjacent crowns of related neighbours would a lso cause inbreeding or self ing. The low heterozygosity, however, may reflect overall low genetic variation at the spec ies northern range due to genet ic drift or founder effects during postglacial recolonizat ion. A c r o s s the sampled range, populations are only weakly differentiated, suggest ing extensive gene flow or recent d ivergence from a common ancestral population. One alternate hypothesis to explain the low differentiation among populat ions may be the ecological similarity between most of the sites samp led . Th is is supported by the non-signif icant correlation between geographic and genet ic d is tances. A n analysis of mating system found that bigleaf maple populat ions are predominantly outcrossing with no ev idence of biparental inbreeding. Th is result was somewhat surprising s ince most populations have significant levels of inbreeding (F t s>0). However, this estimate of outcrossing rates may be b iased upwards in v iew of the fact I used entirely germinated or filled s e e d s which probably did not account for embryonic lethals due to self ing. Another interesting finding is the ev idence of few pollen donors per seed parent, yet the maintenance of high outcrossing rates. Genet ic study of the effects of fragmentation on plant spec ies so far have shown that forest fragmentation can significantly affect population genet ic p rocesses . Resul ts from my study suggest that both seedl ing and adult cohorts in six populat ions p o s s e s s similar levels of genetic variation regardless of whether population habitats were classif ied as fragmented or cont inuous. This finding suggests extensive gene flow among populations of bigleaf maple. Furthermore, the maintenance of genet ic variation in fragmented populat ions could be attributed to the fact that there have not been a sufficient number of generat ions s ince fragmentation to generate a detectable loss of diversity due to genet ic drift and inbreeding. The most important f inding, however, is the ev idence of spatial structuring of genotypes within all three fragmented populat ions as well as one cont inuous population. I attribute the clumping of genetical ly similar individuals mainly to limited seed dispersal resulting in individuals in c lusters being more related than expected by chance. In this study, spatial genet ic structure appeared to be affected by spec ies density. Coincidental ly, the two populat ions that did not show spatial genetic structure s e e m to have higher population density than the four populations that showed spatial genetic structure. Genet ic variation in seedl ing growth and bud phenology was a lso detected both among populat ions, and among famil ies within populat ions. The substantial within-population variation observed in this study, coupled with the moderate heritabilities and moderate genetic correlations among growth traits and bud f lush, suggest an opportunity for genetic improvement and early select ion for these traits. In addit ion, the significant correlation between quantitative traits and cl imatic var iables in this study s e e m s to suggest that bigleaf maple has adapted to varying environmental condit ions, with natural select ion favouring different phenotypes in different environments. Recommendations There is no doubt that management of fragmented populat ions of plant spec ies has become an important element of biological conservat ion, and this issue will continue to grow. Many of the plant spec ies that are currently recognised as threatened are restricted to smal l habitat f ragments and in situ conservat ion of large cont iguous populat ions within a relatively pristine environment is no longer feasible. But the good news with bigleaf maple is that historic levels of genet ic variation have thus far persisted and it does not appear that the spec ies is in need of immediate conservat ion attention. Having said this, integrating our results with other f indings, one issue that remains contentious is whether genetic p rocesses , primarily genetic erosion and inbreeding, actually play a significant role in reducing the viability of smal l f ragmented populations compared to the risk of habitat loss and assoc ia ted demographic factors. Future studies that examine the effects of fragmentation on plant populations should seek a standardized approach to examine habitat subdiv is ion effects. B a s e d on my findings of how fragmentation affects genet ic variation and spatial genetic structure, mating sys tem, and genetic variation at both the quantitative and molecular levels, the following guidel ines are recommended: 1. Where economical ly feasible, compare original un-fragmented (continuous) populat ions and fragmented populations, as fragmentation is a population level p rocess and not an individual based one. If original habitats do not exist, analyze a number of sites or fragments from smal lest and more isolated to largest less isolated. Ideally, one should sample a sufficient number of sites to have the statistical power to separate out the effect of s ize and isolation. 2. Researche rs should focus on critical aspects of the biology of the plants to be studied as they can provide ev idence of how surviving individuals are pass ing on their genes to the next generat ion or the potential for seed and pollen d ispersal in among fragments. 3. Genet ic diversity measures should be carefully s tandard ized because these measures are very sensit ive to sample s izes . From an analyt ical standpoint, new tools are needed to detect changes especial ly in mating patterns and dispersal rates of plant populations in human dominated landscapes that would otherwise go undetected. Hypervariable codominant markers such as microsatell i tes are highly recommended to shed more light on the total genet ic structure and variability of bigleaf maple. 4. The study of quantitative variation in bigleaf maple needs to be extended to include more test sites as well measurements of more traits and more growing s e a s o n s in order to investigate the degree of genotype-by-environment interaction and juvenile-mature correlat ions. Th is knowledge is essent ia l to guide the establ ishment of breeding and deployment zones and to develop further strategies for genetic resource management and utilization in bigleaf maple. R E F E R E N C E S Acker ley, D.D., and. Donoghue, M. J . 1998. 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G . , and Brown, A . H . D . 1999. Paternal bott lenecks in f ragmented populations of the endangered grassland daisy Rutidosis leptorryhynchoides. Genet ica l Resea rch , 73: 111-117. Y o u n g , A . G . , and Boyle , T . J . 2000. Forest fragmentation. In: Forest Conservat ion Genet ics : Pr inciples and Pract ice. Y o u n g , A . , Boshier . D. and Boyle, T. (eds). C S I R O , Austral ia, pp. 123-134. Z a s a d a , J . C . , Tappeiner II, J . C . , and Max, T .A. 1990. Viabil ity of bigleaf maple s e e d s after storage. Western J o u r n a l of Appl ied Forestry, 5: 52-55. Zobe l , B., and Talbert, J 1984. Appl ied Forest tree Improvement. John Wi ley and S o n s . New York. 505 pp. Appendix I Enzyme, buffer systems and recipes for histochemical staining solutions. Enzyme #of loci Gel buffer Stain components Aspirate Aminotransferase (AAT) 2 Sodium Borate (Ridgeway) 50 ml 0.2 M Tris-HCL pH 8.0 1 mg Pyridoxal 5-phosphate 200 mg L-Aspartic acid 100 mg Ketoglutaric acid 200 mg Fast Blue BB salt 6-Phosphogluonate dehydrogenase (6-PGD) 2 Sodium Borate (Ridgeway) 50 ml 0.2 M Tris-HCI pH 8.0 10 mg Phosphogluconic acid 1 ml NADP 1 ml MTT 1 ml P M S Isocitric dehydrogenase (IDH) 1 Sodium Borate (Ridgeway) 50 ml 0.2 M Tris-HCI pH 8.0 100 ml DL-lsocitric acid 1 ml NADP 1 ml MTT 1 ml P M S Phosphoglucose isomerase (PGI) 2 Morpholine 50 ml 0.2 M Tris-HCI pH 8.0 25 mg Fructose-6-phosphate 1 ml NADP 1 ml MTT 1 ml P M S 1 ml MgCI 2 Leucine Aminopeptidase (LAP) 2 Morpholine 50 ml Aminopeptidase buffer pH 6.0 0.4% L-Leucine 30 mg B-naphtylamide 20 mg Black K salt Glutamete Dehydrogenase (GDH) 1 Morpholine 50 ml 0.1 M Tris-HCI pH 8.0 400 mg Glutamic acid 3 ml NADP 3 ml MTT 3 ml P M S Appendix II. Allele f requency distribution of ten loci of bigleaf maple provenance trials. Locus POP1 POP2 POP3 POP4 POP5 POP6 POP7 AAT-1 (N) 89 63 48 79 85 82 67 1 0.719 0.849 0.882 0.715 0.235 0.634 0.743 2 0.281 0.151 0.118 0.285 0.765 0.366 0.257 AAT-2 (N) 88 62 47 80 85 86 61 1 1.000 1.000 1.000 1.000 1.000 1.000 1.000 2 0.000 0.000 0.000 0.000 0.000 0.000 0.000 IDH (N) 96 72 51 70 79 84 64 1 0.708 0.861 0.686 0.700 0.184 0.673 0.706 2 0.292 0.139 0.314 0.300 0.816 0.327 0.294 6PG-1 (N) 90 73 46 78 88 80 59 1 0.822 0.747 0.685 0.763 0.42 0.587 0.795 2 0.178 0.253 0.315 0.237 0.58 0.412 0.205 6PG-2 (N) 92 70 45 78 91 87 60 1 0.891 0.786 1.000 0.904 1.000 1.000 1.000 2 0.109 0.214 0.000 0.096 0.000 0.000 0.000 P G M (N) 88 72 50 80 91 88 60 1 1.000 1.000 1.000 1.000 0.835 0.966 0.975 2 0.000 0.000 0.000 0.000 0.165 0.034 0.025 PGI-2 (N) 96 72 51 80 91 89 61 1 1.000 1.000 1.000 1.000 1.000 1.000 0.878 2 0.000 0.000 0.000 0.000 0.000 0.000 0.122 GDH (N) 82 73 52 80 92 89 63 1 1.000 1.000 1.000 1.000 1.000 1.000 1.000 2 1.000 1.000 0.000 0.000 0.000 0.000 1.000 LAP-1 (N) 87 72 50 79 89 80 63 1 1.000 1.000 1.000 1.000 1.000 1.000 1.000 2 1.000 1.000 0.000 0.000 0.000 0.000 1.000 LAP2 (N) 84 71 47 80 89 84 66 1 0.917 0.782 0.543 0.781 0.966 0.821 1.000 2 0.083 0.218 0.457 0.219 0.034 0.179 0.000 Append ix II (con't). Locus POP8 POP9 POP10 POP11 POP12 POP13 POP14 AAT-1 (N) 65 102 61 79 65 47 53 1 0.777 0.647 0.746 0.734 0.615 0.649 0.557 2 0.223 0.353 0.254 0.266 0.385 0.351 0.443 AAT-2 (N) 70 101 61 80 65 50 53 1 1.000 1.000 1.000 0.975 1.000 1.000 0.745 2 0.000 0.000 0.000 0.025 0.000 0.000 0.255 IDH (N) 70 102 54 74 62 45 50 1 0.779 0.642 0.611 0.649 0.589 0.678 0.340 2 0.221 0.353 0.389 0.351 0.411 0.322 0.660 6PG-1 (N) 69 104 59 78 65 51 45 1 0.812 0.702 0.737 0.840 0.808 0.608 0.733 2 0.188 0.298 0.263 0.160 0.192 0.392 0.267 6PG-2 (N) 73 106 59 80 64 50 44 1 0.863 0.915 0.822 0.850 0.844 0.610 1.000 2 0.137 0.085 0.178 0.150 0.156 0.390 0.000 PGI-1 (N) 72 107 60 80 60 50 48 1 1.000 0.930 1.000 1.000 1.000 1.000 1.000 2 0.000 0.070 0.000 0.000 0.000 0.000 0.000 PGI-2 (N) 72 106 60 80 60 50 46 1 0.847 0.882 1.000 1.000 1.000 1.000 1.000 2 0.153 0.118 0.000 0.000 0.000 0.000 0.000 GDH (N) 72 106 60 80 62 50 49 1 1.000 1.000 1.000 1.000 1.000 1.000 1.000 2 0.000 0.000 0.000 0.000 0.000 0.000 1.000 LAP-1 (N) 71 101 50 79 62 50 45 1 1.000 1.000 1.000 1.000 1.000 1.000 1.000 2 0.000 0.000 0.000 0.000 0.000 0.000 1.000 LAP2 (N) 72 104 61 78 61 49 44 1 1.000 1.000 1.000 1.000 1.000 1.000 1.000 2 0.000 0.000 0.000 0.000 0.000 0.000 0.000 

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