by colorimetry and total nitrogen using the Kjeldahl-Lauro method (see ...... Kratter, A. W. (1992) Montane avian biogeography in southern California and Baja.
KATHOLIEKE UNIVERSITEIT LEUVEN FACULTEIT LANDBOUWKUNDIGE EN TOEGEPASTE BIOLOGISCHE WETENSCHAPPEN DEPARTEMENT LANDBEHEER LABORATORIUM VOOR BOS, NATUUR EN LANDSCHAP
DISSERTATIONES DE AGRICULTURA Doctoraatsproefschrift nr. 598 aan de Faculteit Landbouwkundige en Toegepaste Biologische Wetenschappen van de K.U.Leuven
FOREST PLANT SPECIES IN FRAGMENTED LANDSCAPES : AN ECOLOGICAL AND MOLECULAR GENETIC APPROACH
Proefschrift voorgedragen tot het behalen van de graad van Doctor in de Toegepaste Biologische Wetenschappen door Hans JACQUEMYN
Maart 2004
Doctoraatsproefschrift Nr. 598 aan de Faculteit Landbouwkundige en Toegepaste Biologische Wetenschappen van de K.U.Leuven
KATHOLIEKE UNIVERSITEIT LEUVEN FACULTEIT LANDBOUWKUNDIGE EN TOEGEPASTE BIOLOGISCHE WETENSCHAPPEN DEPARTEMENT LANDBEHEER LABORATORIUM VOOR BOS, NATUUR EN LANDSCHAP
DISSERTATIONES DE AGRICULTURA Doctoraatsproefschrift nr. 598 aan de Faculteit Landbouwkundige en Toegepaste Biologische Wetenschappen van de K.U.Leuven
FOREST PLANT SPECIES IN FRAGMENTED LANDSCAPES : AN ECOLOGICAL AND MOLECULAR GENETIC APPROACH
Promotor: Prof. M. Hermy, K.U.Leuven Leden van de examencommissie: Prof. J. De Baerdemaeker, K.U.Leuven, Voorzitter Prof. H. Gulinck, K.U.Leuven Prof. B. Muys, K.U.Leuven Prof. L.Triest, VUB Prof. J. Van Assche, K.U.Leuven
Maart 2004
Proefschrift voorgedragen tot het behalen van de graad van Doctor in de Toegepaste Biologische Wetenschappen door Hans JACQUEMYN
This web of time – the strands of which approach one another, bifurcate, intersect or ignore each other through the centuries – embraces every possibility.
Jorge Luis Borges
Time present and time past Are both perhaps present in time future, And time future contained in time past.
T.S. Elliot
Dankwoord The desire for knowledge for its own sake is the one which really counts... Some will tell you that you are mad, and nearly all will say, ‘What is the use?’. For we are a nation of shopkeepers, and no shopkeeper will look at research which does not promise him a financial return within a year. And so you will sledge nearly alone, but those with whom you sledge will not be shopkeepers: that is worth a good deal. If you march your Winter Journeys you will have your reward, so long as all you want is a penguin’s egg.1
Deze woorden zijn afkomstig van Apsley Cherry-Garrard, die met zijn 24 levensjaren de jongste deelnemer was aan Scotts dramatische expeditie naar de Zuidpool en die tijdens de barre winter van 1911 deelnam aan de ‘Winterexpeditie’ die tot doel had een beter inzicht te verwerven in de embryologie van de Keizerspinguïn. Ze zijn mij altijd bijgebleven, en ik hoop, zonder de pretentie te hebben evenveel ontberingen en wanhoop te hebben moeten trotseren, dat dit boekwerk mijn eigen klein pinguïnei is. Gelukkig heb ik niet helemaal alleen gesleed, en de mensen die met mij gesleed hebben, waren allerminst winkeliers of boekhouders. Het lijkt me dan ook een gepast moment om ze daarvoor te bedanken. Grote delen van dit doctoraat zijn voortgekomen uit mijn thesis, die ik onder begeleiding van Jan Butaye heb gemaakt. Hierbij heb ik ruimschoots gebruik kunnen maken van data die door Jan en Myriam Dumortier in het kader van VLINA 97/02 verzameld werden. Het was een vliegende start, en het is door de nauwgezetheid waarmee de data verzameld werden en de kwaliteit ervan dat dit werk is kunnen groeien. Jan wil ik verder bedanken voor alle hulp tijdens het tot standkomen van dit werk en de mooie herinneringen aan de lange zomeravonden op één of ander graanveld. Het onderzoek naar de leefbaarheid van de Slanke sleutelbloem heb ik, naast zoveel andere zijprojecten en afsplitsingen van zijprojecten die in de loop van de afgelopen vier jaar het levenslicht gezien hebben (kalkgraslanden, Orchis purpurea, Primula veris, Hottonia palustris, Paris quadrifolia, Ophioglossum vulgatum, ...) altijd in de beste amicale sfeer met Rein Brys uitgevoerd. Maar het zijn ook hier weer vooral de kleine gebeurtenissen die me zullen bijblijven: de bierkaartjes vol ideeën, de
1
Apsley Cherry-Garrard, The Worst Journey in the World (Constable and Co., 1922).
glimwormen die ons welkom heetten na een lange dag veldwerk, de zoveelste Moeder Overste die de dag beëindigde, ... Olivier Honnay was en is nog steeds een grote hulp bij het uitzetten van de bakens, bij het nalezen van de artikels, en het exploreren van nieuwe onderzoeksthema’s. Het is dankzij hem dat de genetische aspecten van versnippering in het onderzoek betrokken werden. Isabel Roldán-Ruiz en alle medewerkers van het CLO en in het bijzonder Nancy Mergan en Katrien Liebaut zou ik willen bedanken voor hun hulp met de genetische analyses. Isabel verdient verder een pluim voor haar constructieve bemerkingen op verschillende versies van het manuscript. Martin Hermy zou ik willen bedanken voor het vertrouwen en de kansen die ik gekregen heb, voor dat laatste duwtje in de rug om de artikels te bundelen tot een geheel en voor alle praktische raadgevingen om dit werk tot een goed einde te brengen. Ik heb ook bijzonder de vrijheid geapprecieerd waarin ik de afgelopen vier jaar heb mogen werken. Ik zou ook graag mijn appreciatie willen uitdrukken voor alle andere collega’s met wie ik op één of andere manier heb mogen samenwerken, in het bijzonder Patrick Endels waarmee ik de wereld van matrixpopulatiemodellen heb ontdekt, en Kris Verheyen en Beatrijs Bossuyt voor interessante discussies over gerelateerde onderzoeken, Viviane en Sofie voor het afhandelen van alle praktische aangelegenheden. Hans Maes zou ik willen bedanken voor de interessante discussies en wijsheden, die niets met dit onderzoek te maken hadden. Mijn ouders hebben mij altijd gesteund, en ik ben hen daar bijzonder dankbaar voor. Zonder hen zou ik nooit dit doctoraat geschreven hebben. Een speciaal woordje van dank aan mijn vader voor het programmeren van al de software nodig voor het analyseren van de data. En tot slot een bijzonder woord van dank aan Elke, voor alle mooie momenten die we al beleefd hebben en nog zullen beleven, voor je steun en luisterbereidheid, voor je liefde en warmte.
Hans.
Table of Contents Samenvatting ……………………………………………………………………
ix
Summary ..............................................................................................................
xiii
CHAPTER 1 ........................................................................................................
p. 1
General Introduction.
CHAPTER 2 ........................................................................................................ p. 21 Forest plant species richness in small, fragmented mixed deciduous forest patches: the role of area, time and dispersal limitation. Adapted from: Jacquemyn, H., Butaye, J. and Hermy, M. (2001) Journal of Biogeography 28: 801-812.
CHAPTER 3 ........................................................................................................ p. 39 Effects of age and distance on the composition of mixed deciduous forest fragments in a fragmented landscape. Adapted from: Jacquemyn, H., Butaye, J., Dumortier, M., Hermy, M. and Lust, N. (2001) Journal of Vegetation Science 12: 635-642.
CHAPTER 4 ......................................................................................................... p. 55 Influence of environmental and spatial variables on regional distribution of forest plant species in a fragmented and changing landscape. Adapted from: Jacquemyn, H., Butaye, J. and Hermy, M. (2003) Ecography 26: 768-776.
CHAPTER 5 ......................................................................................................... p. 73 Patch occupancy, population size and reproductive success of a forest herb (Primula elatior) in a fragmented landscape. Adapted from: Jacquemyn, H., Brys, R. and Hermy, M. (2002) Oecologia 130: 617-625.
CHAPTER 6 ......................................................................................................... p. 93 Genetic structure of the forest herb Primula elatior in a changing landscape. Adapted from: Jacquemyn, H., Honnay, O., Galbusera, P. and Roldán-Ruiz, I. (2004) Molecular Ecology 13: 211-219.
vii
CHAPTER 7 ....................................................................................................... p. 117 Impacts of restored patch density and distance from natural forests on colonization success. Adapted from: Jacquemyn, H., Butaye, J. and Hermy, M. (2003) Restoration Ecology 11: 417-423.
CHAPTER 8 ....................................................................................................... p. 131 Summary, guidelines for restoration and conservation, and perspectives for future research.
References ........................................................................................................... p. 143 Publications ......................................................................................................... p. 164 Appendix ............................................................................................................. p. 168
viii
Samenvatting
Op het niveau van landschappen (i.e. de regionale schaal) is tot op heden slechts weinig geweten over de verspreiding en dynamiek van plantensoorten en de factoren die deze beïnvloeden. Tal van regionale populatiestructuren werden voorgesteld om de verspreiding van plantensoorten op regionale schaal te beschrijven. Daarnaast is duidelijk geworden dat landschappen allesbehalve statisch zijn, maar sterk onderhevig zijn aan verandering. Dit kan op zijn beurt de verspreiding van plantensoorten beïnvloeden. Zeer algemeen kan verondersteld worden dat het voorkomen van een soort in een dynamisch landschap bepaald wordt door een wisselwerking tussen 1) de oppervlakte habitat die geschikt is voor de soort, 2) de snelheid waarmee soorten migreren tussen geschikte habitatfragmenten en 3) de wijze waarop soorten in staat zijn levensvatbare populaties te behouden in bestaande fragmenten. In deze studie werd de verspreiding van bosplanten in een veranderend landschap onderzocht. Het bosareaal in Vlaanderen is reeds van oudsher onderhevig aan ingrijpende veranderingen, wat geleid heeft tot een sterke afname van het aantal bossen en tot een toegenomen versnippering van de overblijvende fragmenten. De laatste 150 jaar echter werden stelselmatig nieuwe bossen aangelegd. Dit doctoraat heeft als doel een beter inzicht te verkrijgen in de wijze waarop de plantendiversiteit van bossen beïnvloed wordt door lokale (e.g. bodemcondities) en regionale (connectiviteit, geschiedenis) factoren. Plantendiversiteit werd daarbij beschouwd op het niveau van bosplantengemeenschappen, op het niveau van individuele soorten en op het niveau van genen (genetische diversiteit). Op basis van bestaand kaartmateriaal werd de landschapsdynamiek van een 80 km² groot gebied minutieus gereconstrueerd, en werd het komen en gaan van bosfragmenten in kaart gebracht. Een spectaculaire afname van de oppervlakte bos werd vastgesteld tussen 1775 en 1856, waarna geleidelijk aan minder bossen gekapt werden en stelselmatig nieuwe (kleine) bossen werden aangelegd. De soortenrijkdom van bossen nam toe met toenemende leeftijd van bossen. In grote bossen steeg het aantal soorten echter sneller met toenemende leeftijd dan in kleine fragmenten. Verschillen in gemeenschapssamenstelling namen toe met de afstand tussen bossen, wat erop wijst dat de beperkte verbreidingscapaciteit van bosplanten inwerkt op de wijze waarop bosgemeenschappen in versnipperde landschappen ontstaan. Dit werd ix
bevestigd door het feit dat ongeveer de helft van alle soorten gegroepeerd in het landschap voorkwamen. De wijze waarop bosgemeenschappen tot stand komen wordt verder beïnvloed door lokale groei-omstandigheden, maar het relatief belang van lokale en regionale factoren verschilde echter naargelang het bostype. Lokale groeiomstandigheden waren minder belangrijk voor voedselrijke Elzen-Essenbossen, voor voedselarme Eikenbossen daarentegen bleek de invloed van lokale omstandigheden dan weer wel van groter belang op de soortensamenstelling van deze bossen. Bosplanten koloniseerden bosfragmenten beduidend trager dan de snelheid waarmee bossen werden aangelegd. Dit impliceert het bestaan van een tijdsverschil tussen het moment waarop een bos wordt aangelegd en daadwerkelijke kolonisatie. Sommige planten slaagden er zelfs helemaal niet in nieuwe bossen te koloniseren. Voor deze planten zijn de geobserveerde veranderingen in bosstructuur problematisch, omdat populaties die door het verlies van bosfragmenten verdwenen zijn, niet vervangen worden door populaties die recente bossen gekoloniseerd hebben. Dit impliceert dat in landschappen waarin bossen regelmatig gekapt en elders opnieuw aangelegd worden, zelfs al zou het totale bosareaal constant blijven, dit zal leiden tot een lagere soortendiversiteit. Dus, de snelheid waarmee bossen gekapt en opnieuw aangelegd worden, bepaalt mee het voorkomen van plantensoorten op regionale schaal, en moet in beschouwing genomen worden bij toekomstige landgebruiksplannen. Planten in kleine populaties van de Slanke sleutelbloem (Primula elatior) produceerden minder vruchten en zaden per plant dan planten in grote populaties. Dit was vermoedelijk het gevolg van onevenredige verhoudingen van kortstijlige en langstijlige individuen in kleine populaties en een lagere activiteit van pollinatoren. De lage zaadzetting in kleine populaties kan bijdragen tot een verhoogde kans op extinctie op lange termijn. Op korte termijn echter zal extinctie waarschijnlijk niet optreden omwille van de lange levensduur van de soort. Door hun lage zaadproductie dragen deze populaties ook weinig bij tot de verbreiding van genetisch materiaal door zaden. Als daarenboven de uitwisseling van pollen tussen populaties beperkt blijft, glijden deze populaties genetisch verder weg van de meer centraal gelegen populaties en kunnen ze op lange termijn genetisch geïsoleerd geraken. Het oprichten van een bosnetwerk kan de verbreiding van genetisch materiaal positief beïnvloeden, als het merendeel van de populaties bij de uitwisseling van genetisch materiaal betrokken worden. Omdat kolonisatie gemakkelijker plaatsvindt in Elzen-Essen bossen, zal
x
genverbreiding makkelijker kunnen bereikt worden in dit type van bossen. In Eikenbossen daarentegen is het veel moeilijker om tot efficiënte bosnetwerken te komen. Het behoud van bestaande bossen is daarom van uiterst belang voor het instandhouden van levensvatbare populaties van soorten die sterk gebonden zijn aan dit bostype.
xi
xii
Summary
At present, the distributions and dynamics of plants at a large scale (i.e. at the level of landscapes) in a spatial context, and the factors generating these patterns are still poorly understood. Numerous kinds of regional population structures have been proposed to explain the regional distribution of plant species. Moreover, it has become clear that landscapes are nothing but static, but rather are subject to change. The dynamic nature of landscapes may also affect distribution patterns of plant species. Generally, one can assume that the distribution of a species in such dynamic landscapes is determined by the interaction between 1) the amount of suitable habitat, 2) the rate at which species migrate between suitable habitat patches and 3) the way species are capable of sustaining viable populations within habitat patches. In this study, the distribution of forest plant species in a changing landscape was investigated. Forest area in Flanders has been subject to radical changes for a long time past, which has led to a strong reduction of the total amount of forest and an increasing fragmentation of the remnant patches. From 1850 onwards, various new forests systematically have been established in the study area. This thesis aimed at gaining better insights in the way local and regional factors affected forest plant diversity and how forest plant communities were generated over time. To do so, a multi-faceted approach was adopted and three different levels of plant diversity were considered: diversity at the community level, diversity at the individual species level and genetic diversity within species. Based on existing historical maps, land use history of an 80-km² area was meticulously reconstructed, all comings and goings of forest patches were documented and the approximate age of each patch was determined. A spectacular decrease in forest area was observed between 1775 and 1856, after which forest loss gradually decreased and progressively more and more new (small) forest patches were established in the study area. Species richness of these forest patches increased with increasing forest age. Large forest patches, however, received faster more species with increasing age than did small patches. Dissimilarities in community composition increased with increasing inter-patch distances. These results suggest that the limited dispersal capacities of forest plant species determine community assembly of forest patches in a fragmented and changing landscape. Aggregated distribution patterns of xiii
almost half of the species support these findings. Local site conditions further determined species accumulation in forest patches, but the relative importance of local and regional factors differed according to the forest type studied. Local site conditions appeared to be less important for productive Alno-Padion forests, but for less productive Quercion forest patches local site conditions were a major factor affecting community composition of these patches. When considering individual plant species, it appeared that colonization rates of forest plant species were significantly lower than rates at which forest patches were established in the region. This implies the existence of significant time lags between patch establishment and effective colonization of forest patches. Some forest plant species did not succeed at all in colonizing forest patches within 200 yrs. For these species, the observed changes in landscape structures may be problematic, because populations that go locally extinct due to forest loss, are not counterbalanced as species have not yet colonized recently established patches. Even if the total forest area would have remained constant, the observed patch destruction and regeneration has most likely led to substantial losses of populations. Hence, the rate at which a landscape changes, is a major factor that should be invoked in studies investigating regional distribution patterns of plant species. For practical purposes, it should also be considered in future landscape planning and forest monitoring programs. More in-depth investigation of the population viability of the forest herb Primula elatior revealed that plants from small populations produced significantly less fruits and seeds per plant than did plants from large populations, probably as a result of imbalances in pin/thrum ratios in small populations and a reduced pollinator activity. The reduced seed set may increase extinction risks in small populations in the long term. In the short term, however, extinctions may be seriously delayed because of the long life span of the species. Because of the low seed production in small populations these populations also contribute little to gene flow by seeds. The low seed set also suggests that gene flow by pollen is restricted. Small peripheral populations may therefore become gradually genetically isolated from the rest of the metapopulation. The establishment of forest networks may increase gene flow, given that the majority of populations is involved in the process of gene exchange. As colonization was frequently observed in Alno-Padion forests, this can be easier achieved in this forest type than in the less productive Quercion forests for which colonization rates were
xiv
much lower. For the latter, conservation of existing forest patches and special population monitoring programs are needed to assure the long-term population survival of species that typically occur in these forests.
xv
1.
GENERAL INTRODUCTION
CHANGES IN LANDSCAPE STRUCTURE AND FOREST FRAGMENTATION
Forests in Flanders as well as in the rest of Europe or Nothern America have undergone drastic changes during the last several centuries. The county of Flanders may be illustrative for this far-reaching modification of the landscape (Tack et al. 1993). In this region, deforestation had already started before the Roman times. Deforestation was mainly the result of reclamations for agricultural land use in order to satisfy the growing demand for food supplies when population numbers increased. As a result of the continuous deforestations, the amount of forest cover reached a first bottomline around 1300 and was less than 10 % of the total area of the region (Tack et al. 1993). After a slight increase between 1300 and 1775, forest area further decreased between 1775 and 1880 to a dramatically low figure of only 6 % in 1880. This reduction in forest area was accompanied by an ever-increasing fragmentation of the remaining forest area, typified by a reduced area and connectivity of the remnant forest patches. Loss of habitat, decreased individual patch area and connectivity, all have been shown to affect the regional survival of both plant and animal species (Saunders et al. 1991, Tilman et al. 1994, Hanski and Ovaskainen 2000, Ney-Nifle and Mangel 2000, Vellend 2003). Although the global history of the forest area in Flanders is mainly one of forest decline, from 1880 onwards numerous new forests have been repeatedly established on abandoned agricultural land, yielding complex mosaic landscapes with forest patches of different ages, sizes and connectivity. At present, 150,000 ha or almost 11 % of the total area of Flanders are covered by forests (De Keersmaeker et al. 2001). Approximately 127,500 ha (85 %) of the current forest area were established after 1775, date of the Ferraris topographical map. As these new forests were mainly established on former agricultural land or heathlands, it remains an open question whether forest plant species succeed in colonizing and establishing viable populations in these new patches and if so, how species distribution patterns, species richness and community composition change over time. Recent research (Butaye et al.
2
Chapter 1
2001, Verheyen and Hermy 2001, Verheyen et al. 2003a) has shown that colonization does take place, even when newly established forest patches are isolated from existing forests, yet the question how species accumulate over time remains largely unanswered. Moreover, turnover of forest patches and accompanying extinction and colonization may also affect genetic properties of plant populations (McCauley 1991, Pannell and Charlesworth 2000), yet have received little attention in forest research programs. Thus, this work is primarily concerned with the spatio-temporal distribution of forest plant species in a fragmented landscape and focuses at three different levels of variation: variation at the community level, variation at the species level, and variation within the species level.
DISTRIBUTION OF PLANT SPECIES IN LANDSCAPES SUBJECT TO CHANGES
Species distribution patterns in landscapes characterized by frequent turnover of suitable habitat patches may reflect the result of a dynamic interplay between i) the amount of suitable habitat for a species, ii) species-specific migration rates between suitable habitat patches and iii) the ability of species to persist in local habitats.
The amount of suitable habitat
Theoretically, habitat loss and fragmentation may seriously reduce the capacity of a landscape to support viable (meta)populations of a species (Hanski and Ovaskainen 2000, Ovaskainen and Hanski 2001). Above a certain threshold value of habitat loss species are no longer capable of sustaining viable populations and species are predicted to go locally extinct. This threshold value has been called the metapopulation capacity of a landscape. However, it remains an open question whether plants act as real metapopulations and whether application of metapopulation theory with regards to plants is justified (Freckleton and Watkinson 2002, 2003, Ehrlén and Eriksson 2003, Pannell and Obbard 2003). Moreover, it may take some time following habitat loss before extinction will occur, which generates a so-called extinction debt (Hanski and Ovaskainen 2002, Gu et al. 2002). For forest plant species, which are generally characterized by long life spans (Table 1.1) and most of which are able to reproduce vegetatively (Verheyen et al.
General Introduction
3
2003b), time lags between habitat fragmentation and resulting population decline may be substantial (between 50-100 years) (Eriksson and Ehrlén 2001). Therefore, populations of forest plant species may show a very slow response to habitat loss and associated fragmentation.
Table 1.1 Projected life spans (yrs) of some typical temperate forest plant species (from Ehrlén and Lehtilä 2002). Species
Family
Conditional life span
Allium triococcum
Liliaceae
Asarum canadense
Aristolochiaceae
Cypripedium acaule
Orchidaceae
26 yr
Primula vulgaris
Primulaceae
48 yr
41 yr 7 yr
The amount of the original habitat destroyed does not only influence population persistence in remnant patches, it may also determine the recovery of plant species diversity and the return time to equilibrium patch occupancy when new habitat patches are established. For forest plant species, Vellend (2003) showed a non-linear relationship between the proportion of ancient forest in a landscape and relative species diversity in recent forests (i.e. the mean number of species per recent patch divided by the mean number of species per old patch). In landscapes characterized by relatively high amounts of old forest relative species diversity in recent forests was higher than that in landscapes where only a few patches of the original habitat remained.
Seed dispersal and connectivity of patches
Seed dispersal of most forest plant species is generally low. At a local scale, colonization rates ranging between 0.2 and 1.25 meters per year are typical for forest plants species (Matlack 1994, Brunet and von Oheimb 1998, Honnay et al. 1999, Bossuyt et al. 1999), suggesting that seed dispersal is limited. Despite these low figures, this does not imply that occasional long-distance seed dispersal does not occur. Butaye et al. (2001), for example, showed that colonization of forest patches isolated from source patches occurred within a few decades yielding evidence for the
Chapter 1
4
existence of long-distance seed dispersal. However, probabilities of forest plant species colonizing forest patches significantly decreased with increasing distance between target (recently established) and source (occupied older) patches. Several species showed colonization probabilities less than 10 % when the distance between source and target patches increased to more than 200 meters (Fig. 1.1).
0.5
Anemone nemorosa Lamium galeobdolon Arum maculatum Polygonatum multiflorum
Colonization probability
0.4
0.3
0.2
0.1
0.0 0
200
400
600
800
1000
Distance to the nearest source patch (meters)
Fig. 1.1 Colonization probability curves of some typical forest species (from Butaye et al. 2001).
Landscapes characterized by high connectivity may therefore positively impact on dispersal and colonization of plant species, and hence on their regional distribution. Indeed, as connectivity of patches increases, the distance between patches is most likely to decrease, which in turn may increase the probability of diaspores successfully colonizing currently empty patches. The resulting increased population density may in turn increase total seed production and hence the production of dispersal units. Honnay et al. (2002a) nicely illustrated this positive relationship between connectivity and colonization rates when comparing colonization patterns of forest plant species in both a low and high connectivity landscape. For most species,
General Introduction
5
colonization rates were significantly higher in the high connectivity landscape than in the more fragmented landscape. These findings illustrate the importance of landscape connectivity in determining migration rates of plant species in fragmented landscapes.
Local persistence of plant populations in remnant patches
Local persistence is most likely to depend on the area and connectivity of remnant patches. All other things being equal, small habitat patches can only support small populations, which are more sensitive to size fluctuations and stochastic perturbations (Holsinger 2000). Large yearly fluctuations in population sizes may increase extinction risks. Moreover, populations located in small forest patches may also suffer from lower habitat quality due to edge effects (Honnay et al. 2002b), which may lead to reduced recruitment and growth rates (Jules 1998, Jules and Rathcke 1999, Jacquemyn et al. 2003, Vergeer et al. 2003). Further, limited plant population sizes may affect seed set and long-time survival due to reduced pollination success in small and isolated populations (Groom 1998, Richards et al. 1999, Lennartsson 2002). Finally, small populations may also be subject to genetic erosion and increasing genetic divergence among populations (Barrett and Kohn 1991, Ellstrand and Elam 1993, Young et al. 1996). Genetic erosion may decrease the species potential to adapt to environmental changes, whereas increased levels of inbreeding may affect individual fitness and population viability. Both processes have been shown to increase extinction risks of plant species (Newman and Pilson 1997, Keller and Waller 2002). Connectivity on the other hand may also influence population persistence of plant species. Small distances between extant populations may increase pollen flow, which may positively impact on reproductive success (Groom 1998, Richards et al. 1999, Steffan-Dewenter and Tscharntke 1999) and hence the probability of new individuals recruiting in existing populations, which in turn may increase local population persistence (Groom 1998, Harrison et al. 2000, Lennartsson 2002). Increased pollen and seed flow due to increased connectivity between extant populations may also restore genetic diversity in genetically impoverished populations and, as a consequence, population viability and long-term survival (Richards 2000, Newman and Tallmon 2001, Ebert et al. 2002).
Chapter 1
6
Patch turnover, the amount of habitat, connectivity and dispersal
The amount of habitat in the landscape, the rate of change of the amount of habitat and species dispersal capacity associated with landscape connectivity, all may influence persistence of populations in a landscape (Keymer et al. 2000). With and King (1999) and Hill and Caswell (1999), connecting metapopulation theory with percolation theory, concluded that the amount of habitat loss that a species can tolerate depends upon the spatial pattern of suitable and unsuitable habitat. Landscapes characterized by clusters of patches rather than by a random distribution of patches maintain connectivity for a greater amount of habitat loss, thereby enhancing metapopulation persistence. Keymer et al. (2000) further showed that for landscapes characterized by changing amounts of suitable habitat, not only the amount of habitat destroyed influenced metapopulation persistence, but also the rate at which the landscape changed. Landscapes may change too fast in relation to the scale of colonization-extinction processes. However, the effects of landscape changes op population persistence were dependent on dispersal capacities of the species involved. Species with high dispersal capacities were much more able to maintain viable metapopulations within the landscape than species with very low dispersal capacity (Keymer et al. 2000, Johst et al. 2002). Simulation models (Tilman et al. 1997, Huxel and Hastings 1999, Johst et al. 2002) have shown that in landscapes with frequent patch turnover there may be a significant time lag between patch establishment and colonization, especially when new patches are established in a completely random way. Changes in environmental conditions associated with succession may further delay colonization, as newly established patches may not represent conditions similar to that of ancient forest patches (Johnson 2000). However, Verheyen et al. (2003a) recently concluded that environmental conditions of recent forest patches were of minor importance in explaining plant distribution patterns compared to spatial characteristics. Moreover, several seed introduction experiments (Eriksson and Ehrlén 2000, Turnbull et al. 2000, Verheyen and Hermy 2004) have convincingly shown that most forest plant species were perfectly capable of growing in forest patches, once a sufficient amount of seeds had arrived and germinated.
General Introduction
7
SPECIES RICHNESS, THE PARADIGM OF DISTANCE AND TIME
As new forest patches are being colonized, more and more species are arriving at a given site. At a certain point in time, patches may reach their equilibrium and a relationship between species richness and time may emerge, yielding the so-called species-time relationship (Preston 1960). As the area of a patch also determines the number of species a patch can support, it is reasonable to expect that both time and area affect species accumulation in habitat patches. Indeed, Preston (1960) also showed that the rate of species accumulation in time should vary with the spatial scale of sampling. Therefore, it can be hypothesized that the slope of species area relationships should be greater in older patches than in younger ones, because for a given size older patches will be nearer to their equilibrium, while younger patches are still far from it. When species show limited dispersal capacities, as most forest plant species do, a negative correlation between site distance and community similarity may arise (Nekola and White 1999, 2002). This decrease in similarity with distance has been codified by geographers as distance decay, or as the first law of geography (Tobler 1970) and has been observed in several biological systems ranging from boreal spruce-fir forest floras (Nekola and White 1999), birds and mammals in Baja California (Kratter 1992), fen habitats (Nekola 1999), Calfornian serpentine flora (Harrison and Inouye 2002), tropical canopy tree floras (Condit et al. 2002) and macroscopic animals in ponds (Chase 2003). Distance, however, is not the sole determinant of community similarity. Species turnover may also occur as a result of differences in environmental conditions. Thus, if forest patches of different ages display different environmental conditions, it can be expected that forest patch age will also affect patterns of species turnover.
LANDSCAPE CHANGES AND GENETIC DIVERSITY
Biodiversity refers to a broad spectrum of types and levels of biological variation, ranging from the landscape, community-ecosystem to population-species and genetic level (Noss 1997). Each of these levels may be subject to change. Until now, we have
Chapter 1
8
mainly focussed on the consequences of patch turnover on the distribution of plant species, but it may be clear that the genetic properties of plant populations may be altered too. Within landscapes subject to repeated destruction and regeneration of habitat patches, genetic differentiation of colonizing populations may either increase or decrease depending on the number of founding events, the origin of colonists and the degree of migration after colonization. The conditions under which genetic differentiation increased under colonization and extinction events have been theoretically deduced by Wade and McCauley (1988) and Whitlock and McCauley (1990). It was found that genetic differentiation (FST)2 of recent populations was higher than that of older populations when k < 2Nm / (1-φ) + 1/2
(1)
where k is the number of colonizing individuals, Nm the number of individuals that move between populations and φ the probability of common origin. Hence, the question as to whether genetic differentiation increases or decreases, largely depends on the relative importance of colonization and migration rates, and hence on the ecology of the species (Wade and McCauley 1988). Until now, studies investigating the effects of population turnover on genetic differentiation are quite rare. Antrobus and Lack (1993) found that colonizing populations of the perennial herb Primula veris were more differentiated from each other than older populations were. Similarly, Giles and Goudet (1997a,b) found higher differentiation of colonizing populations of Silene dioica (FST = 0.057 and 0.030, respectively). However, measures of the size of colonizing individuals or the probability of common origin are hard to measure in nature, especially for plants. Whitlock (1992), studying genetic differentiation of colonizing populations of the forked fungus beetle (Bolitotheterus cornutus), found that young populations were more differentiated than older ones (FST = 0.112 and 0.040 respectively). The average number of colonists was estimated to be 10.85, whereas φ was 0.5. Ingvarsson et al. (1997) estimated φ = 0.78 and likewise found higher genetic differentiation among young populations than among older populations for the beetle Phalacrus substriatus 2
FST: Wright’s fixation index, measuring the genetic correlation between pairs of genes sampled within a population relative to pairs of genes sampled within the overall set of populations.
General Introduction
9
(FST = 0.090 and 0.059, respectively). The only plant study that we are aware of and that was actually able to estimate both the size of colonizing populations and the probability of common origin was the study of McCauley et al. (1995). Using data on Fst of colonizing and older populations of the weedy herb Silene alba, φ, the probability of common origin, was estimated to be 0.73. Genetic differentation among young populations was 0.197, whereas that among older populations was 0.126. These studies all showed that genetic differentiation was higher among recently established populations suggesting that most colonization events lead to higher genetic differentiation. Only in case of the marine tidepool copepod Tigriopus californicus, older populations were more differentiated from each other than young populations were (Dybdahl 1994).
AIMS AND OUTLINE OF THIS THESIS
The way plant species react to changing landscape structures depends largely on their capacities to maintain viable populations within habitat patches and their ability to colonize new patches. To assess the impact of changing landscape structures on species richness and community organization of forest patches, regional abundance and genetic diversity of forest plant species, a multi-faceted and hierarchical approach was adopted, inspired by the fact that it is nearly impossible to study each plant species separately in exactly the same detail. Therefore, we chose to study the effects of changing landscape structures at three different levels: •
at the community level;
•
at the species level;
•
at the population level.
The major aims of this study are: •
To determine how plant community organization and species richness of forest patches embedded within a fragmented agricultural landscape change over
Chapter 1
10
time and to assess the relative importance of spatial and environmental conditions in governing these changes. •
To investigate how spatial and environmental conditions affect individual plant species’ response to changing landscape structures, and to see whether variation in responses can be attributed to variation in species-specific traits.
•
To provide detailed information on how patch characteristics influence occurrence, population size and reproductive success and hence population viability of plant species.
•
To study how land use changes and associated forest patch turnover affect genetic structure and diversity of plant species.
•
To use this information to provide guidelines for the conservation and restoration of forest patches in our cultural landscapes in order to maintain or, if possible, to increase plant biodiversity.
To achieve these goals within a single four-year study, one specific area was selected as the focus of study. Within this area, minute examination of all available topographical maps made it possible to reconstruct the evolution of the forest cover in time and to determine the approximate ages of forest patches. Incorporation of all maps in GIS allowed visualizing the comings and goings of forest patches within the last 225 years (see further). At the most detailed level (the population level), Primula elatior was chosen as model plant to study the effects of patch turnover on population viability and genetic diversity of plant species in fragmented landscapes. This species was chosen because it frequently occurred in the study area, was able to colonize recent patches and showed considerable variation in population size. Moreover, it is characterized by a genetically based self-incompatibe (distylous) breeding system, which allows testing the importance of demographic stochasticity on population viability.
Chapter 2
Estimating the number of species in a community or ecosystem is a fundamental issue both in ecology and conservation science. Most studies so far ignored the possibility that species number may be as sensitive to the time period of observation as to the
General Introduction
11
area sampled. As forest turnover leads to an amalgamation of forest patches with a different age and area, this allows testing the possibility that species number should scale as a function of time, area and/or their interaction. In Chapter 2, the way forest plant species accumulate in forest patches is the focus of study. The effects of patch age, area and their possible interaction on forest plant species richness are quantified. As the type of forest may affect these relationships, analyses were performed for both forest types separately.
Chapter 3
Two different mechanisms may explain the distribution and abundance of plant species (summarized in White and Nekola 1992). The first (the niche difference model) asserts that the distribution and abundance of plant species is determined by the physical environment and niche characteristics, whereas the second (the model of spatial and temporal constraint) states that spatial arrangements and histories of organisms and habitats are predominant determinants of species distributions and abundances. In Chapter 3, species turnover (i.e. the extent of change in species composition along predefined gradients) is quantified using similarity-distance curves. The relative importance of spatial and historical determinants in governing species turnover is assessed.
Chapter 4
In the previous chapters, effects of forest turnover were quantified at the community level. However, plant species may vary substantially in their response to changes in landscape structure, most likely as a result of differences in habitat requirements, local persistence and dispersal abilities. The resulting pattern of patch occupancy may therefore differ from one species to the next and may reflect the outcome of different dynamics at the regional scale. In chapter 4, the individual response of 59 forest plant species to changing landscape structures are quantified. To gain better insights in the role of plant specific traits in determining how plant species respond to changing landscape structures, relationships between regional abundances and plant traits are also assessed.
12
Chapter 1
Chapter 5
Because of serious time lags between fragmentation and subsequent extinction (time lags in the magnitude of at least 50-100 years are common among long-lived plant species) and between patch regeneration and colonization, Eriksson and Ehrlén (2001) suggested that patch occupancy patterns in present-day landscapes are not in equilibrium. Therefore, studying patch occupancy patterns may not always be sufficient to reveal the more subtle effects of fragmentation on plant species viability. To get better insights into the processes affecting plant population viability of plant species in fragmented landscapes, in chapter 5 the effects of patch area, connectivity and within-patch habitat variables are related to distribution, population size and population fitness of the perennial forest herb Primula elatior.
Chapter 6
To estimate the effects of patch turnover on genetic diversity and structure of forest plant species, a landscape genetic approach (Manel et al. 2003) was adopted. Most studies so far found an increased genetic differentiation among young populations. Primula elatior was again used as model system. Information on historic landscape changes was combined with a population demographic analysis to retrieve the approximate age of populations. Based on this information, genetic diversity and structure were compared among colonizing and older populations using dominant AFLP markers. Results of these analyses also allowed assessing possible ways the species dispersed through the landscape.
Chapter 7
The essence of ecological restoration consists of 1) identifying the major factors that hamper successful restoration of degraded lands and 2) providing tools for speeding the recovery of these lands (Dobson et al. 1997). Honnay et al. (2002c) outlined in detail the major constraints determining success of forest restoration projects. It appeared that, apart from high nutrient levels and major soil disturbances due to former agricultural land use, the limited dispersal capacity of most forest plant species
General Introduction
13
and the low connecitivity of habitat patches appeared to impede succesful restoration. In theory, landscapes with 30 to 60 % of woodland cover are needed to allow easy migration of forest plant species (Peterken 2000). However, it may be clear that such huge amounts of forest cover usually cannot be obtained in our modern cultural landscapes. In chapter 7, the impact of restored forest patch density and isolation on forest species richness of recently restored forest patches and distribution patterns of forest plant species are investigated.
Chapter 8
In chapter 8 the results of the previous chapters are briefly summarized and discussed. Based on this information, guidelines for forest conservation and restoration are presented. Finally, suggestions for future research are offered.
STUDY AREA
All data were gathered in one specific area, which was located in the central part of Belgium (Vlaams-Brabant; 50°51-54’N, 4°53-59’E) between Diest, Tienen and Leuven and comprises an area of approximately 80 km² (Fig. 1.2). Forests are situated partly on alluvial sediments of the river Velpe and its tributaries, partly on the surrounding hills bordering the valley. Altitude varies between 33 and 85 m above sea level. Soils in the valley are mainly histosols and gleyic luvisols. They are characterized by a silty texture, poor drainage, mesotrophic nutrient conditions and are weakly acidic. Phytosociologically forests in this part of the study area generally can be characterized as Alno-Padion forests. Typical forest plant species for this community are among others Ranunculus ficaria, Stachys sylvatica, Geum urbanum, Adoxa moschatellina, Primula elatior, Circeae lutetiana and Arum maculatum. Most frequent tree species are Populus sp., Fraxinus excelsior, Alnus glutinosa, Salix sp. and Quercus robur. Soils on the hills on the other hand are mostly gleyic albeluvisols, have a more sandy loam texture and are more acidic. Former land use was mostly arable land. Forests generally are classified as Quercion forests. Main tree species are Sorbus aucuparia, Betula pendula, Prunus serotina and Quercus robur.
Chapter 2
14
Bekkevoort
Tielt-Winge Kortenaken
Lubbeek
Glabbeek
Boutersem
Tienen
Legend
N
Municipal border Non-forested land Forest area in 1775 Established between 1775 and 1856
1
1775
0
1
2
3
Kilometers
Established between 1856 and 1956
1856
Fig. 1.2 Changes of woodland cover in the study area between 1775 and 1991.
Established between 1956 and 1991
General Introduction
1956 Fig. 1.2 (continued)
15
1991
Chapter 2
16
Forest plant species typically found in these Quercion forests include among others Polygonatum
multiflorum,
Lonicera
periclymenum,
Teucrium
scorodonia,
Convallaria majalis, Luzula pilosa and Maianthemum bifolium. Reconstruction of the land use history by comparing all maps available for the study area (nine in total, dating from 1775, 1856, 1868, 1896, 1937, 1956, 1970, 1982 and 1991) showed that this area has been characterized by pronounced changes in land use and by relatively high rates of forest turnover (Fig. 1.2, 1.3, 1.5). During the 18th and the 19th century, deforestation rates ranged from 0.56 to 0.17 and were 3 - 7 times as large as afforestation rates. During the 20th century, deforestation continued at a lower rate. As a result, total forest area declined from 1947 ha in 1775 to 1002 ha in 1856 and 538 ha in 1956. Forest area reached its lowest point in 1970 when only 464 ha of forest area were found. Only in the last 30 years have afforestation rates exceeded deforestation rates, resulting in a final total forest area of 510 ha in 1991 (Fig. 1.3).
Deforestation Afforestation Total area
Forest area (ha)
2000
1500
1000
500
0 200
Age (yrs)
Fig. 1.6 a) Patch size (ha) and b) age (yrs) distribution of all investigated forest patches (241 in total) in the study area.
20
Chapter 1
2.
Forest plant species richness in small, fragmented mixed deciduous forest patches: the role of area, time and dispersal limitation
ABSTRACT
This research aimed at investigating how plant species richness of small, fragmented forest patches changed over time. Possible interactions between time and area were also studied in relation to species richness. To investigate the relative importance of plant dispersal limitation on the process of species accumulation we examined how plant species were distributed within a regional landscape. To do so, the land use history of a region of 80 km² was reconstructed using nine historical maps dating from 1775 to 1991. Within a central area of 42 km², 241 forest patches were surveyed for presence / absence of 203 plant species that predominantly occur in forests. A Monte Carlo simulation was used to estimate aggregation of species within the landscape. Spatial and temporal patterns of species richness were investigated by both regression and analysis of variance (ANOVA). Fifty-one of 103 species investigated showed significant spatial aggregation, which may suggest that dispersal limitation interferes with the process of community assembly. Thus, dispersal limitation should be invoked as a major factor determining species distributions and hence community composition of forest patches. Species richness of forest patches significantly increased with patch age. However, the effects of time on species richness could not be separated from area as time and area clearly interacted. Slopes of regression equations for species number on area and patch age were shown to be significantly interrelated. Therefore, patch area and age cannot be treated independently as predictors of plant species richness. We conclude that both local and regional factors drive the processes of forest succession and species accumulation. Thus, in order to fully understand the process of forest community assembly, in the future more attention should be directed to the regional factors determining local community composition.
Chapter 2
22
INTRODUCTION
Diaspore dispersal plays a critical role in the spatial distribution of plant species and succession of plant communities in many ecosystems. It is widely accepted that many forest plant species are characterized by low dispersal capacities (e.g. Peterken and Game 1984, Matlack 1994, Grashof-Bokdam 1997, Bossuyt et al. 1999, Butaye et al. 2001). Hence, it can be expected that differences in characteristic dispersal distances between species should have a strong impact on forest plant community development, since local community structure is believed to be largely influenced by local immigration and extinction rates (MacArthur and Wilson 1967, Hanski and Gilpin 1997). However, the role of dispersal in structuring local forest plant community composition remains largely unexplored and until recently surprisingly few studies have been able to report evidence for dispersal-assembled forest plant communities (Ehrlén and Eriksson 2000, Butaye et al. 2001, Honnay et al. 2001, Verheyen and Hermy 2001). Most studies investigating forest plant communities shared a common preoccupation with habitat diversity as the driving force in regulating forest plant community composition and primarily focused on the relationship between species number, patch area and habitat diversity (Peterken and Game 1984, Dzwonko and Loster 1988, 1989, Zacharias and Brandes 1990, Lawesson et al. 1998, Honnay et al. 1999a, b), thereby explicitly assuming that local communities are not dispersalassembled, but rather niche-assembled (sensu Hubbell 1997). These niche-assembled communities are membership-limited assemblages in which interspecific competition for limited resources determines which and how many species invade into and persist in a local community (Tilman 1982). In the latter, composition of the regional species pool and dispersal abilities of these species do not contribute substantially to local diversity. Moreover, processes external to the community may have an important effect on local community structure (Ricklefs 1987, Ricklefs and Schluter 1993). In this case, local species composition and richness are thought to be dependent on accidental dispersal events and the regional species pool, which in turn is influenced by evolutionary and historical processes. Presence or absence of a species is thought
Species-time-area relationships
23
to be highly dependent on the spatial configuration of suitable habitat and the dispersal properties of populations. MacArthur and Wilson’s (1967) island biogeographical theory assumed that all species are characterized by similar dispersal capacities and that every species has the same probability of reaching a target community. For forest habitat patches however, Butaye et al. (2001) showed that forest plant species are characterized by different dispersal capacities, which, in very recent forest patches, resulted in a nonrandom community structure based on dispersal and colonization mechanisms. Given this variety in views that are used to explain local community structure, one may ask how space and time influence species numbers and how species interact to constitute the observed local community. At present, relatively few studies have studied the relationship between the number of species and time within a regional ecological landscape, nor how species number, time and area may interact to produce different outcomes of species diversity. Such studies are needed if the dynamic nature of species diversity is to be understood as was explicitly expressed by Williamson (1988). He noted that ‘because all phenomena are perpetually changing at all scales of space and time, stability and equilibria are at best relative, and need to be studied in the context of particular, clearly stated scales of time and space’. One way to determine the role that both regional and local processes may play in contributing to the species composition of local communities is to investigate how species are distributed in a fragmented landscape and how local communities have been constructed over time. The present study examines both spatial and temporal variation in species richness in forest patches within a highly modified, agricultural matrix. More specifically, this study addresses the following questions: •
How do species accumulate over time to constitute the present-day community?
•
Does the accumulation of species in ecological time follows species-time curves that are similar to species-area curves as was predicted by Preston (1960)?
•
Are there interactions of time and area that produce different outcomes of species richness?
•
Do species show aggregated distribution patterns in space?
Chapter 2
24
•
And if so, is there any evidence that dispersal limitation interferes with successional development so that communities at present are still undersaturated with species?
Answering these questions is important in understanding the process of community assembly, i.e. ‘the process of fitting together the dynamically variable pieces which comprise a system, pieces operating at disparate levels of spatial and temporal scale’ (Drake et al. 1999).
MATERIAL AND METHODS
Data collection
The forest plant community was sampled twice, once in early spring and a second time during the summer period. Presence or absence was determined for 203 forest plant species (Tack et al. 1993, Honnay et al. 1999). This species list only contains vascular plant species that have their optimum in forest habitat. Infraspecific taxa (e.g. Rubus fruticosus), hybrids, aliens, casual species in forests, garden escapes and usually planted forestry trees were omitted so as to decrease the recording bias often present in botanical surveys (Rich and Woodruff 1992). Forest fragments were systematically surveyed by walking transects of several meters wide (cf. Kirby et al. 1986). Because special forest microhabitats like ditches, ponds or forest borders may be inhabited by specialist species, these were given special attention. In total 113 species of the 203 species present on the list were found.
Data analysis
Relationship between species number, patch age and patch area
The relationship between species number, patch age and area was investigated using linear regression and analysis of variance (ANOVA). First, differences in species number between age classes were investigated using a one-factor analysis of variance test. Multiple comparisons were performed to obtain age groups with equal species
Species-time-area relationships
25
numbers. A Monte Carlo simulation was used to obtain groups with equal species numbers, since in the case of unbalanced data this method may be substantially more efficient than other valid methods (Edwards and Berry 1987). The relationship between the number of species, patch area and age was studied using multiple linear regression. In order to meet statistical assumptions (normality, constancy of variance), both age and area were log-transformed prior to analysis. The interaction term age*area was also included in the model. To exclude major variability in habitat characteristics, analyses were performed separately for alluvial and non-alluvial datasets. To examine the relationship between species number and patch area for different age classes, analysis of covariance (ANCOVA) by means of General Linear Models (GLM) was applied. These models provide both regression analysis and analysis of variance. Forest patches were grouped in four age classes (11-35 yr, 51-81 yr, 116-186 yr and ≥ 223 yr) for alluvial forest patches and three age classes (11-51 yr, 81-186 yr and ≥ 223 yr) for non-alluvial forest patches. The assumption that all forest patches from different age classes have equal regression slopes was then tested by fitting a model containing main effects of age and patch area, as well as the age*patch area interaction. The interaction term provides the test of the null hypothesis of equal slopes. Type III sum of squares were used to obtain F-values. In case of clear evidence of inequality of regression slopes, separate slope estimates were assessed by fitting a model containing a main effect of age and an interaction effect of age*patch area. The age*patch area parameter estimates then gave the individual slope estimates within each age class.
Testing for the existence of significant aggregation of individual species
In order to be able to show the importance of dispersal on species diversity, a Monte Carlo simulation was used (Manly 1997) to test the hypothesis that species distributions were spatially more aggregated than could be expected from a random distribution. For each individual species (103 in total, after excluding species which occurred only once (6) and species that occurred more than 150 times (4)), the average distance between all patches occupied by the species was calculated. Next, species were randomly assigned to forest patches and the average distance between
Chapter 2
26
randomly occupied patches was recalculated. Each such repeat thus represented one iteration of the Monte Carlo model. This procedure was iterated 1000 times to produce a frequency distribution of average distances expected under the null model. Then the observed value was compared with the expected distribution under the null hypothesis. If the observed average distance lay in the critical region (α = 0.05) of the distribution produced by the Monte Carlo procedure, the observed average distance was considered to be significantly smaller from that which could be expected by randomly assigning species to forest patches and its distribution pattern was then considered to be significantly clumped within the regional landscape.
RESULTS
Forest plant richness differs significantly among age classes both for alluvial (F = 15.48, P 0.05
0.001
< 0.001
95 % Confidence interval Lower bound
-5.04
10.28
13.45
13.63
-9.42
5.27
13.54
Upper bound
8.31
23.20
27.58
44.36
16.45
19.70
26.94
32
Chapter 2
Table 2.4 Results of the Monte Carlo analysis to see whether species have significant clumped distribution patterns. Number of occupied forest patches, 95 % confidence intervals based on random distributions of average distances and observed mean distance for 103 investigated species. If the observed average distance lies in the critical region (α = 0.05) of the distribution produced by the Monte Carlo procedure, the observed average distance was considered to be significantly different from that expected by randomly assigning species to forest patches and its distribution pattern was considered to be significanlty clustered (bold figures). Species
Number of occupied
95 %
Observed
confidence mean distance
Dispersal mode$
patches
interval (m)
(m)
Adoxa moschatellina
80
3265
3063
end/orn
Ajuga reptans
47
3161
3701
myr
Alliaria petiolata
41
3140
2976
Allium ursinum
2
154
0
myr
Anemone nemorosa
115
3313
3010
myr
Arum maculatum
61
3210
2966
end/orn
Athyrium filix-femina
98
3283
3314
ane
Blechnum spicant
8
2424
1392
ane
Brachypodium sylvaticum
32
3076
2339
ane
Calamagrostis epigejos
11
2701
3068
ane
Calluna vulgaris
6
2181
2248
ane
Campanula trachelium
4
1807
3913
ane
Cardamine amara
2
154
1281
aut/bar
Cardamine flexuosa
20
2931
2238
ane
Carex pallescens
10
2666
2101
Carex pilulifera
33
3084
2584
myr
Carex remota
34
3093
3082
hydr
Carex sylvatica
37
3116
2521
myr
Centaurium erythraea
3
1218
1874
ane
Chaerophyllum temulum
56
3198
3655
ane
$
end/orn: endozoochorous and ornitochorous, myr: yrmecochorous, epi: epizoochorous, ane: anemochorous, aut/bar: autochorous or barochorous.
Species-time-area relationships
33
Table 5 (continued) Circaea lutetiana
63
3216
2982
epi
Cirsium oleraceum
48
3167
3532
ane
Convallaria majalis
31
3068
3270
end/orn
Cornus sanguinea
76
3251
3095
end/orn
Crataegus laevigata
22
2964
3067
end/orn
Crepis paludosa
2
154
115
ane
Cruciata laevipes
11
2701
2603
Deschampsia cespitosa
83
3272
3358
ane
Deschampsia flexuosa
21
2948
2351
epi
Digitalis purpurea
28
3041
2178
ane
Dryopteris carthusiana
86
3277
3446
ane
Dryopteris dilatata
92
3278
3330
ane
Dryopteris filix-mas
107
3301
3405
ane
Epilobium angustifolium
34
3093
2835
ane
Epipactis helleborine
78
3258
3537
ane
Evonymus europaeus
16
2858
2870
end/orn
Festuca gigantea
10
2666
2858
epi
Fragaria vesca
3
1218
3853
end/orn
Frangula alnus
64
3219
3231
end/orn
Geum urbanum
110
3309
3225
epi
Hieracium lachenalii
12
2736
2655
ane
Hieracium laevigatum
5
2060
1484
ane
Hieracium sp.
7
2303
2320
ane
Holcus mollis
56
3198
2667
ane
Humulus lupulus
48
3167
2109
ane
Hyacinthoides non-scripta
3
1218
3072
aut/bar
Hypericum dubium
19
2913
3505
Hypericum humifusum
8
2424
2636
Hypericum pulchrum
4
1807
1544
Hypericum quadrangulum
7
2303
4323
Ilex aquifolium
31
3068
2722
ane
end/orn
34
Chapter 2
Table 5 (continued) Lamium galeobdolon
74
3246
2821
Lapsana communis
101
3286
3251
Ligustrum vulgare
2
154
1033
end/orn
Listera ovata
6
2181
4554
epi
Lonicera periclymenum
122
3319
3131
end/orn
Luzula multiflora
21
2948
2825
Luzula pilosa
36
3109
2793
myr
Lysimachia nemorum
5
2060
949
ane
Maianthemum bifolium
21
2948
2926
end/orn
Melandrium dioicum
16
2858
1418
ane
Melica uniflora
4
1807
1022
myr
Mercurialis perennis
3
1218
65
myr
Milium effusum
12
2736
3346
ane
Moehringia trinervia
84
3275
3126
myr
Narcissus pseudonarcissus
5
2060
1755
aut/bar
Ornithogalum umbellatum
14
2805
3106
myr
Oxalis acetosella
19
2913
2637
aut/bar
Paris quadrifolia
6
2181
1153
end/orn
Poa nemoralis
98
3283
3057
ane
Polygonatum multiflorum
137
3342
3152
end/orn
Polygonum bistorta
2
154
0
Potentilla sterilis
9
2545
1947
Primula elatior
69
3236
3426
aut/bar
Prunus spinosa
62
3213
3093
end/orn
Pteridium aquilinum
33
3084
2912
ane
Ranunculus auricomus
4
1807
4551
epi
137
3342
3485
myr
Ribes nigrum
6
2181
3781
end/orn
Ribes rubrum
145
3349
3543
end/orn
Ribes uva-crispa
36
3109
3402
end/orn
Rosa canina
61
3210
3388
end/orn
Ranunculus ficaria
myr
Species-time-area relationships
35
Table 5 (continued) Rubus caesius
49
3173
3494
end/orn
Rubus idaeus
64
3219
3284
end/orn
Rumex sanguineus.
33
3084
3162
Sarothamnus scoparius
14
2805
2808
Scirpus setaceus
4
1807
3442
Scirpus sylvaticus
20
2931
3405
hydr
Scrophularia nodosa
123
3321
3403
ane
Sedum telephium
5
2060
3784
ane
Senecio ovatus
29
3050
2741
ane
Senecio sylvaticus
9
2545
2783
ane
Stachys sylvatica
138
3343
3424
end/orn
Stellaria holostea
118
3315
3144
myr
Stellaria media
2
154
67
Teucrium scorodonia
75
3247
2914
Torilis japonica
35
3101
3444
epi
Vaccinium myrtillus
3
1218
1565
end/orn
126
3326
3426
end/orn
Vicia sepium
8
2424
755
Vinca minor
8
2424
2673
myr
Viola riviniana
47
3161
3303
aut/bar
Viburnum opulus
end/orn
Species occurring only once: Equisetum sylvaticum, Equisetum telmateia, Hieracium umbellatum, Melampyrum pratense, Sarothamnus scoparius, Solidago virgaurea; species occurring more than 150 times: Corylus avellana, Crataegus monogyna, Hedera helix, Rubus fruticosus, Sambucus nigra.
In the case of species-area relationships, where curves increase with time, dispersal limitation may for a large part explain the observed patterns. This is because, in early successional stages, only very mobile and often long-distance dispersers will be present in most forest patches, whereas less mobile species only occur in the least isolated patches. In this case, it can be expected that patch isolation should have a significant effect on species numbers (Fig. 2.3). As time increases, gradually more species are able to colonize forest patches. MacArthur and Wilson (1963) stated that the number of species increases with area more rapidly on far
36
Chapter 2
islands than on near ones. Old forest patches will be more isolated in terms of species that are able to colonize that patch than young patches for which nearly all species are in the immediate vicinity of the patch.
45 40
Species number
35 30 25 20 15 10 5 0 0
200
400
600
800
1000
Isolation (m)
Fig. 2.3 The relationship between species number in early successional forest patches (age: < 50 yrs) and patch isolation (calculated as the mean shortest (edge-to-edge) distance to the five nearest older forest patches) (data from Jacquemyn et al. 2000).
Preston (1960) was the first to acknowledge and formalize the relationship between time and species number when he suggested that species-time curves would resemble species-area curves. This forest patch research shows that species-time curves themselves are highly dependent on area and that area and time cannot be treated independently because of the effects of dispersal limitation.
Community Assembly, Saturation and Dispersal Limitation
The role of dispersal limitation in structuring local plant communities has only recently begun to be explored (Tilman 1997, Ouborg et al. 1999, Hubbell et al. 1999,
Species-time-area relationships
37
Ehrlén and Eriksson 2000, Condit et al. 2000, Turnbull et al. 2000, Verheyen and Hermy 2001). Based on an extensive literature review of seed sowing experiments, Turnbull et al. (2000) found that in 53 % of the introduction studies (where seeds are sown in unoccupied sites), the species established in at least one of the introduction sites. Approximately 50 % of seed augmentation experiments (where seeds are added to existing populations) contributed to an increase in population size of the species showing evidence for seed limitation. Direct evidence for dispersal limitation in temperate forest habitats was given by Ehrlén and Eriksson (2000), who found that 6 out of 7 temperate forest herbs showed that the availibility of seeds had an important effect on recruitment. In our study, almost half of the species showed aggregated distribution patterns, which may suggest that distribution patterns of these species are strongly constrained by dispersal abilities of species and spatial configuration of sites. More evidence for dispersal infuencing community composition was given by Jacquemyn et al. (2001) who found for the same study area, using partial Mantel tests, that similarities in species composition decreased with increasing inter-patch distances giving the so-called distance decay of similarity (Nekola and White 1999). They concluded that dispersal limitation had important effects on species composition of temperate mixed deciduous forest patches since dispersal or migration was expected to reduce the β component of diversity (Loreau 2000) and to reduce or even eliminate the distance decay of similarity. These results imply that, in order to be able to adequately predict outcomes of assembly sequences, regional processes, at least, should be considered when trying to disentangle patterns in the mechanics of community development. However, it is very difficult, if not impossible to predict community composition entirely. At this point, the question remains as to whether or not forest plant communities are saturated with species. According to Loreau (2000), community saturation occurs primarily because of physical limitations. However, community saturation may also result from dispersal limitation. The latter implies that the observed upper limit to species richness may be nothing more than an illusion, since, in this case, physical limitations do not restrict species richness. Hence, a clear distinction should be made between saturation through time and saturation with respect to the species pool (Morton and Law 1997). There are no grounds to believe that at present forest patches are saturated with respect to their species pools. Slopes
38
Chapter 2
of species-area curves increase as time increases. This may indicate that, as more species have colonized a given patch (pure time-effect), the more fully its niche space is exploited and the more important area becomes limiting species richness, but there is no evidence that communities are necessarily fully saturated. Tilman (1997), studying seed limitation in grasslands, also found that experimental seed addition increased species richness, indicating that, for the time being, grasslands were still unsaturated. The fact that species-area curves keep rising may indicate that species accumulation is still an ongoing process and the upper-limits dictated by physical limits not yet have been reached. As such, forest succession to mature saturated forest patches may be a process that takes at least 200 years. This may be especially true for non-alluvial forest patches (Table 2).
ACKNOWLEDGEMENTS
We’d like to thank Myriam Dumortier for assisting with the data collection and Etienne Jacquemyn for programming the randomization algorithm. Kris Verheyen and two anonymous referees provided useful comments on an earlier version of this manuscript.
3.
Effects of age and distance on the composition of mixed deciduous forest fragments in a fragmented landscape
ABSTRACT
Forest patches in central Belgium were inventoried twice for the presence or absence of forest plant species to study the effects of age and distance on species composition. All forests in the study area were subdivided based on their land use history. To avoid effects of autocorrelated environmental characteristics on species composition, habitat homogeneity was indirectly investigated using a TWINSPAN classification of the vegetation data. Two major habitats (alluvial and non-alluvial forests) were distinguished and analyzed separately. Patterns of species composition were investigated at the landscape level using Mantel tests. Species composition similarity measures were calculated between all pairs of fragments based on the floristic data. A highly significant correlation was found between species composition similarity and inter-patch distance. Difference in age, which we used as a measure for habitat quality, was less important in explaining species composition patterns. The effects of spatial configuration became significant when difference in age was accounted for, and a partial correlation was calculated between inter-patch distance and species composition similarity. Different results were found between alluvial and non-alluvial forest types. Alluvial forests were more influenced by the spatial configuration than the non-alluvial. For the non-alluvial forest type effects measured with the difference in age between forest fragments obscured the effects of inter-patch distance. Based on our findings we suggest that species composition is not only internally generated, but external processes like differential colonization caused by varying degrees of isolation may be of overriding importance.
Chapter 3
40
INTRODUCTION
Islands and terrestrial habitat fragments isolated by the surrounding landscape traditionally have been studied in terms of species richness (MacArthur and Wilson 1967, Williamson 1988, Rosenzweig 1995). Most forest studies also focused on species richness (e.g. Helliwell 1976, Peterken and Game 1984, Dzwonko and Loster 1988, 1989, Zacharias and Brandes 1990, Lawesson et al. 1998, Honnay et al. 1999a). Patterns of species composition and species turnover – both spatial and temporal – have received far less attention (Kadmon and Pulliam 1993, Motzkin et al. 1999, Nekola and White 1999). Yet, spatial and temporal dynamics are as important as local richness in determining diversity at the regional scale (Harrison et al. 1992). It was not until recently that a theoretical framework for turnover-distance change relationships was derived and several hypotheses have been proposed to explain how the similarity of communities changes with distance across a given landscape (Nekola and White 1999) but empirical data are still lacking. Decay of similarity with distance may result from species turnover along local and regional environmental gradients or among habitats. In this case, spatial variations in species composition are believed to reflect the spatial distribution of suitable environmental conditions for growth and survival of the species. However, following studies on Australian birds, Cody (1993) suggested that species turnover not only happens between clearly distinct habitats, but that within habitats it is also depending on the distance between habitat fragments. In this case spatial configuration, spatial context and time are believed to influence species distributions across landscapes, and hence patterns of species composition. This perspective hypothesizes that dispersal is the main factor determining species composition. According to this view, high dispersal rates are believed to prevent differentiation of local communities since highly mobile species will be able to reach almost all appropriate habitats in a landscape. An intermediate view is provided by the metacommunity perspective (Holt 1997). This emphasizes the interplay between species-specific colonization rates, persistence abilities and habitat requirements as controllers of both local and regional species composition and diversity. In this case, both characteristics of the environment and characteristics of the organisms influence turnover-distance change relationships.
Effects of age and distance on species composition
41
With respect to forest plant species and the habitats they depend on, several authors have already shown that many forest plant species are characterized by limited dispersal capacities (e.g. Matlack 1994, Grashof-Bokdam 1997, Butaye et al. 2001) and great persistence abilities (Simberloff and Gotelli 1984, Dzwonko and Loster 1989 and Honnay et al. 1999b). Hence, it can be expected that distance will have an impact on species composition whereas area is of minor importance. Furthermore, it has been shown that intensity and duration of former land use may have effects on forest habitat quality. Intensive land-use may cause differences in soil conditions, which may affect establishment of forest plant species (Koerner et al. 1997, Bossuyt et al. 1999, Honnay et al. 1999c) suggesting that environmental conditions influence species composition. In this study we wanted to test the relative importance of both spatial and environmental (reflected by forest age) factors in determining patterns of species composition. Since a large part of the forest area in Flanders consists of small forest fragments of different ages and land use histories, embedded in a hostile agricultural landscape matrix they provide excellent opportunities for the study of turnoverdistance change relationships. More specifically, we wanted to test the hypothesis that: (1) among-patch variation in species composition is mainly related to differences in inter-patch distance; (2) among-patch variation is caused by differences in age; (3) among-patch variation in species composition results from a complex interaction of both spatial and temporal aspects of fragmentation. The results of this study are important because understanding the ecological and geographical factors that are responsible for the way species replace each other in a fragmented landscape may provide a framework for the development of more comprehensive, space-oriented guidelines for afforestation and forest conservation.
MATERIAL AND METHODS
Study area
The forests under investigation are situated in the central part of Belgium between Tienen, Diest and Leuven. Forests are situated partly on alluvial sediments of the river Velpe and its tributaries. The remainder of the forests is situated on the surrounding
Chapter 3
42
hills, which have mainly sandy-loam soil textures. Topographical altitude ranges from 33 to 85 m above sea level and slopes range from 1 to 16 %. The study area (80 km²) has a present-day forest cover of 6.3% and 44% of this has been continuously forested since at least 1775 (survey date of the historical Ferraris topographical map). The other 56 % of the forest area has been reforested following intermittent agricultural use. Two prior land use forms of reforested areas were found in the study area: grasslands, used as meadows or pastures and arable land. Most forests in the area are deciduous forests, currently managed as tall forest. Half of the forests are homogeneous planted with Populus x canadensis. Other, less frequent, dominant tree species are Alnus glutinosa and Quercus robur. In the older forests a neglected understorey coppice layer is often found, consisting mainly of Alnus glutinosa and/or Fraxinus excelsior.
Forest fragments as sample units
Since this study aims to investigate the absolute and relative importance of spatial habitat configuration as well as forest age on plant species composition, we divided forests into forest fragments based on historical land use. Land use was reconstructed for all forests from 1775 to 1991 using historical maps published in 1775, 1850, 1868, 1896, 1930, 1956, 1970, 1982 and 1991 (scale 1:20000 or 1:25000). All historical maps were digitized and a Geographical Information System (GIS) approach (using the Arc/Info software; ESRI 1995) was used to compare them, resulting in forest fragments with a homogeneous land use history. Data collection used the forest fragment as the sampling unit. Since the historical land use may still have important effects on present-day environmental conditions and hence on distribution patterns of plant species, subdividing forests in forest fragments allowed us to integrate past historical land use in our analysis.
Floristical and Environmental Data Collection
A 42 km² area was selected to study the effects of age and distance on plant species composition. All investigated forest fragments (241 in total) were inventoried twice,
Effects of age and distance on species composition
43
in early spring and summer of 1998, for the presence or absence of 203 forest plant species (i.e. native species for northern Belgium, mostly occurring in forest habitat) (Tack et al. 1993, Honnay et al. 1999a). Forest fragments were systematically surveyed by walking transects several meters wide. In this way more time was spent in larger forests. Special attention was given to ditches, ponds and forest edges. Environmental data were partly collected by overlaying the forest fragments with the Belgian soil survey map (scale 1:20000) (IWT 1995). Due to financial constraints 100 forest fragments with homogeneous soil morphological characteristics were selected for measurement of soil characteristics. During the summer five bulk soil samples were taken from each selected fragment, one in the center of the forest fragment and four at random distances between the center and the border of the forest fragment in perpendicular directions. After litter removal, samples were taken from the upper 10 cm. Samples were oven dried, crushed in a mortar and sieved through a 1.8-mm sieve. Soil samples were analyzed for pH (potentiometric determination in CaCl2-solution), organic matter (Walkley and Black), total nitrogen (i.e. organic and ammonia nitrogen) (Kjeldahl) and plant available phosphate (extraction with ammonium-lactate; determination with spectrophotometer). To quantify the degree of isolation of forest fragments the shortest (edgeto-edge) distance between all pairs of forest fragments was calculated using the Arc/Info GIS (ESRI 1995).
Data analysis
An important characteristic, which may obscure the effects of inter-patch distance and differences in age on species composition similarity is habitat per se. Since coexisting plant species can be assumed to have more similar habitat requirements than expected from a random sample, classification of all species into broadly defined forest types is a good method to explore habitat homogeneity. Secondly, species significantly confined to a specific forest type will rarely be able to colonize and establish in the other forest type. To derive major forest types in the study area a polythetic divisive classification (Hill 1979) on the floristic data of the 241 forest fragments was performed. Classification was restricted to the first division. Since the lower divisions
Chapter 3
44
were related to differences in forest age, they were not considered further. The first division was used to calculate some mean environmental characteristics allowing interpretation and characterization of the forest types in terms of habitat characteristics. A Mann-Whitney test was used to compare environmental variables among forest types. Mantel’s randomization was used to study the absolute and relative importance of inter-patch distance and age on forest plant species composition. These tests allow estimation of the association between two or three similarity matrices describing the same set of entities and test whether this association is stronger than one would expect from chance even if the elements in each of the matrices X and Y are not independent (Dietz 1983, Smouse et al. 1986, Manly 1997, Motzkin et al. 1999). Three types of square matrices were constructed in which each element represents the similarity between pairs of forest fragments. The matrices contained inter-patch distances (X1), differences in forest age (X2) between all pairs of fragments and similarities in species composition (Y). For the latter, four different matrices of resemblance were calculated, based on forest species composition using different similarity measures. This reduces the likelihood of the results being dependent on the specific properties of the index. Four similarity measures (cf. Kadmon and Pulliam (1993)) were used: Jaccard (J) and Sörensen (S) coefficient of resemblance (1, 2) and the Simple Matching (SM) and Sokal and Sneath I (SS) coefficient of resemblance (3, 4).
J =
a a+b+c
SM =
a+d a +b+c +d
S=
2a 2a + b + c
SS =
2( a + d ) 2 a + b + c + 2d
(1, 2)
(3, 4)
where ‘a’ = number of common species between forest fragment i and j, ‘b’ = number of species in fragment i and absent in fragment j, ‘c’ = number of species in fragment j and absent in fragment i, ‘d’ = number of species occurring in one of the studied fragments but not in fragment i and j.
Effects of age and distance on species composition
45
They differ in two main properties (Janson and Vegelius 1981): first whether only joint presences (J and S) or also joint absences (SM and SS) occur in the numerator, and secondly whether matches are double weighted (S and SS) or not (J and SM). Smouse et al. (1986) developed a technique, which makes it possible to assess the relative degree of fit of two hypothetical matrices, e.g. X1 and X2, to a third,
Y. This method calculates the partial correlation between matrices X1 and Y, keeping the effect of matrix X2 constant, analogous to partial regression coefficients calculated in multiple regression analysis. This technique - referred to as extended Mantel test was used to calculate the effect of isolation on species resemblance controlling for forest age (YX1.X2). Analogously the partial effect of forest age on species resemblance was calculated controlling for isolation (YX2.X1). All Mantel tests were performed with the MacIntosh version of the R package (Legendre and Vaudor 1991). Matrices of resemblance were calculated with the windows version 9.0 of SPSS (SPSS Inc. 1999).
RESULTS
The region is characterized by a complex land-use history resulting in a strongly fragmented landscape with numerous small forests, surrounded by an agricultural matrix. Historical forest area reconstruction starting from 1775 until the present revealed a sharp decrease in forest area during the second half of the 18th and the first half of the 19th century. The percentage of the study area covered by forest decreased from 25% to 13% between 1775 and 1856. During the next 12 yr deforestation continued at an even higher rate resulting in a forest cover of only 6 % by 1868. During the 20th century the forest area remained quite stable compared to earlier dates, showing a slight decline over the last 100 yr. Since 1970 forest area has increased very slowly probably due to forest establishment on abandoned marginal agricultural land. The decline in total forest area is typically characterized by decreasing maximal and average patch area and by increasing isolation. The overlay of all historical maps resulted in forest fragments with a unique forest age and former land use history. Within the selected area of 42 km² a total of 241 fragments covering a total area of 276 ha was inventoried and 113 species
Chapter 3
46
of a possible 203 were recorded. Classification of the vegetation data identified two distinct forest types (Table 3.1), which are discussed below.
Table 3.1 List of differential species for the alluvial and non-alluvial forest types with frequency (%) in the two major forest types. The significance of association was tested by a χ²-test (***: P < 0.001). Species Ranunculus ficaria Stachys sylvatica Geum urbanum Adoxa moschatellina Primula elatior Cornus sanguinea Arum maculatum Circaea lutetiana Prunus spinosa Rubus caesius Chaerophyllum temulum Ajuga reptans Alliaria petiolata Carex sylvatica Brachypodium sylvaticum Rumex sanguineus Scirpus sylvaticus Polygonatum multiflorum Lonicera periclymenum Frangula alnus Teucrium scorodonia Holcus mollis Epilobium angustifolium Convallaria majalis Pteridium aquilinum Ilex aquifolium Luzula pilosa Digitalis purpurea Maianthemum bifolium Carex pilulifera Deschampsia flexuosa Cytisus scoparius Hieracium spp.
Alluvial forests Non-alluvial forests 86 *** 18 81 *** 25 60 *** 27 51 *** 9 47 *** 4 46 *** 12 41 *** 4 38 *** 11 35 *** 14 34 *** 2 33 *** 11 28 *** 8 28 *** 2 23 *** 6 22 *** 3 21 *** 4 14 *** 1 71 *** 47 38 69 *** 16 41 *** 9 61 *** 7 46 *** 7 24 *** 7 22 *** 5 26 *** 5 24 *** 4 30 *** 4 22 *** 17 *** 3 1 31 *** 1 19 *** 1 12 *** 1 11 ***
Effects of age and distance on species composition
47
The first forest type is mainly situated on silty soils close to the river Velpe and its tributaries, hereafter called alluvial forest type. These forests are poorly drained and have high soil moisture content. Soil chemical analyses are shown in Table 3.2. Forests are in general relatively young as a consequence of the large number of recently established forests. Fragments of this forest type were mainly used as grassland before forest established. A relatively high number of forest plant species (17) was highly significantly (P < 0.001) confined to this forest type (Table 3.1). Phytosociologically this forest type is related to the Alno-Padion forests. The second forest type is situated on soils with a sandy loam texture on well-drained elevated sites and is referred to as non-alluvial. Chemical analyses of the soil reveal significantly lower pH-values and total nitrogen content. Plant available phosphate and organic matter are not significantly different, although the mean C/N ratio is higher (Table 3.2). Former land use was mostly arable land. 16 forest plant species were highly significantly (P < 0.001) confined to this forest type and phytosociologically these fragments belong to the Quercion-alliance (Table 3.1).
Table 3.2 Mean values (±
SE)
of several environmental variables characterizing the
alluvial and non-alluvial forest types. Differences between the two forest types were tested using the Mann-Whitney U test. Soil chemical data refer to a depth of 10 cm. Alluvial
Non-Alluvial
(n = 138)
(n = 103)
Patch area (ha)
1.0 ± 1.2
1.3 ± 1.8
-1.37 (n.s.)
Perimeter (m)
431 ± 288
497 ± 366
-1.37 (n.s.)
Forest age (years)
69 ± 59
124 ± 79
-5.10 ***
Period of grassland use (years)
101 ± 80
9 ± 28
-9.44 ***
Period of use as arable land (years)
35 ± 51
58 ± 61
-3.18 **
Soil acidity (pHCaCl2)
4.9 ± 0.9
3.4 ± 0.4
-4.52 ***
Total nitrogen content (ppm)
4012 ± 1585
3192 ± 1986
-3.05 **
Plant available phosphate (ppm)
17.3 ± 13.6
16.9 ± 14.7
-0.89 (n.s.)
C/N ratio
12.4 ± 1.2
16.6 ± 2.9
-6.19 ***
Soil organic matter (%)
9.9 ± 4.2
10.4 ± 4.0
-0.51 (n.s.)
(n.s.): not significant; * : P ≤ 0.05; ** : P< 0.01; *** : P ≤ 0.001
Z value
Chapter 3
48
The results of the Mantel tests show a significant negative correlation between species similarity and inter-patch distance for the alluvial forest type (Table 3.3a). Forests composition resemblance decreases with increasing distance. There is no correlation between floristic similarity and forest age for the alluvial forests. Since the four similarity measures reveal the same conclusions they seem to be reliable, suggesting that at least for the alluvial forest type presence and absence of species are, in the first place, determined by inter-patch distance rather than by forest age. For the non-alluvial forests results vary according to the similarity coefficient used. Only the Jaccard coefficient yielded a significant result showing a decrease in species composition similarity with inter-patch distance (Table 3.3b). No significant results were obtained when similarity coefficients were used that take joint absent species into account (SM and SS). Contrary to alluvial forest patches, species composition similarity appeared to be much more related to differences in forest age for the non-alluvial forests. Indeed, similarity matrices based on both the Jaccard and Simple Matching coefficient were significanty (P < 0.001) related to the age matrix (Table 3.3b).
Table 3.3 Mantel correlation coefficients between species composition similarity (using four similarity measures) and inter-patch distance or age for the two major forest type: (a) alluvial forests (n = 138), (b) non-alluvial forests (n = 103). (a)
Distance
Age
Jaccard
-0.21 ***
-0.01 (n.s.)
Sörensen
-0.19 ***
0.04 (n.s.)
Simple Matching
-0.33 ***
0.01 (n.s.)
Sokal and Sneath I
-0.33 ***
0.01 (n.s.)
(b)
Distance
Age
Jaccard
-0.19***
-0.25 ***
Sörensen
0.01 (n.s.)
-0.01 (n.s.)
Simple Matching
0.05 (n.s.)
-0.19 ***
-0.01 (n.s.)
-0.02 (n.s.)
Sokal and Sneath I
(n.s.): not significant, *** : P ≤ 0.001
Effects of age and distance on species composition
49
After correcting for significant correlations with difference in forest age, all investigated species similarity matrices had a significant correlation with the interpatch distance matrix (Table 3.4). On the other hand, accounting for inter-patch distance still results in a significant age effect on species similarity. Since difference in forest age was not significantly related with species composition similarity of the alluvial forest fragments partial correlation did not affect the distance effect on alluvial forest fragments either.
Table 3.4 Correlation coefficients between species composition in the non-alluvial forests (n = 103) and inter-patch distance (X1) for constant differences in forest age (X2; i.e. YX1 .X2) and age for constant inter-patch differences (X1; i.e. YX2 .X1) using partial Mantel tests. YX1 . X2
YX2 . X1
Jaccard
-0.17 ***
-0.23 ***
Simple Matching
-0.20 ***
0.08 (*)
(*) : P < 0.1; *** : P ≤ 0.001
DISCUSSION
Effects of Distance on Species Composition
This study shows that forest plant species composition, both woody and herbaceous, in a fragmented agricultural landscape is strongly affected by inter-patch distances for alluvial forest fragments and both by inter-patch distances and differences in forest age for non-alluvial forest fragments. After accounting for age we also found significant distance effects for the non-alluvial forests. These results are consistent with the distance decay paradigm as formulated by Nekola and White (1999) and confirm earlier findings of Kadmon and Pulliam (1993) who found tree species composition on islands significantly related to the distance from a common mainland. Specifically interesting is the fact that forest types reacted differently to the effects of age and distance.
Chapter 3
50
Studying the distribution of plants on granite outcrops in Western Australia with respect to substrate, location and a broader classification in eastern and western outcrops Burgman (1987) found that matrix correlations were in accord with the hypothesis of environmental control that predicts distributions reflecting spatially autocorrelated environmental variables. In our case, by subdividing fragments in alluvial and non-alluvial forest fragments, a large part of the effects of autocorrelated environmental variables were excluded prior to analysis. This, however, does not mean that the observed decrease in similarity with increasing distance is not a consequence of spatially autocorrelated environmental variables. It does indicate that large-scale spatial structure and dispersal may be of great importance in shaping the components of forest plant diversity in a highly fragmented landscape. Nekola and White (1999) showed that distance decay rates were slow for species with high dispersal rates. Species with low dispersal capacities tend to have significantly clumped distribution patterns resulting in higher among-patch diversity and significant turnover-distance change relationships. Further evidence for dispersal affecting species diversity was given by Butaye et al. (2001) who showed, for the same area, that forest plant species are characterized by different dispersal capacities which led, for very recent forest fragments, to a colonization-based, non-random community structure. Hence, the local community structure resulted from a hierarchical ordering in species arrivals in forest fragments in relation to the geographic placement of the forest fragments within a landscape.
Effects of Age on Species Composition
The effect of distance on species composition cannot be separated from age for the non-alluvial forest fragments. Many authors (Goovaerts et al. 1990, Richter et al 1994, Messing et al. 1997, Koerner et al. 1997, Bossuyt et al. 1999, Verheyen et al. 1999) have shown that historical land-use may have long-term effects on the presentday environmental conditions of forests, which may hamper or slow down colonization processes of many forest plant species. Besides, forest age also represents the colonization period, which is directly related to the probability of a species reaching a forest fragment. This effect may be less clear since it is also strongly influenced by species composition in the surrounding forests. Generally, a
Effects of age and distance on species composition
51
longer colonization period will be related to a higher probability of all species reaching the forest. The colonization probability will be most affected by forest age for species with intermediate dispersal capacities. Species with very low or very high dispersal capacities will not be influenced. Our results indicate that difference in forest age is far less important than the difference in inter-patch distance for the alluvial forests. None of the four similarity matrices had a significant relationship with the age matrix. For the nonalluvial forests difference in forest age revealed significant results for two similarity coefficients. It could be that in alluvial ecosystems historical land-use had little impact on environmental conditions and therefore not on species composition. The relatively high nutrient status of alluvial soils and the rapid turnover of organic matter in these forests may partly explain the floristic similarity between recent and older forest soils. Based on chemical soil analyses we found that plant available phosphate, total nitrogen content and organic matter content were not significantly different between different forest age classes. Only pH was significantly different between forest age classes of the alluvial forest fragments (Kruskal-Wallis χ² = 22.4; P = 0.002). Relatively extensive agricultural practices (meadows and pastures) may explain the small difference in soil properties between young and old forest fragments. Also the fact that alluvial forest fragments are younger than non-alluvial fragments may explain why age is less important in explaining species composition in alluvial forest fragments. Human disturbances may have had a greater impact on environmental conditions of the more nutrient poor non-alluvial forest fragments since these sites were mostly converted to arable land. Several studies (e.g. Koerner et al. 1997, Bossuyt et al. 1999, Verheyen et al. 1999) already documented pronounced differences in soil properties between forests on former arable land and forest that were never cleared.
Landscape Configuration and Spatial Context
A possible point of criticism is that we only took the actual inter-patch distance into account, which is a huge simplification of the reality since forest plant species
Chapter 3
52
composition of older forest fragments may also be influenced by historical landscape configurations. Fig. 1.2 clearly shows that connectivity in the past was much larger than it is now, therefore it can be expected that historical landscape configurations may have influenced both the initial and present species composition in the older forests. We have tried to incorporate this shortcoming by including the age effect on species composition in the analysis. Secondly, it must be mentioned that the intermediate agricultural landscape may also affect colonization rates. Landscapes with a high connectivity are thought to have lower resistance to the movement of species than landscapes with low levels of connectivity leading to higher levels of species overlap and hence slower rates of distance decay. Several authors have shown that small landscape elements such as hedgerows, rows of trees or river borders can allow temporal survival of forest plant species (Pollard et al. 1974, Burel and Baudry 1990, Corbitt et al. 1999, Honnay et al. 1999c). Hence, colonization from these small landscape elements may have reduced the distance effect on species composition. However, based on a random survey of 22 linear elements in the study area we concluded that the intermediate agricultural landscape is unsuitable for supporting most forest plant species although there are some exceptions such as Rubus fruticosus, Sambucus nigra, Stachys sylvatica, Chaerophyllum temulum and Lapsana communis which were frequently found. The survival of a number of forest plant species in the agricultural landscape may ultimately result in lower distance effects.
ACKNOWLEDGEMENTS
This research could be performed thanks to a project financed by the Flemish Community, AMINAL, Departement Nature: VLINA/C97/04b: ‘Development of indicators and indices for biodiversity and the effects of fragmentation on forest plant species in Flemish forests.’ Two anonymous referees provided useful comments on an earlier version of the manuscript.
Effects of age and distance on species composition
53
54
Chapter 3
4.
Influence of environmental and spatial variables on regional distribution of forest plant species in a fragmented and changing landscape
ABSTRACT
During the past several centuries, forests in Europe and large parts of western North America have been subject to extensive forest clearance. The last several decades, however, numerous new forest patches have been established onto former agricultural land. As a result, the present forest area often consists of a mixture of small forest patches of different age, area, habitat quality and connectivity embedded within a hostile agricultural landscape. In these patchy landscapes, distribution patterns of plant species may be affected by both regional and local factors, although the relative importance of both is still poorly understood. In this study, we investigated distribution patterns of 113 forest plant species in a fragmented landscape. Species abundances at the regional scale conformed to a clearly unimodal abundance distribution which we believe to be related to 1) environmental heterogeneity due to succession and 2) inequality in migration rates. Patch incidence was significantly related to life form, which in turn was correlated to seed mass and dispersal mechanism. Multiple logistic regressions showed that presence / absence of 59 species studied was significantly affected by patch connectivity, patch area and age for 35, 30 and 34 species, respectively. The results of this study indicate that distribution patterns of forest plant species are influenced by both local and regional factors. Moreover, they also demonstrate that next to spatial aspects of fragmentation, temporal patterns of landscape change may have far-reaching effects on presence/absence patterns of plant species and therefore should be incorporated in studies dealing with regional population structures of plants.
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INTRODUCTION
During the past hundreds of years, large complexes of natural habitat have been converted into agricultural, industrial or urbanized landscapes, leading to severe loss of the original habitat and an increasing fragmentation of the remnant patches. Recently, it has become clear that this massive destruction of natural land may be detrimental for future biodiversity (Wilcox and Murphy 1985, Wilcove et al. 1986, Tilman et al. 1994) and numerous initiatives have been undertaken to stop the ongoing process of habitat destruction and fragmentation by establishing new patches on formerly degraded land (Dobson et al. 1997). However, it remains an open question whether these new patches will be colonized in the near future and how spatial and environmental conditions affect species recovery and patch occupancy patterns in these newly emerging landscapes (Huxel and Hastings 1999). In addition, it is reasonable to expect that the response to changing landscape structures differs between species, most likely as a result of different habitat requirements and dispersal abilities, and that therefore a large variation in local abundance and regional distribution between species may exist (Quintana-Ascencio and Menges 1996, Villard et al. 1999). In general, species regional distribution in a constantly changing landscape may reflect the result of a dynamic interplay between i) the amount and configuration of suitable habitat for a species, ii) species-specific migration rates between suitable habitat patches and iii) the ability of species to persist in local habitats. Local persistence is most likely to depend on the area of the habitat since small habitat patches can only support small populations, which are more sensitive to size fluctuations and stochastic perturbations (Pimm et al. 1988, Menges 1991). Migration rates, on the other hand, will depend on species dispersal capacity and the spatial configuration of patches. A significant time lag between patch establishment and species recovery may occur when new habitat patches are established in a completely random way (Tilman et al. 1997, Huxel and Hastings 1999, Johst et al. 2002). If, on the other hand, restored patches are established adjacent to existing, occupied patches, the time lag between patch restoration and species recovery was shown to decrease and species were found to reach substantial higher frequencies of occupancy (Huxel and Hastings 1999, Johst et al. 2002). Furthermore, it can be supposed that it may take
Regional distribution of forest plant species
57
some time before newly established habitats display environmental conditions comparable to that of the original habitats (i.e. a time-lag between patch regeneration and the origination of suitable habitat conditions due to succession), which may further delay colonization of new habitat patches (Johnson 2000). At present, the relative importance of local and regional factors determining regional abundance of plant species is still poorly understood (Ehrlén and Eriksson 2003, Freckleton and Watkinson 2003) and only a few studies have investigated the role of area and isolation in determining plant distribution patterns (e.g. Quintana-Ascencio and Menges 1996, Groom 1998, Bastin and Thomas 1999, Harrison et al. 2000, Jacquemyn et al. 2002, Johansson and Ehrlén 2003). Eriksson et al. (1995) and Dupré and Ehrlén (2002) concluded that habitat quality might be of overriding importance compared to habitat configuration in determining plant species distribution patterns. Quintana-Ascencio and Menges (1996), Groom (1998) and Harrison et al. (2000) on the other hand found significant effects of spatial configuration on plant population occurrence. Recent seed introduction experiments seem to support the latter as they have amply shown that the occurrence of species in suitable habitat was limited by the availability of seeds (Primack and Miao 1992, Ehrlén and Eriksson 2000, Turnbull et al. 2000). Starting point of this study is the fact that forest history in much of Europe and large parts of western North America is characterized by extensive forest clearance. The last several decades, numerous new forest patches have been established onto former agricultural land. In this study, we addressed three specific questions concerning the spatial and temporal distribution of forest plant species: (1) How do patch dynamics influence patch occupancy patterns of forest plant species? (2) What are the mechanisms behind these patterns? (3) Which plant traits are important in determining the relationships between patch dynamics and regional dynamics of plant populations? We hypothesized that plant species with specific dispersal mechanisms or high seed production should have higher patch occupancy than plants lacking specific dispersal mechanisms or low seed production. We also hypothesized that species that are more generalized in their niche requirements may find it less difficult to cope with changing landscape structures than species with narrow ecological amplitudes.
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MATERIALS AND METHODS
Vegetation data
A continuous area of 42 km² was selected to study the effects of patch dynamics on species occurrence. Forests consisting of parts with a different land use history and hence a different age were divided into forest patches with a uniform land use history. In this way, we were able to estimate the approximate age of each forest patch. In total, 240 forest patches were distinguished. For each patch, individual patch area was determined and for each pair of forest patches, inter-patch (edge-to-edge) distances were calculated using the Arc/Info software. All forest fragments were inventoried twice, once in early spring and a second time during the summer period. Presence / absence of 203 forest plant species was determined systematically by walking transects of several meters wide so that the whole patch was completely sampled (cf. Kirby et al. 1986). All inventoried species are vascular plant species that grow optimally in forests (Honnay et al. 1999).
Life history traits
Information on life history traits was collected from a variety of sources (Grime et al. 1988, Hodgson 1995, Kleyer 1995, Ecological Flora of the British Isles) (Table 4.1). These traits were selected on the basis of their possible relevance for persistence in a dynamic landscape. They relate to dispersal (dispersal mode, mean plant height, seed number per pant and seed mass) and to survival (diaspore bank longevity, clonal propagation and life form) (Table 4.1). Next, since Ellenberg indicator values were generally considered accurate predictors for species occurrence (Dupré and Diekmann 1998, Butaye et al. 2002), habitat requirements were determined based on the Ellenberg values for nutrient, soil moisture, light requirements and soil acidity (Ellenberg et al. 1992). Following Blomqvist et al. (2003), values were converted to nominal variables, including indifference as a class (Table 4.1).
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59
Table 4.1 Description of life history traits used in the analysis. Trait
Description
Traits related to colonization Dispersal modea
Three
types:
zoochory
(1),
barochory
(2),
anemochory (3). Mean plant heightb
Continuous b
Seed number per plant
1: < 25, 2: 26-100, 3: 101-1000, 4: 1001-10000, 5: > 10000.
b
Seed mass
Continuous
Traits related to survival Life form
Woody (1) or non-woody perennial (2)
Diaspore bank longevityc
Transient (1), short-term persistent (< 5 years) (2), long-term persistent (> 5 years) (3)
Environmental requirements Light requirements (L)d
Indifferent (X), shade demanding (1-4), shade tolerant (5-6), light demanding (7-9)
Moisture (F)d
Indifferent (X), dry (4-6), moist (7-8), wet (9-11)
Nutrient requirements (N)d
Indifferent (X), oligotrophic (1-4), mesotrophic (56), eutrophic (7-9)
Acidity (R)d a
Indifferent (X), acid (1-4), neutral (5-7), basic (7-9)
Grime et al. (1988) ; b Ecological Flora of the British Isles, Kleyer (1995) ; c Hodgson et al. (1995) ; d
Ellenberg et al. (1992).
Statistical analysis
For all species (113 in total), patch occupancy was calculated as the number of occupied patches divided by the total number of patches. To evaluate the importance of life history traits on species occurrence patterns, life history traits were related to patch occupancy patterns using one-way ANOVA and Pearson product moment correlations. To obtain normality and constancy of the error terms, patch occupancy was arcsine square root transformed prior to analysis. To investigate the possible existence of a time lag between forest restoration and species recovery, two-way tests of independence were used. Patch age
60
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was converted to four classes, which correspond approximately with the following ages: 1) 0-50 yr; 2) 51-100 yr; 3) 101-200 yr, 4) more than 200 yr. For the results to be statistically relevant, analyses were performed for only those species occurring in at least 10 % of all forest patches (59 species in total). For the same set of species, a more detailed analysis of the importance of environmental conditions, patch area, age and spatial isolation in explaining patch occupancy patterns was performed. For each species and for each patch, species-specific isolation was measured using the connectivity measure defined in Hanski (1994) as Si = ∑Ajexp(-αdij), which takes into account distances to all possible source patches. Aj is the area of each source patch j occupied by the focal species, dij is the distance from the edge of the focal patch i to the edge of each other patch j occupied by the focal species. Since patch occupancy patterns may be largely affected by environmental characteristics of the habitat, for each patch we calculated mean Ellenberg values for soil moisture, soil nutrient conditions, light conditions and soil acidity. For each species, patch occupancy patterns were analysed by analysis of deviance of multiple logistic regression models (McCullagh and Nelder 1989). Initially, seven model variables were included in the model. Both Mallow’s Cp statistic and model deviances were used to determine the model with the best fit.
RESULTS
Patch occupancy patterns
The number of patches occupied by a given species ranged from 1 (0.4 %) to 215 (89.6 %) (median: 28 (11.7 %)) (Fig. 4.1). Frequency distribution of occupied patches was highly skewed: 11 (10 %) and 41 species (36 %) occurred in less than 1 % and 5 % of all patches respectively. Relating plant traits with patch occupancy patterns revealed that patch occupancy is mainly determined by species habitat preferences, dispersal mechanisms, plant height and life form (Table 4.2). Woody species had significantly higher occupancy than non-woody species. Species preferring acid or basic soil conditions had significantly lower patch occupancy than species with no clear affinity with soil acidity, or species preferring acid-neutral soils. Anemochores and zoochores showed significantly higher patch occupancy patterns than species
Regional distribution of forest plant species
61
lacking specific dispersal mechanisms. As dispersal mechanism and plant height were significantly related to life form (Table 4.2), data were reanalysed for woody and nonwoody perennials separately. For non-woody plants, patch occupancy was significantly related to plant height (r77 = 0.28, p < 0.05), but not to dispersal mechanism (F2,82 = 1.29, p > 0.05). Plant height was not significantly (r17 = 0.28, p > 0.05) related to patch occupancy for woody perennials. No relationships between dispersal mechanism and patch occupancy were calculated for woody perennials, as all species were zoochorous.
Number of species
20
15
10
5
0 0
20
40
60
80
100
120
140
160
180
200
220
Number of patches occupied
Fig. 4.1 Species frequency distribution of 113 forest plant species in 240 forest patches in Central Belgium.
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62
Table 4.2 Interrelations of the studied life history traits as determined by one-way ANOVA, Pearson product moment correlation and χ²statistics. Seed mass
Seed number per plant
Mean plant height Dispersal mode
Seed number per plant
F4,39 = 0.69
Mean plant height
r78 = 0.43 ***
F4,42 = 1.29
Dispersal mode
F2,73 = 0.64
χ2 = 17.28 *
F2,85 = 2.10
Diaspore bank longevity
F2,40 = 1.08
χ2 = 12.32
F2,80 = 2.52
χ2 = 4.25
Life form
F1,77 = 8.01 ***
χ2 = 4.48
F1,91 = 10.21 **
χ2 = 17.43 ***
*** : p < 0.001, ** : p < 0.01, * : p < 0.05
Diaspore bank longevity
χ2 = 1.37
Regional distribution of forest plant species
63
Table 4.3 Univariate relationships between species occurrence and ecological traits. Pearson correlations were used for the correlation between incidence and continuous variables. One-way ANOVA was used to investigate relationships between patch occupancy and categorical variables. Values of patch occupancy were arcsine square root transformed prior to analysis.
Trait
Value
Significance
Diaspore bank longevity
F2, 83 = 0.77
P = 0.47
Dispersal mode
F2,99 = 3.68
P = 0.03
Seed number per plant
F4, 42 = 1.86
P = 0.13
Mean plant height
r = 0.37
P < 0.001
Seed mass
r = 0.27
P = 0.02
Life form
t111 = -4.141
P < 0.001
Light requirements (L)
F3, 109 = 1.40
P = 0.25
Moisture (F)
F3, 109 =1.16
P = 0.33
Acidity (R)
F3, 109 = 4.47
P = 0.005
Nutrient requirements (N)
F3, 109 = 1.09
P = 0.36
50 out of 59 investigated species had a significantly higher occurrence in older forest patches indicating that there is a time lag between patch regeneration and effective colonization of restored patches (Fig. 4.2). Results of the multiple logistic regression analyses showed that, depending on the species, environmental variables explained some part of the variation in the occurrence of forest plant species. Soil acidity had a significant effect for 24 species, soil nutrient content for 18 species, and soil moisture content for 12 of the 59 investigated species. Patch occupancy was significantly affected by light conditions for 28 (47 %) species. Significant effects for isolation were obtained for 35 (59 %) species (Table 4.3). Area on the other hand was significant for 30 of 59 (51 %) species, whereas age had a significant effect for 34 (58 %) species (Table 4.3).
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64
35
Number of species
30
25
20
15
10
5
0 p > 0.05
0.01 < p < 0.05
0.001 < p < 0.01
p < 0.001
Significance level
Fig. 4.2 The number of species of which the distribution is significantly affected by patch age. Significance levels are based on χ²-statistics.
DISCUSSION
Patch occupancy and life history traits
Patch occupancy patterns of forest plant species in the studied system are mainly determined by life form (woody vs. non-woody perennials), which was strongly correlated to plant height, by seed weight, dispersal mechanism and habitat requirements. It appeared that only 12 species (11 %) occurred in more than 50 % of all patches. Most of these species were tall shrub species, which are characterized by a large seed production and large seed mass, show little or no specific habitat requirements and are generally dispersed by birds (e.g. Corylus avellana, Sambucus nigra, Crataegus monogyna, Rubus fruticosus coll, Viburnum opulus). Gaudet and Keddy (1988) showed that plant traits such as plant height and seed size are related to competitive abilities, which may suggest that both competitive abilities and dispersal were important in determining patch occupancy patterns.
Regional distribution of forest plant species
65
Table 3 Multiple logistic regression of patch occupancy on environmental variables, patch area, patch connectivity and patch age for 59 species occurring in more than 10 % of all patches. Both estimated regression coefficients and significances are shown. Model selection was based on analysis of deviance and Mallow’s Cp statistic. Connectivity was measured using Hanski’s (1994) Incidence Function Measure.
Environmental variables Incidence (%)
Light
Adoxa mosschatellina
33
Ajuga reptans
20
Alliaria petiolata
17
Anemone nemorosa
48
Arum maculatum
25
Athyrium filix-femina
41
-2.043 ***
Brachypodium sylvaticum
14
-1.821 **
Carex pilulifera
14
Carex remota
14
-1.622 **
Carex sylvatica
16
-1.184 *
Chaerophyllum temulum
23
Circaea lutetiana
27
Cirsium oleraceum
20
Nutrients
Acidity
Humidity
Area
2.503 *** 1.487 ***
0.304 **
2.592 *** -1.695 *** 1.856 ***
Age
0.080 ***
0.017 ***
0.056 ** 0.090 **
0.997 ***
0.085 ***
0.022 ***
1.022 **
0.084 ***
0.016 ***
1.554 *** 1.195 ** -1.276 *
Connectivity
0.723 *** 0.531 ***
0.009 *** 0.117 ***
-0.729 ***
0.013 *** 2.319 **
1.320 *
0.285 **
0.084 ***
0.491 ***
0.084 ***
0.011 **
0.321 **
0.066 ***
0.006 *
1.177 *** -1.129 *** 1.360 ** 1.245 **
3.318 ***
0.130 ***
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66
Table 3 (continued) Environmental variables Incidence (%)
Light
Convallaria majalis
13
-2.002 ***
Cornus sanguinea
32
Corylus avellana
72
Crataegus monogyna
74
Deschampsia cespitosa
35
Digitalis purpurea
12
Dryopteris carthusiana
36
Dryopteris dilatata
Nutrients
Acidity
Humidity
Area
Connectivity
0.263 *** 1.033 *
0.729 **
-1.705 ***
0.010 ** 0.068 ***
0.346 * 2.312 ***
Age
0.014 ***
0.048 *** 0.048 **
0.007 *
0.445 ***
0.041 **
0.008 ***
-0.929 **
0.301 **
0.135 ***
-1.816 ***
-1.005 *** 2.253 ***
0.324 *
0.009 ***
39
-2.316 ***
-1.023 **
0.636 ***
0.006 ***
Dryopteris filix-mas
45
-1.103 ***
Epilobium angustifolium
14
0.737 *
Epipactis helleborine
33
0.702 ***
0.468 ***
Frangula alnus
27
-1.128 *** 1.152 *
0.456 ***
Geum urbanum
46
-1.154 *** 1.616 ***
Hedera helix
67
-1.766 ***
Holcus mollis
23
-0.980 *
Humulus lupulus
20
1.378 **
1.642 **
0.410 *** 0.011 ***
-0.889 *
0.045 **
0.009 ***
0.051 ***
0.007 ** 0.009 ***
-1.925 *** 0.897 *
0.828 *
0.330 *
0.118 ***
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67
Table 3 (continued) Environmental variables Incidence (%)
Light
Nutrients
Acidity
Humidity
Area
Connectivity
-1.426 **
0.319 **
0.147 ***
Ilex aquifolium
13
Lamium galeobdolon
31
-1.699 ***
0.655 *
Lapsana communis
43
0.579 *
-0.990 **
Lonicera periclymenum
51
-0.794 *
Luzula multiflorum
9
Luzula pilosa
15
Maianthemum bifolium
9
Moehringia trinervia
35
-0.966 **
Poa nemoralis
41
-0.666 *
Polygonatum multiflorum
58
-3.093 *** -0.977 **
Primula elatior
29
2.102 ***
Prunus spinosa
26
Pteridium aquilinum
14
Ranunculus ficaria
57
Ribes rubrum
61
-1.873 ***
Ribes uva-crispa
15
-1.234 **
-1.340 ***
0.117 ***
0.012 **
0.050 ***
0.007 ***
0.466 *** 0.512 *
-1.707 *** -2.019 ***
0.012 *** -1.058 ***
0.347 *
0.016 ***
-0.955 ***
0.681 ***
0.019 ***
0.332 *** -1.037 *
2.088 *** 0.532 **
-1.081 ***
Age
0.338 **
0.066 *** 0.089 ***
0.009 ***
0.051 **
0.028 ***
0.060 ***
0.011 ***
0.044 **
-1.279 * 0.878 *
1.258 ***
1.994 ***
0.805 *
1.991 ***
0.040 *
0.009 *** 0.008 **
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Table 3 (continued) Environmental variables Incidence (%)
Light
Nutrients
Acidity
Humidity
Area
Rosa sp.
25
Rubus caesius
20
Rubus fruticosus
90
Rubus idaeus
27
Rumex sanguineus
14
-0.966 *
Sambucus nigra
84
-0.892 **
Scrophularia nodosa
52
0.426 **
Senecio ovatus
12
0.382 ***
Stachys sylvatica
58
Stellaria holostea
50
Teucrium scorodonia
32
Torilis japonica
15
Viburnum opulus
53
Viola riviniana
20
*** : p < 0.001, ** : p < 0.01, * : p < 0.05
1.404 ***
0.539 *
0.210 *
1.211 **
0.286 **
-0.832 **
Connectivity
0.009 ** 0.047 ***
0.251 ** 1.402 *
-1.248 ***
0.010 *
0.077 *** 0.065 **
1.528 ***
-1.457 **
Age
0.007 *** 0.133 ** 0.029 *
0.006 **
0.089 ***
0.006 **
0.350 *
0.007 **
1.036 *** -0.751 *** 1.025 ***
0.030 * 0.439 ***
0.100 ***
0.010 **
Regional distribution of forest plant species
69
Rare species (i.e. species occurring in less than 5 % of all patches) on the other hand generally showed extreme requirements with regards to soil acidity, most of them occurring on acid soils or basic soils. This may indicate the importance of suitable habitat availability in shaping regional distribution patterns of especially rare species. Other studies (e.g. Dupré and Diekmann 1998, Dzwonko 2001, Dupré and Ehrlén 2002) came to similar conclusions. No relationship between patch occupancy and seed number was found probably because several of these rare species generally have a large seed production (e.g. Listera ovata, Campanula trachelium, Centaurium erythraea, Solidago virgaurea).
Spatial components of patch occupancy
We found significant distance and area effects for 35 and 30 of 59 investigated species respectively, indicating that habitat configuration may have important effects on species occurrence. These results strongly contrast with those obtained by Dupré and Ehrlén (2002), who found significant effects for only four and eleven of 57 species studied, respectively. As Dupré and Ehrlén (2002) did not use species-specific isolation measures, this may explain the observed differences between both studies (see also Bastin and Thomas 1999). Studying distribution patterns of both woody and herbaceous species in scrub communities in Florida, Quintana-Ascencio and Menges (1996) on the other hand used isolation measures similar to those reported here and came to similar conclusions. The significant distance effects confirm the findings of Tilman et al (1997), Huxel and Hastings (1999) and Johst et al. (2002), who showed that clumped distribution patterns of habitat patches may lead to higher survival rates in dynamic landscapes for species with restricted dispersal abilities. Indeed, afforestation led to a non-random pattern of patch distribution with older reforested patches often being established adjacent to existing forests, whereas most recent forest patches were often established isolated from existing patches. These new patches adjacent to older patches could be easily colonized by a relatively large number of species (see also Brunet and von Oheimb 1998), whereas good dispersers only colonized isolated patches.
70
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Temporal patterns of patch occupancy
Fahrig (1992) stressed that both spatial and temporal heterogeneity are important determinants of landscape structure. Whereas the spatial component of landscape structure has received far more attention than temporal heterogeneity (e.g. Hanski 1994, Hiebeler 2000), the present study demonstrates that both may have far-reaching effects on species occurrence and species diversity and more attention should be directed to the temporal effects of habitat fragmentation. Keymer et al. (2000) have shown that for a given species with a particular life history, there may exist a critical value for habitat life span below which the landscape changes too fast in relation to the scale of colonization-extinction processes. Vellend (2003) on the other hand showed that the initial extent of habitat loss affected both patch occupancy patterns and the time needed to reach equilibrium conditions. Our results support both of these findings. The observed patch dynamics allowed only a small amount of species to migrate easily through the landscape as the occurrence of most species was significantly related to patch age. This implies that colonization is a slow process that may take several decades, and in some species colonization of new habitat rarely occurred. Given that at present only 19 % of the original forest habitat of 1775 remains, the observed colonization rates may be too low to establish new viable populations in recent forest patches, especially when they are isolated. Therefore, species will go extinct in the long term if the remaining old forest habitat cannot be protected from further destruction, even when forest area increases.
Regional population structure
Large-scale spatial dynamics of plant populations have recently received increased attention (reviewed in Eriksson 1996, Husband and Barrett 1996, Thomas and Kunin 1999, Freckleton and Watkinson 2002). Whereas the metapopulation perspective has been the most appealing theory to predict regional population dynamics, its use has been questioned with regard to plants. Based on an extensive review, Freckleton and Watkinson (2002) proposed several different mechanisms generating regional plant population structures.
Regional distribitution of forest plant species
71
Our results demonstrate that the mechanisms generating regional populations structures in plants depend in part on dispersal capacities on the one hand and habitat tolerances on the other hand. Therefore, it may come as no surprise that species abundances at the regional scale conformed to a unimodal, rather than to a bimodal abundance distribution as two of the critical assumptions of models expecting bimodality are not fulfilled: environmental homogeneity and equality of species immigration and extinction rates (Scheiner and Rey-Benayas 1997). At one end of the dispersal and / or habitat tolerance spectrum are species that rarely colonize recent patches. These species may represent island populations sensu Freckleton and Watkinson (2002) or remnant populations sensu Eriksson (1996, 2000) as there is hardly any immigration or emigration and there may be no other suitable habitat patches than those already occupied. Although the latter needs to be confirmed, these species may ultimately go locally extinct in the long term. At the other end of the spectrum, species with high dispersal capacities and / or no specific habitat requirements frequently colonized forest patches and may be categorized as metapopulations, since occupied patches produced migrants that colonized unoccupied patches or, if long-distance dispersal was a rare event, as spatially structured populations which spread through local dispersal. As we were able to clearly distinguish discrete suitable patches, and colonization was frequently observed based on the historical data, the metapopulation model may be suitable to describe the regional structure of mobile plant species. However, because of the clonal growth and long life span of many forest plant species (Eriksson 1996, Eriksson and Ehrlén 2001, Ehrlén and Lehtilä 2002), local extinction rates will be extremely low, which may invalidate the metapopulation assumption (Freckleton and Watkinson 2002). The observed significant area effects on the other hand suggest patch area dependent extinctions and support the metapopulation perspective.
ACKNOWLEDGEMENTS
This work was financially supported by the Flemish Fund for Scientific Research (FWO) and by the Flemish Community, AMINAL, Dept. Nature. Olivier Honnay and Beatrijs Bossuyt provided useful comments on an earlier version of the manuscript.
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5.
Patch occupancy, population size and reproductive success of a forest herb (Primula elatior) in a fragmented landscape
ABSTRACT
Forest fragmentation is expected to affect patch occupancy patterns, population size and population viability of plant populations through changes in both patch area and isolation. We tested the hypothesis that patch area had a significant effect on patch occupancy and population size of Primula elatior, a common forest herbaceous plant species in Flanders, Belgium. The hypothesis that plants from small populations have lower fitness as reflected by several characteristics related to reproduction was also tested. Finally the probability of P. elatior colonizing presently empty patches was investigated. Patch area proved to be the most important factor explaining population size. Patch area, spatial isolation and within-patch habitat characteristics all contributed significantly to the explanation of the distribution pattern of P. elatior. Plants from small populations had a significant lower individual fitness than plants from large populations. Small populations produced significantly fewer seeds per fruit and per plant than did large populations. Individual seed mass decreased with increasing population size, but total seed mass increased with increasing population size. Plant-to-plant variability in proportion of flowers setting fruit, number of seeds per fruit and number of seeds per plant decreased with increasing population size. Skewed pin-thrum ratios and lower pollination intensity may explain the reduced fecundity in small populations. Geographical isolation had a significant effect on the probability of P. elatior colonizing empty patches. The results show that patch area and isolation may influence regional persistence of plant populations through altered colonization probabilities and reduced reproductive success of small populations.
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74
INTRODUCTION
Forest fragmentation – the process by which formerly continuous forest area turns into forest patches of different size, isolated from each other by non-forested land (Haila 1999) usually includes three major components (Andrén 1996) all of which may have distinctive effects on populations: (1) the pure loss of habitat eliminating all species that occurred only locally in the habitats destroyed by development; (2) reduction in habitat fragment size and (3) increasing isolation between the remnant fragments. The distribution of species in such fragmented landscapes is a function of local extinction and colonization rates (Hanski and Gilpin 1991) both of which are believed to have a great impact on regional-scale species persistence. For continuously changing landscapes, persistence will depend both on the ability of a species to maintain viable populations within individual habitat fragments and on its capacity to (re)colonize habitat patches that are presently empty. Since most forest plant species have rather limited powers of dispersal (Matlack 1994, Grashof-Bokdam 1997, Butaye et al. 2001), it can be expected that the probability of (re)colonization of currently empty patches decreases with increasing isolation (Hanski 1994). It is also believed that, other things being equal, small habitat fragments can only support small populations, which have a higher risk of extinction than large populations (Shaffer 1981, Pimm et al. 1988, Menges 1991, Menges and Dolan 1998). Small populations in combination with increased isolation may be subject to increased inbreeding and to the loss of genetic variation (Barrett and Kohn 1991, Ellstrand and Elam 1993), both of which are expected to result in reduced individual fitness and population viability. The reduced fitness may cause further rapid decline in population size and hence an increased risk of population extinction (Fischer and Matthies 1998). Although the relationship between population size and its effects on plant fitness has been studied by several authors (e.g. Widén 1993, Ouborg and van Treuren 1993, Oostermeijer et al. 1994, Heschel and Paige 1995, Morgan 1999), contradicting results are often obtained and it is difficult to find predictable patterns. Although there is a growing interest in theoretical studies of habitat fragmentation (Hiebeler 2000), empirical studies relating plant population traits with landscape characteristics remain rare. Incidence functions based on presence / absence
Patch occupancy and reproductive success of Primula elatior
75
data have recently been used to predict changes in occupancy patterns of spatially defined metapopulations (Hanski 1994, Andrén 1996). Although only rarely put into practice for plants (Quintana-Ascencio and Menges 1996), at least two major criticisms have been levelled at the incidence approach. First, the metapopulation model does not take into account differences in environmental conditions which may have a strong impact on colonization probabilities (Honnay et al. 1999, Turnbull et al. 2000), or population viability (Widén 1993, Oostermeijer et al. 1994, Fischer and Matthies 1998, Oostermeijer et al. 1998). Secondly, patch area-dependent stochastic extinctions have rarely been demonstrated for plant species (Schemske et al. 1994, Eriksson 1996, Husband and Barrett 1996). In this study we investigated the effects of fragmentation on regional-scale population persistence of Primula elatior, a forest herbaceous plant species common on rich soils in Flanders. The main purpose of this study was to provide insights into how this plant species is distributed geographically in a highly fragmented landscape typical of Flanders and many other parts of Europe, and how environmental and geographical factors may affect extinction and colonization rates. More specifically, four major objectives were addressed in this study. First, we investigated the spatial structure of patch occupancy of P. elatior in this highly fragmented landscape, and the factors generating this pattern. The relative importance of patch area vs. spatial isolation and environmental conditions as driving forces influencing presence / absence were studied. Secondly, the hypothesis that population size was related to patch area was tested. We further tested the hypothesis that plants in small populations are characterized by lower fitness, making small populations more extinction-prone than large populations. If the latter is true, minimum patch areas and minimum population sizes are needed to maintain viable populations within patches. Finally, the probability of P. elatior colonizing presently empty patches, and hence maintaining or even enlarging its regional population structure in a continuously changing landscape was studied.
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MATERIALS AND METHODS
Study area
The study area is situated in central Belgium (Vlaams-Brabant) 25 km east of Leuven (Glabbeek) and covers a total area of 42 km². In the past the area has been subjected to intensive deforestation, resulting in a highly fragmented forest landscape with numerous small forests embedded in an agricultural matrix (Fig. 1.2 illustrates the forest fragmentation process for an area of 80 km²). Forest area increased recently again with 57 ha (20.6 % of the actual forest area) becoming established between 1956 and 1998. At present only 6.3 % of the area is forested. Most forests are small (range 0.7 to 117.8 ha) and isolated (distance to nearest forest ranging from 0.1 to 1.2 km). One part of the study area is situated in the valley of the river Velpe while the other part covers the hills bordering the valley. Altitude varies from 33 m above sea level in the river valley to 85 m above sea level on the surrounding hills. Slope varies from 1 to 16%. Soils in the valley are loamy, poorly drained and have a moist to wet soil moisture content, whereas soils on the hills are sandy loamy and on average well drained. Most forests in the valley were established on former pastures and hayfields and the tree layer consists of Populus sp., Fraxinus excelsior, Alnus glutinosa, Salix sp. and Quercus robur (Alno-Padion-forest) whereas Quercus robur, Sorbus aucuparia, Betula pendula and Prunus serotina mainly occur on the hill tops (Quercion-forest). The agricultural matrix is typically composed of arable fields and pastures, which are inhospitable habitat for most forest plant species. Land use history was reconstructed by digitising all available historical maps for this study area (dates 1775, 1850, 1864, 1893, 1930, 1956, 1970, 1982, 1991) using the Arc/Info software (ESRI 1995). Forests were subdivided into forest patches with a consistent land use history and were used as sampling units for data collection. In total, 241 forest patches ranging in size from 0.1 to 11.4 ha could be distinguished. The pronounced changes in land use and the resulting high degree of fragmentation (decreased patch area, increased isolation) provide excellent opportunities to study the effects of fragmentation on colonization probability, patch occupancy and population viability within remnant patches.
Patch occupancy and reproductive success of Primula elatior
77
Study species
P. elatior is a small perennial rosette plant showing a clear affinity for moist habitats (Whale 1984). Although P. elatior is considered a typical forest plant, it can occasionally be found in meadows outside forests (Valentine 1948, Dierschke 1968). However, in our study area the species occurred almost exclusively in forest habitats.
Fig. 5.1 Oxlip (Primula elatior).
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78
In early spring up to twenty inflorescences are formed per plant, with often more than 20 flowers per inflorescence. P. elatior, like most other Primula species, is distylous (pin plants are long-styled and their anthers are near the base of the corolla, in thrum plants anthers are positioned above the short style) and self-incompatible (Boyd et al. 1990). Flowers are mainly pollinated by bees and bumblebees, although other insects may be involved in the pollination process. Only pollination between pin and thrum morphs results in seed set. At the end of May – beginning of June fruits are produced (Salisbury 1942). The number of seeds per fruit ranges from 35 to 61, with an average of 51 seeds per fruit (Salisbury 1942). Plants are long-lived (10 – 30 years).
Distribution pattern, population size and reproductive success
In April 2000 all forest patches (241 in total) were inventoried for presence / absence of P. elatior. In all forest patches in which P. elatior occurred population size was determined as the number of flowering rosettes. To determine the effect of fragmentation on several characteristics related to reproduction a subsample of 18 populations of P. elatior (ranging in size from 4 to 2271 flowering adults and occurring in patches strongly differing in size and isolation) was selected. For each population we determined the pin-thrum ratio as the difference in numbers of thrum and pin plants divided by the total number of flowering plants. In each population, 20 flowering rosettes were randomly selected and individually marked with flags. When there were fewer than 20 flowering individuals, all individuals were marked. A total of 332 plants was studied. In April, the number of inflorescences and the total number of flowers were recorded for each plant. At the same time, the lengths of the three tallest inflorescences and of the three longest leaves were measured to the nearest millimeter. At the beginning of June all locations were visited for a second time. Most plants were relocated. A small number of plants had been predated by snails and had produced no fruits; these were excluded from further analysis. The number of fruits was counted for each plant. 5 fruits were collected at random and their seeds were counted and weighed. The number of seeds per plant was calculated as the average number of seeds per fruit multiplied by the total number of fruits per plant. Mean
Patch occupancy and reproductive success of Primula elatior
79
individual seed mass was calculated for each plant as the total seed mass from 5 fruits divided by the total number of seeds.
Environmental variables and soil samples
For every forest patch, age, area and geographic isolation (edge-to-edge distance to the nearest P. elatior population) were determined in GIS using the Arc/Info software (ESRI 1995). Based on data from a previous survey (1998) a set of environmental variables was determined. For each patch we calculated mean Ellenberg indicator values for light, soil moisture, pH and soil nutrient status (Ellenberg et al. 1992). The use of Ellenberg values has recently received increased attention as a valuable tool for predicting the occurrence of plant species along environmental gradients (Dupré and Diekmann 1998). For all locations where reproductive variables were determined on P. elatior five soil samples (10 cm deep cores) were taken in the immediate neighbourhood of the sampled plants, after the litter was removed. The five soil samples per plot were thoroughly mixed and pH, organic carbon and the nutrients P, K, Mg and Ca were determined on the composite samples per plot. These samples were air dried and passed through a 2mm sieve. Carbon was determined using the modified Walkley and Black method and pH was measured in a 0.1 molar KClsolution. The nutrients were extracted using an ammonium lactate (pH=3.75) solution. K, Mg and Ca were measured with Atomic Absorption Spectrophotometry (A.A.S.), P by colorimetry and total nitrogen using the Kjeldahl-Lauro method (see Hendrickx et al. (1992) for details).
Data analysis
The distribution pattern of patches occupied by P. elatior was related to patch habitat characteristics (mean Ellenberg indicator values for light, pH, soil moisture and soil nutrient status) and geographical characteristics (geographical isolation and patch area) using Generalized Linear Models (GLM, McCullagh and Nelder 1989). A backward stepwise procedure was used to select all variables significantly contributing to the distribution pattern of P. elatior. Significance was determined
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80
using a likelihood ratio chi-square test. Colonization capacity of P. elatior was determined using GLM. In order to be able to study colonization properly one has to unequivocally distinguish source from target patches. Since forest area started to increase from the early seventies, all forest patches originating after 1970 (84 forest patches in total) were considered target patches whereas all occupied older forest patches were treated as source patches. Presence / absence of P. elatior was then related to geographic isolation (edge-to-edge distance from target patches to the nearest occupied patch). Population size was correlated with patch area, geographic isolation, patch shape and mean Ellenberg indicator values using Spearman rank correlations. A Kruskal-Wallis test was used to see whether population size was significantly correlated with forest patch age. Multiple linear regressions were used to study the relative importance of habitat variables, plant size and population size for several characteristics related to reproduction. Backwards elimination was used to determine the combination of variables giving the best model and to see whether population size was retained in the subset of variables that all contribute significantly to the regression. Akaike’s Information Criterion was used to find the ‘best’ model that balances the reduction of estimated variance error and the numbers of parameters being fitted. Population size, plant size (mean length of the three longest leaves and height of the three longest inflorescences) and habitat characteristics were used as independent variables. Number of flowers, number of seeds per fruit, number of seeds per plant and mean and total seed weight were used as dependent variables. Since soil variables were partly interdependent, we performed a Principal Component Analysis (PCA) to reduce the set of variables to components that were uncorrelated. Sample scores for the rotated principal components were used as habitat variables for further analysis in the regression model. All calculations were based on population means. Population size was log transformed prior to analysis. Finally, in order to be able to quantify pollination succes within and among populations we calculated the coefficient of variation (CV = 100⋅1
SD/mean)
for the
proportion of flowers setting fruit, number of seeds per fruit and number of seeds per plant for each population. This measure yields an estimate of the (un)certainty of reproduction in relation to the mentioned variables (Oostermeijer et al. 1998).
Patch occupancy and reproductive success of Primula elatior
Because
CV
81
standardizes for the mean, it is less dependent on the mean than the
standard deviation, and provides an index of plant-to-plant variation in reproductive variables relative to the mean. Pearson moment correlations were then used to analyse whether population size had an influence on the population coefficients of variation for the different variables related to reproduction.
Results
P. elatior occurred in 69 of the 241 (29 %) inventoried forest patches. The results of the Generalized Linear Modelling approach showed that patch occupancy of P. elatior was determined by both habitat and patch characteristics. Starting from four Ellenberg indicator values, it proved that especially soil moisture content and soil nutrient status were determinant variables explaining presence / absence of P. elatior. Probability of occurrence decreased significantly when patch isolation increased and patch area decreased (Table 5.1).
Table 5.1 Parameter estimates and significances for patch occupancy of P. elatior as derived by a Generalized Linear Modelling (GLM) approach. Initially 6 candidate variables (patch area, patch isolation, mean Ellenberg values for light, soil moisture, soil nutrient status and pH) were included in the model. Estimated
P-value
Coefficient Patch area
0.27
0.03
Patch isolation
-3.31
< 0.001
Mean Ellenberg value for soil nutrient status 1.14
0.01
Mean Ellenberg value for soil moisture
< 0.001
1.50
P. elatior occurred sixteen times (19 %) in young forest patches (n = 84). Colonization probability of P. elatior in these recently-established forest patches decreased significantly with increasing geographical isolation (χ² = 11.89 ; P < 0.001). Colonization probability decreased to less than 10 % when distance to the nearest occupied forest patch was larger than 350 m (Fig. 5.2).
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Colonization probability (%)
40
χ2 = 11.89, p < 0.001, Nagelkerke R2 = 0.25
30
20
10
0 0
200
400
600
800
1000
Distance (m)
Fig. 5.2 Predicted colonization probability of P. elatior for target patches at a certain distance to nearest source patch. Source patches are forest patches originating before 1970 and occupied by P. elatior, whereas all forest patches originating after 1970 (84 forest patches in total) were considered target patches.
Most populations were rather small : 60 % of all populations contained fewer than 200 flowering adults (Fig. 5.3). Population size was significantly correlated with patch area. No correlations between population size and isolation or habitat characteristics could be found (Table 5.2).
Table 5.2 Spearman rank correlations (rs) between population size, patch area, geographic isolation and mean Ellenberg values for light (L), soil moisture (F), pH (R) and soil nutrient status (N) in 69 forest patches in which P. elatior occurred.
Population size
(a)
Patch area
Isolation (b)
L
F
R
N
0.37
-0.11
-0.01
0.18
0.04
-0.02
**
(ns)
(ns)
(ns)
(ns)
(ns)
** : P < 0.01 ; (ns) : not significant (a)
Population size was measured as the number of flowering rosettes.
Patch occupancy and reproductive success of Primula elatior
(b)
83
Geographic isolation was distance to the nearest P. elatior population.
No differences in population size between forest age classes could be found (KruskalWallis χ² = 0.96, P > 0.05). Looking at the frequency distribution of pin and thrum morphs in small populations, we clearly see an under-representation of one of the morphs in the smallest populations (Fig. 5.4). When population size increases, inequality in numbers decreases, resulting in more equal proportions of pin and thrum morphs within larger populations.
0.6
Frequency
0.5 0.4 0.3 0.2
Fig. 5.3 Frequency distribution 0.1
of population size (number of flowering rosettes) for the 69
0.0 0
400
800
1200
1600
2000
2400
Population size (number of flowering rozettes)
inventoried populations within the study area.
1 0,8
Pin-thrum ratio
0,6 0,4 0,2 0 -0,2 1
10
100
1000
10000
-0,4 -0,6
Fig. 5.4 Relationship between
-0,8 -1
population size and the pinPopulation size
thrum ratio for the 18 studied populations of P. elatior.
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84
The factor loadings of the Principal Component Analysis on all soil variables resulting from a VARIMAX rotation are shown in Table 5.3. Three axes with an eigenvalue > 1 and explaining 84.3 % of the variance were withdrawn for further analysis.
Table 5.3 PCA factor loadings for the 8 soil variables determined on soil samples from 18 populations after a VARIMAX rotation. Rotated principal component (a) PCA 1
PCA 2
PCA 3
Ca
0.87
-0.09
0.02
Mg
0.85
0.15
-0.26
PH
0.82
0.23
0.36
Na
0.70
0.51
-0.25
N
0.33
0.90
0.04
C
0.28
0.95
-0.01
K
-0.21
0.76
0.17
P
-0.04
0.10
0.97
Variance explained
3.76
1.83
1.15
(in % of total)
47.00
22.88
14.37
Soil variable
(a)
Factor loadings in bold indicate which variables are significantly correlated with a given principal
component. Only principal components with an eigenvalue of 1 or above are presented. Together, the three principal components explained 84.3 % of the total variance.
The number of flowers per plant is mainly determined by plant size (mean heigth of infloresences) and by habitat characteristics (Table 5.4, Fig. 5.5). There is, however, also a correlation between number of flowers per plant and population size. Mean number of seeds per plant and mean number of seeds per fruit depended significantly on population size, but values were also influenced by habitat characteristics (Table 5.4). Large populations have significantly more seeds per fruit and more seeds per plant than do small populations (Fig. 5.5). Mean seed mass on the other hand decreased with increasing population size, but increased with plant size.
Patch occupancy and reproductive success of Primula elatior
85
Total seed mass however increased significantly with population size because of the
70
Number of seeds per fruit
higher number of seeds produced per plant in large populations (Fig. 5.5).
r = 0.52 *
Number of flowers
60 50 40 30 20 10
70
50 40 30 20 10
0 1
10
100
1000
r = 0.87 ***
60
0 1
10000
10
1.3
r = 0.68 ***
Mean seed mass (mg)
Total seed mass (mg)
2400
100
1000
10000
Population size
Population size
2000 1600 1200 800 400 0
r = -0.68 **
1.1
0.9
0.7
0.5 1
10
100
1000
Population size
10000
1
10
100
1000
10000
Population size
Fig. 5.5 Univariable relationships between population size and a) mean number of flowers per plant; b) mean number of seeds per fruit; c) mean seed mass and d) total seed mass per plant. Both regression lines and Pearson correlation coefficients are shown.
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Table 5.4 Multiple linear regression of several variables related to reproduction on plant size, population size and habitat characteristics as derived by backwards elimination. Plant size is determined by (1) mean height of the three tallest inflorescences and (2) mean length of the three largest leaves. Sample scores for the rotated principal components were used as habitat parameters. (A) Fitted parameter values (intercept; slopes otherwise) and t test for H0 that each regression parameter differs from 0. (B) ANOVA on overall multiple regression. A) Parameter estimates Variable
B) Analysis of variance Parameter
t
P
Source
df
Mean square
F value
P
R2
7.78
0.003
0.63
10.38
0.001
0.58
estimate Number of flowers per plant Intercept
-14.89
-0.72 0.483
Regression
3
895.25
Mean length of inflorescences
2.55
2.72
0.016
Residual
14
115.11
PCA1
7.74
2.97
0.010
PCA3
7.12
2.72
0.017
Intercept
-794.61
-1.92 0.07
Regression
2
2939324.21
Population size
858.45
4.52
Residual
15
283245.85
PCA 2
-264.71
-1.95 0.07
Mean number of seeds per plant
35-year-old). Because this information, together with a population demographic analysis, allowed us to estimate the approximate age
95
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96
of oxlip populations, this study provides excellent opportunities to examine the genetic properties of colonizing and older populations in a dynamic landscape. While this species was found to be able to colonize recent forest patches, colonization probability decreased fast with increasing distance to the nearest occupied source patch, suggesting limited seed dispersal distances and colonization from a limited number of source populations (Jacquemyn et al. 2002). Furthermore, as most recent forest patches in the study area were mainly established in the central part of the valley, oxlip populations occurring in young patches were almost exclusively located in this part of the study area (see below). In contrast, older patches occupied by the species were also found in the higher parts of the study area, mainly along small tributaries of the main river. Specific questions addressed in this study are: •
Does the observed patch turnover and resulting population age structure affect genetic properties of the forest herb P. elatior?
•
Does genetic diversity within populations located in young forest patches significantly differ from diversity within populations located in older patches?
•
Are young populations more or less differentiated from each other than old ones are?
•
Does the spatial arrangement of populations in the landscape affect genetic differentiation?
MATERIAL AND METHODS
Species
Primula elatior L. (oxlip) is a small, long-lived (more than 10 years), herbaceous, diploid (2n = 22) perennial growing in mixed deciduous temperate forests. It is widely distributed in western and central Europe and extends its distribution range to Denmark and further eastwards to central Asia (Woodell 1969). It appears as a late successional species, typically occurring in ancient forests although colonization of recent forest patches has frequently been observed (Jacquemyn et al. 2002). Its presence in forest patches is strongly related to soil moisture content thereby showing
Genetic structure in a changing landscape
97
a clear affinity for moist habitats (Whale 1983; Jacquemyn et al. 2002). In early spring, up to 20 inflorescences per plant, with often more than 20 flowers per inflorescence, are formed. Like most other species of the genus Primula, P. elatior is distylous and self-incompatible (Richards 1997). Flowers are mostly pollinated by Hymenoptera (mostly bumblebees) and Diptera, although other insects (e.g. moths) may be involved in the pollination process (Schou 1983). Only cross-pollination between thrum and pin morphs results in seed set. At the end of May, beginning of June, up to 176 fruits per plant are produced. In large populations with equal proportions of pin and thrum morphs, the number of seeds per fruit ranges from 35 to 61 (Salisbury 1942). In small populations, mean seed set per fruit may be substantially less (Jacquemyn et al. 2002). Seeds lack specific dispersal mechanisms and are generally dispersed in the immediate vicinity of the mother plant. However, as fruit predation by roe deer was frequently observed in the study area, occasional longdistance dispersal may occur through seed transfer by roe deer (cf. Higgins et al. 2003, Vellend et al. 2003).
Studied populations and sampling procedure
Fifteen populations of P. elatior, randomly chosen from all populations occurring in the study area (Fig. 1), were sampled in April 2002. Seven populations were situated in very recent forest patches (less than 35 years old) and eight populations in forest patches older than 35 years. In each population, the total number of flowering individuals was recorded and young leaf material was collected from 20 randomly chosen plants (or all plants in cases where fewer than 20 plants per population were found) and immediately frozen in liquid nitrogen. Afterwards leaf material was lyophilised for 48 h and homogenised with a mill (Retsch MM 200) to a fine powder. Total DNA was extracted from 30 mg of lyophilised leaf material using methods described in Lefort and Douglas (1999). DNA concentrations were estimated on 1.5% (w/v) agarose gels.
97
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98
AFLP protocol
AFLP analysis (Fig. 6.2) was carried out following Vos et al. (1995), using a commercial kit (GIBCO-BRL) and following the protocol of Roldán-Ruiz et al. (2000). The enzymes EcoRI and MseI were used for DNA digestion. Each individual plant was fingerprinted with three primer combinations. The primer extensions used were EcoACC/MseCAA, EcoACC/MseCAT and EcoACT/MseCAT. Fragment separation and detection took place on an ABI Prism 377 DNA sequencer on 36 cm denaturing gels using 4.25% polyacrylamide (4.25% acrylamide/bisacrylamide 19/1, 6M Urea in 1X TBE). GeneScan 500 Rox labelled size standard (Perkin Elmer) was loaded in each lane.
O3 #
#
O4 #
#
O6
O5
Y5 #
Y2
Y1 O1 #
#
#
Y4
#
O7
Y7
#
#
#
#
#
O8
Y6
Y3 #
N O2
1
0
1
2 Kilometers
Fig. 6.1 Map of the study area and location of the fifteen sampled populations of P. elatior. Each polygon represents a single forest patch with a unique land use history. Shaded polygons represent recent forest fragments.
Genetic structure in a changing landscape
99
Fig. 6.2 Overview of the AFLP process (from Mueller and Wolfenbarger 1999). AFLP involves the restriction of genomic DNA by both a rare (EcoRI) and frequent (MseI) cutting enzyme, followed by ligation of adaptors complimentary to the restriction sites and selective PCR amplification of a subset of the adapted restriction fragments. These fragments are visualized on denaturing polyacrylamide gels either through autoradiographic or fluorescence methodologies.
99
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100
DATA ANALYSIS
The fluorescent AFLP patterns were scored using Genotyper (Perkin Elmer 1996). The presence or absence of each marker in each plant individual was scored. Each individual displayed a unique banding pattern. The differences among populations were thus attributable to frequency differences in variable markers, rather than to private markers. Pairwise genetic distances (Nei’s D) were calculated for all populations after Lynch and Milligan (1994)3:
Jˆ Var ( Hˆ ' ) Var ( Hˆ ) Var ( Hˆ k ) jk jk j ˆ − D jk = − ln + + Jˆ Jˆ 2 Jˆ 2jk 4 Jˆ 2j 4 Jˆk2 j k
(3)
A principal coordinate (PCO) analysis was performed based on this matrix using NTSYS (Rohlf 2000) and the first two components were plotted graphically. Evidence of isolation by distance among populations was obtained by examining correlations between matrices of pairwise genetic distances and pairwise geographical distances (Slatkin 1993). Significance of the observed relationships was obtained by using a Mantel test (Mantel 1967). A total of 5000 random permutations were performed. The same test was also used to look for isolation by distance separately among populations in young patches and among populations in old patches. Genetic diversity and differentiation were investigated using
AFLPSURV
V.1.0 (Vekemans et al. 2002). Estimates of allelic frequencies at AFLP loci were calculated using the Bayesian method with a nonuniform prior distribution of allele frequencies following Zhivotovsky (1999), assuming some deviation from HardyWeinberg genotypic proportions. Prior information on the level of inbreeding (FIS) was taken from Van Rossum et al. (2002) who investigated, based on allozymes, levels of inbreeding of P. elatior in nine populations from a nearby geographical area. After estimating allele frequencies, statistics of gene diversity and genetic structure were computed according to Lynch and Milligan (1994). For each population, we
3
See Lynch and Milligan (1994) for explanation of the different terms.
Genetic structure in a changing landscape
101
calculated the number (NLP) and proportion of polymorphic loci (PLP) at the 5 % level and Nei’s gene diversity (Hj). Following Lynch and Milligan (1994), the latter was calculated as:
1 L Hˆ j = ∑ Hˆ j (i ) L i =1
(4)
where Hˆ j (i ) = 2qˆ j (i )[1 − qˆ j (i )] + 2Var[qˆ j (i )]
(5)
and q(i) the frequency of the null allele at the locus involving the ith marker.
Differences among young and old populations in the percentage of polymorphic loci and Nei’s gene diversity were investigated using t-tests. To investigate whether population size affected the percentage polymorphic loci and Nei’s gene diversity, Pearson product moment correlations were used.
Estimates of FST were calculated using formulae given in Lynch and Milligan (1994):
Hˆ Hˆ Var ( Hˆ W ) − Hˆ W Var ( Hˆ B ) + ( Hˆ B − Hˆ W )Cov( Hˆ B , Hˆ W ) FˆST = B x 1 + B Hˆ T Hˆ B Hˆ T2
n n 1 1 with Cov ( Hˆ B , Hˆ W ) = ∑ Hˆ j ∑ Hˆ jk − ( Hˆ W ⋅ Hˆ B ) n n(n − 1) j =1 k =1 k≠ j
and HT, HW
−1
(6)
(7)
and HB the total gene diversity, the mean gene diversity within
populations and the average gene diversity among populations in excess of that observed within populations, respectively. F-statistics and their 99 % confidence interval obtained by random permutation (5000 permutations) of individuals among populations were computed among all populations, between populations in young and old patches, and within populations in young and old patches. Additionally, total genetic diversity was 101
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102
partitioned among groups of populations (populations located in young vs. old patches), among populations within groups, and within populations by carrying out a hierarchical analysis of molecular variance (AMOVA) on Euclidean pairwise distances among individuals using
GENALEX
(Peakall and Smouse 2001). Significances were
determined using permutation tests. A Mantel test was also used as a non-parametric equivalent of
ANOVA
(Sokal and Rohlf 1995) to test the hypothesis that FST values
among populations located in patches of the same age class were significantly different from FST values among populations located in patches of different age classes. Therefore we constructed a dissimilarity matrix with zeroes in the triangular within-group submatrices and ones in the square between-group submatrix. If differences between groups are greater than those within groups, then the ones in the design matrix will be associated with the larger differences.
RESULTS
The population size of the 15 populations studied ranged from nine to more than 2000 flowering individuals (Table 6.1). Patch age had no effect on population size (t-test = -1.88, P = 0.10), although the sample variances were significantly different between populations occurring in young patches and old patches (F = 13.11, P < 0.01). The former had population sizes ranging from 13 to 104 flowering individuals and showed little variance, whereas populations located in older patches showed a much larger variance in population size. Some of these older populations were small and decreasing in size, whereas others were quite large and still expanding (H. Jacquemyn, unpublished data).
Overall characteristics of the AFLP dataset and between-population relationships
The three AFLP primer pairs detected a total of 157 scorable bands. No monomorphic bands were scored. Marker frequencies ranged from 1 to 99% (mean = 48.8, median = 47.7) (Fig. 6.1). For the primer combination EcoACC/MseCAA, 74 polymorphic bands were scored ranging from 75 to 450 base pairs, whereas for the primer combination EcoACC/MseCAT, 48 polymorphic bands were scored. Finally, 35 polymorphic bands were scored for the EcoACT/MseCAT primer combination.
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103
Number of markers
20
15
10
5
0 0
10
20
30
40
50
60
70
80
90
100
Marker frequency
Fig. 6.3 Histogram of the AFLP marker frequencies for the 235 P. elatior samples.
The first two principal coordinates, which explained 70 and 16 % of the total variance, clearly separated the populations according to patch age (Fig. 6.4). No grouping of the populations according to geographical distances is apparent in this plot. In Fig. 6.5, pair-wise genetic similarities between populations have been represented in a different way. In this case, the populations were plotted according to their geographical location, and each population is connected by an arrow to ‘the most similar population’. This kind of representation could highlight inter-population relationships, which might be obscured in the principal-coordinate plot. It is clear from this figure that populations located in young and in old forest patches form two independent networks of genetic interrelationships, as most arrows connect populations located in patches of the same age category. Remarkable exceptions are the relationships between populations O1, O2, Y1 and Y3, located in close geographical neighbourhoods, and the arrow connecting O4 and Y4.
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15
Population Young forest Old forest
10
O3 O7
Y2
5
PCO2
O6
Y1
O4 O1 0
Y4
O5 Y3
O8
Y7
Y5
-5
Y6
O2
-10
-15 -40
-30
-20
-10
0
10
20
PCO1
Fig. 6.4 PCO-plot of first two coordinates calculated based on 157 polymorphic AFLP markers and using Nei genetic distances (Nei’s D) between populations.
Isolation by distance
A significant, positive relationship between genetic and geographical distances was found for old populations (rm = 0.313, P = 0.02), whereas no such relationship was detected for recent populations (rm = -0.145, P = 0.69) (Fig. 6.6a). The overall Mantel test based on all 15 populations was not significant (rm = 0.064, P = 0.73) (Fig. 6.6b).
Genetic structure in a changing landscape
105
O3 O4
O5 O6
Y5
Y4 Y6 Y2
O1 Y3
Y7
O7
O8
Y1 O2
Fig. 6.5 Schematic representation of the relationships between geographic location and genetic similarity based on pairwise Nei distances (Nei’s D). Each population is connected by an arrow (solid lines: old populations, dashed lines: young populations) to the most similar population (smaller Nei’s D) in the dataset. For example, the population most similar to O3 is O4. The population most similar to O4 is Y4.
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a)
Population Old forest Young forest 0.06
FST
0.04
0.02
0.00
0
2
4
6
8
Geographic distance (km)
b)
0.12
FST
0.08
0.04
0.00
0
2
4
6
8
Geographic distance (km)
Fig. 6.6 Relationship between geographical distance and FST for a) young and old populations separately and b) for all populations.
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Genetic variation within populations
The per population percentage of polymorphic loci (PLP) and expected heterozygosity (Hj) ranged from 72.6 to 91.1 and from 0.245 to 0.321, respectively (Table 6.1). No significant (P > 0.05) relationships were found between population size and percentage of polymorphic loci or expected heterozygosity (Pearson r = -0.11 and –0.09 respectively) (Fig. 6.7). Furthermore, both the percentage polymorphic loci and expected heterozygosity were not significantly (P > 0.05) different among populations located in young (mean: 79.53 and 0.278 respectively) and old patches (mean: 79.38 and 0.288) (t13 = 0.05 and –0.95, respectively).
Table 6.1 Statistics of gene diversity within 15 populations of P. elatior for 157 AFLP loci. N = number of flowering individuals, n = number of individuals for which scorable patterns were obtained for the three AFLP primer combinations used, NLP = number of polymorphic loci, PLP = proportion of polymorphic loci at the 5% level, Hj = expected heterozygosity. Population
N
n
NPL
PLP
Hj
Y1
104
20
124
79.0
0.247
Y2
42
18
125
79.6
0.253
Y3
31
19
123
78.3
0.298
Y4
88
17
126
80.3
0.286
Y5
53
15
119
75.8
0.274
Y6
13
11
138
87.9
0.295
Y7
25
16
119
75.8
0.291
O1
837
18
130
82.8
0.302
O2
1200
17
116
73.9
0.273
O3
76
19
114
72.6
0.262
O4
9
6
114
72.6
0.267
O5
257
15
115
73.2
0.278
O6
12
9
135
86.0
0.314
O7
2273
20
130
82.8
0.289
O8
49
15
143
91.1
0.320
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1.0
Genetic diversity
0.8
0.6
0.4
0.2 PLP Hj
0.0 1
10
100
1000
10000
Population size
Fig. 6.7 Relationship between population size and percentage of polymorphic loci and expected heterozygosity in fifteen populations of Primula elatior.
Between-population differentiation and genetic structure
A low but significant level of genetic differentiation was found when all populations were compared (FST = 0.0355, P< 0.001). Total gene diversity and mean withinpopulation gene diversity were higher for old than for young populations (Ht = 0.2987 and 0.2828 and Hw = 0.2881 and 0.2776 respectively) (Table 6.2). Average gene diversity among populations (Hb) was higher for old populations (Table 6.2). The FST value among populations in young patches was significantly smaller than that among populations in old patches (FST = 0.0185 and 0.0353 respectively, Table 6.2). Although relatively low, both values were significantly (P < 0.0001) different from zero. Results of the
ANOVA-like
Mantel test indicated that FST values between groups
of populations (populations in old and young patches, respectively) were also significantly (P < 0.0001) greater than those within groups.
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109
AMOVA revealed that most of the genetic variation (93 %) was found between individuals within populations, whereas 5 % was attributed to variation between populations within groups and 2 % to variation between groups (Table 6.3). All values were highly significant (P < 0.001, Table 6.3).
DISCUSSION
Because of their high replicability and ease of use, amplified fragment length polymorphism (AFLP) markers have been frequently used in analysis of genetic variation below the species level, especially in investigations of population structure and differentiation (reviewed in Mueller and Wolfenbarger 1999). Recently, AFLP markers have also been successfully used in population assignment studies (Campbell et al. 2003). In the context of metapopulations, AFLP markers have been used to study the distribution of genetic variation within and between subpopulations. Tero et al. (2003), for example, studied genetic structure and differentiation in a metapopulation of the endangered perennial Silene tatarica. Contrary to the results presented here, they found relatively high FST values indicating clear subpopulation differentiation, although no hierarchical regional structuring in the metapopulation was found. Other studies estimating FST values based on AFLP markers have also reported rather high FST values (see e.g. Travis et al. 1996, Schmidt and Jensen 2000, Gaudeul et al. 2000, Després et al. 2002, but see Coart et al. 2002). In most cases, however, species were sampled over a wide distribution range, a factor shown to generate large FST values (Nybom and Bartish 2000). On the other hand, studies investigating metapopulation genetic structure using isozymes have demonstrated FST values similar to the values reported here. In the classical study of genetic metapopulation structure of the dioecious perennial Silene dioica in the Skeppsvik archipelago in Sweden, the FST value over all loci for the entire archipelago was 0.038. In a patchy population of the circumpolar tundra species Silene acaulis, little genetic differentiation between populations was found (the mean FST values was 0.007) (Gehring and Delph 1999). Moreover, the FST value found in this study largely corresponds with the value reported in Van Rossum et al. (2002) (FST = 0.058), who investigated genetic diversity of nine P. elatior populations over a larger area in Flanders. 109
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Table 6.2 Analysis of the genetic diversity among populations based on 157 AFLP markers scored in 235 individuals from 7 young and 8 old populations. n = number of populations, Ht = total gene diversity, Hw = mean gene diversity within populations, Hb = average gene diversity among populations, FST = differentiation between populations, Low 99% FST and High 99% FST = critical values at the 99 % level of the randomisation distribution of FST assuming no population structure, based on 5,000 random permutations. All calculations were performed with AFLPsurv 1.0 (Vekemans et al. 2002). Comparison
n
All populations
15
Between young and old 2
Ht
Hw
SE(Hw)
Hb
SE(Hb)
Fst
Low
99 High
% FST
% FST
99 P-value
0.2936
0.2832
0.005475
0.0104
0.001502
0.0355
-0.0094
0.0102
< 0.0001
0.2810
0.2755
0.009364
0.0054
0.000000
0.0193
-0.0091
0.0061
< 0.0001
0.2828
0.2776
0.007698
0.0052
0.000335
0.0185
-0.0130
0.0102
< 0.0001
0.2987
0.2881
0.007782
0.0105
0.001381
0.0353
-0.0118
0.0188
< 0.0001
populations Within
young 7
populations Within old populations
8
110
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111
Table 6.3 Hierarchical analysis of molecular variance (AMOVA) based on 157 AFLP loci in 15 populations. Populations were assigned to two groups: young and old populations.
Sum
of Variance
% of the total P-value
Source of variation
df
squares
components
variance
Between groups
1
80.011
0.369
1.77
< 0.001
Among populations within groups
13
461.473
1.038
4.99
< 0.001
Among individuals within populations
220
4268.881
19.404
93.24
< 0.001
Total
234
4810.366
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Previous research (Jacquemyn et al. 2002) has shown that spatial, temporal and environmental factors significantly affected distribution patterns and demographic properties of the forest herb P. elatior. In this study, we have demonstrated that landscape changes may also have substantial effects on genetic properties of this plant species. Consistent with models of metapopulation genetics (Gilpin 1991; Hastings and Harrison 1994; Harrison and Hastings 1996), the data presented in this study indicate that patch turnover can result in a reduction of genetic variation within populations. Consequently, when landscape changes continue, the metapopulation as a whole may lose a part of its genetic variation. Further, contrary to most other studies examining genetic differentiation between colonizing and older populations (Whitlock 1992; Antrobus and Lack 1993; McCauley et al. 1995; Giles and Goudet 1997a,b; Ingvarsson et al. 1997), we found lower genetic differentiation among populations located in recent forest fragments than populations found in older patches. These results suggest that colonization dynamics not always increase genetic variance among populations. Wade and McCauley (1988) and Whitlock and McCauley (1990) described the conditions under which extinction and recolonization would decrease or increase genetic differentiation (i.e. FST), relative to the equilibrium case with no extinction. It was found that propagule pool colonization (i.e. all founding individuals are drawn from the same source population) always increased genetic differentiation relative to an island model at equilibrium. In the migrant pool model (i.e. colonists originate from different source populations), genetic differentiation increased if the number of colonists was less than twice the number of migrants (Wade and McCauley 1988, Whitlock and McCauley 1990). Based on ecological information on the colonization pattern of P. elatior, it was concluded that colonizing individuals were most likely to originate from a few, nearby populations as colonization probability decreased significantly with increasing interpatch distance (see Fig. 2 in Jacquemyn et al. 2002). Further evidence is provided by the fact that no significant relationships between colonization probability and mean isolation were found when more than 5 source patches were taken into account, whereas strong relationships were found between colonization probability and the distance to the nearest source patch (H. Jacquemyn unpublished results). For the weedy plant Silene alba, McCauley et al. (1995) found that colonization events were of the propagule pool colonization type. Based on
112
Genetic structure in a changing landscape
113
information on the number of colonists and the levels of differentiation among recently established populations and older populations, the authors were able to show that relatively little mixing of individuals from different source populations occurred during colonization. A similar situation was expected in P. elatior as whole fruits or parts of fruits rather than individual seeds are dispersed over larger distances, as was also the case in the related P. vulgaris (Valverde and Silvertown 1995). Therefore, a higher differentiation among recently established populations was expected by the start of the present study. This, however, was not the case and may be explained by the facts that each extant population was not equally likely to be a source of colonists for the patches being restored, and that each patch was not equally likely to receive incoming seeds. In the studied system, most recent patches were established mainly in the central part of the valley. Older populations on the other hand were also found in the higher parts of the study area along small tributaries of the main river. As a consequence, colonization occurred mainly in the central part in the valley, whereas relatively less gene flow was likely to occur in the higher parts of the study area. In this case, the higher FST values among older populations may reflect historic, rather than current levels of gene flow, as these populations were more isolated than young populations are now and as a result contributed less to migration. Similar results have been reported by Dybdahl (1994), who found that older populations of the bottomdwelling marine copepod Tigriopus californicus were genetically more differentiated than the younger populations were. This hypothesis was confirmed by the fact that isolation by distance was found for older populations, whereas young populations showed no such relationships. This may suggest that older populations are moving toward an equilibrium between migration and drift. According to Dybdahl (1994), newly founded populations should not display higher differentiation, given that they are mainly drawn from younger central populations, as they are not influenced by accumulated differentiation among old populations. Inspection of Fig. 3 suggests that this might be the case. Older populations and young populations seem to form two superimposed networks of populations, suggesting that recently established populations might be tightly linked by gene flow, whereas peripheral, mostly older populations contribute little to gene flow.
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Moreover, gene flow may occur through the transfer of both seeds and pollen. Although seed dispersal is the primary source of long-distance gene flow in colonizing populations, it is reasonable to expect that, after colonization, pollen flow among populations may also have contributed to the observed relationships. Given their central location and the decreased distances between occupied patches, the significantly lower FST values at nuclear loci observed among pairs of young populations may partly reflect high levels of pollen flow among these populations, which can be expained by several factors. First of all, these populations may be more accessible to pollinators than populations in older patches because these older forests generally have a well-developed forest edge structure, which may impede exchange of pollinators among patches. In the second place, most young forest patches were established in the central part of the study area and are separated by relatively small geographical distances, which may increase pollen flow (Steffan-Dewenter and Tscharntke 1999; Richards et al. 1999). Furthermore, as the size of a population may also affect gene flow by pollen (Groom 1998; Richards et al. 1999), small, peripheral populations located in older patches are likely to receive only a small amount or even no incoming pollen from other populations. Small populatons indeed showed a reduced reproductive success, most likely as a result of pollinator deficits (i.e. Allee effect) (Jacquemyn et al. 2001c, 2002). Because of their low seed set and limited exchange of pollen, it can be assumed that these populations are less closely connected to each other and to the rest of the metapopulation, which makes them more sensitive to the effects of genetic drift. Only a more thorough analysis of actual rates of pollen and seed flow in the studied system would allow confirming the validity of these conclusions. In conclusion, the results of this study suggest that historical changes in landscape structure may be important in shaping the genetic diversity of plant populations and therefore should be incorporated in studies examining genetic structure at the landscape scale.
ACKNOWLEDGMENTS The study was funded by two grants from the Flemish Fund for Scientific Research (FWO). Many thanks to Nancy Mergan and Katrien Liebaut for technical assistance in
Genetic structure in a changing landscape
115
the laboratory. We are most grateful to three anonymous referees who provided very useful comments on an earlier version of the manuscript.
115
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7.
Impacts of restored patch density and distance from natural forests on colonization success
ABSTRACT
The reduction and fragmentation of forest habitats is expected to have profound effects on plant species diversity as a consequence of the decreased area and increased isolation of the remnant patches. To stop the ongoing process of forest fragmentation, much attention has been given recently to the restoration of forest habitat. The present study investigates restoration possibilities of recently established patches with respect to their geographical isolation. Since seed dispersal events over 100 m are considered to be of long distance, a threshold value of 100m between recent and old woodland was chosen to define isolation. Total species richness, individual patch species richness, frequency distributions in species occurrences and patch occupancy patterns of individual species were significantly different among isolated and non-isolated stands. In the short term, no high species richness is to be expected in isolated stands. Establishing new forests adjacent to existing woodland ensures higher survival probabilities of existing populations. In the long term however, the importance of long-distance seed dispersal should not be underestimated since most species showed occasional long-distance seed dispersal. A clear distinction should be made between populations colonizing adjacent patches and patches isolated from old woodland. The colonization of isolated stands may have important effects on the dynamics and diversity of forest networks and more attention should be directed towards the genetic traits and viability of founding populations in isolated stands.
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INTRODUCTION
Forests in Western Europe and many other parts of the world are becoming more and more fragmented typically resulting in numerous small patches isolated from each other by non-forested land (e.g. Curtis 1956; Burgess and Sharpe 1981; Jacquemyn et al. 2001a; Rudd et al. 2002). As a consequence of decreased patch area, species richness is most likely to decrease. Other things being equal, small patches can only support small populations (Jacquemyn et al. 2002), which have a higher risk of extinction than large populations due to the effects of environmental, genetic and demographic stochasticity (Shaffer 1981; Pimm et al. 1988; Menges 1991). Increased isolation on the other hand makes (re)colonization highly improbable so that in the long term small forest patches are most likely to lose some species and forests gradually to become impoverished. To stop the ongoing process of forest fragmentation and resulting species’ loss, much attention has been given recently to the theory and practice of restoring forest habitats (e.g. Peterken 2000; Bossuyt and Hermy 2000; Honnay et al. 2002). Clearly, typical forest species, in particular so-called ancient-forest species, are very slow to colonize recent forest patches (e.g. Peterken and Game 1984, Matlack 1994; Brunet and von Oheimb 1998; Bossuyt et al. 1999; Dzwonko 2001; Singleton et al. 2001) or are not able to colonize recent forest patches when they are isolated from old woodland (Grashof-Bokdam 1997; Grashof-Bokdam and Geertsema 1998; Butaye et al. 2001). The low colonization capacity of these forest plant species can be explained partly by unsuitable habitat conditions of recent patches (recruitment limitation) (Koerner et al. 1997; Honnay et al. 1999a) and partly by low dispersal capabilities (dispersal limitation) (Bossuyt et al. 1999; Butaye et al. 2001). Low seed production of many forest plant species appears to be a major problem (Verheyen et al. 2003). These findings pose severe questions about the effectiveness and ultimate success of reforestation projects. In a recent paper, Peterken (2000) applied percolation theory to suggest landscape structures required to allow successful migration of plant species among forest patches. He showed that values of thirty and sixty percent of woodland appeared to be important threshold values. Beneath thirty percent of woodland, most new forest patches are isolated from old forests thus hampering migration among
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119
patches. Above 60 % most forest patches are interconnected to form large forest complexes enabling successful migration of plant species within these forest structures. Given the high demands of other land uses (e.g. agriculture, urban and industrial infrastructure), such huge amounts of forest cover cannot be obtained in our cultural landscapes. Therefore, Peterken (2000) concluded that planning forest habitat networks has to be conducted in the most efficient way possible, i.e. ‘the network benefits must be sought with a minimal amount of new woodland’. At present, quantitative data are lacking about the amount of new forests needed to ensure that a relatively large number of forest plant species succeed in colonizing recent forest patches. Moreover, the interaction between geographical location and the amount of new forest habitat necessary for colonization success is not yet known. In this study we present a simple randomization algorithm that allows investigation, based on empirical data, of the number of patches (a measure of patch density) needed to ensure high overall colonization success on a landscape scale and predictions about forest habitat networks. More specifically, we address the following questions: •
Does the number of recently restored patches affect the total number of species encountered in these patches?
•
Does isolation affect total species richness of recently restored forest patches?
•
Is there a difference in species richness between isolated and non-isolated patches?
•
Are there differences in distribution patterns of forest plant species among isolated and non-isolated patches?
MATERIAL AND METHODS
Study area
The study area is situated in the central part of Belgium (Vlaams-Brabant) (central point 50°52’N, 4°50’E), between Diest, Tienen and Leuven and contains a large river valley bordering the surrounding hills. Topographical altitude ranges from 33 m
119
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above sea level in the valley to 85 m on the hills. Soils in the valley have a silty texture, are poorly drained, are mesothropic and weakly acid, whereas soils on the hills are more acid, well drained and have a sandy loam texture. For a more comprehensive outline of the study area we refer to Jacquemyn et al. (2001a; 2001b). In this article we exclusively focus on forests in the alluvial part of the study area. These forests generally belong to the Alno-Padion alliance. Characteristic species in these communities include among others Adoxa moschatellina (moschatel), Circaea lutetiana (enchanter’s nightshade), Geum urbanum (herb Bennet), Primula elatior (oxlip) and Ranunculus ficaria (lesser celandine). The studied alliance is rather rare in Flanders; at present, only 3,000 ha (or 2 % of the total forest area in Flanders) of welldeveloped Alno-Padion communities can be found, despite an estimated potential area of 135,000 ha (De Keersmaeker et al. 2001). Based on historical maps, previous work (Jacquemyn et al. 2001a; 2001b) has demonstrated intensive changes in land use during the last two centuries with forests constantly appearing and disappearing resulting in a complex forest history. At present, 139.2 ha or 19.9 % of the valley area is afforested. During the last decades, numerous new forest patches have been established mainly in the valley of the study area. Between 1956 and 1998, a total of 60 forest patches (44.84 ha or 32.22 % of the total forest area in the valley) were established. Since all forest patches were established on former grasslands, it can be assumed that these sites do not contain persistent seed banks of typical woodland species (Brown and Warr 1992), or relict populations. As such they provide excellent opportunities to study the restoration possibilities of forest plant communities in an agricultural landscape. Most of these patches are small: area ranges from 0.12 to 3.91 ha (mean: 0.74, SD: 0.65). Isolation (measured as the shortest (edge-to-edge) distance to the nearest old forest) ranges from 0 to 955 m (mean: 236, SD: 312).
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Data collection and analysis
Presence / absence of 203 plant species typically confined to forest habitat (Honnay et al. 1999b) was determined twice, once in early spring and a second time between July and August in 1998. All forest patches were surveyed by systematically walking transects several meters wide. A total of 69 species was found in these recent forest patches. A simple randomization algorithm was used to evaluate the importance of patch density (number of recent forest patches) on overall colonization success in young forest patches. Therefore, for each density (ranging from 2 to 59 forest patches) forest patches were randomly selected and the total number of different species was calculated. For each density, this procedure was repeated 499 times to produce a frequency distribution of total number of species. To determine species accumulation in young forest patches, mean values were plotted against patch density. A 90 % threshold value (the number of forest patches / km needed to obtain 90 % of all species occurring in recent forest patches) was determined graphically. Next, to investigate the impact of geographical location of forest patches on total species accumulation in recent forest patches, the same analysis was repeated for isolated and non-isolated recent patches. A threshold value of 100m between recent and old woodland was chosen to define isolation since seed dispersal events over 100 m are considered to be long distance (Cain et al. 2000). The results were compared with the real situation in which both isolated and non-isolated patches were considered simultaneously. To show the importance of isolation on individual species occurrences, we compared species distribution patterns among isolated and non-isolated recent patches. The Kolgomorov-Smirnov two-sample test (Sokal and Rohlf 1995) was used to test the hypothesis that the two samples are distributed equally. A Wilcoxon rank sum test was used to test whether the mean number of patches occupied by a species differed among isolated and non-isolated patches. Since the size of a patch may significantly affect species richness, we also related patch area to species richness using Spearman rank correlation. Finally, G-tests of independence (Sokal and Rohlf 1995) were used to see whether species occurrences were significantly related to isolated or non-
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isolated patches. G-tests were only performed for species occurring in at least 5 (~10 %) patches. All statistical analyses were performed in S-PLUS 4.5 (MathSoft 1998).
RESULTS
Species number in recent alluvial forest patches ranged from 1 to 40 species (mean: 14, SD: 7). The total number of species encountered in recent patches increased rapidly when patch density is low, but a gradually slower rate when patch density was high (Fig. 7.1). Only thirty patches (50 %) were needed to comprise 90 % of all species occurring in young forest patches (Fig. 7.1).
100
Percentage of total species
90 80 70 60 50 40 30 20 10 0
10
20
30
40
50
60
Number of patches
Fig. 1 The relationship between accumulated number of plant species (expressed as % of all species occurring in recent patches) and patch density (the number of patches). Species number was calculated by randomly selecting forest patches and counting the total number of species encountered in these patches. For each number of patches, values represent means of 499 combinations of randomly selected patches. The dashed line represents the 90 % threshold patch density (i.e. the number of forest patches needed to obtain 90 % of all species occurring in recent forest patches).
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Species richness in isolated patches was also significantly lower than in non-isolated patches (Wilcoxon Z = 2.6, p = 0.009) (Fig. 7.2). No significant relationship, however, was found between patch area and species richness (rs = 0.02, p > 0.05).
25
Species richness
20
15
10
5
0 Isolated
Non-Isolated
Fig. 7.2 Species richness of isolated (distance to old woodland > 100m, n=30) was significantly different from that of non-isolated (distance to old woodland < 100m, n=30) patches (Wilcoxon Z = 2.6, P < 0.01).
Rates of species accumulation and total species richness in recent forest patches was significantly higher for non-isolated than isolated forest patches (Fig. 7.3). Surprisingly, there were no differences in accumulation rate and total species richness when all forest patches (both isolated and non-isolated patches) were compared with non-isolated patches (Fig. 7.3).
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Total number of species
60
50
40
30
20
10
0
5
10
15
20
25
30
Number of patches
Fig. 7.3 Cumulative number of species for isolated (*), non-isolated (■) and all young (○) forest patches. Cumulative number of species was obtained by randomly selecting forest patches and calculating the total number of different species occurring in forest patches. Values represent means of 499 combinations of randomly selected patches.
Species distribution patterns differed significantly among isolated and non-isolated forest patches (Kolmogorov-Smirnov test statistic = 0.26, p = 0.02) (Fig. 7.4). The number of patches occupied by an individual species ranged from 1 to 28 for nonisolated patches (median = 7) and from 1 to 24 patches for isolated patches (median = 2). Patch occupancy was significantly higher in non-isolated patches than in isolated patches (Wilcoxon Z = 2.20, p = 0.028). Of the 44 species for which statistical testing was possible, the occurrence of 10 species was significantly (p ≤ 0.05) related to the isolation of the recent forest patch. Anemone nemorosa (wood anemone), Lamium galeobdolon (yellow archangel), Rubus idaeus (raspberry), Adoxa moschatellina and Athyrium filix-femina (lady fern) showed the highest significant difference (Fig 7.5).
Percentage of species
Restoration of forest habitat networks
125
a)
b)
50
50
40
40
30
30
20
20
10
10
0
0 0
5
10
15
20
25
30
0
5
10
15
20
25
30
Number of occupied patches
Fig. 7.4 Frequency distributions of species occurrence in a) non-isolated (distance to old woodland < 100m, n=30) and b) isolated (distance to old woodland > 100m, n=30) forest patches.
Discussion
Overall colonization success of forest plant species can be quantified by 1) the total number of species colonizing recent forest patches, 2) the number of patches occupied by a species, and 3) species richness of young forest patches. Whereas most other studies focused on species richness of recently established forests or on individual colonization patterns (e.g. Dzwonko 1993; Grashof-Bokdam 1997; Grashof-Bokdam and Geertsema 1998; Butaye et al. 2001), our study incorporated all three components of colonization success.
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**
Rubus fruticosus Stachys sylvatica Crataegus monogyna Sambucus nigra Ranunculus ficaria Corylus avellana Hedera helix
*
Viburnum opulus
Non-isolated Isolated
*
Stellaria holostea Dryopteris filix-mas Rosa sp.
* (*)
Geum urbanum Torilis japonica Chaerophyllum temulum
**
Adoxa moschatellina
**
Rubus idaeus Prunus spinosa Primula elatior Lapsana communis
Species
Epipactis helleborine
(*)
Ajuga reptans Scrophularia nodosa
(*)
Ribes rubrum Cirsium oleraceum Moehringia trinervia
(*)
Lamium galeobdolon
**
Deschampsia cespitosa
(*)
Cornus sanguinea Athyrium filix-femina
**
Rubus caesius Lonicera periclymenum
(*) (*) (*) *
Frangula alnus Arum maculatum Anemone nemorosa Alliaria petiolata
* * (*)
Humulus lupulus Circaea lutetiana Poa nemoralis Melandrium dioicum Rumex sanguineus Polygonatum multiflorum Hypericum dubium Cardamine flexuosa Scirpus sylvaticus
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Fig. 7.5 Individual species occurrence in isolated (distance to old woodland > 100m, n=30) and non-isolated (distance to old woodland < 100m, n=30) forest patches for 44 species found in at least 5 patches. Asterisks indicate significance of G-tests of independence: (*) P < 0.1; * P < 0.05, ** P < 0.01.
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Total species richness rapidly increased at low patch densities, but decreased at high densities. There is relatively little overlap in species diversity between forest patches when only a few random patches are selected. Given the fact that species similarity over a landscape decreases with increasing inter-patch distance (Jacquemyn et al. 2001b) and species diversity of recent patches is significantly related to the surrounding old patches (Butaye et al. 2002), establishment of more patches distributed over a landscape will result in a larger number of species that have been able to colonize recent patches. However, when a certain number of patches have been established, only occasionally will additional species be able to colonize these patches. Although the difference was not substantial, we found that total species richness was always higher for non-isolated patches than for isolated patches. Indeed, notwithstanding the number of occurrences, 92 % of the species present in nonisolated patches was also present in isolated patches. This means that only a limited number of species was unable to colonize recent patches when the distance between the old woodland and recent forests was larger than 100 meter. Species such as Brachypodium sylvaticum (false brome), Carex remota (remote sedge), Mercurialis perennis (dog’s mercury), Millium effusum (wood millet), and Rumex sanguineus (wood dock) had distributions limited typically to only a few old forest patches. However, when comparing total species of both isolated and non-isolated patches with that of non-isolated patches alone, no difference in total species richness was found. Thus, both isolated and non-isolated patches received almost the same number of species compared to the case when only non-isolated patches were considered. Patch species richness and species distribution patterns were significantly different between isolated and non-isolated forest patches. Non-isolated patches contained significantly more species than isolated patches. These results confirm earlier findings of Dzwonko (1993) and Butaye et al. (2001) who demonstrated that the distance to existing old woodland was largely determined species richness of recent forest patches. Frequency distribution patterns were also significantly different between isolated and non-isolated patches. Ten of the 44 investigated species showed higher occupancy in non-isolated than isolated forests suggesting that most species are able to disperse occasionally over distances larger than 100m. For species such as Anemone nemorosa, Lamium galeobdolon and Adoxa moschatellina, these findings
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confirm earlier results suggesting that some forest plant species are characterized by limited dispersal abilities and are only able to colonize forest patches adjacent to existing woodland. Hence, establishing new forests adjacent to existing woodland may be valuable to enlarge existing populations in order to decrease risks of extinction due to environmental, genetic or demographic stochasticity (Shaffer 1981; Pimm et al. 1988; Menges 1991). However, in terms of a forest habitat network, these populations contribute little to the regional population structure since most populations are nothing more than extensions of existing populations. Dispersal over distances larger than 100 m occasionally occurred, resulting in the colonization of isolated forest patches. Although species richness of isolated forest patches is significantly lower than that of non-isolated patches, the fact that forest plant species occasionally do cross distances larger than 100 m may have important consequences for forest habitat networks. Indeed, Clark (1998), Clark et al. (1999) and Higgins and Richardson (1999) demonstrated that even a very small proportion (0.001) of seeds dispersing over long distances can lead to a significant increase in predicted spread of plant species. Since these data were based on presence / absence data and no inferences were made about the size of colonizing populations, it remains still unclear whether colonization of recent, isolated stands results in viable populations of forest plant species. It is however reasonable to expect that most of these colonization events are one-off events, which are typified by small population sizes. McCauley et al. (1995), for example, demonstrated that, based on information from allozymes and cpDNA, relatively little mixing of individuals from source populations of Silene alba occurred during colonization. Due to their isolation, newly founded populations have been shown to be particularly sensitive to chance extinction as a consequence of their small effective population size (Newman and Pilson 1997). Therefore it can be expected that both the quantity and the quality of founding individuals may determine the viability of new populations. Although there is some evidence that genetic rescue by occasional influx of pollen or seeds may have significant effects on the persistence of isolated colonies (Richards 2000; Newman and Tallmon 2001), little is known about gene flow events across heterogeneous landscapes (Sork et al. 1999), or the demographic consequences of gene flow (McCauley et al. 2001). It must be stressed,
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however, that as a consequence of their clonal growth, most forest plant species may be less sensitive to chance extinction due to small effective population size.
Conclusion
The conclusion of this research is that no high species richness is to be expected in isolated stands in the short term. When land and money are limited, management should be directed in such a way that a maximum of existing old forests are conserved (Thomas et al. 1997). If this is achieved, supplemental efforts can be invested in restoration projects. In the case of the latter, restoration of forest habitat adjacent to existing woodland is the best option because it ensures higher survival probabilities of existing populations. Maximizing the number of forests and establishing new forests isolated from existing woodland therefore may seem rather pointless because 1) 90 % of all species already occurred in half of the forest patches established in the study area, 2) distribution patterns of typical forest plant species were significantly higher in non-isolated patches and 3) species richness of non-isolated forests was significantly higher than of isolated patches. In the long term, however, and in landscapes with relatively small amounts of existing forest, both isolated and non-isolated new forests should be established if forest restoration seeks to construct a network with a minimum amount of new woodland. The benefits of this policy can be found in the fact that occasional long-distance seed dispersal leading to the colonization of isolated stands does occur. In terms of the total number of species, establishing both isolated and non-isolated patches has little effect on the overall number of species able to colonize recent patches compared with the situation in which only non-isolated forest patches are established. Moreover, a clear distinction has to be made between populations colonizing adjacent patches and patches isolated from old woodland. The former refers to expanding, existing populations, whereas the latter are essentially new populations. The importance of these new populations must not be underestimated. Indeed, they can affect gene flow in a landscape by increasing pollen and seed flow and strengthen metapopulation structure (Thrall et al. 1998). However, several models have shown that the genetic consequences of founding events depend both on the number of individuals involved in the founding event and on the number of source populations from which they are 129
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drawn (Slatkin 1977; Whitlock and McCauley 1990). More research focusing on the genetic traits and the viability of founding populations are needed to fully understand the effectiveness of forest habitat networks.
Acknowledgments We’d like to thank Myriam Dumortier for assisting with the data collection. Etienne Jacquemyn kindly provided the randomization algorithm. H. J. received financial support from the Flemish Institute for Scientific Research (FWO) and J. B. from the Flemish Community, AMINAL, Dept. Nature.
8. Summary, guidelines for restoration and conservation, and perspectives for future research Natuur is voor tevredenen of legen. En dan: wat is natuur nog in dit land? Een stukje bos, ter grootte van een krant, Een heuvel met wat villaatjes ertegen.
J. C. Bloem
Summary
In 1991, the British ecologist Jonathan Silvertown remarked that for decades plant population ecology had stubbornly stuck to the individual population as the focus of study, treating different parts of the population as replicates or as entirely separate entities, and not as potentially interdependent units. The recent rise of metapopulation theory as the new paradigm in population ecology (Hanski and Simberloff 1997) has led to an increasing number of studies (e.g. Menges 1990, van der Meijden et al. 1992, Ouborg 1993, Antonovics et al. 1994, Giles and Goudet 1997a,b, Valverde and Silvertown 1997, Husband and Barrett 1998, Harrison et al. 2000) that have emphasized the importance of regional dynamics in structuring plant populations. However, it is not clear whether all studies mentioned above were actually able to show clear metapopulation structures. Recent reviews by Bullock et al. (2002) and Freckleton and Watkinson (2002) suggest that this might not have been the case. In addition, Freckleton and Watkinson (2002) have shown that a large variation in types of regional plant population structure may exist, and that the metapopulation is just only one form of large-scale plant population distribution. The general aim of this study (Chapter 1) was to get a better understanding of the factors affecting regional distribution patterns of forest plant species in a fragmented landscape subject to change and to suggest guidelines for forest conservation and restoration in our present-day, cultural landscapes. In order to achieve this aim, a multifaceted approach was adopted. First, landscape changes were
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meticulously reconstructed using all available historical, topographic maps (Chapter 1). This allowed us to determine the approximate age of each forest patch present in the landscape. The effects of area, age and distance on species richness and community composition of forest patches were then assessed (Chapter 2 and 3). Further, the impact of patch area, connectivity and age, and local site conditions on regional distribution patterns were compared among 59 forest plant species and the importance of individual plant traits in determining patch occupancy was assessed (Chapter 4). Finally, more in-depth knowledge of the effects of fragmentation and landscape changes on population viability and genetic diversity and structure of forest plant species was determined for the perennial herb Primula elatior (Chapter 5 and 6). The amount of forest cover in the study area showed considerable changes during the last three centuries (Chapter 1). Whereas forest area decreased spectacularly between 1775 and the first half of the nineteenth century, from 1850 onwards deforestation rates gradually decreased and progressively more and more new patches have been established in the study area, mostly on former agricultural lands. These new forests have been steadily colonized by forest plant species, as was shown by the relationships between species number and patch age (Chapter 2). For both forest types (viz. Quercion and Alno-Padion forests), large patches received faster more species than did small patches. This might be explained by the fact that larger patches are easier colonized than smaller patches, and once colonized, support larger population sizes (i.e. the area per se argument) or by the fact that larger patches provide a wider range of habitat types, thus allowing successful colonization by and supporting viable populations of a larger number of ecologically dissimilar species than smaller patches (i.e. the habitat diversity argument) (Ricklefs and Lovette 1999, Honnay et al. 1999a). Almost half of the species showed aggregated distribution patterns (Chapter 2), which suggests that dispersal processes interfere with the process of community assembly. The importance of dispersal in structuring local plant communities was further confirmed by the fact that for both Quercion and AlnoPadion forests, community dissimilarity significantly increased with increasing distance (Chapter 3). Indeed, dissimilar community composition in environmentally similar sites should be most prominent in systems with large species pools, low levels
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of dispersal and low rates of disturbance (Chase 2003). Hence, dispersal limitation should be invoked as a major factor determining species distributions in a landscape. Species richness and composition therefore cannot be explained without accounting for dispersal processes occurring at the regional scale. Studying colonization patterns of forest plant species, Verheyen et al. (2003b) recently reached similar conclusions. These findings imply that temporal and spatial organization of forest patches within a landscape, species composition of neighbouring forest patches and characteristic dispersal distances of plant species, all contribute to local community composition of forest patches. However, this does not mean that local factors should not be important in structuring local plant communities, but the results of this study show that the magnitude of the effects of local site conditions on community assembly and forest succession differ according to the type of forest studied. Local site conditions appeared to be more important in the less productive Quercion forests than they were in the more productive Alno-Padion forests. Indeed, local species richness of forest patches increased at a greater rate with increasing age in Alno-Padion forest than in Quercion forests (Chapter 2). Within less than a century, species numbers in AlnoPadion forests were comparable to that of forests that were already present since 1775. In Quercion forests on the other hand, species accumulation occurred at a much lower rate and even after 150 years species numbers were significantly different of those in old forests. Moreover, for Quercion forests, significant effects of age on community similarity were found, even after distance effects were excluded, whereas no such effects were detected for Alno-Padion forests. These results suggest that local site conditions had a large impact on succession and community assembly of Quercion forests, but to a lesser extent of Alno-Padion forests. Recently, Verheyen et al. (2003c) came to similar conclusions when comparing species richness and composition of plots located in Quercion and Alno-Padion forests, respectively. Consistent with the results presented here, few changes in community properties were found after 70 yrs of colonization in Alno-Padion forests. In Quercion forests on the other hand, species richness and plot dissimilarity still changed 120 years after forests were established. The results presented so far were further confirmed when distribution patterns of individual forest plant species were considered (Chapter 4). For most
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species, distribution patterns were determined by both local and regional variables. In most cases, the occurrence of forest plant species was affected by local soil conditions (soil moisture content and nutrients), in some cases also by light conditions related to successional changes in crown development of forest trees. Significant connectivity effects again confirm our earlier findings that most forest plant species have limited colonization capacities and that dispersal is a limiting factor in determining the occurrence of forest plant species in a fragmented landscape. The significant area effects suggest that the capacity of forest plant species to maintain viable populations within the remaining patches is related to the area of a patch. Finally, age significantly affected species occurrence of half of the species studied, which may suggest the existence of time lags between patch generation and effective colonization (Tilman et al. 1997, Huxel and Hastins 1999, Johnson 2000, Johst et al. 2002). Due to this time lag, landscapes may have been changing too fast to allow species to maintain viable (meta)populations within this landscape. This may be especially true for species with limited dispersal capacities, as these species have great difficulties in colonizing recently restored patches, especially when they are isolated from existing old forests (Keymer et al. 2000, Johst et al. 2002). Species that were most capable of dealing with altered landscape structures were tall shrub species that produced large-bodied, fleshy fruits, which are mainly dispersed by birds (e.g. Sambucus nigra, Crataegus monogyna, Viburnum opulus, etc.). In addition, it was shown that altered landscape structures also affect reproductive success and genetic diversity and structure of forest plant species (Chapter 5 and 6). Small populations produced significantly less fruits and seeds per individual than large populations, which may increase their risk of local extinction and reduce the probability of their seeds being effectively dispersed. These results also suggest that local extinctions of plant populations may not be as enigmatic as previously thought. In Chapter 6, it was further demonstrated that not all populations were equally likely to contribute to gene flow, and that some populations (especially those at higher locations in the study area) were isolated from the rest of the metapopulation. Loss of gene flow between populations may imply that they become genetically isolated, which in the long run may lead to lower fitness and ultimately to increased risks of extinction (Richards 2000).
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Guidelines for conservation and restoration
Restoration ecology has at its core the assumption that habitat degradation is to a certain extent temporary, and that some proportion of habitat loss and population decline is recoverable (Young 2000). In the context of forests, forest loss seems reversible as it was shown in this and other studies that newly established patches slowly recovered to their original state, although the trajectory and time needed to reach this point may differ from one system to the next. One of the more specific aims of restoration ecology is to identify the major factors that hamper successful restoration of degraded lands and to provide tools for speeding the recovery of these lands (Dobson 1997). Conservation biology, on the other hand, is mainly concerned with the conservation of existing populations and mainly focuses on individual species (e.g. Caughley 1994, Schemske et al. 1994, Young 2000). It has become clear that this narrow-minded species view will not be sufficient to save species from extinction in the long term and that therefore whole ecosystems are to be conserved (Franklin 1993). However, restoration ecology and conservation biology often go hand in hand as the restoration of degraded lands cannot be obtained if the conservation of existing habitats (i.e. source habitats), from which colonization is most likely to take place, cannot be guaranteed. Results from the present study have shown that both local and regional processes determined community composition and species richness of forest patches, but that their relative importance differed according to the forest type studied. This has important implications for restoration and conservation. Lockwood and Pimm (1999) and Young et al. (2001) have argued that if local conditions are the sole determinants of community composition, restoration efforts should mainly focus on returning the initial site conditions. If, on the other hand, regional processes or both local and regional processes determine community composition, it is clear that both local conditions and regional processes should be restored. This means that, in the case of less-productive Quercion forests, restoration can only be successful if local site conditions are improved (e.g. by removing the thick litter layer, diminishing competition of ruderal species and providing suitable places for germination and establishment). Moreover, as the complete regeneration of these forests to their original habitat may take many centuries (Pigott 1977, Peterken 1974, 1977, Honnay
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et al. 1999a, Hermy et al. 1999), restoration seems pointless when the conservation of old forests cannot be guaranteed. As several species typically confined to this type of forest (e.g. Convallaria majalis, Luzula pilosa, L. multiflora, Maianthemum bifolium) hardly showed any colonization, conservation of the remaining populations seems crucial for their long-term survival. Diminishing stochastic perturbations (e.g. due to forest practices), providing conditions that increase population growth rates (e.g. by creating suitable germination conditions), decreasing deterministic threats due to edge effects or deteriorating environmental conditions (e.g. vigorous growth of competitive species such as Rubus fruticosus coll. and Deschampsia flexuosa), all may reduce extinction risks of local populations of these species. In contrast and despite the smaller amounts of old forests in the alluvial part of the study area (contradicting results of Vellend 2003), recently established forest patches in the alluvial part of the landscape were colonized quite easily, especially when they were located in the vicinity of existing old forests (Chapter 7). Forests reached species numbers comparable to those of old forest patches within less than 200 years. Hence, the restoration of new forest patches in this part of the landscape may be a successful measure to maintain or even enlarge existing viable (meta)populations of forest plant species and to create forest networks, which allow easy migration of plant species. Increased patch density may also increase gene flow (Chapter 6), and diminish the probability that populations become genetically isolated. However, for most regionally rare species (e.g. Allium ursinum, Campanula trachelium, Crepis paludosa, Cardamine amara, Fragaria vesca, Paris quadrifolia, Polygonum bistorta and Ranunculus auricomus), these measures will not be sufficient and a more species-specific approach will be needed. Determining most important problems and defining conservation targets for each species specifically are needed to guarantee their long-term survival. Distance decay of similarity further implies that each individual forest has an almost unique community composition. Because of this uniqueness, the importance of small forest patches for the conservation of forest plant species in highly fragmented landscapes must not be underestimated (Nekola and White 1999). Rather than conserving a few, very large forest areas, we would therefore recommend the preservation of a wide range of smaller, dispersed forests that represent a broad range of topographic and hydrologic variation. Honnay et al. (1999a), using a different
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approach, came to similar conclusions. If the colonization of recently established forest patches is not hampered by local site conditions, maximizing the number of patches may increase gene flow by both pollen and seeds and strengthen regional population structure. However, if local site conditions are the prevailing conditions determining colonization success, more efforts should be invested in the conservation of existing forests and their populations.
Perspectives for future research
It may be clear that the results of this study have only revealed a glimpse of the possible effects of landscape fragmentation on plant species distribution and viability, and that several other and important aspects of plant population ecology have received very little attention in this study. In the remainder of this chapter, I will draw attention to some shortcomings in current research of regional plant dynamics and distributions and offer some suggestions for further research. Our understanding of the complex effects of landscape dynamics and fragmentation on the regional abundance of plant species is most likely to increase when we step beyond simple pattern studies and delve ourselves into the processes that have caused them. The latter relate to among-population processes, genetic processes as well as to within-population processes (Fig. 8.1). Whereas presence / absence data may be suitable to make inferences on colonization patterns and species accumulation in forest patches and to provide practical guidelines for forest restoration and conservation, they give little information on more specific questions as: Where did seeds originate from? How and how many of them were dispersed? How did populations establish and survive in habitat patches the first years after colonization? Do all colonization events succeed and do all of them lead to viable plant populations? How many colonization events are needed before a population becomes a viable population? Do populations, once established, survive forever or are local extinctions common phenomena among forest plant species? These questions urge a more process-oriented approach and a broadening of the plant population ecologists view beyond the single population (Silvertown 1991). Fig. 8.1 gives an overview of the processes (and their relationships) likely to affect regional persistence
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of plant populations in landscapes characterized by frequent turnover of habitat patches.
Changes in land use
Changes in area, connectivity and habitat quality of forest patches and the total amount of forests in the area
Among-population processes
Genetic processes (e.g. gene flow, inbreeding, ...)
(e.g. colonization, extinction, migration, ...)
Within-population processes (e.g. pollination, recruitment, ...)
Regional persistence of plant populations
Species richness and community composition of forests patches Fig. 8.1 Schematic representation of the effects of changing land use practices on forest structure and the processes that affect regional persistence of plant populations in dynamic landscapes.
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Regional population structure of plants: are plant populations metapopulations?
The basic formulation of metapopulation theory is based on three simple facts: (i) populations are prone to local extinction, (ii) patches are recolonized after extinction, and (iii) populations exhibit some degree of independent dynamics (Eriksson 1996, Hanski 1999, Bullock et al. 2002, Freckleton and Watkinson 2002). Thus, the regional distribution of populations in a landscape are controlled by a balance of local dynamics and linked landscape-level colonization and extinction events. For plant species, the metapopulation concept has been proven problematic (Bullock et al. 2002, Freckleton and Watkinson 2002) and several other types of regional population structures have been proposed, the most prominent being spatially extended populations (a basically continuous population existing on a large area of suitable habitat in which migration is nonexistent and dynamics are the product of local processes) and regional ensembles (a series of local populations that show hardly any or no migration between sites and may be highly persistent) (Freckleton and Watkinson 2002). Applying this terminology to the results from this study (Chapter 4) suggests that different plant species show different regional population structures. The regional structure of the clonal forest herb Maianthemum bifolium may be best described as a ‘regional ensemble’, as this species showed no migration or colonization and because local populations may be highly persistent. The forest shrub Rubus fruticosus coll. on the other hand may be a good example of a plant species with a regional population structure that can be best described as a ‘spatially extended system’. The species shows no specialized habitat requirements and colonization was spatially dependent (see Butaye et al. 2001). Within-patch dynamics are most likely to be determined by local density-dependence. A species most likely to show a metapopulation structure is the forest herb Primula elatior. Suitable habitat patches occur in discrete patches (Chapter 5), local populations do not have completely synchronous dynamics and local habitat patches are not too isolated to become colonized. Ehrlén and Eriksson (2003) have criticized this typology by arguing that conflation of these separate notions is most likely to engender some confusion in ecological literature, rather than resolving the problems. Most prevalent problems relate to difficulties of quantifying extinction rates and determining the proportion of recruits that are derived by immigration.
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A) Quantification of extinction rates
One of the challenges in understanding the regional population structure of plant species lays in the quantification of extinction rates. The easiest way to measure extinction is by monitoring large numbers of natural or artificially created populations. One of the few studies to date found that local populations of the annual plant Eichhornia paniculata showed frequent extinction and that patch occupancy values varied considerably from one year to the next (Husband and Barrett 1998). However, because of the long life spans of most forest herbs, extinction may be seriously delayed and monitoring a large number of populations may not be sufficient to assess extinction rates. In this case, other methods are to be used. One such method consists of applying matrix population models and stochastic simulations (Caswell 2001, Fieberg and Ellner 2001). These models make substantial use of matrix algebra and require data collected for three or more consecutive years. Simulation of population projections provides a means of estimating the probability of extinction, which in turn can be used in ecological models in which spatial structure is explicit. At the same time, these models allow estimating the lag between fragmentation and local extinction (Ehrlén and Eriksson 2001, Gu et al. 2002, Hanski and Ovaskainen 2002).
B) Seed dispersal and colonization
The importance of long-distance seed dispersal in determining patch occupancy of plant species has been widely acknowledged. However, in reality measures of longdistance seed dispersal may be extremely hard to obtain, especially when it has recently become clear that a poor relationship existed between the morphology of dispersal units and their standard means of dispersal (Higgins et al. 2003). In several cases, non-standard mechanisms of dispersal appeared to be more important in explaining dispersal patterns of plant species than their expected means of dispersal based on seed morphology (reviewed in Higgins et al. 2003). Hence, in reality it is hard to make general predictions about long-distance seed dispersal based on morphological dispersal syndromes alone. Nor Butaye et al. (2001) or Verheyen et al. (2003) could explain colonization capacity of forest plant species based on
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morphological dispersal syndromes. These findings urge more detailed knowledge about the exact mechanisms generating long-distance dispersal. Vellend et al. (2003) found that deer were involved in the long-distance seed dispersal of the myrmecochorous Trillium grandiflorum. Using an animal-movement – seed-retention model, it was shown that seeds moved by deer should travel at least several hundred meters from parent plants. A more spatially explicit model could be obtained if home ranges and movement of dispersal vectors are combined with information on source and target patches of the plants being dispersed. Next, it is reasonable to expect that not all populations produce the same amounts of seeds, as was clearly shown in Chapter 5. In two different regions, Honnay et al. (unpublished results) found no colonization of the forest herb Maianthemum bifolium. In both regions, mature plants produced a very low amount of seeds (only 0.006 % of all flowers produced a berry). In this case, the absence of M. bifolium in recently established forest patches can be almost entirely explained by fecundity limitation (sensu Bolker et al. 2003), and not by dispersal limitation. Thus, a more realistic model of long-distance seed dispersal would be obtained if fecundity were included in statistical and mechanistic models of seed dispersal. Modeling studies, however, only provide indirect extrapolations of observed patterns. Two other approaches that may be very useful to facilitate direct measurements of actual dispersal are stable isotope analysis and molecular genetic techniques (reviewed in Ouborg et al. 1999, Sork et al. 1999, Cain et al. 2000, Wang and Smith 2002 and Nathan et al. 2003). Molecular markers that are specifically inherited only through seeds (e.g. chloroplast DNA) may be particularly valuable to reveal the importance of seed flow in determining genetic relationships between populations, and to reveal estimates of dispersal distances. Comparing both chloroplast and nuclear markers, on the other hand, may show the relative contributions of seed and pollen flow on gene flow in plant populations (McCauley 1994).
Conclusion
We agree that the study of the regional distribution of plant species is still in its infancy (Ehrlén and Eriksson 2003) and that most progress is likely to be made when different disciplines (plant population ecology, molecular genetics, landscape ecology,
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etc.) are combined in a synergetic way to more efficiently quantify dispersal, colonization, recolonization and extinction processes in a range of plant populations. Only if these processes are better understood, more specific guidelines for land use planning, restoration and conservation can be offered.
References Ågren, J. (1996) Population size, pollinator limitation, and seed set in the self-incompatible herb Lythrum salicaria. Ecology 77: 1779-1790. Aizen, M. and Feinsinger, P. (1994) Forest fragmentation, pollination, and plant reproduction in a chaco dry forest, Argentina. Ecology 75: 330-351. Andrén, H. (1996) Population responses to habitat fragmentation : statistical power and the random sample hypothesis. Oikos 76: 235-242. Antonovics, J., Thrall, P., Jarosz, A. and Stratton, D. (1994) Ecological genetics of metapopulations: the Silene-Ustilago plant-pathogen system. In: Real, L. A. (ed.) Ecological Genetics. Princeton University Press, Princeton, New Jersey. pp. 146-170. Antrobus, S. and Lack, A. J. (1993) Genetics of colonizing and established populations of Primula veris. Heredity 71: 252-258. Barrett, S. C. H. and Kohn, J. R. (1991) Genetic and evolutionary consequences of small population size in plants : implications for conservation. In: Falk, D. A. and Holsinger, K. H. (eds.) Genetics and Conservation of Rare Plants. Oxford University Press, Oxford, pp 3-30. Bastin, L. and Thomas, C. D. (1999) The distribution of plant species in urban vegetation fragments. Landscape Ecology 14: 493-507. Berge, G., Nordal, I. and Hestmark, G. (1998) The effect of breeding systems and pollination vectors on the genetic variation of small plant populations within an agricultural landscape. Oikos 81: 17-29. Bierzychudek, P. (1982) Life histories and demography of shade-tolerant temperate forest herbs: a review. New Phytologist 90: 757-776. Blomqvist, M. M., Vos, P., Klinkhamer, P. G. L. and ter Keurs, W. J. (2003) Declining plant species richness of grassland ditch banks – problem of colonisation or extinction? Biological Conservation 109: 391-406. Bossuyt, B, Hermy, M. and Deckers, J. (1999) Migration of herbaceous plant species across ancient-recent forest ecotones in central Belgium. Journal of Ecology 87: 628-638. Bossuyt, B. and Hermy, M. (2000) Restoration of the understorey layer of recent forest bordering ancient forest. Applied Vegetation Science 3: 43-50. Brown, A. H. F. and Warr, S. J. (1992) The effect of changing management on seed bank in ancient coppices. In Buckley, G. P. (ed.) Ecology and management of coppice woodland. Chapman and Hall, New York. pp. 147-166
143
144
References
Brunet, J. and von Oheimb, G. (1998) Migration of vascular plants to secondary woodlands in southern Sweden. Journal of Ecology 86: 429-438. Bullock, J. M., Moy, I. M., Pywell, R. F., Coulson, S. J., Nolan, A. M. and Caswell, H. (2002) Plant dispersal and colonization processes at local and landscape scales. In: Bullock, J. M., Kenward, R. and Hails, R. (eds.). Dispersal Ecology. Blackwell Science, London. pp. 279-302. Burel, F. and Baudry, J. (1990) Hedgerow networks as habitat for species : Implications for colonising abandoned agricultural land. In: Bunce, R.G.H. and Howard, D.C. (eds.) Species dispersal in agricultural habitats, pp. 82-97. Belhaven press, London. Burgess, R. L. and Sharpe, D. M. (1981) Introduction. In: Burgess, R. L. and Sharpe, D. M. (eds). Forest island dynamics in man-dominated landscapes. Springer-Verlag, NewYork. pp. 1-5 Burgman, M. A. (1987) An analysis of the distribution of plants on granite outcrops in southern Western Australia using Mantel tests. Vegetatio 71 : 79-86. Butaye, J., Jacquemyn, H. and Hermy, M. (2001) Differential colonization causing nonrandom forest plant community structure in a fragmented agricultural landscape. Ecography 24: 369-380. Butaye, J., Jacquemyn, H., Honnay, O. and Hermy, M. (2002) The species pool concept applied to forests in a fragmented landscape: dispersal limitation versus habitat limitation. Journal of Vegetation Science 13: 27-34. Boyd, M., Silvertown, J. and Tucker, C. (1990) Population ecology of heterostyle and homostyle Primula vulgaris: growth, survival and reproduction in field populations. Journal of Ecology 78: 799-813. Cain, M. L., Milligan, B. G. and Strand, A. E. (2000) Long-distance seed dispersal in plant populations. American Journal of Botany 87: 1217-1227. Campbell, D., Duchesne, P. and Bernatchez, L. (2003) AFLP utility for population assignment
studies:
analytical
investigation
and
empirical
comparison
with
microsatellites. Molecular Ecology 12: 1979-1991. Caswell, H. (2001) Matrix population models. Construction, analysis and interpretation. Sinauer. Caughley, G. (1994) Directions in conservation biology. Journal of Animal Ecology 63: 215244. Chase, J. M. (2003) Community assembly: when should history matter? Oecologia 136: 489498. Clark, J. S. (1998) Why trees migrate so fast: confronting theory with dispersal biology and the paleorecord. American Naturalist 152: 204-224.
References
145
Clark, J. S., Silman, M., Kern, R., Macklin, E. and RisLambers, J. H. (1999) Seed dispersal near and far: patterns across temperate and tropical forests. Ecology 80: 1475-1494. Coart, E., Lamote, V., De Loose, M., Van Bockstaele, E., Lootens, I. and Roldán-Ruiz, I. (2002) AFLP markers demonstrate genetic differentiation between indigenous oak species (Quercus robur L. and Quercus petraea (Matt.) Liebl.) in Flemish populations. Theoretical and Applied Genetics 105: 431-439. Cody, M. L. (1993) Bird diversity components within and between habitats in Australia. In : Ricklefs, R. E. and Schluter, D. (eds.) Species Diversity in Ecological Communities, pp. 147 -158. The University of Chicago Press, Chicago. Condit, R., Ashton, P. S., Baker, P., et al. (2000) Spatial patterns in the distribution of tropical tree species. Science 288: 1414-1418. Condit, R., Pitman, N., Leigh, E. G., Chave, J., Terborgh, J., Foster, R. B., Nunez, P., Aguilar, S., Valencia, R., Villa, G., Muller-Landau, H. C., Losos, E. and Hubbell SP (2002) Bèta-diversity in tropical forest trees. Science 295: 666-669. Corbit, M., Marks, P. L. and Gardescu, S. (1999) Hedgerows as habitat corridors for forest herbs in central New York, USA. Journal of Ecology 87: 220-232. Curtis, J. T. (1956) The modification of mid-latitude grasslands and forests by man. In: Thomas, W. L. (ed.). Man’s role in changing the face of the earth. University of Chicago Press, Chicago, Illinois. pp. 721-736. De Keersmaeker, L., Rogiers, N., Lauriks, R. and De Vos, B. (2001) Ecosysteemvisie bos in Vlaanderen, ruimtelijke uitwerking van de natuurlijke bostypes op basis van bodemgroeperingseenheden en historische boskaarten. VLINA C97/06, Brussel. DeMauro, M. M. (1993) Relationship of breeding system to rarity in the lakeside daisy (Hymenoxys acaulis var. glabra). Conservation Biology 7: 542-550. Després, L., Loriot, S. and Gaudeul, M. (2002) Geographic pattern of genetic variation in the European globeflower Trollius europaeus L. (Ranunculaceae) inferred from amplified fragment length polymorphism markers. Molecular Ecology 11: 2337-2347. Didham, R. K., Ghazoul, J., Stork, N. E. and Davis, A. J. (1996) Insects in fragmented forests : a functional approach. Trends in Ecology and Evolution 11: 255-260. Dierschke, H. (1968) Zur synsystematischen und syndynamischen Stellung einiger CalthionWiesen mit Ranunculus auricomus L. und Primula elatior (L.) Hill in Wümme-Gebiet. Mitteilungen der Floristisch-soziologischer Arbeitsgemeinschaft Neue Fassung 13: 5970. Dietz, E. J. (1983) Permutation tests for association between two distance matrices. Systematic Zoology 32: 21 – 26.
145
146
References
Dobson, A. P., Bradshaw, A. D. and Baker, A. J. M. 1997. Hopes for the future: Restoration ecology and conservation biology. Science 277: 512-522. Drake, J. A., Zimmermann, C. R., Purucker, T. and Rojo, C. (1999) On the nature of the assembly trajectory. Ecological Assembly Rules: Perspectives, advances, retreats. (ed. by Weiher, E. and Keddy, P.). pp. 233-250. Cambridge University Press. Dupré, C. and Diekmann, M. (1998) Prediction of occurrence of vascular plants in deciduous forests of South Sweden by means of Ellenberg indicator values. Applied Vegetation Science 1: 139-150. Dupré, C. and Ehrlén, J. (2002) Habitat configuration, species traits and plant distributions. Journal of Ecology 90: 796-805. Dybdahl, M. F. (1994) Extinction, recolonisation, and the genetic structure of tidepool copepod populations. Evolutionary Ecology 8: 113-124. Dzwonko, Z. (1993) Relations between the floristic composition of isolated young woods and their proximity to ancient woodland. Journal of Vegetation Science 4: 693-698. Dzwonko, Z. (2001) Assessment of light and soil conditions in ancient and recent woodlands by Elenberg indicator values. Journal of Applied Ecology 38: 942-951. Dzwonko, Z. and Loster, S. (1988) Species richness of small woodland on the Western Carpathian foothills. Vegetatio 76: 15-27. Dzwonko, Z. and Loster, S. (1989) Distribution of vascular plants species in small woodlands on the Western Carpathian foothills. Oikos 56: 77-86. Ebert, D., Haag, C., Kirkpatrick, M., Riek, M., Hottinger, J. W. and Pajunen, V. I. (2002) A selective advantage to immigrant genes in a Daphnia metapopulation. Science 295: 485488. Edwards, D. and Berry, J. J. (1987) The efficiency of simulation-based multiple comparisons. Biometrics 43: 913-928. Ehrlén, J. and Eriksson, O. (2000) Dispersal limitation and patch occupancy in forest herbs. Ecology 81: 1667-1674. Ehrlén, J. and Eriksson, O. (2003) Large-scale spatial dynamics of plants: a response to Freckleton and Watkinson. Journal of Ecology 91: 316-320. Ehrlén, J. and Lehtilä, K. (2002) How perennial are perennial plants? Oikos 98: 308-322. Eisto, A., Kuitunen, M., Lammi, A., Saari, V., Suhonen, J., Syrjäsuo, S. and Tikka, P. M. (2000) Population persistence and offspring fitness in the rare bellflower Campanula cervicaria in relation to population size and habitat characteristics. Conservation Biology 14: 1413-1421. Ellenberg, H., Weber, H. E., Düll, R., Wirth, V., Werner, W. and Pauliβen, D. (1992) Zeigerwerte von Pflanzen in Mitteleuropa. Scripta Geobotanica 18: 1-248.
References
147
Ellstrand, N. C. and Elam, D. R. (1993) Population genetic consequences of small population size : implications for plant conservation. Annual Review of Ecology and Systematics 24: 217-242. Eriksson, A., Eriksson, O. and Berglund, H. (1995) Species abundance patterns of plants in Swedish semi-natural pastures. Ecography 18: 310-317. Eriksson, O. (1996) Regional dynamics of plants : a review of evidence for remnant, sourcesink and metapopulations. Oikos 77: 248-258. Eriksson, O. (2000) Functional roles of remnant plant populations in communities and ecosystems. Global Ecology and Biogeography 9: 443-449. Eriksson, O. and Ehrlén, J. (2001) Landscape fragmentation and the viability of plant populations. In: Silvertown, J. and Antonovics, J. (eds.) Integrating Ecology and Evolution in a Spatial Context. Blackwell, Oxford. pp. 157-175. ESRI (1995) Arc/Info 7.1 Environmental Systems Research Institute, Redlands, California. Fahrig, L. (1992) Relative importance of spatial and temporal scales in a patchy environment. Theoretical Population Biology 41: 300-314. Fieberg, J. and Ellner, S. P. (2001) Stochastic matrix models for conservation and management. Ecology Letters 4: 244-266. Fischer, M and Matthies, D. (1998) Effects of population size on performance in the rare plant Gentianella germanica. Jurnal of Ecology 86: 195-204. Franklin, J. F. (1993) Preserving biodiversity: species, ecosystems or landscapes? Ecological Applications 3: 202-205. Freckleton, R. P. and Watkinson, A. R. (2002) Large-scale spatial dynamics of plants: metapopulations, regional ensembles and patchy populations. Journal of Ecology 90: 419-434. Freckleton, R. P. and Watkinson, A. R. (2003) Are all plant populations metapopulations? Journal of Ecology 91: 321-324. Gaudet, C. L. and Keddy, P. A. (1988) A comparative approach to predicting competitive ability from plant traits. Nature 334: 242-243. Gaudeul, M., Taberlet, P. and Till-Bottraud, I. (2000) Genetic diversity in an endangered alpine plant, Eryngium alpinum L. (Apiaceae), inferred from amplified fragment length polymorphism markers. Molecular Ecology 9: 1625-1637. Gehring, J. L. and Delph, L. J. (1999) Fine scale genetic structure and clinal variation in Silene acaulis despite high gene flow. Heredity 82: 628-637. Giles, B. E. and Goudet, J. (1997a) Genetic differentiation in Silene dioica metapopulations: estimation of spatiotemporal effects in a successional plant species. American Naturalist 149: 507-526.
147
148
References
Giles, B. E. and Goudet, J. (1997b) A case study of genetic structure in a plant metapopulation. In: Hanski, I. and Gilpin, M. E. (eds.) Metapopulation biology: ecology, genetics, and evolution. Academic Press, San Diego, California. pp. 429-454. Gilpin, M. (1991) The genetic effective size of a metapopulation. Biological Journal of the Linnean Society 42: 165-175. Godt, M. J. W. and Hamrick, J. L. (1991) Genetic variation in Lathyrus latifolius (Leguminosae). American Journal of Botany 78: 1163-1171. Goovaerts, P., Frankart, R and Gérard, G. (1990) Effet de la succession de differente affectations sur les propriétés de pedons en Fagne de Chimay (Belgique). Pédologie XL-2 : 179 – 194. Grashof-Bokdam, C. (1997) Colonization of forest plants: the role of fragmentation. Wageningen, DLO Institute for forestry and Nature Research (IBN-DLO). Grashof-Bokdam, C. (1997) Forest species in an agricultural landscape in the Netherlands : Effects of habitat fragmentation. Journal of Vegetation Science 8: 21-28. Grashof-Bokdam, C. J. and Geertsema, W. (1998) The effect of isolation and history on colonization patterns of plant species in secondary woodland. Journal of Biogeography 25: 837-846. Grime, J. P., Hodgson, J. G. and Hunt, R. (1988) Comparative Plant Ecology. A functional approach to common British species. Unwin-Hyman. Groom, M. J. Allee effects limit population viability of an annual plant. American Naturalist 151: 487-496. Gu, W., Heikkilä, R. and Hanski, I. (2002) Estimating the consequences of habitat fragmentation on extinction risk in dynamic landscapes. Landscape Ecology 17: 699710. Haila, Y. (1999) Islands and fragments. In: Hunter ML (ed) Maintaining Biodiversity in Forest Ecosystems. Cambridge University Press, Cambridge, pp 234-236 Hanski, I. (1994) Patch-occupancy dynamics in fragmented landscapes. Trends in Ecology and Evolution 9: 131-135. Hanski, I. (1999) Metapopulation Ecology. Oxford University Press, Oxford. Hanski, I. and Gilpin, M. (1991) Metapopulation dynamics : brief history and conceptual domain. Biological Journal of the Linnean Society 42: 3-16. Hanski, I. and Gyllenberg, M. (1993) Two general metapopulation models and the coresatellite species hypothesis. American Naturalist 142: 17-41. Hanski, I. and Gilpin, M. E. (1997) Metapopulation biology: ecology, genetics and evolution. Academic Press, San Diego, CA.
References
149
Hanski, I. and Simberloff, D. (1997) The metapopulation approach, its history, conceptual domain, and application to conservation. In: Hanski, I. and Gilpin, M. E. (eds.). Metapopulation biology ecology, genetics and evolution. Academic Press, San Diego, CA. pp. 5-26. Hanski, I. and Ovaskainen, O. (2000) The metapopulation capacity of a fragmented landscape. Nature 404: 755-758. Hanski, I. and Ovaskainen, O. (2002) Extinction debt at extinction threshold. Conservation Biology 16: 666-673. Harrison, S., Ross, S. J. and Lawton, J. H. (1992) Beta diversity on gradients in Britain. Journal of Animal Ecology 61: 151-158. Harrison, S. and Hastings, A. (1996) Genetic and evolutionary consequences of metapopulation structure. Trends in Ecology and Evolution 11: 180-183. Harrison, S., Maron, J. and Huxel, G. (2000) Regional turnover and fluctuation in populations of five plants confined to serpentine seeps. Conservation Biology 14: 769-779. Harrison, S. and Inouye, B. D. (2002) High beta diversity in the flora of Californian serpentine ‘islands’. Biodiversity and Conservation 11: 1869-1876. Hastings, A. and Harrison, S. (1994) Metapopulation dynamics and genetics. Annual Review of Ecology and Systematics 25: 167-188. Helliwell, D. R. (1976) The effects of size and isolation on the conservation value of wooded sites in Britain. Journal of Biogeography 3: 407 – 416. Hendrickx, H., Boon, W., Bries, J., Kempeneers, L., Vandendriessche, H., Deckers, S. and Geuypens, M. (1992) De chemische bodemvruchtbaarheid van het Vlaamse akkerbouw- en weilandareaal. Bodemkundige dienst van België, Heverlee. Hermy, M., Honnay, O., Firbank, L., Grashof-Bokdam, C. J. and Lawesson, J. E. (1999) An ecological comparison between ancient and other forest plant species of Europe, and the implications for forest conservation. Biological Conservation 91: 9-22. Heschel, M. S. and Paige, K. N. (1995) Inbreeding depression, environmental stress, and population size variation in Scarlet Gilia (Ipomopsis aggregata). Conservation Biology 9: 126-133. Hiebeler, D. (2000) Populations on fragmented landscapes with spatially structured heterogeneities : landscape generation and local dispersal. Ecology 81: 1629-1641. Higgins, S. I. and Richardson, D. M. (1999) Predicting plant migration rates in a changing world: the role of long-distance dispersal. American Naturalist 153: 464-475. Higgins, S. I., Nathan, R. and Cain, M. L. (2003) Are long-distance dispersal events in plants usually caused by non-standard means of dispersal? Ecology 84: 1945-1956.
149
150
References
Hill, M. F. and Caswell, H. (1999) Habitat fragmentation and extinction thresholds on fractal landscapes. Ecology Letters 2: 121-127. Hill, M. O. (1979) TWINSPAN - a FORTRAN Program for arranging Multivariate Data in an Ordered Two Way Table by Classification of the Individuals and Attributes. Cornell University, Departement of Ecology and Systematics, Ithaca, New-York. 50p. Hobson, K. A., Bayne, E. M. and Van Wilgenburg, S. L. (2002) Large-scale conversion of forest to agriculture in the boreal plains of Saskatchewan. Conservation Biology 16: 1530-1541. Hodgson, J. G., Grime, J. P., Hunt, R. and Thompson, K. (1995) The electronic comparative ecology. Chapman and Hall. Holsinger, K. (2000) Demography and extinction in small populations. In: Young A. G. and Clarke G. M. (eds.) Genetics, demography and viability of fragmented populations. Cambridge University Press, Cambridge. pp. 55-74. Holt, R. D. (1997) From metapopulation dynamics to community structure. Some consequences of spatial heterogeneity. In: Hanski and Gilpin (eds.) Metapopulation Biology: Ecology, Genetics, and Evolution, pp. 149-164. Academic Press, San Diego. Honnay, O., Hermy, M. and Coppin, P. (1999a) Effects of area, age and diversity of forest patches in Belgium on plant species richness, and implications for conservation and reforestation. Biological Conservation 87: 73-84. Honnay, O., Hermy, M. and Coppin, P. (1999b) Nested plant communities in deciduous forest fragments: species relaxation or nested habitats? Oikos 84: 119-129. Honnay, O., Hermy, M. and Coppin, P. (1999c) Habitat quality factors affecting the colonization of forest plant species. Forest Ecology and Management 115: 157-170. Honnay, O. Verhaeghe, W. and Hermy, M. (2001) Plant community assembly along dendritic networks of small forest streams. Ecology 82: 1691-1702. Honnay, O., Verheyen, K., Butaye, J., Jacquemyn, H. Bossuyt, B. and Hermy, M. (2002a) Possible effects of habitat fragmentation and climate change on the range of forest plant species. Ecology Letters 5: 525-530. Honnay, O., Verheyen, K. and Hermy, M. (2002b) Permeability of ancient forest edges for weedy plant species invasion. Forest Ecology and Management 161: 109-122. Honnay, O., Bossuyt, B., Verheyen, K., Butaye, J., Jacquemyn, H. and Hermy, M. (2002c) Ecological perspectives for the restoration of plant communities in European temperate forests. Biodiversity and Conservation 11: 213-242. Hubbell, S. P. (1997) A unified theory of biogeography and relative species abundance and its application to rain forests and coral reefs. Coral Reefs 16 (suppl), S9-S21.
References
151
Hubbell, S. P., Foster, R. B., O’Brien, S. T., Harms, K. E., Condit, R., Wechsler, B., Wright, S. J. and Loo de Lao, S. (1999) Light-gap disturbances, recruitment limitation, and tree diversity in a neotropical forest. Science 283: 554-557. Husband, B. C. and Barrett, S. C. H. (1996) A metapopulation perspective in plant population biology. Journal of Ecology 84: 461-469. Husband, B. C. and Barrett, S. C. H. (1998) Spatial and temporal variation in population size of Eichhornia paniculata in ephemeral habitats: implications for metapopulation dynamics. Journal of Ecology 86: 1021-1031. Huxel, G. R. and Hastings, A. (1999) Habitat loss, fragmentation and restoration. Restoration Ecology 7: 309-315. Inghe, O. and Tamm, C. O. (1985) Survival and flowering of perennial herbs IV. The behaviour of Hepatica nobilis and Sanicula europaea on permanent plots during 1943 – 1981. Oikos 45: 400-420. Ingvarsson, P. K., Olsson, K. and Ericson, L. (1997) Extinction-recolonization dynamics in the mycophagous beetle Phalacrus substriatus. Evolution 51: 187-195. Jacquemyn, H., Butaye, J. and Hermy, M. (2000) The colonization of recent forest patches in a fragmented landscape: the role of geographical isolation and implications for forest expansion (in Dutch). Landschap 17: 165-176. Jacquemyn, H., Butaye, J. and Hermy, M. (2001a). Forest plant species richness in small, fragmented mixed deciduous forest patches: the role of area, time and dispersal limitation. Journal of Biogeography 28: 801-812. Jacquemyn, H., Butaye, J., Dumortier, M., Hermy, M. and Lust, N. (2001b) Effects of age and distance on plant species composition of mixed deciduous forest patches in a fragmented, agricultural landscape. Journal of Vegetation Science 12: 635-642. Jacquemyn, H., Brys, R. and Hermy, M. (2001c) Within and between plant variation in seed number, seed mass and germinability of Primula elatior: effect of population size. Plant Biology 3: 561-568. Jacquemyn, H., Brys, R. and Hermy, M. (2002) Patch occupancy, population size and reproductive success of a forest herb (Primula elatior) in a fragmented landscape. Oecologia 130: 617-625. Jacquemyn, H., Van Rossum, F., Brys, R., Endels, P., Hermy, M., Triest, L. and De Blust, G. (2003) Effects of agricultural land use and fragmentation on genetics, demography and population persistence of the rare Primula vulgaris, and implications for conservation. Belgian Journal of Botany 136: 5-22. Janson, S. and Vegelius, J. (1981) Measures of ecological association. Oecologia 49: 371-376.
151
152
References
Jennersten, O. (1988) Pollination in Dianthus deltoides (Caryophyllaceae): effects of habitat fragmentation on visitation and seed set. Conservation Biology 2: 359-366. Johansson, P. and Ehrlén, J. (2003) Influence of habitat quantity, quality and isolation on the distribution and abundance of two epiphytic lichens. Journal of Ecology 91: 213-221. Johnson, M. P. (2000) The influence of patch demographics on metapopulations, with particular reference to successional landscapes. Oikos 88: 67-74. Johst, K., Brandl, R. and Eber, S. (2002) Metapopulation persistence in dynamic landscapes: the role of dispersal distance. Oikos 98: 263-270. Jules, E. S. (1998) Habitat fragmentation and demographic change for a common plant: Trillium in old-growth forest. Ecology 79: 1645-1656. Jules, E. S. and Rathcke, B. J. (1999) Mechanisms of reduced Trillium recruitment along edges of old-growth forest fragments. Conservation Biology 13: 784-793. Kadmon, R. and Pulliam, H. R. (1993) Island biogeography: Effect of geographical isolation on species composition. Ecology 74: 977-981. Kahmen, S. and Poschlod, P. (2000) Population size, plant performance , and genetic variation in the rare plant Arnica montana L. in the Rhön, Germany. Basic and Applied Ecology 1: 43-51. Keller, L. F. and Waller, D. M. (2002) Inbreeding effects in wild populations. Trends in Ecology and Evolution 17: 230-241. Kéry, M., Matthies, D. and Spillmann, H. (2000) Reduced fecundity and offspring performance in small populations of the declining grassland plants Primula veris and Gentiana lutea. Journal of Ecology 88: 17-30. Keymer, J. E., Marquet, P. A., Velasco-Hernández, J. X. and Levin, S. A. (2000) Extinction thresholds and metapopulation persistence in dynamic landscapes. American Naturalist 156: 478-494. Kirby, K. J., Bines, T., Burn, A., Mackintosh, J., Pitkin, P. and Smith, I. (1986) Seasonal and observer differences in vascular plant records from British woodland. Journal of Ecology 74: 123-131. Kleyer, M. (1995) Biological traits of vascular plants. A database. Arbeitsberichte Institut für Landschafstplanung und Ökologie, Universität Stuttgart. Koerner, W., Dupouey, J. L., Dambrine, E. and Benoit, M. (1997) Influence of past land use on the vegetation and soils of present day forest in the Vosges mountains, France. Journal of Ecology 85: 351-358. Kratter, A. W. (1992) Montane avian biogeography in southern California and Baja California. Journal of Biogeography 19: 269-283.
References
153
Kwak, M. M., Velterop, O. and van Andel, J. (1998) Pollen and gene flow in fragmented habitats. Applied Vegetation Science 1: 37-54. Lambinon, J., De Langhe, J. E., Delvosalle, L. and Duvigneaud, J. (1998) Flora van België, Groothertogdom Luxemburg, Noord-Frankrijk en aangrenzende gebieden. Patrimonium van de Nationale Plantentuin van België. Meise. Lawesson, J.E., de Blust, G., Grashof, C., Firbank, L., Honnay, O., Hermy, M., Hobitz, P. and Jensen, L.M. (1998) Species diversity and area-relationships in Danish beech forests. Forest Ecology and Management 106: 235-245. Le Corre, V. and Kremer, A. (1998) Cumulative effects of founding events during colonisation on genetic diversity and differentiation in an island and stepping-stone model. Journal of Evolutionary Biology 11: 495-512. Lee, T. D. (1988) Patterns of fruit and seed production. In Lovett Doust J, Lovett Doust L (eds) Plant Reproductive Ecology: Patterns and Strategies. Oxford University Press, New York, pp 179-201. Lefort, F. and Douglas, G. C. (1999) An efficient micro-method of DNA isolation from mature leaves of four hardwood tree species Acer, Fraxinus, Prunus and Quercus. Annals of Science 56: 259-263. Legendre, P. and Vaudor, A. (1991) The R Package: Multidimensional analysis, spatial analysis. Département de sciences biologiques, Université de Montréal. Lennartsson, T. (2002) Extinction thresholds and disrupted plant-pollinator interactions in fragmented plant populations. Ecology 83: 3060-3072 Les, D. H., Reinartz, J. A. and Esselman, E. J. (1991) Genetic consequences of rarity in Aster furcatus (Asteraceae), a threatened, self-incompatible plant. Evolution 45: 1641-1650. Lockwood, J. L. and Pimm, S. L. (1999) When does restoration succeed? In: Weiger, E. and Keddy, P. A. (eds.). Ecological assembly rules: perspectives, advances, retreats. Cambridge University Press, Cambridge. pp. 363-392. Loreau, M. (2000) Are communities saturated? On the relationship between α, β and γ diversity. Ecology Letters 3: 73-76. Lynch, M. and Milligan, B. G. (1994) Analysis of population genetic structure with RAPD markers. Molecular Ecology 3: 91-99. MacArthur, R. H. and Wilson, E. O. (1963) An equilibrium theory of insular zoogeography. Evolution 17: 373-387. MacArthur, R. H. and Wilson, E. O. (1967) The theory of island biogeography. Princeton University Press, Princeton, New-Jersey, USA.
153
154
References
Manel, S., Schwartz, M. K., Luikart, G. and Taberlet, P. (2003) Landscape genetics: combining landscape ecology and population genetics. Trends in Ecology and Evolution 18: 189-197. Mantel, N. (1967) The detection of disease clustering and a generalized regression approach. Cancer Research 27: 209-220. Manly, B. F. J. (1997) Randomization, Bootstrap and Monte Carlo methods in biology. 2nd edition. Chapman and Hall. Matlack, G. R. (1994) Plant species migration in a mixed-history forest landscape in eastern North America. Ecology 75: 1491-1502. MathSoft (1998) S-PLUS 4 Guide to Statistics. Data Analysis Products Division. MathSoft, Seattle, Washington. McCauley, D. E. (1991) Genetic consequences of local population extinction and recolonization. Trends in Ecology and Evolution 6: 5-8. McCauley, D. E. (1994) Contrasting the distribution of chloroplast DNA and allozyme polymorphism among local populations of Silene alba: implications for studies of gene flow in plants. Proceedings of the National Academy of Sciences USA 91: 8127-8131. McCauley, D. E., Raveill, J. and Antonovics, J. (1995) Local founding events as determinants of genetic structure in a plant metapopulation. Heredity 75: 630-636. McCauley, D. E., Richards, C. M., Emery, S. N., Smith, R. A. and McGlothin, J. W. (2001) The interaction of genetic and demographic processes in plant metapopulations: a case study of Silene alba. In: Silvertown, J. and Antonovics, J. (eds.) Integrating Ecology and Evolution in a Spatial Context. Blackwell Science. pp. 177-196. McCullagh, P. and Nelder, J. A. (1989) Generalized Linear Models. Chapman and Hall, London. McElroy, T. C., Kandl, K. L., Garcia, J. and Trexler, J. C. (2003) Extinction-colonization dynamics structure genetic variation of spotted sunfish (Lepomis punctatus) in the Florida Everglades. Molecular Ecology 12: 355-368. van der Meijden, E., Klinkhamer, P. G. L., de Jong, T. J. and van Wijk, C. A. M. (1992) Metapopulation dynamics of biennial plants: how to exploit temporary habitats. Acta Botanica Neerlandica 41: 249-270. Menges, E. S. (1990) Popuation viability analysis for an endangered plant. Conservation Biology 4: 52-62. Menges, E. S. (1991) The application of minimum viable population theory to plants. In Falk DA, Holsinger KE (eds) Genetics and Conservation of Rare Plants. Oxford University Press, New York, pp 47-61.
References
155
Menges, E. S. and Dolan, R. W. (1998) Demographic viability of populations of Silene regia in midwestern prairies : relationships with fire management, genetic variation, geographic location, population size and isolation. Journal of Ecology 86: 63-78. Messing, I., Alriksson, A. and Johansson, W. (1997) Soil physical properties of afforested and arable land. Soil Use and Management 13: 209-217. Morgan, J. W. (1999) Effects of population size on seed production and germinability in an endangered, fragmented grassland plant. Conservation Biology 13: 266-273. Morton, R. D. and Law, R. (1997) Regional species pools and the assembly of ecological communities. Journal of Theoretical Biology 187: 321-331. Motzkin, G., Wilson, P., Foster, David R. and Allen, A. 1999. Vegetation patterns in heterogeneous landscapes : The importance of history and environment . Journal of Vegetation Science 10: 903-920. Mueller, U. G. and Wolfenbarger, L. L. (1999) AFLP genotyping and fingerprinting. Trends in Ecology and Evolution 14: 389-394. Nathan, R., Perry, G., Crnin, J. T., Strand, A. E. and Cain, M. L. (2003) Methods for estimating long-distance dispersal. Oikos 103: 261-273. Nekola, J. C. (1999) Paleorefugia and neorefugia: the influence of colonization history on community pattern and process. Ecology 80: 2459-2473. Nekola, J. C. and White, P. S. (1999) The distance decay of similarity in biogeography and ecology. Journal of Biogeography 26: 867-878. Nekola, J. C. and White, P. S. (2002) Conservation, the two pillars of ecological explanation, and the paradigm of distance. Natural Areas Journal 22: 305-310. Ney-Nifle, M. and Mangel, M. (2000) Habitat loss and changes in the species-area relationship. Conservation Biology 14: 893-898. Newman, D and Pilson, D. (1997) Increased probability of extinction due to decreased genetic effective population size: Experimental populations of Clarkia pulchella. Evolution 51: 354-362. Newman, D. and Tallmon, D. A. (2001) Experimental evidence for beneficial fitness effects of gene flow in recently isolated populations. Conservation Biology 15: 1054-1063. Nicholson, S. A. and Monk, C. D. (1974) Plant species diversity in old-field succession on the Georgia Piedmont. Ecology 55: 1075-1085. Noss, R. F. (1997) Hierarchical indicators for monitoring changes in biodiversity. In: Meffe, G. K. and Carroll, C. R. (eds.) Principles of Conservation Biology. Sinauer Associates, Sunderland, Massachusetts. pp. 88-89.
155
156
References
Nybom, H. and Bartish, I. V. (2000) Effects of life history traits and sampling strategies on genetic diversity estimates obtained with RAPD markers. Perspectives in Plant Ecology, Evolution and Systematics 3: 93-114. Oostermeijer, J. G. B., van Eijck, M. W. and den Nijs, J. C. M. (1994) Offspring fitness in relation to population size and genetic variation in the rare perennial plant species Gentiana pneumonanthe (Gentianaceae). Oecologia 86:359-367. Oostermeijer, J. G. B., Luijten, S. H., Krenová, Z. V. and den Nijs, H. C. M. (1998) Relationships between population and habitat characteristics of the rare Gentiana pneumonanthe L. Conservation Biology 12: 1042-1053. Oostermeijer, J. G. B., Brugman, M. L., de Boer, E. R. and den Nijs, J. C. M. (1996) Temporal and spatial variation in the demography of Gentiana pneumonanthe, a rare perennial herb. Journal of Ecology 84: 153-166. Ouborg, N. J. (1993) Isolation, population size and extinction: the classical and metapopulation approaches applied to vascular plants along the Dutch Rhine-system. Oikos 66: 298-303. Ouborg, N. J. and van Treuren, R. (1995) Variation in fitness-related characters among small and large populations of Salvia pratensis. Journal of Ecology 83: 369-380. Ouborg, N. J., Piquot, Y. and van Groenendael, J. M. (1999) Population genetics, molecular markers and the study of dispersal in plants. Journal of Ecology 87: 551-568. Ovaskainen, O. and Hanski, I. (2001) Spatially structured metapopulation models: global and local assessment of metapopulation capacity. Theoretical Population Biology 60: 281302. Pannell, J. R. and Charlesworth, B. (2000) Effects of metapopulation processes on measures of genetic diversity. Philosophical Transactions of the Royal Society London B 355: 1851-1864. Pannell, J. R. and Obbard, D. J. (2003) Probing the primacy of the patch : what makes a metapopulation. Journal of Ecology 91 : 485-488. Peakall, R. and Smouse, P. E. (2001) GenAlEx V5: Genetic Analysis in Excel. Population genetic software for teaching and research. Australian National University, Canberra, Australia. http://www.anu.edu.au/BoZo/GenAlEx/ Peterken, G. F. (1974) A method of assessing woodland flora for conservation using indicator species. Biological Conservation 6: 239-245. Peterken, G. F. (1977) Habitat conservation priorities in British and European woodlands. Biological Conservation 11: 223-236.
References
157
Peterken, G. F. and Game, M. (1984) Historical factors affecting the number and distribution of vascular plant species in the woodlands of central Lincolnshire. Journal of Ecology 72: 155-182. Peterken, G. F. (2000) Rebuilding networks of forest habitats in lowland England. Landscape Research 25: 291-303. Peterson, M. A. and Denno, R. F. (1998) The influence of dispersal and diet breadth on patterns of genetic isolation by distance in phytophagous insects. American Naturalist 152: 428-446. Pickett, S. T. A., Parker, V. T. and Fiedler, P. L. (1992) The new paradigm in ecology : Implications for conservation biology above the species level. In: Fiedler, P. L. and Jain, S. K. (eds.) Conservation Biology : The Theory and Practice of Nature Conservation Preservation and Management. Chapman and Hall, New York. pp. 65-88 Pigott, D. C. (1977) The scientific basis of practical conservation: aims and methods for conservation. Proceedings of the Royal Society London B 197: 59-68. Pimm, S. L., Jones, H. L. and Diamond, J. (1988) On the risk of extinction. American Naturalist 132: 757-785. Pollard, E., Hooper, M. D. and Moore, N. W. 1974. Hedges. Collins, London. Primack, R. B. and Miao, S. L. (1992) Dispersal can limit local plant distribution. Conservation Biology 6: 513-519. Preston, F. W. (1960) Time and space and the variation of species. Ecology 41: 611-627. Quintana-Ascencio, P. F. and Menges, E. S. (1996) Inferring metapopulation dynamics from patch-level incidence of florida scrub plants. Conservation Biology 10: 1210-1219. Reinartz, J. A. and Les, D. H. (1994) Bottleneck-induced dissolution of self-incompatibility and inbreeding consequences in Aster furcatus (Asteraceae). American Journal of Botany 81: 446-455. Rich, T. C. G. and Woodruff, E. R. (1992) Recording bias in botanical surveys. Watsonia 19: 73-95. Richards, A. J. (1997) Plant Breeding Systems. Allen & Unwyn, London. Richards, C. M., Church, S. and McCauley, D. E. (1999) The influence of population size and isolation on gene flow by pollen in Silene alba. Evolution 53: 63-73. Richards, C. M. (2000) Inbreeding depression and genetic rescue in a plant metapopulation. American Naturalist 155: 383-394. Richter, D. D., Markewitz, D., Wells, C. G., April, R., Heine, P. R. and Urrego, B. (1994) Soil chemical change during three decades in an old-field loblolly pine (Pinus taeda L.) ecosystem. Ecology 75: 1463 – 1473.
157
158
References
Ricklefs, R. E. (1987) Community diversity: relative roles of local and regional processes. Science 235: 167-171. Ricklefs, R. E. and Schluter, D. (1993) Species diversity in ecological communities. Historical and geographical perspectives. The University of Chicago Press, Chicago. Ricklefs, R. E. and Lovette, I. J. (1999) The roles of island area per se and habitat diversity in the species-area relationships of four Lesser Antillean faunal groups. Journal of Animal Ecology 68: 1142-1160. Rohlf, F. J. (2000) NTSys: numerical taxonomy and multivariate analysis system, Version 2.1. State University of New York, New York. Roldán-Ruiz, I., Dendauw, J., Van Bockstaele, E., Depicker, A. and De Loose, M. (2000) AFLP markers reveal high polymorphic rates in ryegrasses (Lolium spp.). Molecular Breeding 6: 125-134. Rosenzweig, M. L. (1995). Species Diversity in Space and Time. Cambridge University Press, Cambridge. Rudd, H., Vala, J. and Schaefer, V. (2002) The importance of backyard habitat in a comprehensive biodiversity conservation strategy: a connectivity analysis of urban green spaces. Restoration Ecology 10: 368-375. Salisbury, E. J. (1942) Reproductive capacity of plants. Bell & Sons, London. Saunders, D. A., Hobbs, R. J. and Margules, C. R. (1991) Biological consequences of ecosystem fragmentation – a review. Conservation Biology 5: 18-32. Scheiner, S. M. and Rey-Benayas, J. M. (1997) Placing empirical limits on metapopulation models for terrestrial plants. Evolutionary Ecology 11: 275-288. Schemske, D. W., Husband, B. C., Ruckelshaus, M. H., Godwillie, C., Parker, I. M. and Bishop J. G. (1994) Evaluating approaches to the conservation of rare and endangered plants. Ecology 75: 584-606. Schmidt, K. and Jensen, K. (2000) Genetic structure and AFLP variation of remnant populations in the rare plant Pedicularis palustris (Scrophulariaceae) and its relation to population size and reproductive components. American Journal of Botany 87: 678689. Schou O (1983) The distyly in Primula elatior (L.) Hill (Primulaceae), with a study of flowering phenology and pollen flow. Botanical Journal of the Linnean Society 86: 1679-1690. Shaffer, M. L. (1981) Minimal population sizes for species conservation. Bioscience 31: 131134. Sih, A. and Baltus, M. S. (1987) Patch size, pollinator behavior, and pollinator limitation in catnip. Ecology 68: 1679-1690.
References
159
Silvertown, J. (1991) Dorothy’s dilemma and the unification of plant population biology. Trends in Ecology and Evolution 6: 346-348. Silvertown, J., Franco, M. and Menges, E. S. (1996) Interpretation of elasticity matrices as an aid to the management of plant populations for conservation. Conservation Biology 10: 591-597. Simberloff, D. and Gotelli, N. (1984) Effects of insularisation on plant species richness in the prairie forest ecotone. Biological Conservation 29: 27-46. Singleton, R., Gardescu, S., Marks, P. L. and Geber, M. A. (2001) Forest herb colonization of postagricultural forests in central New York State, USA. Journal of Ecology 89: 325338. Slatkin, M. (1977) Gene flow and genetic drift in a species subject to frequent local extinctions. Theoretical Population Biology 12: 253-262. Slatkin, M. (1993) Isolation by distance in equilibrium and non-equilibrium populations. Evolution 47: 264-279. Smouse, P. E., Long, J. C. and Sokal, R. R. (1986) Multiple regression and correlation extensions of the Mantel test of matrix correspondence. Systematic Zoology 35: 627632. Sokal, R. R. and Rohlf, F. J. (1995) Biometry. The Principles and Practice of Statistics in Biological Research. W. H. Freeman, New York. Sork, V. L., Nason, J., Campbell, D. R. and Fernandez, J. F. (1999) Landscape approaches to historical and contemporary gene flow in plants. Trends in Ecology and Evolution 14: 219-224. Steffan-Dewenter, I. and Tscharntke, T. (1999) Effects of habitat isolation on pollinator communities and seed set. Oecologia 121: 432-440. Steinberg, E. K. and Jordan, C. E. (1997) Using molecular genetics to learn about the ecology of threatened species: The allure and illusion of measuring genetic structure in natural populations. In: Fiedler, P. L. and Kareiva P. M. (eds.). Chapmann and Hall, New York. pp. 440-460. Sun, M. (1997) Population genetic structure of yellow starthistle (Centaurea solstitialis), a colonizing weed in the Western United States. Canadian Journal of Botany 75: 14701478. Tack, G., Van den Bremt, P. and Hermy, M. (1993) Bossen van Vlaanderen, een historische ecologie. Davidsfonds, Leuven. Tero, N., Aspi, J., Siikamaki, P., Jakalaniemi, A. and Tuomi, J. (2003) Genetic structure and gene flow in a metapopulation of an endangered plant species, Silene tatarica. Molecular Ecology 12: 2073-2085.
159
160
References
Thomas, C. D. and Kunin, W. E. (1999) The spatial structure of populations. Journal of Animal Ecology 68: 647-657. Thomas, R. C., Kirby, K. J. and Reid, C. M. (1997) The conservation of a fragmented ecosystem within a cultural landscape: the case of ancient woodland in England. Biological Conservation 82: 243-252. Thrall, P.H., Richards, C. M., McCauley, D. E. and Antonovics, J. (1998) Metapopulation collapse: consequences of limited gene flow in spatially structured populations. In Bascompte, J. and Solé, R. V. (eds.). Modeling spatiotemporal dynamics in ecology. Academic Press, New York, USA. pp. 83-104. Tilman, D. (1982) Resource competition and community structure. Princeton University Press, Princeton, New-Jersey. Tilman, D. (1997) Community invasibility, recruitment limitation, and grassland biodiversity. Ecology 78: 81-92. Tilman, D., May, R. M., Lehman, C. L. and Nowak, M. A. (1994) Habitat destruction and the extinction debt. Nature 371: 65-66. Tilman, D., Lehman, C. L. and Kareiva, P. (1997) Population dynamics in spatial habitats. In: Tilman, D. and Kareiva, P. (eds.) Spatial Ecology. Princeton University Press, pp. 3-20. Tobler, W. R. (1970) A computer movie simulating urban growth in the Detroit region. Economic Geography 46: 234-240. Travis, S. E., Maschinski, J. and Keim, P. (1996) An analysis of genetic variation in Astragalus cremnophylax var. cremnophylax, a critically endangered plant, using AFLP markers. Molecular Ecology 5: 735-745. Turnbull, L. A., Crawley, M. J. and Rees, M. (2000) Are plant populations seed-limited? A review of seed sowing experiments. Oikos 88: 225-238. Valentine, D. H. (1948) Studies in British Primulas – II. Ecology and taxonomy of primrose and oxlip (Primula vulgaris Huds. and P elatior Schreb.). New Phytologist 47: 111130. Valverde, T. and Silvertown, J. (1995) Spatial variation in the seed ecology of a woodland herb (Primula vulgaris) in relation to light environment. Functional Ecology 9: 942950. Valverde, T. and Silverton, J. (1997) A metapopulation model for Primula vulgaris, a temperate understory herb. Journal of Ecology 85: 193-210. Van Rossum, F., Echchgadda, G., Szabadi, I. and Triest, L. (2002) Commonness and longterm survival in fragmented habitats: Primula elatior as a case study. Conservation Biology 16: 1286-1295.
References
161
Van Treuren, R., Bijlsma, R., Ouborg, N. J. and van Delden, W. (1993) The significance of genetic erosion in the process of extinction. 4. Inbreeding depression and heterosis effects caused by selfing and outcrossing in Scabiosa columbaria. Evolution 47: 16691680. Vekemans, X., Beauwens, T., Lemaire, M. and Roldán-Ruiz, I. (2002) Data from amplified fragment length polymorphism (AFLP) markers show indication of size homoplasy and of a relationship etween degree of homoplasy and fragment size. Molecular Ecology 11: 139-151. Vellend, M. (2003) Habitat loss inhibits recovery of plant diversity as forests regrow. Ecology 84: 1158-1164. Vellend, M., Myers, J. A., Gardescu, S., Marks, P. L. (2003) Dispersal of Trillium seeds by deer: implications for long-distance migration of forest herbs. Ecology 84: 1067-1072. Vergeer, P., Rengelink, R., Copal, A. and Ouborg, N. J. (2003) The interacting effects of genetic variation, habitat quality and population size on performance of Succisa pratensis. Journal of Ecology 91: 18-26. Verheyen, K., Bossuyt, B., Hermy, M. and Tack, G. (1999) The land use history (1278-1990) of a mixed hardwood forest in western Belgium and its relationship with chemical soil characteristics. Journal of Biogeography 26: 1115-1128. Verheyen, K and Hermy, M. (2001) The relative importance of dispersal limitation of vascular plants in secondary forest succession in Muizen Forest, Belgium. Journal of Ecology 89: 829-840. Verheyen, K., Guntenspergen, G. R., Biesbrouck, B. and Hermy, M. (2003a) An integrated analysis of the effects of past land use on forest herb colonization at the landscape scale. Journal of Ecology 91: 731-742. Verheyen, K., Honnay, O., Motzkin, G., Hermy, M. and Foster, D. R. (2003b) Response of forest plant species to land-use change: a life-history trait-based approach. Journal of Ecology 91: 563-577. Verheyen, K., Bossuyt, B., Honnay, O. and Hermy, M. (2003c) Herbaceous plant community structure of ancient and recent forests in two contrasting forest types. Basic and Applied Ecology 4: 537-546. Verheyen, K. and Hermy, M. (2004) Recruitment and growth of herb-layer species with different colonizing capacities in ancient and recent forests. Journal of Vegetation Science 15: 125-134. Villard, M. A., Trzcinski, M. K. and Merriam, G. (1999) Fragmentation effects on forest birds: relative influence of woodland cover and configuration on landscape occupancy. Conservation Biology 13: 774-783.
161
162
References
Vos, P., Hogers, R., Bleeker, M. et al. (1995) AFLP: a new technique for DNA fingerprinting. Nucleic Acids Research 23: 4407-4414. Wade, M. J. and McCauley, D. E. (1988) Extinction and recolonization: their effects on the genetic differentiation of local populations. Evolution 42: 995-1005. Wang, B. C. and Smith, T. B. (2002) Closing the dispersal loop. Trends in Ecology and Evolution 17: 379-385. Whale, D. M. (1983) The response of Primula species to soil waterlogging and soil drought. Oecologia 58:272-277. White, P. S. and Nekola, J. C. (1992) Biological diversity in an ecological context. In: Barker, J. R. and Tingey, D. T. (eds.) Air Pollution and Biodiversity. Van Nostrand Reinhold, New York. pp. 10-29. Whitlock, M. C. (1992) Nonequilibrium population structure in forked fungus beetles: extinction, colonization, and the genetic variance among populations. American Naturalist 139: 952-970. Whitlock, M. C. and McCauley, D. E. (1990) Some population genetic consequences of colony formation and extinction: genetic correlations within breeding groups. Evolution 44: 1717-1724. Widén, B. (1993) Demographic and genetic effects on reproduction as related to population size in a rare, perennial herb, Senecio integrifolius. Biological Journal of the Linnean Society 50: 179-195. Wiens, D. (1984) Ovule survivorship, brood size, life history, breeding system, and reproductive success in plants. Oecologia 64: 47-53. Wilcove, D. S., McLellan, C. H. and Dobson, A. P. (1986) Habitat fragmentation in the temperate zone. In: Soulé, M. S. (ed.) Conservation Biology. Sinauer Associates, p. 237-256. Wilcox, B. A. and Murphy, D. D. (1985) Conservation strategy: the effects of fragmentation on extinction. American Naturalist 125: 879-887. Williamson, M. (1988) Relationship of species number to area, distance and other variables. In: Myers, A. A. and Giller, P. S. (eds.). Analytical Biogeography: An integrated Approach to the Study of Animal and Plant Distributions. Chapman & Hall, London. pp. 91-114. With, K. A. and King, A. W. (1999) Critical thresholds for species in fractal landscapes. Conservation Biology 13: 314-326. Woodell, S. R. J. (1960) What pollinates Primulas? The New Scientist 8: 568-571.
References
163
Young, A. (1995) Landscape structure and genetic variation in plants: empirical evidence. In: Hansson, L., Fahrig, L. and Merriam, G. (eds.) Mosaic Landscapes and Ecological Processes. Chapmann and Hall, New York, pp. 153-177. Young, A., Boyle, T. and Brown, T. (1996) The population genetic consequences of habitat fragmentation for plants. Trends in Ecology and Evolution 11: 413-418. Young, T. P. (2000) Restoration ecology and conservation biology. Biological Conservation 92: 73-83. Young, T. P., Chase, J. M. and Huddleston, R. D. (2001) Community succession and assembly: comparing, contrasting, and combining paradigms in the context of ecological restoration. Ecological Restoration 19: 5-18. Zacharias, D. and Brandes, D. (1990) Species area-relationships and frequency - Floristical data analysis of 44 isolated woods in northwestern Germany. Vegetatio 88, 21-29. Zhivotovsky, L. A. (1999) Estimating population structure in diploids with multilocus dominant DNA markers. Molecular Ecology 8: 907-913.
163
Publications Peer reviewed articles 1.
Jacquemyn, H., Butaye, J. and Hermy, M. (2000) Kolonisatie van jonge bosfragmenten. De rol van ruimtelijke isolatie en implicaties voor bosuitbreiding. Landschap 17: 165-176.
2.
Butaye, J., Jacquemyn, H. and Hermy, M. (2001) Differential colonization causing non-random forest plant community structure in a fragmented agricultural landscape. Ecography 24: 369-380.
3.
Jacquemyn, H., Butaye, J., Dumortier, M., Hermy, M. and Lust, N. (2001) Effects of age and distance on plant species composition of mixed deciduous forest fragments in an agricultural landscape. Journal of Vegetation Science 12: 635-642.
4.
Jacquemyn, H., Butaye, J. and Hermy, M. (2001) Forest plant species richness in small, fragmented mixed deciduous forest patches: the role of area, time and dispersal limitation. Journal of Biogeography 28: 801-812.
5.
Jacquemyn, H., Brys, R. and Hermy, M. (2001) Within and between plant variation in seed number, seed mass and germinability of Primula elatior: effect of population size. Plant Biology 3: 561-568.
6.
Dumortier, M., Butaye, J., Jacquemyn, H., Van Camp, N., Hermy, M and Lust, N. (2002) Predicting vascular plant species richness of fragmented forests in agricultural landscapes in central Belgium. Forest Ecology and Management 158: 85-102.
7.
Honnay, O., Bossuyt, B., Verheyen, K., Butaye, J, Jacquemyn, H. and Hermy, M. (2002) Ecological perspectives for the restoration of plant communities in European temperate forests. Biodiversity and Conservation 11: 213-242.
8.
Endels, P., Jacquemyn, H., Brys, R., Hermy, M. and De Blust, G. (2002) Temporal changes (19861999) in populations of primrose (Primula vulgaris Huds.) in an agricultural landscape and implications for conservation. Biological Conservation 105: 11-25.
9.
Butaye, J., Jacquemyn, H., Honnay, O. and Hermy, M. (2002) The species pool concept applied to forests in a fragmented landscape: dispersal limitation versus habitat limitation. Journal of Vegetation Science: 13: 27-34.
10. Jacquemyn, H., Brys, R. and Hermy, M. (2002) Patch occupancy, population size and reproductive success of a forest herb (Primula elatior) in a fragmented landscape. Oecologia 130: 617-625. 11. Jacquemyn, H., Brys, R. and Hermy, M. (2002) Flower and fruit production in small populations of Orchis purpurea Huds. and implications for management. In: Kindlmann, P., Willems, J. H. and Whigham, D. F. (eds). Trends and fluctuations and underlying mechanisms in terrestrial orchid populations. Backhuys, The Netherlands. pp. 67-84. 12. Honnay, O., Verheyen, K., Butaye, J., Jacquemyn, H., Bossuyt, B. and Hermy, M. (2002) Possible effects of habitat fragmentation and climate change on the range of forest plant species. Ecology Letters 5: 525-530. 13. Endels, P., Jacquemyn, H., Brys, R. and Hermy, M. (2002) Changes in pin-thrum ratios in populations of the heterostyle Primula vulgaris Huds.: Does imbalance affect population persistence? Flora 197: 326-331. 14. Jacquemyn, H., Brys, R. and Hermy, M. (2003) Short-term effects of different management regimes on the response of calcareous grassland vegetation to increased nitrogen. Biological Conservation 111: 137-147.
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Publications
165
15. Jacquemyn, H., Van Rossum, F., Brys, R., Endels, P., Hermy, M., Triest, L. and De Blust, G. (2003) Genetic, demographic and ecological effects on population persistence of Primula vulgaris and implications for conservation. Belgian Journal of Botany 136: 5-22. 16. Jacquemyn, H., Butaye, J. and Hermy, M. (2003) Impacts of patch density and distance from natural forests on colonization success. Restoration Ecology 11: 417-423. 17. Brys, R., Jacquemyn, H., Endels, P., Hermy, M. and De Blust, G. (2003) The relation between reproductive success and demographic structure in remnant populations of Primula veris. Acta Oecologica 24: 247-253. 18. Jacquemyn, H., Butaye, J. and Hermy, M. (2003) Patch occupancy patterns of forest plant species in a fragmented landscape. Ecography 26: 768-776. 19. Jacquemyn, H., Honnay, O., Galbusera, P. and Roldán-Ruiz, I. (2004) Genetic structure of the forest herb Primula elatior in a changing landscape. Molecular Ecology 13: 211-219. 20. Brys, R., Jacquemyn, H., Endels, P., Van Rossum, F., Hermy, M., Triest, L. De Bruyn, L., and De Blust, G. (2004) Reduced reproductive success and mate availability in small populations of the self-incompatible Primula vulgaris. Journal of Ecology 92: 5-14. 21. Endels, P., Jacquemyn, H., Brys, R. and Hermy, M. Response of Primula veris populations to ecological restoration: linking fitness-related characteristics with demography. Plant Ecology: in press. 22. Endels, P., Jacquemyn, H., Brys, R. and Hermy, M. Performance of primrose (Primula vulgaris) in an agricultural landscape: impact of management regime and habitat chracteristics on demographic traits on multiple scales. Applied Vegetation Science: in press
Submitted articles 23. Brys, R., Jacquemyn, H., Endels, P., De Blust, G. and Hermy, M. The effect of grassland management on plant traits and demographic variation in the perennial herb Primula veris. Submitted to Journal of Applied Ecology 24. Honnay, O., Jacquemyn, H., Roldán-Ruiz, I. and Hermy, M. Effects of remant population dynamics on the genetic structure of the clonal plant Maianthemum bifolium in ancient relic forest fragments. Submitted to Oecologia 25. Jacquemyn, H., Brys, R., Hermy, M. and Willems, J. H. Are non-rewarding orchids more extinction prone than rewarding species: an assessment using historical records from Belgium and the Netherlands. Submitted to Biological Conservation. 26. Brys, R., Jacquemyn, H., Endels, P., De Blust, G. and Hermy, M. Habitat deterioration affects population dynamics and increases extinction risks in a previously common perennial. Submitted to Conservation Biology.
Others 27. Jacquemyn, H., Brys, R., Hermy, M., De Blust, G. and Zeevaert, A. (2001) De effecten van beheer en bemesting op de soortendiversiteit van kalkgraslanden van het type Galio-Trifolietum. Natuurhistorisch Maandblad 90: 151-157. 28. Brys, R., Jacquemyn, H., Van Rossum, F., De Blust, G., Hermy, M. and Triest, L. (2001) Kwetsbare plantenpopulaties in agrarisch gebied: verspreiding, verbreiding, en genetische diversiteit als basis voor functionele habitatnetwerken. Eindverslag van VLINA-project 98/03.
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29. Jacquemyn, H., Brys, R., Hermy, M. and Zeevaert, A. (2002) Vruchtzetting bij de Purperorchis (Orchis purpurea) (Orchidaceae) : Effecten van habitat, plantgrootte, populatiegrootte en densiteit. Natuurhistorisch Maandblad 91: 1-6. 30. Jacquemyn, H., Brys, R., Van Rossum, F., Endels, P., Hermy, M., Triest, L. and De Blust, G. (2002) Behoud van zeldzame plantensoorten in kleine landschapselementen: Sleutelbloemen als voorbeeld. Natuur.focus 1: 19-24. 31. Jacquemyn, H., Verheyen, K., Muys B. and Hermy, M. (2002) Wanneer een exoot invasief wordt: het geval van Spireae douglasii Hook. in het universiteitsbos in Hamont-Achel. Natuur.focus 1: 92-96. 32. Jacquemyn, H., Brys, R. and Hermy, M. (2003) Bestuiving bij orchideeën. Over bloemen en bijen, verleiding en bedrog. Natuur.focus 2: 109-114.
Appendix Allelic frequencies in all populations of P. elatior. Locus
Population Y1
Y2
Y3
Y4
Y5
Y6
Y7
O1
O2
O3
O4
O5
O6
O7
O8
1 0,263 0,530 0,353 0,318 0,276 0,272 0,499 0,261 0,277 0,148 0,302 0,481 0,421 0,231 0,363 2 0,263 0,338 0,437 0,278 0,276 0,111 0,390 0,337 0,134 0,090 0,198 0,321 0,265 0,263 0,316 3 0,033 0,225 0,121 0,072 0,045 0,018 0,144 0,041 0,102 0,090 0,106 0,116 0,137 0,062 0,231 4 0,085 0,261 0,279 0,169 0,276 0,111 0,257 0,226 0,134 0,062 0,198 0,321 0,198 0,088 0,316 5 0,085 0,261 0,180 0,203 0,192 0,161 0,109 0,160 0,134 0,035 0,106 0,192 0,137 0,201 0,272 6 0,059 0,066 0,065 0,072 0,080 0,111 0,180 0,098 0,041 0,062 0,302 0,192 0,421 0,062 0,052 7 0,900 0,904 0,872 0,876 0,888 0,872 0,732 0,742 0,868 0,883 0,827 0,870 0,517 0,670 0,686 8 0,230 0,338 0,279 0,203 0,372 0,214 0,390 0,128 0,238 0,148 0,198 0,192 0,265 0,088 0,052 9 0,764 0,904 0,661 0,737 0,630 0,571 0,638 0,584 0,643 0,484 0,302 0,624 0,517 0,497 0,468 10 0,199 0,530 0,437 0,407 0,485 0,571 0,390 0,298 0,317 0,314 0,302 0,481 0,517 0,297 0,363 11 0,140 0,298 0,065 0,135 0,153 0,063 0,180 0,128 0,071 0,035 0,022 0,116 0,198 0,062 0,052 12 0,900 0,904 0,746 0,876 0,888 0,872 0,879 0,874 0,868 0,883 0,827 0,870 0,816 0,878 0,600 13 0,169 0,261 0,212 0,203 0,372 0,063 0,109 0,379 0,134 0,148 0,022 0,275 0,137 0,231 0,272 14 0,297 0,380 0,315 0,361 0,322 0,403 0,343 0,525 0,643 0,148 0,106 0,080 0,198 0,171 0,231 15 0,900 0,904 0,746 0,737 0,888 0,872 0,879 0,742 0,868 0,664 0,827 0,870 0,637 0,753 0,686 16 0,609 0,530 0,437 0,457 0,888 0,872 0,732 0,654 0,455 0,314 0,580 0,547 0,338 0,497 0,272 17 0,085 0,126 0,092 0,013 0,045 0,018 0,076 0,226 0,134 0,035 0,022 0,013 0,079 0,036 0,020 18 0,085 0,191 0,150 0,203 0,322 0,334 0,144 0,226 0,134 0,009 0,022 0,192 0,198 0,231 0,363 19 0,140 0,126 0,038 0,103 0,153 0,063 0,109 0,192 0,041 0,035 0,022 0,080 0,198 0,088 0,085 20 0,900 0,756 0,872 0,646 0,730 0,872 0,732 0,654 0,571 0,664 0,424 0,624 0,637 0,497 0,529 21 0,551 0,904 0,535 0,737 0,630 0,571 0,638 0,423 0,643 0,664 0,827 0,481 0,421 0,604 0,272 22 0,263 0,191 0,092 0,135 0,080 0,334 0,044 0,128 0,102 0,035 0,106 0,321 0,338 0,263 0,192 23 0,900 0,904 0,872 0,737 0,888 0,872 0,879 0,742 0,868 0,883 0,827 0,870 0,816 0,878 0,413 24 0,085 0,038 0,150 0,072 0,080 0,161 0,044 0,098 0,134 0,009 0,022 0,080 0,137 0,201 0,272 25 0,900 0,904 0,746 0,876 0,888 0,872 0,879 0,654 0,868 0,883 0,580 0,481 0,637 0,878 0,529 26 0,230 0,298 0,245 0,169 0,322 0,161 0,076 0,379 0,102 0,393 0,302 0,232 0,637 0,604 0,192 27 0,008 0,010 0,012 0,013 0,012 0,018 0,044 0,041 0,012 0,009 0,022 0,013 0,079 0,012 0,085 28 0,008 0,038 0,012 0,013 0,012 0,018 0,013 0,041 0,041 0,009 0,022 0,046 0,025 0,036 0,020 29 0,199 0,225 0,245 0,072 0,116 0,063 0,180 0,098 0,202 0,062 0,106 0,423 0,338 0,142 0,085 30 0,453 0,475 0,279 0,737 0,552 0,272 0,442 0,379 0,405 0,118 0,827 0,870 0,517 0,604 0,413 31 0,140 0,066 0,065 0,103 0,080 0,111 0,044 0,069 0,202 0,035 0,106 0,046 0,025 0,036 0,052 32 0,085 0,096 0,150 0,103 0,153 0,214 0,217 0,226 0,202 0,118 0,106 0,080 0,137 0,062 0,085 33 0,230 0,096 0,150 0,135 0,276 0,480 0,217 0,261 0,071 0,035 0,022 0,192 0,025 0,201 0,119 34 0,370 0,338 0,279 0,240 0,630 0,272 0,343 0,226 0,134 0,090 0,106 0,192 0,265 0,115 0,052 35 0,332 0,530 0,245 0,240 0,322 0,214 0,390 0,160 0,134 0,090 0,106 0,370 0,198 0,171 0,363 36 0,900 0,904 0,872 0,876 0,730 0,480 0,879 0,874 0,868 0,883 0,827 0,870 0,637 0,878 0,686 37 0,551 0,380 0,535 0,407 0,425 0,687 0,563 0,337 0,238 0,210 0,198 0,232 0,338 0,201 0,192 38 0,112 0,126 0,121 0,169 0,153 0,272 0,217 0,128 0,134 0,035 0,022 0,046 0,198 0,115 0,119 39 0,033 0,010 0,038 0,013 0,012 0,111 0,013 0,069 0,102 0,009 0,022 0,046 0,079 0,062 0,085 40 0,112 0,066 0,092 0,072 0,080 0,161 0,044 0,261 0,012 0,035 0,022 0,046 0,025 0,062 0,020 41 0,332 0,338 0,393 0,457 0,233 0,687 0,390 0,654 0,733 0,210 0,106 0,080 0,198 0,201 0,272 42 0,500 0,426 0,353 0,361 0,485 0,571 0,499 0,584 0,455 0,393 0,580 0,192 0,265 0,332 0,155 43 0,033 0,010 0,012 0,013 0,045 0,063 0,013 0,069 0,041 0,009 0,022 0,046 0,025 0,088 0,085 44 0,297 0,475 0,279 0,457 0,372 0,403 0,442 0,298 0,041 0,353 0,424 0,153 0,137 0,201 0,119
167
168
Appendix (continued) Locus
Population Y1
Y2
Y3
Y4
Y5
Y6
Y7
O1
O2
O3
O4
O5
O6
O7
O8
45 0,453 0,591 0,484 0,407 0,425 0,480 0,499 0,423 0,360 0,484 0,198 0,275 0,338 0,263 0,192 46 0,008 0,010 0,012 0,013 0,080 0,063 0,044 0,041 0,041 0,035 0,022 0,013 0,025 0,012 0,020 47 0,008 0,010 0,012 0,013 0,012 0,018 0,013 0,014 0,012 0,009 0,022 0,013 0,025 0,012 0,052 48 0,008 0,010 0,012 0,013 0,012 0,018 0,013 0,014 0,041 0,009 0,022 0,013 0,025 0,012 0,020 49 0,008 0,010 0,012 0,135 0,012 0,111 0,180 0,014 0,041 0,009 0,022 0,013 0,079 0,012 0,052 50 0,370 0,475 0,279 0,361 0,630 0,872 0,390 0,423 0,405 0,437 0,580 0,275 0,198 0,409 0,192 51 0,677 0,664 0,661 0,574 0,630 0,872 0,638 0,742 0,571 0,595 0,827 0,370 0,338 0,604 0,316 52 0,008 0,010 0,038 0,072 0,045 0,161 0,013 0,041 0,071 0,090 0,022 0,046 0,025 0,062 0,272 53 0,033 0,010 0,065 0,013 0,012 0,161 0,013 0,041 0,041 0,009 0,022 0,013 0,025 0,142 0,231 54 0,169 0,096 0,180 0,169 0,116 0,272 0,390 0,261 0,238 0,009 0,022 0,232 0,265 0,088 0,155 55 0,140 0,475 0,315 0,278 0,045 0,111 0,390 0,379 0,277 0,484 0,424 0,275 0,265 0,497 0,119 56 0,059 0,096 0,038 0,042 0,012 0,063 0,013 0,192 0,041 0,243 0,198 0,046 0,079 0,263 0,052 57 0,677 0,756 0,746 0,876 0,730 0,872 0,638 0,584 0,571 0,664 0,580 0,721 0,421 0,753 0,231 58 0,764 0,664 0,535 0,876 0,630 0,687 0,638 0,742 0,571 0,751 0,580 0,721 0,421 0,878 0,272 59 0,609 0,756 0,872 0,574 0,730 0,872 0,732 0,654 0,571 0,664 0,827 0,481 0,421 0,753 0,272 60 0,199 0,380 0,437 0,407 0,552 0,687 0,298 0,298 0,238 0,314 0,580 0,370 0,198 0,369 0,468 61 0,140 0,191 0,353 0,203 0,153 0,334 0,298 0,226 0,238 0,118 0,198 0,116 0,137 0,332 0,192 62 0,332 0,475 0,315 0,407 0,485 0,571 0,343 0,471 0,202 0,393 0,198 0,370 0,338 0,547 0,155 63 0,169 0,066 0,038 0,278 0,372 0,214 0,109 0,160 0,317 0,035 0,022 0,046 0,338 0,297 0,020 64 0,033 0,096 0,038 0,240 0,485 0,111 0,109 0,098 0,238 0,437 0,198 0,423 0,079 0,297 0,085 65 0,085 0,225 0,038 0,135 0,153 0,111 0,076 0,098 0,102 0,278 0,106 0,046 0,338 0,088 0,052 66 0,008 0,038 0,012 0,042 0,012 0,018 0,013 0,069 0,012 0,009 0,106 0,013 0,137 0,036 0,020 67 0,008 0,010 0,012 0,013 0,012 0,018 0,044 0,014 0,012 0,009 0,022 0,013 0,025 0,012 0,052 68 0,008 0,010 0,038 0,013 0,012 0,018 0,013 0,014 0,012 0,009 0,022 0,046 0,025 0,036 0,052 69 0,008 0,010 0,038 0,135 0,012 0,063 0,044 0,014 0,041 0,009 0,022 0,013 0,025 0,036 0,085 70 0,008 0,038 0,012 0,013 0,012 0,018 0,044 0,069 0,012 0,009 0,022 0,013 0,025 0,012 0,020 71 0,008 0,066 0,038 0,013 0,012 0,018 0,013 0,098 0,012 0,035 0,022 0,013 0,025 0,297 0,052 72 0,033 0,066 0,065 0,072 0,012 0,063 0,109 0,128 0,012 0,090 0,106 0,013 0,079 0,115 0,052 73 0,008 0,096 0,038 0,013 0,012 0,161 0,044 0,041 0,167 0,009 0,022 0,046 0,025 0,088 0,119 74 0,112 0,038 0,180 0,042 0,080 0,063 0,144 0,098 0,202 0,035 0,022 0,013 0,025 0,036 0,020 75 0,900 0,904 0,746 0,876 0,888 0,872 0,879 0,874 0,868 0,664 0,827 0,870 0,816 0,753 0,600 76 0,370 0,530 0,535 0,361 0,730 0,571 0,499 0,379 0,643 0,178 0,302 0,481 0,338 0,297 0,468 77 0,900 0,904 0,872 0,876 0,888 0,872 0,879 0,874 0,868 0,883 0,827 0,870 0,816 0,878 0,686 78 0,900 0,904 0,746 0,876 0,888 0,872 0,879 0,874 0,868 0,883 0,827 0,870 0,816 0,753 0,686 79 0,677 0,904 0,593 0,457 0,485 0,872 0,732 0,584 0,733 0,353 0,302 0,547 0,421 0,409 0,363 80 0,900 0,904 0,872 0,876 0,888 0,872 0,879 0,742 0,868 0,883 0,827 0,870 0,816 0,878 0,600 81 0,900 0,904 0,746 0,876 0,888 0,872 0,879 0,874 0,868 0,883 0,827 0,870 0,816 0,753 0,600 82 0,764 0,904 0,746 0,876 0,888 0,872 0,879 0,742 0,733 0,751 0,827 0,870 0,517 0,604 0,686 83 0,609 0,904 0,593 0,407 0,730 0,687 0,563 0,423 0,455 0,437 0,827 0,721 0,421 0,369 0,468 84 0,764 0,904 0,661 0,876 0,888 0,687 0,879 0,654 0,733 0,751 0,827 0,721 0,517 0,547 0,803 85 0,900 0,904 0,746 0,737 0,730 0,872 0,638 0,584 0,643 0,751 0,580 0,870 0,816 0,753 0,600 86 0,764 0,904 0,661 0,876 0,888 0,872 0,879 0,742 0,868 0,484 0,827 0,870 0,816 0,753 0,600 87 0,140 0,158 0,245 0,203 0,485 0,571 0,257 0,226 0,238 0,148 0,302 0,423 0,265 0,142 0,231 88 0,169 0,380 0,180 0,407 0,233 0,272 0,217 0,261 0,071 0,314 0,198 0,370 0,517 0,451 0,231
169
Appendix (continued) Locus
Population Y1
Y2
Y3
Y4
Y5
Y6
Y7
O1
O2
O3
O4
O5
O6
O7
O8
89 0,677 0,298 0,437 0,407 0,630 0,214 0,298 0,742 0,202 0,278 0,198 0,275 0,517 0,369 0,316 90 0,764 0,904 0,661 0,737 0,888 0,687 0,879 0,742 0,868 0,751 0,827 0,870 0,637 0,878 0,686 91 0,140 0,096 0,121 0,013 0,192 0,063 0,044 0,160 0,202 0,210 0,106 0,192 0,079 0,062 0,085 92 0,900 0,904 0,661 0,876 0,888 0,872 0,879 0,874 0,868 0,664 0,827 0,721 0,637 0,753 0,529 93 0,112 0,096 0,150 0,169 0,192 0,403 0,217 0,098 0,134 0,090 0,022 0,080 0,265 0,036 0,119 94 0,263 0,380 0,315 0,318 0,485 0,480 0,563 0,298 0,238 0,210 0,106 0,321 0,198 0,201 0,192 95 0,199 0,096 0,315 0,278 0,425 0,687 0,298 0,226 0,455 0,278 0,198 0,370 0,338 0,231 0,155 96 0,764 0,904 0,661 0,876 0,888 0,687 0,879 0,742 0,868 0,664 0,827 0,870 0,637 0,670 0,600 97 0,008 0,010 0,065 0,042 0,012 0,161 0,013 0,041 0,012 0,009 0,022 0,046 0,079 0,088 0,231 98 0,085 0,096 0,065 0,042 0,080 0,111 0,013 0,192 0,041 0,210 0,022 0,153 0,137 0,263 0,272 99 0,169 0,225 0,245 0,407 0,425 0,403 0,390 0,226 0,405 0,148 0,106 0,321 0,265 0,171 0,155 100 0,008 0,010 0,012 0,013 0,012 0,018 0,013 0,014 0,012 0,035 0,022 0,013 0,025 0,062 0,192 101 0,008 0,010 0,038 0,013 0,012 0,018 0,013 0,041 0,041 0,009 0,106 0,013 0,079 0,012 0,020 102 0,008 0,010 0,038 0,072 0,012 0,161 0,013 0,069 0,102 0,009 0,022 0,046 0,079 0,142 0,231 103 0,059 0,010 0,012 0,042 0,080 0,272 0,180 0,069 0,041 0,035 0,022 0,080 0,137 0,036 0,020 104 0,033 0,010 0,038 0,013 0,012 0,063 0,013 0,014 0,012 0,009 0,022 0,013 0,079 0,012 0,085 105 0,085 0,038 0,038 0,103 0,045 0,063 0,109 0,041 0,041 0,090 0,022 0,013 0,079 0,012 0,119 106 0,085 0,096 0,038 0,103 0,012 0,063 0,013 0,041 0,041 0,009 0,022 0,116 0,025 0,036 0,052 107 0,410 0,380 0,535 0,278 0,276 0,571 0,180 0,192 0,360 0,090 0,198 0,481 0,265 0,171 0,316 108 0,059 0,261 0,150 0,203 0,012 0,111 0,144 0,261 0,202 0,243 0,198 0,232 0,198 0,369 0,119 109 0,033 0,010 0,038 0,013 0,045 0,018 0,013 0,014 0,041 0,062 0,022 0,046 0,025 0,012 0,155 110 0,033 0,096 0,038 0,013 0,012 0,063 0,013 0,098 0,041 0,090 0,106 0,013 0,079 0,142 0,119 111 0,169 0,380 0,212 0,240 0,322 0,272 0,109 0,379 0,041 0,278 0,302 0,275 0,421 0,369 0,363 112 0,085 0,191 0,121 0,072 0,153 0,272 0,044 0,160 0,102 0,210 0,106 0,153 0,137 0,201 0,231 113 0,112 0,038 0,121 0,135 0,045 0,018 0,044 0,041 0,012 0,009 0,106 0,013 0,079 0,036 0,020 114 0,140 0,530 0,353 0,407 0,485 0,687 0,442 0,379 0,405 0,353 0,424 0,624 0,338 0,409 0,192 115 0,008 0,038 0,012 0,013 0,012 0,111 0,044 0,098 0,012 0,009 0,022 0,013 0,025 0,036 0,020 116 0,008 0,010 0,012 0,013 0,012 0,018 0,013 0,041 0,012 0,118 0,106 0,116 0,079 0,088 0,119 117 0,033 0,010 0,065 0,013 0,012 0,063 0,013 0,014 0,012 0,035 0,022 0,013 0,079 0,036 0,052 118 0,764 0,756 0,484 0,737 0,730 0,687 0,638 0,471 0,733 0,595 0,827 0,721 0,421 0,547 0,529 119 0,059 0,066 0,038 0,278 0,116 0,063 0,180 0,041 0,041 0,035 0,106 0,013 0,079 0,036 0,020 120 0,008 0,038 0,012 0,013 0,012 0,018 0,013 0,014 0,012 0,035 0,022 0,013 0,079 0,036 0,085 121 0,008 0,096 0,065 0,203 0,045 0,018 0,180 0,128 0,041 0,062 0,106 0,046 0,079 0,088 0,052 122 0,059 0,191 0,065 0,318 0,116 0,018 0,257 0,128 0,041 0,090 0,106 0,153 0,137 0,201 0,052 123 0,900 0,904 0,746 0,876 0,888 0,872 0,879 0,874 0,868 0,751 0,827 0,870 0,816 0,878 0,803 124 0,900 0,904 0,746 0,737 0,730 0,872 0,732 0,874 0,868 0,883 0,827 0,547 0,816 0,878 0,600 125 0,677 0,904 0,593 0,876 0,730 0,872 0,442 0,654 0,643 0,664 0,827 0,624 0,816 0,670 0,686 126 0,410 0,380 0,393 0,574 0,425 0,571 0,499 0,471 0,360 0,243 0,424 0,275 0,517 0,409 0,413 127 0,609 0,591 0,484 0,574 0,485 0,687 0,499 0,654 0,317 0,278 0,580 0,481 0,637 0,409 0,529 128 0,764 0,591 0,593 0,512 0,630 0,687 0,563 0,742 0,360 0,484 0,580 0,423 0,816 0,497 0,600 129 0,900 0,904 0,746 0,876 0,888 0,687 0,879 0,874 0,868 0,751 0,827 0,870 0,816 0,878 0,803 130 0,900 0,904 0,872 0,876 0,888 0,872 0,879 0,874 0,868 0,751 0,827 0,870 0,816 0,878 0,600 131 0,677 0,904 0,661 0,574 0,888 0,687 0,563 0,874 0,868 0,664 0,827 0,721 0,816 0,753 0,529 132 0,453 0,475 0,437 0,512 0,425 0,571 0,499 0,654 0,360 0,664 0,580 0,275 0,637 0,451 0,413
169
170
Appendix (continued) Locus
Population Y1
Y2
Y3
Y4
Y5
Y6
Y7
O1
O2
O3
O4
O5
O6
O7
O8
133 0,764 0,904 0,661 0,512 0,888 0,872 0,732 0,874 0,868 0,664 0,198 0,870 0,637 0,878 0,529 134 0,900 0,904 0,872 0,737 0,730 0,872 0,879 0,742 0,733 0,751 0,302 0,624 0,816 0,878 0,803 135 0,263 0,664 0,661 0,407 0,630 0,571 0,390 0,525 0,510 0,484 0,580 0,547 0,816 0,753 0,529 136 0,112 0,158 0,315 0,278 0,153 0,111 0,390 0,261 0,317 0,178 0,302 0,116 0,198 0,171 0,155 137 0,500 0,475 0,437 0,457 0,425 0,480 0,499 0,584 0,277 0,353 0,302 0,547 0,338 0,332 0,272 138 0,900 0,756 0,661 0,646 0,552 0,687 0,732 0,654 0,360 0,883 0,827 0,624 0,517 0,604 0,363 139 0,900 0,904 0,872 0,876 0,888 0,872 0,732 0,874 0,868 0,751 0,827 0,870 0,816 0,878 0,529 140 0,900 0,904 0,872 0,737 0,888 0,687 0,879 0,874 0,643 0,883 0,827 0,870 0,816 0,753 0,529 141 0,900 0,904 0,661 0,876 0,730 0,687 0,638 0,742 0,643 0,751 0,580 0,870 0,637 0,878 0,686 142 0,609 0,664 0,484 0,646 0,425 0,571 0,499 0,584 0,360 0,595 0,424 0,481 0,517 0,547 0,316 143 0,900 0,904 0,661 0,737 0,888 0,687 0,638 0,742 0,510 0,751 0,827 0,870 0,816 0,753 0,686 144 0,677 0,756 0,279 0,646 0,485 0,480 0,638 0,525 0,238 0,537 0,302 0,370 0,517 0,604 0,231 145 0,609 0,530 0,484 0,407 0,322 0,571 0,563 0,584 0,405 0,484 0,580 0,721 0,816 0,670 0,316 146 0,551 0,530 0,437 0,646 0,630 0,334 0,390 0,423 0,405 0,393 0,302 0,370 0,265 0,604 0,363 147 0,900 0,904 0,746 0,737 0,888 0,687 0,732 0,742 0,571 0,751 0,827 0,721 0,816 0,878 0,600 148 0,677 0,756 0,484 0,512 0,485 0,687 0,442 0,584 0,277 0,537 0,424 0,481 0,421 0,753 0,231 149 0,410 0,591 0,315 0,318 0,372 0,403 0,563 0,379 0,238 0,178 0,302 0,370 0,338 0,263 0,231 150 0,677 0,261 0,437 0,574 0,276 0,571 0,638 0,584 0,510 0,353 0,580 0,275 0,637 0,604 0,468 151 0,677 0,904 0,535 0,646 0,730 0,687 0,638 0,874 0,643 0,751 0,580 0,624 0,816 0,753 0,600 152 0,230 0,591 0,393 0,278 0,425 0,272 0,442 0,525 0,202 0,595 0,106 0,232 0,079 0,297 0,363 153 0,230 0,298 0,245 0,407 0,153 0,334 0,563 0,337 0,167 0,393 0,302 0,232 0,198 0,369 0,085 154 0,551 0,664 0,535 0,361 0,630 0,571 0,343 0,525 0,510 0,664 0,580 0,624 0,816 0,670 0,468 155 0,059 0,010 0,065 0,042 0,045 0,063 0,076 0,069 0,041 0,090 0,022 0,046 0,025 0,036 0,085 156 0,263 0,158 0,212 0,135 0,233 0,480 0,144 0,041 0,238 0,178 0,302 0,275 0,198 0,231 0,119 157 0,008 0,010 0,092 0,072 0,080 0,063 0,013 0,041 0,012 0,009 0,022 0,046 0,079 0,088 0,052
